No.  71
        1,l-Sichlorethylene


  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone 'scrutiny to
ensure its technical accuracy.

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                             1,1-DICHLOROETHYLENE
                                   Summary  .
     Ambient levels  of 1,1-dichloroethylene have  not been determined.   The
primary effect  of  acute and chronic  occupational  exposure to  1,1-dichloro-
ethylene is depression  of  the  central nervous system.  In  experimental  ani-
mals, both  liver and kidney damage have  been  noted after  exposure,  regard-
less of the route of administration.   1,1-Oichloroethylene  has  been  shown to
be a  mutagen  in bacterial  systems and  a carcinogen in  mice.   Both  kidney
adenocarcinomas and  mammary adenocarcinomas were 'produced  after exposure to
1,1-dichloroethylene  by  inhalation.    No  teratogenic effects  have been  ob-
served.
     For  freshwater  fish,   the   reported  96-hour  LC5Q  values  range  from
73,900  to  108,000  ug/1  1,1-dichloroethylene.   Reported 48-hour EC5Q  values
for  Daphnia magna  range  from 11,600  to 79,GOO ug/1.   96-Hour LC,-n  values
       —•        - -                                                  >u
of over 22^,000 ug/1 have  been  observed  for  saltwater  fish  and  inverte-
brates.  An embryo-level test with  freshwater  fish  resulted  in  an  adverse
effect occurring at  2,800 ug/1.   Algae,  both  fresh  and  saltwater,  apparently
are  not  affected   by  concentrations  of  1,1-dichloroethylene  as  high  as
716,000 ug/1.
                                 7/-3

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                             1,1-OICHLOROETHYUENE
I.   INTRODUCTION
     This  profile  is based  on the  Ambient  Water Quality  Criteria Document
for Dichloroethylenes (U.S. EPA, 1979a).
     1,1-Dichloroethylene  (C2H2C12;  molecular  weight  96.95)  is  a  clear
colorless liquid used as a chemical  intermediate  in  the synthesis of methyl-
chloroform  and  in  the production  of  polyvinylidene chloride  copolymers
(PVCDs).   Prior  to 1976, annual  production  of 1,1-dichloroethylene  was ap-
proximately  120,000 metric  tons  (Arthur 0.  Little,  Inc.t  1976).   1,1-Qi-
chloroethylene has  the  following physical/chemical  properties:   water solu-
bility  of 2,500 ug/ml,  vapor  pressure  591  mm Hg,  and  a melting  point of
-122.1°C.   For  more  general  information  regarding  the  dichloroethyienes,
the  reader  is referred to the EPA/ECAO  Hazard Profile  on Dichioroethyler.es
(U.S. EPA, 1979b).
II.  EXPOSURE
     A.  Water
         The  National  Organic Monitoring Survey  (U.-S. EPA,  1978a)  reported
detecting 1,1-dichloroethylene in  finished drinking  waters;  however,  neither
the amount nor the occurrence was quantified.
     8.  Food
         Pertinent data could  not be located  in  the available literature on
the  ingestion of  1,1-dichloroethylene  in  foods.  The  U.S.  EPA  (1979a)  has
estimated  the weighted, bioconcentration  factor  for 1,1-dichloroethylene- to
be 6.9  for the edible portions of fish and  shellfish  consumed by Americans.
                                                    *•
This estimate was  based on  the  octanol/water partition coefficient  of 1,1-
dichloroethylene.                                                     '

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     C.  Inhalation
         The population  at  risk due to vinylidene  chloride  exposure is com-
posed  primarily  of workers in  industrial  or commercial  operations manufac-
turing or using  it.  Airborne  emissions  of vinylidene chloride are not like-
ly to  pose  a significant risk  to  the general  population.   Emissions during
production,  storage,  and transport can  be controlled by methods  similar to
those planned for control of vinyl chloride (Hushon and Kornreich,  1978).
III.  PHARMACOKINETICS
     A.  Absorption
         Specific  data  en  the  absorption  of dichloroethylenes  are unavail-
able.  However,  a  recent study by McKenna,  et  al.  (1978b)  suggests  that in
rats  most,  if not  all,  of the orally administered dcss is  absorbed  at two
dose levels:  1 and 50 mg/kg.
     3.  Distribution
         Distribution of 1,1-dichlorcethylene v;ss  studied  in rats follcv;ir.g
inhalation  (Jaeger, et  al.  1977).   The largest  concentrations  were found in
kidney,  followed by  liver, spleen,  heart,  and  brain,  and  fasting  made no
difference  in  the  distribution  pattern.   At  the  subcellular  level 1,1-di-
chloroethylene or  its metabolites  appear  to bind  to rnacromolecules  of the
iTiicrosomes and mitochondria  (Jaeger, et al.  1977).   There is also  some asso-
ciation with the lipid fraction.
     C.  Metabolism
         In  the  intact  animal,  a large portion  of  the systemically absorbed
1,1-dichloroethylene  is  metabolically converted, with 36  percent appearing
in the  urine of rats within 26 hours (Jaeger,   et  ai.  1977).  The essential

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feature of 1,1-dichloroethylene metabolism  is  the  presence  of epoxide inter-
mediates, which are  reactive  and  may form covalent bonds with  tissue macro-
molecules (Henschler, 1977).  In  rats  and mice,  covalently  bound metabolites
ars found in  the  kidney and liver  (McXenna,  et  al. 1978b).   Interaction  of
1,1-  dichloroethylene  with the microsomal  mixed function oxidase  system  is
not clear, since both inhibitors  (dithiocarbamate)  and  inducers (phenobarbi-
tal)  decreased the  toxic effects  of  the  compound  (Anderson and- Jenkins,
1977;  Reynolds, et  al.  1975; Jenkins,  et  al. 1972).   However,  Carlson  and
Fuller (1972)  reported  increased  mortality  from  1,1-dichloroethylene in rats
following phenobarbital pretreatment.   There  is  evidence  that  the  1,1-di-
chloroethylene metabolites are conjugated with glutathione,  which presumably
represents a detoxification step  (McXenna,  et al. 197Sa).
     0.  Excretion
         It  is speculated  that  1,1-dichloroethylene  has  a  rapid  rate  of
elimination,  sir.ce a substantial  fraction of the total absorbed  oose may  be
recovered in  the  urine  within 26 to 72 hours  (Jaeger,  et al.  1977; McKanna,
et al. 1978a).  Also, disappearance  of  covalently  bonded metabolites of 1,1-
dichloroethylene  (measured  as TCA-insoluble fractions) appears  to  be fairly
rapid, with a reported half-life  of 2 to 3 hours  (Jaeger,  et al. 1977).
IV.  EFFECTS
     A.  Carcinogenicity
         1,1-Oichloroethylene has  been shown  to produce kidney  adenocarci-
nomas in male  mice  and mammary adenocarcinomas  in female mice  upon inhala-
tion  of  100  mg/m3  (Maltoni,  1977;  Maltoni,  et  al. 1977).   In  similar  ex-
periments with  Sprague-Oawley rats exposed  up to  800 mg/m  ,  no  significant
increase in tumor incidence  was  noted.  Also, hamsters  exposed  to  t'he  same
                                    71-6

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conditions as  the  mice failed to exhibit  an  increased tumor incidence (Mai-


toni, et al. 1977).   In  rats  exposed  to 1,1-dichloroethylene in their drink-


ing  water  (200 mg/1), there  was  no evidence of  increased  tumors (Rampy, et


al.  1977).'  There  was an increased incidence of  mammary tumors  in  rats  re-


ceiving 20  mg  of  1,1-dichloroethylene  by  gavage 4  to 5 days  a  week for 52


weeks.  The  incidence was 42 percent in the  treated animals and  34 percent


in the  controls;  however,  the data was  not  analyzed statistically (Maltoni,


et al. 1977).

     B.  Mutagenicity


         1,1-Dichloroethylene has  been  shown to  be  mutagenic  in  S^ typhimu-


rium  (Bartsch,  et  al. 1975) and E^ coli K12  (Greim, et al. 1975).   In both


systems,  mutagenic activity  required  microsomal activation.   In mammalian


systems,  1,1-dichloroethylene was  negative  in  the  dominant  lethal  assay


(Short, et al.  1977b; Andersen,  et al. 1977).


     C.  Teratcgenicity


         A  study  by  Murrary, et  al.  (1979)  failed to  shcv/ tsratogenic  af-


fects  in  rats  or  rabbits  inhaling  concentrations cf  up to 160  ppm 1,1-di-


chloroethylene  for 7  hours per  day or  in  rats  given drinking  water contain-


ing 200 ppm  1,1-dichloroethylene.

     0.  Other Reproductive Effects


         Pertinent data could not be located in the available literaure.


     E.  Chronic Toxicity


         In  animal studies,  liver  damage is  associated with exposure, either

in  the air  or water, to  1,1-dichloroethylene  (6 jdg/m   or 0.79  pg/1)  with


transitory damage  appearing as vacuolization  in liver  cells.  In  both guinea
                                                                      »
pigs  and  monkeys,  continuous exposure  to 1,1-dichloroethylene produced  in-

creased mortality,  while intermittent exposure  to the same  concentration in

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air produced no  increase  in mortality (U.S..EPA, 1979a).  Less attention has
been  paid  to the  renal  toxicity of  1,1-dichloroethylene  despite the occur-
rence  of histologically  demonstrated damage at  exposures equal  to or less
than  those  required for  hepatotoxicity  (Predergast,  et al.  1967; Short,  at
al. 1977a).
     F.  Other Relevant Information
         Alterations in  tissue glutathione concentrations affect  the hepato-
toxicity of '1,1-dichloroethylene,  with  decreased  tissue  glutathione   asso-
ciated  with greater  toxicity and  elevated glutathione associated  with de-
creased toxicity (Jaeger, et al. 1973,1977).
V.   AQUATIC TOXICITY
     A.  Acute Toxicity
         Dill, et  al.  calculated, for  the fathead  minnow, Pimephales prome-
las,  96-hour  1C--,  values  of  169,000  jug/1   using   static   tachnicues and
1CS,CCO ug/1 using flow-through  tasts with measured concentrations.   The re-
ported  96-hour LC--, value  for the bluegill, Leoomis  mactochirjs, is 73,900
ug/1  in a  static  test (U.S.  EPA,  1978b).  Two  48-hour  tests  with Daphnia
maona  resulted  in  ECcg values of  11,600  and  79,000  jjg/1,   respectively
(Dill,  et  al.;   U.S.  EPA, 1978b).  The  96-hour LC5Q  values  for  the sheeps-
heao  minnow,  Cyorinodon  varieoatus,  and  the   tidewater  silversioe, Menidia
beryllina,  are 249,000 and  250,000 ug/1,  respectively (U.S. EPA,   1978b; Daw-
son,  et  al.  1977).   The  96-hour  LC^,,   for   the  mysid  shrimp,  Mysidoosis
bahia, is reported to be  224,000 ug/1 (U.S. EPA, 1978b).
     B.  Chronic Toxicity
         An  embryo-larval test  with  the  fathead minnow resulted in  no ad-
                                                                      •
verse effects occurring  at 2,800 /jg/1, the highest test concentration  (U.S.
EPA, 1978b).

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     C.  Plant Effects

         The  96-hour  EC--  value based  on  cell  numbers  of  the  freshwater

alga, Selenastrum capricornutum, is reported to be  greater than  798,000 ug/1

(U.S.  EPA,-1978b).   The  effective  concentration of  1,1-dichloroethylene  on

the  saltwater  alga,  Skeletonema costatum,  was observed  to be  712,CGO  ug/1

(U.S. EPA, 1978b).

     D.  Residues

         Pertinent data could not be located in the available  literature.

VI.  EXISTING GUIDELINES AND STANDARDS

     A.  Human

         The  American  Conference  of   Governmental   Industrial  Hygienists

(ACGIH,  1977)  threshold  limit  value (TLV)  for  1,1-dichloroethylene  is  40

mg/m ,  with  calculated daily exposure  limits  of 286  mg/day.   1,1-Oichloro-

ethylene is  suspected  of  being  a human  carcinogen;  and using the  "one-hit"

model,  the U.S.  EPA  (1979a) has estimated levels of  l,l-dichlcrcethyler:e  in

ambient water which will result  in soecified risk levels of human cancer:

Exposure Assumptions         Risk Levels with Corresponding Draft Criteria
     (per day)
                                      10-7           ig-6           ig-5

2 liters of drinking water         0.013 ug/1      0.13-.ug/1      1.3 jjg/1
and consumption of 18.7
grams fish and shellfish.

Consumption of fish and            0.21  ug/1       2.1 ug/1       21 jug/1
shellfish only.


     8.  Aquatic

         For  1,1-dichloroethylene,  the   drafted  criterion to protect  fresh-

water  aquatic  life  is  530  ug/1 as a 24-hour average,  not to  exceed  1,200
                                                                     »
/jg/1 at  any  time.  No  saltwater  criterion has been proposed because of in-

sufficient data.
                                   7'-?

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                             1,1-OICHLOROETHYLENE

                                  REFERENCES
American Conference  of Governmental  Industrial  Hygienists.   1977.  Documen-
tation of the threshold limit values.  3rd ed.

Anderson, 0.,  et al.   1977.   Dominant  lethal studies with  the  halogenated
olefins  vinyl  chloride and vinylidene  dichloride in male CD-I mice.   Envi-
ron. Health Perspect.  21: 71.

Anderson, M.E.  and  L.J.  Jenkins,  Jr.   1977.  Enhancement  of  1,1-dichloro-
ethylene hepatotoxicity by  pretreatment with  low  molecular  weight epoxides.
Proc. Soc. Toxicol.  41.

Arthur D. Little,  Inc.  April, 1976.   Vinylidene Chloride monomer emissions
from the monomer, polymer,  and  polymer  processing industries, Arthur 0. Lit-
tle, Inc., for the U.S. Environ. Prot. Agency, Research Triangle Park, N.C.

Bartsch,  H.,   et  al.   1975.   Tissue-mediated  mutagenicity  of  vinyiidene
chloride and 2-chlorobutadiene in Salmonella tyohirnurium.  Nature.  255: 641.

Carlson, G.P.  and  G.C. Fuller.   1972.   Interactions  of  modifiers of hepatic
microscmal drug  metabolism  and  the inhalation toxicity of 1,1-dichloroethyl-
ene.  Res. Comm. Chem. Pathol. Pharmacoi.  4:  553.

Dawson,  G.w.,  et al.  1977.   The acute toxicity  of  47 industrial chemicals
to fresh and saltwater fisn.es.  Jour. Hazard Mater.   1: 303.

Oiil,  O.C.,  at al.   Toxicity of  1,1-dicnlcroethylene  (vinylidene chloride)
to aquatic organisms.  Dow Chemical Co.   (Manuscript)

Greim, H.,  et  al.    1975.   Mutagenicity  in vitro  and potential carcinogeni-
city of  chlorinated  ethylenes as  a function  of  metabolic oxirane formation.
Siochem. Pharmacoi.  24: 2013.

Henschlar, D.   1977.   Metabolism  and  mutagenicity of halogenated  olefins - A
comparison of structure and activity.  Environ'. Health Perspect.   21: 61.

Hushon,  J. and  M.  Kornreich.   1978.  Air  pollution  assessment of vinylidene
chloride.  EPA-450/3-78-015.  U.S. Environ. Prot. Agency, Washington, O.C.

Jaeger,  R.J.,  et al.  1973.  Diurnal variation of hepatic  glutathione con-
centration and its correlation  with  1,1-dichloroethylene  inhalation toxicity
in rats.  Res.  Comm. Chem. Pathol. Pharmacoi.  6: 465..

Jaeger,  R.L.,  et al.   1977.   1,1-Oichloroethylene hepatotoxicity:  Proposed
mechanism of  action  of  distribution and  binding of ^C-radioactivity^ f0j__
lowing inhalation exposure in rats.  Environ. Health  Perspect.  21: 113!

Jenkins, L.I., et al.  1972.   Biochemical  effects of 1,1-dichlorpethylene in
rats: Comparison with carbon tetrachloride and  1,2-dichloroethylene.   Toxi-
col. Appl. Pharmacoi.  23: 501.

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Maltoni,  C.   1977.   Recent  findings  on  the. carcinogenicity  of chlorinated
olefins.  Environ. Health Perspect.  21: 1.

Maltoni,  C.,  et  al.   1977.   Carcinogenicity bioassays  of  vinylidene chlor-
ide.  Research plan and early results.  Med. Lav.  68: 241.
McKenna,  M.J.,   et  al.   1978a.    The  pharmacokinetics  of  (14C)  vinylidene
chloride  in  rats following  inhalation  exposure.  Toxicol.  Appi.  Pharmacol.
45: 599.

McKenna,  M.J.,  et  al.   1978b.  Metabolism  and pharmacokinetics  profile of
vinylidene chloride  in rats  following  oral administration.   Toxicol.  Appl.
Pharmacol.  45: 821.

Murray, F.J.,  et al.  1979.   Embryotoxicity and fetotoxicity  of  inhaled or
ingested  vinylidene  chloride  in  rats  and  rabbits.   Toxicol:  Appl.  Pharma-
col.  49: 189.

Prendergast,   J.A.,  et al.   1967.   Effects  on experimental  animals  of long-
term inhalation of  trichloroethylene,  carbon tetrachloride, 1,1,1-trichloro-
ethane,  dichlbrodifluoromethane,   and  1,1-dichloroethylene.   Toxicol.  Appl.
Pharmacol.  10: 270.

Rampy,  L.W.,  et  al.  1977.   Interim  results  of  a two-year toxicological
study in  rats of vinylidene  chloride  incorporated  in the  drinking  water or
administered by repeated inhalation.  Environ. Health Perspect.  21~: 33.

Reynolds, E.S.,  et al.   1975.  Hepatoxicity  of vinyl  chloride  and  1.1-di-
chioroethylene.  Am. Jour. Pathol.  81: 219.

Short,  R.D.,  et al.   1977a.  Toxicity of   vinylidene  cnloride in mice  and
rats  and its alteration  by  various  treatments.   Jour.  Toxicoi.  Environ.
Health.  3:  913.

Short,  R.D.,  et  al.   1977b.  A dominant  lethal  study in  male rats after re-
peated  exposures  to vinyl  chloride  or  vinylidene chloride.   Jour.  Toxicol.
Environ, health.  3: 965.

U.S.  EPA.   1973a.   Statement of basis and  purpose  for  an  amendment  to cne
National interim primary  drinking  water regulations  on  a  treatment technique
for  synthetic organics.   Off. Drinking  Water.  U.S. Environ.  Prot.  Agency,
Washington,  D.C.

U.S. EPA.   1978b.   In-depth  studies on  health  and  environmental  impacts of
selected  water pollutants.   Contract  No.  68-01-4646,  U.S.  Environ.  Prot.
Agency.

U.S.  EPA.   1979a.   Dichloroethylenes:  Ambient Water Quality  Criteria Docu-
ment. (Draft)

U.S. EPA.   1979b.   Environmental  Criteria and Assessment Office.   Dichloro-
ethylenes: Hazard Profile.   (Draft)

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                                      No.  72
     trana-l,2-Oichloroethylene


  Health and environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracv.

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                  TRANS-1,2-DICHLQRETHYLENE




                           SUMMARY




     There is little specific information available on trans-



1,2-dichloroethylene.   This  compound is  quantitatively  less



toxic  than  the   1,1-dichloroethylene  isomer;  however,   the



toxicity  appears  qualitatively  the  same  with  depression



of  the  central nervous  system  as well  as liver  and kidney



damage.    Trans-l,2-dichloroethylene  has  been   shown  to  be



a  mutagen in  bacterial  systems.   The  teratogenicity  and



carcinogenicity of this compound have not been evaluated.



     In  the  only aquatic  study  reported,  the   observed  96-



hour LCeg value for  the bluegill  is  135,000 pg/1 in  a static



bioassay.
                             7*-}

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                  TRANS-1,2-DICHLORETHYLENE


I.   INTRODUCTION


     This  profile  is  based  on  the  Ambient  Water  Quality


Criteria Document for Dichloroethylenes  (U.S. EPA, 1979).


     Trans-1,2-dichloroethylene   (traans  1,2-DCE;  C-tUCl-;


molecular weight  96.95)  is  a clear  colorless liquid.   Since


the early  1960's trans-l,2-dichloroethylene  has  had  no wide



industrial usage  (Patty, 1963).   Trans-l,2-dichloroethylene


has  the  following  physical/chemical  properties:     water


solubility of  6,300 ug/ml,  a  vapor  pressure of  324  mm Hg,


and a melting point of -50°C (Patty,  1963).



II.  EXPOSURE


     A.   Water


          Trans-l,2-dicnloroethylene  was found at a  concen-


tration of 1 ug/1 in Miami  drinking water  (U.S.  EPA,  1975,


1978) .


     B.   Food


          Pertinent data could  not be  located  in the avail-


able literature on the ingestion of trans-1,2-dichloroethylene


in  foods.   The U.S. EPA (1979)  has  not  estimated a  biocon-


centration factor for trans-1,2-dichloroethylene.


     C.   Inhalation


          Pertinent  information  could  not  be   located  in



the available literature.
                          7 a.-4/
                         - V/J*
                          "^5^7

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III. PHARMACOKINETICS
     A.   Absorption
          Animal  or human  studies  do  not  appear  to exist
which  specifically  document the  degree of  systemic  absorp-
tion of trans-l,2-dichloroethylene by any route.
     B,   Distribution
          Pertinent data  could  not be  located  in the avail-
able literature.
     C.   Metabolism
          Trans-l,2-dichloroethylene  is metabolized  through
an  epoxide  intermediate   to  either  a  dichloroacetaldehyde
or  monochloroacetic acid   (Liebman  and Ortiz,  1977).   The
epoxide  intermediate  which  is  reactive,  may  form covalent
bonds  with  tissue  macromolecules  (Henschler,  1977).   Meta-
bolism  of  the  cis-isomer  relative  to  the  amount  tak'en up
by the liver was much greater than the trans-isomer  (McKenr.a,
et al. 1977).
     D.   Excretion
          Pertinent data  could  not be  located  in the avail-
able literature.
IV.  EFFECTS
     A.   Carcinogenicity
          Pertinent data  could  not be  located  in the avail-
                                            *
able literature.
     B.   Mutagenicity
                                                           »
          Trans-l,2-dichloroethylene  has  been  shown to be
negative  in  the  E^  coli  K12  and  Salmonella  mutagenicity
assays  (Greim, et al.  1975; Cerna and Kypenova, 1977).

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     C.   Teratogenicity and Other Reproductive Effects
          Pertinent  information  could  not  be  located  in
the available literature.
     D.   Chronic Toxicity
          Although  little  data  is  available  specifically
on trans-1,2-dichloroethylene,  it appears  that chronic expo-
sure  results in  kidney  and  liver  damage  similar   to  that
noted with  1,1-dichloroethylene (U.S.  EPA,  1979).   Jenkins,
et al. (J.972) found  trans-l,2-dichloroethylene  to be consider-
ably less potent than 1,1-dichloroethylene.
V.   AQUATIC TOXICITY
     A.   Acute Toxicity
          The reported  96-hour  LC^Q value  for  the bluegill,
Lepomis  macrochirus,   exposed   to   1,2-dichloroethylene  is
135,000 ug/1 (U.S. EPA, 1979)  in a static test procedure.
     B.   Chronic Toxicity, Plant Effects and Residues
          Pertinent  information  could  not  be  located  in
the available literature.
VI.  EXISTING GUIDELINES AND STANDARDS
     A.   Human
          The American  Conference of Governmental Industrial
Hygienists  (ACGIH,   1977)  threshold  limit  value  (TLV)  for
1,2-dichloroethylene  is  790  mg/m  ,  with  calculated  daily
exposure limits of  5,643  mg/day.   The U.S.  EPA  (1979)  draft
Water Quality  Criteria Document for  Dichloroethylene stat.es
that  human  health  criterion could not  be  derived  due  to
the lack of sufficient data on which to base a criterion.
                          74-6

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     B.   Aquatic
          Guidelines  do not  exist  for  salt  water  species
because of  insufficient data.    The  draft criterion  to pro-
tect freshwater  aquatic  life  is 530 ^g/1 as a  24-hour  aver-
age and not to exceed 1200  pq/1 at any time (U.S. EPA, 1979).

-------
                 TRANS 1,2-DICHLOROETHYLENE

                         REFERENCES

American Conference of Governmental Industrial Hygienists.
1977.  Documentation of the threshold limit value.  3rd. ed.

Cerna, M. , and H. Kypenova.  1977.  Mutagenic activity of.
chloroethylenes analyzed by screening system tests.  Mutat.
Res.  46: 214.

Greim, H., et al.  1975.  Mutagenicity in vitro and potential
carcinogenicity of chlorinated ethylenes as a function of
metabolic oxirana formation.  Biochem. Pharmacol. 24: 2013.

Henschler, D.  1977.  Metabolism and mutagenicity of halo-
genated olefins - A comparison of structure and activity.
Environ. Health Perspect.' 21: 61.

Jenkins, L.J., et al.  1972.  Biochemical effects of 1,1-di-
chloroethylene in rats: Comparisons with carbon tetrachloride
and 1,2-dichloroethylene.  Toxicol. Appl. Pharmacol. 23: 501.
Leibman, K.C., and E. Ortiz.  1977.  Metabolism of halogen-
ated ethylenes.  Environ. Health Perspect. 21: 91.

McKenna, M.J., et al.  1977.  The pharmacokinetiLcs of  i^c]
vinylidene chloride  in rats following inhalation exposure.
Toxicol. Appl. Pharmacol. 45: 599.

Patty. F.A.  1963.  Aliphatic halogenated hydrocarbons.  Ind.
Hyg. Tox. 2: 1307.

U.S. EPA.  1975.  Preliminary assessment of suspected  carcin-
ogens in drinking water.  Rep. to Congress.  Off. Toxic
Subst.  U.S. Environ. Prot. Agency, Washington, D.C.

U.S. EPA.  1978.  List of organic compounds identified in
U.S. drinking water.  Health Effects Res. Lab.  U.S. Environ.
Prot. Agency, Cincinnati, Ohio.

U.S. EPA.  1979.  Dichloroethylenes: Ambient Water Quality
Criteria. (Draft).

-------
                                  No. 73
         Dlchloroethylanes


  Health and Environmental Effaces
U.S. ENVIRONMENTAL PROTECTION AGENCY
      WASHINGTON, D.C.  20460

          APRIL 30, 1980
       73-1

-------
                          DISCLAIMER
     This report represents a survey of the.potential health
and environmental hazards from exposure  to  the  subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources   and  available  reference documents.
Because of the limitations of such sources,  this short profile
may not reflect  all  available information  including all the
adverse health  and  environmental  impacts  presented  by the
subject chemical.  This  document has undergone  scrutiny to
ensure its technical  accuracy.
                        73-3-

-------
                            DICHLOROETHYLENES




                                 Summary




     Of the three dichloroethylene  isomers,  cis 1,2-dichloroethylene,




trans 1,2-dichloroethylene,  and 1,1-dichloroethylene,  only the 1,1-dichloro-




ethylene isomer is produced in large quantities.   Most of the health




effects information available is related to  the 1,1-dichloroethylene




isomers; however, qualitatively the toxicity of the  1,2-dichloroethylene




isomers appears to be similar, with depression of the  central nervous




system and liver and kidney damage.   Of the  three isomers, 1,1-dichloro-




ethylene is the most toxic.   Both 1,1-dichloroethylene and trans  1,2-




dichloroethyler.e are autagenic in bacterial  systems.   Only 1,1-dichloro-




ethylene has been shown to be a carcinogen.




     All of the available aquatic data, with one  exception,  are for 1,1-




dichlcroethyiene. Reported 96-hour  [,€50 values for the bluegill are 73,900




and 135,500 ug/1, respectively, for 1 , 1-dicnlorcehtylene and  1,2-di-



chloroethylene. Two observed 46-hcur ^€50 values for  Daphnia exposed to



1,1-dichloroethylene range were 11,600  and 79,000 ug/1.   All  saltwater



fish and invertebrates tasted with  1,1-dichioroethylene showed 96-hour



LC5Q values over 224,000 ug/1, and  all  algae tested  both in fresh  and



saltwater, had 96-hour EC^g values  (based on cell numbers) of 716,000




and over.  In the only reported chronic study,  no adverse effects  were



observed at the highest test concentration of 2,800  ug/1 for  fathead




minnows exposed to 1,1-dichloroethylene.
                               73-3

-------
                            DICHLOROETHYLENES




I.   INTRODUCTION




     This profile is based on the draft Ambient Water Quality Criteria




Document for Dichloroethylenes (U.S.  EPA,  1979).




     The dichloroethylenes (^2^2^2'i  molecular weight 96.95)  consist of




the three isomers:  1,1-dichloroethylene,  cis- 1,2-dichloroethylene, and




trans-l,2-dichloroethylene.  Dichloroethylenes are clear colorless liquids




with water solubilities between 2,500 and  6,300 ug/1, vapor pressures

                                                v

between 591 and 208 mm Hg, and melting points between -50°C and -122°C




(U.S. EPA, 1979).  The 1,1-dichloroethylene isomer is the most extensively




used in industry, with annual production prior to  1976  of approximately




120,000 metric tons (Arthur D. Little, Inc.,  1976).   The 1,1-dichloroethylene




isomer is used as a chemical intermediate  in the synthesis of methylchloroform




ana in the production of pclyvinylidene chloride copciymers (FVDCs).



II.  EXPOSURE



     I.
     A.   water



          The National Organic Monitoring  Survey (U.S. EPA, 1973a) reported




detecting 1,1-dichloroethylene in finished drinking waters; however,




neither the amount nor the occurrence was  quantified.  Both cis and trans-




1,2-dichlcroethylene were found at concentrations  of 16 and 1 ug/1,




respectively, in Miami drinking water (U.S.  EPA, 1975, 1978b).




     B.   Food




          Pertinent data could not be located on the ingestion of dichloro-




ethylene in foods.  The U.S.  EPA (1979) has estimated the weighted bioconcen-




tration factor for 1,1-dichloroethylene to be 6.9  for the edible portions of
                               73-/

-------
fish and shellfish consumed by Americans.  This estimate is based on the



octanol/water partition coefficients of  1,1-dichloroethylene.  There is no



estimate for a bioconcentration factor for the other iscrners.



     C.   Inhalation



          The population at risk due to vinylidene chloride exposure is composed



primarily of workers in industrial or commercial operations manufacturing or



using it.  Airborne emissions of vinylidene chloride are not likely to pose a



significant risk to the general population.  Emissions during production,



storage, and transport can be controlled by methods similar to those planned



for control of vinyl chloride (Hushon and Kornreich, 1978)
     •


III. PHARMACOKINETICS



     A.   Absorption



          Specific data on the absorption of dichloroethylenes are unavailable.



However, a recent study by McXenna, et al. (1973b) suggests that in rats most,



if not all, of the orally administered dose is absorbed at two dose levels: 1



and 50 35/kg.



     B.   Distribution



          Distribution of 1,1-dichloroethylene was studied in rats following



inhalation (Jaeger, et al. 1977).  The largest concentrations were found in



kidney, followed by livsr, spleen, heart, ar.d brain; and fasting made no



difference in the distribution pattern.  At the subcellular level 1,1-dichloro-



ethylene or its metabolites appear to bind to macromolecules of the microsomes



and mitochondria (Jaeger, et al. 1977).  There is also some association with



the lipid fraction.  Distl,2-dichloroethylene isomers, are not available.



     C.   Metabolism



          The essential feature of all dichloroethylene metabolism is the



presence of epoxide intermediates which are reactive and may form covalent

-------
bonds with tissue raacromolecules (Henschler, 1977).  In rats and mice,

covalently bound metabolites of 1,1-dichloroethylene are found in the

kjdney and liver (McKenna, et al. 1978b).  Interaction of dichloroethylenes

with the microsomal mixed function oxidase system is not clear,' since

both inhibitors (dithiocar'oamate) and inducers (phenobarbital) decreased

the toxic effects of 1,1-dichloroethylene (Anderson and Jenkins, 1977;

Reynolds, et al. 1975; Jenkins, et al. 1972).  Carlson and Fuller (1972),

however, reported increased mortality from 1,1-dichloroethylene in rats

following phenobarbital pretreatment.  There is evidence that the 1,1-

dichloroethylene metabolites are conjugated with gluthathione, which

presumably represents a detoxification step  (McKenna, et al. 1978b).

     B.   Excretion

          The only information available on elimination pertains to the

1,1-dichlorcethylene isc~er.  It is postulated that the 1,1-dichloro-

ethylane isorner has a rapid rate of elimination since a substantial

fraction of the total absorbed dose may be recovered in urine within 26

to 72 hours (Jaeger, et al. 1977; McKenna, et al. 1978a).  Also, dis-

appearance of covalently bonded metabolites of 1,1-dichloroethylene

(measured as TCA-insoluble fractions) appears to be fairly rapid, with a

reported half-life of 2 to 3 hours (Jaeger, et al. 1977).

IV.  EFFECTS

     A.   Carcinogenicity

          There is only data on the carcinogenicity of the 1,1-dichloro-

ethylene isomer.  This isomer has been shown to produce kidney adeno-

carcinomas in male mice and mammary- adenocarcinomas in female mice upon
                                                                    «
inhalation of 100 mg/m3 (Maltoni, et al. 1977; Maltoni, 1977).  In

-------
similar experiments with Sprague-Dawley rats exposed as high as 800 mg/m^ f



no significant increase in tumor incidence was noted.  Hamsters exposed



to the same conditions as the mice failed to exhibit an increased tumor



incidence-(Maltoni, et al. 1977).  In rats exposed to 1,1-dichloroethylene



in their drinking water (200 mg/1) there was no evidence of increased



tumors (Rampy, et al. 1977).  There was an increased incidence of mammary



tumors in rats receiving 20 mg of 1,T-dichloroethylene by gavage U to 5



days a week for 52 weeks.  The incidence was 42 percent in the treated



animals and 31* percent in the controls; however, the data was not analyzed



statistically (Maltoni, et al. 1977).



     B.   Mutagenicity



          1,1 -Dichloroethylene has been shown to be mutagenic in S_._ typhimuriuj



(Bartsch, et al. 1975) and Z^ coli K12 (Greim, et al. 1975); however,



both the cis and trans iscaers of 1,2-dichioroethylene were non-mucagenic



when assayed with £_._ coli K"2.  In order to demonstrate mutagenic activity,



1,1-dichloroethylene needed microsomal activation.  In addition, cis



1,2-dichloroethylene was tnutagenic in Salmonella tester strains, and



promoted chromosomal aberrations in cytogenic analysis of bone marrow



cells (Cerna and Kypenova, 1977).  In mammalian systems,  1,1-dichloroethylene



was negative in the dominant lethal assay (Short, et ai.  1977b; Anderson,



et al. 1977).



     C.   Teratogenicity



          A study by Murray, et al. (1979) failed to show teratogenic

                                                  ,-
effects in rats or rabbits inhaling concentrations of up to 160 ppm 1,1-di-



chloroethylene for 7 hr/day or in rats given drinking water containing
                                                                    »


200 ppm 1,1-dichloroethylene.
                                 ? 1-

-------
     D.   Other Reproductive Effects



          Pertinent data could not be located in the available literature.



     E.   Chronic Toxicity



          In- animal studies, liver damage is associated with exposure



either in the air or water, to dichloroethylenes (6 mg/m3 or 0.79  mg/1)



with transitory damage appearing as vacuolization in liver cells  (U.S.



EPA, 1979).  Jenkins,  et al. (1972) found both cis and trans 1,2-dichloro-



ethylene to be considerably less potent than 1,1-dichloroethylene  as  a


hepatotoxin.  Less attention has been paid to the renal toxicity of the



dichloroethylenes despite the occurrence of histologically demonstrated



damage at 1,1-dichloroethylene exposures equal to or less than those



required for hepatoxicity (Prendergast, et al. 1967; Short,  et al.  1977a).



     F.   Other Relevant Information



          Alterations in tissue glutathione concentrations affect  the



hepatotoxicity of 1,1-dichloroethylene, with decreased tissue slutathicr.e



associated with greater toxicity and elevated gluthachione associated



with decreased toxicity (Jaeger, et al. 1973, 1977).



V.   AQUATIC TOXICITY



     A.   Acute Toxicity



          All of the available data for dichloroethyiene,  with one exception,


are for 1,1-dichloroethylene.  The data on acute static tests with bluegill,



Lepomis macrochirus,  under similar conditions show a correlation between


the degree of chlorination and toxicity.  The 96-hour LC5Q values  for the



bluegill are 73,900 and 135,000 ug/1 for 1,1- and 1,2-dichloroethylene,


respectively.  Additional data for other ethylene chlorides are as follows:  44,700
                                                                      »

ug/1 for trichloroethylene, and 12,900 ug/1 for tetrachloroethylene (U.S.



EPA, 1978c).  These results indicate an increase in the lethal effect on


bluegills with an increase in chlorine content.
                                73-?

-------
     The 96-hour LC5Q value for the sheepshead minnow, Cypuimocen variegatus,


tidewater silverside, Menidia beryllina, and mysid shrimp, Mysidepsis


behia, following exposure to 1,1-dichloroethylenes are all over 224,000


ug/1 (U.S. EPA, 1978c).


     B.   Chronic Toxicity


          In the only reported chronic study, an embryo-larval test  in


fathead minnows, no adverse effects were observed at the highest test


concentration of 1,1-dichloroethylene, 2800 _ug/l (U.S. EPA,  1979).


     C.   Plant Effects


          The 96-hour £050 values based on cell numbers of the freshwater


algae, Salenestr-un: capriecrr.utur. and the saltwater algae,  Sksletcns-a


costatum, are 798,000 and 712,000 ug/1, respectively, for exposure to


1,1-dichloroethylsne (U.S. EPA,  1978c).


     D.   Residues


          Pertinent information could not be located in the  available


literature.


VI.  EXISTING GUIDELINES AND STANDARDS


     A.   Human


          The American Conference of Governmental Industrial Hygienists


(ACGIH, 1977) threshold limit values (TLV) are 40 mg/m3 (1,1-dichloro-


ethylene) and 790 mg/ni3 (1,2-dichloroethylene).  These values allow daily


exposures of 286 ag 1,1-dichloroethylene per day and 5,643 eg 1,2-di-


chloroethylene per day.  The U.S.  EPA (1979) draft water criteria document
                                                 f

for dichloroethylene states that no human health criterion could be derived


for cis- and trans-l,2-dichloroethylene due  to the lack of sufficient


data on which to base a criterion.  1,1-dichloroethylene is  suspected of

-------
being a human carcinogen, and using the "one-hit" model, the U.S. EPA

(1979) has estimated levels of 1,1-dichloroethylene in ambient water

which will result in specified risk levels of human cancer:
Exposure Assumptions
    (per day)

2 liters of drinking water
 and .consumption of 18.7
 grams fish and shellfish

Consumption of fish and
 shellfish only.
 Risk levels and Corresoondine Draft Criteria
      10'
10
                  -6
    0.013 ug/1  0.13 ug/1


    0.11  ug/1  2.1  ug/1
  IP"3

•1.3 ug/1


21   ug/1
     B.   Aquatic

          The -proposed draft criterion to protect freshwater species

from dichlcroethylene toxicity are as follows (U.S. EPA._ 1979):
Compound


1,1-dichloroethylene
1,2-dichloroethylene

For saltwater species:

1,1-dichioroethylene
1,2-dichloroethylene
2*4-hr.  Average

   530  us/I
   620  ug/1
 1,700 ug/1
Not available
      Concentration not to be
      exceeded at anytime

             1,200 us/1
             1,400 ug/1
             3,900 ug/1
            Not available

-------
                               DICHLOROETHYLENES

                                  References

American  Conference  of  Governmental  Industrial  Hygenists.   1977.   Docu-
mentation of.the threshold limit values.  3rd ed.

Anderson, p.,  et al.   1977.   Dominant  lethal  studies with  the  halogenatea
olefins  vinyl  chloride  and   vinylidene   dichloride  in  male  CO-i  mice.
Environ. Health Perspect.  21: 71.

Anderson, M.E.  and  L.J.  Jenkins,  Jr.   1977.   Enhancement  of 1,1-dichloro-
ethylene hepatotoxicity  by  pretreatment with low molecular  weight epoxides.
Proc. Soc. Toxicol.  41.

Arthur  D.  Little,   Inc.  April,  1976.   Vinylidene chloride  monomer emissions
from  the monomer,  polymer,   and  polymer  processing  industries,  Arthus  D.
Little,  Inc.,  for  the  U.S.  Environ.  Prot.  Agency,  Research  Triangle Park,
N.C.

Bartsch,  H.,  'et al.,   1975.   Tissue-mediated  mutagenicity   of  vinylidene
chloride and 2-chiorobutadiene in Salmonella tv^himurii:". . Mature 255: 641.

Carlson, G.P. and  G.C.  Fuller.  1972.   Interactions  of modifiers  of hepatic
microsomal  drug metabolism  and  the   inhalation toxicity  of  1,1-dichloro-
ethylene.  Res. Comm. Chem.  Pathol. Pharmacol.   4:  553.

Carna,  M.,  and  H.  Xypenova.    1977.   The  acute toxicity  of  47  industrial
chemicals to fresn and saltwater fishes.  Jour. Hazsrc. Mater,  i: 3C3.

Dill, O.C., et  al.  Toxicity  of  1,1-dichlcroethylene  (vinylidene chloride)
to aquatic organisms.  Dow Chemical Co.  (Manuscript).

Greim,  H.,  et  al.   1975.    Mutagenicity  in  vitro  and potential  carcino-
genicity of chlorinated  ethylenes as  a  function  of metabolic  oxirane forma-
tion.  Biochem. Pharmacal.  24: 2013.

Henschler, D.   1977.  Metabolism  and  mutagenicity of halogenated  ol'efins - a
comparison of structure ana activity.   Environ. Health Fersoect.  21: 6i.

Hushon,  J.  and  M.  Kornreich.   1978.  Air pollution assessment of vinylidene
chloride.  EPA-480/3-78-015.  U.S. Environ.  Prot. Agency, Washington, D.C.  "

Jaeger,  R.J.   1973.   Diurnal  variation  of  hepatic  glutathions concentration
and  its  correlation with 1,1-dichloroethylene  inhalation toxicity  in rats.
Res. Comm. Chem. Pathol. Pharmacol.  6:  465.

Jaeger,  R.L., et  al.  1977.   1,1-Dichloroethylene  hepatotoxicity:   Proposed
mechanism of  action of  distribution  and binding of ^C  radioactivity  fol-
lowing inhalation exposure in  rats.  Environ. Health Perspect.   21: 113,

-------
Jenkins, L.J., et  al.   1972.   Biochemical effects of 1,1-dichloroethylene in
rats:   Comparison  with  carbon  tetrachloride   and  1,2-dichloroethylene.
Toxicol. Appl. Fharmacol.  23: 501.

Maltoni, C.   1977.   Recent  findings  on the carcinogenicity  of chlorinated
olefins.  Environ. Health Perspect.  21: 1.

Maltoni,  C.,   et  al.   1977.    Carcinogenicity  bioassays  of  vinylidene
chloride.   Research plan and early results.  Med. Law.  63: 241.

McKenna,  M.J..  et  al.   1978a.   The  pharmokinetics  of  [14C]  vinylidene
chloride in  rats  following  inhalation, exposure.  Toxicol.  Appl.  Pharmacol.
45: 599.

McKenna,  M.J.,  et  al.   1978b.   Metabolism  and  pharmokinetic profile  of
vinylidene chloride  in rats  following oral administration.   Toxicol. Appl.
Pharmacol.  45: 821.

Murray, F.J.,  et al.   1979.   Embryotoxicity and fetotoxicity  of  inhaled or
ingested   vinylidene  chloride   in    rats   and   rabbits.    Toxicol.   Appl.
Pharmacol.  49: 139.

Prendergast,   J.A.,  et al.   1967.  Effects on experimental  animals  of long-
term inhalation of trichloroethylane,  carbon tetrachloride, 1,1,1-trichioro-
ethans, dichlorodifluorcmethane,   and  1,1-dichloroethylene.   Toxicol. Appl.
Pharmacol.  10: 270.

Rsmpy,   L.W.,   et  =1.   1977.   Interim  results  of  a tv/o-yesr   toxicologies!
study  in rats of vir,yiidene chloride  incorporated  in the  drinking  water or
administered by repeated inhalation.    Environ. Health Psrspect.  21: 33.

Reynolds.    E.S.,   et   al.    1975.    Hepatoxicity  of  vinyl   chloride  and
1,1-dichloroethylene.  Am. Jour. Pathol.  81: 219.

Short,   R.D.,  et al.   1977a.   Toxicity  of vinylidene chloride in  mice  and
rats  and   its alteration  by  various  treatments.   Jour,  Toxicol.  Environ.
Health  3: 913.

Short,   R.O.,  et  al.   1977b.   A  dominant  lethal study  in male  rats after
repeated exposure  to  vinyl  chloride  or  vinylidene chloride.   Jour.  Toxicoi.
Environ. Health  3: 965.

U.S.   EPA.    1975.    Preliminary   assessment  of  suspected  carcinogens  in
drinking water,  Rep.  to  Congress.  Off.  Toxic  Subst.  U.S.  Environ. Prot.
Agency, Washington, D.C.

U.S. EPA.   1978a.  Statement  of  basis  and  purpose  'for  an  amendment  to  the
National interim primary drinking  water  regulations  on a  treatment technique
for  synthetic organics.  Off.  Drinking  Water.   U.S. Environ.  Prot.   Agency,
Washington, O.C.

U.S. EPA.   1978b.   List  of  organic  compounds  identified in  U.S.  drinking
water.   Health Effects  Res.  Lab. U.S.  Environ.  Prot.   Aoency,  Cincinnati,
Ohio.

-------
U.S.  EPA.   1978c.   In-depth studies  on  health  and  environmental  impacts of
selected  water  pollutants.  Contract NO. 68-01-4646.   U.S.  Environ.  Prot.
Agency.

U.S.   EPA.    1979.   Dichloroethylenes:   Ambient  Water  Quality  Criteria.
(Draft).
                              73-'?

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                                         SJ-46-04
                                         DRAFT
                                         10-21-80
                                       No.  74
          Dichloromethane
  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY

      WASHINGTON, B.C.  20460


          October 30, 1980

-------
                            DISCLAIMER
     This report represents a survey of the potential health and
environmental hazards from exposure to the subject chemical.  The
information contained in the report is drawn chiefly from secondary
sources and available reference documents.  Because of the limi-
tations of such sources this short profile may not reflect all
available information including all the adverse health and
environmental impacts presented by the subject chemical.  This
document has undergone scrutiny to ensure its technical accuracy.
                               74-2

-------
                      DICHLOROMETHANE (DCM)




                             SUMMARY






     In humans, DCM is a central nervous system depressant




resulting in narcosis at high concentrations, and impaired task




performance.  Dichloromethane is metabolized to carbon monoxide




and causes an increase in carboxyhemoblogin, placing persons with




cardiovascular disease, and perhaps those-who are pregnant, at




increased risk of disease.  On the basis of present evidence, DCM




cannot be firmly identified as an animal or human carcinogen.




DCM has been shown to be mutagenic to Salmonella, but not to




S.  cerevisia and Drosophi'la, and causes cell transformation.




     Aquatic organisms are fairly resistant to dichloro-




methane, with acute toxicity values ranging from 193,000 to 331,000




ug/1.
                               74-?

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                          DICHLOROMETHANE









 I.<   INTRODUCTION




      This  profile  is  based  on  the  Ambient Water  Quality  Criteria




 Document for  Ha'lomethanes (U.S.  EPA,  1979a).




      Dichloromethane  (CH2C12,  methylene chl'oride, methylene




 dichloride, and methylene bichloride; molecular  weight 84.93)  is




 a  colorless liquid with  a melting  point of -95.18C,  a boiling




 point of 40°C, a. specific gravity  of  1.327 g/ml  at 20°C, a vapor
                                             e



 pressure of 362.4 mm  Hg  at  20*C, and  a solubility in water of




 13.2  g/1 at 25°C.  Dichloromethane is a common industrial solvent




 found in insecticides, metal cleaners, -paints, and paint and




 varnish removers (Balmer, et al.,  1976).  In 1976, 244,129 metric




 tons  were  imported (U.S.  EPA,  1977).  For additional information




 regarding  the halomethanes  as  a  class, the reader is referred  to




 the Hazard Profile on Halomethanes (U.S. EPA, 1979b).




 II.   EXPOSURE




      A.   Water




          The U.S. EPA (1975)  has  identified dichloromethane in




 finished drinking waters  in the U.S.  in R of 83  sites, with a




maximum level of 0.007 mg/1 and a median of less than 0.001




mg/1.  The dichloromethane  in  drinking water is  not  a product  of




water chlorination (U.S. EPA,  1975; Morris and McKay, 1975).   In




the national organics monitoring survey, dichloromethane was




detected in 15 of 109 sites, with a mean concentration (positive




results only)  of 0.0061 mg/1 (U.S.  EPA,  1978).






                               74-4

-------
      B.    Food


           Pertinent  information  could not be  located  in the


available  literature.


      C.    Inhalation


           Reported background  concentrations  of dichloromethane


in both continental  and  saltwater atmospheres were about 0,0001?.


mg/m^, and urban  air concentrations ranged from less  than 0.00007


to 0.00005 mg/m^.  Local indoor  concentrations can be high due to


the use of aerosol sprays or solvents (Natl.  Acad. Sci., 1978).


III.  PHARMACOKINETICS


      A.    Absorption


           Efficiences of absorption of dichloromethane by the


lungs are  between 30 to  75 percent, depending on length of


exposure,  concentration, and activity level (Natl. Acad. Sci.,

                                                             I
1978; Natl. Inst. Occup. Safety  a.nd Health, 1976).


      B.    Distribution


          Upon inhalation and  absorption, dichloromethane levels


increase rapidly in  the blood  to equilibrium  levels that depend


primarily upon atmosphere concentration (Natl. Acad.  Sci.,  1978).


Carlsson and Hultengren (1975) reported that  dichloromethane


and its metabolites  were in highest concentrations in white


adipose tissue, followed in descending order by levels in brain


and liver.


     C.   Metabolism


          Dichloromethane is metabolized to carbon monoxide.


Some of this carbon monoxide is  exhaled, but a significant  amount



                               74-5

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 is  involved  in  the  formation  of  carboxyhemoglobin  (Natl.  Inst.




 Occup. .Safety and' Health,  1976).   Cardiorespiratory  stress  from




 elevated  carboxyhemoglobin may be  greater  as  a  result  of




 dichloromethane exposure  than from exposure  to  carbon  monoxide




 alone due to the continued formation  of  carbon  monoxide  following




 cessation'of dichloromethane  exposure (Stewart  and Hake,  1976).




 As  shown  by  animal  experiments,  other possible  human metabolites




 of  dichloromethane  include carbon  dioxide, formaldehyde,  and




 formic acid  (Natl.  Acad.  Sci., 1978).



     D.    Excretion




           A  large proportion  of  absorbed dichloromethane  is ex-




 creted unchanged, primarily via  the lungs, with some in  the urine.




 DiVincenzo,  et  al.  (1972)  have reported  that  about 40  percent of




 absorbed  dichloromethane  undergoes  some  reaction and decomposition




 process in the  body.



 IV.  EFFECTS




     A.    Carcinogenicity




          Friedlander et  al.  (1978) analyzed  the mortality of



 Eastman-Kodak male employees  exposed  to  low levels of  methylene




 chloride.  No significant  neoplastic  risk  factors were identified.




     Theiss and  coworkers  (1977) examined  the tumorigenic activity




 of dichloromethane in strain A mice.  Dichloromethane  at  the low




 dose (1:5 dilution of the maximum  tolerated dose) produced




marginally significant increases in tumor  response.  Shimkin and



 Stoner (1975) did not report a positive carcinogenic response for




 the strain A mouse bioassay system.






                               74-6

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        Although the data base is inadequate, there is a basis to




   suspect' the potential carcinogenicity of DCM based on the (marginally




   positive) pulmonary adenoma response in strain A mice, on positive




   responses for mutagenicity in the Ames test, and on the ability




   to transform rat embryo cells (see below).




        B.    Mutagenicity




             Simmon, et al.  (1977)  reported that dichloromethane




   tfas mutagenic to Salmonella typhimurium strain TA1QO when assayed




   in a dessicator whose atmosphere contained the test compound.




   Metabolic activation was  not required, and the number of revertants




   per plate was directly dose-related.  A linear dose response curve




   was observed.  Dichloromethane did not increase mitoti.c recombination




   in S_.  cerevisia D3 (Simmon, et al.,  1977), and it was reported




   negative  on testing for mutagenicity in Drosophila (Filippova, et




   al.,  1967).  Positive results for dichloromethane in the Ames




   assay  were recently confirmed by Jongen,  et al. (1978) with vapor




   phase  exposures (5,700 ppm) of strains TA98 and TA100.




        C.    Teratogenicity
4



             Schwetz et al.,  (1975)  showed that DCM can affect




   embryonal and fetal development  in rats and mice as evidenced by




   the increased incidence of  extra  sternebrae.   DCM also affects the




   development of  chick embryos,  causing  a 2  to 3-fold increase in mal-




   formation frequencies (Elsavaara  et  al.,  1979).




       D.    Other Reproductive Effects




             Gynecologic problems in femal workers exposed for
                                  74-7

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 long  period  to  gasoline  and  dichloromethane  vapors  were  reported




 by Vozovaya  (1974).   Also, inhalation  exposures  of  rats  and mice




 to vapor  levels  of  4,342 mg/m^  for  seven  hours daily  on  gestation




 days  6  to  15  produced evidence  of feto- or embryotoxicity  (Schwetz,




 et al., r975; Natl. Inst. Occup. Safety and  Health, 1976).




      E.    Chronic Toxicity




           Acute  exposures to  dichloromethane produce  central




 nervous system disfunction,  are irritating to mucous  membranes,




 and increase  the level of carboxyhemoglobin  (Natl.  Acad. Sci.,




 1978).  Price, et al.  (1978)  reported  that Fischer  rat embryo




 cells (F1706) were  transformed  by dichloromethane at  high




 concentrations (1.6 x 10"^M)  in the growth medium.  However,




 Sivak (1978)  indicated the presence of carcinogenic contaminants




 in the  dichloromethane and could not demonstrate transformation



 in the  BALB/C-3T3 assay system  with highly purified food grade




 dichloromethane.




 V.    AQUATIC TOXICITY




      A.    Acute Toxicity




           Acute toxicity values have been obtained  for two species




 of freshwater fish and one species of  freshwater invertebrates.




 LC5Q values for the fathead minnow (Pimephales promelas) ranged




 from 193,000 ug/1 in  a flowthrough assay to  310,000 ug/1 in a




 static assay.  An LC^Q value  of 224,000 ug/1 was obtained  for the




bluegill (Lepomis Macrochirus)  in a static assay.  Daphnia magna




were reported as having an "LC^Q value of 224,0^0 ug/1 (U.S. EPA,

-------
 1979a).   For  the  marine fish,  the  sheepshead  minnow  (Cyprinodon




 variegatus) ,  an 1650  of 331,000  ug/1  was  obtained.   The  marine




 mysid  shrimp  was  reported  as having an  LC5Q value  of 256,000




 ug/1.




     B.    Chronic Toxicity




           Chronic tests for freshwater  or marine species  could




 not  he located  in the available  literature.




     C .    Plant Effects




           Both  species of  freshwater  algae, Selenastrum  capricor-




 nortum  and marine  algae,  Skeletonema cornutum, were equally




 resistant  to  dichloromethane, with l>C$r\ values in  excess  of




 662,000 ug/1.




 VI.  EXISTING GUIDELINES AND STANDARDS




     Neither  the  human health nor the aquatic criteria derived by




 the U.S. .EPA  (1979a),  which are  summarized below, have gone




 through the process of public review, therefore, there is a




 possibility that  these criteria  will be changed.




     A.    Human




           OSHA  (1976)  has  established an eight-hour, time-weighted




 average for dichloromethane of 1,737 mg/m^; however, NIOSH (1976)




 has recommended a ten-hour, time-weighted average exposure limit




 of 261 mg/m^.   The U.S. EPA (1979a) draft water quality criterion




 for dichloromethane is  2 ug/1.    The reader is referred to the




Halomethanes Hazard Profile for  discussion of criteria derivation




 (U.S. EPA, 1979b).
                               74-9

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     B.«   Aquatic



          The criterion for protecting freshwater aquatic life has




been drafted as 4,000 ug/1, not to exceed 9,000 ug/1, while the




marine criterion has been drafted as 1,900 ug/1, not to exceed




4,400 ug/1.
                              74-10

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                          DICHLOROMETHANE
                             References
 Balmer, M.  F . •,  et  al.   1976.   Effects  in  the  liver  of
 methylene  chloride inhaled  alone  and with  ethyl  alcohol.
 Am.  Ind. Hyg. Assoc.  Jour.   37:345.

 Carlsson,.A., and  M.  Hultengren.   1975.   Exposure to
 methylene  chloride,  III.  Metabolism of 14C-labeled
 methylene  dichloride _in rat.  Scand. Jour.  Work Environ.
 Health  1:104.

 DiVincenzo, G.  D.,  et  al.   1972.  ^Hutnan and canine  exposures
 to methylene  chloride  vapor.   Am.  Ind. Hyg. Assoc.  Jour.
 33:125.

 Elovaara,  E., K. Hemminki,  and H.  Vaimio.  1979.  Effects of
 methylene  chloride,  trichloroethane, trichloroethylene and
 tolerance  on  the development  of chick  embryos.   Toxicology
 2:111-119.

 Filippova, L. M.,  et  al. 1967.  Chemical mutagens.  IV.
 Mutagenic  activity  of  geminal  system.  Genetika  8:134.

 Friedlander,  B. R., T. Hearne  and  S. Hall.  1978.  Epedemi-
 ologic investigation  of employees  chronically exposed to
 methylene  chloride.   J. Occup. Med.  20:657-666.

 Jongen, W.  M. F.,  et  al.  1978.  Mutagenic effect of di-
 chloromethane on Salmonella typhimurium.   Mutat. Res.  56:245.

 Morris, J.  C.,  and  G.  McKay.   1975.  Formation of halogenated
 organics by chlorination of water  supplies.  EPA 600/1-75-002.
 PB 241-511.  Natl.  Tech. Inf.  Serv., Springfield, Va.

 National Academy of Sciences.  1978.   Nonfluorinated halometh-
 anes in the environment.  Washington,  B.C.

 National Institute  of  Occupational Safety  and Health.  1976a.
 Criteria for a  recommended standard:   Occupational exposure
 to methylene chloride.  HEW Pub. NO. 76-138.   U.S. Dep. Health
Educ. Welfare,  Cincinnati, Ohio.

 Occupational Safety and Health Administration.   1976.  General
 industry standards.  OSHA 2206, revised January  1976.  U.S.
Dep.  Labor.  Washington, D.C.
                              74-11

-------
 Price,  P.  J.,  et  al.   1978.   Transforming  activities  of  tri-
 chloroethylene and  proposed  industrial  alternatives.   In Vitro
 14:290*.

 Schwetz, B.  A., et  al.   1975.   The  effect  of maternally  in-
 haled trichloroethylene,  perchloroethylene, methyl  chloro-
 form, and  methylene chloride  on embryonal  and  fetal develop-
 ment in mice and  rats.   Toxicol. Appl.  Pharmacol.   32:84-96.

 Shimkin, M.  B., and G. D.  Stoner.   1975.   Lung  tumors  in mice:
 application  to carcinogenesis bioassay.  A'dv.  Cancer  Res.  21:1

 Simmon, V. F.  et  al.   1977.   Mutagenic  activity  of  chemicals
 identified in  drinking water.   In;  S. Scott, et  al.,  eds.
 Progress in  Genetic Toxicology.

 Sivak, A.  1978.  BALB/C-3T3  neoplastic transformation
 assay with methylene chloride (food grade  test  specification).
 Rep. Natl. Coffee Assoc.,  Inc.

 Stewart, R.  D., and C. L.  Hake.  1976.  Paint  remover  hazard.
 Jour. Am. Med.  Assoc.  235:398.

 Theiss, J. C.,  et al.  1977.  Test  for  carcinogenicity of
 organic contaminants of United  States drinking waters  by
 pulmonary  tumor response  in strain  A mice.  Cancer  Res.
 37:2717.

 U.S. EPA.  1975.  Preliminary assessment of suspected  carcino-
 gens in drinking water, and appendices.  A report to  Congress,
 Washington,  D.C.

 U.S. EPA.  1977.  Area 1.  Task 2.  Determination of  sources
 of selected  chemicals in waters and amounts from these sources.
 Draft final  rep.  Contract No.  68-01-3852. Washington, D.C.

 U.S. EPA.  1978.  The National  Organic Monitoring Survey.
 Rep. (unpubl.).  Tech. Support  Div., Off. Water  Supple.
Washington,  D.C.

U.S. EPA.  1979a.    Halomethanes:  Ambient Water  Quality,
 (Draft).

U.S. EPA.  1979b.    Evironmental Criteria and Assessment
Office.   Halomethanes:   Hazard  Profile.  (Draft).

Vozovaya, M.  A.  1974.   Gynecological illnesses  in workers
of major industrial rubber products plants occupations.  Gig.
Tr. Sostoyanie Spetsificheskikh Funkts.  Tab.  Neftekhim.  Khim.
Prom-sti.  (Russian).56  (Abstract)
                              74-12

-------
                                   LB-45-01
                                   No.  75
       2,4 - Dichlorophenol

 Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, B.C. 20460

          October 1, 1980

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                            DISCLAIMER

     This report represents a survey of the potential health and
environmental hazards from exposure to the subject chemical.  The
information contained in the report is drawn chiefly from secondary
sources and available reference documents.  Because of the
limitations of such sources, this short profile may reflect all
available information impacts presented by the subject chemical.
This document has undergone scrutiny to ecsure its technical accuracy
                               75-2

-------
                         2, 4-DICHLOROPHENOL




                              Summary






      Insufficient data exist to indicate that 2,4-dichlorophenol




 is a carcinogenic agent.  2 ,4-Dichlorophenol appears to act as a




 nonspecific irritant in promoting tumors in skin painting studies.




 No information on mutagenicity, teratogenicity, or chronic toxicity




 is available.  In a subacute study, the only adverse effect noted




 in mice was microscopic nonspecific liver changes.  2,  4-




 Dichlorophenol app.ears to be a weak uncoupler of oxidative




 phosphorylation.




      Acute and chronic toxic effects of 2,4-dichlorophenol have




 been observed at  a concentrations as low as 2,200 and 365 ug/1




 respectively.  Mortality to early life stages of one species of




 fish occurs at 70 ug/1.  Flavor impairment  studies indicate'




 that the highest  concentrations of 2,4-dichlorophenol in water




 which would not cause tainting of the  edible portions of fish




 range from 0.4 to 14 ug/1 depending on the  species of fish




• consumed.
                                75-3

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                   2,4-DICHLOROHENOL (2,4 DCP)

I.  INTRODUCTION
    «
     This profile is in large part based on  the Ambient Water

Quality Criteria Document for 2,4-dichlorophenol (U.S. EPA,1980).

     2,4-Dichlorophenol is a colorless, crystalline solid having

the empirical formula CgH^^O and a molecular weight of 163.0

(Weast, 1975).  It has the following physical and chemical

properties (Sax, 1975; Aly and Faust, 1965; Weast, 1975; Kirk and

Othmer, 1964):

          Melting Point:                   45° C
          Boiling Point:                   210° C at 760 mm Hg
          Vapor Pressure:     :             1.0 mm Hg at 53.0° C
          Solubility:                      slightly soluble in water
                                           at neutral pH; dissolves
                                           readily in ethanol and
                                           benzene

     2,4-DCP is a commercially produced, substituted phenol used

entirely as an intermediate in the manufacture of industrial and

agricultural products such as the herbicide 2,4-dichlorophenoxyacetic

acid (2,4-D), germicides, and miticides.

     Little data exists regarding the persistence of 2,4-

dichlorophenol in the environment.  It is a product resulting

from degradation of many commercial products by plants, micro-

organisms, and sunlight.  Its low vapor pressure cause it to be

only slowly removed from surface water via volatilization (U.S.

EPA, 1980).  Studies have indicated low absorption of 2,4-DCP

from natural surface waters by various clays (Aly and Faust,

1964).   2,4-DCP is photolabile in aqueous solutions (Aly and Faust,

1964;  Crosby and Tutass, 1966) and can be degraded to succinic
                               75-4

-------
 acid  by  microorganisms  in  soil  and water  (Alexander  and Aleera,




 1961;  Ingols,  et .al., 1966; Loos, et  al.,  1967).   In lake water,




 under  laboratory conditions,  the half  life  of  2,4-DCP  is 8-9




 days  in  aerated waters  and  17 days under  anaerobic conditions




 (U.S.  EPA  1980).




 II.  EXPOSURE




     A.  Water




     Sources of 2,4-DCP in  water are  agricultural  run-off (as a




 contaminant' and metabolic breakdown product of biocides) and




 manufacturing waste discharges  (U.S. EPA,  1980).   Recent




 experiments under conditions  simulating the natural environment




 have not demonstrated that  2,4-dichlorophenol  is a significant




 product  resulting from  chlorination of phenol-containing wastes




 (Glaze,  et  al. 1978; Jolley,  et al. 1978).  The worst-case exposure




 to 2,4 DCP  from drinking water, as calculated  from 2,4-DCP level




 in water downstreams from a 2,4-DCP manufacturing,  facility,




 has been estimated as 36 ug/kg body weight/day.




     B.  Food




     Contamination of food with 2,4-DCP could  be an indirect




 result from use of the herbicide 2,4-D (U.S. EPA,  1980).  The




worst use estimate for the degree of human exposure to 2,4-DCP from




consumption of contaminated meat is about 4 ug 2,4-DCP/kg body




weight.




     The U.S.  SPA (1980) has estimated the weighted average




bioconcentration factor for 2,4-dichlorophenol to be 41 for the




edible portions of fish and shellfish consumed by Americans.




This estimate is based on the octanol/water partition coefficient.






                               75-5

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      C.   Inhalation


      Pertinent  information  regarding  direct  evidence  indicating
       *

 that  humans  are  exposed  to  significant amounts of  2,4-dichlorohenol


 through  inhalation has not  been  found in  the  available  literature.



 III.   Absorption


      Pertinent  information  regarding  the  Absorption of  2,4-


 dichlorophenol  in humans or animals was not  found  in  the available


 literature,  although data on  toxicity indicate that 2,4-


 dichlorophenol  is absorbed  after oral administration  (Deichmann, 1943;


 Kobayashi, et al. 1972).  Due, to its high lipid solubility and


 low ionization at physiological pH, 2,4-dichlorophenol  is expected


 to be  readily absorbed after oral administration (U.S.  EPA, 1979).


 B.  Distribution


     Pertinent information  dealing directly with tissue distribution


 after  2,4-dichlorophenol exposure was not found in the  available


 literature.  Feeding of 2,4-D (300 - 2000 ug/g feed)  to cattle


 and sheep (Clark, et al. 1975) and Nemacide (50 -  800 ug/g feed)


 to laying hens (Sherman, et al. 1972) did not produce detectible


 residues of 2,4-dichlorophenol in muscle or fat.    Cattle and


 sheep had high levels of 2,4-dichlorophenol in kidney and liver;


hens had detectible levels of 2,4-dichlorophenol in liver and yolk.


C.  Metabolism


     Pertinent information dealing directly with metabolism of


administered 2,4-dichlorophenol was not  found in the available


literature.  In mice, urinary metabolites of ^C-labelled gamma
                               75-6

-------
 or beta benzene  hexachloride  (hexachlorocylohexane) included 2,4-

 dichorophenol  and  its  glucuronide and  sulfate conjugates (as 4-6

 percent of  total metabolites) (Kurihara,'1975) .

 D.  Excretion

     Pertinent information dealing with  excretion of administered

 2,4-dichlorophenol was not found in the  available literature.

 After oral  administration of  1.6 mg Nemacide to rats over a. 3-day

 period, 67  percent of  that compound appeared in urine as 2,4-

 dichlorophenol within  3 days.  With a  dosage of 0.16 mg Nemacide,

 70 percent  of  the  compound appeared in urine as 2,4-dichlorophenol

 within 24 hours  (Shafik, et al. 1973).

 IV.  EFFECTS

     A.  Carcinogenicity

         Existing  data are not sufficient to indicate whether

 2,4-dichlorophenol is  a carcinogen.  The only study performed

 (Boutwell and  Bosch, 1959) indicate that 2 ,4-dichlorophenol may

 promote skin cancer in mice after initiation with dimethylbenz-

 anthracene.  An analysis of the data of Boutwell and Bosch using

 the Fisher Exact Test  indicated that the incidence of papillomas.

 in 2,4-DCP-treated groups was significantly elevated over controls,

while the incidence of carcinomas was not (U.S.  EPA, 1980) .

     B.  Mutagenicity, Teratogenicity and Other Reproductive
         Effects

         No studies addressing the mutagenicity, teratogenicity

or other reproductive effects of 2,4-DCP in mammaliam systems

were found in the available literature.  However, genotoxic

effects of 2,4-DCP have been reported in plants.  Exposure of


                               75-7

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flower buds or root cells of vetch (Vicla fabia) to solutions of




2,4-DCP, 01M and 62.5 mg/1, respectively, caused meiotic and




mitotic changes including alterations of chromosome stickiness,




lagging chromosome anaphase bridges and fragmentation (Amer and




Ali, 1968, 1969, 1974).  The relationship of such changes in




plant cells .to potential changes in mammalian cells has not been




established (U.S. EPA, 1980).




     C,  Chronic Toxicity




         One report (Bleiberg, et al. 1964) suggested that 2,4-




dichlorophenol was involved in the induction of chloracne and




porphyria cutanea tarda in workers manufacturing 2 ,4-dichlorophenol




and 2 , 4 , 5-trichlorophenol.  Since various chlorinated dioxins




(powerful chloracnegens) have been implicated as contaminants of




2,4,5-trichlorophenol, the specific role of 2,4-dichlorophenol in




causing chloracne and porphyria is not conclusive (Huff and




Wassom, 1974).




         In a study (Kobayaski, et al. 1972) in which male mice




were fed 2,4-dichlorophenol at estimated daily doses of 45, 100,




and 230 mg/kg body weight, no adverse effects were noted except




for some microscopic nonspecific- liver changes after the maximum




dose.  Parameters evaluated included body and organ weights and




food consumption,  as well as hematological and histological




changes•



     D.  Other  Relevant Information




         2,4-DCP is a weak uncoupler of oxidative phosphorylation




(Farquharson,  et al.  1958;  Mitsuda, et al.  1963).   Values on odor
                               75-8

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 threshold, for  2,4-DCP  in  water  range  from  0.65  to  6.5 ug/1,




 depending on the  temperature  of  water  (Hoak,  1957).




 V.    AQUATIC TOXICITY




      A.  Acute  Toxicity  (U.S.  EPA,  1980)




          Two 96-hour assays have been  performed  examining the




 acute  effects  of  2,4-dichlorophenol in freshwater  fish.  An LC^O




 value  of  2,020 ug/1  for the bluegill,  Lepomis macrochirus, and an




 LC5Q  value of  3,230 ug/1  for  the juvenile  fathead  minnow,




 Pimpephales promelas, have been  reported.  Two  studies on the




 freshwater cladoceran, Daphnia magna,  have produced 48-hour static




 LC5Q  values of 2,610 and  2,600 ug/1.




          Only one marine  fish or invertebrate species has been




 tested for the acute effects  of 2,4-DCP: the mountain bass, a




 species endemic to Hawaii is  poisoned  at 20 mg 2,4-DCP/l.




     B.   Chronic Toxicity




          Data for the chronic effects  of 2,4-DCP for either




 freshwater or marine organisms were not located  in the available




 literature .




     C.   Plant Effects




          Concentrations of 2,4-DCP causing a 56  percent reduction




 in photosynthetic oxygen production or a complete  destruction of




 chlorophyll were 50 or 100 mg/1, respectively, in  algal assays




with Chlorella pyrenoidosa.   An earlier study reported that 58.3




mg 2,4-D/l caused a 50 percent reduction in Chlorophyll in the




duckweed, Lemna minor.   No marine plant species have been examined.
                               75-9

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     D.  Residues




         A bioconcentration factor of 130 has been estimated from
     *



the octanol-water partition coefficient of 2,4-dichlorophenol for




aquatic organisms having a lipid content of  eight percent.  The




estimated weighted average bioconcentraion factor for the edible




portion 'of aquatic organisms is 41.




     E.  Miscellaneous




         Flavor impairment studies indicated that the highest




concentration of 2,4-DCP in the exposure water which would not




cause tainting of the edible portion of fish ranged from 0.4 ug/1




for the largemouth bass (Microbterus salmoides), to 14 ug/1 for




the bluegill (Lepomis marcrochirus).  The value for the rainbow




trout (Salmo gairdneri) was 1 ug/1.




     A.  Human



         Based upon the prevention of adverse organoleptic




effected,  the criterior for 2,4-DCP in water recommended by the




U.S. EPA (1980) is 0.3 ug/1.  This level is  far below minimal no-



effect concentrations determined in laboratory animals (U.S. EPA,




19RO).  3.09 mg/1 is the criterion based on toxicity' data (U.S.




EPA, 1980).




     B.  Aquatic



         The criterion for protecting freshwater organisms is




2020 ug/1  (acute)  and 365 ug/1 as a chronic exposure value.   No



criterion  was derived for marine organisms (U.S.  EPA,  1980).
                              75-10

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                         2,4-DICHLOROPHENOL

                             REFERENCES
 Alexander,  M.  and  M.I.H.  Aleera.   1961.   Effect  of  Chemical
 structure  on microbial  decomposition  of  aromatic herbicides.
 Jour.  Agric.' Food  Chera.  9;44.

 Aly, O.M.  and  S.D.  Faust.   1963.   Studies  on  the fate  of 2,4-D
 and ester  derivatives in  natural  surface waters.   Jour. Agric.
 Food Chem.   12:541.

 Amer,  S.M.  and E.M.  Ali.   1968.   Cytological  effects of pesticides.
 II.  Meiotic effects of  some phenols.  Cytologia 33:21.

 Amer,  S.M.  and E.M.  All.   1969.   Cytological  effects of pesticides.
 IV.  Meiotic effects of  some phenols.  Cytologia 34:533.

 Amer,  S.M.  and E.M.  Ali.   1^74.   Cytological  effects of pesticides.
 V. Effect  of some  herbicides on Specia f aba.  Cytologia 33:633.

 Bleiberg,  J.M., et  al..   1964.  Industrially acquired prophyria.
 Arch.  Dermatol.  89:793.

 Boutwell,  R.K. and  D.K. Bosch.  1959.  The  tumor-promoting action
 of phenol  and related compounds for mouse  skin.  Cancer Res.
 19:413.

 Clark, D.E., et al.  1975.   Residues of chlorophenoxy acid herbicides
 and their  phenolic  metabolites in  tissues  of  sheep and cattle.
 Jour.  Agric. Food  Chem.   23:573.

 Crosby, D.G. and H.O. Tutass.  1966.  Photodecomposition of 2,4-
 dichlorophenoxyacetic acid.  Jour. Agric.  Food  Chem.   14:596.

Deichmann, W.B. 1943.  The  toxicity of chlorophenols for rats.
Fed . Proc .2:76.

Farquharson, M.E.,  et al.   1958.  The biological action of
chlorophenols.   Br. Jour. Pharmacol.  13:2D.

Glaze, W.H., et al.  1978.  Analysis of new chlorinated organic
compounds  formed by chlorination of municipal wastewater.   Page
139 In: R.L. Jolley, (ed.)  Water chlorination  - environmental
impact and health effects.  Ann Arbor Science Publishers.

Hoak, R.D. 1957.  The causes of tastes and odors in drinking
water.  Water and Sew. Works.  104:243.
                              75-11

-------
 Huff,  J.E.  and  J.S.  Wassom.   1974.   Health  hazards  from  chemical
 impurities:   chlorinated  debenzodioxins  and  chlorinated  dibenzofurans
 Int. Jour.  Environ.  Studies  6:13.
     »
 Ingols,  R.S., et  al.   1966.   Biological  activity  of  halophenols.
 Jour.  Water  Pollut.  Control.  Fed.  38:629.

 Jolley,  R.L., et  al.   1978.   Chlorination of  organics  in cooling
 waters and  process effluents.   In  Jolley, R.L., Water  chlorination
 environmental impact  and  health  effects. 1:105.   Ann Arbor  Science
 Publishers.

 Kirk,  R.E.,  and D.F.  Othmer.   1964.  Kirk-Othmer  encyclopedia of
 chemical  technology.   2nd ed.  Interscience Publishers,  New York.

 Kobayashi,  S.,'et al.   1972.   Chronic  toxicity of 2,4-dichlorophenol
 in mice.  Jour. Med.  Soc. Toho,  Japan.   19:356.

 Kurihara, N. 1975.   Urinary metabolites  fromjfand £-BHC in the
 mouse:   chlorophenolic  conjugates.   Environ.  Qual.  Saf.  4:56.

 Loos,  M.H.,  et  al.   1967b.  Phenoxyacetate herbicide detoxication
 by bacterial enzymes.   Jour.  Agrlc.  Food Chem.  15:858.

 Mitsuda, W., et al.   1963.  Effect of  chlorophenol  analogues on
 the oxidative phosphorylation  in rat liver mitochondria.  Agric.
 Biol.  Chem.  27:366.

 Sax, N.I.   1975.  Dangerous properties of industrial materials.
 4th ed. Van  Nostrand  Rheinhold Co.,  New York.

 Shafik, T.M., et al.  1973.  Multiresidue procedure  for halo and
 nitrophenols.  Measurement of  exposure to biodegradable  pesticides
 yielding these  compounds as metabolites.  Jour. Agric. Food Chem.
 21:295.
 •
 Sherman, M., et al.   1972.  Chronic  toxicity  and residues from
 feeding Nemacide  [o-(2,4-dichlorophenol)-o,o-diethylphosphoro-
 thioate] to  laying hens.  Jour.  Agric. Food Chem.   20:617.

 U.S. EPA.  1978.  In-depth studies on health  and environmental
 impacts of selected water pollutants.  Contract No.  68-01-4646.
 U.S. Environ. Prot. Agency.

U.S. EPA.  1980.  Ambient Water Quality Criteria for 2,4-dichloro-
 pehnol:  EPA 440/5-80-042.

Weast,  R.C., ed.  1975.  Handbook of chemistry and physics.  55th
ed.  CRC Press, Cleveland, Ohio.
                              75-12

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                                             No.  76
         2,6-Dichlorophenol

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
          WASHINGTON, B.C.

          OCTOBER 30, 1980

                76-1

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                          DISCLAIMER
     This report represents a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary sources and available reference documents.
Because of the limitations of such so.urces , this short profile
may not reflect all available information including all the
adverse health and environmental impacts presented by the
subject chemical.  This document has undergone scrutiny to
ensure its technical accuracy.
                             76-2

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                      2,6-DICHLOROPHENOL




                           SUMMARY








     There is 'no available information on the possible




carcinogenic, teratogenic, or adverse reproductive effects




of 2,6-dichlorophenol.




     The compound did not show mutagenic activity in the Ames




assay.   A single report has indicated that 2,6-dichlorophenol



produced chromosome aberrations in rat bone marrow cells;




details of this study were not available for evaluation.




     Prolonged administration fo 2,6-dichlorophenol may




produce hepatoxic effects.  Pertinent data on the toxicity of




2,6-dichlorophenol to aquatic organisms were not found in the




available literature.  However, EPA/ECAO Hazard Profiles on




related compounds may be consulted, including metachlorophenol,




2,4,5-trichlorophenol,  and 2,3,4,6-tetrachlorophenol.
                             76-3

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I.     INTRODUCTION

       2, 6-Dichlorophenol  (2,6-DCP), CAS registry number 87-65-0,

exists' as white needles and has a strong penetrating odor

resembling o-chlorophenol .  It has the following physical and

chemical constants (Weast, 1972; Hawley, 1971):
         Formula:
         •Molecular Weight:             163
         Melting Point:                68AC - 69°C
         Boiling Point:                219°C - 220°C (74n torr)
         Vapor Pressure:               1 torr @ 59.5°C
         pH:             .              6.79
         Production:                   unknown
2,6-DCP is produced as a by-product from the direct chlorination

of phenol.  It is used primarily as a starting material for

the manufacture of trichlorophenols , tetrachlorophenols , and

pentachlorophenols (Doldens, 1964).

II.   EXPOSURE

      A.  Water

          Phenols occur naturally in the environment and

chlorophenols are associated with bad taste and odor in tap

water (Hoak, 1957).  2,6-DCP has a taste and odor threshold
 •
of 0.002 mg/1 and 0.003 mg/1, respectively (McKee and Wolf,

1963).  Piet and DeGrunt (1975) found unspecified dichlorophenols

in Dutch surface waters at 0.01 to 1.5 ug/1, and Burttschell,

et al. (1959) demonstrated that chlorination of phenol-

containing water produced, among other products, 2,6-DCP in a

25-percent yield after 18 hours of reaction.
                             76-4

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       B.  -Food




          Pertinent  data  could  not  be  located  in  the  available




literature.




       C.  Inhalation




          Olie, et al.  (1977) reported  finding dichlorophenols




in flue gas•condensates from municipal  incinerators.  The




levels were not quantified.




       D.  Dermal




          Pertinent  data  could  not  be  located  in  the  available




literature; however, it is known  that  dichlorophenols are




less toxic by skin contact than mono-chlorophenols and less




likely to be absorbed through the skin  (Doldens,  1964).




III.   PHARMACOKINETICS




     .  A.  Absorption




          Pertinent  data  could  not  be  located  in  the  available




literature.   By comparison with other  chlorophenols,  it is




expected that 2,6-DCP is  absorbed through the  skin and from




the gastrointestinal tract, and rapidly eliminated (U.S. EPA,




1980).




       B.  Distribution




          Pertinent  data  could not  be  located in  the  available




literature.   The high lipid solubility of the compound would




suggest that the unexcreted and unmetabolized compound distributes




to adipose tissues.




      C.  Metabolism and Excretion




          Pertinent data  could not  be located in  the  available




literature.   By comparison with other chlorophenols,  it is






                             76-5

-------
 expected  that  2,6-DCP  is  rapidly  eliminated  from  the  body,


 primarily as urirvary sulfate and  glucuronide  conjugates  (U.S.
       •

 EPA,  1980).


 IV.    EFFECTS-


       A.  Carcinogencity


          • Pertinent data  could not be  located  in  the  available


 literature.


       B.  Mutagenicity


          2,6-DCP did not show mutagenic activity in  the Ames


 assay  (Rasanen, et al. 1977).  Chromosome aberrations in rat


 bone marrow cells have been observed following compound


 administration (route and dosage  not indicated) (Chung,  1978).


       C.  Teratogenicity  and Other Reproductive Effects


          Pertinent data  could not be  located  in  the  available


 literature.


       D.  Chronic Tcxicity


          Administration  of 2,6-DCP to rats  (route and dosage


 not specified)  has been reported  to produce hepatic degeneration


 (Chung, 1978).


       E.  Other Relevant  Information


          In vitro tests have indicated that 2,6-DCP  inhibits


 liver mitochondrial respiration (level not specified) (Chung,


 1978).  At relatively high concentrations 2,6-DCP affects the


nervous system  (U.S. EPA, 1980).
                             76-6

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V.    AQUATIC TOXICITY


      A.  Acute
      *
          McLeese., et al. (1979) reported a 52-hour lethal


threshold limit of 19,100 ug/1 for marine shrimp (Crangon


septemspinosa) exposed to 2,6-DCP.

      B.  Chronic Toxicity, Plant Effects and Residues


          Pertinent data could not be located in the available


literature.

VI.   EXISTING GUIDELINES AND STANDARDS


      A.  Human

          Based on the organoleptic properties of 2,6-DCP, a

water quality criterion of 0.2 ug/1 has been recommended by


the U.S. EPA (1980).


      B.  Aquatic

          No existing criteria to protect fresh and saltwater


organisms were found in the available literature.
                             76-7

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                           REFERENCES
 Burttschell,  R.H.,  et  al.  1959.   Chlorine  derivatives  of
 phenol  causin'g  taste and  odor.   Jour.  Amer.  Water  Works Assoc.
 51:205.

 Chung,  Y.  1978.   Studies  on  cytochemical toxicities  of
 chlorophenols to  rat.  Yakhak Hoe  Chi  22:175.

 Doldens, J.D. 1964.  Chlorophenols.  In; Kirk-Othmer Encyclopedia
 of  Chemical Technology.   John Wiley and Sons,  Inc.,  New York.
 p.  325.

 Hawley, G.G.  (ed.)  1971.   The Condensed Chemical Dictionary,
 8th ed.  Van Nostrand  Reinhold Co., New York.

 Hoak, R.D. 1957.  The  causes of  tastes and odors in  drinking
 water.  Purdue  Eng. Exten. Service.  41:229.

 McKee,  J.E. and H.W. Wolf.   1963.  Water quality criteria.
 The Resources Agency of California, State Water Quality
 Control Board.

 McLeese, D.W.,  V. Zitko and M.R. Peterson.   1979.  Structure-
 lethality  relationships for phenols, anilines, and other
 aromatic compounds  in  shrimp and clams.  Chemosphere 8:53.

 Olie, K.,  et al.  1977.   Chlorodibenzo-p-dioxins and
 chlorodibenzofurans are trace components of  fly ash  and flue
 gas of  some municipal  incinerators in  the Netherlands.
 Chemosphere  8:445.

 Pfet, G.J. and  F. DeGrunt.  1975.  Organic chloro  compounds
 in surface and drinking water of the Netherlands in -problems
 raised by  the contamination of nan and his environment.
 Comm. Eur. Communities, Luxembourg, p.81.

 Rasanen, L., M.L. Hattula and A. Arstila.  1977.   The
 mutagenicity of MCPA and its soil metabolites, chlorinated
 phenols, catechols and some widely used slimicides in Finland.
 Bull.  Environ. Contam. Toxiciol.  18:565.

 U.S. EPA,  1980.  Ambient Water Quality Criteria for Chlorinated
 Phenols, EPA 440/5080-032.

Weast, R.C. 1972.  Handbook of Chemistry and Physics, 53rd
 ed.  Chemical Rubber Co., Cleveland, Ohio.
                             76-8

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                                      No. 77
2,4-Dlchlorophenoxy3cetic Acid (2,4-D}


   Health and  Environmental Streets
 U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980
             77-/

-------
                          DISCLAIMER
                     •

     This report represents a  survey  of  the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the  report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including  all the
adverse health  and   environmental  impacts  presented  by  the
subject chemical.   This  document  has  undergone scrutiny  to
ensure its technical accuracy.
                           77-2.

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                        2 , 4-OICHLOROPHENOXYACETIC ACID


                                    Summary





     Oral  administration of 2,4-Oichlorophenoxyacetic acid (2,4-0) failed to


produce  carcinogenic effects  in  mice  or dogs;  however,  feeding technical


grade 2,4-0  did produce tumors in a  study with rats.  Subcutaneous adminis-


tration of the  isooctyl ester  of  2,4-0 has been  reported to  produce reticu-


lum cell sarccmas in mice.


     A single  study has indicated  that 2,4-0 produced mutagenic  effects in


Saccharomyces.  Other  investigations  have failed  to show  mutagenic  effects


of  the  compound  Salmonella,   Drosoohila,  Saccharomyces ,  or  the  dominant


lethal assay with mice.


     2,4-0 and  several of  its  esters  failed  to shew  teratcgsnic  effects in


mice; the propylene glycol butyl ether  ester  of the  compound  produced an in-


crease in  cleft  palates  in  this st'jdy.   Studies  in  hamsters  orally adminis-


tered 2,4-0 and derivatives showed teratogenic  effects.   Oral administration


of 2,4-0 to rats failed to  indicate teratogenicity  in one study;  another in-


vestigation using oral administration of  2,4-0  to rats  found  teratogenic ef-


fects.  A  three-generation  feeding study of  2,4-0  to rats  indicated  feto-


toxic effects at a dosage  of 1,500 ppm.


     Toxicity tests on a variety  of aquatic  organisms generally  have demon-


strated that various esters  of 2,4-0 are  more toxic  than  the  2,4-0 acid,  di-


methyl amine,  or sodium salt.   Freshwater trout and bluegill sunfish  were


adversely affected  by  the  propylene glycol butylether (PGSE) ester  at  con-
                                                       s

centrations of  900  to  2,000 ug/1.   Daphnids  and  freshwater seed  shrimp  were


sensitive to the PGBE  ester at concentrations of 100 to 300 ug/1.   Chronic


exposure of several species of  fish to  concentrations  up  to 310 jug/1  has not


demonstrated any toxic effect.
                                 77-3

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                        2,4-QICHLOROPHENOXYACETIC ACID
I.   INTRODUCTION
     2,4-Oichlorophenoxyacetic acid,  CAS Registry  number 94-75-7,  commonly
known as 2,4-0, is a whits  or slightly yellow crystalline compound  which  is
odorless when  pure.   2,4-0  has the  following physical and chemical proper-
ties (Herbicide Handbook, 1979}:
                Density:
                Vsc-or Pressure:
                Solubility:
                Formula:                  C^
                Molecular Weight:         221.0
                Msltina Point:            135°C-133°C (technical);
                                          140°C-141°C (purs)
                                          160  C S 0.4 torr
                                          1.56530
                                          0.4  torr 1 lc'GcC
                                          Acetone  alcohol, oioxane ether,
                                          isopropyl alcohol;  slightly
                                          soluble in benzene,  solubility  in
                                          water 0.09c/100g, H20
                Production:               Unknown
     2,4-0 is used  as  an herbicide along with  its  various salts and esters,
which vary its  solubility  properties.   It is  used  mainly to  control  broad-
leafed  plants  in pastures,  and right-of-ways,  and, and to  keep  lakes  and
ponds free of unwanted submersed and emersed weeds.
II.  EXPOSURE
     A.    Water
          No  estimates  of  average  daily uptake  of  2,4-0   from  water   are
available; however, after  treatment  for water milfoil  in  reservoirs  in
                                    /v »  i
                                  ' J W
                                 77-

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Alabama  and Tennessee, the  Tennessee Valley Authority  found the  concentra-
tion  at  downstream monitoring stations to  be  2 ppb.  2,4-0 was not  found  in
the harvested beans  of red  Mexican bean plants after irrigation with contam-
inated water (Gangst,  1979).
     B.   Food
          The Food  and Drug Administration, in monitoring  milk  and meat for
residues  of 2,4-0.  from  1963 to  1969,  found  no  trace  of the  herbicide  in
13,000  samples  of  milk  and  12,000 samples  of  meat  (Day,  et  al.  1978).
Cattle and sheep which were  fed 2,000 ppm of 2,4-0  for 28 days had less  than
0.05  ppm 2,4-0  in  the  fat and  muscle tissue and no  detectable  amount  of
2,4-dichlorophenol.  After  seven  days withdrawal from the  2,4-0 diet,  these
tissue levels were drastically reduced  (Clark,  et al.  1975).   Six species of
fish were monitored for three weeks after the water  in a pcnd was  treated
with a  2.4_o  ester.   The highest tissue concentration  reached  was 0.24 ppm
sight days sftsr application.  Subssc'jently, ths  herbicide  cr its ^etacclite
•"as eliminated  raoidlv.  Clams  and  ovsters  acc'.jrr"jiste  more  2,4-Q  than dc
fish and crabs.  Residue peaks occur from  1 to 9 days  after  application and
then rapidly decline (Gangst, 1979).
     C.    Inhalation
          Pertinent  data  were not  found in  the  available literature;  how-
ever,  some  2,4-0  esters which are  much more volatile  than the  parent  com-
pound have  been monitored  in  air  up  to 0.13 pg/nv5 (Farwell,  et  al.  1976;
Stanley,  et  al. 1971).
     0.    Dermal
          Pertinent data were not found in the available literature.
                                  77-J-

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III. PHARMACQKINETICS
     A.   Absorption
          Human  absorption  of  2,4-0   following  oral  intake  is  extensive;
Kohli  et  al.  (1974) have  determined absorption of  75  to 90 percent  of the
total  dietary  intake of  the compound.   Animal  studies have  indicated that
the gastrointestinal  absorption of  2,4-0  esters  may be  less  efficient than
that of the free acid or salt form of the compound (NRCC, 1978).
     3.   Distribution
          The phenoxy herbicides  are readily distributed throughout the body
tissues of mammals.  Tissue  levels  of  herbicide may be  higher  in  the  kidney
than in the blood;  liver  and muscle show  levels  lower  than those determined
in  the  bleed  (NRCC, 1973).  withdrawal of dietary  compound  produced  almost
complete tissue loss of residues in seven cays (Clark,  at al.  1975).
          Small  amounts  of phenoxy  herbicides  are  passed   to  the  young
through the mother's milk  (Blerke.  e^  al.  ^972).   Transolacsntal transfer  of
2.4-0 has been renorted in mice (Lindquist and Ullberg,  1971).
     C.   Metabolism
          Sauerhoff, et al.  (1976)  determined  that following  oral adminis-
tration of 2,4-0 to human volunteers,  the  major amount  excreted in the urine
was free  compound: a  smaller  amount  was excreted  as  a conjugate.   Tissue
analysis of  sheep and  cattle  fed  2 4-0  have shown unchanged  compound  and
2,4-dichlorophenol to be present (Clark, et al.  1975).
     D.   Excretion
          Elimination of  orally  administered  2,4-0 by  humans  is  primarily
through the urine  (95.1 percent of  the initial dose);  the half-life  of  the
compound in  the body  has been  estimated  as 17.7  hours (Sauerhoff,  et.al.
1976).   Clark,  et al. (1964) have reported urinary  elimination  of  96 percent

-------
of  an  oral dose of  labelled 2,4-0 within  72  hours by  sheep;  approximately
1.4 percent of the administered dose was eliminated in the feces.
          The plasma  half-life  of 2,4-0 has been  estimated  to be- from  11.7
to 33 hours in' humans (NRCC, 1978).
IV.  EFFECTS
     A .   Carcinogenicity
          Innes, et  al.  (1969)  reported no significant increase  in  tumors
following  feeding  of mice  with  2,4-0  for 18  months.   A  two-year  feeding
study in rats did indicate  an increase  in  total  tumors in females and malig-
nant tumors in males  following  feeding  of -technical 2,4-0; a  parallel study
with dogs  fed technical compound  did not show carcinogenic effects (Hansen,
et al.  1971).
          Mice were  administered  maximum  tolerated doses of  2,4-0 and  its
butyl,  isoprcoyl, and isooctyl  esters  in a long-terrr,  carcir.ogenicity  study.
Carcinogenic effects were seen  after subcutaneous acministraticr,  of the  iso-
octyl ester (reticulum cell  sarcomas)  (NiCI,  1963).
     3.    Mutansnicity
          No  mutagenic   effects   of   2,4-0   in   tests  with   Salmonella,
Saccharomvces,  or  Orosophila  were observed  (Fahrig,  1974).    Siebert   and
Lemperle  (1974)  have  reported  m.'jtaaenic  effects  following   tr3=t~ert  of
Saccharomvces  csrevisiae strain  04 with  aqueous 2,4-0  solution  (1,000 mg/1).
          Gavage or  intraperitoneal  administration of  2,4-0  to mice  failed
to show  mutagenic  effects  in   the dominant lethal  assay (Epstein,  et   al.
1972) .
                                                       t
     C.    Teratogenicity
          Testing of  2,4-0  and  its n-butyl, isopropyl,  and  isooctyl  esters
in pregnant mice produced no significant teratogenic  effects.   There was a
                                77-7

-------
significant  increase  in  cleft palate deformities after administration of  the
propylene glycol butyl ether  ester of 2,4-0  (Courtney,  1974).
          Subcutaneous  injection of the  two  isopropyl  esters and  the iso-
octyl  ester  of 2,4-0 in  pregnant mice has  been  reported to produce terato-
genic  effects  (Caujells,  et  al.  1967),  although  the DMSC  vehicle  used  is,
itself, a teratogen.  Sage,  et al.  (1973)  have also reported teratogenic  ef-
fects  in mice following injection of 2,4-0.
          Oral  administration of 2,4-D  to hamsters resulted  in  the produc-
tion of some terata (Collins  and Williams,  1971).   Studies with rats report-
ed that oral administration  of the  parent  compound or its isooctyl and butyl
esters, and butoxy ethanol and dimethylamine salts,  produced teratogenic  ef-
fects  (Khera and  McKinlay,  1972).  However, Sclr.vetz,  et  al. (1971)  were  un-
able to  show teratogenic effects in rats following  the  oral administration
of 2,4_o or its isoocytoi or  propylene giycol butyl ether esters.
     0.   Other Ssorccuctivs  Effects
          Embryctoxic effects following  subcutaneous administration of 2,4-0
to pregnant  mice 'nave  been   reported  (Caujolle,  et al.  1967;  Sage, et  al.
1973).
          Fetotoxic effects of the compound  and its  esters have been report-
ed after  oral  aciTiiniSursLiGn  or  rnaxii7iaij.y  L0i=rauec doses  vjcruvcuz,  ec  5^.
1971; Khera and McKinely, 1972).
          Results  of  a  three-generation  study of  rats  fed  2,4-0  indicate
that at dietary levels up  to 500 ppm,  no  reproductive  effects  are produced;
at levels of 1,500 ppm, a  decrease  in  survival and body weights of weanlings
was observed (Hansen, et  al.  1971)-.  Bjorklund and  Erne  (1965)  reported  no
adverse reproductive effects  in rats fed 1,000 mg/1 2,4-0 in drinking water.

-------
     E.   Chronic Toxicity
          Animal  studies with prolonged oral  administration of 2,4-0 or  its
amine  salt  have  indicated renal  and hepatic  effects  (Bjorklund  ana  Erne,
1971; Sjorn and Northen,  1948);  the chemical purity of  the  material  adminis-
tered is not  known.   A feeding study  in  rats  has reported  histcpatholcgical
liver changes at  dietary  levels  of 2,4-0 equivalent  to 50 mg/kg (Oow  Chem-
ical, 1962).
V.   AQUATIC TOXICITY
     A.   Acute Toxicity
                                                    \
          The  National Research  Council  of Canada  (1978)  has  reviewed  the
toxic  effects  of   2,4-0   to  fish.    For  the  blusgill   sunfish   (Lescmis
macrcchirus),  2,4-0  acid and  2,4-0 dimethyl amine produced toxic affects at
concentrations greater than 100,000 ug/1.  At  2,4-0  concentrations of 50,000
ug/1 or  less,  no  increased mortalities were reported  except in pink salmon.
The  isoprcpyi,  butyl,  ethyl,  butoxy etnanoi,  and  FG5E  esters  produced
48-hour  LC5Q  values of  900,  1,300,  1,400.  2,100,  and  from 1,000  to  2,100
ug/1, respectively.
          For other  fish species,  the  results  follow  a similar trend in that
the esters  tend  to  be more  toxic than other  formulations.  Meehan,  st  al.
(1974)  ccnductac  tssts of various  formulations  of  2,4-0  en echo salmon  fry
and fingerlings (Oncorhycus Kitutch),  chum salmon fry  (0_.  keta),  pink salmon
fry  (0_.  gorbuscha),   sockeye  salmon  smolts  (0_.   nerka),   Dolly  Varden
(Salvelinus malma),  and rainbow  trout (Salmo gairdneri).   The  butyl ester
was  the most toxic  ester  tested,  with  concentrations  of  1,000  ug/1  or
greater producing nearly 100  percent mortalities  in  all  species  tested.   The
PGBE ester  was  similar in  toxicity  to the butyl ester.   Rainbow trout  were
reported to have  shown  a  48-hour LC5Q vaiue  Of 1,100  /ug/i on  exposure  to

-------
the  PG8E ester  of 2,4-0.   Harlequin  fish  (Rasbora  heteromorpha)  showed  a
48-hour  LC5Q value  of 1,000  ug/1  on  exposure  to the  butoxyethyl  ester of
2,4-0  (National  Research Council of Canada 1978).   Rehwoldt,  et al.  (1977)
have  observed  96-hour  LC50  values   of  26,700;  40,000;  70,100;  70,700;
94,600;  96,500;  and  300,600 ug/1 for  banded  killifish (Fundulus diaphanus),
white  perch (Roccus  americanus),  stripped bass  (Morone sazatilis), guppies
(Libistes   reticulatus),   bumpkinseed   sunfish   (Lepomis   gibbosus),   carp
(Cyprinus  carpio),  and  American  eel  (Anguilla  rostrata),  respectively,
exposed to commercial technical grade 2,4-0.
          Sanders  (1970)  conducted  a comparative  study  on  the toxicities of
various  formulations  of  2,4-0 for  six species  of  freshwater  crustaceans.
The PG8E ester was  generally  most  toxic,  while  the  dimethylamine  salt was
least toxic.  The  crayfish  (Orconectes  nails)  was the most resistant species
tested,  with  46-hour static 1C    values  areater than 100.COO uo/1  for all
                                ^u          -                     ••  -
formulations  tested.   The  water flea  (Dap'nnia  ~.acna)   and   seed  shrimp
(Cypridopsis  vidua)  were  most sensitive  to  the FGBE  ester,  with  43-hour
LC=g   values   of   100  and   320  jug/1,   respectively.    Scuds   (Gammarus
fasciatus),   scwbugs   (Ascellus brevicaucus),  and  freshwater  grass  shrimp
(Palaemonetes  kadiakensis)  were also  moderately sensitive,  with  48-hour
LC5Q  values  ranging  from  2,200 to  2,700 ,ug/l.   Sanders  and Cope (1968)
reported  a   96-hour  LC5Q  value   of  1,600   ug/1   for   stonefly   naiads
(Pteronarcv californica) exposed to  the butoxyethanol ester of 2,4-0.  Tech-
nical grade 2,4-0  produced a  96-hour  LC5Q vaiue  of  14,000 ug/1.   Robertson
                                                       .•
and  Bunting  (1976)   reported  96-hcur  LC5Q  values  ranging  from  5,320  to
11,570 /jg/1 for copepods (Cyclops vernalis) nauplli exposed to  2,4-0 as free
acid.  The  range  of  96-hour LC5Q values for  nauplli  exposed to  2,4-0  alko-
nolamine salt was 120,000 to 167,000 Aig/1.

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          Among  marine invertebrates, those  of commercial significance have
been  examined  for toxic effects  on exposure to  2,4-0 formulations.  Butler
(1965)  determined the 96-hour median effective concentration based on shell
growth  for oysters as  140  ug/1 for the PGEE ester of  2,4-0.   The 2,4-0 acia
had  no  detectable effect  at exposures  of 2,000 jug/1  for 96-hours.  Butler
(1963)  observed  paralysis  of brown shrimp (Penaeus aztecus) exposed to 2,4-0
acid  at a  concentration of 2,000 (ug/l for 48-hours.   Sudak and Claff (1960)
found   a  96-hour  LC5Q  vaiue  of  5,000,000  jug/1  for  fiddler  crabs  (yea
pugmax) exposed  to 2,4-0.
          McKee  and  Wolf (1963)  have reviewed  the toxic effects  of 2,4-0 to
aquatic organisms.   Toxic  concentrations as lev; as 1,000  ug/I  produced  -3. 40
percent mortality for  fingerling  bluegills exposed to  2,4-0 butyl ester.   In
general, esters of 2,4-0 were reported to  be  more  toxic than sodium salts of
2,4-0.
     E.   Chronic Tcxicity
          Rehwoldt!,  et al.  (1970)  exposed several  species  of fish  to  100
jug/1  2,4-0  for  ten  months  and  observed  no   overt  effects  to  any  tasted
species.  The  percent  reduction of brain  acetylcolinestarase  ranged from 16
percent in white  perch  to  35 oercent in  American eels.  In breeding exoeri-
ments with guppies,  a  100  ug/1  concentration  of 2,4-0  had  no significant ef-
fect  on the  reproductive  process of the  species  under experimental  ccnc-i-
tions.   Cope,  et al.  (1970)  examined the  chronic effects of  FGEE  ester  of
2,4-0 to bluegill sunfish.  Fish  were exposed to  the herbicide in one-eighth
                                                       .-
acre pcr.ds containing  initial concentrations  of up to  10,CCO  pg/1.   Altera-
tions in  spawning activity,  and  the occurrence  of  pathological  lesions  of
the liver, brain, and  vascular  system were reported  for a  period  of up to 84
                                  -*»•

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days.   Mount and Stsphan  (1967)  exposed  1-inch  fathead minnows  (Pimeohales
promelas)  to  a  continuous series  of  concentrations  of  the  butoxyethanol
ester of 2,4-0  ranging  from 10 to 310 ug/1 for a 10-month  period.   No deaths
of  deleterious  effects,   including  abnormal  spawning  activity and  reduced
survival of  eggs  from exposed  fish, were  observed.
          In static-renewal tests, Sigmon (1979)  reported that the  percent
pupation and the percent  emergence  of Chironomus  larvae were  significantly
reduced by exposure  to  1,000 or 3,000 ug/1 1,4-0 (acid  equivalent  in  Wesdone
LV-4 formulation).
                                                     ».
     C.   Plant Effects
          The  genera  Microcystis,  Scanedesmus,   Chlorella,   and   Nitzschia
shewed  no tcxic response  when  exposed to 2,OCO jug/1  2,4-0 Lawrence  (1952).
Poorman  (1973)  treated cultures  of Euglena  gracilis  with concentrations  of
50,OCO  .ug/1  2,4-0  for  24   hours  and   observed   cecressed   growth   rates.
Valentine and Einc°?.m  (1974)  cemonstrateo that  at 100,000  ug/1,  2,4-D  re-
duced  the  ceil  nurnoers of  Scenecesmus  to  one  percent of  control  levels,
Chlai7iydo~.or.3s  to  43 percent  of  control  levels,  Chlorella  to 66 percent  of
control levels,  and  Euglena to 90 percent of control levels  within 4  to  12
days.   The  bluegreen algae (Nostoc  muscor.ji~1  displayed s.  68-percent  reduc-
•^-.cpi — n  grew ^i .  n)icr< cxposec ^c  -LWW  ug/j. *_,-T— ^  ^w^.-w-^  o.iv- wav'ci«.j — •>— ,  — ^«^/.
Singh (1974)  exposed Cylindrosoermum to  2,4-0  sodium salt at  concentrations
ranging  from  100,000  to  1,200,000  jjg/1 and  reported  that  concentrations
above 800,000 ug/1  caused  growth to  cease completely.  McKee and Wolf (1963)
reviewed  the  effectiveness of  2,4-0 in  control of  emergent aquatic  plants
and  reported  that  concentrations ranging  from 6,000  to  100,000 jug/1  have
been effective in controlling a number of species.
                                   • ? 7 ?~
                                  ^^^^^^^^7

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     0.   Residue
          Cope,  et al.  (1970)  examined residues of  the PG3E ester of  2,4-0
 in  the freshwater vascular plant, Potamogeten  nodosus,  in a one-eighth  acre
 pond treated  with  single 100 to 10,000 ug/1 applications  of  the  chemical.   A
 gradual depeletion of the  herbicide  to insignificant levels was•demonstrated
 within three months.
          Schultz  and Gangstad  (1976)  reported that  the flesh  of fish ex-
 posed to 2,4-0 dimethyl  sodium salt  in ponds treated  with from 2.24 to  8.96
 kg  (as an acid equivalent)  of the chemical did not attain the 100  ug/1  level
 realized in the water two weeks after application.
          The National Research  Council of Canada (NRCC)  (1978) has reviewed
 the bioconcsntraticn  data  and associated  residues  of 2,4-0  in  a  number of
 studies.   NRCC indicated that a relatively short half-life  of less than two
 days is  found for fish  and oyster.   At water  concentrations of  100  to 2CQ
 ug/1, the bicconcentration  of  2,4-0  various  aquatic invertebrates v/as one to
 two orders  of magnitude greater  than  lin  the  water.   Oysters   (Crassostica
 viroinica)  were  reported to have  a  bioconcentration  factor  of  150 when ex-
oosed to  the butoxyethanol  ester of  2,4-0.   The freshwater oluegill and mos-
quito fish (Garnbusia  affinls)  had  bioccncsntration  factors ranging from 7 to
55,  respective  to water concentrations.'  fish  fee;  a diet  containing  2,4-G
bicconcentrated the 2,4-0 acid by less than 0.2.
VI.   EXISTING GUIDELINES
     A.   Human
          The acceptable daily intake  of  2,4-0  for  humans has  been  estab-
 lished at 0.3 mg/kg (FAO, 1969).
     B.   Aquatic
          Pertinent data were not found in the available literature.

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                         2,4-OICHLOROPHENOXYACETIC ACID

                                   References
 Bage,  at al.   1973.  Teratogenic and embryotoxic effects of herbicides  diand
 trichlorophenoxyacetic  acids (2,4-0 and  2,4,5-T).   Acta Pharmacol.  Toxicol.
 32:  408.

 Sjerke,  E., et  al.   1972.   Residue studies of phenoxy herbicides  in  milk and
 cream.   Jour. Agric. Food Chem.  20: 963.

 Bjorklund,  N.  and K. Erne.   1966.  Toxicological  studies  of  phenoxy acetic
 herbicides in animals.  Acta Vet.  Scand.  7: 364.

 Bjorklund, M. and K. Erne.   1971.   Phenoxy-acid-induced renal changes in the
 chicken.  I. Ultra structure.  Acta Vet. Scand.  12: 243.

 Bjorn,  M.  and H. Northen.   1948.   Effects  of 2,4-dichlorophenoxyacetic acid
 on chicks.  Science  108: 479.

 Butler,  P.A.  1963   Commercial  Flshary  Investigations.   U.S.   Oept.  Interior
U.S. Fish anc Wildlife Service Circ.  167: 11.

Butler,  P.A.    1965.   Effects   of herbicides   on  estusrir.e  fauna.   Prcc.
Southern Weed Conference  18: 576.

Caujoile, F.,  et al.   1967.   Limits cf  toxic and tsratcgsnic  tolerance  of
ci~3thyl sulfc:
-------
 Dow  Chemical Company.  1962.  Results  of  90-day dietary feeding of  the  pro-
 pylene  glycol isobutyl ether  ester of  silvex  (Dowco 171)  to rats.   Unpub-
 lished  Report.  Oow Chemical Co., Midland, MI.

 Epstein,  S.,  et al.  1972.  Detection  of chemical  mutagens  by the  dominant
 lethal  assay  in the mouse.  Toxicol. Appl. Fharmacol.  23: 283.

 Food  and  Agriculture  Organization  of the United Nations (FAO).  1969.  Work-
 ing  party of experts on  pesticide residues.   Evaluations  at some pesticide
 residues  in  food, the monographs.   FAQ/WHO PL 1968/m/9/l.

 Fahrig, R.   1974.  Comparative mutagenicity  studies with pesticides.  Chem-
 ical  Carcinogenesis Assays.  IARC Scientific Publication  10:  161.  Lyoh.

 Farwell,  S.O.  et al.   1976.   Survey   of  airborne  2,4-0  in  south  central
 Washington.   Jour. Air Pollut.  Control Assoc.   26: 224.

 Gangst, E.O.   1979.   Herbicide Residue  of 2,4-0,  Office of Chief of Engine-
 ers,  Washington, O.C.  NTIS AD-67160.

 Hansen, w. et al.   1971.   Chronic  toxicity of 2,4-dichlorophenoxyacetic acid
 in rats and dogs.  Toxicoi. Api. Fharmacoi.  20: 122.

 Herbicide  Handbook.   1979.   4th   ed.   Weed   Science   Society  of  America,
 Champaign, IL.  p. 129.

 Innes,  J. ,  et ai.   1969.   Bicassay of pesticides  and  industrial  chemicals
 for tumorigencity in  mice:   A  preliminary note.   Jour.  Nstl.  Cancer Instit.
 42: 1101.

 p'nera,  K.  ana w.  McKiniey.  1572.   Pre and postnatal  srjci^s en 2,4,5-tri-
 chlorophenoxyacetic acid,  2,4-dichiorophenoxyacetic  acid  and  their  deriva-
 tives in rats.  lexical.  Appl. Fharmacol.   22:  14.

 Konli,  J., et al.  1974.   Absorption  and  excretion  of  2,4-dichlorophenoxy-
 acetic acid in man.   Xenobiotica,  4: 97.

 Lawrence,  J.M.   1962.   Aquatic Herbicide  Data.   U.S. Dept.  of Agriculture,
                                  p.
Lindquist, N. and S. Ullberg.   1971.   Distribution  of the herbicides 2,4,5-T
and  2,4-0  in  pregnant  mice.   Accumulation  in the yolk  sac  epithelium.
Experientia, 27:  1439.

McKee,  J.E.  and H.w.  Wolf.   1963.   Water Quality  Criteria.   Calif.  State
water Quality Board Publication 3-A.

Meehan, W.R.,  et al.   1974.   Toxicity of  various  formulations of  2,4-0  to
saimcnids in southeast Alaska.  Jour. Fish Red. 3d.  Canada  31: 480.

Mount,  O.I.  and  C.E. Stephen.  1967.   A  method  for  establishing  acceptable
toxicant limits for  fish -  malathion and  the butoxy  ethanol ester of 2,4-0.
Trans. Am. Fish Soc.  96:  185.

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National  Research Council Canada  (NRCC).   1978.   Phenoxy Herbicides -  Their
Effects   on   Environmental  Quality.    Associate   Committee  on  Scientific
Criteria  for Environmental  Quality  NRCC No.  16075,  ISSN 0316-0114.  Avail-
able:  Publications NRCC/CNRC Ottawa K1A OR6.

National  Cancer  Institute.  1968.  Evaluation of carcinogenic, teratogenic,
and  mutaaenic. activities  of selected  pesticides and  industrial chemicals.
National Cancer Institute, PB-223 159.

Poorman,  A.E.   1973.   Effects  of  pesticides on  Euglena gracilis  I growth
studies.  Bull. Environ. Contain. Toxicol.  10: 25.

Rehwoldt, R.E.,  et al.   1977.   Investigations  into  the acute  toxicity and
some chronic  effedts  of selected  herbicides and pesticides  on  several fish
species.  Bull. Environ. Contam. Toxicol.  18: 361.

Robertson, E.B. and D.L.  Bunting.   1976.  The acute,  toxicity of four herbi-
cides  to 0-4 hour  Nauplli  of  Cyclops  vernalis fishes.   Bull.  Environ.
Contam. Toxicol.   16: 682.

Sanders, H.O.  1970.  Toxicities of  some herbicides  to six species of fresh-
v/ater crustaceans.  Int. Jcur. Water Pcllut. Ccntrcl Fed.  42: 1544.

Sanders, H.O. and O.B.  Cope.  1963.   The relative toxicities of several pes-
ticides  to  naiads  of  three  species  of stoneflies.   Limnol and  Oceanogr.
13: 112.

Sauerhcff, M..,  et  ai.    1976.   The  fate  of  2,i-dichicrcpr,er,oxyacetic  acia
(2,4-C)  fallawina oral  administration  to  rr,an.   Toxicoi.  Apcl.  Fhsr~iacci.
37: 136.

Scnwetz,  3.,  et  al.    1571.   The effect  of  2,4-tiichlorcphenoxyacetic  acid
(2,4-0) and  esters  of 2,4-0  on rat srcbrycnal,  foetal,  sr.d  necnatal growth
and development.   Food Cosmet. Toxicol.  9: 301.

Schultz, D.P. and E.O. Gangstad.   1976.  Dissipation  of  residues of  2,4-0 in
water,  hydrosoil,  and fish.  Jour.  Aquat. Plant Managa.  14: 43

Sisoer^". 0.   and P.  Lsn^c^rle,  1974.   Genetic  effects  of herbicides:  inc'jc—
ticn of  mitotic  gene  conversion  in  Saccharcrnyres  cerevisiae.  Mut.  Res.
22: 111.

Sigmon, C.F.   1979.   Influence  of  2,4-0 and  2,4,5-T  on  life history charac-
teristics  of  Chironomus  (Diptera  Chironomidae).   Bull  Environ.   Contam.
Toxicol.  21: 596.

Singh,   P.K.   1974.   Algicidal  effect of  2,4-dichlorophenoxyacetic   acid  on
blue-green algae.   Cylindrosperum sp. Arch. Microbiol. ' 97:  69.

Stanley,  C.W., et al.   1971.   Measurement  of atmospheric  levels of pesti-
cides.   Environ.  Sci.  Technol.  5:  430.

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Sudak, F.N.  and C.L. Claff.   1960.   Survival of Uca  pugnax  in sand, watsr,
and  vegetation  contaminated  with  2,4-dichlorophenoxyacetic  acid.   Proc.
Northeast Weed Cont. Conf.  14: 5Q8.

Valentine,  J.P.  and  S.'.v. Bingham.   1974.   Influence of  several  algae  on
2,4-0 residues in water.  Weed Sci.  22: 358.

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                                      No.  78
        l,2-Dichloropco,pane


  Health and Environmental Effects
U.S. ENVKOT.ENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such  sources, this short profile
may not  reflect  all available  information  including all the
adverse health  and   environmental  impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical acc-uracy.
                             - ?~J cr~

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                              1,2-DICHLOROPROPANE
                                    Summary

     The major  environmental  source of dichloropropane  is  from the use of a
mixture  of dichlorcpropanes  and dichloropropenes  as a soil  fumigant.   On
chronic exposure  of  rats to dichloropropanes  the  only observed effect was a
lack of normal  weight  gain.   There is no  evidence that dichloropropanes are
carcinogens or  teratogens.  Oichloropropanes have produced mutations in bac-
teria and caused chromosomal aberrations in  rats. ••
     Aquatic toxicity  tests of 1,2-dichloropropane are limited to  four acute
investigations.   Two  observed  96-hour  LC--,  values  for  the  bluegill  are
230,000  and  320,000 jjg/1  and  the  43-hour  LC50  value  for Osphnia magna is
52,500  jug/1.   A  saltwater  fish  has  a   reported  96-hour   LC5Q   value  of
240,000 Jjg/i.

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                               1.2-OICHLOROPRQPANE
' I.   INTRODUCTION
      This profile  is based  on  the Ambient  Water Quality  Criteria Document
 for Dichloropropanes/Dichloropropenes (U.S. EPA, 1979).
      1,2-Oichloropropane  (1,2-rDC,  molecular weight  112.99)  is  a  liquid at
 environmental  temperatures.   This  isomer  of dichloropropane  has  a boiling
 point of  96.4°C,  a density  of  1.156 g/ml,  a vapor pressure of  40 mm Hg at
 19.4°C  and   a  water  solubility  of 270  mg/100  at 20°C  (U.S.  EPA,  1979).
 Mixtures  of  1,2-dichloropropane  and  cis-trans-L, 3-dichloropropene  are  used
 as  soil  fumigants.  For  the purposes of  discussion in this  hazard profile
 document, dichloropropane refers  to  the  1,2-dichlorcpropane  iscmer.   When
 heated to decomposition  temperatures,  l,2-dichioropropane  emits highly toxic
 fumes of phosgene (Sax,  1975).
 II.  EXPOSURE
      A.   ',','3 tar
          Oichlorooropane  can enter the  aquatic  environment   as  discharges
 from  industrial  and manufacturing processes,   as  run-off  frcm agricultural
 land, and  from municipal  effluents.   This  compound  was identified  but  not
 quantified in New Orleans drinking water (Oowty, et al. 1975).
      8.   Food
          Information was  not found, concerning the concentration of dichloro-
 propane  in commerical foodstuffs;  therefore, the  amount of  this compound in-
 gested by humans  through  food  is not  known.  The U.S. EPA  (1979)  has esti-
 mated the bioconcentration factor  (BCF) of  dichloropropane to be  20.   This
 estimate  is  based on the octanol/water partition coefficients of  dichloro-
                                                                       9
 propane.  The weighted average  BCF  for edible portions  of  all aquatic organ-
 isms consumed by Americans is calculated to be 5.3.

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     C.  Inhalation
         Atmospheric  levels  of  aichloropropane  have  not  been  positively
determined.  However,  it  is known  that  5-10 percent of  the  dichloropropane
which  is  applied  to the  soil  as a  fumigant  is released to  the  air (Thc~as
and McXeury, 1973).
III. PHARMACOKINETICS
     A.  Absorption, Distribution and Metabolism
         Pertinent  data  could  not  be   located   in   available   literature
searches regarding the absorption of dichloropropane.
     B.  Excretion
         Pertinent  data  cculd   not be   located   in   available   literature
searches regarding  excretion of  dichloropropane.  in the  rat,  approximately
50 percent  of  an  orally administered dose of  dichloropropane was eliminated
in the urina in 24 hours (Hutson, et ai.  1971).
     A.  Carcinogenicity
         Only  one  study  is  reported  on  the  carcinogenicicy of  dichloro-
propane.   Heppel,  et  al.   (1948)  repeatedly  exposed  mice  (37  exposure
periods) to  1.76  mg dichloropropane per liter  of air.   Of  the 80 mice,  only
three survived  the  exposure  and subsequent observation  period; however,  the
three survivors had  multiple hepatomas at the  termination  of the experiment
(13 months of  age).  Due  to  the high mortality, an evaluation  based  on' this
study cannot be made.
     B.  Mutagenicity
         DeLorenzo,  et   al.   (1977)   and  Bignami,   et  al.   (1977)   showed
                                                                      »
dichloropropane to  be mutagenic  in S.  typhimurium  strains  TA  1535 and  TA
100.  Dichloropropane has also been shown to cause mutations in  A.  nidulans

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(Bignami, et  al.  (1977),  and to cause  chromosomal aberrations  in  rat bone
marrow (Dragusanu and Goldstein, 1975).
     C.  Teratogenicity
         Pertinent information  ccuid not be  located  in available literature
searches regarding teratogenicity.
     0.  Other Reproductive Effects
         Pertinent information  could not be located  regarding  other repro-
ductive effects.
     E.  Chronic Toxicity
         Pertinent information  could not be  located  in available literature
searches regarding chronic tcxicity  studies  of dichloroprooane  exposure in
humans.  In  one study by  Heppel,  et al. (1948) rats,  guinea  pigs,  and dogs
were exposed  to 400  ppm of dichloropropane  for 128 to 140  daily seven hour
•period (given  five aays  per week).   The  only effect observed was a decreased
weight in rats.
V.   AQUATIC TOXICITY
     A.  Acute Toxicity
         Two   observed   96-hour  LC5Q   values   for  the   bluegill,   Lepcmis
macrochirus,  upon  exposure to  1,2-dichloropropane  were 280,000-and 320,000
ug/1 (Oawscn,  et al.  1977; U.S.  EPA, 1978).   In the only freshwater inverte-
brate  study  reported,  the 48-hour  LCjQ for Daphnia  macna is  52,500 ug/1
(U.S.  EPA,   1979).   Tidewater  silverside,   (Menidia bewllina),   has  an
observed 96- hour LC^ of 240,000/jg/l (Oawson, et al.  1977).
     8.  Chronic Toxicity
         Chronic  data  are not  available  for any  saltwater or freshwater
species.

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     C.  Plant Effects
         The phytotoxicity  of  1,2-dichloropropane has not been investigated.
     0.  Residues
         No information available.
VI.  EXISTING GUIDELINES AND STANDARDS
     Neither the human  health nor  the aquatic criteria derived by  the  U.S.
EPA  (1979),  which  are  summarized below,  have gone  through  the process  of
public review; therefore, there is  a possibility  that these criteria will be
changed.
     A.  Human
         The TLV for  dichloropropane  is  75 pom (350 mg/m  )  (Am. Conf.  Gcv.
Ind. Hyg.,  1977).  The draft  water  criteria  for  dichloropropane is  203  ug/i
(U.S. EPA,  1979).
         ;-cr  i.2-dichlorcprcpane,  the  proposed  draft  criteria  to  protect
                                                                        i
frsshv/ater aquatic life are 92C ug/1 a 24-hour average-and the concentration
should not  exceed 2,ICO ug/1  at  any  time.   Criteria are not  available  for
saltwater species (U.S.  EPA,  1979).
                               ni'i
                               7 ' i *S4f
                              71-7

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                     1,2-DICHLOROPROPANE

                         REFERENCES

American Conference of Governmental Industrial Hygienists.
1977.  Documentation of the threshold limit values.  3rd.
ed.

Bignami, M. , et al.  1977.  Relationship between chemical
structure and mutagenic activity in some pesticides:  The use
of Salmonella typhimurium and Aspergillus nidulans.  Mutag.
Res. 46: 3.   ~*"~~~~~~~~~

Dawson, G.W., et al.  1977.  The acute toxicity of 47 indus-
trial chemicals to fresh and saltwater fishes.  Jour. Hazard.
Mater. 1: 303.

DeLorenzo, F., et al.  1977.  Mutagenicity of pesticides
containing 1,3-dichloropropene.  Cancer Res. 37: 6.

Dowty, B., et al.  1975.  Halogenated hydrocarbons in New
Orleans drinking water and blood plasma.  Science 87: 75.

Dragusanu, S., and I. Goldstein.  1975.  Structural and nu-
merical changes of chromosomes in experimental intoxication
with dichloropropane.  Rev. Ig. Bacteriol. virusol.  Parazi-
tol. Epidemiol. Pr.euir.cf itziol. Ig 24: 37.

Heppel, L.A., =t al.  1943-.  Toxicology of 1,2-d ichloroprc-
pane (propylene dichloride) IV. Effect of repeated exposures
to a lev; concentration of the vapor.  Jcur. Ind. Hyg.  Tex i-
col. 30: 139

Hutson, D.H., et al.  1971.  Excretion and retention of com-
ponents of the soil fumigant D-D^R^ and their metabolites
in the rat.  Food Cosmet. Toxicol. 9: 677.

Leistra, M.  1970.  Distribution of 1,3-Dichloropropene over
the phase in soil.  Jour. Agric. Food Chem.  18: 1124.

Roberts, R.T., and G. Staydin.  1976.  The degradation of (2)-
and (E)-l,3-dichloropropenes and 1,2-dichloropropanes in
soil.  Pestic. Sci.  7: 325.

Sax, N.I.  1975.  Dangerous properties of industrial mate-
rials.  Reinhold Book Corp., New York.

Thomason, I.J., and M.V. McKenry.  1973.  Movement and fate
as affected by various conditions in  several soils.  Part I.
Hallgardia 42: 393.
                                                          »
U.S. EPA.  1978.  In-depth studies on health and environmen-
tal impacts of selected water pollutants.  Contract No.  68-
01-4646.
                              7F-?

-------
U.S. EPA.   1979a.   Dichloropropenes/Dichloropropanes:  Ambient
Water Quality Criteria.  (Draft).

U.S. EPA.   1979b.   Dichloropropenes/Dichloropropanes:  Hazard
Profile.
                          7*-?
                          '

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                                    No. 79
                «•
  Dichloropropane/Dichloropropenes



  Health and Rnvlrotnaental Effects
u.s.  ENvmoNMEmL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980
            77-t

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not  reflect  all available  information  including all the
adverse health  and   environmental  impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                DICELGROPROPANES/DICHLOROPROPENES
                             SUMMARY
      The major environmental source of dichloropropanes  and
dichloropropenes is from the use of these compounds  as  soil fumi-
gants.  Some mild kidney damage has been observed  in rats chroni-
cally exposed to 1,3-dichlorpropene.  Both dichloropropane and
dichloropropene have been shown to be mutgenic  in  the Ames assay
test.  Data are not available to prove conclusively  that  these
compounds are chemical carcinogens.
      Aquatic toxicity studies suggest that  the acute toxicity
of the dichloropropanes decreases as the distance  between the
chlorine atoms increases.  As an example, the reported  96-hour
LCen values for the bluegill, Lepomis macrochirus, for  1,1-,
1,2-, and 1,3-dichloropropane are 97,900, 280,000, and  greater
than 520,000 ug/1, respectively.  For Daphnia itiagna,  the  corres-
ponding reported 48-i-iour LC-fl values are 23,000, 52,000,  and
282,000 ug/1, respectively.  Similar results have  been  obtained
with marine organisms.
      The dichloropropenes ara considerably more toxic  in acute
exposure than the dichloropropanes.  For 1,'3-dichlorpropene,
the 96-hour LC-Q value for the bluegill is 6,060 pg/1 compared
to 520,000 ug/1 for 1,3-dichloropropane.  For Daphnia magna,
the corresponding values are 6,150 and 282,000 pg/1,  respectively.
The ECSQ, based on chlorophyll a for a freshwater  alga, is 4,950
ug/1 for 1,3-dichloropropene, and 48,000 for 1,3-dichloropropane.
                                                           »
Data on measured residues could not be located  in  the available
literature for any saltwater or freshwater species.

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I.    INTRODUCTION

      This profile is based on the Ambient Water Quality Criteria

Document for Dichloropropanes/Dichloropropenes  (U.S. EPA, 1979).

      Dichloropropanes (molecular weight 112.99) and dichloropro-

penes (molecular weight 110.97) are liquids at environmental

temperatures.  Their boiling points range from 76 to 120.4°C

depending on the compound and the isomer.  They are slightly

denser than water, with densities ranging from 1.11 to 1:22.

The principal uses of dichloropropanes and dichloropropenes are

as soil fumigants for control of nematodes, in oil and fat sol-

vents, and in dry cleaning and degreasing processes (Windholz,

1976).  When heated to decomposition temperatures, 1,2-dichloropro-

pane emits highly toxic fumes of phosgene, while 1,3-dichloropro-r

pane gives off toxic fumes of chlorides  (Sax, 1975).  Production

of mixtures of dichloropropanes/dichioropropenes approached 50

million pounds per year prior to 1975  (U.S. EPA, 1979).

II.   EXPOSURE

      A.   Water

           Dichlorcpropanes and dichloropropenes can enter the

aquatic environment in discharges from industrial and manufactur-

ing processes, as run-off from agricultural land, and from munici-

pal effluents.  These compounds have been identified but not  .

quantified in New Orleans drinking water (Dowty/ et al.  1975).

      B.   Food

           Information was not found in the available literature
                                                            »
concerning the concentrations of dichloropropanes and dichloro-

propenes in commercial food stuffs.  Therefore, the amount of

these compounds ingested by humans is not known.  The U.S. EPA
                               7T-Y

-------
(1979) has estimated the weighted average bioconcentration  fac-
tors  (BCFs) of dichloropropanes and dichioropropenes to range
between 2.9 and 5.8 for the edible portions of fish and shellfish
consumed by Americans.  This estimate is based on the octanol/
water partition coefficients of these compounds.
      C.   Inhalation
           Atmospheric levels of dichloropropanes and dichioro-
propenes are not known.  However, from information on loss  of
these- compounds to the air after land application, it was esti-
mated that, in California alone, about 72 tons  (8 percent of
the pesticide used) were released to the atmosphere in 1971  (Calif,
State Dept.'Agric.  1971).
Ill.  PHARMACOKINETICS
      A.   Absorption, Distribution and Metabolism
           ?3rtinent information regarding tha absorption,  dis-
tribution, and metabolism of the dichloroprcpanes and dichioropro-
penes could not be located in the available information.
      B.   Excretion
           No human data are available on the excretion of  dichlor-
cpropanes or dichloroprcpenes.  In the rat, 30 to 90 percent  '
of an orally administered dose of dichloropropane or dichloropro-
pene was eliminated by all routes within 24 hours (Hutson,  et
al.  1971).  Approximately 50 percent of the administered dose
was eliminated in the urine within 24 hours."'
IV.   EFFECTS
      A.   Carcinogenicity
           Information concerning the carcinogenicity of mixtures
of dichloropropanes and dichioropropenes could not be located
                           79-f

-------
in the available literature.  However, cis-l,3-dichloropropane

has produced local sarcomas at the site of repeated subcutaneous

injections  (Van Duuren, et al., in press).  No remote  treatment-

related tumors were observed.

      B.   Mutagenicity

           Mixtures of 1,2-dichloropropane and 1,3-dichloropro-

pene are mutagenic to S^_ typhimurium strains TA 1535 and TA  100,

as are the individual compounds.  The mixture, but not the in-

dividual compounds, is also mutagenic to TA 1978  (in the presence

of microsomal activation) indicating a frame-shift mutation  not

capable of being produced by the individual compounds.

      C.   Teratogenicity and Other Reproductive Effects

           Pertinent information could not be located  in the

available literature.

      D.   Chronic Toxicity

           Inhalation exposure of rats, guinea pigs, and decs

to 400 ppm of 1,2-dichloropropane for 128 to 140 daily 7-hour

periods (5 days per week) decreased normal weight gain in rats

(Keppel, et al., 1948).  Inhalation exposures of rats  to 3 ppm

of 1,3-dichloropropene, 4 hours a day, for 125 to 130 days pro-

duced cloudy swelling in renal tubular epithelium which disap-

peared by 3 months after exposures ended (Torkelson and Oyen,

1977) .

V.    AQUATIC TOXICITY

      A.   Acute Toxicity
                                                           »
           Exposures of bluegill, Lepomis macrochirus, to 1,1-,

1,2-, and 1,3-dichloropropane under similar conditions yielded

96-hour LC5Q values of 97,900,  280,000, and greater than 520,000

-------
rag/1, respectively (U.S. EPA, 1978).  These data suggest that
toxicity decreases as the distance between the chlorine atoms
increases.  A reported 96-hour LC-Q for 1,3-dichloropropene is
5,060 ug/1 for the bluegill, approximately two orders of magni-
tude lower than for 1,3-dichloropropane (U.S. EPA, 1979).  Under
static test conditions, reported 48-hour LC5Q values for 1,1-,
1,2-, and 1,3-dichloropropanes are 23,000, 52,500 and 282,000
ug/1, respectively, (U.S. SPA, 1978)  for the only freshwater
invertebrate species tested, Daphnia magna.  The 48-hour LC-Q
value for 1,3-dichloropropene and Daphnia magna under static
conditions is 6,150 ug/1 (U.S. SPA, 1978).
           The 96-hour -Ceg values for the saltwater sheepshead
minnow, Cyprinodon variegatus, exposed to 1,3-dichloropropane
and 1,3-cichioroprcpane ware 36,700 jjg/'l and 1,770 ug/1, respec-
tively (U.S. SPA, 1973).  Dawson, et al. -(1977)  obtained a 96-
hour LCSO of 240,000 ug/1 for the tidewater silvsrsida, Manidia
beryllina, for exposure to 1,2-dichloropropane.
           For Mysidopsis sahia, the 96-hour LC^Q for 1,3-dichlcro-
propene was one-thirteenth thac for 1,3-dichloropropane, i.e.,
790 ug/1 and 10,300 ug/1, respectively (U.S. SPA, 1978)
      B.    Chronic Toxicity
           Chronic studies are limited to one freshwater study
and one saltwater study.  In an embryo-larval test,  the chronic
value for fathead minnows, Pimephales promelas,  exposed to 1,3-
dichloropropene was 122 ug/1 (U.S. EPA, 1978).   The  chronic value
                                                           »
for mysid shrimp, Mysidopsis bahia, was 3,040 ug/1 for 1,3-di-
chloropropane in a life cycle study (U.S.  EPA,  1978)
                             77-7

-------
      C.   Plant Effects


           For 1,3-dichloropropene, the 96-hour EC5Q values,


based on chlorophyll a concentrations and cell numbers of  the


freshwater alga, Selenastrum capricornutum, were  4,950 ug/1 and


4,960 ug/1, respectively.  The respective values  obtained  for


1,3-dichloropropane were 48,000 and 72,200 ug/1.  Thus,  the pro-


pene compound is much more toxic than the propane compound, as


is true for the bluegill and Daphnia magna.


      0.   Residues


           Measured steady-state bioconcentration factors  (BCF)


are not available for any dichloropropane or dichloropropene


in any fresh or saltwater species..  Based on octanol/water coef-


ficients of dichloropropanes and dichloropropenes, the U.S. EPA;


(1979) has estimated the bioconcentration factors for these com-


pounds to range between 10 and 35.


VI.   Other Pertinent Information


      In the non-aquatic environment, movement of 1,2-dichloro-


propane in the soil results from diffusion in the vapor  phase,


as these compounds tend to establish an equilibrium between the


vapor phase, water and absorbing phases (Leistra, 1970).   1,2-


dichloropropane appears to undergo minimal degradation in  soil


with the major route of dissipation appearing to  be volatiliza-


tion  (Roberts and Staydin, 1976).


      Following field application, movement'of 1,3-dichloropro-


pene in soil results in vapor-phase diffusion (Leistra,  1970).
                                                           »

The distribution of 1,3-dichloropropene within soils depends


on soil conditions.  For example, cis-l,3-dichlorobenzene  is


chemically hydrolyzed in moist soils to the corresponding  cis-

-------
3-chloroalkyl alcohol, which can be microbially  degraded  to  car-

bon dioxide and water by Pseudomonas sp.  (Van Dijk,  1974).

VII.  EXISTING GUIDELINES AND STANDARDS

      Neither the human health nor the aquatic criteria derived

by U.S. EPA (1979), which are summarized  below,  have gone  through

the process of public review; therefore,  there is  a  possibility

that these criteria may be changed.

      A.   Human

           The TLV for dichloropropane is 75 ppra (350 rag/m )

(Am. Conf. Gov. Ind. Hyg., 1977).  The draft water criterion

(U.S. EPA, 1979) for dichloropropane is 203 ug/I.  The draft

water criterion for dichloropropenes is 0.53 pg/1  (U.S.  EPA,

1979) .

      3.   Aquatic

           The dr'aft criteria for the dichioroprcpanes and gi-

chloropropsr.as to prctact frashwatar aquatic life  are as  follows

(U.S. EPA, 1979):


                                               Concentration not
                                                 to be exceeded
Compound                  24-Hour Average          at any  time

1,1-dichloropropane           410 pg/1                930 ug/1

1,2-dichloropropane           920 ^ug/1             2,100 ug/1

1,3-dichloropropane         4,800 ug/1             11,000 ug/1

1,3-dichloropropene            L8 pg/1                250 ug/1


The draft criteria to protect saltwater species  are  as follows
                                                           #
(U.S. EPA, 1979):

-------
                                               Concentration  not
                                                to be exceeded
Compound                  24-Hour Average         at any time

1,1-dichloropropane         not derived          not derived

1,2-dichloropropane          400 jig/1              910 ug/1

1,3-dichloropropane           79 pg/1              180 pg/1

1,3-dichloropropene          5.5 ng/1               14 pg/1

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                       DICHLOROPROPANES/DICHLOROPROPENE5

                                  REFERENCES
American  Conference  of Governmental  Industrial  Hygienists.   1977.  Documen-
tation of the threshold limit values.  3rd. ed.

California  State  Department of  Agriculture.   1971.   State  pesticide  use
report.

Oawson, G.W.,  et al.  1977.  The acute toxicity  of 47 industrial chemicals
to fresh and saltwater fishes.  Jour. Hazard. Mater.  1: 303.

Oowty, 8.,  et al.   1975.   Halogenated  hydrocarbons  in  New  Orleans drinking
water and blood plasma.  Science  37: 75.

Heppel, L.A.,  et  al.   1948.   Toxicology  of  1,2-dichloropropane  (propylene
dichloride).  IV. Effect of  repeated  exposures  to  a low concentration of the
vapor.  Jour. Ind. Hyg. Toxicol.  30: 189.

Hutson, D.H.,  et al.  1971.  Excretion and  ratantion  of comocnents  of the
soil  fumigant  0-OAR) and their metabolites  in the  rat.   Food Cosrnet. Toxi-
col.  9: 677.

Leistra,  M.   1970.  Distribution of 1,3-dichloropropene  over the pnase in
soil.  Jour. Agric. Food Chem.  13: 1124.

Roberts,  R.T.   and  G.   Stoydin.    1975.   The   degradation  of  (2)- arc
(E)-l,3-di-  chioropropenes  and  l,2-dichioroproper.5s  in soil.  Psstic.  Sci.
7: 325.

Sax,  N.I.   1975.  Dangerous properties of  industrial  materials.   Reinhoid
Book Corp.,  New York.

Torkelson, R.R.  and  F.  Oyen.  1977.  The  toxicity of 1,3-dichloropropene as
determined by  repeated  exposure of laboratory animals.  Jour.  Am.  Ind.  Hyg.
Assoc.  38:  217.

U.S.  EPA.   1978.  In-depth  studies on  health  and environmental  impacts of
selected water pollutants.  Contract No. 68-101-4646.

U.S.  EPA.   1979.  Dichloropropanes/Dichloropropenes:   Ambient Water Quality
Criteria.   (Draft).

Van  Oijk,  J.    1974.   Degradation  of 1,3-dichloropropenes   in   the  soil.
Agro-Ecosystems.  1: 193.

Van Ouuren,  8.L., et al.   1979.   Carcinogenicity of halogenated olefinic and
alipahtic hydrocarbons.  (In press).                                    •

Windholz,  M.,  ed.  1976.   The  Merck Index.   9th  ed.  Merck  and  Co.,  Inc.,
Rahway, N.J.
                                 77-//

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                                      No. 80
          Dichloropropanol


  Health and Envirorraental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not  reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                               DICHLOROPROPANOL


                                    Summary






     There  was  no  evidence  found  in  the  available  literature  to  indicate


that exposure to dichloropropanol produces  carcinogenic  effects.   Conclusive


evidence  of mutagenic,  teratogenic, or  chronic effects  of dichloropropanol


was not found in  the available literature.   Acute exposure  results  in toxi-


city similar  to  that induced by carbon  tetrachloride,  including  hepato- and


nephrotoxicity.   Data concerning the effects of dichloropropanol  to aquatic
                                                           •

organisms was not found in the available literature.

-------
I.   INTRODUCTION
     This profile is based  on computerized searches of Toxline,  Biosis,  and
Chemical Abstracts, and  review of other  appropriate information sources  as
available.  Oichloropropanol  (molecular weight 128.9), a  colorless,  viscous
liquid with a chloroform-like odor, refers  to four iscmers  with the  mole-
cular  formula  C-3H6OC12.   The  physical  properties  of   each   isomer  are
given below.
                        Boiling Point Density     Solubility (Weast.  1976)
                                                 Water      Alcohol    Ether
2,3-Oichloro-l-propanol    182QC      1.368    slight      miscible   miscible
l,3-Oichloro-2-propanol    1740c      1.367    very        very      miscible
3 , 3-Oichloro-l-pfooanol   32-33°C     1.316    not listed
1 , l-Oichioro-2-prcpanol  146-1 AO^C     1.3334   slight      very      very
     Additional physical data  and synonyms of  the  above isomers are  avail-
able in Heilbron  (1965), Fairchild (1979)-,  Sax (1979), Windholz  (1976),  ana
Verscnueren (1977).
     Oichioropropanol is prepared from glycerol,  acetic acicj, and  hydrogen
chloride.   It  is  used as a  solvent for hard  resins  and nitrocellulose,  in
the manufacture  of photographic  and  Zapon lacquer,  as  a cement for  cellu-
loid,  and  as  a binder for water colors (Windholz, 1976).   The compound  is
considered to ba a moderate  firs  hazard when  exposed  to  neat,  flame, or oxi-
dizers, and a disaster hazard  in  that it may decompose at high  temperatures
to phosgene gas (Sax, 1979).
II.  EXPOSURE
     Dichloropropanol was detectable  in  the air of a  glycerol manufacturing
plant in  the  U.S.S.R. (Lipina and  Selyakov,  1975).  Unreacted  dichloropro-
panol was  also  found in  the  wastewater effluent of a  halohydrin  manufactur-
ing plant  (Aoki and  Katsube,  1975).   No  monitoring  data are  available  to
indicate ambient air or water levels of the compound.

-------
     Human exposure  to  dichloropropanol from  foods  cannot be assessed,  due
to a lack of monitoring data.
     Bioaccumulation data on dichloropropanol  was not  found  in the available
literature.
III. PHARMACOKINETICS
     Pertinent data  could not  be located in the available literature  on the
metabolism, distribution, absorption, or excretion of dichloropropanol.
IV.  EFFECTS
     A.   Carcinogenicity
          Pertinent data could not be located in the available literature.
     8.   Mutagenicity
          2,3-OichloropropanoI  and  1,3-dichloroprcpanol were evaluated  for
mutagenicity by  a modified  Ames assay using  S_._ typhimurium  strains.   Some
evidence of  rnutagenic  activity was  seen,  but  the authors fait  that  further
evidence and clarification  of the metabolic activation pathway to  mutagens
             I
via halcalkanols were r.ecessary (Nskamura,  et al. 1579).
     C.   Teratogenicity, Other Reproductive Effects and Chronic Toxicity
          Pertinent data could not be located in the available literature.
     0.   Acute Toxicity
          2,3-Dichlcropropanol was  found to  have  an  oral LD^  j_n the  rat
of  90  mg/kg.  The  lowest  published lethal  concentration  (LCLQ) in rats is
500 ppm  by inhalation for 4  hours.   A dose of 6,800 ug  in  the eye  of  the
rabbit caused  severe irritation (Fairchild,  1979).   1,3-Dichloropropanol was
found  to have an oral LD5Q  in the rat  of 490 mg/kg'and lowest  published
lethal concentration for inhalation  exposure in rats of 125  ppm/4 hrs.   Ten
                                                                          »
mg  applied to  the skin of the rabbit  for 24 hours  produced  mild irritation,
and  800 mg/kg was  the LD50  for  the  same  route  and species  (Fairchild,
1979).

-------
            Several  references  report  the  clinical  indications  of acute  di-
  chloropropanol intoxication as being  similar to carbon tetrachloride poison-
  ing,  i.e.,   central  nervous  depression;  hepatotoxicity,  including" hepatic
  cell  necrosis and fatty  infiltration; and  rsnal toxicity,  including  fatty
  degeneration  and  necrosis of the  renal  tubular epithelium (Sax,  1979;  Gos-
  selin, et al. 1976).
  V.   AQUATIC TOXICITY
       Data concerning the  effects  of  dichloropropanol  to aquatic  organisms
  were not fou.nd in the available literature.
  VI.  EXISTING GUIDELINES AND STANDARDS
       A.   Human
            The  maximum  allowable  concentration  of  dichloropropanol in  the
  working  environment  air  in  the U.S.S.R.  is 5  mg/m3 (lipina and  Belyakov,
  1975).
            The maximum allowable ccncar.tration in Class  I  waters  for  the  pro-
  duction of drinking water is 1 mg/1 (verschueren, 1577).
       B.   Aquatic
            The organoleptic limit in water set in the U.S.S.R.  (1970)  is 1.0
.  mg/1 (Verschueren, 1977).

-------
                                  REFERENCES
Aoki, S. and E.  Katsube.   1975.  Treatment of  waste  waters from haiohydrin
manufacture.  Chem. Abs.  CA/083/15875D.

Fairchild,   E.  (ed.)   1979.    Registry  of  Toxic  Effects  of  Chemical  Sub-
stances.  U.S.  Department of  Health, Education  and  Welfare, National Insti-
tute for Occupational Safety  and Health,  Cincinnati, Ohio.

Gosselin, et al.   1976.   Clinical Toxicology of  Commercial Products.   Wil-
liam and wilkins  Publishing Co., Baltimore,  Maryland.

Heilbron, I.  (ed.)  1965.  Dictionary of Organic Compounds.   4th  edition.
University Press, Oxford.

Lipina, T.G. and  A.A.  Belyakov.  1975.-   Determination of allyl alcohol, al-
lyl chloride, epichlorohydrin and dichlorohydrin in the air.  Gig. Tr.  Prof.
Zabol.  5:  49.

Nakamura. A.,  et al.  1979.   The mutagenicity  of halogenated  alkanols and
their  phosphoric  acid  esters  for  Salmonella   tyohimurium.   Mutat.   Res.
66: 373.

Sax,  N.I.   1979.  Dangerous  Properties  of Industrial  Materials.   Van  Nos-
trand Reinhold  Co., New York.

Verschueren, !<.   1977.  Handtcc-k of Environmental Data on Organic Chemicals.
Van Ncstrand Reinhold Co., New York, p. 659.
                                  i
V/east,  R.C.  (ed-.)   1976.  Handbook  of  Chemistry  and Physics.   CRC Press,
Cleveland,  Ohio,  p. c-^54.

Windholz, M. (ed.)  1976.  The  Merck Index.  9th ed.  Merck and Co., Rahway,
New Jersey.
                                       Sro-7

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                                      No. 81
        1,3-Dichloropropene


  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                              1.3-OICHLOROPROPENE
                                    SUMMARY

     The major environmental  source  of dichlorcpropenes is from the use of  a
mixture  of dichloropropenes  and dichloropropanes  as  a soil  funigant.   On
chronic exposure of  rats  to  dichloropropene mild kidney damage was observed.
Dichloropropene has  produced subcutaneous  tumors  at the  site  of injection,
and has  been  shown to be mutagenic  in bacteria.  However, not enough  infor-
mation is available to classify this compound as a carcinogen.
                                                      o
     The bluegill  (Leoomis  macrochirus)  has  a reported 96-hr  LC5Q value of
oCoG uc/1;  Saohnia magna has a reported  43-hr  LC5Q of 5130 ug/i.   ^or the
saltwater  invertebrate,   Mysidoosis  bahia,  a reported  96-hr LC--  value is
790 /jg/1.   In the  only   long-term  study  available,  the  value  obtained for
1,3-dichloropropene  toxicity  to fathead  minnows  (Pimeohales oromeles) in an
embryo-larval  test  is 122 jug/1.   Based en chlorophyll  a  concentrations and
cell  numbers,  the  96-hr EC-- values  for  the  frsshwazar  alga  Selsnastrun
caorieornutum  are  4,950  and  4,960 ug/1,  respectively:  for  the' marine alga
Skeletonema costatum, the respective values are- 1,000 and  1,040 ;jg/l.

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                              1,3-DICHLOROPROPENE

I.   INTRODUCTION

     This  profile  is based  on the  Ambient  Water Quality  Criteria Document

for Dichloropropanes/Dichloropropenes (U.S. EPA, 1979a).

     1,3-dichloropropene  (molecular  weight 110.97)  is a  liquid  at environ-

mental temperatures.  The  isomers  of 1,3-dichloropropene  have boiling points

of  104.3°C  for the trans-isomer  and  112°C  for  the cis-isomer,  and  the

densities  are  1.217  and 1.224 g/ml,  respectively.   The water solubility for

the   two  isomers   is   approximately   0.275   percent.    When   heated  to

decomposition  temperatures,   1,3-dichloropropene gives  off  toxic  fumes  of

chlorides  (Sax,  1975).   Mixtures  of cis-  and trans- 1,3-dichloropropene and

1,2-dichloropropane  are   used  as   soil   fumigants.    In   this  document,

dichloropropene will refer to either cis- or trans-1,3-dichloropropene.  For

more  information  regarding the dic'nloroprcoanss,  tha reader is  rsfsrrsd to

the EPA/ECAC  Hazard  Profile  on  Dichlorcpropanes/Dichloropropenes  (U.S. EFA,

1979b).

II.  EXPOSURE

     A.  Water

         Dichloropropene  can  enter the  aquatic environment in the  discharges

from  industrial and manufacturing  processes,  in  run-off  from  agricultural

land,  or from  municipal  effluents.   This compound  has  been  identified but

not quantified  in New Orleans drinking water  (Dowty, et al. 1975).

     B.  Food

         Information was   found  in  the  available  literature  concerning the

concentration of dichloropropene in  commercial  foodstuffs.   Thus,  the amount
                                                                       »
of this  compound  ingested by humans  is  not  known.   The U.S.  EPA (1979a) has

estimated  the weighted  average  bioconcentration factor (BCF)  of dichloropro-

pene  to  be 2.9 for  the  edible portions  of  fish  and  shellfish  consumed by

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Americans.   This  estimate  is  based  on  the dctanol/water  partition coeffi-
cient of dichloropropene.
     C.  Inhalation
         Atmospheric levels  of  dichlcrcprcpene have not been measured.  How-
ever, it  is  estimated  that about 8  percent of the  dichloropropene which is
applied to  the soil as  a fumigant  is released to  the  atmosphere  (U.S. EPA,
1979a).
III. PHARMACOKINETICS
                                                    ^.
     A.  Absorption
         Data  on  the absorption,  distribution and metabolism of dichloropro-
pene could not be located in the available  literature.
         Data  on  the  excretion of  dichloropropene by  humans could  not be
located in  the available literature.  In the  rat,  however, approximately 80
percent of  an  orally  aoninistered dose of  dicnioropropene  was eliminated in
the urine within 24 hcurs (Hutson, et al. 1971).
IV.  EFrECTS
     A.  Carcinogenicity
         Van Ouuren, et al. (1979)  investigated  the ability of dichloropro-
pene  to  act as a tumor  initiator  or promoter  in mouse  skin,  or to cause
tumors  after  subcutaneous  injection.   Dichloropropene  showed  no  initiation
or  promotion activity, and  only  local  sarcomas developed  in mice  following
subcutaneous administration.   In  none of the  studies were treatment-related
remote tumors  observed.
                                                       ^
     8.  Mutagenicity
         DeLorenzo, et al.  (1977) and Neudecker,  et al. (1977) reported tjhat
dichloropropene was mutagenic in  S.  typhimurium strains TA1535 and  TA100 but
not  in  TA1978, TA1537, or TA98.  Results did  not differ with or without the

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addition of  liver microsomal fraction.   Neudecker,  et al.  (1977)  found the


cis-isomer to be twice as reactive as the trans-isomer.


     C.  Teratogenicity  and Other Reproductive Effects


         No  pertinent  information  regarding  the  teratogenicity  and  other


reproductive effects could not be located in the available literature.


     D.  Chronic Toxicity


         On exposure  of rats to 3 ppm  dichloropropene for period of 0.5,  1,


2 or 4  hours/day,  5  days a week for 6 months  (Torkelson  and Oyen,  1977),  or


rats, guinea pigs, and  rabbits  to  1  or 3 ppm of dichloropropene, 7 hours per


day for 125-130 days over a  180-day period,  only rats exposed 4 hours/day  at


3.0 ppm  showed  an effect (U.S. EPA,  1979a).   The only effect observed was a


cloudy  swelling  of  the  renal  tubular  epithelium  which  disappeared  by  3


months after exposures ended.


V.   AQUATIC TOXICITY


     A.  Acute Toxicity


         Tests  on the  bluegill,  Legcmis .Tiacrochirus, yielded  a  96-hr LC5Q


value of 6060 )jg/l for  1,3-dichloroprcpene  exposure.  For Daphnia maona, the


48-hr  LC5Q  value  is  6,150 jug/1  (U.S.  EPA,  1978).   The  observed  96-hr


LC5Q  for the  saltwater my rid  shrimp,  Mysidopsis  bania,  is  790  jjg/1 (U.S.


EPA, 1973).


     B.  Chronic Toxicity


         An  embryo-larval test  has  been conducted  with  the  fathead minnow


(Pimephales  promeles)  and 1,3-dichloropropene.   The.  observed  chronic value


was 122 jjg/1 (U.S. EPA, 1979a).


     C.  Plant Effects
                                                                         »

         Based  on chlorophyll  a  concentrations and cell  numbers,  the 96-hr


EC5Q  values  for  the freshwater alga,  Selenestrum  caoricornutum,  are 4,950

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and  4,960  pg/1, respectively  (U.S.  EPA,  1978).   The  respective  values for
the saltwater alga Skeletonema ccstatum  were  1,000 and 1,040 jug/1 (U.S. EPA,
1978).
     0.  Residues
         Measured steady-state bioconcentration  factors (8CF) are not avail-
able for 1,3-dichloropropene.  A BCF of 19 has  been estimated  based on the
octonol/water coefficient for 1,3-dichloropropene (U.S. EPA, 1979a).
     E.  Other Relevant Information
         Following  field  application,   movementsof 1,3-dichloropropene  in
soil results  in vapor-phase diffusion (Leistra, 1970).   The distribution of
1,3-dichloropropene within  soils depends  on  soil  conditions.   For example,
cis-l,3-dichlcroprcpane is  chemically hydroiyzed in moist soils  to  the cor-
responding cis-3-chloroalkyl  alcohol,  which  can be  microbially degraded to
carbon dioxide and water by Pseudomonas so.  (Van Dijk  1974).
VI.  EXISTING GUIDELINES AND STANDARDS
     Neither  the  human health nor the  aquatic criteria  derived by  U.S. EPA
(1979a), which are summarized below, have  gone  through the process of public
review;  therefore,  there   is  a  possibility  that  these criteria  will  be
changed.
     A.  Human
         The  draft water   criterion  for  1,3-dichloropropene  is 0.63  /jg/1
(U.S. EPA,  1979a).
     B.  Aquatic
         The  draft criterion  to protect freshwater'species is  18 /jg/1  as a
24-hr average not  to  exceed 250 jug/1 at any  time.   For  marine  species, the
                                                                      »
value is 5.5-jug/l as  a 24-hr  average not to exceed 14 ;jg/l at any time  (U.S.
EPA, 1979).

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                              1,3-DICHLOROPROPENE

                                  REFERENCES
OeLorenzo, p.,  et al.  1977.  Mutagenicity of  pesticides containing 1,3-di-
chloropropene.  Cancer Res. 37: 6

Dowty, 8.,  et al.  1975.   Halogenated  hydrocarbons in  New  Orleans drinking
water and blood plasma.  Science 87: 75.

Hutson,  O.H.,  et al.   1971.   Excretion  and  retention of components  of the
soil, fumigant  QrQ^>  and  their   metabolites   in   the   rat.   Food  Cosmet.
Toxicol.  9: 677.

Lsistra,  M.  .197,0.   Distribution  of 1,3-dichloropropene over the  phase in
soil.  Jour. Agric. Food Chem.  18: 1124.

Neudecker, T.-,  et al.  1977.  In  vitro  mutagenicity of  the  soil nematocide,
1,3-dichloropropene.  Experientia 33: 8.

Sax,  M.I.   1975.   Dangerous  Properties  of  Industrial  Materials.   Reinhold
Book Corp., New York.

Torkelson, R.R. and F. Oyen.   1977.   The toxicity of 1,3—dichloroprcpene as
determined by  repeated  exposure  of laboratory animals.   Hour.  Am.  Ind. Hyg.
Asscc.. 38: 217.     .

U.S.  EPA.   1978.   In-depth  studies on  health,  and  environmental  impacts of
selected water pollutants.  Contract No. 63-01-4646.

U.S.  EPA.   1979a.  Oichloropropanes/Oichloropropenes: Ambient  Water Quality
Criteria (Draft)

U.S.  EPA.   1979b.   Dichloropropanes/Dichloropropenes:  EPA/ECAO  Hazard Pro-
file.

Van Dijk, J.   1974.   Degradation of 1,3-dichloropropenes in  the soil.   Agro-
Ecosystems.  1: 193.

Van  Duuren,   et  al.   1979.   Carcinogenicity  at  halogenated  olefinic  and
aliphatic hydrocarbons.  (In press).

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                                       No. 82
              Dieldrin


  Health and EirrLroTaaental  Effects
U.S. ENVIRONMENTAL PROTECTION  AGENCY
       WASHINGTON, D.C.   20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

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                       SPECIAL NOTATION









U.S. EPA1s Carcinogen Assessment Group (CAG) has evaluated



dieldrin and has found sufficient evidence to indicate that



this compound is carcinogenic.

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                           DIELDRIN
                           SUMMARY
      Dieldrin is a compound belonging  to  the  group  of  cyclodiene
insecticides.  The chronic toxicity of  low doses  of  dieldrin
includes shortened life span, liver changes  ana  teratogenic effects.
The induction of hepatocellular carcinoma  in mice by dieldrin
leaas to the conclusion that it Is lively  to" be  a human carcinogen.
Dieldrin has been found to be non-mutagenic  in several  test sys-
tems.  The WHO'S acceptable daily  intake for dieldrin is 0.0001
mg/kg/day.
      The toxicity of disidrin to  aquatic organisms  has been
investigated in numerous studies.  The  96-hour i->Ccn  values for
the common freshwater fish range from 1.1  to 360  ug/1.   The acute
toxicity is considerably more varied for ft.esnwacer  invertebrates,
with 96-hcur I^Q values ranging frcrr, 0.5 ug/1 for  the  stonefiy
to- 7^0 ,ug/l for the crayfish.  Acute i-Ccn values  for eight salt-
water fish species range from 0.66 to 24.0 ug/1  in  flow-through
tests; LC,-Q values for estuarine invertebrates range from 0.70
to *:40 ug/1.  The only reported chronic values are  0.11 jug/I
for steel head trout  (Salmo guirdnes) in an  emoryolarval study
ana 0.4 ug/1 for the guppy (Poecilia reticulata)  in  a life-cycle
test.  Both fresh and salt water algae  are less  sensitive to
dieldrin toxicity than the corresponding fish  ana inverteorates.
Bioconcentration factors were 128  for a freshwater  alga,  1395
for Daphnia magna, 2993 for the channel catfish,  ana 8000 for
        — -       	                                            #
the edible tissues of the Eastern  oyster.

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                             DI2LDRIN

I.     INTRODUCTION

      This  profile  is  based on  the draft Ambient  Water Quality

Criteria  Document  for  Aldrin  and  Dieldrin   (U.S.  EPA,  1979).

Dieldrin  is a  white  crystalline  substance  with  a  melting point

of 176-177°C and  is soluble  in  organic solvents (U.S. EPA, 1979).

The  chemical  name  for  dieldrin  is  1,2,3,4,10,10-hexachlor-6,7-

epoxy-1,4,4a,5,6,7,8,8a-octohydro-endo, exo-1,4:5,3-dimethanonaph-

thalene.                                    v

      Dieldrin  is extremely  stable and persistant in the  environ-

ment.   Its  pe'rsistance is  due  to  its extremely low  volatility

(1.78  x   10"7  mm  Hg  at 20°C)  and  low  solubility  in  water   (186

ug/1 at 25-29°C).   The time required for 95 percent of the  dieldrin

to disappear  from soil  has  been  estimated  to vary  from  5  to 25

years  depending  on  the  microbial  flora  of  the  soil  (Edwards..

1966).   Patil, et  al.  (1972)  reported that  dieldrin was not  de-

graded or metabolized in sea water or  polluted water.

      Dieldrin was  primarily used as a broad  spectrum  insecticide

until  1974,  when  the  U.S.  EPA  restricted  its  use  to termite  con-

trol by direct soil injection,  ana non-food seed  and plant treat-

ment  (U.S.  EPA, 1979).   From 1966 to 1970, the amount of  dieldrin

used  in  the  United  States  decreased  from 500  to  approximately

335,000  tons  (U.S. EPA,  1979).   This decrease  in  use has   been

attributed  primarily  to  increased  insect resistance to  dieldrin

and to development  of  substitute materials.   Although the produc-
                                                              »
tion  of  dieldrin is  restricted in  the  United States,   formulated

products  containing  dieldrin  are  imported  from  Europe  (U.S.

EPA, 1979).

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II.   EXPOSURE

      A.   Water

           Dieldrin  has  been  applied to  vast areas  of agricul-

tural land  and aquatic  areas  in the  United  States,  and  in most

parts of  the  world.   As  a  result,   this  pesticide  is  found  in

most  fresh  and  marine  waters.    Dieldrin has  been  measured   in

many  freshwaters  of  the United States,  with  mean  concentrations

ranging from 5  to  395  ng/1 in surface water and  from  1 to 7 ng/1

in  drinking  water  (Epstein,  1976) .   Levels  as  high  as  50 ng/1

have  been  found in  drinking water  (Karris,  et  al.  1977).   The

half-life of  dieldrin in water, 1  meter  in depth,  has been esti-

mated to be 723 days  (i-iacKay and Wolkoff, 1373).

      B.   Food

           Dieldrin  is  one  of  the  most  stable  and persistant

organcchiorine  pesticides  (Nash and  Woolson,  1957} ,  and  because

it  is  lipophilic,  it  accumulates  in the  food  chain (Wurster,.

1971).   Its persistence in  soil  varies  with  the  type of  soil.

(Matsumura and Boush, 1967).

           The  U.S.   EPA  (1971)  estimated  that  99.5  percent   of

all  human  beings  have dieldrin  residues  in their  tissue.   These

residues  are  primarily  due  to  contamination  of  foods of animal

origin.    The  overall concentration  of  dieldrin  in  the  diet  in

the  United  States  has  been  calculated  to  be  approximately   43

ng/g of food consumed  (Epstein,  1976).  The U.S. EPA has estimated

the  weighted  average  bioconcentration  factor   for  dieldrin   to
                                                             »
be  4,50.0  in  the  edible  portion of  fish  and  shellfish consumed

by Americans (U.S. EPA,  1979).   This  estimate  is based  on measured

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steady-state bioconcentration  studies  in  several  species of  fish
and shellfish.
      C.   Inhalation
           Dieldrin  enters  the  air  through  various mechanisms,
such  as spraying,  wind  action,  water  evaporation,  and  adhesion
to  particulates.    The  U.S. EPA detected  dieldrin  in  more  than
85  percent of  the  air  samples tested  between  1970-1972,   with
the  mean  levels  ranging  from  1  to 2.8  ng/m   (Epstein, 1976).
From these levels,  the average daily intake of dieldrin by respi-
ration was calculated to be 0.035 to 0.098 ug.
           Although  dieldrin  is no longer  used  in  the United
States,  there  is  still  the possibility of airborne  contamination
from ocher parts of the world.
      D.   Cerraai
           Dermal .exposure to dieldrin is  limited  to  those  in-
volved  in  its manufacture  or  application as  a  pesticide.   Wolfe,
et al.  (1972)  rapcrtad that exposure in workers was mainly through
dermal absorption rather than  inhalation.  The ban on the manufac-
ture  of dieldrin  in  the  United States  has greatly  reduced  the
cis'< of e:-:pc:s_re.
III. PHARMACOKINETICS
      A.   Absorption
           The   absorption  of  dieldrin  by  the  upper  gastrointes-
tinal  tract  begins almost  immediately  after  oral administration
in  rats and has  been found to  vary with  the amount  of solvent
used  (Heath and Vandekar,  1954).  These authors also demonstrated
that absorption takes place via  the portal vein, and  that dieldrin

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could be recovered from the stomach, small  intestine,  large  intes-

tine and feces one hour after oral administration.

      B.   Distribution

           The distribution of dieldrin has been  studied  in  numer-

ous  feeding  experiments.   Dieldrin  has  an affinity  for  fat,  but

high  concentrations  are  also reported  in  the  liver  and kidney,

with moderate concentrations  in  the  brain one and two hours  after

administration  in rats   (Heath  and Vandekar,  1964).    Deichman,

et  al.  (1968)  fed dieldrin  to  rats  for a-  period of  183   days.

The mean concentration in the fat was 474 times  that in the  blood,

while the  concentration  in  the  liver  was  approximately  29   times

the blood concentration.

           Additional animal  studies on  the distribution of  diel-

drin  have  shown that concentrations in   tissues  are  dose related

and  IT.ay  vary with the  sex cf the  animal  ("yalk-=r:;  et al.   136^).

Matthews, et  al.  (1971)  found that  female  rats administered oral

doses of  dieldrin had  higher  tissue levels  of  the compound than

male rats.   The  females  stored  the  compound predomir.atly as  diel-

drin.   In males,  other metabolites,  identified as keto-dieidrin

trans-hydro-aidrin and a polar metabolite, were  detected.

           The  concentrations  of dieldrin  in  human  body  fat were

found to be  0.15  + 0.02 ^ag/g for the  general population and 0.36

ug/g  in  one   individual  exposed  to   aldrin  (aldrin is metabolized

to  dieldrin)  (Dale and  Quinby,  1963).   The  tnean concentrations

of  dieldrin   in  the  fat,  urine,  and plasma  of  pesticide workers
                                                              »
were 5.67, 0.242  and  0.0185  mg/g,  respectively  (Hayes and Curley,

1963).   Correlations  between, the   dose  and  length  of  exposure

to  dieldrin  and  the  concentration   of dieldrin  in  the  blood  and

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other tissues have been reported  (Hunter, et  al.   1969).   Dieldrin
residues  in  the  blood  plasma  of workers  averaged  approximately
four  times  higher  than that  in the  erythrocytes  (Mick,  et  al.
1971).
      C.   Metabolism
           The epoxidation of aldrin to dieldrin  has  been  reported
in many organisms,  including man (U.S. EPA,  1979).   The  reaction
is  NADPH-dependent,  and the enzymes  have  been  found  to be  heat
labile  (Wong and Terriere, 1965).
           The metabolism  of dieldrin  has been studied in several
species,  including  mice,  rats,  rabbits,  and  sheep.     Dieldrin
metabolites  have  been  identified  in  the  urine  and  faces in  the
form  of several  compounds  more  polar  than  the  parent   compound
(u.S. E?A,  1979).   Bedford  and  Hutscr. (1975) summarized  the  four
.
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as in  the  feces.   Robinson,  et  al.  (1969)  found  that 99 percent



of the  dieldrin  fed  to rats  for  8  weeks was  excreted  during a



subsequent 90-day  observation period.  The  half-life of dieldrin



in the liver and blood was 1.3 days for the period of  rapid elimi-



nation and  10.2 days  for  a  later, slower period.   The half-life



of dieldrin  in  adipose tissue and brain  were 10.3  and  3.0 days,



respectively.



           The  concentration  of   dieldrin in  the  urine  of   the



general  human  population  is  0.3   mg/1  for man  and   1.3  mg/1   for



women as  compared to  5.3,  13.8,   or  51.4  mg/1  for  men  with low,



medium, or high exposure  (Ceuto  and Biros,  1967).   Tha half-life



for  dieidrin in  the  blood  of  humans ranges  from  141-592  days



with  a  mean of 369  days  (Hunter,  et al.  1969).    Jager   (1970)



reported the half-iifa tc  be 265   days.  Because there is a rela-



tionship between  the  concentration of  dieidrin in  che  blood   and



that  in adipose and  other  tissues, it seems  likely that che half-



life  in  the blood  may  reflect  the  over-all half-life  in  other



tissues (U.S. EPA, 1979).



IV    EFFECTS



      .-..    Care i nog en ic i ty



           Dieldrin  has  produced   liver tumors  in  several strains



of mice according to  six reports of chronic  feeding  studies (NCI,



1976  (43  FR 2450);  Davis  and Fitzhugh,   1962;  Davis,  1965;  Song



and Harville, 1964; Walker, et al.  1972; Thorpe, and Walker, 1973).



In rats,  dieldrin has  failed to induce a statistically significant



excess of  tumors  at  any site in three  strains  during six chronic



feeding  studies  (Treon  and  Cleveland,   1965;  Cleveland,  1966;

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Fitzhugh,  et'  al.  1964;  Deichman,   et  al.  1967;  Walker,  et  al.



1969; Deichmann, et al. 1970).



           The only information concerning  the  carcinogenic  poten-



tial  of dieldrin  in  man is  an  occupational  study  by  Versteeg



and Jager  (1972) .   The workers had  been  employed  in a  plant  pro-



ducing aldrin and dieldrin with a mean exposure time  of  6.6  years.



An  average  of 7.4  years had  elapsed since  the  end of  exposure.



No permanent adverse effects,  including cancer, were  observed.



      B. Mutagenicity
                                               •             .


           Microbial assays  concerning  the mutagenicity  of  aldrin



and  dieldrin  have yielded  negative  results  even when  some  type



of  activation system  was added  (Fahrig,  1S73; Bidwell,  et  al.



1575; Marshall,  et  al. 1976) .   A host-mediated assay and a domi-



nant  lethal  test,  also  yielded  negative  results   (Bidwell,   et



ai.  1975).    L-iajumdar,  et  al.   (1977),  however,   fcund  dial-rin



to  be  mutagenic in  S.  typ.h iph imur i urn,  although  these  positive



results  wera  questioned  because  several differences existed  be-



tween their procedures  and those recommended  (U.S. EPA,  1979).



           A  decrease  in the  mitotic index  was  observed in  vivo



v;itli  .T.CUSS  bone  marrow  cells  and _in y i erg  wicn  human  iung cells



treated with 1 mg/kg and  1 ^ig/ml dieldrin,  respectively  (Majumdar,



et al. 1976).



      D.   Teratogenicity


                                                          14
           In  1967,  Hathaway,   et  al. established  that   C-diel-



drin could cross  the placenta in rabbits.  Dieldrin caused signifi-

                                                              »

cant  increases  in  fetal death in  hamsters, and  increased fetal



anomalies  (i.e.  open  eye, webbed foot,  cleft palate, and others)

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in hamsters  and  mice when  administered. in single oral doses  dur-

ing  gestation  (hamsters  50,  30,  5 mg/kg  and  mice  25,  15,   2.5

mg/kg) (Ottolenghi, et al.  1974).

           However,  in subsequent  studies  no  evidence  has  been

found  that  dieldrin causes  teracogenic  effects  in mice and  rats

(Chernoff, et al. 1975) or  mice  (Dix, et al. 1977).

      D.   Other Reproductive Effects

           Deichmann  (1972)  reported  that  aldrin   and  dieldrin

(25  mg/kg/diet)  fed to mice  for six generations affected  ferti-

lity , gestation,  viability, lactation,  and survival of the  young.

However,  no  changes in weight  or  survival of  fetuses  were found

in mice  administered dieldrin  for  day 5  through 14  of gestation

at doses already  mentioned in  this report   (Ottolenghi,  et   al.

1374).

      E.   Chronic  Toxicity

           The other effects  produced  by  chronic administr -sticr.

of dieldrin  to mice,  rats, and dogs  include  shortened  life span,

increased  liver  to  body  weight  ratio,  various  changes  in liver

histology, and the  induction of hepatic enzymes  (U.S. EPA,  1979).

      F.   Gch=r Relevanc  Information

           Since  aldrin  and  dieldrin are  metabolized  by  way of

the  mixed function  oxidase  (MFC)  system  and  dieldrin  has  been

found  to  induce the  production  of these  enzymes,   any  inducer

or  inhibitor of the  MFO  enzymes   should  affect  the  metabolism

of dieldrin  (U.S.  EPA,  1979).   Dieldrin  fed  in low dose9-^i£_ior
                                                               »
to  an  acute dose  of dieldrin alters  its  metabolism  (Baldwin,

et al.  1972).   Dieldrin can effect  the  storage of DDT  (U.S.  EPA,

-------
1979) and  induce a. greater number of  tumors in mice when  admini-
stered with DDT as compared to DDT alone  (Walker, at  al.  1372).
V.    AQUATIC TOXICITY
      A.   Acute Toxicity
           The  acute  toxicity of  dieldrin  has  been investigated
in numerous studies.  Reported 96-hour LC5Q values for  freshwater
fish are 1.1 to 9.9 ug/1 for rainbow trout,  Salmo gairdneri  (Katz,
1961; Macek,  et al.  1969); 16  to 36  ug/1 for  fathead minnows,
Pimephales promelas (Henderson, et al. 1959; Tarzwell and  Henderson,
1957) ;  and 8  to 32 pg/1   for  the bluegill,  Lepomis macrochirus
(Henderson, et al.  1959; Macek, et al. 1969; Tarzwell and  Henderson,
1957) .   Freshwater invertebrates  appear  to be more  variable  in
their sensitivity  to  acute dieldrin  toxicity.   The 96-hour  LC
values  range  from  0.5  ug/1 for  the  stone  fly  (Sanders  and Cope,
i:>66) to 740 ug/1 for the crayfish (Sanders, 1372) .
           The acute LCr« values for  eight saltwater fish  species
                       -* u
range from  0..55  to  24.0  ug/1  in  flow-through tests  (Butler, 1963;
Earnest  and  Benville,  1972;  Korn  and  Earnest,   1974; Parrish,
et al.  1973;  Schoettcer, 1970;  and  Lowe,  undated).   LC5Q  values
ranging from  0.7 to 240.0  ug/1  have  oeen  reported for  estuarian
invertebrates  species,   with  the  'most  sensitive  species   tested
being  the  commercially  important  pink  shrimp,  Penaeus duorarum
(U.S. EPA, 1978).
      B.   Chronic Toxicity
           Chronic  toxicity has been  studied  in  two  species of
                                                              »
freshwater  fish.    The  chronic  value  for  steelhead  trout  (Salmo
gairdneri)  from  an embro-larval  study  is  0.11  ug/1   (Chadwick

-------
and  Shumway,  1969).   For  the guppy,  Poecilia  reticulata,  in  a

life-cycle test, the chronic value  is 0.4 pg/1  (Roelofs,  1971).

      C.   Plant Effects


           Freshwater  plants  are less  sensitive to dieldrin  than


freshwater fish  or invertebrates.   For  example,  a concentration


of 100  ^ig/1  caused a  22  percent reduction  in  the  biomass of  the


alga  Scenedlesmus  quadricaudata  (Stadnyk  and  Campbell,   1971) ,

and  12,800 ug/1  reduced  growth by 50 percent in  the diatom,  Navi-

cula  seminulum  after  5  days  of  exposure  (Gairns,  1968).   In  a

saltwater plant  species  growth rate was reduced  at concentrations

of approximately 950 ug/1  (Batterton, et al.  1971).

      D.   Residues


           Bioconcentration  factors  (BCF)  have been determined


for  9  freshwater  species  (U.S.  EPA,  1976).   Representative  BCF


values  are 128  for the alga,  Sc_ence=_smus_ gbl^guus  (Reinert,  1972,

1395 for p_a£hn_i_a mag_na (Reinert,  1972) ,  2385-2993 for  the  channel

catfish,  Ictjilurus pu_nctatus_  (Shannon,  1977a;  i»77b)  and  63,268


for  the yearling  lake  trout,  Salvelinus  namaycush   (Reinert,  et


al.  1974).   The edible  tissue of the Eastern  oyster,  Crassostrea

vjLrgj^nica, had  a  3CF  value of  3000 after  392  days  of  exposure

(Parrish, 1974).   Spot,  Leiostomus  xanthurus,  had  a BCF of  2,300

after 35 days exposure to  dieldrin  (Parrish,  et al. 1973).

VI.  EXISTING GUIDELINES AND STANDARDS

      Neither the  human health  nor the  aquatic criteria  derived

by U.S.  EPA  (1979),  which are summarized below,  have  gone  through
                                                              »
the  process  of  public  review; therefore,  there is a  possibility

that these criteria will be changed.

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      A.    Human
           The  current  exposure  level for  dieldrin set  by OSKA
is an air  time-weighted  average of 250 ug/m  for skin absorption
(37 FR 22139).   In 1959, the  U.S.  Public Health Service Advisory
Committee  recommended  that  the drinking  water  standard for diel-
drin  be  17  ug/1   (Mrak,   1969) .    The  U.N.  Food  and Agricultural
Organization/World Health  Organization's  acceptable  daily  intake
for dieldrin is 0.0001 rag/kg/day  (Mrak, 1969).
           The  carcinogenicity  data  of   Walker,  et   al.   (1972)
were  used  to calculate the draft  ambient water quality criterion
for dieldrin of 4.4  x 10~2  ng/1  (U.S.  EPA,  1979).   This level
keeps the lifetime cancer risk for  humans  below iO'3.
      B.    Aquatic
           The  draft  criterion  to  protect  freshwater   life   L=
0.0019 ug/1  as a  24-hcur average;  the concentration  should  noc
exceed 1.2  ug  at  any time.   To protect  saltwater  acu?.cic life..
she  draft  criterion  is   0.0069  ^ug/1  as   a  24-hour   average;  the
concentration should not exceed 0.16 ug/1  at any  time.

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                          DIELDRIN

                         REFERENCES

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-------
Davis, K.J.  1965.  Pathology report on  mice  for  aldrin,
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Davis, K.J., and O.G. Fitzhugh.   1962.   Tumorigenic  potential
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Deichmann, W.B., et al.  1968.  Retention of dieldrin  in
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Deichmann, W.B., et al.  1970.  Tumor igenicity of aldrin,
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Earnest, R.D., and P.E. Benviile, Jr.  1972.  Acute  toxicity
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Edwards, C.A.  1966.  Insecticide residues  in soils.   Residue
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,
            1972.  Comparative -utagenicity studies with pes-
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Fitzhugh, O.G., et al.  1964.  Chronic oral toxicity of al-
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2: 551.

Harris, R.H., et al.  1977.  Carcinogenic hazards of organic
chemicals in drinking water.  Page 309 in H.H. Hiath,  et al .
eds.  Origins of human cancer.  Cold Sp"rTngs Harbor Lab. Mew
York .

Hathaway, D.E., et al.  1967.  Transport of dieldrin from
mother to blastocyst and from mother to foetus in pregnant
rabbits.  Eur. Jour. Pharmacol.  1: 167.

-------
Hayes, W.J.,  and A.  Curley.   1968.  Storage and excretion  of
dieldrin and  related  compounds:  Effect oif occupational  expo-
sure.  Arch.  Environ.  Health  16:  155.

Heath, D.F.,  and M.  Vandekar.   1964.  Toxicity and metabolism
of dieldrin  in  rats.   Br.  Jour.  Ind. Med.   21: 269.

Henderson, C.,  at  al.   1959.  Relative toxicity of ten chlor-
inated hydrocarbon insecticides  to four species of fish.
Trans. Am. Fish. Soc.   88:  23.

Hunter, C.G., et al.   1969.   Pharmacodynamics of dieldrin
(HEOD).  Arch.  Environ.  Health   18:  12.

Jager, K.W.   1970.   Aldrin,  dieldrin, endrin and telodrin:  An
epidemiological and  toxicological  study of long-term occupa-
tional exposure.   Elsevier Publishing Co., Amsterdam.

Katz, M.  1961.  Acute toxicity  of some organic insecticides
to three species of  salmonids  and  to the threespine sticle-
back.  Trans. Am.  Fish.  Soc.  90:  264.

Korn, 5., and R. Earnest.   1974.   Acute toxicity of twenty
insecticides  to striped  bass,  Morone saxatilis.  Calif.  Fish
Game.  60: 128.

Lowe, J.I.   Results  of toxicity  tes-ts with fishes and macro
i^*r3v*tebrcit':ic.  D^ts  sbeets  —T?.~ "' ^ s'c 13 frcrr. 'T. S. E^v^^cr.
O *~o i~  LI rj a rvo * 7  ^r> TT i y-/^n   Oac  ^".^•^    ("I'iT^ O >- p o -7 a  ^ 1 a

Macek, K.J.,  et al.   1969.   The  effects of temperature on  the
encr*o»^***^^ 1 i*-t»  r\£  V-.1 «•• o *"• ^ T 1 ^  ^r\ A  v- -3 i ^ l-> *•>». T *-•*-.-> i •»*- *-<•> r^o^i'^1-^^
— UOw^.j—'— — -^—-i.~s_)-  ^_ -^  *-/j_— ^-^— — — J  C* * . \_*  i. w — - . *- ^-^ •' w.v^v. w — O— -.^.^.w — ^*
pesticides.   Bull.  Environ.  Contam.  Toxicol.   4: 174.

HacKay, D.,  and A.W.  Wolkoff.   1973.  Rate of evaporation  of
low-solubility  contaminants  from water bodies to atmosphere.
Environ. Sci. Technol.  7:  611.

Majuir.dar, S.K., et al.  1976.   Dieldrir,- induced chromosome
damage in mouse bone-marrow  and  WI-38 human lung cells.
Jour. Hered.  67:  303.

Majumdar, S.K., et al.  1977.   Mutagenicity of dieldrin  in
the Salmonella-microsome test.   Jour. Herd.  68: 194.

Marshall, T.C., et al.  1976.   Screening of pesticides for
mutagenic potential  using  Salmonella typhimurium mutants.
Jour. Agric.  Food  Chem.   24: 560.

Matsumura, F.,  and G.M.  Bousch.   1967.  Dieldrin: Degradation
by soil microorganisms.   Science  156: 959.                •
                               _ A ft —

-------
Matthews, H.B., et al.   1971.   Dieldrin  metabolism,  excre-
tion, and storage in male  and  female  rats.   Jour.  Agric.
Food Chem.  19: 1244.

Mick, D.L., et al.  1971.  Aldrin  and  dieldrin  in  human  blood
components.  Arch. Environ. Health   23:  177.

Mrak, E.M.  1969.  Chairman 1969 report  on  the  secretary's
commission on pesticides and their  relationship to environ-
ment health.  U.S. Dept. Health Edu.  Welfare, Washington,
D.C.

Nash, R.G., and E.A. Woolson.   1967.   Persistence  of chlori-
nated hydrocarbon insecticides  in  soils.  Science   157:   924.

Ottolenghi, A.D., et al.   1974.  Teratogenic effects of  al-
drin, dieldrin and endrin  in hamsters  and mice.  Teratology
9: 11.

Parrish, P.R.  1974.  Arochlor  1254,  DDT and DDD,  and diel-
drin: accumulation and loss by  American  oysters, Crassostrea
virginica exposed continuously  for  56  weeks.  ?roc.  Naci.
Shellfish Assoc.  64.

Parrish, P.R., et al.  1973.   Dieldrin:  Effects  en several
estaurine organisms.  Pages 427-434  in Proc. 27th  Annu.  Conf.
S.E. Assoc. Game Fish Comn.

Patil, K.C., et al.  1972. -Metabolic  trar.sfsrnaticn of  DITT,
dieldrin, aldrin, ar.d endrin by -.r.arine microorganisms.   Envi-
ron. Sc'i. Technol.  5: 631.

Reinert, R.E.  1972.  Accumulation  of  dieldrin  in  an alga
Scenedesrnus obi iqus, Daphnia magna  and the  guppy,  Poecilia
reticulata.  Jour. Fish  Res. Board  Can.  29: 1413.

Reinert, R.E., et al.  1974.  Dieldrin and  DDT:  Accumulation
from water and food by lake trout,  Salvelinus namaycush,  in
the laboratory.  Proc. 17th Conf. Gr^=t  Lakes Res.  52.

Robinson, J., et al.  1969.  The pharmacokinetics  of HEOD
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Roelofs, T.D.  1971.  Effects of dieldrin on the intrinsic
rate of  increase of the  guppy,  Peocilia  reticulata Peters.
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Sanders, H.O.  1972.  Toxicity of some insecticides  to four.
species  of malacostracan crustaceans.  Bur. Sport  Fish. Wild.
Tech. Pap. No. 66.

-------
Sanders, H.O., and O.B. Cope.   1968.   The  relative  toxicities
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Schoettger, R.A.  1970.  Progress  in  sport  fishery  research.
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                                -W-
                                y

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Walker, A.I.T., et al.  1972.  The toxicology of dieldrin
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                                      No. 83
 o,o-Dlethyl Dithiophosphoric Acid


  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.   20460

           APRIL 30, 1980
                            •/

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                          DISCLAIMER
     This report represents a survey  of  the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained  in  the report is drawn chiefly
from secondary  sources  and   available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all  available  information  including  all  the
adverse health  and   environmental  impacts  presented  by  the
subject chemical.  This  document  has  undergone scrutiny  to
ensure its technical accuracy.
                              £3-3-

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                       0,0-DIETHYL DITHIOPHOSPORIC ACID


                                    Summary






     There  is  no available information to  indicate  that o,o-diethyl dithio-



phosphoric  acid produces  carcinogenic,  mutagenic,  teratogenic,   or  adverse



reproductive effects.



     A  possible metabolite  of  the  compound,  o,o-diethyl  dithiophosphoric


acid, did  not   show  mutagenic  activity in  Drosophila,  E.  coli,   or  Saccha-



romyces.



     The pesticide  phorate,  which  may release o,o-diethyl  dithiophosphoric
                                                    «,

acid as a metabolite, has shown some  teratogenic  effects in  developing chick



embryos and adverse reproductive effects in mice.



     An acute value of 47.2 /jg/1 has  been  reported for rainbow  trout  exposed



to  a diethyl   dithiophosphoric  acid  analogue,   dioxathion.   A  synergistic



toxic effect with the latter chemical and  malathicn is suaaested.

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 I.    INTRODUCTION
      o,o-0iethyl hydrogen  dithiophosphate,  CAS registry number 298-06-6,  al-
 so called  o,o-diethyl  phosphorodithioic acid or o,o-diethyl dithiophosphoric
 acid,  is used  primarily  as an intermediate in the synthesis of several pest-
 icides:  azinphcsmethyl,  carbcphenothion, dialifor,  dioxathion,  disulfoton,
 ethion, phorate, phosalone and terbufos.  It  is  made from phosphorus penta-
 sulfide (SRI,  1976).
 II.  EXPOSURE
     A.   Water                                     v
          Pertinent  data were  not found in the available  literature;  how-
 ever,  if found in  water, its  presence would most likely be  due to microbial
 action on phorate  or disulfoton (Daughton,  et al. 1575),  or as a contaminant
 of any of the above pesticides  for which it is a starting compound.
     8.   Food
          Pertinent  data were  net found in  t.u;2 available  literature;  how-
 ever,  if  present  in food,  the compound would  probably  originate  from  the
 same  sources discussed above.   Organcphosphorus  pesticide  residues have been
 found in feed  (Vettorazzi,  1975).
     C.   Inhalation
          Pertinent  data  were  not found in  the available  literature;  how-
ever,  major  exposure could  come  from   fugitive  emissions  in  manufacturing
 facilities.
     0.   Dermal
          Pertinent data were not found in the available literature.

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III. PHARMACOKINETICS
     A.   Absorption
          Information  relating  specifically to the absorption of o,o-diethyl
dithophosphoic  acid  was  not found in  the  available  literature.   Acute toxi-
city studies  with the pesticides disulfoton and  phorate indicate that these
related  organophosophorous  compounds are  absorbed following oral  or dermal
administration  (Gaines, 1969).
     B.   Distribution
          Pertinent  data  were not  found  in the  available  literature.   Oral
administration  of labelled  phorate, the  S-(ethyl thio)methyl derivative  of
o,o-diethyl  dithiophosphoric acid,  to cows  accumulated  in liver,  kidney,
lung,  alimentary  tract,  and glandular  tissues;  fat  samples showed  very  low
residues (Bowman and Casida, 1553).
     C.   Metabolism
          Pertinent data were net  found in the available literature.  Metab-
olism  studies  with disulfoton (3ull,  1965)  and phorate  (Sc.vman  and Casida,
1958}  indicate  that  both compounds  are converted  to  diethyl phosphorodithic-
ate, diethyl phorphorcthicate, and diethyl phosphate.
     0.   Excretion
          Pertinent data were  net  round in the available  literature,   c^sed
on animal studies  with related  organophosphorous  compounds, the  parent  com-
pound and its  oxidative  metabolites may be expected to  eliminated  primarily
in the urine (Matsumura,  1975).
IV.  EFFECTS
                                                       .-
     A.   Carcinogenesis
          The  dioxane  s-s  diester  with  o,o-diethyl dithiophosphoric  acid,
dioxathion,   has  been  tested  for  carcinogenicity   in   mice  and   rats   by

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long-term  feeding.  No  carcinogenic effects  were  noted in  either  species
(NCI,  1978).
     Q.   Mutagenicity
          Diethyl  phosphorothioate,  a possible metabolite of the parent com-
pound,  did  not show mutagenic  activity in  Orosophila,  £.  coli,  or  Saccha-
romyces (Fahrig, 1974).
     C.   Teratogenicity
          Pertinent  data were  not found  in the available  literature.  In-
jection  of  phorate into  developing  chick  embryos,  has been reported  to
produce malformations (Richert and Prahlad, 1972).
     0.   Other Reproductive Effects
          Pertinent  data were  not  found  in  the available  literature.   An
oral feeding study conducted in mice with  pnorate (0.6 to 3.0 ppm) indicated
that the  hicrest  level  of  compound  did  produce sc^e  adverse rsprcd'jctive
effects (American  Cyanaaiid,  1966}.  Chronic  feeding of mics  with technical
dicxatnion  at  levels of 430 to  6CG ppm  produced  seme  testiscular atrophy
(NCI, 1978).
     E.   Chronic Toxicity
          Chronic  feeding  of  technical  dioxathicn  produced  hyperplastic
nodules "'^  ^iv°rs  of  ^s1?  mice,   o c-Oisthyl  dit~iocncsrhc!ric scid   like
other  organophosphates,  is  expected  to  produce cholinesterase  inhibition
(MAS, 1977).
V.   AQUATIC TOXICITY
     A.   Acute
          Marking  (1977)  reports  on  LC5Q  value of 47>2 jug/i f0r  rainbow
trout   (Salmo   gairdneri)   exposed   to   the   dithiodioxane   analogue •  of
bis(o,o-diethyl  dithiophosphoric  acid),  dioxathion,  and  an  LC5Q  value  of

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3.44 pg/1  when  this  latter  compound  is  applied  in  combination with  mal-
athion.  The synergistic action with malathion  suggests  that the combination
is more than eight times as toxic as either of the individual chemicals.
     B.   Chronic, Plant Effects,  and Residues.
          Pertinent data were not found in the available literature.
VI.  EXISTING GUIDELINES
     Existing guidelines or  standards  were not  found  in the available  lit-
erature.

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                                  REFERENCES
American  Cyanamid   1966.   Toxicity  data  on  15 percent  Thimet  granules.
Unpublished  report.   In:   Initial  Scientific  and  Minieconomic Review  of
Phorate (Thimet) Office of Pesticide Programs, Washington.

Bowman, J. and  J. Casida   1958.   Further studies on the metabolism of Thimet
by plants, insects,  and mammals.  J. Econ. Entomol.  51: 338.

Bull, 0.  1965.   Metabolism of  di-systox by insects,  isolated cotton leaves,
and rats.  J. Econ.  Entcmol.  58: 249.

Daughton, C.G.,  A.M.  Cook, M.  Alexander  1979.   Phosphate and  soil binding
factors   limiting  bacterial   degradation  of   ionic   phosphorus-containing
pesticide metabolites.  App. Environ. Micrcbio.  37: 605.
                                                  «.
Fahrig,  R.    1974.    Comparative  mutagenicity  studies   with   pesticides.
Chemical Carcinogsnesis Assays,  IARC Scientific Publication #10,  p. 161.

Gaines, T.   1969.   Acute  toxicity of pesticides.  Toxicol. A.ppl. Pharmacol.
14: 515.

Marking,  L.I.   1977.   Method for asssessing  additive toxicity  of  chemical
mixtures.   In:    Aquatic  Toxicology and Hazard  Evaluation.   STP 634  ASTM
Special Technical Publication.  p. 99.

Matsumura, F.   1975.   Toxicolocy of Insecticides.  Mew  York:  Plenum Press,
o. 223.      '                  "                                     •   •

National  Academy  of  Sciences  1977.  Drinking  Water and  Health.   National
Sesearcn Council, wasnington,  p. 615.

National  Cancer  Institute   1973.    3icassay  of  Oioxathion  for  Possible
Carcinogenicity.  U.S.  OHEW,   NCI  Carcinogenesis  Technical  Report  Series
0125, 44 pp.

Richert, E.  and '<.  Prahlad  1972.  Effect of the organophcsohate o,o-diethvl
s-t(ecnylthio)metnyij pnospnoroaitnioate on tne cnick.   Pouit.  Sci.  51: 513.

SRI   1976.   Chemical  Economics  Handbook.   Stanford  Research  Institute.
Pesticides, July 1976.

vettorazzi,  G.    1975.   State  of  the  art  on  the  tcxicolcgical  evaluation
carried  out  by  the joint   FAO/WHO  meeting  on  pesticide  residues.   II.
Carbamate  and  organophosphorus  pesticides  used in  agriculture and  public
health.  Res. Rev.  63:  1.
                                     PT(9—
                                  -  /  /  v

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                                        No. 84
            3
o,o-Diethy1-4-methyl Phosphorodithioate
    Health and Environmental Effects
  U.S. ENVIRONMENTAL PROTECTION AGENCY

         WASHINGTON, D.C.  20460




             APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documen-ts.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

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                    o,o-OIETHYL-S-METHYL PHOSPHORODITHIOATE
                                    Summary

     There  is  no available  information on the  possible  carcinogenic,  muta-
genic,  teratogenic  or adverse  reproductive effects  of  o,o-diethyl-S-methyl
phosphorodithioate.    Pesticides  containing   the   o,o-diethyl   phosphoro-
dithioate moiety  did not show  carcinogenic effects  in  rodents  (dioxathion)
or teratogenic  effects in chick embryos  (phorate).   The possible metabolite
of  this  compound,   o,o-diethyl phosphorothioate,  did  not  show  mutagenic
activity  in Drosophila,  E.  coli,  or Saccharomyces.   o,o-Oiethyl-S-methyl
phosphorodithioate,  like other organophosphate  compounds,   is  expected  to
produce chclinesterase inhibition in humans.
     There is no available data on the aquatic toxicity of this compound.

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                    0,0-0IETHYL-S-METHYL  PHOSPHORODITHIOATE
 I.   INTRODUCTION
      o,o-Oiethyl-S-methyl .phosphorodithioate (CAS  registry number 3288-58-2)
 is described in  German  patents 1,768,141  (CA  77:151461s) and  1,233,390 (CA
 66:115324p).    The  latter  states  the  compound  has  "partly  insecticidal,
 acaricidal and  fungicidal  activity"  and is useful as  an intermediate  for
 organic synthesis.  It has  the following  physical  and chemical properties:

                    Formula:               C5H13
                    Molecular Weight:      200
                    Boiling Point:         ICCOQ  to 102°C  (4 torr)
                    (CA  55:8335h)
                    Density:               1.192420
                    (CA  55:S335n)
      Pertinent  data were not  found in  the available literature with  respect
 tc production, consumption  or the current  use of this ccmcound.
 II.   EXPOSURE
      Pertinent data were not  found in the  available  literature.
 III.  PHARMACGKINETICS
      A.   Absorption
          Information  relating  specifically to the  absorption  of c,o-di-
ethyl-S-methyl  phosphcrcdithioats  was  not  found  in  the available  liter-
ature.  Oral  administration of the S-ethylthio  derivative of  this ccmpcund,
the insecticide phorate, indicates that this derivative is absorbed from the
gastrointestinal tract (Bowman and Casida, 1958).
                                                     «•
     3.   Distribution
          Pertinent  data   were  not   found   in   the  available  literature.
Studies with  22p  radiolabelled phorate in the  cow  indicated  that following
oral  administration,   residues  were   found  in   the  liver,   kidney,   lung,

-------
 alimentary   tract,   and   glandular   tissues;   fat   samples   showed   very   low
 residues  (Bowman  and Casida,  1958).
      C.   Metabolism
          Pertinent  data were not found  in the available literature.  Based
 on  metabolism  studies with various  crganophosphatss in mammals,  o,o-diethyl-
 S-methyl  phosphorodithioate may  be  expected to undergo hydrolysis to diethyl
 phosphorodithioic acid,  diethyl  phosphorothioic acid,  and diethyl phosphoric
 acid  (Matsumura,  1975).
      0.   Excretion
          Pertinent   data  were   not  found  in  the  available  literature.
 Related  metabolites   (o,o-diethyl  phosphorodithioic,  phosphorothioic,  and
 phosphoric  acids) have  been  identified  in  the urine  of   rats  fed phcrate
 (Bowman and Casida,  1958).
 IV.   EFFECTS
      A.   Carci.noqenicity
          Pertinent  data were not  found in  the available  literature.   The
dioxane-S-S-diester  with  o,o-diethyl  phosphorodithioate,   dioxathion,  has
been  tested  for carcinogenicity in  mice  and rats  by  long-term feeding.   No
carcinogenic effects wers noted in either species (NCI, 1978).
     3.   Mutagenicity
          Pertinent  data   were  not   found  in  the  available  literature.
Diethyl phosphorothioate, a  possible  metabolite of the parent  compound,  did
not show mutagenic activity  in Orosphila, E_.  coli,  or  Saccharomyces  (Fahrig,
1974).

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     C.   Teratogenicity
          Pertinent  data were  not  found  in the  available literature.   In-  •
jection of  phorate into developing  chick  embryos has  been reported to  pro-
duce malformations (Richert and Prahlad, 1972).
     0.   Other Reproductive Effects
          Pertinent  data were  not  found  in  the available- literature.   An
oral feeding study conducted in mice with  phorate (0.6 to  3.0 ppm)  indicated
that the highest  level  of  compound did produce some adverse reproductive ef-
fects  (American Cyanamid,  1966).   Chronic  feeding of rats  with  technical
dioxathion  at  levels from  450 to  600 ppm  produced  some  testicular atrophy
(NCI, 1978).
     E.   Chronic Toxicity
          Pertinent  data   were   not  found  in  the  available   literature.
Chronic feeding of technical dioxathion  produced hyperplastic  nodules in  the
livers  of   ,7,aie  mice.    o,o-Oietnyi-S-methyI onospnoroaichioats,  like  other
organophosphates,  is expected  to  produce  choiinesterase  inhibition  (NAS,
1977).
V.   AOJATIC TOXICITY
     Pertinent data were not found in the available literature.
VI.  EXISTING GUIDELINES AND STANCAF.CS
     Existing  guidelines  and  standards  were  not  found  in  the available
literature.

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                    0,0-OIETHYL-S-METHYL PHOSPHORODITHIOATE

                                  References
American Cyanamid.  1966.   Toxicity  data  on 15 percent Thimet granules.  Un-
published report.  In:  Initial Scientific and Minieconomic Review of
Phorate (Thimet) Washington, DC:  Office of Pesticide Programs.

Bowman,  J.  and  J. Casida.   1958.   Further  studies  on  the metabolism  of
Thimet by plants, insects, and mammals.  Jour. Econ. Entom.  51: 838.

Fahrig, R.  1974.   Comparative  mutagenicity studies with  pesticides.   Chem-
ical Carcinogenesis Assays, IARC Scientific. Publication No. 10.  p. 161.  •

Matsumura, F.   1975.   Toxicology of  Insecticides.   Plenum Press,  New York
p. 223.

National Academy  of Sciences.   1977.   Drinking Water and  Health.   National
Research Council, Washington, DC.  p. 615.

National Cancer  Institute.   1578.   3ioassay of Dioxathion  for Possible Car-
cinogenicity.    QHEW.   National  Cancer Institute.   Carcinogenesis  Tecnnical
Report Series No. 125: 44.

Richert,  E.P.  and  K.V.   Prahlad.    1972.   Effect  cf  the  crganophcspnate
o,o-dirthyl-5-C(sthyithio)methyi]  phospnorocithiste on  the  chick.   Fault.
Sci. 31: 513.

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                                      No. 85
         Dlethyl Phchalate


  Health and Eavironnental Effaces
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.   20460

           APRIL 30,  1980

-------
                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is  drawn chiefly
from secondary  sources  ana  available  reference  documents.
Because of the limitations of such sources, this  short profile
may not  reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by the
subject chemical.   This  document  has undergone  scrutiny  to
ensure its technical accuracy.
                          _ L^^"r .
                          ') I/ U ) ""

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                      DIET3YL PHTHALATE
                           SUMMARY"
     Diethyl  phthalate  has  been shown  to produce  rautagenic
effects in the Ames Salmonella assay.
     Teratogenic  effects  were  reported foJ .owing  i.p.  admin-
istration of  diethyl  phthalate  to pregnan  rats.   This  same
study has also indicated  fetal toxicity aa  increased  resorp-
tions after i.p. administration of DEP.
     Evidence  that diethyl  phthalate  prc .uces carcinogenic
effects has not been found.
     A single clinical  report  indicates t. at the  development
of hepatitis  in  several  hemodialysis  pati nts  may have  been
related  to  leaching  of  diethyl  phthalate  from  the  plastic
tubings utilized.
     Diethyl  phthalate  appears  to  be mor   toxic  for  marine
species acutely  tested,  with a  concentra  :on  of  7,590  ug./l
being  reported  as  the  LC-n  in  marine i  vertebrates.    The
data base  for  the  toxic  effects of  die  lyl  phthalates  to
aquatic  organisms  is   insufficient  to  d  ift  criterion   for
their protection.
                                 * "/!?l/ Cr"

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                      DIETHYL PHTHALATE

I.   INTRODUCTION

     This  profile  is based  on  the  Ambient  Water  Quality

Criteria Document for Phthalate Esters  (U.S.  EPA,  1979a).

     Diethyl phthalate  (DEP)  is a  diester  of the ortho  form

of benzene dicarboxylic  acid.   The  compound  has a molecular

weight  of  222.23,  specific  gravity of  1.123,  boiling point

of 296.1°C", and is  insoluble in water  (U.S. EPA,  1979a) .

     DEP is  used  as a plasticizer  for cellulose ester plas-

tics and as a carrier for perfumes.

     The 1977  current production  of  diethyi phthalate  was:

3.75 x 10  tons/year  (U.S.  EPA, 1979a).

     Phthalates have  been detected  in soil,  air,  and water

samples; in animal  and human  tissues;  and in certain vegeta-

tion.   Evidence  from in  vitro  studies indicate  that certain

bacterial  flora  may  be   capable  of  metabolizing phthalates

to the monoester form (Engelhardt, et al. 1975).

II.  EXPOSURE

     Phthalate  esters  appear  in  all  areas  of  the  environ-

ment.   Environmental  release of  the  phthalates may  occur

through leaching  of plasticizers  from  plastics, volatiliza-

tion of  phthalates  from  plastics,  and  the  incineration of

plastic items.   Human exposure to phthalates includes contami-

nated  foods  and  fish, dermal  application -'in cosmetics,  and

parenteral  administration  by  use  of  plastic   blood  bags,
                                                          »
tubings, and  infusion devices  (mainly  DEHP  release)  (U.S.

EPA, 1979a).

-------
     Monitoring studies have indicated that most water phthal-

ate concentrations  are in the  ppra range,  or  1-2 pg/1  (U.S.

EPA,  1979a).    Industrial air  monitoring  studies  have  mea-

sured air levels of  phthalates  from 1.7 to 66 mg/m   (Milkov,

ec  al.  1975) .   Information  on  levels of  DEP  in  foods  is

not available. The-U.S. EPA (1979a) has estimated the  weighted

average bioconcentration  factor  for DEP  to  be  270  for  the

edible portions of  fish and  shellfish consumed  by Americans.

This  estimate is  based  on measured  steady-state bioconcen-

tration studies in bluegills.

III. PKARMACOKINETICS

     Specific  information is  not  available  on  the   absorp-

tion, metabolism,  distribution,  or excretion  of DEP.    The

reader is  referred to a general coverage of onthalate  metabo-

lism in the phthalate ester hazard  profile  {"J.5.  SPA,  137Sb) .

IV.  EFFECTS

     A.    Carcinocenicity

          Pertinent  information  could  not  be  located   in.

the available literature.

     3.    Mutagenici*:y

          Diethyl phthalate  has been  shown  to  produce muta-

genic effects  in the  Ames  Salmonella  assay  (Rubin,  et  al.

1979) .

     C.    Teratogenicity

          Administration  of   DEP   to  pregnant  rats   by   i.p.
                                                          »
injection  has  been  reported  to produce  teratogenic   effects

(Singh,  et al. 1972).
A
                                    -i /> / * -
                                     u i J

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     D.   Other Reproductive Effects



          Fetal   toxicity   and  increased  resorptions   were



produced  following  i.p.  injection  of  pregnant   rats   with



DEP  (Singh, et al. 1972).



     E.   Chronic Toxicity



          A  single  clinical   report  has  been  cited  by  the



U.S. EPA  (1979a)  which  correlated  leaching of DEP  from  hemo-



dialysis  tubing  in  several  patients with  hepatitis.    Char-



acterization  of  all  compounds  present  in  the hemodialysis



fluids was not done.



V.   AQUATIC TOXICITY



     A.   Acute Toxicity



          Among  aquatic  organisms,  the  bluegill sunfish,  ;



Lepomis macroc'nirus,  has beer, shown  to be acutely sensitive



to  diethyl  phthalate;  a 96-hour static  LCCQ of  98,200  pg/i



is  reported  (U.S. EPA.  1978).   For  the freshwater inverte-



brate, Daphnia  magna,  a 48-hour static  LC,-Q of  51,100  pg/1



was obtained.  Marine organisms  proved  to be more  sensitive,



with  the  sheepshead  minnow,  Cyprinodon  variegatus,   showing



a 96-hour static LC5Q of 29,600  jjg/1, while the mysid  shrimp,



Mysidopsis  bahia,  showed  an  96-hour  static  LC   of  7,590



ug/1 (U.S. EPA, 1978).



     B.   Chronic Toxicity



          Pertinent   information  could'  not  be  located   in



the available literature.



     C.   Plant Effects



          Effective  concentrations  based  on  chlorophyl  a



content  and  cell  number  for  the  freshwater  alga,   Selena-

-------
strum  capcicornutum,  ranged  from  35,600  to  90,300  ug/1,
while the marine  alga,  Skeletonema costatum, was more  sensi-
tive,  with  effective  concentrations  ranging   from   65,500
to 35,000 ug/1.
     D.   Residues
          A  bioconcentration  of  117  was  obtained  for  the
freshwater invertebrate, Daphnia magna.
VI.  EXISTING GUIDELINES AND STANDARDS
     Neither the  human health  nor the aquatic  criteria  de-
rived by  U.S.  EPA  (1979a) , which  are  summarized below,  have
gone  through  the  process  of  review;  therefore,   there  is
a possibility that  these criteria  will be changed.
     A.   Human
          Based  on  "no  effect"  levels  observed  in chror.ic
feeding studies  with rats or  dogs,  the U.S.  E?A has  calcu-
lated an  acceotable daily  intake  (ADI)  level of 438  me/day
for DEP.
          The  recommended  water   quality  criterion   level
for  protection  of  human health   is 50  mg/1  for  DE?   (U.S.
EPA, 1979a).
     B.   Aquatic
          Data are insufficient  to  draft criterion  for  the
protection of  either  freshwater  or  marine  organisms   (U.S.
EPA, 1979a).

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                      DIETHYL PHTHALATES

                          REFERENCES

Engelhardt,  G.  et  al.    1975.   The  raicrobial  metabolism of
di-n-butyl phthalate  and  related dialkyl  phthalates.   Bull.
Environ. Contain. Toxicol.   13: 32.

Milkov, L.E.,  et al.   1975.  Health status of  workers  ex-
posed to phthalate plasticizers in  the manufacture of artifi-
cial leather and  films  based on  PVC resins.   Environ. Health
Perspect. Jan. 1975.

Rubin, R.J., et  al.   1979.   Ames mutagenic assay of a series
of phthalic  acid  esters:   Positive  response  of  the dimethyl
and diethyl  esters  in TA  100.  Abstract. Soc. Toxicol. Annu.
Meet.  March 11, 1979, New Orleans.

Singh, A.  et al.   1972.   Teratogenicity of  phthalate esters
in rats.  Jour. Pharm. Sin. Gl, 51.

U.S. EPA.    1978.   In-depth  studies on  health  and environ-
mental impacts  of  selected water pollutants.   U.S. Environ.
Prot.  Agency, Contract No. 68-01-4646.

U.S. EPA.   1979a.   Phthalate Esters:   Ambient  Water Quality
Criteria (Draft).

U.S.  EPA.    I979b.   Environmental Criteria and  Assessment
Office.  Hazard Profile:  Phthalate Esters (Draft).

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                                      No. 36
        DimethyInitrosamine


  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents a survey of  the  potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained  in the report is drawn chiefly
from secondary  sources  and   available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all  available  information including  all  the
adverse health  and  environmental  impacts presented  by  the
subject chemical.  This  document has  undergone scrutiny  to
ensure its technical  accuracy.
                            ¥<-

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                     DIMETHYLNITROSAMINE
                           SUMMARY
     Dimethylnitrosamine produces liver and kidney  tumors
in rats.  It is mutagenic in several assay systems.  No
information specifically dealing with the teratogenicity,
chronic toxicity or other standard toxicity tests of dimethyl-
nitrosamine was available for review.
     Hepatocellular carcinoma has been induced  in rainbow
trout administered 200 to 800 fig dimethylnitrosamine in
their diet.
                                  -j &/ 3 "

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                      DIMETHYLNITROSAMINE


I.   INTRODUCTION


     This profile  is  based on the Ambient Water Quality


Criteria Document  for Nitrosamines (U.S. EPA, 1979a).


     Specific  information on the properties, production,


and use of dimethylnitrosamine was not available.  For general


information on dimethylnitrosamine, refer to the ECAO/EPA


Hazard Profile for  Nitrosamines (U.S. EPA, 1979b).


     Dimethylnitrosamine  can exist for extended periods


of time in the aquatic  environment (Tate and Alexander,


1975; Fine, et al.,  1977a).


II.  EXPOSURE


     A.   Water


          Dimethylnitrosamine has beer, detected at a concen-


 tration cf 2  to  4  ug/1  ir. wastawatar samples from waste


 treatment plants adjacent to, or receiving effluent from,


 industries using nitrosamines or secondary amines in produc-


 tion operations  (Fine,  et al., 1977b).


      3.    Feed


           Dimethyinitrosamine was found to oe present in


Na variety of foods (including smoked, dried or salted fish,
   ^

 chees£v<3alaiai,  frankfurters, and cured meats)  in the 1


 to 10  u/kg "tfange and occasionally at levels up to 100  ug/kg


 (Montesano anc>Bartsch, 1976).


          The  U.\S.  SPA  (1979a)  has estimated  the  weighted


 average bioconce.ntrati°r;  factor for dimethylnitrosamine


 for  the edible pc?rtions of  fish and shellfish  consumed  by

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Americans to be 0.06.   This  estimate is based on the n-octanol/
water 'partition coefficient  of dimethylnitrosamine.
     C.   Inhalation
          Dimethylnitrosamine  has been detected in ambient
air samples collected  near  two chemical plants, one using
the amine as a raw material  and the other discharging it
as an unwanted byproduct  (Fine, et al., 1977a).
          Tobacco smoke contains dimethylnitrosamine.  The
intake of dimethylnitrosamine  from smoking 20 cigarettes
per day has been estimated at  approximately 2 ug/day (U.S.
EPA, 1979a).
III. JHA3MACCKINETICS
     A.   Absorption
          Pertinent data could  not  be  located in  the  avail-
able literature.
     3.   Distribution
          Following intravenous  injection  into rats, dimetnyi-
nitrosamine is rapidly  and rather uniformly distributed
throughout the body (Mages,  1972).
     C.   i-ietaboiism and Excretion
          In vitro studies have demonstrated that  the  organs
in the rat with the major capacity  for metabolism  ct dimethyl-
nitrosamine are the liver and kidney  (Montesanc and  Magee,
1974).  After administration of  14C-labeled-- "imethylnitro-
samine to rats oc mice, about 60  percent o: fc^e  isotope
appears as I4C02 within 12 hours, while 4  tarcent  is axcretaa"
                                  .„ / a ^*"

-------
in the urine  (Magee, et al., 1976).  Dimethylnitrosamine

is excreted in the milk of female rats  (Schoental, et al.,

1974}.

IV.  EFFECTS

     A.   Carcinogenicity

          Chronic feeding of dimethylnitrosamine at doses

of 50 mg/kg induces liver tumors in rats  (Magee and Barnes,

1956; Rajewski, et al., 1966).  Shorter, more acute expo-

sures to dimethylnitrosamine ranging from 100 to 200 mg/kg

produce kidney tumors in rats and liver tumors in hamsters

(Magee and Barnes, 1959; Tomatis and Cafis, 1967).  A single

unspecified intraperitoneal dose given  to newborn mice in-

duced hepatocellular carcinomas (Toth,  et al., 1964).

     3.   Mutacenicity

          Dimethylnitrosamine and diethylnitrosamine have

been reported to induce forward and reverse mutations in

3_. typhimurium, E. coli, iMeurospora crassa and other organisms;

gene recombination and conversion in Saccharomyces cerevisiae;

"recessive lethal mutation" in Drosophila; and chromosome

aoerracions in mammalian cells (Montesano and Bartsch, 1976).

Nitrosamines must be metabolically activated to be mutagenic

in microbial assays (U.S. EPA, 1979a).  Negative results

were obtained in the mouse dominant lethal test (U.S. EPA,

1979a).

     C.   Teratogenicity and Other Reproductive Effects
                                                           *
          Pertinent information could not be located in

the available literature on the teratogenicity and other

reproductive effects of dimethylnitrosamine.

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     D.   Chronic Toxicity

          Pertinent information could not be  located  in

the available literature on the chronic activity of dimethyl-

nitrosamines.

     E.   Other Relevant Information

          Aminoacetonitrile, which inhibits the metabolism

of dimethylnitrosamine, prevented the toxic and carcinogenic

effects of dimethylnitrosamine in rat livers  (Magee,  et

al., 1976).

          Ferric oxide, cigarette smoke, volatile acids,

aldehydes, methyl nitrite, and benzo(a)pyrene have been

suggested to act in a cocarcinogenic manner with dimechyl-

nitro-samine (Stenback, et al., 1973; Magee,  et al.,  1976).

V.   AQUATIC TOXICITY

     Pertinent information about acute and chronic aquatic

toxicity was not found in the available literature.   Addition-

ally, -c mention was made in any reports ariout plane  effects

or residues.

     One study reported that Shasta strain rainbow trout

(Salmo gaircneri), fed dimethylnitrosamine in their diet

for 52 weeks, developed a dose-response incidence of  hepato-

cellular carcinoma during a range of exposures from 200

to 300 mg dimethylnitrosamine per kg body weight 52 to 78

weeks after dosing (Grieco, 1978).

VI.  EXISTING GUIDELINES AND STANDARDS
                                                          »
     Neither the human health nor aquatic criteria derived

by U.S. EPA  (1979a),  which are summarized below, have gone


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through the process of public review; therefore, there is
a possibility that these criteria may be changed.
     A.   Human
          The U.S. EPA (1979a) has estimated that the water
concentrations of dimethylnitrosamine corresponding to life-
time cancer risks for humans of 10~  , 10  , or 10   are
0.026 ug/1, 0.0026 ug/1,  and 0.00026 pg/1, respectively.
     3.   Aquatic
          Data are insufficient to draft freshwater marine
criteria for dimethylnitrosamine.
                             /

-------
                      DIMETHYLNITROSAMINE

                          REFERENCES

Fine, D.H., et al.   1977a.   Human  exposure  to N-nitroso com-
pounds  in  the environment.   In:  H.H.  Hiatt,  et  al.,  eds.
Origins of human cancer.  CoTd Spring Harbor Lab.,  Cold
Spring  Harbor, New York.

Fine, D.H., et al.   1977b.   Determination of dimethylnitrosa-
mine  in air, water and soil  by thermal energy analysis:  mea-
surements  in Baltimore, Md.   Environ.  Sci.  Technol.   11:
581.

Grieco, M.P., et al.  1978.   Carcinogenicity and  acute toxic-
ity of dimethylnitrosamine  in rainbow trout  (Salmo  gaird-
neri).  Jour. Natl.  Cancer Inst.   60:  1127.

Magee,  P.N.  1972.   Possible  mechanisms  of  carcinogenesis and
mutagenesis by nitrosamines.  Inr  W.  Nakahara,  et al., eds.
Topics  in chemical carcinogenesTs.  University  of Tokyo
Press, Tokyo.

Magee, P.N., and J.M. Barnes.  t956.   The producrion  of ma-
lignant primary hepatic tumors in  the  rat by feeding  dimethyl-
nitrosamine.  Br. Jour. Cancer   10: 114.

Magee, P.N., and J.M. Barnes.  1959.   The experimental pro-
duction of tumors in  the  rat  by  dirnechylnitrosamine  (N-nitro-
sodimethyiamine).  Acta.  CJn.  Int.  Cancer  15: 137.

Magee, P.N1., ac al.   1976.  N-Nitroso  compounds and  related
carcinogens.  In; C.S. Searie, ed.  Chemical Carcir.ccens.
ACS Monograph No. 173.  Am. Chem.  Soc.,  Washington,  D.C.

Mcntesano, R., and H. Sartsch.   1976.  Mutagenic and  carcino-
genic N-nitroso compounds: possible environmental hazards.
Mutat. Res.  32: 179.

Montesano, R., and P.N. Magee.   1974.  Comparative metacoiism
_in vitro of nitrosamines  in various animal species  including
man.  In: R. Montesano, et al. ,  eds.   Chemical  carcinogenesis
essays.  IARC Sci. Pub. No. 10.  Int..  Agency  Res. Cancer,
Lyon, France.

Rajewsky, M.F., et al.  1966.  Liver  carcinogenesis  by di-
ethylnitrosamine in  the rat.  Science  152c  83.

Schcental, R., et al.  1974.  Carcinogens in  milk: transfer
of ingested diethylnitrosamine into milk lactating rats.  Br.
Jour. Cancer  30: 238.

-------
Stenback,, P., et al.  1973.  Synergistic effect of ferric
oxide on dimethylnitrosamine carcinogenesis in the Syrian
golden hamster.  Z. Krebsforsch.  79: 31.

Tate, R.L., and M. Alexander.  1976.  Resistance of
nitrosamines to microbial attack.  Jour. Environ. Qual.  5:
131.

Tomatis, L., and F. Cefis.  1967.  The effects of multiple
and single administration of dimethylnitrosamine to hamsters,
Tumori  53: 447.

Toth, B.f et al.  1964.  Carcinogenesis study with dimethyl-
nitrosamine administered orally to adult and subcutaneously
to newborn BALBC mice.  Cancer Res.  24: 2712.

U.S. EPA.  1979a.  Nitrosamines: Ambient Water Quality Cri-
teria. (Draft).

U.S. EPA.  1979b.  Environmental Criteria and Assessment Of-
fice.  Nitrosamines: Hazard Profile.
                                 -SO

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                                    No. 37
       2,4-Dlmethylphenol


 Health and Environmental iffacts
U.S.  ENVIRONMENTAL PROTECTION AGENCY
      WASHINGTON, D.C.   20460

          APRIL 30, 1980
            -103.1-

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
niay not  reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny  to
ensure its technical accuracy.

-------
                              2.4-OIMETtiYLPhENOL
                                    Summary
     2,4-Oimethylphenol  (2,4-OMP)  is  an  intermediate in  a number of  indus-
trial and  agricultural  products.  The  main route of  exposure for humans  is
dermal with 2,4-OMP being readily absorbed  through the skin.
     Little data is available on the mammalian effects of  2,4-OMP.   Tests  on
mice conclude  that  the  compound may  be a promoting agent  in  carcinogenesis.
2,4-OMP  inhibits  vasoconstriction in isolated rat  lungs; this  ability may
cause adverse health effects in chronically exposed humans.
     A  reported 96-hour  LC-j-,  value  for  fathead  minnows  is  16,750 jug/1;
chronic value using embryo-larval  stages of the  same  species  is 1,100 ug/1.
Oacnnia  magna  has  an   observed  48-hour  LC-Q  value  of 2,120 ,ug/l.    In
limited  testing,  one  aquatic  alga  ana  auckweed are over  100  times less
                                                                              •«
sensitive  than  the  Caonnia in acute  exposures.   The  bioconcsntration  factor
for  2,4-  diiTietnyiphehoi is  150 for  the  bluegili.   rrom  half-life  stuoies,
residues of the chemical are noc a potential hazard for aquatic species.

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I.   INTRODUCTION


     This profile  is based primarily  on the Ambient  Water Quality Criteria


Document for 2,4-Oimethylphenol (U.S. EPA, 1979).


     2,4-Oimethylphenol  (2,MM5)  is  derived from coal and petroleum sources


and  occurs  naturally  in  some  plants.   2,4-DMP  (C-H,gO)  is  usually  found


with the five other  dimethylphenol  and  three methylphenol isomers.  It has a


molecular weight  of 122.17  and normally  exists as a colorless  crystalline


solid.   2,4-OMP has a  melting point  of  27 to  28°C,   a  boiling  point  of


210°C  (at  760 mm  Hg),  a vapor pressure  of 1 mm  Hg at  52.8°Cr  and a dens-


ity of 0.0965 g/ml at 20°C (U.S. EPA, 1979).


     2,4-QMP  is a  weak  acid   (pk_-10.6)  and is  soluble in  alkaline  solu-


tions.  It  readily  dissolves  in organic -solvents  and  is  slightly soluble in


water (Weast, 1976).                                                          •


     2,4-OM?  is  a chemical  intermediate in  the manufacture of  a number of


industrial  and  agricultural  products, including  phenolic antioxidants,  dis-


infectants,   solvents,   Pharmaceuticals,  insecticides,   fungicides,  plasti-


cizers,  rubber chemicals,  polyphenylene  oxide,  wetting  agents,  and  dye-


stuffs.  It  is  also found in lubricants,  gasolines,  and  cresylic acid (U.S.


EPA, 1979).


     Very little information exists on  the environmental  persistence of 2,4-


DMP.  Complete biodegradation of  2,4-OMP occurs in  approximately  two  months


(U.S. EPA,  1979);  however, no environmental conditions were described.


II.  EXPOSURE


     A.  Water


         U.S. EPA (1979)  reported that no  specific data are available  on the
                                                                        •

amounts of  2,4-OMP  in drinking  water.  The concentrations of 2,4-OMP present


in drinking  water  vary depending on  the amounts  present in  untreated  water

-------
and on  the efficiency of  water treatment systems  in removing phenolic  com-
pounds.   In  the U.S., the gross  annual  discharge of 2,4-OMP  into  waters was
estimated  to be  100  tons  in  1975 (Versar, 1975).  Manufacturing was  the  lar-
gest  source  of  the  discharge.   Leachates  from  municipal  and  industrial
wastes also contain  the compound  (U.S. EPA, 1979).
         Hoak  (1957) determined  that, at 30°C, the  odor threshold  for  2,4-
OMP was 55.5>jg/l.
     8.  Food
         DMP's  occur naturally  in  tea (Kaiser,  1967)T  tobacco (Baggett and
Marie, 1973; Spears,  1963),  marijuana (Hoffmann, et al. 1975), and a conifer
(Gcrncstasva,   et   al.   1977).   There   is   no  evidence  to  suggest   that
dimethylphenols  occur in  many  plants used  for  food;   however,  it  may  be
assumed that trace amounts are  ingested (U.S. EPA, 1979).                     ;
         The  U-S.  E?A  (1979)  has   estimated  the  weighted  average   biocon-
centraticn  factor  for 2,4-QMP to be  340  for  tna edible portions of  fish ana
shellfish  consumed  by Americans.  This  estimate is  based on  the  measured
steady-state bioconcentration studies in  the  bluegill.
     C.  Inhalation
         2,4-Oimethylphenol  has  been  found  in  commercial  degreasing agents
(NIQSH,  1978),   cresol  vapors  (Corcos,   1935),  cigarette  smoke  condensatss
(Baggett  and  Morie,  1973;  Hoffmann  and  Wynder,  1963;  Smith  and  Sullivan,
1964), marijuana cigarette smoke  (Hoffmann, et  al.  1975) and vapors  from the
combustion  and  pyrolysis  of building materials (Tsuchiya and  Sumi, 1975).
Concentrations  in  smoke  condensates  from six  different brands  of American
cigarettes ranged  from 12.7  to 20.3 mg/cigarette without filters and 4.4  to
                                                                        »
9.1 mg/cigarette with filters (Hoffman and Wynder, 1963).


-------
         There  is no  evidence  in the  available literature  indicating  that
humans are  exposed to 2,4-DMP other  than  as components of complex mixtures.
Adverse  health  effects have been  found  in workers  exposed  to mixtures  con-
taining  amounts  of  2,4-DMP;  however,  the  effects  were  not  attributed  to
dimethylphencl exposure per se (NIOSH, 1978).
     0.  Dermal
         Absorption  through the skin is  thought to  be  the  primary route  of
human exposure to complex mixtures containing 2,4-OMP (U.S. EPA, 1979).
III. PHARMACOKINETICS
     A.  Absorption
         2,4-DMP  is  readily  absorbed  through the skin (U.S.  E?A, 1979).  The
dermal  LD^g for  molten 2,4-OMP  is  1,040  mg/kg in  the rat  (Uzhdovini,  et
al. 1974).
     8.  Distribution
         U.S. EPA  (1979)  found  no  pertinent data on the distribution of  2,4-
OMP in humans or  animals  in the  available literature.  2,6- or 3,4-CMP given
orally to rats  for eight  months  caused  damage to the  liver, spleen, kidneys,
and heart (Maazik, 1963).
     C.  Metabolism
         Urinary  metabolites,  resulting from  oral  administration  of  350  mg
of 2,4-OMP  to  rabbits, were primarily ether-soluble  acid  and  ether glucuro-
nide, with  lesser amounts  of ethereal sulfate,  ester glucuronide  and  free
non-acidic  phenol (Bray,  et al.  1950).    Similar  metabolism  of  the  other
dimethylphenol positional isomers was reported.
     0.  Excretion
                                                                        »
         A  study  done  on  rabbits  by Bray,  et  al.   (1950)  indicates  rapid
metabolism and excretion of 2,4-OMP.

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IV.  EFFECTS
     A.  Carcinogenicity
         Epidemiologic  studies  of workers exposed  to  2,4-OMP were not  loca-
ted in the available literature.
         In a  Carcinogenicity bioassay, 26  female  Sutter mice were dermally
exposed  to 25 jul of  20 percent  2,4-OMP in  benzene  twice  weeekly  for 24
weeks.   Twelve percent  of the  exposed mice  developed carcinomas; however,
benzene  was not  evaluated by  itself  in this  study  Ooutwell  and  Bosch,
1959).   In a related  study,  Boutwell and Bosch  (1959) applied 25 ul of 20
percent  2,4-OMP  in benzene to  the skin of  female  Sutter mice  twice  a week
for 23 weeks  following  a single  acclication  of a  sub-carcinogenic dose (75
ug) of CMBA.   Papillomas  or carcinomas developed in 18 percent  of the mice,
indicating that 2,4-OMP may be a promoting agent for carcinogenesis.
         Fractions of  cigarette smoke condensate containing  phenol,  methyl-
phenols  and 2,4-OMP  have  been snown  to promote  carcinogenesis in mouse skin
bioassays  (Lazar, et al. 1566; Sock, et al. 1971; Roe, et al. 1559).
     B.  Mutagenicity, Teratogenicity and Other Reproductive Effects
         Pertinent data  could  not be. located in  the  available  literature
regarding mutagenicity, teratogenicity and other reproductive effects.
     C.  Chronic Toxicity
         Pertinent information  concerning the chronic  effects  of 2,4- OMP
was not  located  in the available  literatureHU.S.  EPA,  1979);  however,  data
was available  on  other  positional  isomers.   Examination  of  rats  treated
orally with 6  mg/kg  of 2,6-dimethylphenol or  14 mg/kg of 3,4-dircethylphencl
for eight  months  revealed fatty dystrophy and atrophy  of the hepatic  ceJ^ls,

-------
hyaline-droplet  dystrophy  in  the  kidneys,   proliferation  of  mycloid  and



reticular cells, atrophy of  the  lymphoid  follicles  of the spleen,  and paren-



chymatous dystrophy of the heart cells (Maazik, 1968).



     0.  Other Relevant Information



         Tests on isolated rat lungs  indicate  that  2,4-OMP may inhibit vaso-



constriction, most  likely due to its ability to block  ATP (Lunde,  et  al.



1968).  Because of  2,4-OMP's physiological  activity,  U.S. EPA (1979) reports



that  chronic  exposure to the  compound may cause  adverse  health  effects in



humans.



V.   AQUATIC TOXICITY



         Pertinent data cculd net  be  located  in the available literature re-



garding any saltwater species.



     A.  Acute Toxicity



         A  reported  96-hour LC-g value  for  juvenile  fathead  minnows  is



16,750  ,ug/l  (U.S.  EPA,  1979).    For the  freshwater  invertebrate  Dacnnia



maqna, the observed 48-hour LC5Q is 2,120 jjg/1 (U.S. EPA,  1579).



     3.  Chronic Toxicity



         3ased on an  embryo-larval test  with  the fathead minnow,  Pimeohales



oromelas,   the  derived chronic  value  is  1,100 pg/1  (U.S. .EPA,  1978).   NO



chronic values are available for  invertebrate  species.



     C.  Plant Effects



         Based on chlorosis  effects,  the  reported  LC,-Q  for  duckweed,  Lemna



minor,  is  292,800 jjg/1  for  2,4-dimethylphenol exposure  (Blackman,  et  al.



1955).



     D.  Residues
                                                                          »


         A  bioconcentration  factor  of 150 was obtained  for the  bluegill.



The biological half-life  in  the  bluegill  is   less  than one  day,  indicating
                                           -j^r\ iLr



                                           17'?

-------
that  2,4-dimethylphenol residues  are  probably not  a potential  hazard  for
aquatic organisms (U.S. EPA, 1978).
VI.  EXISTING GUIDELINES AND STANDARDS
     Standards have  not been promulgated  for 2,4-CMP  for  any sector of  the
environment or workplace.
     A.  Human
         The draft  criterion for 2,4-dimethylphenol  in water recommended by
the  U.S.  EPA  (1979)    is  15.5 jug/1  based  upon the  prevention  of adverse
effects attributable to the organoleptic properties of 2,4-OMP.
     8.  Aquatic
         For  2,4-dimethylphenol,  the draft  criterion  to  protect  freshwater
aquatic life  is  38  ug/1 as a  24-hour average; the concentration should  not
exceed 86 jug/1 at any  time.   No  criterion exists for  saltwater species (U.S.
EPA, 1575).

-------
                              2.4-OIMETHYLPHENOL

                                  Referencss
Baggett,  M.S.,  and G.P. Morie.   1973.   Quantitative determination of phenol
and  alkylpnenols  in cioarette  smoke and  their removal  by  various  filters.
Tob. Sci.' 17: 30.

Blackman,  E.G.,  et  al.   1955.   The  physiological activity  of substituted
phenols.   I.  Relationships  between  chemical  structure and  physiological
activity.  Arch. Biochem. Biophys.   54: 45.

Bock,  F.G.,  et al.   1971.   Composition  studies on tobacco.   XLIV.  Tumor-
promoting  activity  of  subfractions  of the  weak acid  fraction  of cigarette
smoke condensate.  Jour. Natl. Cancer Inst.  47: 427.

Boutwell, R.K., and O.K. Bosch.   1959.   The tumor-producing action of phenol
and related compounds for mouse skin.  Cancer Res.   19: 413.

Bray,  H.G..  et al.   1950.   Metabolism  of derivatives of  toluene.   5.   The
fate of  the  xylenois  in the rabbit  with further observations on the metab-
olism of  the xyienes.  Biochem. Jour. _ 47: 395.

Corcos,  A.   1939.   Contribution  to  the  study  of  occupational  poisoning ;by
cresols.  Dissertation.  Vigot Freres Editeurs.  (Fre).

Gornostaeva,  L.I.,  et  al.   1977.   Phenols from  abies  sibirica sssentaial
oil.  Khim. Pirir.  Soedin:  ISS 3, 417-413.

Hcak, R.D.   1957.  The  causes  of tastes and odors  in  drinking water.  Proc.
llth Ind. Waste Conf.  Purdue Univ. Eng. Bull.    41:  229.

Hoffmann,  0.,   et  al.   1975.   On  the  carcinogenicity of  marijuana smoke.
Recent Adv. Phytochem.  9:  63.

Hoffmann, 0.,  and E.L. Wyncer.   1963.   Filtration of phenols from cigarette
smoke.  Jour. Natl. Cancer Inst.  30: 67.

Kaiser,  H.E.   1967.   Cancer-promoting  effects  of  phenols  in  tea.   Cancer
20: 614.

Lazar,  P.,  et  al.   1966.   8enzo(a)pyrene, content and  carcinogenicity of
cigarette  smoke  condensate  -  results  of  short-term  and  long-term tests.
Jour. Natl. Cancer Inst.  37: 573.

Lunde,  P.K.,  et  al.    1968.   The inhibitory  effect  of  various phenols on
ATF-induced  vasoconstricticn  in  isolated  perfused  rabbit  lungs.   Acta.
Physiol. Scand.  72: 331.
                                                                      »
Maazik,  I.K.  1963.   Oimethylphenol  (xylenoi)  isomers  and  their  standard
contents in water bodies.   Gig. Sanit. 9: 18.

-------
National  Institute of Occupational  Safety and  Health.   1978.  Occupational
exposure  to  cresol.   OHEW (NIOSH) Publ.  No.  78-133.  U.S.  Dep. Health Edu.
Welfare, Pub. Health Ser., Center for Dis. Control.
Roe, r.j.c.,  et al.  1959.   Incomplete carcinogens  in  cigarette  smoke con-
densate:   tumor-production  by  a  phenolic  fraction.   Br.  Jour.  Cancer
13: 623.
Smith, G.A.,  and  P.2. Sullivan.   1964.  Determination of the steam-volatile
phenols present in cigarette-smoke condensate.  Analyst  89: 312.
Spears,  A.W.   1963.    Quantitative   determination  of  phenol  in  cigarette
smoke.  Anal. Chem.  35: 320.
Tsuchiya,  Y.,   and  K. Sumi.   1975.    Toxicity  of  decomposition products -
phenolic  resin.   Build.  Res. Note-Natl.  Res.  Counc. Can.,  Oiv. Build. Res.
106.
U.S. EPA.   1978.   In-depth  studies  on health  and  environmental  impacts  of
selected  water  pollutants.   Contract NO.  68-01-4646.   U.S.  Environ.  Prot.
Agency, Washington, O.C.
U.S.  EPA.    1979.    2,4-Oimethylphenol:    Ambient  Water  Quality  Criteris
(Draft).
Uzhdovini, E.R.,  et  al.   1974.  Acute  toxicity of  lower  phenols.   Gig.  fr.
Prof. Zaboi.   (2): 58.
Versar,  Inc.   1975.   Identification  of organic compounds in  effluents frcm
industrial sources.  EPA-560/3-75-002.  U.S. Environ. Prat. Agency.
Weast,  R.C.   1576.  Handbook  of chemistry and physics.   57th  ed.  CRC Press,
Cleveland, Ohio.

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                                      No. 88
         Dimethyl Phthalate


  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical acc-uracv.

-------
                      DIMETHYL PHTHALATE



                           SUMMARY



     Dimethyl  phthalate  has been  shown to produce mutagenic



effects in the Ames Salmonella assay.



     Administration  of dimethyl  phthalate to  pregnant rats



by  i.p.  injection has been  reported  to  produce teratogenic



effects in  a  single study.   Other reproductive effects pro-



duced  by  dimethyl  phthalate included  impaired implantation



and parturition in rats following  i.p.  administration.



     Chronic  feeding  studies  in  female  rats  have  indicated



an  effect  of  dimethyl  phthalace   on  the  kidneys.    There is



no evidence to indicate that dimethyl phthalate has carcino-



genic effects.



     Among  the  four   aquatic  species  examined,  freshwater



fish  and  invertebrates appeared   to  be more  sensitive than



their marine  counterparts.   Acute toxicity values  at concen-



trations of  49,500  pg/1 were  obtained for  freshwater  fish.



Criterion could  not  be drafted because of insufficient daca



concerning the toxic effects  of dimethyl phthalates to aquatic



organisms.
                                 **/$ 3 7

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                      DIMETHYL PHTHALATE

I.   INTRODUCTION

     This  profile  is  based  on  the  Ambient  Water  Quality

Criteria Document for Phthalate Esters  (U.S.  EPA,  1979a).

     Dimethyl  phthalate  (DM?)   is  a  diester  of  the  ortho

form of benzene  dicarboxylic  acid.   The compound has  a  mole-

cular  weight of  194.18,  specific gravity  of 1.189,  boiling

point  of  282°C,  and  a solubility  of 0.5  gms in  100 ml  of

water  (U.S. EPA, 1979a).

     DMP is  used as a  plasticizer  for cellulose ester  plas-

tics and as an insect repellant.

     Current  Production:  4.9 x _10"  tons/year  in 1977  (U.S.

EPA, 1979a).

     Phthalates  have  been  detected  in soil,  air,  and  water

samples; in  animal  and  human  tissues;  and in certain  vegeta-

tion.   Evidence  from  _in_ vitro studies indicates  that  certain
                                             i
bacterial  flora  may  be  capable of  metabolizing  DMP  to  the

monoester form  (Englehardt, et al. 1975).

     For  additional   information   regarding  the   phthaiate

escers  in  general,  the reader  is  referred  to  the EPA/SCAO

Hazard Profile on Phthalate Esters (U.S. EPA, 1979b).

II.  EXPOSURE

     Phthalate  esters appear  in  all  areas of  the environ-

ment.   Environmental  release of  phalates may  occur  through

leaching of  the compound  from  plastics,  volatilization from
                                                           »
plastics,  or the  incineration of  plastic  items.    Sources

of  human  exposure  to  phthalates include  contaminated  foods

and  fish,  dermal application, and  parenteral administration

-------
by use  of plastic  blood  bags,  tubing,  and  infusion devices
(mainly  DEHP  release).   Relevant  factors in  the migration
of phthalate  esters  from  packaging  materials  to  food  and
beverages are:   temperature,  surface area contact, lipoidial
nature of the food, and length of contact  (U.S. EPA, 1979a).
     Monitoring studies have indicated that most water phtha-
late   concentrations are  in the ppm range,  or  1-2 pg/liter
(U.S.  EPA, 1979a).   Industrial air  monitoring  studies have
measured  air  levels of  phthalates  from 1.7 to 66 mg/m  (Mil-
kov,   et  al.  1973) .   Information on  levels  of DMP  in  foods
is not available.
     The  U.S. SPA  (1979a)  has estimated the weighted average
bioconcentration  factor for  BMP  to  be  130  for  the  edible .
portions  of fish and  shellfish consumed  by  Americans.   This
estimate  is  based  on  the  measured  steady-state  bioconcen-
tra^ion studies in bluegills.
III.  PHARMACGKINETICS
     Specific  information   is  not   available  on  the  absorp-
tion, distribution,  metabolism, or excretion of  DMP.    The
reader is referred to a general coverage of phthalate metabo-
lism  in the phthalate ester hazard profile (U.S.  EPA, 1979b).
IV.  EFFECTS
     A.    Carcinogenicity
          Pertinent data could  not be located  in  the  avail-
able  literature.
     B.    Mutagenicity
          Dimethyl phthalate has been  shown  to produce  muta-
genic effects  in the  Ames  Salmonella  assay  (Rubin, et  ai.
1979) .

-------
     C.   Teratogenicity

          Administration  of  DMP  to  pregnant  rats  by  i.p.

injection  has  been  reported  to produce  teratogenic  effects

(Singh,  et  al.   1972).    Intraperitoneal  administration  of

DMP  to pregnant  rats  in  another  study did  not  result  in

teratogenic effects  (Peters and Cook, 1973).

     D.   D.   Other Reproductive Effects

          Adverse effects by DMP on implantation and parturi-

tion were  reported  by Peters and  Cook  (1973)  following  i.p.

administration of the compound to rats.

     E.   Chronic Toxicity

          Two-year  feeding  studies  with  dietary  DMP  have

produced some kidney effects  in  female  rats and minor  growth

effects (Draize, et al. 1943).

7.  . AQUATIC TOXICITY

     A.   Acute Toxicity

          Two  freshwater  species   were  examined   for  acute

toxicity  from  dimethyl phthalate  exposure.    The   48-hour

static  "C_Q  for  the Cladcceran,  Daphnia rnagna,  was   33,000

ug/1  (U.S.  EPA,  1978).   The 96-hour  static LC-Q  value  for

the  bluegill,  Lepomis  macrochirus,  was 49,500  pg/1.   For

marine  species,  96-hour static  kC^Q  values for  the   sheeps-

head minnow, Cyprinodon  variegatus,  and mysid shrimp,  Mysid-

opsis bahia, were 58,000 and 73,700 pg/1, respectively.

     B.   Chronic Toxicity
                                                          »
          Pertinent  information  could  not  be  located   in

the available literature.

-------
     C..   Plant  Effects

           Effective   concentrations  based  on  chlorophyl  a

content and  cell number  for  the  freshwater  algae  Selena-

.strum  capricornutum and  the marine algae  Skeletonema costa-

tum  ranged from  39,800  to 42,700  pg/1  and 26,100  to 29,800

ug/1,  respectively.

     D.    Residues

           A  bioconcentration  factor of  57 was  obtained  for

the  freshwater bluegill,  Lepomis macrochirus.

VI.  EXISTING GUIDELINES  AND STANDARDS

     Neither the  human health nor the aquatic criteria derived

by  U.S. SPA  (1979a),  which  are  summarized below,  have  gone

through the process  of  public review;  therefore,   there  is

a possibility that  these  criteria  will be changed.

     A.    Human

           Based  on  "no   effect"  levels  ooserved in  chronic

^do^ * -~ e*-i-J-a=   i— >- = -c  3->r* [•'r«e=   h h P  fl ^  ^T^A  MQ7Qa\  U-^
^.c^C^xii^ OUM^*^.G>^  ^.^. j.
-------
                     DIMETHYL PHTHALATES
                          REFERENCES

Draize,  J.K.,  et  al.    1948.    Toxicological investigations
of compounds  proposed  for  use  as insect  repellents.   Jour.
Pharmacol. Exp. Ther. 93: 26.

Engelhardt,  G.,   et  al.    1975.    The  raicrobial  metabolism
of  di-n-butyl  phthalate  and  related  dialkyl  phthalates.
Bull. Environ. Contain.  Toxicol. 13: 342.

Milkov, L.2., et al.  1973.  Health status of workers exposed
to phthalate  plasticizers  in  the manufacture  of  artificial
leather  and  films  based  on PVC resins.    Environ.  Health
Perspect. Jan. 175.

Peters,  J.W.,  and R.M.  Cook.    1973.    Effects  of phthalate
esters on  reproduction  of  rats.   Environ.  Health Perspect.
Jan.  91.

Rubin, R.J., et al.  1979.   Ames rautagenic assay of a series
of phthalic acid  esters:   positive response  of  the dimethyl
and diethyl esters  in TA  100.   Abstract. Sec. Toxicoi. Annu.
Meet. Naw'orleans, March 11.

Singh, A.,  at  al.-  1972.   Terat:cgenicity of phthalate esters
in ra^s.   Jour. Pharin.  Sci. 61: 51.

U.S.   EPA.   1978.   In-depth studies  on  health  and environ-
mental impacts of selected water  pollutants.   "J.5. Environ.
Prot.  Agency, Contract No. 68-01-4646.

U.S.  SPA.   1979a.   Phthalate Esters:    Ambient  Water Quality
Criteria (Draft).

U.S.   EPA.    1979b.   Environmental  Criteria and  Assessment
Office.  Hazard Profile:  Phthalate Esters  (Draft).

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental  impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                                      No.  89
          Dinitrobenzenes
  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                                DINITROBENZENE5
                                    Summary

     Due to  ths lack of  available  information  no sssessmsnt  of ths °ctsn—
tial of dinitrobenzenes  to  produce  carcinogenic effects,  mutagenic  effects,
teratogenic effects, or adverse reproductive effects can be made.
     Oinitrobenzene  is  the most  potent methemoglobin-forming  agent of  the
nitroaromatics and rapidly produces cyanosis in exposed populations.
                                                  x
     Fish have been  acutely affected  by  exposure to non-specified isomers of
dinitrobenzene at concentrations ranging from 2,000 to 12,000 ug/1.

-------
                                OINITR08ENZENE-
I.  INTRODUCTION
     This  profile  is  based  on  the  Investigation  of  Selected  Potential
Environmental Contaminants:  Nitroaromatics (U.S. EPA, 1976).
     The dinitrobenzenes  exist as the ortho, meta,  or para isomers, depend-
ing on  the  position of the  nitro  group substitutents.  Ortho-dinitrobenzene
(1,2-dinitrobenzene, M.W.  168.1)  is a white, crystalline  solid with a boil-
ing  point  of  319°C,  a  melting point,  of 118°C,  and a specific  gravity of
1.57.   Meta-dinitrobenzene  (1,3-dinitrobenzene)  is  a  yellow,  crystalline
solid  that  melts  at  89-90°C,  boils  at  300-303°C,  and   has  a  density of
1.55.   Para-dinitrobenzene  (1,4-dinitrobenzene)  is  a  white,  crystalline
solid with  a coiling  point of  299°C,  a melting  point of  173-174°C,  and a
density  of  1.63  (windholz,  1976).   The  dinitrobenzenes  have low  aqueous
sclubiiitv =nd z~°  soluble in  alcohol.
     The dinitrcbenzanes are  used in  organic  synthesis-,   the  orccuc'ion of
ayes, and as a camphor substitute  in celluloid prediction.
     The  ccmestic  production  volume  of meta-dinitrcbenzsne   in  1572  was
aporoximately 6 x Id^ tons (U.S. EPA, 1976).
     Oinitrobenzenes are  generally stable in  neutral aqueous  solutions; as
tne medium oecomes  more  alkaline  they may undergo  hydrolysis  (Murto, 1966).
Para-dinitrobenzene  will  undergo  photochemical  reduction  in  isoproparol
under nitrogen,  but this  reaction is  quenched  when  the  solvent  is  aerated
(Hashimoto and Kano, 1972).
     Biodegradation  of  dinitrobenzenes  has  been   reported  for  acclimated
microorganisms (Chambers, et al. 1963; Bringmann and Kuehn, 1959).
                                                                         »
     Based on  the  octanol/water partition coefficient,  Neely  et  al. (1974)
have estimated a low bioconcentration potential  for the dinitrobenzenes.

-------
II.  EXPOSURE
     Industrial  dinitrobenzene  poisoning  reports  have- shown  that  workers
will develop  intense cyanosis  with only  slight  exposure  (U.S. EPA,  1976).
Exposure  to  sunlight  or  ingestion  of  alcohol  may  exacerbate  the  toxic
effects of dinitrobenzene exposure  (U.S. EPA,  1976).
     Monitoring  data on  levels of dinitrobenzenes  in  water,  air,  or  food
were  not  found  in  the  available literature;  human  exposure  from  these
sources cannot be evaluated.
III. PHARMACOKINETICS
     A.  Absorption
          Metherccglobin  formation  in  workers exposed to dinitrobenzene indi-
cates  that  absorption  of the  compound  by  inhalation/dermal  routes  occurs.
Animal  studies demonstrate  that .dinitrobenzene is  absorbed  following  oral
„ ,j_ .• _ .• _ j— i. .• - „
awl—i lj.iv-.Lac-i.ui i.
     3.  Distribution
          Pertinent  information on districution  of  dinicrcbenzenes \vas  not
found  in the available  literature.
     C.  Metabolism
          Dinitrobenzene  undergoes both  metabolic  reduction  and  cxidaticn.
.-.nri2_  s^ucies  indicate ... .sti •„.
-------
     0.  Excretion
          Oral   administration  of   radiolabelled   meta-dinitrobenzene   to
rabbits was  followed  by elimination of  65-93%  of the dose  within two  days.
Excretion was  almost  entirely via the urine;  1-5% of the administered  label
was determined in the feces (Parks, 1961).
IV.  EFFECTS
     A.  Carcinogenicity
          Information on  the Carcinogenicity of  the dinitrobenzenes was not
found in the available literature.
     B.  Mutagenicity
          Information  on  the mutagenicity  of the  dinitrobenzenes  was not
found in  the available  literature.  The  possible dinitrobenzene metabolite,
dinitropnenoi  (U.S. EPA,  1979),  has  been reported to induce chromatid breaks
in bone ."arrow cells of injected mice (Micra ana Manna, 1971).
     C.  Teratogenicity
          Information on  the  ceratogenicity  of the  oinitrobenzenes  was not
fc-jrd in  the available  literature.  The  possible dinitrcbenzene metacoiite,
dinitropnenoi  (U.S.  EPA,  1979), has produced  developmental  abnormalities  in
the sea urchin (Hagstrom  and  Lonning,  1966).   No  effects were seen follcv.-i.-g
injection cr era! administration of dinitropnenol to mice (Gioson, 1973).
     0.  Other Reproductive Effects
          Pertinent information was not found in the available literature.
     E.  Chronic Toxicity
          Oinitrobenzene  is  the most  potent methemoglobin-forming  agent of
the  nitroaromatics.   Poisoning  symptoms  in  humans  may be potentiated  by
                                                                          »
exposure to sunlight or ingestion of alcohol (U.S. EPA. 1976).

-------
V.   AQUATIC TOXICITY



          A.  Acute Toxicity



               McKee and Wolf  (1963)  have presented a brief  synopsis  of the



toxic effects of dinitrobenzenes to aquatic  life.   A  study by LeClsrc  (1960)



reported lethal  doses  of non-specific isomers of  dinitrobenzene  for minnows



(unspecified) at concentrations  of 10,000 to 12,000 ug/1  in  distilled water



or 8,000 to  10,000 ug/1  in  hard  water.   Meinck et al. (1956)  reported  lethal



concentration of 2,000 pg/1 for  unspecified dinitrobenzenes  for  an unspeci-



fied fish species.



     8.  Chronic Toxicity



          Pertinent  data could  not  be  found  in  the 2vail3ble  litsrsturs



regarding aquatic toxicity.



     C.  Plant Effects



          Howard et al.  (1975) rsport that  the  algae  Chlorella sp.  dispiayec



inhibited  photosynthetic activity  upon  exposure  to  n-dinitrobenzene at  a



concentration of 10~4 M.



VI.  EXISTING GUIDELINES



     The 8-hour  time-weighted-averace  (TWA) occupational  exoosure  limit for



dinitrobenzenes is 0.15 ppm(ACGIH, 1974).

-------
                                DINITR08ENZENES

                                  References
ACGIH.   1974.   Committee  on threshold  limit values:  Documentation  of the
threshold limit values for substances in the  workroom air.  Cincinnati, Ohio.

3ringmann, G.  and R. Kuehn.   1959.   Water toxicity  studies  with protozoans
as test organisms.  Gesundh.-Ing.  80: 239.

Chambers, C.W.,  et al.  1963.   Degradation of aromatic  compounds by pheno-
ladopted bacteria.  Jour. Water Pollut. Contr. Fedr. 35: 1517.

Gibson, J.£.   1973.   Teratology  studies  in mice with 2-sec-Butyl-4,- 6-dini-
trophenol (Dinoseb).  Fd. Cosmet. Toxicol.  11: 31.,.

Hagstrom, 8.E.  and S. Lcnning.  1966.  Analysis  of the effect of  -Qinitro-
phenol on  cleavage and  development  of the sea urchin  embryo.  Protoplasms.
42(2-3): 246.  •

Hashimoto, 5.  and K.  Xano.   1572.  Fhotocnsmical  reduction  of nitrobenzene
and  reduction  intermediates.    X.   Photochemical   reduction  of  the  mono-
substitutsc1 nitrcbenzanes in 2-propanol.  Bull. Chem. Soc. Jap.  45(2): 549.

Howard,  P.M.,  et  al.   1975.  Investigation  of selected  potential  environ-
mental  contaminants:  Nicrcaromatico.   Syracuse,  N.Y.:    Syracuse  Research
Corccration,  T° 76-573.

LeClerc, E.   1960.   Self  purification .of  streams  and the  relationship  ce-
tv/een  chemical arc1  biological  tests.   2nd  Symposium on  the  Treatment  of
Waste Waters.  Pergamon Press, p. 282.

McKee,  J.E.  and  H.w. Wolf.   1963.   Water  quality criteria.   The  Resource
Agency of California  State Water Quality Control Board Publication NO. 3-A.

Meinck, ,-.,  et al.   1956.   Industrial waste water.  2nd  ed.  Gustav Fisher
Verlag Stutosrt, o. 536.

Micra,  A.3.  and  G.K. Manna.   1971.   Effect of  some phenolic compounds  on
chromosomes of bone marrow cells on mice.  Indian J. Med.  Res.  59(9): 1442.

Murto, J.  1966.   Nucleophilic reactivity.   Part  9.  Kinetics  of the  reac-
tions  of  hydroxide   ion and  water  with  picrylic compounds.   Acta  Chem.
Scand.  20: 310.
                                                      s
Neely,  W.B.,   et   al.   1974.   Partition  coefficient  to  measure  bicconcan-
tration  potential of organic  chemicals  in  fish.    Environ.  Sci.  Technoi.
3: 1113.
                                                                        «

Parke,  O.W.    1961.   Detoxication.   LXXXV.   The  metabolism  of  m-dinitro-
benzene-C^ in the rabbit.   Biochem. Jour.  78: 262.

-------
U.S. EPA.   1976.   Investigation of  selected  potential environmental contam-
inants:  Nitroaromatics.

U.S. EPA.   1979.   Environmental  Criteria and Assessment  Office.   2,4-Dini-
trophenol:  Hazard Profile (Draft).

Windholz, M.  (ed.)   1975.  The  Merck  Index.   9th ed.   Merck  and  Co.,  Inc.,
Rahway, N.J. p. 3269.

-------
                                     No. 90
        4,6-Dinitro-o-cresol
  Health and  Environmental Effects
U.S.  ENVIRONMENTAL PROTECTION AGENCt
       WASHINGTON, D.C.  20460

           APRIL 30, 1980
                 Jo-/

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical acc-aracy.

-------
                    4,6-DINITRO-O-CRESOL



                           SUMMARY



     There is no available evidence to  indicate  that  4,6-



dinitro-ortho-cresol (DNOC) is carcinogenic.



     This compound has produced some DNA damage  in  Proteus



mirabilis but failed to show mutagenic  effects in the Ames



assay or in E. coli.  Available information does not



indicate that DNOC produces teratogenic or adverse



reproductive effects.



     Human exposure incidents have shown that DNOC  produces



an increase in cataract formation.

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                     4 ,6-DINITRO-O-CRESOL



I.   INTRODUCTION



     This profile  is based  on  the Ambient  Water  Quality Cri-



teria Document for Nitrophenols  (U.S.  EPA,  1979a).



     Dinitrocresols  are  compounds closely  related  to  the di-



nitrophenols; they bear  an  additional  2-position methyl



group.  The physical properties  of  4,6-dinitro-ortho-cresol



(DNOC, M.W. 198.13)  include a  melting  point of 85.8°C  and  a



solubility of 100  mg/1 in water  at  20°C  (U.S. EPA,  1979a).



     Dinitro-ortho-cresol is used primarily as a blossom



thinning agent on  fruit  trees  and as a fungicide,  insecticide



and miticide on the  fruit trees  during the dormant  season.



There is no record of current  domestic manufacture  of  DNOC



(U.S. EPA, 1979a).   For  additional  information regarding the



nitrophenols in general, the reader  is referred  to  the  Hazard



Profile on Nitrophenols  (U.S.  EPA,  1979b).



II.  EXPOSURE



     The lack of monitoring data makes it difficult to  assess



exposure from water, inhalation, and foods.  DNOC has  been



detected at 13 mg/1  in effluents from  chemical plants  (U.S.



EPA, 1979a).



     Exposure to DNOC appears  to be primarily through  occupa-



tional contact (chemical manufacture,  pesticide  application).



Contaminated water may result  in isolated poisoning inci-



dents.



     The U.S. EPA  (1979a) has  estimated a weighted average



bioconcentration factor  for DNOC to be 7.5 for the edible



portions of fish and shellfish consumed by Americans.   This



estimate is based on the octanol/water partition coefficient.
                             X

-------
III. PHARMACOKINETICS



     A.   Absorption



          WOC  is  readily  absorbed  through  the skin,  the res-



piratory tract, and the gastrointestinal  tract (NIOSH,



1978).



     B.   Distribution



          DNOC  has been found  in  several  body  tissues;  how-



ever, the compound may be  bound to  serum  proteins,  thus pro-



ducing non-specific organ  distribution  (U.S. EPA,  1979a).



     C.   Metabolism



          Animal studies on  the metabolism  of  DNOC indicate



that like the nitrophenois,  both  conjugation of  the  compound



and reduction of the nitro groups to  amino  groups  occurs.



The metabolism  of  DNOC to  4-amino-4-nitro-c-crescl  is a de-



toxification mechanism that  is effective  only  when  toxic



doses of DNOC are  administered (U.S.  EPA, 19T9a).   The



metabolism of DNOC Ls very slow in  man  as compared  to that



observed in animal studies (King  and  Harvey, 1953).



     D.   Excretion



          The experiments  of Parker and coworkers  (1951)  in



several animal  species indicates  that DNOC  is  rapidly ex-



creted following injection;  however,  Harvey, et  al.  (1951)



have shown slow excretion of DNOC in  volunteers  given the



compound orally.  As in metabolism, there is a substantial



difference in excretion patterns of humans  vs. experimental



animals.

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IV.  EFFECTS



     A.   Carcinogenicity



          Pertinent data could not  located  in  the  available



1iterature.



     B.   Mutagenicity



          Adler, et al.  (1976) have  reported  that  DNOC  shows



some evidence of producing DNA damage  in Proteus mirabilis.



Testing of this compound in  the Ames Salmonella system



(Anderson, et al., 1972) or  in JB. coli  (Nagy,  et al., 1975)



failed to show any mutagenic effects.



     C.   Teratogenicity and Other Reproductive Effects



          Pertinent data could n.ot be  located  in the



available literature regarding teratogenicity  and  other



reproductive effects.



     D.   Chronic Tcxicity



          Human use of DNOC as a dieting aid has produced



poisoning cases at accepted thereputic dose levels, as well



as some cases of cataract development  resulting from



overdoses (HIOSH, 1978).



     E.   Other Relevant Information



          DNOC is an uncoupler of oxidative phosphorylation,



an effect which accounts for its high acute toxicity in



mammals.



V.    AQUATIC TOXICITY



     Pertinent information could not be located in the



available literature.

-------
VI.  EXISTING GUIDELINES AND STANDARDS



     A.   An eight-hour TLV exposure  limit  of  0.2  mg/m^  has



been recommended for DNOC by the ACGIH  (1971).



          A preliminary draft water criterion  for  DNOC has



been established at 12.8 ug/1 by the  U.S. EPA  (1979a).   This



draft criterion has not gone through  the process of  public



review; therefore, there is a possibility that  the criterion



may be changed.



     B.   Aquatic



          Criteria for the protection of freshwater  and



marine aquatic organisms were not drafted due  to lack of



tcxicoiogical evidence (U.S. SPA, 1979a) .

-------
VI.  EXISTING GUIDELINES AND STANDARDS



     A.   An eight-hour TLV exposure  limit  of  0.2  mg/m-3  has



been recommended for DNOC by the ACGIH  (1971).



          A preliminary draft water criterion  for  DNOC has



been established at 12.3 ug/1 by the  U.S. SPA  (197Sa).   This



draft criterion has not gone through  the  process of  public



review; therefore, there is a possibility that  the criterion



may be changed.



     B..   Aquatic



          Criteria for the protection of  freshwater  and



marine aquatic organisms were not drafted due  to lack of



toxlcological evidence (U.S. EPA, 1979a).

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                                       No. 91
         2,4-Dlnitrophenol
  Health and Environmental Effects
CJ.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980
                _^^*^--A.
                 ) U U ^

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental  impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                        2,4-DINITROPHENOL


                             Summary


     There  is  no  evidence to indicate that  2,4-dinitrophenol  pos-


sesses carcinogenic activity.


     Genetic toxicity  testing  has shown positive effects  in mouse


bone marrow cells and in  E_._ coli.    In_  vitro cell culture assays


failed to show the potential for mutagenic activity of 2,4-dinitro-


phenol as measured by unscheduled ONA synthesis.


     Teratogenic  effects  have  been  observed  in  the  chick embryo


following administration of  2,4-dinitrophenol.  Studies  in mammals


failed to show that the compound produced any  taratocenic  effacts.


At the levels of compound used  in these mammalian studies, embryo-
                                                                 *

toxic effecrs were observed.


     Human  use of  2,4-dinitrophenol  as  a 'dieting aid  has produced


seme cases  cf  agranulccytosis,  neuritis,  functional  heart da.~age,


and cataract development.


     For aquatic  organisms  LC^g  values  ranged from  620 ug/1  for


the biuegill tc 15,700 j:g/l  for the fathead minnow.

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                        2,4-DINITROPHENOL

I.    INTRODUCTION

     This profile  is based on  the  Ambient Water Quality Criteria

Document for Nitrophenols  (U.S. EPA, 1979a).

     The dinitrophenols are  a family of compounds composed of  the

isomers resulting from nitro-group substitution of phenol at  vari-

ous positions.  2,4-Dinitrophenol has a molecular weight of  184.11,

a melting point of  114-115°C,  a density of 1.683 g/ml  and  is  sol-

uble in water at 0.79 g/1  (U.S. EPA, 1979a).

     The  dinitrophenols  are  used as  chemical  intermediates  for

sulfur  dyes,  azo  dyes, photochemicais,  pest  control agents,  wood

preservatives,  and  explosives  (U.S. EPA,  1979a).   The 1963  pro-

duction of  2,4-dinitroohenol was 4.3  x 10^  tons/yr.    (U.S.  EPA^
                       *"                                           *

1979a).

     For  additional  ir.fcrniaticn  regarding  the  nitrochenols as

a class,  the' reader  is referred  to  the Hazard  Profile on Nitro-

phenols (1979b).

II.  EXPOSURE

     The  lack of  monitoring data  for  the nitrophenols makes it

difficult co  assess  exposure  from  water,   inhalation,  and foods.

Nitrophenols  have  been detected in  effluents  from chemical plants

(U.S. SPA,  1979a) .    Dermal  absorption of  the  dinitrcphenols  has

been reported  (U.S. EPA,  1979a).

     Exposure  to   nitrophenols appears  to  be   primarily   through

occupational  contact  (chemical  plants,   pesticide  application)  .
                                                             »
Contaminated water may contribute to isolated poisoning  incidents.

The  U.S.  EPA   (1979a) has  estimated  the weighted  average  biocon-

centration  factor  for 2,4-dinitrophenol to be  2.4  for  the edible

-------
portions of  fish and shellfish consumed by Americans.   This esti-

mate  was based  en  the. cctanol/water  partition  coefficients  of

2,4-dinitrophenol.

III. PHARMOCOKINETICS

     A.   Absorption

          The dinitrophenols  are readily absorbed  following oral,

inhalation, or dermal administration  (U.S. EPA,  1979a).

     3.   Distribution

          Dinitrophenol  blood  concentrations rise  rapidly  after

absorption, with little subsequent distribution  or  storage at tis-

sue sites (U.S. EPA, 1979a).

     C.   Metabolism

          Metabolism of  the nitrophenols occurs through  conjuga-

tion and reduction of nitro-grcups to  amino—groups,  or oxidation to

dihydric-nitropher.Qls (U.S. EPA,  1979a) .

     D.   Excretion

          Experiments .with several  animal  species  indicate  that

urinary clearance of dinitrophenols is  rapid  (Harvey, 1959).

"TT   r~--n r^T">/-*fT*r^
>/l.  ixriCTS                                             •**

     A.   Carcinogenicity-

          2,4-Dinitrophenol  has  been  found  not to  promote  skin

tumor  formation  in mice following  DMBA initiation (Bautwell  and

Sosch, 1959) „

     B.   Mutagenicity

          Testing  of  2,4-dinitrophenol  has  indicated   mutagenic
                                                             »
effects  in  E. coli  (Demerec,  et al.. 1951) .   In  vitro  assays  of

unscheduled  DNA  synthesis  (Friedman  and  Staub,  1976)  and  DNA

-------
damage  induced  during  cell culture (Swenberg, et al. 1976)  failed

to show the potential for mutagenic activity  of  this compound.

     C.   Teratogenicity

          2,4-Dinitrophenol has been shown  to produce development-

al abnormalities  in  the  chick  embryo  (Bowman, 1967; Miyatmoto,  et

al. 1975) .   No teratogenic effects were seen following intragastric

administration  to rats (Wulff, et al. 1935) or intraperitoneal ad-

ministration to mice (Gibson, 1973).

     D.   Other Reproductive Effects

          Feeding of  2,4-dinitrophenol to  pregnant rats produced

an increase  mortality  in  offspring  (Wulff,  et  al. ,  1935); simi-

larly,  intraperitoneal  administration of  the  compound  to mice

induced  embryotoxicity  (Gibson,  1973).   The  influence  of this

compound or. maternal health may have contributed to these effects.

     S.   Chronic Toxicity

          Use of 2,4— dir.itroohenol as a human distinc aid has pro-

duced  some  cases of  agranulocytosis,  neuritis,  functional heart

damage, and  a  large number  of patients  suffering  from cataracts

(Horner, 1342).

     ?.   Other Relevant Information

          2,4-Dinitrophenol is a  classical  uncoupler  of oxidative

phosphorylation,  an effect  which  accounts  for  its   high acute

toxicity in mammals.

          A  synergistic  action  in  producing ' teratogenic   effects

in the  developing chick  embryo has been  reported  with  a combina-
                                                             »
tion of 2,4-dinitrophenol and insulin  (Landauer and Clark,  1964).

-------
V.   AQUATIC TOXICITY


     A.   Acute


          The  bluegill  (Lepomis macrochirus)  was  the most  sensi-


tive aquatic organism tested, with an LC^Q of 620 pg/1 in  a  static,


96-hour assay  (U.S.  EPA,  1978).   Juvenile fathead minnows  (Pime-


phales promelas)  were more  resistant in  flow through tests, with


an I*CCQ of  16,720 pg/1  (Phipps, et  al.   manuscript).  The  fresh-


water  cladoceran  (Daphnia  magna)  displayed  a range  of  observed


LC5Q  values of  4,090   to  4,710  pg/1  (U.S.  EPA,  1979a)  .   Acute


values for  the marine  sheepshead minnow  (Cyprinodon variegatus)


are  LC--,  values  ranging  from  5,500  to  29,400   pg/1 (Rosenthal


and  Stelzer,  1970).   The marine mysid  shrimp (Mysidopsis  bahia)


had an LC5Q of 4,350 ug/1 (U.S. EPA, 1978).                       j.


     3.   Chronic Toxicity


          Pertinent  data  could  not  be  located  in  the   available


literature.


     C.   Plant Effects


          Effective  concentrations  for  freshwater  plants   ranged


from  1,472  pg/1  for duckweed  (Lemna minor)  to  50,000  /ag/1  for


the  alga  (Chlorella  pyranoidosa)  (U.S.  EPA,  1979a).   The  marine


alga  (Skeletonema costatum)  was more  resistant  with  a  reported


96-hour SC-Q value based on  cell numbers of 98,700 pg/1.


     D.   Residues


          Based on the octanol/water partition coefficient,  a bio-


concentration  factor of 8.1  has  been estimated  for   2,4-dinitro-
                                                             »

phenol for aquatic organisms with a lipid content of 8 percent.
                                  ?f-7

-------
V.   EXISTING GUIDELINES AND STANDARDS



     Neither the human health nor aquatic criteria derived by U.S.



EPA (1979a)  which are  summarized below have undergone the process of



public review;  therefore,  there  is a  possibility  that  these criter-



ia will be changed.



     A.   Human



          The  draft  water,  criterion   for  dinitrophenols,  based



on  data  describing adverse effects,  has  been  estimated  by  the



U.S. EPA (1979a) as 68.6 ug/1.



     B.   Aquatic



          For protecting  freshwater  aquatic  life,  the draft cri-



terion is 79 ug/1 as a 24-hour avetage concentration  not to exceed



180  ug/1.    The marine  criterion has  been  proposed as  37  ug/1



as  a  24-hour average not  to  exceed  84  ug/1  at  any  time  (U.S.



EPA, 1979aj .



          To protect  saltwater  life,  the  draft criterion  is  37



ug/1 as a 24-hour average  not  to  exceed  84  ug/1 at any time (U.S.



EPA, 1979a).

-------
                         2,4-DINITROPHENOL

                            REFERENCES
Bautwell,  R.,  and  D.  Bosch.   1959.   The tumor-promoting  action
of  phenol  and  related  compounds  for  mouse  skin.    Cancer  Res.
19: 413.

Bowman, P.  1967.   The effect of  2,4-dinitrophenol on  the  develop-
ment of early chick embryos.  Jour. Embryol. Exp. Morphol. 17: 425.

Demerec, M., et al. 1951.  A survey  of  chemicals  for rautagenic ac-
tion on E.  coli.  Am. Natur. 85:  119.

Friedman, M.A., and J. Staub. 1976.  Inhibition of mouse testicular
DNA synthesis  by mutagens  and  carcinogens as  a potential  simple
mammalian assay for mutagenesis.   Mutat.  Res.   37: 67.

Gibson, J.E. 1973.  Teratology  studies  in mice with 2-secbutyl-4,
5-dinitrcphenol (dinoseb).  Food  Cosmet.  Toxicol.  11:  31.

Harvey, D.G. 1959.  On  the  metabolism of some  aromatic  nitro  com-
pounds by different species of animal.  Part  III.  The  toxicity of
the dinitrophenols, with a note  on the effects of  high environment^
al temperatures.  Jour.  Pharm. Pharmacoi.   II:  462.

Homer, W.D. 1942.  Dinitroohenol  and ics  relation to formation of
cataracts.  Arch. Ophthal.  27: 1097.

Landauer, W., and  E.  Clark.  1964.   uncoupiers  of oxidative phos-
phorylation and teratogenic activity  of insulin.  Nature 204:  235.

Miyamoto, K., ec al. 1975.  Deficient myelination by 2,  4-dinicro-
phencl administration  in early stage of  development.    Teratology
12: 204.
                      .-nu,A -. *^ . . ^ — <_—.,,,_,;_.. ,— —  -> « — — -s '  -. „ -i  — ,-*.— i. : »«
                      J. ** ^ AW M <_ w wOA.iO.fcwy <*J A.  ^llCliw^  dli**J  d U«^ M *- X W
phenols to the fathead minnow.  (Manuscript).

Rosenthal, H., and  R. Stelzer.   1970.   Wirkungen von 2,4-und 2,5-
dinitrophenol  auf  die   Embryonalentwicklung des  Herings  Clupea
harengus.  Mar. Biol.  5: 325.                                  "

Swenberg, J.A., et  al.  1976.   In  vitro DNA  damage/akaline elution
assay  for  predicting carcinogenic  potential.,   Biochem.  Biophys.
Res. Commun.  72:  732.

U.S. EPA.   1979a.   Nitrophenols:   Ambient water quality criteria.
(Draft).

U.S. EPA.   I979b.   Nitrophenols:   Hazard  Profile.   Environmental
Criteria and Assessment Office (Draft).
                                *    LJ.  '&-
                                   * i D i) ^

-------
U.S. EPA.   1978.   In-depth  studies on  health  and  environmental
impacts of selected  water pollutants.  Contract No. 68-01-4646.

Wulff,  L.M.B., et al. 1935.   Some effects of alpha-  dinitrophenol
on pregnancy  in the white rat.  Proc. Soc.  Exp. Biol. Med.  32:  678.
                               9/-/0

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                                       No.  92
           Dinitrotoluene
  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a  survey of the potential  health
and environmental hazards from exposure  to  the  subject  chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources,  this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented  by the
subject chemical.   This  document  has undergone  scrutiny  to
ensure its technical accuracy.
                              ,„{,{} ^7 /
                            *  I V ) ''"

                              7

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                        DINITROTOLUENE



                           SUMMARY



     Most of the information on the effects of dinitrotoluene



deals with 2,4-dinitrotoluene.  2,4-Dinitrotoluene  induces



liver cancer and mammary tumors in mice and is mutagenic



in some assay systems.  Information on teratogenicity was



not located in the available literature.  Chronic exposure



to 2.,4-dinitrotoluene induces liver damage, jaundice, raethemo-



globinemia and anemia in humans and animals.



     Acute studies in freshwater fish and invertebrates



suggest that 2,3-dinitrotoiuene is much more  toxic  than



2,4-dinitrotoluene.

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                        DINITROTOLUENE



I.   INTRODUCTION



     This profile is based on the Ambient Water Quality



Criteria Document for Dinitrotoluene  (U.S. EPA, 1979).



     There are six isomers of dinitrotoluene  (CH-,CgH3  (N02)2;



molecular weight 182.14), with the 2,4-isomer being  the



most important commercially.  2,4-Dinitrotoluene has a melt-



ing point of 71°C, a boiling point of 300°C with decomposi-



tion, and a solubility in water of 270 mg/1 at 22°C.  It



is readily soluble in ether, ethanol, and carbon disulfide



(U.S. EPA, 1979).  2,6-Dinitrotoluene has a melting point



of 66°C and is soluble in alcohol.  Production in 1975 was



273 x 10  tons per year for the 2,4- and 2,6- isomers com-



bined (U.S. S?A, 1979) .



     Dinitrotcluene is an ingredient of explosives for commer-



cial anc military use,  a chemical stabilizer  in the manufac-



ture of smokeless powder, an intermediate in  the manufacture



of toluene diisocyanates used in the production of urethane



polymers, and a raw material for the manufacture of dyestuffs.



Dinitrotoluenes are relatively stable at ambient tempera-



tures (U.S. EPA, 1979).



II.  EXPOSURE



     A.    Water



          Data on concentration levels for dinitrotoluene



were not available.  Dinitrotoluene waste products are dumped



into surface water or sewage by industries that manufacture



dyes, isocyanates, polyurethanes and munitions (U.S. SPA,



1979) .

-------
     B .   Food



          According to  the U.S.  EPA  (1979),  the likelihood



of dinitrotoluene existing in  food is minimal  since it is



not used as a pesticide or herbicide.



          The U.S. EPA  (1979)  has estimated  the weighted



average bioconcentration  factor  for  2,4-dinitrotoluene to



be 5.5 for the edible portions of fish  and shellfish consumed



by Americans.  This estimate is  based on  the octanol/water



partition coefficient.



     C.   Inhalation



          Exposure to dinitrotoluene by inhalation  is most



likely to occur occupationally  (U.S. EPA, 1979).  However,



pertinent data could not  be located  in  the available litera-



ture en atmospheric concentrations of dinitrctcius-2 and,



thus, ocssible human •=.v<'oosure canr.ct be sstimatac.
     A.   Absorption


                            14
          The absorption of   "C-labeled  iscmers of  dinitrctol



uene after oral administration to rats was essentially com-



plete within 24 hours, with 60 to 90 percent of the dose



being absorbed.  The 2,4- and 3,4-isomers were absorbed



to a greater extent than the 3,5- and 2,5- isomers, which



in turn were absorbed to a greater extent than the  2,3-



and 2,5-isomers (Hodgson, et al. 1977) .  2/4-Dinitrotoluene



is known to be absorbed through the respiratory tract and



skin (U.S.  EPA, 1979) .

-------
     B.   Distribution .



          Tissue/plasma ratios of radioactivity after adminis-


           14
tration of   C-labeled dinitrotoluene to rats  indicated


             14
retention of   C DMT in both the liver and kidneys but not



in other tissues (Hodgson, et al./ 1977).  A similar experi-



ment with tritium-labeled 2,4-dinitrotoluene  ( H-2,4-DNT)



in the rat showed relatively high amounts of radioactivity



remaining in adipose tissue, skin, and liver seven days



after administration (Mori, et al., 1977).



     C.   Metabolism



          No studies characterizing the metabolism of dinitro-



toluene in mammals are available.  However, on the basis



of a comparison of the metabolism of 2,4-dinitrotoluene



and 2,4,5-trinitrotoluene in microbial sys-ems, and the



known metabolism of 2,4,5-trinitrotoluene in mammals, the



U.S. Z?A (1379) speculated that tiie metabolites of 2,4-di-



nitrotoluene in mammals would be either tcxic  and/or car-



cinogenic .



     D.   Excretion

                                                      T  |

          In studies involving oral administration of ~"C-



dinitrotoluene or  H-2,4-dinitrotoluene to rats (Hodgson,



et al., 1977; Mori, et al., 1977), elimination of radioactiv-



ity occurred mainly in urine and feces.  No radioactivity

                                           .•

was recovered in the expired air.  About 46 percent of the



administered dose in the latter study was excreted in the



feces and urine during the seven days following administration.

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IV.  EFFECTS



     A.   Carcinogenicity



          2,4-Dinitrotoluene fed to rats and mice  for  two



years produced dose-related increases  in fibromas  of the



skin in male rats and fibroadenomas of the mammary gland



in female rats.  All of these were benign tumors.  No  statis-



tically significant increase in tumor  incidence was noted



in mice (Natl. Cancer Inst., 1978).



          In a second bioassay of rats and mice fed 2,4-



dinitrotoluene for two years, the findings in rats included



a significant increase of hepatccailuiar carcinomas and



neoplastic nodules in the livers.of females, a significant



increase of mammary gland tumors in females, and a suspicious



ir.craase of hepaeocellular carcincrr.as of che liver in  males.



Male mice had a highly signi'f icar.c ir.craase of kidney  tumors



(Lee, et ai., 1975; .



     2.   Mutagenicicy



          2,4-Dinitrotoluene was mutagenic in the  dominant



lethal assay and in Salmonella typhimurium strain  TA1535



(Hodgson,  et al. 1976).   Cultures of lymphocytes and Kidney



cells derived from rats fed 2,4-dinitrotoluene had signifi-



cant increases in the frequency of chromatid gaps  but  not



in translocations or chroraatid breaks  (Hodgson, et al.,



1976).



          The mutagenic effects of products from ozonation
                                                          »


or chlorination of 2,4-dinitrotoluene and other dinitrotoluenes
                               7*1-7

-------
were negative in one study  (Simmon, et al.,  1977),  and,


for products of ozonation alone, were ambiguous  in  another


study (Cotruvo, et al., 1977).


     C.    Teratogenicity and other Reproductive  Effects


          Pertinent data could not be located in the avail-


able literature.


     D.    Chronic Toxicity


          Chronic exposure  to 2,4-dinitrotoluene may produce


liver damage, jaundice, methemoglobineraia  and reversible


anemia with reticulocytosis in humans and  animals  (Linch,


1974; Key, et al. 1977; Proctor and Hughes,  1978; Kovalenko,


1973).


     E.    Other Relevant Information                 N


          Animals were more resistant, to the toxic  effects


of 2,4-dinitrotoluene administered in the  diet when given


diets high in fat or protein  (Clayton and  Baumann,  1944,


1948; Shils and Goldwater,  1953) or protein  (Shils  and Gold-


water, 1953) .


          Alcohol has a synergistic effect on the toxicity


of 2,4-dinitrotoluene (Friedlander, 1900; McGee, et al.,


1942).


          In subacute studies (13 weeks),  2,4- and 2,6-dini-


trotoluene caused methemoglobinemia, anemia with reticulocyto-


sis, gliosis and demyelination in the brain', and atrophy


with aspermatogenesis of the testes in several species  (Ellis,
                                                          *

et al.,  1976).

-------
V.   AQUATIC TOXICITY
     A.   Acute Toxicity
          Static assays with  the  freshwater  bluegill (Lepomis
macrochirus) produced a 96-hour LCen  value of  330  ug/1 for
2,3-dinitrotoluene  (U.S. EPA, 1978),  while the same assay
with the fathead minnow (Pimephales promelas)  produced a
96-hour LC5Q value of 31,000 pg/1  for  2,4-dinitrotoluene
(U.S. Army, 1976).  The greater toxicity  of  2,3-dinitrotoluene
when compared to that of 2,4-dinitrotoluene, was demonstrated
in 48-hour static assays with the  freshwater cladoceran,
Daphnia magna, with I*CCQ values of 660 jjg/l(U.S. EPA,  1978)
and 35,000 ug/1 (U.S. Army, 1976)  being reported.   A single-
marine fish, sheepshead minnow  (Cyprinodon variegatus),
has been tested for adverse acute  effects of 2,3-dinitro-
toluene.  A 96-hour static assay LC5Q- value of 2,280 ug/1
was reported (U.S. EPA, 1978).  For marine invertebrates
a 96-hour static LC5Q value of 590 jig/1 was obtained for
the mysid shrimp (Mysidopsis bahia) with  2,3-dinitrotoluene.
     3.   Chronic Toxicity
          The sole chronic study examining the effects of
2,3-dinitrotoluene in an embryo-larval assay on the fathead
minnow produced a chronic value of 116 ug/1 based  on reduced
survival of these stages.  No marine chronic data  were pre-
sented  (U.S. EPA, 1979).
     C.    Plant Effects
           Concentrations of 2,3-dinitrotoluene that caused
50  percent  adverse effects in cell numbers or  chlorophyll

                              ^^^^^2OL-_
                                lu ) i!!"*

-------
a in the freshwater algae, Selenastrum capricornutum, were

1,370 or 1,620 ug/1, respectively.  These same effects mea-

sured in the marine algae, Skeletonema costatum,  showed

it to be more sensitive.  ECco values were 370 or  400 ug/1,

respectively.

     D.   Residues

          A bioconcentration factor of 19 was obtained for

aquatic organisms having a lipid content of 8 percent  (U.S.

EPA, 1979).

VI.  EXISTING STANDARDS AND GUIDELINES

     Neither, the human health nor aquatic criteria derived

by U.S. EPA  (1979), which are summarized below, have gone

through the process of public review; therefore,  there is

a possibility that these criteria may be changed.

     A.   Human

          Based on the induction of fibroadenomas of the

mammary gland in female rats (Lee, et al., 1978), and using

the "one-hit" model, the U.S. EPA (1979) has estimated levels

of 2,4-dinitrotoluene in ambient water which will result

in specified risk levels of human cancer:


Exposure Assumptions         Risk Levels and Corresponding Draft Criteric
      (Per day)                     £    ^7      ^-6^o         '

2 liters of drinking water and       7.4 ng/1   74.0 mg/1  740 ng/1      ;
consumption of 18.7 grams fish                                           •
and shellfish.

Consumption of fish and shell-      .156 ^ug/1  1.56 pg/1  15.6
fish only.                                                '

-------
          The American Conference of Governmental Industrial
Hygienists (1978)  recommends a TLV-time weighted average
for 2,4-dinitrotoluene of 1.5 mg/m  with a short term expo-
sure limit of 5 mg/m .
     3.   Aquatic
          A criterion to protect freshwater life has been
drafted as 620 ug/1 for a 24-hour average not to exceed
1,400 pg/1 for 2.4-dinitrotoluene and 12 ug/1 not to exceed
27 pg/1 for 2,3-dinitrotoluene.  For marine environments
a criterion has been drafted for 2,3-dinitrotoluene as a
4.4 pg/1 as a 24-hour average not to exceed 10 pg/1.  Data
was insufficient to draft a criterion for 2,4-dinitrotoluene
for marine environments.
                              ft    ' ~ r> *
                               '  I U v **
                           72-11

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                                DINITROTOLUENE
                                  REFERENCES
American  Conference  of  Governmental  Industrial Hygienists.   1978.   TLV's:
Threshold  limit values  for  chemical substances  and physical  agents in  the
workroom environment with intended changes for 1978.

Clayton, C.C.  and C.A. Baumann.  1944.   Some  effects of diet  on the  resis-
tance of mice toward 2,4-dinitrotoluene.  Arch. Biochem.  5: 115.

Clayton, C.C.  and C.A.  Baumann.   1948.   Effect  of  fat and calories on  the
resistance of mice to  2,4-dinitrotoluene.  Arch.  Biochem.  16:  415.

Cotruvo, J.A.,  et  al.  1977.   Investigation  of mutagenic effects of  products
of ozonation reactions in water.  Ann. N.Y. Acad. Sci.  298: 124.

Ellis, H.V., III,  et  al.  1976.   Subacute toxicity of 2,4-dinitrotoluene  and
2,6-dinitrotoluene.   Toxicol.  Appl.  Pharmacol.   37:  116.   (Abstract from
15th Ann. Meet. Soc. Toxicol., March 14-18.)

Friedlander, A.  1900.  On  the clinical picture of poisoning with benzene
and toluene  derivatives  with  special  reference  to  the  so-called anilinism.
Neurol. Centrlbl.  19: 155.

Hodgson,  J.R.,  et  al.    1976.   Mutation  studies  on  2,4-dinitrotoluene.
Mutat. Res.  38: 387.  (Abstract  from the 7th Ann.  Meet. Am.  Environ. Muta-
gen. Soc., Atlanta, March 12-15.)

Key, M.M., et  al.  (eds.)  1977.  Pages  278-279  In;  Occupational diseases: A
guide to  their recognition.    U.S. Dept.  Health  Edu. Welfare.   U.S. Govern-
ment Printing Office, Washington, O.C.

Kovalenko, I.I.   1973.  Hemotoxicity of nitrotoluenes  in relation to number
and positioning of nitro groups.  Farmakol.  Toxicol. (Kiev.)  8: 137.

Lee, C.C., et  al.  1978.  Mammalian toxicity  of munition  compounds.  Phase
III: Effects of lifetime exposure.   Part  I:  2,4-dinitrotoluene.   U.S. Army
Med. Res.  Dev.  Command.   Contract No.  OAMO-17-74-C-4073.   Rep.  No.  7, Sep-
tember.

Linch, A.L.  1974.   Biological monitoring for industrial exposure to cyano-
genic  aromatic  nitro  and  amino  compounds.   Am.   Ind.  Hyg.  Assoc.  Jour.
35: 426.

McGee, L.C.,  et al.   1942.    Metabolic  distrubances in  workers  exposed  to
dinitrotoluene.  Am. Jour.  Dig. Ois.   9:  329.
                                                                      »
Mori,  M., et al.   1977.   Studies on the metabolism  and  toxicity of dinitro-
toluenes — on   excretion and  distribution of tritium-labeled  2,4-dinitroto-
luene (^H^^-ONT) in the rat.  Radioisotopes   26: 780.

-------
National Cancer Institute. '  1978.   Bioassay of 2,4-dinitrotoluene for possi-
ble  carcinogenicity.   Carcinogenesis Tech.  Rep.  Ser. No.  54.   USOHEW  (NIH)
Publ. No. 78-1360.  U.S. Government Printing Office, Washington, O.C.

Proctor, N.H.  and J.P.  Hughes.   1978.   Chemical hazards  of the workplace.
J.8. Lippincott Co., Philadelphia/Toronto.

Shils, M.S. and L.J. Goldwater.   1953.   Effect of diet on  the susceptibility
of the  rat  to  poisoning by  2,4-dinitrotoluene.   Am. med.  Assoc.  Arch. Ind.
Hyg. Occup. Med.  8: 262.

Simmon, V.F., et  al.   1977.   Munitions  wastewater treatments: does chlorina-
tion  or ozonation   of  individual  components  produce  microbial  mutagens?
Toxicol. Appl. Pharmacol.  41: 197.   (Abstract  from  the  16th Ann. Meet. Soc.
Toxicol., Toronto, Can., March 27-30.)

U.S. Army  Research  and Development Command.   1976.   Toxicity of TNT waste-
water (pink water)  to  aquatic organisms.  Final  report,  Contract DAM017-75-
C-5056.  Washington, D.C.

U.S. EPA.   1978.   In-depth  studies on  health  and environmental  impacts  of
selected water pollutants.  Contract No. 68-01-4646.

U.S. EPA.  1979.  Oinitrotoluene: Ambient Water Quality Criteria.  (Draft) ^

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                                      No. 93
         2,4-Dinitrotoluene


  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents a survey of  the  potential health
and environmental hazards from exposure  to  the  subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources   and  available  reference documents.
Because of the limitations of such sources,  this short profile
may not reflect  all  available information including all the
adverse health  and  environmental  impacts  presented  by the
subject chemical.  This  document has undergone scrutiny to
ensure its technical  accuracy.
                            93-2-

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                       SPECIAL NOTATION










U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated



2,4-dinitrotoluene and has found sufficient evidence to



indicate that this compound is carcinogenic.

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                              2.4-OINITROTOLUENE
                                    Summary

     2,4-Qinitrotoluene  induces  liver cancer and  mammary tumors in mice  and
is mutagenic  in some assay  systems.   Information on  teratogenicity was  not
located in  the  available literature.   Chronic exposure to  2,4-dinitrotoluene
induces liver damage,  jaundice, methemoglobinemia and anemia  in humans  and
animals.
     Two acute  studies,  one on  freshwater fish and  the  other on  freshwater
invertebrates, provide the only  data  of 2,4-dinitrotoluene's adverse effects
on  aquatic  organisms.    Acute  LC5Q   values were  reported  as  17,000  and
30,000 jug/I.  No marine data are available.
                                   •*f D 
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                              2,4-OINITROTOLUENE
I.   INTRODUCTION
     This  profile  is  based  on the  Ambient Water  Quality Criteria  Document
for Dinitrotoluene (U.S. EPA, I979a).
     2,4-Oinitrotoluene  (2,4-ONT) has  a  melting  point  of  71°C,  a  boiling
point  of 300°C  with decomposition,   and  a solubility  in water  of 270  mg/1
at  22°C.   It  is  readily  soluble in ether,  ethanol,  and  carbon disulfide
(U.S. EPA, 1979a).
     Production  in   1975   was  273  x   10   tons/year  for  the  2,4-   and
2,6-isomers combined  (U.S. EPA, 1979a).   2,4-Oinitrotoluene is an  ingredient
in explosives  for  commercial and military  use,  a chemical stabilizer in  the
manufacture of  smokeless powder,  an  intermediate  in  the  manufacture of  tol-
uene diisocyanates used in  the production  of urethane  polymers,  and a  raw
material for the manufacture of dye-stuffs.   Dinitrotoluenes are  relatively
stable  at  ambient  temperatures  (U.S.  EPA,  I979a).   For  additional infor-
mation  regarding the dinitrotoluenes in  general,  the  reader is referred to
the EPA/ECAO Hazard Profile on Oinitrotoluenes (U.S. EPA, 1979b).
II.  EXPOSURE
     A.  Water
         Data  on concentration levels  of 2,4-ONT  in  water  were  not avail-
able.  Dinitrotoluene  waste  products  are  dumped  into  surface water or sewage
by  industries  that  manufacture dyes,  isocyanates, polyurethanes  and muni-
tions (U.S. EPA, 1979a).
     8.  Food
         According to  the U.S. EPA  (1979a),  the  likelihood  of  2,4-dinitro-
                                                                       •
toluene  existing in  food is minimal   since  it  is  not  used as a  pesticide or
herbicide.
                                    ,n/ ~f
-------
         The  U.S.  EPA  (1979a)  has estimated  the  weighted average  biocon-
centration  factor  for 2,4-dinitrotoluene  to be  5.5  for  edible portions  of
fish and  shellfish consumed  by Americans.   This estimate  was based  on  the
octanol/water partition coefficient.
     C.  Inhalation
         Exposure  to  dinitrotoluene  by inhalation  is most Likely to  occur
occupationally  (U.S.  EPA,  1979a).   However,  pertinent   data  could  not  be
located in  the available  literature on atmospheric concentrations of  dini-
trotoluene; thus, possible human exposure cannot  be  estimated.
III. PHARMPCOKINETICS
     A.  Absorption
         The  absorption  of    u-labeled  isomers   of dinitrotoluene  after
oral administration to  rats was  essentially complete  within 24 hours, with
60  to  90  percent of  the  dose being absorbed.   The  2,4-and  3,4-isomers were
absorbed to  a greater extent  than the 3,5-  and  2,5-isomers,  which in turn
were absorbed_to a greater  extent  than the 2,3- and 2,6-isomers (Hodgson,  et
al.  1977).   From  toxicity  studies,  2,4-Oinitrotoluene is  known   to  be ab-
sorbed through the respiratory tract and skin (U.S.  EPA, I979a).
     B.  Distribution
         Tissue/plasma  ratios  of  radioactivity  after  administration   of
  C-labeled   dinitrotoluene   (DNT)  to  rats   indicated   retention  of    C
2,4-QNT in  both  liver and  kidneys  but  not in other  tissues  (Hodgson,  et al.
1977).    A   similar  experiment   with   tritium-labeled   2,4-dinitrotoluene
(rl-2,4-QNT)  in  the  rat  showed  relatively  high  amounts  of  radioactivity
remaining  in adipose  tissue,  skin,  and  liver seven days after administration
(Mori,  et  al.  1977).
                                    17* 3 0"u "

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     C.  Metabolism
         No  studies  characterizing  the  metabolism of  2,4-dinitrotoluene  in
mammals are  available.  However,  on  the  basis of a comparison  of the metab-
olism  of  2,4-dinitrotoluene and  2,4,6-trinitrotoluene  in  microbial systems,
and the metabolism of  2,4,6-trinitrotoluene  in mammals,  the U.S.  EPA (1979a)
speculated  that the metabolites  of  2,4-dinitrotoluene  in mammals  would  be
either toxic and/or carcinogenic.
     D.  Excretion
                                                        14
         In  studies  involving  oral  administration of   C-dinitrotoluene  or
 H-2,4-dinitrotoluene  to  rats  (Hodgson,  et  al.  1977;   Mori,  et  al,  1977),
elimination  of  radioactivity  occurred mainly in urine and  feces.   No radio-
activity was recovered in  the  expired air.   About 46 percent  of the admin-
istered dose in the  latter  study  was excreted in the feces and urine during
the seven days following administration.
IV.  EFFECTS
     A.  Carcinogenic!ty
         2,4-Dinitrotoluene  fed  to  rats  and mice  for two  years  produced
dose-related  increases in  fibromas  of  the  skin  in male rats  and  fibro-
adenomas of  the mammary gland in  female  rats.   These  tumors were  benign.   No
statistically significant  reponse was noted  in  mice  (Natl.  Cancer  Inst.,
1978).
         In  a  second bioassay  of rats and  mice  fed 2,4-dinitrotoluene  for
two years, the  findings  in rats  included  a  significant increase  of  hepato-
                                                   *
cellular carcinomas and neoplastic nodules in the livers of females,  a sig-
nificant increase of mammary  gland  tumors in  females,  and a suspicipus in-
crease  of  hepatocellular  carcinomas  of the  liver  in  males.    Mice  had a
highly significant increase of kidney tumors  in males  (Lee,  et al.   1978).
                                     93-7

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     8.  Mutagenicity
         2,4-Oinitrotoluene  was mutagenic  in  the dominant  lethal assay and
in  Salmonella  typhimurium strain TA  1535 (Hodgson, et  al.  1976).  Cultures
of lymphocytes and kidney  cells  derived from rats fed 2,4-dinitrotoluene had
significant increases  in  the frequency of  chromatid gaps but  not in trans-
locations or chromatid breaks (Hodgson, et al. 1976).
         The mutagenic effects of  products  from  ozonation or chlorination of
2,4-dinitrotoluene  and  other dinitrotoluenes  were  negative  in  one  study
(Simmon, et al.  1977)  and, of products from  ozonation  alone,  were ambiguous
in another study (Cotruvo, et al. 1977).
     C.  Teratogenicity and Other Reproductive Effects
         Pertinent data could not be located in the available literature.
     0.  Chronic Toxicity
         Chronic exposure  to  2,4-dinitrotoluene may  produce  liver  damage,
jaundice, methemoglobinemia  and reversible  anemia  with  reticulocytosis  in
humans and animals (Linen, 1974; Key,  et  al.  1977;  Proctor and Hughes,  1978;
Kovalenko., 1973).
     E.  Other Relevant Information
         Animals were  more  resistant  to the.  toxic effects of 2,4-dinitro-
toluene administered in the  diet when given diets  high in  fat (Clayton  and
Baumann,  1944,   1948;  Shils  and  Goldwater,   1953)  or  protein  (Shils  and
Goldwater, 1953).
     Alcohol has a synergistic  effect on the toxicity  of  2,4-dinitrotoluene
(Friedlander,  1900; McGee,  et al. 1942).
                                73 -8

-------
     In subacute  studies  (13  weeks)  of several species,  1,2,4-dinitrotoluene



caused methemoglobinemia, anemia with  reliculocytasis, gliosis,  and demyeli-



nation in  the  brain,  and atrophy with aspermatogenesis of  the  testes (Ellis



et al., 1976).



V.   AQUATIC TOXICITY



     A.  Acute Toxicity



         The  only toxicity data  available for  the  effects of  2,4-dinitro-



toluene in aquatic animals are  from a single  freshwater fish  and  inverte-



brate  species  (U.S.  Army, 1976).  A 96-hour  static  LCcn value  for  the  fat-
                                                       j\j


head minnow (Pimephales promelas) was  reported as 31,000 pg/1 and  a 48-hour



static  LC5Q  value  for  the  cladoceran,  Daohnia  magna,  was  reported  as



35,000 ;jg/l.



     8.  Chronic Toxicity and Plant  Effects



         Pertinent data could not be located in the available literature.



     C.  Residues



         A bioconcentration factor of 19 was obtained for  2,4-dinitrotoluene.



VI.  EXISTING GUIDELINES AND STANDARDS



     Neither  the  human  health  nor  aquatic  criteria derived   by  U.S.   EPA



(1979a),  which are summarized below, have gone through the  process  of public



review; therefore, there is a possibility  that these  criteria may be changed.



     A.  Human



         Based on the  induction  of fibroadenomas of the  mammary  gland  in



female rats (Lee,  et al.  1978), and  using  the "one-hit"  model,   the  U.S.  EPA

                                                    s

(1979a) has  estimated  levels  of 2,4-dinitrotoluene   in ambient  water which



will result in specified risk levels  of human  cancer:
                                    , ~| O> -
                                   fV ) l




                                     73-?

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Exposure Assumptions                 Risk Levels  and  Corresponding Criteria
     (per day)
                                      0       1C.-7        10-*         10-5
Consumption of 2 liters of drink-          7.4 ng/1     74.0  ng/1    740 ng/1
ing water and 18.7 grams fish and
shellfish.
Consumption of fish and shellfish            .156 jjg/1   1.56 jjg/1    15.6/jg/l
only.
         The  American  Conference   of  Governmental  Industrial   Hygienists
(1978) recommends  a TLV-time-weighted average  for 2,4-dinitrotoluene of  1.5
mg/m  with a short term exposure limit of 5 mg/m  .
     8.  Aquatic
         A  criterion  has  been drafted  for  protecting  freshwater life  from
the toxic effects of 2,4-dinitrotoluene.   A  24-hour average concentration  of
620 jug/1, • not to  exceed  1,400 ;jg/l,  has  been proposed.   Data are insuffi-
cient for drafting a marine criterion.

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                      2,4-DINITROTOLUENE

                         REFERENCES

American Conference of Governmental Industrial  Hygienists.
1978.  TLV'sR: Threshold limit values  for  chemical
substances and physical agents in  the  workroom  environment
with intended changes for 1978.

Clayton, C.C., and C.A. Baumann.   1944.  Some effects of  diet
on the resistance of  mice toward 2,4-dinitrotoluene.  Arch.
Biochem.  5: 115.

Clayton, C.C., and C.A. Baumann.   1948.  Effect of--fa-fe-and	
calories on the resistance of mice to  2,4-dinitrotoluene.
Arch. Biochem.  16: 415.

Cotruvo,, J.A., et al.  1977.  Investigation of mutagenic
effects of products of ozonation reactions  in water.  Ann.
N.Y.  Acad. Sci.  298: 124.

Friedlander, A.  1900.  On the clinical picture of poisoning
with benzene and toluene derivatives with  special reference
to the so-called anilinism.  Neurol. Centrlbl.  19: 155.

Hodgson, J.R., et al.  1976.  Mutation studies  on 2,4-dini-
trotoluene.  Mutat. Res.  38: 387.  (Abstract from the  7th
Annu. Meet. Am. Environ. Mutagen Soc., Atlanta, March 12-15).

Hodgson, J.R., et al.  1977.  Comparative  absorption, distri-
bution, excretion, and metabolism of 2,4,6-trinitroluene
(TNT) and isomers of dinitrotoluene (DNT)  in rats.  Fed.
Proc.  36: 996.

Key, M.M., et al. (eds.)  1977.  Pages 278-279  Ijn:
Occupational diseases: A guide to  their recognition.  U.S.
Dept. Health, Edu. Welfare.  U.S. Government Printing Office,
Washington, D.C.

Kovalenko, I.I.  1973.  Hemotoxicity of nitrotoluene in rela-
tion to number and positioning of nitro groups.  Farmakol.
Toxicol. (Kiev.) 8: 137.

Lee, C.C., et al.  1978.  Mammalian toxicity of munition  com-
pounds.  Phase III: Effects of life-time exposure.  Part  I:
2,4-Dinitrotolune.  U.S. Army Med. Res. Dev. Command.  Con-
tract No. DAMD-17-74-C-4073.  Rep. No. 7,  September.

-------
Linch, A.L.  1974.  Biological monitoring  for  industrial  ex-
posure to cyanogenic aromatic nitro and amino  compounds.   Am.
Ind. Hyg. Assoc. Jour.  35: 426.

McGee, L.C., et al.  1942.  Metabolic disturbances  in workers
exposed to dinitrotoluene.  Am. Jour. Dig. Dis.  9:  329.

Mori, M., et al.  1977.  Studies on the metabolism  and  tox ic-
ity of dinitrotoluenes — on excretion and distribution of
tritium-labelled 2,4-dinitrotoluene (3H-2,4-DNT) in  the
rat.  Radioisotopes  26: 780.

National Cancer Institute.  1978.  Bioassay of 2,4-dinitro-
toluene for possible carcinogenicity.  Carcinogenesis Tech.
Rep- Ser. No. 54.  U.S. DHEW (NIH) Pufal. No. 78-1360.   U.S.
Government Printing Office, Washington, D.C.

Proctor, N.H., and J.P. Hughes.  1978.  Chemical hazards  of
the workplace.  J.B. Lippincott Co., Philadelphia/Toronto.

Shils, M.E., and L.J. Goldwater.  1953.  Effect of diet on
the susceptibility of the rat to poisoning by  2,4-dinitro-
toluene.  Am. Med. Assoc. Arch. Ind. Hyg. Occup. Med. 8:
262..

Simmon, V.F., et al.  1977.  Munitions wastewater treatments:
dose chlorination or ozonation of individual components pro-
duce microbial mutagens?  Toxicol. Appl. Pharmacol.  41:  197.
(Abstract from the 16th Annu. Meet. Soc. Toxicol.,  Toronto,
Can., March 27-30) .

U.S. Army Research and Development Command.  1976.   Toxicity
of TNT wastewater (pink water) to aquatic organisms.  Final
Report, Contract DAMD 17-75-C-5056.  Washington, D.C.

Q.S. EPA.  1979a.  Dinitrotoluene: Ambient Water Quality  Cri-
teria. (Draft).

U.S. EPA.  1979b.  Dinitrotoluene: Hazard Profile.   Environ-
mental Criteria and Assessment Offica.

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                                      No.  94
         2,6-Dinitrotoluene


  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental  impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny  to
ensure its technical accuracy.
                        9Y-.

-------
                                                             Ill
                        2,6-Dinitrotoluene




SUMMARY


     2,6-Dinitrotoluene  is known to cause methemoglobinemia in


cats, dogs, rats, and mice.  When administered orally to these


animals for a maximum of thirteen weeks, the major effects seen


in addition to the blood effects were depressed spermatogenesis,


degeneration of the liver, bile duct hyperplasia, incoordination


and rigid paralysis of the hind legs, and kidney degeneration.


     Positive results were obtained with mutagenicity testing in


a number of Salmonella typhimurium strains.


     2,6-DNT has been found in tap water in the United States.


The nitro groups on the aromatic ring retard degeneration so


there is a potential for it to accumulate in the aquatic environ-


ment.




I.   INTRODUCTION


     This profile is based on the Ambient Water Quality Criteria


Document for Dinitrotoluene (U.S. EPA, 1979b) and a U.S. EPA


report entitled "Investigation of Selected Potential Environ-


mental Contaminants:  Nitroaromatics" (1976).


     2,6-Dinitrotoluene  (2,6-DNT; CyHgNjO^; molecular weight


182.14) is a solid at room temperature.   It is in the shape of


rhombic needles and is soluble in ethanol.  Its melting point is
                                                            »

66°C and its density is 1.28 at lll'C (Weast, 1975).


     A review of the production range (includes importation)


statistics for 2,6-dinitrotoluene (CAS.  No. 606-20-2) which is

-------
listed in the  initial TSCA  Inventory  (1979a)  has  shown that

between 50,000,000 and  100,000,000  pounds  of  this chemical were

produced/imported in 1977._/

     Mixtures  of the dinitrotoluene isomers are  intermediates in

the manufacture of toluene  diisocyanates,  toluene diamines and

trinitrotoluene  (Wiseman, 1972).  Dinitrotoluene  (both 2,4- and

2,6-) is an ingredient  in explosives  for commercial  and military

use and is also used as  a chemical  stabilizer in  the manufacture

of smokeless powder (U.S. EPA,  1979b).



II.  EXPOSURE

     A.   Environmental  Fate

     Based on  the photodecomposition  of trinitrotoluene (TNT)

described by Burlinson  et al.  (1973), 2,6-dinitrotoluene would be

expected to react photochemically.  Decomposition of 65% of the

TNT had occurred when the decomposition products  were  examined.

     2,6-Dinitrotoluene  would  be expected  to  biodegrade to a

limited extent.  The nitro  groups retard biodegradation and

studies with soil microflora have shown that  mono- and di-

substituted nitrobenzenes persist for more than 64 days

(Alexander and Lustigmann,  1966).   McCormick _et_ _al_.  (1976)  and

Bringmann and Kuehn (1971)  reported microbial  degradation of

2,6-DNT by anaerobic and aerobic bacteria, respectively.


—/ This production range information does  not  include  any   ,
   production/importation data  claimed as  confidential by the
   person(s) reporting  for  the TSCA inventory, nor does it
   include any information  which would compromise Confidential
   Business Information.  The  data  submitted  for  the TSCA
   Inventory,  including production  range information,  are subject
   to the limitations contained'in  the Inventory  Reporting

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     B.   Bioconcentration



     In general nitroaromatic compounds do not have high biocon-



centration potential based on calculations using their octanol-



water partition coefficients.  They are not expected to



biomagnify based on their water solubility (U.S. EPA, 1976).



     C.   Environmental Occurrence



     2,6-Dinitrotoluene has been identified in tap water in the



United States (Kopfler and Melton, 1975).  Its environmental con-



tamination would come almost exclusively from the chemical plants



where it is produced.  It was detected in the water effluent from



a TNT plant in Radford, Virginia at concentrations of 3.39 to



56.3 ppm.  It was also found in the raw waste of a DNT plant at



150 ppm.  The raw effluent contained 0.68 ppm and the pond efflu-



ent 0.02 ppm (U.S. EPA, 1976).








III. PHARMACOKINETICS



     2,6-Dinitrotoluene can enter the body through inhalation of



vapors or dust particles,  ingestion of contaminated food,  and



absorption through the skin (EPA, 1979b) .  Hodgson _et_ _al_.  (1977)



traced the pathway of * C labeled di- and tri-substituted  nitro-



toluenes after oral administration of the compounds to rats.  All



of the compounds were well absorbed with 60 to 90% absorption



after 24 hours.   The radiolabel was found in the liver,  kidneys



and blood but not in other organs; none was found in the expired



air indicating that the aromatic ring was not broken down  through



metabolism of the compounds.  Most of the labeled compounds were






   Regulations (40 CFR 710).
                               W-5

-------
eliminated in the urine as metabolites; biliary  excretion was



also an important elimination pathway.








IV.  HEALTH EFFECTS



     A.   Carcinogenicity



     No carcinogenicity testing of 2,6-DNT has been  reported  in



the literature.  The National Cancer  Institute conducted  a bio-



assay to determine the carcinogenicity of 2,4-DNT by administer-



ing it to rats and mice in their diet.  2,4-DNT  induced benign



tumors in male and female rats, however, the benign  tumors were



not considered a sufficient basis for establishing carcinogen-



icity.  The assay produced no evidence of carcinogenicity of  the



compound in mice (NCI, 1978).



     B.   Mutagenicity



     Simmon _et^ _al_. (1977) tested 2,6-dinitrotoluene  for



mutagenicity in Salmonella typhimurium.  Positive results were



obtained with strains TA1537, TA1538, TA98, and  TA100, but not



TA1535.  These results were obtained without metabolic activa-



tion.



     C.   Other Toxicity



          1.   Chronic



     The subchronic toxicity of 2,6-dinitrotoluene was determined



by oral administration to dogs, rats, and mice for about  13



weeks.  The primary effects were on red blood cells,  the  nervous



system, and the testes.  Both dogs and rats had  decreased mu^cu-



lar coordination primarily in the hind legs, rigidity in  exten-



sion of the hind legs, decreased appetite,  and weight loss.   The

-------
mice experienced only the decreased appetite and weight  loss.



All of the animals had methemoglobinemia, and anemia with reticu-



locytosis.  The tissue lesions seen were extramedullary  hemato-



poeisis in the spleen and liver, gliosis and demyelination in the



brain, and atrophy with aspermatogenesis in the testes (Ellis et



al.,  1976>.  Methemoglobinemia was also found in cats adminis-



tered 2,6-DNT (U.S. EPA, 1979b).



          2.   Acute



     Oral LDSO's have'been reported for rats and mice.   They are



180 mg/kg and 1,000 mg/kg respectively (Vernot et al., 1977).  A



mixture of 2,4-DNT and 2,6-DNT was applied to the skin of rabbits



in a series of 10 doses over a two week period and no cumulative



toxicity was found (U.S. EPA, 1976).








VI.  EXISTING GUIDELINES



     The OSHA standard for 2,6-DNT in air is a time-weighted



average of 1.5 mg/m3 (39 PR 23540).

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                           BIBLIOGRAPHY

Alexander, M. and B.K. Lustigmann.  Effect of chemical  structure
on microbial degradation of substituted benzenes.   J. Aor.  Food.
Chem. 14(4), 410-41, 1966.  (As cited in U.S. EPA,  1976).

Bringmann, G. and R. Kuehn.  Biological decomposition of  nitro-
toluenes and nitrobenzenes by Azotobacter Agilis.   Gesundh.-Ing.,
92(9), 273-276, 1971.  (As cited  in U.S. EPA, 1976).

Burlinson, N.E. _et_ jal^.  Photochemistry of TNT:  investigation  of
the "pink water" problem.  U.S. NTIS AD 769-670,  1973.   (As cited
in U.S. EPA, 1976).

Ellis, H.V., III j|t_ ^1^.  Subacute toxicity of 2,4-dinitrotoluene
and 2,6-dinitrotoluene.  Toxicol. Appl. Pharm.  37," 116, -1976.

Hodgson, J.R. et al.  Comparative absorption, distribution,
excretion, and metabolism of 2,4,6-trinitrotoluene  (TNT)  and
isomers of dinitrotoluene  (DNT) in rats.  Fed.  Proc. 36,  996,
1977.

Kopfler, F.C. and R.G. Melton.  1977.  Human exposure to  water
pollutants.  In Advances in Environmental Science  and Technology,
Vol. 3.  Fate of Pollutants in the Air and Water  Environments.
Part 2.  Chemical and Biological Fate of Pollutants in  the
Environment.  Symposium at the 165th National American  Chemical .
Society Meeting in  the Environmental Chemistry  Division.  Phila-
delphia, PA.  April 1975.  John Wiley and Sons, Inc., New York.

McCormick, N.G. _et_ _al_.  Microbial transformation  of 2,4,6-trini-
trotoluene and other nitroaromatic compounds.   Appl. Environ.
Microbiol. 31(6), 949-958, 1976.

National Cancer Institute.  Bioassay of 2,4-dinitrotoluene  for
possible carcinogenicity.  PB-280-990, 1978.

National Institute  of Occupational Safety and Health.   Registry
of Toxic Effects of Chemical Substances, 1978.

Simmon, V.F. et al.  Mutagenic activity of chemicals identified
in drinking water.  Dev. Toxicol. Environ. Sci. 2,  249-258,  1977.

U.S. EPA.  Investigation of Selected Potential  Environmental
Contaminants:  Nitroaromatics.  PB-275-078, 1-976.

U.S. EPA.  Toxic Substances Control Act Chemical  Substance
Inventory, Production Statistics  for Chemicals  on  the Non-Confi-
dential Initial TSCA Inventory, 1979a.

U.S. EPA.  Ambient  Water Quality Criteria:  Dinitrotoluene.
PB-296-794, 1979b.

-------
Vernot, E.H. et_ _al^.  Acute toxicity  and' skin  corrosion data for
some organic and inorganic compounds  and  aqueous  solutions.
Toxicoi. Appl. Pharmacol. 42(2),  417-424,  1977.

Weast, R.C., ed. 1978.  CRC Handbook  of  Chemistry and Physics.
CRC Press, Inc., Cleveland, Ohio.

Wiseman, P.  1972.  An Introduction  to  Industrial Organic
Chemistry.  Interscience Publishers,  John-Wiley and Sons,  Inc.,
New York.
                                y
                             -11 * 3-
                             *f r o J

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                                     No.  95
        Di-n-octyl Phthalate




  Health and Environmental  Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY

       WASHINGTON,  D.C,   20460



           APRIL 30,  1980
         ^^^-j±M-
          -yyw

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                          DI-n-OCTYL PHTHALATE



            ' •                   Summary







      Di-n-octyl phthalate has produced teratogenic effects following



i.p. injection, of pregnant rats.  This same study has also indicated



some increased resorptions and fetal toxicity.



      Evidence is not available indicating mutagenic or carcinogenic



effects of 'di-n-octyl phthalate.



      Data pertaining to the aquatic toxicity of di-n-octyl phthalate



is not available.
                                 . / /> z.—
                                ) )w v

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                           DI-n-OCTYL  PHTHALATE


 I.     INTRODUCTION


     .  This profile is based on the Ambient Water Quality Criteria Document


 for Phthalate Esters (U.S. EPA, 1979a).


       Di-n-octyl phthalate (OOP) is a diester of the ortho form of


 benzene dicarboxylic acid.  The compound has a molecular weight of


 391.0,  specific gravity of 0.978,  boiling point of 220°C at 5 mm Hg,


 and is insoluble in water.


       DOP is used as a plasticizer in the production of certain plastics.


       Current Production:   5.8 x 103  tons/year in 1977 (U.S.  EPA,  1979a).


       Phthalates have been detected in soil, air, and water samples; in


 animal and human tissues,  and in certain vegetation.   Evidence from in


 vitro studies indicates that certain  bacterial flora may be capable of


 metabolizing DOP to the monoester  form (Engelhardt,  et al» 1975).   For


 additional information regarding the  phthalate esters in general,  the


 reader is referred to the  EPA/ECAO Hazard Profile on Phthalate Esters


'(U.S.  EPA 1979b).



 II.    EXPOSURE


       Phthalate esters appear in all  areas of the environment.  Environmental


 release of phthalates may  occur through leaching of  the compound from


 plastics,  volatilization from plastics,  or the incineration of plastic


 items.  Sources of human exposure to phthalates include contaminated


 foods  and fish,  dermal application, and parenteral administration  by


 use  of plastic blood bags,  tubings, and infusion devices (mainly DEHP


 release).   Relevant factors in the migration of phthalate esters from


 packaging materials to food and beverages are:   temperature,  surface


 area contact,  lipoidal nature of the  food,  and length of contact (U.S.


 EPA,  I979a).
                                   / / fl s
                                 ^7 
-------
      Monitoring studies have indicated that most water  phthalate  concentrations



are in the ppm range, or 1-2 jug/liter  (U.S. EPA, I979a).   Industrial



air monitoring studies have measured air levels of  phthalates  from 1.7



to 66 mg/m3-(Milkov, at al. 1973).



      Information on levels of OOP in  foods is not  available.   Bio-



concentration factor is not available  for OOP.



III.  PHARMACOKINETICS



      Specific information could not be located on  the absorption.,




distribution, metabolism, or excretion of DOP.  The reader  is  referred



to a general coverage of phthalate metabolism (U.S. EPA,  1979b).



IV.   EFFECTS



      A.     Carcinogenicity



        Pertinent data could not be located in the  available literature.



      B.     Mutagenicity



        Pertinent data could not be located in the  available literature.








      C.     Teratogenicity



        Administration of DOP to pregnant rats by i.p. injection has



been reported to produce some teratogenic effects,  although less so



than several other phthalates tested (Singh, et al. 1972).



      D.     Other Reproductive Effects



        An increased incidence of resorption and fetal toxicity was



produced following i.p. injection of pregnant rats  with.DOP (Singh,  et



al. 1972).



      E.     Chronic Toxicity



        Pertinent data could not be located in the  available literature.

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V.    AQUATIC TOXICITY




      Pertinent data could not be located in the available literature.




VI.   EXISTING GUIDELINES AND STANDARDS




      Neither the human health nor the aquatic criteria derived by U.S.




EPA (1979a), which are summarized below, have gone through the process




of public review; therefore, there is a possibility that these criteria




will be changed.




      A.     Human



             Pertinent data concerning the acceptable daily intake




(ADI) level in humans of DOP could not be located in the available




literature.



             Recommended water quality criterion level for protection




of human health is not available for DOP.




      B.     Aquatic



             Pertinent data is not available pertaining to the aquatic




toxicity of di-n-octyl phthalate; therefore, no criterion could be




drafted.

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                             DI-N-OCTYL PHTHALATE
Engelhardtj • G., at  al.   1975,   The microbial rnetabolism of di-n-butyl  phtha-
late.  and  related  dialkyl  phthalates.   Bull.  Environ.  Contam.  Toxicol.
13: 342.

Milkov, L.E.,  at  al.   1973.   Health status of  workers exposed to phthalata
plasticizers in the manufacture  of artificial  leather and films based  on  PVC
resins.  Environ.  Health Perspect.   (Jan.): 175.

Singh,  A.R.,  st  al.   1972.   Teratogenicity  of phthalate  esters  in  rats.
Jour. Pharm.  Sci.   61:  51.

U.S. EPA.  1979a.   Phthalate Esters: Ambient Water Quality Criteria.   (Draft)

U.S. EPA.  1979b.   Environmental Criteria and  Assessment Office.  Phthalate
Esters: Hazard Profile.  (Draft)

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                                      No. 96
       1,2-Dlphenylhydrazine


  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented  by the
subject chemical.   This document has undergone  scrutiny to
ensure its technical accuracy.

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                      SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (CAG)  has evaluated
1,2-diphenylhydrazine and has found sufficient evidence to
indicate that  this compound is carcinogenic.
                            16-3

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                      1, 2-DIPHENYLHYDRAZINE
                             Summary
     The adverse  effects  of exposure to 1,2-diphenylhydrazine  in-
clude damage to both the kidney and liver.  Acute  LD5Q  values have
ranged from 300 to 960  rag/kg in experimentally dosed rats.   No data
concerning the absorption,  distribution,  or  excretion of the 1,2-
diphenylhydrazine have been generated.  Benzidine  has been  identi-
fied as  a metabolite  in  urine of  rats exposed to the chemical.
Diphenylhydrazine is carcinogenic in both sexes of  rats  and in  fe-
male mice.
     The only  aquatic  toxicity data for diphenylhydrazine  are  for
freshwater organisms.  Acute toxicity levels of 270  and  4,100 ug/3,'
were reported  for  bluegill and Daphnia magna, respectively, and  a
single chronic value of 251 jag/1 was reported  for  Daphnia magna.

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                      1,2-DIPHENYLHYDRAZINE


I.   INTRODUCTION


     This profile  is  based primarily on the Ambient Water Quality


Criteria Document for Diphenylhydrazine.


     Diphenylhydrazine  (DPH)  has a molecular  weight  of 184.24, a


melting point of 131°C and  a boiling point of 220°C. DPH is slight-


ly  soluble  in  water  and  is  very  soluble  in  benzene,  ether  and


alcohol.


     The  symmetrical isomer  of  diphenylhydrazine,  1,2-diphenyl-


hydrazine is  used  ia the synthesis of  benzidine  for  use in  dyes,


and in the synthesis of phenylbutazone, an anti-arthritic drug.


     The  reported  commercial  production of more  than  1000 pounds


annually, as  of 1977,  is  most  lively an  underestimation  of  the


total amount of diphenylhydrazine available.   Diphenylhydrazine  is


produced .in  several synthetic processes as  an intermediate and a


contaminant,  but there  is  no  way of  estimating these  substantial


quantities.


II.  EXPOSURE


     A..   Water


          The highest reported concentration of 1,2-diphenylhydra-


zine in drinking water is one ug/1  (U.S. EPA,  1975).


     B.   Food


          The U.S.  EPA  (1979)  has  estimated  the  weighted average


bioconcentration factor  for  diphenylhydrazine to  be   29  for  the
                                             ,-

edible portions of fish and shellfish consumed by Americans.  This


estimate  is  based  on the  octanol/water partition  coefficient of


diphenylhydrazine.

-------
     C.   Inhalation
          Pertinent  data  could  not  be  located in  the available
literature.
III. PHARMACOKINETICS
     Pertinent information  could not be located  in the available
literature regarding absorption, distribution and excretion.
     A.   Metabolism
         - Various metabolites, including the known carcinogen ben-
zidine, -have been identified in  the urine of rats.  1,2-Diphenylhy-
drazine was administered orally  (200,400 mg/kg), intraperitoneally
(200 mg/kg),  intratracheally (5r10  mg/kg)  and  intravenously (4,8
mg/kg).  The metabolites detected were not dependent upon the base
or route of administration  (Williams, 1959).                       4
IV.  EFFECTS
     A.   Carcinogenicity
          Diphenylhydrazine  has  been   identified   as  producing
significant  increases  in  hepatocellular carcinoma  at  5 ug/kg/day
and 18.3 ug/kg/day in both sexes of rats; Zymbal's gland squamous-
cell  tumors  in  male  rats   at  13.8  ug/kg/day; neoplastic  liver
nodules  in  female   rates  at  7.5 ug/kg/day;   and  hepatocellular
carcinomas in female mice  at 3.75  ug/kg/day (NCI, 1978).  Diphenyl-
hydrazine was not carcinogenic in male mice.
     B.   Mutagenicity
          No microbial mutagenetic assays witir'or without metabolic
activation have been conducted on diphenylhydrazine.  An intraperi-
toneal dose of 100 mg/kg had an  inhibitory effect on the incorpora-
tion of  (  H)-thymidine into testicular  DNA  of  experimental mice
(Sieler, 1977).
                             76-7

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     C.   Teratogenicity
          Pertinent information could not be located  in the avail-
able literature.
     D.   Toxicity
          One study reported an LDcg of  959  mg/kg for  male  rats ad-
ministered DPH  as a  five  percent  solution.   In  the Registry of
Toxic Effects of Chemical Substances,  the  oral  LD5Q  is listed as
301 mg/kg.  Neoplasms resulted in rats  after 52  weeks with a  total
dose  of  16- g/kg  DPH  administered  subcutaneously.    In  2  mice
studies, neoplasms resulted after  25 weeks with topical application
of  5280  mg/kg  and after  38 weeks  with subcutaneous  injection of
8400 mg/kg DPH.   Liver and kidney  damage have  been implicated in
the adverse effects  of  diphenylhydrazine chronically administered.
to rats.  No experimental or epidemiological studies have been con-
ducted on the effects of diphenylhydrazine  in humans.
V.   AQUATIC TOXICITY
     A.   Acute
          Ninety-six-hour  LCen  values   for freshwater  organisms
have been  reported as 270 pg/1  for the bluegill,  Lepomis macro-
chirus,  and  the 48-hour LCcg  for  the  cladoceran,  Daphnia magna,
is  4,100 ug/1  (U.S.  EPA,  1978).    No   toxicity data  for marine
animals could be located in the available literature.
B.   Chronic
          A chronic  value  of  251  ^ag/1  has  been obtained  for the
freshwater cladoceran, Daphnia Magna  (U.S. EPA,  1978).  No chronic
                                                              »
tests of diphenylhydrazine ace available for marine organisms.

-------
     C.   Plants
          Pertinent  data could  not be  located  in the  available
literature.
     0.   Residues
          Based on  the octanol/water partition coefficient  of  870
for  1,2-diphenylhydrazine,  a  bioconcentration factor  of 100  has
been estimated for aquatic organisms with a lipid content of 8 per-
cent.
VI.  EXISTING GUIDELINES AND STANDARDS
     Neither  .the  human  health  nor aquatic  criteria  derived  by
U.S.  EPA  (1979),  which  are   summarized  below have  gone  through
the  process  of public  review; the'refore,  there  is a  possibility
that these criteria may be changed.
     A.   Humans
          *
          No standards were found for humans exposed to  diphenylhy-
drazine in occupational or ambient  settings.
          Recommended  draft  criteria for  the protection of  human
health are as follows:
Exposure Assumptions        Risk Levels and Corresponding Criteria
                            0  10f7         lOf6         10_~5
2 liters of drinking water  0  4     ng/1   40    ng/1   400   ng/1
and consumption of 18.7
grams fish and shellfish  (2)
Consumption of fish and     0  .019  ug/1   0/19  ug     1.9
shellfish only.
                           ~*} / t V "

-------
     B.   Aquatic


          Criterion to protect  freshwater  aquatic life from toxic


effects of diphenylhydrazine have been drafted  as a 24-hour aver-


age concentration  of 17  ug/1 and  not  to  exceed 38 ug/1  at  any


time.
                               ./ / ft
                             *i) I • /*"

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                        DIPHENYLHYDRAZINE

                            REFERENCES
NCI Publication NO.  (NIH) 78-1342.  1978.  Bioassay of  hydrazoben-
zene for possible carcinogenicity.

Sieler, J.P.   1977.   Inhibition  of  testicular  DNA  synthesis by
chemical  rautagens  and  carcinogens..   Preliminary results  in  the
validation of a novel short term test.  Mutat. Res.  46: 305.

U.S. EPA.    1375.    Primary assessment  of  suspected   carcinogens
in drinking water.   Report to Congress.

U.S. EPA.   1978.   In-depth studies  on health  and environmental
impacts of selected water pollutants.  Contract No. 68-01-4646.

U.S. EPA.   1979.   Diphenylhydrazine:   Ambient  Water  Quality Cri-
teria.   (Draft).

Williams,  R.   1959.   Detoxication Mechanisms.   New  York:   John
Wiley and Sons. p.  480.

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                                      No. 97
             Disulfoton
  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                          Disclaimer  Notice
Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.

-------
                                  DISULFOTON
                                   Summary

     Oisulfotcn is a highly toxic organophosphorous insecticide used on many
agricultural  crops,   the  human  oral LDLQ  is  estimated  at  5 'mg/kg  body
weight.   Exposure  results  in central  nervous  system  toxicity.   The  LD5Q
for several animal species ranges from 3.2 to  6 mg/kg.   Carcinogenic, muta-
genic, and  teratogenic  studies  were not  found  in  the available literature.
The occupational threshold limit  value for disulfoton is  10 ug/m5.  Allow-
able residue tolerances for agricultural  commodities  range  from 0.3 to 11.0
ppm.
     Although disulfoton is considered toxic to aquatic  organisms, specific
studies on aquatic toxicity were not located in  the available literature.
                                   I I  *] //
                                 ') I  A. * •
                                  77-y

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 I.    INTRODUCTION
      Disulfoton  is  a  highly  toxic  organophosphorous  insecticide used  in
 agriculture  to  control mainly sucking insects such  as aphids and plantfeed-
 ing mites.   Small  amounts are used on home plants and gardens in the form of
 dry granules with  low  content of active  ingredient (U.S. EPA, 1974).  Disul-
 foton was  introduced   in 1956  by Bayer  Leverkusen  (Martin and  Worthing,
 1974), and today it  is produced by only one  U.S.  manufacturer,  Mobay  Chemi-
 cal Corporation, at  its Chemogro Agricultural Division  in  Kansas City,  Misr
 souri (Stanford  Research Institute (SRI),  1977).   An estimated  4500  tonnes
 were • produced in  1974  (SRI,  1977).   Disulfoton  is made by  interaction  of
 0,0-diethyl  hydrogen   phosphorodithioate   and  2-(2-ethylthio)ethylchloride
                                           *
 (Martin  and  Worthing,   1974).   Disulfoton  is  slightly  soluble in  water  and
 readily  soluble in  most organics.   Its   overall  degradation  constant  is
 0.02/day.  Disulfoton  has a  bioconcentration factor of  1.91  and  an octanol/
water partition coefficient of 1.0 (see Table 1).
 II.   EXPOSURE
     A.   Water
          Disulfoton concentrations  are  highest during  the  production pro-
cess.   Concentrated  liquid wastes  are barged  to  sea  (150-200  mi; 240-320
km), and sludge wastes are disposed in landfills.
          Agricultural  application  rates  normally  range  from 0.25  to  1.0
Ib/acre  (0.28-1.1  kg/ha); to a maximum of 5.0 Ib/acre  (5.5  kg/ha)  for some
uses.   Target crops  include  small grains,  sorgum,  corn, cotton,  other  field
crops; some vegetable,  fruit and nut crops; ornamentals (Fairchild,  1977).
          Disulfoton is considered stable  in groundwater.  Less than 10 per-
cent  is estimated to decompose  in  five days  (equivalent  to  50-250 mi; 80-400
                                    tjn  IT t
                                 * It X -!»'

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           TABLE 1.  PHYSICAL AND CHEMICAL PROPERTIES OF OISULFQTON
Synonyms:  0,0-Qiethyl S-(2-(ethylthio)ethyl) phosphorodithioate;
           0,0-Oiethyl S-(2-(ethylthio)ethyl) dithiophosphate;   Thiodemeton;
           Frumin;  Glebofos;  Ethylthiometon  B;   VUAgT  1964;  Di-Syston  G;
           Disipton; ENT-23437;  Ethyl  thiometon;  VUAgT 1-4; Bay 19639;  M 74
           [pesticide]; Ekatin  TO;  CAS Reg. No. 298-Q4-4;  M  74 (VAN); Bayer
           19639;  Oi-Syston; Dithiodemeton;  Oithiosystox; Solvirex;  Frumin
           AL; Frumin G

Structural Formula:

Molecular Weight:  274.4

Description:   Colorless  oil;   technical  product  is  a  dark  yellowish  oil;
               readily soluble in most organics

                                   2Q
Specific Gravity and/or Density:, d,   = 1.144

Melting and/or Boiling Points:  bp 620Q at 0.01 mm Hg

Stability:   Relatively stable to hydrolysis at pH below 8
             Overall degradation rate constant (0.02/day)

Solubility (water):  25 ppm at room temp.

                   sediment . .5
                     H20    '  1
Vapor Pressure:  1.8 x 10-4 mm Hg at 2QQC

Bioconcentration Factor (BCF) and/or
Octanol/water partition coefficient (Kow): •  KQW =  1.91
                                            BCF =1.0
Source:  Martin and Worthing,  1974;  Fairchild,  1977;  Windholz,  1976;
         U.S. EPA, 1980; Berg, et al.  1977.
                                    ,' |  ~\ f
                                    T^&*^

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km) in a  river  environment.   Decomposition  in  a lake environment is estimat-
ed to be  near 90 percent in one year (U.S. EPA 1980).     —-  ""•  -" ~~r -'  -
     B.   Food
          In a  study by  Van  Dyk  and Krause  (1978),  disulfoton was applied as
a granular  formulation at 2  g/m length in rows during cabbage  planting  (5
percent active  ingredients,  rows one meter apart,  plants 0.5  meters  apart).
The disulfoton sulphone concentration reached a maximum in 18  to  32 days and
decreased to between 0.3 and 6.4 mg/kg 52  riays after application.  The  cab-
bage residue of disulfoton at harvest  time was below  the  maximum limit  of
0.5 mg/kg.
          Disulfoton applied  at  about 1.5  kg/10 cm-ha  (hectare  slice)  per-
                                           •
sisted for the  first week, and residue levels declined slowly the  following
week.  After  one month,  only 20 percent of the amount  applied was  found.
Disulfoton  was  not  found  to  translocate   into  edible  parts  of lettuce,
onions, and carrots  (less than  5 ppb),  but was present  at  about 20 ppb  in
the root system of lettuce (Belanger and Hamilton, 1979).
     C.   Inhalation and Dermal
          Data are not available indicating the number  of people subject  to
inhalation or  dermal exposure  to  disulfoton.   The primary  human exposure
would  appear  to  occur  during production  and application.   The  U.S.  EPA
(1976)  listed  the frequency  of  illness, by  occupational groups caused by
exposure to organophosphorous pesticides.  In 1157 reported cases, most  ill-
nesses  occurred among ground  applicators  (229)  and  mixer/loaders (142);  the
lack of or refusal to  use safety equipment, was a  major factor of this  con-
tamination.   Other groups  affected  were gardeners   (101), field  workers  ex-
posed to pesticide residues (117),-nursery and greenhouse workers (75),  soil
fumigators in agriculture (29), equipment cleaners  and mechanics  (28), trac-
                                  ? 7-7

-------
 tor drivers  and irrigators  (23),  workers exposed  to pesticide drift  (22),
.pilots (crop dusters)  (17),  and flaggers  for  aerial application  (6).   Most
 illnesses were a result  of  carelessness, lack of  knowledge of the  hazards,
 and/or lack of safety  equipment,   under dry,  hot conditions, workers tended
 not to wear  protective clothing.   Such conditions  also  tended to  increase
 pesticide levels and dust on  the crops.
 III.  PHARMACOKINETICS
      A.    Absorption,  Distribution,  and  Excretion
           Pertinent data  could not  be located in  the available literature.
      B.    Metabolism
                                                    •
           Oisulfoton is metabolized  in  plants to  sulfoxide and sulfone and
 the corresponding derivatives  of the phosphorothioate  and demeton-S.   This
 is also  the  probable  route  in  animals (Martin  and  Worthing,  1974; Menzie
 1974;  Fairchild,  1977).
 IV.   EFFECTS
      A.    Carcinogenic!ty, Mutagenicity  and Teratogenicity
           Pertinent data could not be located in  the available literature.
     8.    Chronic Toxicity and Other  Relevant Information
           Oisulfoton is highly toxic to all  terrestrial and aquatic fauna.
 Human  oral  LD^  is  estimated  to  be  5 mg  disulfoton  per kilogram  body
 weight (5 mg/kg).  The symptoms  produced by  sublethal doses are  typical of
 central  and peripheral nervous-system toxicity (Gleason, et al. 1969).   The
 reported  LD^Q concentrations  for other  species  are summarized  below (Fair-
 child, 1977).
                                   -*f) Si 2 "•

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        Species            Exposure Route              LD5Q (mg/kg)
          rat                   oral                        5
          rat                   dermal                      6
          rat              intraperitoneal                  5.4
          rat              intravenous                      5.5
         mouse                  oral                        5.5
         mouse             intraperitoneal                  7
          bird        .          oral                        3.2
Rats survived  for  60 days at 0.5 mg/kg/day (Martin and Worthing 1974).  The
no-effect level  in the diet was  2 ppm for rats  and  1 ppm  for  dogs  (Fair-
child, 1977).
          In rats, single injections of 1.2 mg disulfoton per kg body weight
caused 14 percent  reductions of hippocampal norepinephrine within 3 hours of
exposure.  Norepinephrine returned to control levels within 5 days (Holt and
Hawkins, 1978).  In  female chicks  administered  with disulfoton  intraperito-
neally  (single dose  8.6 mg/kg),  the total  lipid content of  the  sciatic
nerve,  kidney and  skeletal muscles  increased  whereas  that  of the  brain and
spinal cord remained  the  same  or  decreased.   When female chicks were orally
administered with disulfoton  (0.29 mg/kg  daily for 71  days), the total lipid
content in  all the  organs  except  the  liver and  sciatic  nerves  decreased.
Although degenerative changes  were indicated in  both exposure  studies,  no
adverse, effect on the growth  of chicks was noted (Gopel and Ahuja,  1979).
          Disulfoton applied at 1  to 1.5 kg/ha very  markedly decreased the
populations of soil bacteria  (Tiwari, et al. 1977). "
V.   AQUATIC TOXICITY
          The  96-hour  Tl_m   (equivalent  to  a  96-hour  1X50)   for  fathead
minnows was found  to  be  2.6  mg/1  in hard water and 3.7 mg/1 in  soft water.
                                   77-?

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Both  tests were conducted  at 25°C.   The corresponding  value for bluegilis
is estimated to be Q.07 mg/1 (McKee and Wolf, 1963).
VI.  EXISTING GUIDELINES AND STANDARDS
     A.   Human
          The  occupational  threshold  limit  value  for  air has  been estab-
lished  as  100 Ajg/m'.   Established residue  tolerance  for  crops  range  from
0.3 to 12.0 ppm; 0.75 ppm for most (Fairchild, 1977).
     B.   Aquatic
          Pertinent data could not be located in the available literature.

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                                  REFERENCES
Belanger, A. and H.A. Hamilton.   1979.   Determination of disulfpton and per-
methrin  residues  in an  organic soil and  their translocation  into lettuce,
onion and carrot.  Jour. Environ. Sci. Health.  814: 213.

Berg, G.L.,  et  al.  (ed.)  1977.  Farm  Chemicals  Handbook.   Meister Publish-
ing Company, Willoughby, Ohio.

Fairchild,  E.J.,   (ed.)   1977.   Agricultural chemicals  and pesticides:   A
subfile of  the  NIOSH  registry of toxic  effects of chemical substances., U.S.
Dept. of HEW, July.

Gleason,  M.N.,  et  al.   1969.   Clinical Toxicology of  Commercial Products.
Acute Poisoning, 3rd ed.

Gopal,  P.K.  .and S.P.  Ahuja.   1979.  LLpid  and  growth changes  in organs of
chicks Gallus domesticus during  acute and  chronic toxicity with disyston and
folithion.

Holt, T.M.  and  R.K. Hawkins.   1978.  Rat hippocompel norepinephrine response
to cholinesterase inhibition.  Res. Commun. Chem..  Pathol. Pharmacol  20: 239.

Martin and Worthing, (ed.)  1974.   Pesticide Manual, 4th ed.  p. 225

McKee,  J.E.  and H.W.  Wolf.   1963.   Water  Quality Criteria.  2nd  ed.   Cali-
fornia State Water Quality Control  Board.  Publication 3-A.

Menzie, C.M.  1974.  Metabolism  of  Pesticides:  An Update.   U.S. Dept.  of the
Interior Special Scientific Report  — Wildlife No. 184, Washington, D.C.

Stanford Research  Institute.   1977.  Directory  of Chemical Producers.   Menlo
Park, California.

Tiwari,  J.K.,  et al.    1977.   Effects of  insecticides  on  microbial flora of
groundnut field soil.   Ind. Jour. Micro.   17: 208.

U.S.  EPA.   1974.   Production,  Distribution, Use, and  Environmental  Impact
Potential of  Selected  Pesticides.   Report No. EPA 540/1-74-001.   U.S. Envi-
ronmental Protection Agency,  Office of  Water and Hazardous Materials,  Office
of Pesticide Programs.

U.S.  EPA.   1976.   Organophosphate  Exposure from Agricultural Usage, EPA 600/
1-76-025.

U.S.  EPA.   1980.   Aquatic Fate and Transport Estimates for Hazardous Chemi-
cal  Exposure  Assessments.   Environmental  Research Laboratory,  Athens^ Geor-
gia.

Van  Dyk,  L.P. and  M.  Krause  1978.  Persistence  and efficacy  of disulfoton
on Cabbages.  Phytophylactica  10:  53.

Windholz, M.,  (ed.) 1976.   The Merck  Index, 9th ed.  Merck and  Co., Inc.,
Rahway, New Jersey.

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                                      No.  98
             Endosulfan


  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, B.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.
                            L^^^^^,
                          • 4) J->

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                                  ENOOSULFAN
                                    Summary
     Endosulfan is  an  insecticide and is  a  member of the  organochlorocyclo-
diene insecticides.  Endosulfan  does  not appear to be carcinogenic,  mutagen-
ic or teratogenic.   In humansr chronic  toxic  effects have not been  observed
when endosulfan has been  properly  handled occupationally.  Chronic  feeding
of endosulfan to  rats  and  mice produced kidney damage, parathyroid hyperpla-
sia, testicular atrophy, hydropic change of the liver, and lowered survival.
Oral administration  of endosulfan to  pregnant rats increased fetal mortality
and resorptions.  Sterility  can  be induced in  embryos in sprayed bird  eggs.
At  very  high levels of acute exposure,  endosulfan is toxic  to the  central
nervous system.   The U.S.  EPA has  calculated an  ADI  of 0.28 mg  based on  s
NOAEL of 0.4 mg/kg  for mice  in a chronic  feeding  study.  The ADI established ;
by the food  and  Agricultural Organization (1975)  and World Health Organiza-
tion is 0.0075 mg/kg.
     Ninety-six  hour  LC5Q  values  ranged  from 0.3  to  11.0  ug/1  for  five
freshwater fish; from 0.09 to 0.6 ug/1  for five saltwater fish in 48- or  96-
hour  tests;  from 0.04  to  380  ug/1  (EC50  and  t-C5Q)  for  seven  saltwater
invertebrate species;  and from  62  to  166 pg/1  for Daphnia  maqna  (48-hour
LC3Q).  In  the  only chronic aquatic  study involving  endosulfan,  no  adverse
effects on fathead minnows were observed at 0.20 jug/1.

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I.   INTRODUCTION
     Endosulfan        (6,7,8,9,10,10-hexachloro-l,5,5a,6,9,9a-hexahydro-6,9-
methano-2,4,3-benzodioxathiepin-3-oxide;         CgClgHgO,S;         molecular
weight  406.95)  is a light  to dark brown  crystalline  solid  with a terpene-
like odor.  Endosulfan  is a broad  spectrum insecticide of the group of poly-
cyclic chlorinated hydrocarbons  called  cyclodiene insecticides.   It also has
uses as an  acaricite.   It  has  a  vapor pressure of  9 x  10"  mm Hg  at 80
degrees centigrade.  It exhibits a solubility  in water of 60 to 150 pg/1 and
is readily  soluble in organic solvents  (U..S. EPA, 1979).  The trade names of
endosulfan  include Beosit,  Chlorithiepin, Cyclodan,  Insectophene, Kop-Thio-
dan, Malix, Thifor, Thisnuml, Thioden, and. Thionex (Berg, 1976).
     Technical grade endosulfan  has a purity  of 95 percent  and is composed
of a mixture  of two  stereoisomers. referred  to as alpha-endosulfan and beta-  4
endosulfan  or  I  and  II.  These  isomers are present in- a ratio  of 70 parts
alpha-endosulfan  to  30  parts beta-endosulfan.   Impurities  consist mainly of
the degradation  products  and may not exceed 2  percent  endosulfandiol  and 1
percent endosulfan ether (U.S. EPA, 1979).
     Production:  three million pounds in 1974 (U.S~ EPA, 1979).
     Endosulfan  is  presently on the Environmental  Protection  Agency's  re-
stricted list.   However,  significant commercial use  for  insect  control on
vegetables, fruits, and tobacco continues (U.S. EPA,  1979).
     Endosulfan  is stable to  sunlight  but  is susceptible to oxidation  and
the formation  of endosulfan  sulfate  in the presence  of  growing vegetation
(Cassil and Drummond, 1965).  Endosulfan is  readily adsorbed  and absorbed by
sediments  (U.S.  EPA, 1979).   It is  metabolically  converted  by microorgan-
isms,  plants, and animals  to endosulfan sulfate, endosulfandiol,  endosulfan
ether,  endosulfan hydroxyether and endosulfan  lactone  (Martens,  1976;  Chopra

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 and Mahfouz,  1977; Gorbach, et al. 1968; Miles  and  May,  1979)..  The end-pro-
 duct,  endosulfan  lactone,  disappears quickly once  formed.   Accumulation  of
 endosulfan sulfate may  be favored in  acidic soils (Miles and Moy, 1979).
 II.  EXPOSURE  '
     A.   Water
          Endosulfan has  been detected  in  water samples  from  some of  the
 streams,  rivers,  and lakes in the United  States and  Canada and  in Ontario
 municipal water supplies.  The maximum concentration  of  endosulfan monitored
"in'municipal  water was  0.083 jug/1,,  which was  found  in  Ontario- municipal
 water  samples but 68 jjg/1 has been measured in  irrigation  run-off (U.S. EPA,
 1979).   Endosulfan contamination of  water  results from  agricultural runoff,
 industrial effluents,  and spills.   One  serious  accidental  industrial  dis-
 charge in  Germany in  1969 caused a massive fishkill  in  the Rhine  River.
 Most of  the  river  water samples contained less than 500 ng/1  endosulfan.
 Residues  in run-off  water from  sprayed  fields  can be as  high  as  220 jug/1
 (U.S.  EPA,  1979).
     B.   Food
          An average  daily intake (ADI)  less  than  or equal  to  0.001 mg  of
 endosulfan  and endosulfan sulfate was estimated for 1965-1970 from  the mar-
 ket  basket study  of  the  FDA  (Duggan and Comeliussen, 1972).  The  U.S. EPA
 (1979) has estimated the  weighted average  bioconcentration factor for  sndo-
 sulfan to be  28 for the  edible  portions  of fish and shellfish consumed .by
 Americans.  This estimate is based on measured steady-state- bioconcentration
 studies  with  mussels.  The  processing of  leafy  vegetables causes  endosulfan
 residues  to decline from  11 jug/kg to  6 pg/kg (Comeliussen,  1970).
                                      7

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     C.   Inhalation
          In  1970,  air samples from 16 states showed  an  average level of 13.0
ng/m   alpha-endosulfan  and  0.2  ng/m   beta-endosulfan.   None of  the  air
samples collected  in  1971 or 1972 contained detectable  levels  of either iso-
mer  (Lee, 1976).   Endosulfan  residues  (endosulfan  and endosulfan  sulfate)
have  been detected in  most types of  U.S.  tobacco products  in  recent  years
(U.S.  EPA,  1979).  Average  residue  levels  range  from 0.12  mg/kg to  0.83
mg/kg  for 1971-1973 (Domanski, et al.  1973,1974;  Oorough and  Gibson,  1972)..
The  extent to  which  endosulfan  residues  in tobacco products  contribute  to
human  exposure  is not  known.   Spray operators  can  be exposed  up  to  50
jug/hour  of  endosulfan  from  a  usual application of  a 0.08  percent  spray
(Wolfe,  et  al. 1972).    Non-target  deposition  on  untreated  plants  after-
spraying may lead  to  residues  of  up to 679 ug/kg (Keil,  1972).
     0.  Dermal
         Wolfe,  et al.  (1972)  estimated  that sprayers  applying a 0.08  per-
cent aqueous solution are exposed dermally -to 0.6 to 98.3 mg/hour.   Endosul-
fan can persist on the hands for 1 to 112 days after exposure  (Kazen,  et al.
1974).
III. PHARMACOKINETICS
     A.  Absorption
         Undiluted endosulfan is  slowly  and incompletely  absorbed  from  the
mammalian  gastointestinal tract,  whereas endosulfan  dissolved  in cottonseed.
oil is readily though not  completely  absorbed  (Boyd and Dobo.s,  1969; Maier-
Bode,  1968).   The  beta-isomer is  more readily absorbed than the alphaisomer.
Alcohols,  oils, and  emulsifiers  accelerate  the absorption  of endosulfan  by
the skin  (Maier-Bode,  1968).  Inhalation is  not  considered to be an impor-
tant route of  absorption for endosulfan  except in spray operators (U.S.  EPA,
1979).

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     8.  Distribution
         After  ingestion by  experimental animals,  sndosulfan is  first  dis-
tributed to  the liver and then  to the other organs  of the body and  the re-
mainder  of  the gastrointestinal  tract  (Boyd  and  Oobos,  1969;  Maier-Sode,
1968).  In cats,  endosulfan  levels  peaked in brain,  liver,  spinal cord and
plasma, with the  brain and liver  retaining  the highest concentrations  after
administration  of a 3 mg/kg dose (Khanna,  et al. 1979).
         In  mice,  24  hours   after  oral  administration  of    C-endosulfan,
residues were detected in  fat,  liver, kidney,  brain, and  blood  (Deema,  et
al. 1966).
         Data  from  autopsies  of three suicides  show levels of endosulfan  in
brain which  were  much lower  than  those  in liver and  kidney,  which in  turn,
were lower  than levels  in  blood  (Coutselinis,  et  al. 1978).   Data from an-
other suicide  indicate  higher levels of endosulfan in liver and  kidneys  than
in blood (Demeter, et al. 1977).
     C.  Metabolism
         Endosulfan sulfate  is the metabolite most commonly present  in  tis-
sues, feces,  and  milk  of  mammals  after  administration of endosulfan  (Whit-
acre, 1970;  Oemma,  et  al.  1966;. FMC, 1963).  The largest amounts of endosul-
fan sulfate  are found in small  intestine  and visceral  fat  with only traces
in skeletal  muscle  and kidney (Deema, et  al.  1966).  Endosulfan sulfate has
been detected in the brains of two humans who committed suicide  by ingesting
endosulfan  (Demeter and  Heyndrickx, 1978),  but not--in the-brains  of  mice
                                                         ,•
given nonfatal  doses of  endosulfan.   However,, it has been detected in liver,
visceral fat and small  intestines  of mice (Deema,  et al. 1966).  Other meta-
bolites of endosulfan are endosulfan lactone,  endosulfandiol,  andosulfan hy-
droxyether,  and endosulfan ether (Knowles, 1974; Menzie, 1974).  These meta-
bolites have also been found  in microorganisms and plants (U.S. EPA, 1979).

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     D.  Excretion
         The principal  route  of excretion for endosulfan and endosulfan  sul-
fate is in  the  feces  (U.S.  EPA, 1979).   Other metabolites are also  excreted
in the feces 'and  to a'small extent in the urine, the metabolites  in  the  lat-
ter  being  mainly in  the form  of  endosulfan  alcohol  (U.S. EPA,  1979).    In
studies with sheep  receiving a single oral  dose of radiolabeled  endosulfan,
92 percent  of the dose  was  eliminated in 22 days.   The  organ with the high-
est  concentration of radiolabeled  endosulfan after 40  days was  the  liver.
Major metabolites did not persist in the fat  or in the  organs (GorbachT  et
al. 1968).  After a single  oral dose, the half-life of radiolabeled  endosul-
fan  in the feces and urine of sheep was approximately two  days  (Kloss,  et
al. 1966).  Following 14 days of dietary exposure  of female rats, the half-
life of  endosulfan residues  was approximately  seven  days  (Dorough, et al.«
1978).
IV.  EFFECTS
     A.  Carcinogenicity
         In bioassays on both mice and  rats,  orally administered endosulfan
was not carcinogenic  even though doses were high enough  to produce  symptoms
of toxicity (Kotin,  et al.  1968;  Innes,  et  al..  1969;  Weisburger,  et  al.
1978).
     B.  Mutagenicity
         Data from  assays with Salmonella typhimurium (with and without mi-
crosomal activation)  (Dorough, et al. 1978), Sacchaxomyces. cerevisiae, Esch-
ericia coli, and  Serratia marcescens  (Fahrig,  1974) indicate that endosulfan
is not mutagenic.

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     C.  Teratogenicity
         Endosulfan  did  not  produce  teratogenic effects  in  rats  (Gupta,
1978).
     0.  Other Reproductive Effects
         In  rats,  endosulfan  produced dose-related  increases  in  maternal
toxicity  and caused  increases in  fetal  mortality and. resorptions  (Gupta,
1978).   Doses of 100  mg/kg'  reduce  hatchability  of  fertile, white  leghorn
chicken eggs  by  54 percent,  but this was  dependent on carrier (Ounachie  and
Fletcher,  1969).  Alterations  in  the  gonads  of the  embryos  within  sprayed
hens' _eggs were  noted  and the progeny  of  hens and quails, Cotumix  Cotumix
japonica, were sterile  (U.S. EPA, 1979).
     E.  Chronic Toxicity
         In the  NCI bioassays (Xotin,  et  al.  1968; Weisberger, et al.  1978)
endosulfan was toxic to the kidneys  of rats of both  sexes, and  to the kid-
neys of  male mice.   Other signs of  toxicity  were parathyroid hyperplasia,
testicular atrophy in male rats, and  high  early death  rates in male mice.
         In  a  two-year  feeding  study  with  rats (Hazelton 'Laboratories,
1959), endosulfan  at 10 mg/kg diet reduced  testis weight in  males and low-
ered survival in females; at  100  mg/kg diet,  renal tubular  damage and some
hydropic changes in the liver were induced.
         In humans,  there has been an  absence of  toxic effects with proper
handling of endosulfan in the occupational setting  (Hoechst, 1966).
     F.  Other Relevant Information               "  	
         The  acute  toxicity  of endosulfan sulfate  is  about  the same as that
of  endosulfan.   The LD50 for technical  endosulfan in  rats  is *— 22 to  &6
mg/kg and  6.9 to 7.5 mg/kg in mice  (Gupta,  1976).  Reagent grade  a- and  £-
endosulfan are less  toxic to rats (76  and 240 mg/kg,  respectively; Hoechst,

                                      .  I . I A_
                                   -  I I 7<>
                                      /,

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 1967).   The  inhalation 4-hour  LC5Q values  for rats  have been  reported  as
 350 and  80 ug/1  for males  and females,  respectively  (Ely,  et al.  1967).
 Acute  toxicities  of other metabolites  (endosulfan lactone,  endosulfandiol,
 endosulfan hydroxyether  and endosulfan  ether)  are less  than  that of  the
 parent  compound (Dorough,  et  al. 1978).
          At very high  levels of acute exposure,  endosulfan is  toxic  to the
 central nervous  system (U.S.  EPA,  1979).   Endosulfan  is  a convulsant  and
 causes  fainting, tremors, mental confusion,  irritability,  difficulty in uri-
 nation,  loss of  memory and  impairment of  visual-motor  coordination.   Acute
 intoxification  can be relieved by diazepam but  chronic effects  are  manifest-
 ed  in central nervous system  disorders  (Aleksandrowicz, 1979).
          There  appear  to  be sex differences (see previous  Chronic Toxicity
 section)  and species differences in sensitivity to endosulfan.   Of the spe-
 cies tested with endosulfan,  cattle are the  most  sensitive to  the neurotoxic
 effects of  endosulfan and appear   to be  closer  in sensitivity to  humans.
 Dermal  toxicity of endosulfan-sprayed cattle is also high.   Typical symptoms
'are listlessness, blind  staggers,  restlessness, hyperexcitability,  muscular
 spasms, goose-stepping and convulsions  (U.S.  EPA,  1979).
          Endosulfan  is  a nonspecific inducer  of  drug metabolizing  enzymes
 (Agarwal,  et al. 1978).   Protein deficient  rats are somewhat more  suscepti-
 ble to  the toxic effects of  endosulfan than  controls  (Boyd and  Oobos,  1969;
 Boyd, et  al.  1970).
 V.   AQUATIC TOXICITY
     A.   Acute  Toxicity
          Ninety-six   hour LC5g  values,  using  technical  grade  endosulfan,
 for five  species of  freshwater fish range  from  0.3  jjg/1  for  the  rainbow
 trout,  Salmo qairdneri,  (Macek,  et  al.  1969)  to  11.0  ;jg/l for  carp  finger-

                                   - I [ >/..{—
                                  ' j ) I >
                                      f

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lings, Cyprinus  caraio (Macek,  et  al. 1969;  Schoettger,  1970; Ludemann  acid-
Neumann,  1960;  Pickering  and Henderson,  1966).  Among  freshwater  inverte-
brates, Oaohnia  magna is  reported  to have  48-hour  LC5Q values ranging, from-
62 to 166 ug/r (Macek,  et al. 1976; Schoettger,  1970),  with three other  in-
vertebrates  yielding 96-hour  LC5Q values  of 2.3  (Sanders  and  Cope,  1968)
to 107 jjg/1  (Sanders, 1969; Schoettger,  1970).   Levels of  400 and 800  ng/1
of technical endosulfan damaged the kidney,  liver,  stomach and intestine  of
Gvmonocorymbus  ternetzi.   The  96-hour LCep value was  1.6 ug/L- (Amminikutty—
and Rege, 1977,1978).
         Of  the  five saltwater fish species  tested,  the reported 48- or  96-
hour  LCeg values  ranged  from  0.09  (Schimmel,  et  al.  1977) to  0.6" ^g/1
(Butler,  1963,1964;  Korn  and Earnest,  1974; Schimmel,  et  al.  1977).    The
most sensitive species was the spot (Leiostomus xanthurus).
         The  seven saltwater  invertebrate species tested showed a wide  range
of sensitivity  to endosulfan.   The range  of EC5Q  and LC5Q  values  is  from
0.04 (Schimmel, et al. 1977)  to  380 jug/1  with the most sensitive species  be-
ing the pink shrimp  (Penaeus duorarum).
     3.  Chronic Toxicity
         Macek,  et  al. (1976)  provided  the  only aquatic  chronic  study  in-
volving endosulfan.  No adverse  effects on fathead minnow, Pimephales orome-
las,  parents or offspring were  observed  at 0.20 jug/1.   Gvmonocorvmbus ter-
netzi chronically exposed  to  400 and 530 ng/1 for 16 weeks evinced necrosis
of intestinal mucosa cells,  ruptured  hepatic cells "and  destruction  of pan-
creatic islet cells  (Amminikutty and Rege, 1977,1978).
     C.  Plant Effects
         Little  data is  available  concerning the  effects  of  endosulfan  on
aquatic  micro/macrophytes.   Growth  of  Chlorella  vuloaris   was  inhibited
 >2000ug/l (Knauf and Schulze, 1973).
                                      4
                                       ?*•//

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     0.  Residues
         Schimmel, et  al.  (1977)  studied the uptake, depuration, and metabo-
lism of endosulfan by  the striped mullet, Mugil  cephalus.   When the-eoncen--
trations  of endosulfans  I and  II and  endosulfan sulfate  were combined  to
determine  the  bioconcentration factor  (BCF), an  average whole-body  BCF  of
1,597  was  obtained.   Nearly all  the  endosulfan was in  the  form of the  sul-
fate.  Even though the duration  of the study was  28 days,  this investigator
questioned  whether- a  steady-state condition was  reached.   Complete-- depura-- -
tion occurred  in  just  two days in an endosulfan-free environment.  Residues
in pond  sediments may  be  as high as 50 pg/kg B-endosulfan  and 70 pg/kg  of
endosulfan  sulfate 280 days after insecticidal  endosulfan  application  (FMC,
1971).
         Dislodgable residues  on  cotton foliage  in Arizona  declined  to  10
percent and one-third  for the low-melting, and  high-melting  isomers, respec-
tively, 24  hours  after application of 1.1 kg/ha endosulfan.   However, though
residues had declined  to 4 percent and  11 percent respectively, 4 days  after
application endosulfan sulfate residues on  the  leaves  increased markedly  to
0.14 jug/cm2 (Estesen,  1979).
VI.  EXISTING GUIDELINES AND STANDARDS
     Neither the  human health nor the  aquatic criteria, derived by U.S. EPA
(1979), which are summarized  below,  have gone through  the process of public
review;  therefore,   there  is  a  possibility  that  these criteria will   be
changed.
                                                          j
     A.  Human
         The U.S. EPA  (1979)  has recommended a draft criterion for endosul-
                                                                            »
fan in ambient  water of 0.1 mg/1  based  on  an ADI of 0.28 mg/day.   This ADI
was calculated  from a NOAEL  of  0.4 mg/kg  obtained for mice  in  a chronic
feeding study (Weisburger, et al. 1978) and an uncertainty factor of 100.
                                     • ••it
                                 ^}) I  J

-------
         The  American   Conference   of  Governmental  Industrial  Hygienists
(ACGIH, 1977)  TLV time  weighted average  for endosulfan  is  0.1 mg/m .  The
tentative value  for  the TLV  short-term exposure  limit  (15  minutes)  is 0.3
mg/m .
         The  ADI for  endosulfan established by  the Food  and Agricultural
Organization and the World Health Organization is 7.5 ug/kg (FAO, 1975).
     B.  Aquatic
         For  endosulfan,  the draft  criterion to  protect  freshwater aquatic
life is 0.042 ug/1  in a 24-hour average  and not to exceed 0.49 ug/1 at any
time.  Saltwater criteria cannot be developed because  of insufficient data
(U.S. EPA, 1979).

-------
                                  ENDOSULFAN
                                  REFERENCES
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agents in the workroom environment  with intended changes for 1977.  1977 TLV
Ariborne  Contaminants Committee,  American Conference  of  Government  Indus-
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Agarwal, O.K.,  et  al.  1978.  Effect  of  endosulfan  on  drug metabolizing en-
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Aleksandrowicz,  O.R.  1979.   Endosulfan  poisoning  and  chronic  brain  syn-
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Amminikutty, C.K.  and M.S.  Rege.   1977.  Effects  of acute  and  chronic ex-
posure to pesticides, Thioden-35 E.C.  and Aoallol "3" on the liver of widow
tetra (Gymonocorymbus ternetzi).  Boulenger-Indiana Jour. Exp. Biol.  15: 97.

Amminikutty, C.K. and M.S. Rege.  1978.   Acute and chronic  effect of Thioden
35 E.C.  and Aoallol "3"  on kidney, stomach and intestine of the widow- tetra
(Gymonocorymbus temetzi).  Boulenger-IndiaRa Jour.  Exp. Biol.  16: 202.

Berg,  H.    1976.    Farm   chemicals   handbook.   Meister   Publishing  Co.,
Willoughby, Ohio.

Boyd, E.M.  and  I. Dobos.   1969.  Protein deficiency  and tolerated oral doses
of endosulfan.  Arch. Int. Pharmacodyn.  178: 152.

Boyd, E.M.,  et  al.    1970.   Endosulfan toxicity and dietary  protein.   Arch.
Environ. Health  21: 15.

Butler, P.A.  1963.   Commercial fisheries investigations,  pesticide-wildlife
studies.  A review  of Fish  and  Wildlife Service Investigations  during 1961
and 1962.  U.S. Oept. Inter. Fish Wildl. Circ.  167:  11.

Butler, P.A.  1964.   Pesticide-wildlife  studies,  1963.   A review  of Fish and
Wildlife Service Investigations during the calendar  year.   U.S. Oept.  Inter.
Fish Wildl. Circ.  199: 5.

Cassil, C.C. and P.E. Orummond.  1965.  A plant surface oxidation product of
endosulfan.  Jour.  Econ.  Entomol.  58: 356.

Chopra, N.  and A.  Mahfouz.   1977.   Metabolism of  eridosulfa'h  I,  endosulfan
II, and endosulfan sulfate in tobacco leaf.  Jour. Agric^ Food Chem.  25: 32.

Comeliussen, P.E.   1970.   Residues in food and  feed:  pesticide  residues in
total diet samples (V).  Pestic. Monit. Jour.   4: 89.

Coutselinis, A., et al.   1978.  Concentration levels of endosulfan in bio-
logical material (report of three cases).  Forensic  Sci.  11: 75.

-------
Oeema,  P.,  et  al.   1966.  Metabolism,  storage,  and  excretion of  i4C-endo-
sulfan  in the mouse.   Jour. Econ.  Entomol.   59:  546.

Oemeter,  3.  and A.  Heyndrickx.   1978.  Two  lethal endosulfan  poisonings  in
man.  Jour. Anal. Toxicol.  2: 68.

Oemeter,  J.,  et al.   1977.   Toxicological analysis  in a case  of endosulfan
suicide.  Bull. Environ.  Contain. Toxicol.   18:  110.

Ocmanski,  J.J., et  al.  1973.   Insecticide  residues on  1971  U.S.  tobacco
products.  Tobacco Sci.   17: 80.

Oomanski,  J.J., et  al.  1974.   Insecticide  residues on  1973  U.S.  tobacco
products.  Tobacco Sci.   13: 111.

Oorough,  H.w.  and J.R.  Givson.   1972.  Chlorinated  insecticide  residues  in
cigarettes purchases in 1970-72.  Environ.  Entomol.   1:  739.

Dorough,  H.W.,  et- al.   1978.   Fate of  endosulfan in rats and  toxicological
considerations of apolar  metabolites.  Pestic. Biochem.  Physiol.  8: 241.

Ouggan,  R.E,  and  P.E.   Comeliussen.   1972.   Dietary intake  of  pesticide
chemicals  in. the  United States  (III),  Jutfe 1968 to April  1970.   Pestic.
Monit.  Jour.  5: 331.

Ounachie, J.F.  and w.w.  Fletcher.   1966.  Effect of some insecticides on the
hatching rate of hens' eggs.  Nature   212:  1062.

Ely,  T.S.,  et  al.   1967.   Convulsions  in  Thiodan  workers:  a  preliminary
report.  Jour. Qccup. Med.  9: 36..

Estesen,  B.J.,  et  al.   1979.   Oislodgable  insecticide residues on  cotton
foliage: Permethrin, Curocron, Fenvalarate, Sulprotos, Oecis and  Endosulfan-
Bull. Environ. Contain. Toxicol.  22: 245.

Fahrig, R.   1974.   comparative mutagenicity studies  with  pesticides.  Int.
Agency Res. Cancer Sci. Publ.  10: 161.

FAQ.  1975.  Pesticide  residues in food: report of the 1974 Joint Meeting  of
the FAQ Working Party of Experts  on  Pesticide  Residues and  the  WHO  Expert
Committee  on Pesticide  Residues.   Agricultural  Studies  NO.   97,  Food and
Agriculture Organization  of the United States, Rome.

FMC Corp.   1963.  Unpublished  laboratory  report of.  Niagara  Chemical  Divi-
sion,  FMC Corporation, Middleport, New York.  In: Maie"r-8ode,'' 1968'..
                                              ^™"         ,/

FMC Corp.  1971.   Project 015:  Determination of  endosulfan I,  endosulfan  II
and endosulfan  sulfate residues in  soil,  pond, mud  and water.    Unpublished
report.   Niagara  Chemical  Division,  FMC  Corp.,  Richmond, Cal.   In:  Nati.
Res. Council, Canada, 1975.

Gorbach, S.G., et al.   1968.  Metabolism of endosulfan in milk  sheep.   Jour.
Agric. Food Chem.  16: 950.

-------
Gupta,  P.K.   1976.   Endosulfan-induced  neurotoxicity  in  rats  and mice.
Bull. Environ. Contain. Toxicol.  15: 708.

Gupta, P.K.   1978.   Distribution  of endosulfan in plasma and brain after  re-
peated oral administration to rats.  Toxicology  9: 371.

Hazleton  Laboratories.   1959.  Unpublished report,  May 22.   Falls  Church,
Virginia.  In: ACGIH, 1971.

Hoechst.   1966.   Unpublished  report of  Farbwerke Hoechst  A.G., Frankfurt,
West Germany.  In: Maier-Bode, 1968.

Hoechst.   1967.   Oral  1050 values  for white  rats.   Unpublished  report of
Farbwerke  Hoechst A.G.,  Frankfurtr,  West Germany.   Cited  in  Demeter   and
Heyndrickx, 1978.  Jour. Anal. Toxicol.  2: 68.

Innes,  J.R.M.,  et al.   1969.   bioassay  of pesticides  and  industrial chem-
icals  for tumorigenicity in  mice: a  preliminary  .note.  Jour.  Natl. Cancer
Inst.  42: 1101.

Kazen,  C.,  et al.   1976.  Persistence of pesticides  on the hands  of some.
occupationally exposed people.  Arch. Environ, health  29: 315.

Keil,  J.E.,  et  al.   1972.   Decay of  parathion and endosulfan  residues on
field-treated tobacco, South Carolina,  1971.  Pestic. Monit. Jour.  6: 73.

Khanna, R.N., et  al.  1979.   Distribution of endosulfan in cat brain..  Bull.
Environ. Contam. Toxicol.  22: 72.

Kloss, G.,  et al.   1966.   Versuche an Schaffen mit  C^-markierten Thiodan.
Unpublished.  In: Maier-8ode, 1968.

Knaut, W.  and C.F.  Schulze.   1973.  New  findings on the toxicity  of endo-
sulfan and  its  metabolites to  aquatic organisms.   Meded.  Fac.  Landlouwwey.
Kijksuniv. Gent.  38: 717.

Knowles,  C.O.   1974.  Detoxification  of acaricides by  animals.   Pages 155-
176  In:  M.A.  Kahn and  J.P.  Bederka, Jr.,  eds.   Survival in  toxic environ-
ments.  Academic Press,  New York.

Korn,  S., and  R.  Earnest.   1974.   Acute  toxicity  of 20  insecticides  to
striped bass Morone saxatilis.  Calif. Fish Game  69: 128.

Kotin, P., et al.  1968.  Evaluation of  carcinogenic,  teratogenic and muta-
genic activites  of  selected pesticides and industrial' chemicals.   Pages   64,
69  In:  Vol.   1:  carcinogenic  study.  Bionetics  Research .Laboratories report
to Natl. Cancer Inst.  NTIS-PB-223-159.

Lee, R.L., Jr.   1976.  Air pollution  from pesticides and  agricultural pro-
cess.  CRC Press, Inc.,  Cleveland, Ohio.

Ludemann,  D.  and H.  Neumann.   1960.   Versuche   uber   die  akute  toxische
Wirkung  neuzeitlicher Kontaktinsektizide  auf  einsommerige Karfen  (Cyprinum
carpio L.)  Z. Angew. Zool.   47: 11.

-------
Macek, K.J.,  et al.   1969.   The effects of temperature  on the susceptibility
of  bluegills  and rainbow  trout  to  selected  pesticides.   Bull..  Environ.
Contam. Toxicol.  4: 174.

Macsk, K.J.,  et 'al.   1976.   Toxicity of  four  pesticides to water  fleas and
fathead minnows.  EPA-600/3-76-G99.  U.S.  Environ.  Prot.  Agency.

Maier-Sode,  H.   1968.   Properties,  effect,  residues  and  analytics of  the
insecticide endosulfan (review).  Residue  Rev.   22: 2.

Martens,  R.   1976.   Degradation  of  (8,9,-C-14)  endosulfan  by  soil  micro-
organisms.  Appl. Environ. Microbiol.  31: 853.

Menzie,  C.M.    1974.   Metabolism of  pesticides: an  update.  Special  scien-
tific  report.   Fish  and Wildlife Service, Wildlife 184.  U.S. Department of
Interior, Washington, D.C.

Miles, J.R.W.  and P.  Moy.   1979.   Degradation  of  endosulfan and its  metab-
olites by a mixed" culture of  soil microorganisms.  Bull~ Environ. Contam..
Toxicol.  23: 13.

Pickering, Q.H. and  C. Henderson.  1966.  The  acute toxicity of some  pesti-
cides to  fish.  Ohio Jour. Sci.  66: 508.

Sanders,  H.O.   1969.   Toxicity  of  pesticides  to the  crustacean  Gammarus
lacustris.  U.S. Bur. Sport,Fish Wildl. Tech. Pap.  25.

Sanders,  H.O.   and- O.B.  Cope.   1968.   The  relative  toxicities  of several
pesticides  to  naiads  of  three species  of  stoneflies.   Limnol.   Oceanogr.
13: 112.

Schimmel,  S.C., et  al.   1977.   Acute toxicity  to and  bioconcentration  of
endosulfan by  estuarine  animals.   Aquatic toxicology and hazard evaluation.
ASTM STP  634, AM. Sac. Test. Mat.

Schoettger,  R.A.  1970.   Fish-pesticide  research  laboratory,  progress  in
sport  fishery  research.  U.S.  Dept.   Inter.  Bur.  sport  Fish Wildl. Resour.
Publ. 106.

U.S. EPA.  1979.  Endrin: Ambient Water Quality Criteria.   (Draft)

Weisburger, J.H.,  et al.   1978.   Bioassay of  endosulfan  for  possible car-
cinogenicity.   National  Cancer  Institute  Division  of  Cancer  Cause   and
Prevention,  National  Institutes   of   Health,   Public "Health" Service,  U.S.
Department  of  Health,  Education  and  Welfare,  Bethesda,  Maryland,  Pub.
78-1312.   Report by Hazleton Laboratories to NCI, NCI-CG-TR-62.  54  pp.

Whitacre,  O.M.   1970.  Endosulfan  metabolism  in  temperature-stressed  rats.
Diss. Abstr. Int.  30: 44358.

Wolfe, H.R.,   et  al.   1972.    Exposure  of  spraymen to  pesticides.   Arch.
Environ.  Health  25:  29.

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                                       No. 99
               Endrin
  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980
           	,-i ./ r>~
             11  ) 7 -*

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                           EN DRIM




                           SUMMARY



     Endrin does not appear to be carcinogenic.   Endrin  is



teratogenic and embroytoxic in high doses and produces gross



chromosomal abnormalities when administered  intratesticu-



larly.  Chronic administration of endrin causes damage to  the



liver, lung, kidney, and heart of experimental animals.  No



information about chronic effects in humans  is available.



The ADI established by the Food and Agricultural  Organization



and World Health Organization is 0.002 mg/kg.



     Endrin has proven to be extremely toxic to aquatic  orga-



nisms.  In general, marine fish are more sensitive  to endrin



with an arithmetic mean LC^Q value of 0.73 ug/1,  than



freshwater fish with an arithmetic mean LC^Q value  of



4.42 ug/1.  Invertebrate species tend to be more  resistant



than fish with arithmetic mean LC^Q values of 3.80  and



58.91 ug/1 for marine and freshwater invertebrates,  respec-



tively.


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                           ENDRIN



I.   INTRODUCTION



     Endrin (molecular weight 374) is a broad  spectrum  insec-



ticide of- the group of polycyclic chlorinated  cyclodiene  hy-



drocarbons of which the insecticides aldrin and dieldrin  are



also members.  Endrin is isomeric with dieldrin and  is  used



as a rodenticide and ovicide.  The endrin sold  in  the U.S.  is



a technical grade product containing not less  than 95 percent



active ingred'ient.  The solubility of endrin in water at  25 °C



is about 200 ug/1 (U.S.  EPA, 1979).  Its vapor pressure  is 2



x 10"7 mm Hg at 25°C (Martin, 1971).



     Endrin is used primarily as an insecticide and  also  as a



rodenticide and avicide.  Over the past several years,  endrin



utilization has been increasingly restricted (U.S. EPA, 1979.



Endrin production in 1978 was approximately 400,000 ..pounds



(U.S. EPA, 1978).  Endrin persists in the soil  (U.S. EPA,



1979) .



II.  EXPOSURE



     A.   Water



          Occasionally, groundwater may contain more than 0.1



ug/1.  Levels as high as 3 ug/1 have been correlated with



precipitation and run off following endrin applications (U.S.



EPA, 1978).



          Concentrations of endrin in finished drinking water



have been decreasing.  In a study of ten municipal water



treatment plants on the Mississippi or Missouri Rivers, the*



number of finished water samples containing concentrations of



endrin exceeding 0.1 ug/1 decreased from ten percent in 1964-
                           99-f

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1965 to zero Ln 1966-1967  (Schafer,  et  al.,  1969)..  The  high-


est concentration of endrin  in drinking water  in New Orleans,


Louisiana measured by the U.S. EPA  in 1974 was  4 ng/1 (U.S.


EPA, 1974).


     B.   Food


          The general population  is  rarely exposed  to endrin


through the diet.  In the market  basket study  by the FDA,  the


total average daily intake from food ranged  from approximate-


ly 0.009 ug/kg body weight in 1965  to 0.0005 ugAg  body


weight in 1970 (Duggan and Lipscomb, 1969; Duggan and Corne-


liussen, 1972).


          The U.S. EPA (1979) has estimated  the weighted av-


erage bioconcentration factor of  endrin at 1,900 for the edi-


ble portions of fish and shellfish  consumed  by  Americans.


This estimate is based on measured  steady-state bioconcentra-


tion studies in six species  (both freshwater and saltwater).


     C.   Inhalation


          Exposure of the general population to endrin via


the air decreased from a maximum  level  of 25.6  ug/ra3 in


1971 to a maximum level of 0.5 ug/m3 in 1975 (U.S.  EPA,


1979).


          Tobacco products are contaminated  with endrin  cesir-


dues.  Average endrin residues for  various types of  tobacco


products have been reported  in the  range of'0.05 ug/g to 0.2


ug/<3 (Bowery, et al., 1959;  Domanski and Guthrie, 1974).
                                                           •

          Inhalation exposure of  users  and manufacturers of


endrin sprays may be around  10 ug/hour  (Wolfe,  et al. 1967)


but use of dusts can produce levels  as  high  as  0.41  mg/hour


(Wolfe, et al. 1963).

-------
     D.   Dermal

          Dermal exposure of spray operators .can  range up to

3 mg/body/hour even for operators wearing  standard  protective

clothing (Wolfe, et al. 1963, 1967).  The  spraying  of dusts

can lead to exposures of up to 19 mg/hour  (Wolfe, et  al.

1963).

III. PHARMACOKINETICS

     A.   Absorption

          Endrin is known to be absorbed through  the  skin,

lungs, and gut, but data on the rates of absorption are  not

available (U.S.  EPA, 1979).

B.   Distribution

          Endrin is not stored in human tissues  in  signifi-

cant quantities.  Residues were not detected  in plasma,  adi-

pose tissue, or urine of workers exposed to endrin  (Hayes and

Curley, 1968).  Measurable levels of endrin have  not  been de-

tected in human subcutaneous fat or blood, even  in  persons

living in areas where endrin is used extensively  (U.S.  EPA,

1979).  Endrin residues have been detected in  the body tis-

sues of humans only immediately after an acute exposure  (U.S.

EPA, 1979; Coble, et al. 1967).

          In a 128 day study, dogs were fed 0.1 mg/endrin/kg..

body weight/day.  Concentrations of endrin, in  the tissues at

the end of the experiment were as follows: adipose  tissue,

0.3 to 0.8 ug/g; heart, pancreas, and muscle,  0.3  ug/1;
                                                            •
liver, kidney and lungs, 0.077 to 0.085 ug/g;  blood,  0.002  to

0.008 ug/g (Richardson, et al., 1967).  In a  six  month feed-

ing study with dogs at endrin  levels of 4  to  8 ppm  in the

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diet, concentrations of endrin were  1 ug/g  in  fat,  1  ug/g  in

liver, and 0.5 ug/g in kidney  (Treon, et al.,. 1955).

     C.   Metabolism

          In rats, endrin  is readily metabolized  in the  liver

and excreted as hydrophilic metabolites  including hydroxyen-

drins, and 12-ketoendrin  (also known as  9-ketoendrin).   Hy-

droxyendrins and especially 12-ketoendrin have  been reported

to be more acutely tox ic  to mammals  than the parent compound

(Bedford, et al., 1975;  Hutson, et al.,  19.75).  The 12-keto-

endrin is also more persistent in tissues.  Female  rats  me-

tabolize endrin more slowly than males  (Jager,  1970).

     D.   Excretion

          Endrin is one of the least persistent chlorinated

hydrocarbon pesticides (U.S.. EPA, 1979).  Body  content of  en-

drin declines fairly rapidly after a single dose  or when a

continuous feeding experiment  is terminated (Brooks,  1969).

In rats, endrin and its metabolites  are  primarily excreted

with the feces (Cole, et  al., 1968; Jager,  1970).   The major

metabolite in rats is anti-12-hydroxyendrin which is  excreted

in bile as the glucuronide.  12-Ketoendrin was  observed  as a

urinary metabolite in male rats; the major  urinary  metabolite

in female rats is anti-12-hydroxyendrin-O-sulfate (Hutson,  et

al., 1975).

          In rabbits, excretion is primarily urinary.  In  fe-

males, endrin excretion also occurs  through the milk.  Al-
                                                            »
though endrin is rapidly  eliminated  from the body,  some  of
   !
its metabolites nay persist for longer periods  of time (U.S.

EPA, 1979).

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IV.  EFFECTS



     A.   Carcinogenicity



          In lifetime feeding studies with Osborne-tMendel



rats, endrin was neither tumorigenic nor carcinogenic  (Deich-



mann, et al., 1970; Deichmann and MacDonald, 1971;  Deichmann,



1972).  A recent MCI bioassay concluded that endrin was  not



carcinogenic for Qsborne-Mendel rats or for B6C3F1 mice



(DHEW, 1979).  However, a different conclusion has  been
                                              •

reached by Reuber  (1979) based only on one study  (National



Cancer Institute, 1977), compared with eight other  inconclu-



sive or unsatisfactory studies.



     B.   Mutagenicity



          Endrin (1 mgAg) administered intratesticularly



caused chromosomal aberrations in germinal tissues of  rats,



including stickiness, bizarre configurations, and abnormal



disjunction (Dikshith and Datta, 1972, 19731) .



     C.   Teratogenicity



          An increased incidence of club foot was found  in



fetuses of mice that had been treated with endrin (0.58  mg/



kg) before becoming pregnant (Nodu, et al., 1972).



          Treatment of pregnant hamsters with endrin (5  mg/



kg) produced the following congenital abnormalities: open



eye, webbed foot, cleft palate, fused ribs, and. meningoen-



cephalocele (Ottolenghi, et al., 1974; Chernoff,  et al.,



1979).  Treatment of pregnant mice with endrin (2-5 mg/kg)
                                                           »

produced open eye and cleft palate in the offspring (Otto-



lenghi, et al., 1974).  Single doses which produced terato-

-------
genie effects in hamsters and mice were one-half  the LD^g



in each species (Ottolenghi, et al., 1974).



     D.   Other Reproductive Effects



          •Endrin given to hamsters during gestation produced



behavioral effects in both dams and offspring  (Gray, et  al.,



1979).  In another study endrin produced a high incidence of



fetal death and growth retardation (Ottolenghi, et al.,



1974).



     E-   Chronic Toxicity



          Mammals appeared to be sensitive to  the toxic  ef-



fects of endrin at low levels in their diet.   Significant



mortality occurred in deer mice fed endrin at  2 mg/kg/day in



the diet (Morris, 1968).  The mice exhibited symptoms of CNS



toxicity including convulsions.  Lifetime feeding of endrin



to rats at 12 mg/kg/day in the diet decreased  viability  and



produced moderate increases in congestion and  focal hemor-



rhages of the lung;  slight enlargment, congestion and mott-



ling of the liver, and slight enlargement, discoloration or



congestion of the kidneys (Deichmann, et al.,  1970).  After



19 months on diets containing 3 mg/kg/day endrin, dogs had



significantly enlarged kidneys and hearts (Treon, et al.,



1955).



          Chronic administration of relatively small doses of



endrin to monkeys produced a characteristic .change in the



electroencephalogram (EEC);  at higher doses, electrographic



seizures developed.   EEC and behavior were still abnormal



three weeks after termination of endrin administration;  sei-

-------
zures recurred under stress conditions months after  termina-
tion of endrin administration  (Kevin, 1968).
     P.   Other Relevant Information
          Endrin is more toxic, in both acute and chronic
studies, than other cyclediene insecticides  (U.S. EPA,
1979).
          Female rats metabolize and eliminate endrin more
slowly than males (Jager, 1970) and are more sensitive  to en-
drin toxicity (U.S. EPA, 1979).  Dogs and ..monkeys are more
susceptible to endrin toxicity than other species (U.S. EPA,
1979).
          Endrin, given in equitoxic doses with delnav, DDT,
or parathion gave lower than expected LD5Q values, sug-
gestive of antagonism.  Endrin given in equitoxic doses with
aldrin (a closely related compound) or chlordarie gave higher
than expected LD50 values suggestive of synergism (Kep-
linger and Deichmann, 1967).  Humans poisoned acutely exhibit
convulsions, vomiting, abdominal pain, nausea, dizziness,
mental confusion, muscle twitching and headache.  Such  symp-
toms have been elicited by doses as low as 0.2 mg/kg body
weight.  Any deaths have usually occurred through respiratory
failure (Brooks, 1974).
V.   AQUATIC TOXICITY
     A.   Acute
          The toxic effects of endrin have been extensively
studied in freshwater fish.  LCgQ values for static
bioassays ranged from 0.046 ug/1 £or carp fry (Cyprinus
carpio) fry to 140.00 ug/1 for adult carp (lyatomi, et  al.,

                              y
                            Co-//)

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1958).  Excluding the results of age factor differences  for


this species, adjusted static LCjQ values ranged  from


0.27 ug/1 for large mouth bass  (Microptecus salmoides)


(Fabacler, 1976) to 8.25 ug/1 for the bluegill  (Lepomis


macrochirus) (Katz and Chadwick, 1961).  The LC50 values


for flow-through assays were 0.27 ug/1 for the  bluntnose


minnow (Pimeplales notatus) to  2.00 ug/1 for the  bluegill


(U.S. EPA, 1979).  Twenty-five  LC5Q values for  17 species


of freshwater invertebrates were reported', and  ranged  from


0.25 ug/1 for stoneflies (Pteronarcys californica) to  500.0


ug/1 for the snail, (Physa gyrina) (U.S. EPA, 1979).


          For marine fish, LCgQ values ranged from 0.005


ug/1 for the Atlantic silversides (Menidia menidia)  (Eisler,


1970) to 3.1 ug/1 for the northern puffer (Sphaeroides macu-


latus).  A total of 17 species  were tested in 33  bioassays.


The most sensitive marine invertebrate tested was the  pink


shrimp, (Penaeus duordrum) with an LC50 value of  0.037


ug/1, while the blue crab (Callinectes sapidus) was  the most


resistant, with an LC50 of 25 ug/1.


     B.   Chronic


          Freshwater fish chronic values of 0.187 ug/1 and


0.257 ug/1 were reported for fathead minnows (Pimephales


promelas) (Jarvinen and Tyo, 1978) and flagf ish.. (Jordanella


floridae) Hermanutz, 1978), respectively, in life cycle


toxicity tests.  No freshwater  invertebrate species  have been
                                                           •

chronically examined.  The marine fish, the sheepshead minnow


(Cyprinodon variegatus) has provided a chronic  value of 0.19


ug/1 from embryolarval tests (Hansen, et al., 1977).   The

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grass shrimp (Palaemonetes pugio) must  be  exposed  to less


than a chronic concentration of 0.038 ug/1  for  reproductive


success of this marine invertebrate  species (TylerShroeder,


in press)


     C.   Plants


          Toxic effects were elicited at concentrations  for


freshwater algae ranging from 475 ug/1  for  Anacystis nidu-


laras (Batterton, 1971) to >20,000 ug/1 for Scenedesmus  quad-


ricauda and Oedogonium sp.  Marine algae appeared.  more_ sensi-


tive with effective concentration ranging  from  0.2 ug/1  for


the algae, Agmenellum quadruplicaturn (Batterton, 1978),  to


1,000 ug/1 for the algae Dunaliella  tertiotecta (U.S.  EPA,


1979).


     D.   Residues


          Bioconcentration factors ranged  from  140 to  222 in


four species of freshwater algae.  Bioconcentration factors


ranging from 1,640 for the channel catfish  Ictalurus puncta-


tus (Argyle, et al. 1973) to 13,000  for the flagfish Jordan-


ella floridae (Hermanutz, 1978) have been  obtained.   Among


four marine species, bioconcentration factors ranging  from


1,000 to 2,780 were observed for invertebrates  and from  1,450


to 6,400 for marine fish.  Residues  as  high as  0.5 ppm have


been found in the mosquito fish, Gambusia-.affinjls  (Finley, et


al. 1970) and fish frequently have contained' levels  above 0.3


ppm (Jackson, 1976).
                                                            »

VI.  EXISTING GUIDELINES AND STANDARDS


     Both the human health and aquatic  criteria derived  by


U.S. EPA  (1979), which are summarized below,  have  not  gone

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through the process of public  review;, therefore,  there  is  a



possibility that these criteria may be changed.



A.   Human



          The U.S. EPA (1979)  has calculated  an ADI  for en-



drin of 70 ug from a NOAEL of  0.1 mg/kg  for dogs  in  a 128  day



feeding study and an uncertainity factor of 100.   The U.S.



EPA (1979) draft criterion of  1 ug/1  for endrin in ambient



water is based on the 1 ug/1 maximum  allowable concentration



for endrin in drinking water, proposed  by the  Public  Health



Service in 1965 (Schafer, et al., 1969)  and on the calcula-



tions by EPA.  Human exposure  is assumed to come  from drink-



ing water and fish products only.



          A maximum acceptable level  of  0.002 mg/kg  body



weight/day (ADI) was established by the  Food  and  Agricultural



Organization (1973) and the World Health Organization.



          A time weighted average TLV for endrin  of  100



ug/m3 has been established by  OSHA  (U.S.  Code of  Federal



Regulations, 1972) and ACGIH (Yobs, et al., 1972).



          The U.S. EPA (40 CFR Part 129.102)  has  promulgated



a toxic pollutant effluent standard for  endrin of 1.5 ug/1



per average working day calculated over  a period  of  one



month, not to exceed 7.5 ug/1  in any  sample representing one.



working-day's effluent.  In addition,  d iseharge.. is not  to  ex-



ceed 0.0006 kg per 1,000 kg of production.

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     B.   Aquatic



          The draft criterion  for  the protection  of fresh-



water aquatic life is 0.0020 ug/1  as a  24  hour  average  con-



centration not to exceed 0.10  ug/1.  For marine organisms,



the draft criterion is 0.0047  ug/1 as a 24  hour average not



to exceed 0.031 ug/1.

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                            ENDRIN

                          REFERENCES

Argyle, R.L., et al.  1973.  Endrin uptake and release by
fingerling channel catfish, Ictaluras punctatus.  Jour.
Fish Res. Board Can. 30: 1743.

Batterton, J.C., et al.  1971.  Growth response of bluegreen
algae to aldrin, dieldrin, endrin and their metabolites.
Bull. Environ. Contam. Toxicol. 6: 589.

Bedford, C.T., et al.  1975.  The acute toxicity of endrin
and its metabolites to rats.  Toxicol. Appl. Pharmacol.
33: 115.

Bowery, T.G., et al.  1959.  Insecticide residues on tobacco.
Jour. Agric. Food Chem.  7: 693.

Brooks, G.T.  1969.  The metabolism of diene-organochlorine
(cyclodiene) insecticides.  Residue Rev.  28: 81.

Brooks, G.T.  1974.  Chlorinatedlnsecticides.  Vol. II.
Biological and environmental aspects.  CRC Press, Cleveland,
Ohio.

Chernoff, N., et al.  1979.  Perinatal toxicity of endrin
in rodents.  I. Fetotoxic effects of prenatal exposure in
hamsters.  Manuscript submitted to Toxicol. Appl. Pharmacol.
and the U.S. Environ. Prot. Agency.

Colde, Y., et al.  1967.  Acute endrin poisoning.  Jour.
Amer. Med. Assoc. 203.: 489.

Cole, J.F.,  et al.  1968.  Endrin and dieldrin: A comparison
of hepatic excretion rates in the rat.  (Abstr.) Toxicol.
Appl. Pharmacol.  12: 298.

Deichmann, W.B.  1972.  Toxicology of DDT and related chlorin-
ated hydrocarbon pesticides.  Jour. Occup. Med.  14: 285.

Deichmann, W.B., and W.E. MacDonald.  1971.  Organochlorine
pesticides and human health.  Food Cosmet. Toxicol.  9:
91.

Deichmann, W.B., et al.  1970.  Tumorigenfcity of aldrin,
dieldrin, and endrin in the albino rat.  Indt Med. Srug.
39: 37.

Dikshith, T.S.S., and K.K. Datta.  1972.  Effect of intra-  .
testicular injection of lindane and endrine on the testes
of rats.  Acta Pharmacol. Toxicol.  31: 1.

-------
Dikshith, T.S.S., and K.K. Datta.  1973.  Endrin  induced
cytological changes in albino rats.  Bull. Environ. Cotam.
Toxicol.  9: 65.

Domanski, J.J., and F.E. Guthrie.  1974.  Pesticide residues
in 1972 cigars.  Bull. Environ. Contain. Toxicol.   11:  312.

Duggan, R.E., and G.Q. Lipscomb.  1969.  Dietary  intake
of pesticide chemicals in the United States  (II),  June 1966-
April 1968.  Pestic. Monitor. Jour.  2: 153.

Duggan, R.E., and P.E. Corneliussen.   1972.  Dietary  intake
of pesticide chemicals in the United States  (III) , June
1968-April 1970.  Pestic. Monitor. Jour.  5: 331.

Eisler, R.  1970.  Acute toxicities of organochlorine  and.
organophosphorous insecticides to estuarirve  fishes.  Tech.
Pap. 46.  Bur. Sport Fish. Wildl. U.S. Dep.  Inter.

Fabacher, D.L. 1976. ' Toxicity of endrin and an endrinmethyl
parathion formulation to largemouth bass fingerlings.  Bull.
Environ. Contain. Toxicol.  16: 376.

Finley, M.T., et al.  1970.  Possible  selective mechanisms
in the development of insecticide resistant  fish.  Pest.
Monit. Jour. 3: 212.

Gray, L.E., et al.  1979.  The effects of endrin  administra-
tion during gestation on the behavior of the golden hamster.
Abstracts from the 18th Ann. Meet. Soc. of Tox. New Orleans
p. A-200.

Hansen, D.J., et al.  1977.  Endrin:  Effects on  the  entire
lifecycle of saltwater fish, Cyprinodon variegatus.  Jour.
Toxicol. Environ. Health  3: TZT~.

Hayes, W.J., and A. Curley.  1968.  Storage  and excretion
of dieldrin and related compounds.  Arch. Environ. Health
16: 155.

Hermanutz, R.O.  1978.  Endrin and malathion toxicity  to
flagfish  (Jordanella floridae).  Arch. Enviorn. Contam.
Toxicol. 1: 159.

Hutson, D.H., et al.  1975..  Detoxification  and bioactiva-
tion of endrin in the rat.  Xenobiotica  Ii: 69-7-.

lyatomi, K.T., et al.  1958.  Toxicity of endrin  to fish.
Prog. Fish.-Cult.  20: 155.
                                                            »
Jackson, G.A.  1976.  Biologic half-life of  endrin in  chan-
nel catfish tissues.  Bull. Environ. Contam. Toxicol.  16:
505.

-------
Jager, K.W.  1970.  Aldrin, dieldrin, endrin, and  telodrin.
Elsevier Publishing Co., Amsterdam.

Jarvinen, A.W., and R.M. Tyo.  1978.  Toxicity to  fathead
minnows of endrin in food and water.  Arch. Environ. Contain.
Toxicol. 7: 409.

Katz, M., and G.G. Chadwick.  1961.  Toxicity of endrin
to some Pacific Northwest fishes.  Trans. Am. Fish. Soc.
90: 394.

Keplinger., M.L., and W.B. Deichmann.  1967.  Acute toxicity
of combinations of pesticides.  Toxicol. Appl. Pharmacol.
10: 586.

Martin, H.  1971.  Pesticide manual, 2nd ed.  Brit.  Crop
Prot. Council.

Morris, R.D.  1968.  Effects of endrin feeding on  survival
and reproduction in the deer mouse, Peromyscus maniculatus.
Can. Jour. Zool.  46: 951.

National Cancer Institute.  1977.  Bioassay of endrin for
possible carcinogenicity.  NCI Technical Report Series,
No. 25.

National Cancer Institute.  1979.  Bioassay of endrin for
possible carcinogenicity.  HEW Pub. No.  (NIH) 79-812.  U.S..
Dept. of Health, Education and Welfare, Bethesda, Md.

Nodu, et. al.  1972.  Influence of pesticides on embryos.
On the influence of organochloric pesticides  (in Japanese)
Oyo Yakuri  6: 673.

Ottolenghi, A.D., et al.  1974.  Teratogenic effects of
aldrin, dieldrin, and endrin in hamsters and mice.  Terato-
logy 9: 11.

Reuber, M.D.  1979. . Carcinogenicity of endrin.  Sci.  Tot.
Environ. 12: 101.

Revin, A.M.  1968.  Effects of chronic endrin administration
on brain electrical activity in the squirrel monkey.  Fed.
Prac.  27: 597.

Richardson, L.A., et al.  1967.  Relationship of dietary
intake to concentration of dieldrin and endrin in  dogs.
Bull. Environ. Contam. Toxicol.  2: 207.

Schafer, M.L., et al.  1969.  Pesticides in drinking water
- waters from the Mississippi and Missouri Rivers.  Environ*.
Sci. Technol.  3: 1261.

-------
Treon, J.F., et al.  1955.  Toxicity of endrin for labora-
tory animals.  Agric. Food Chem.  3: 842.

Tyler-Schroeder, D.B.  Use of grass shrimp, Palaemonetes
pugio, in a life-cycle toxicity test.  In Proceedings of
Symposium on Aquatic Toxicology and Hazard Evaluation.
L.L. Marking and R.A. Kimerle, eds. Am. Soc. Testing and
Materials (ASTM), October 31-November 1, 1977.   (In press).

U.S. EPA.  1974.  Draft analytical report—New Orleans area
water supply study.  Lower Mississippi River facility, Sur-
veillance and Analysis Division, Revion VI, Dallas. Texas.

U.S. EPA.  1978.  Endrin-Position Document 2/3.  Special
Pesticide Review Division.  Office of Pesticide  Programs,
Washington, D.C.

U.S. EPA.  1979.  Endrin:  Ambient Water Quality Criteria.
(Draft).                    '         •

Wolfe, H.R., et al.  1963.  Health hazards of the pesticides
endrin and dieldrin.  Arch. Enviorn. Health 6: 458.

Wolfe, H.R., et al.  1967.  Exposure of workers  to pesti-
cides.  Arch. Environ. Health 14: 622.

Yobs, A.R., et al.  1972.  Levels of selected pesticides
in ambient air of the United States.  Presented  at the National
American Chemical Society—Symposium of Pesticides in Air.
Boston,  Maine.

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                                        No. 100
Epichlorohydrin (l-chloro-2,3-epoxypropane)

      Health and Environmental Effects
    U.S.  ENVIRONMENTAL PROTECTION AGENCY
          WASHINGTON, D.C.   20460

              APRIL 30,  1980
                100"/

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                          DISCLAIMER
     This report represents  a survey of the potential' health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is  drawn chiefly
from secondary  sources  and  available reference documents.
Because of the limitations of such sources, this  short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented  by the
subject chemical.   This document  has undergone scrutiny to
ensure its technical accuracy.

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                       SPECIAL NOTATION
U.S. EPA1s Carcinogen Assessment Group (CAG) has evaluated
epichlorohydrin and has found sufficient evidence to in-
dicate that this compound is carcinogenic.

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                          l-CHLDRO-2,3-EPOXYPROPANE
                              (Epichlorohydrin)
                                   Summary

     The adverse health  effects  associated with exposure to epichlorohydrin
are extreme irritation to the eyes, skin, and respiratory tract.  Inhalation
of  vapor  and  percutaneous  absorption of  the  liquid  are the  normal human
routes of  entry.   Exposure  to epichlorohydrin usually  results  from occupa-
tional contact with the chemical, especially in glycerol and epoxy resin op-
erations.   Pulmonary effects have been well documented.  Recent studies have
demonstrated epichlorohydrin  to  be a potent  carcinogen to nasal  tissue  in
experimental animals.   Cytogenic  studies  both in vitro and in vivo in humans
and  experimental  animals  have   indicated  epichlorohydrin to  be  an  active
clastogenic agent.  NO data on the concentration of epichlorohydrin in drink-
ing water or foods have been  reported.  Studies on the  effects of epichloro-
hydrin to aquatic  organisms  could not be  located in the  available literature.
                                100-V

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 I.    INTRODUCTION


      This  profile is  based  primarily on  a  comprehensive review  compiled  by


 Santodonato,  et al.  (1979).   The health hazards  of  epichlorohydrin  have also


 been reviewed, by  the  National Institute  for  Occupational Safety and  Health


 (NIOSH,  1976)  and the  Syracuse Research Corporation  (SRC, 1979).


      Epichlorohydrin  (OLOCHOLCl;   molecular  weight  92.53)   is  a  color-


 less liquid  at  room  temperature with  a distinctive chloroform-type  odor.


 The boiling point of  epichlorohydrin is  116.4°C,  and its vapor  pressure  is
                   •

 20  mm  Hg  at  29°C.   These  factors  contribute  to  the  rapid evaporation  of


 the chemical  upon release into the environment.


      Epichlorohydrin is a reactive molecule forming covalent bonds  with bio-


 logical macromolecules.   It  tends   to  react more  readily  with  polarized


 groups,  such  as sulfhydryl groups.


      The total U.S.  production for epichlorohydrin  was estimated  at 345 mil-


• lion pounds in 1973 (Oesterhof, 1975), with 160  million  pounds used as feed-


 stock for the  manufacture of glycerine   and  180 million pounds  used  in the


 production of epoxy resins.   Production levels  for the  year 1977  have been


 estimated  at  400 million pounds.


 II.  EXPOSURE


      A.  Water


         No ambient monitoring data  on  epichlorohydrin  are available  from


 which reliable conclusions on the potential exposure from drinking  water may


 be  made.  However,  if  a major release of epichlorohydrin were  realized, the


 chemical is stable enough to be transported significant  distances.   The rate


 of  evaporative loss would give  an estimated half-life of about two days for
                                                                          »

 epichlorohydrin  in  surface  waters (to  a depth  of   1m).  The  only  reported


 contamination of  a  public water supply  resulted from a  tank car derailment


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 and  subsequent spillage  of 20,000 gallons  (197,000 pounds) of epichlorohy-
 drin  at Point  Pleasant,  West Virginia  on January  23,  1978.   Wells at  the
 depth  of 25  feet were heavily  contaminated.  More  specific information  is
 not yet available.
     B.  Food
         Epichlorohydrin  is used as a cross-link  in molecular sieve  resins,
 which  are,  in turn, used in the treatment  of foods (21  CFR 173.40).  Food
 starch  may be  etherified with  epichlorohydrin,  not  to  exceed  0
 alone  or  in combination with propylene  oxide,  acetic anhyd           cc
 anhydride  (21 CFR 172.892).   No  data concerning concentrations of epichloro-
 hydrin in foodstuffs has been generated.
     C.  Inhalation
         Numerous  environmental  sources  of epichlorohydrin have been  identi-
 fied  (SRC,  1979).  Epichlorohydrin is released into the  atmosphere  through
 waste ventilation  processes  from a  number of industrial operations which re-
 sult in volatilization .of the chemical.  ' No quantitative monitoring informa-
 tion  is  available on  ambient epichlorohydrin  concentrations.   High concen-
 trations have been observed  in the  immediate  vicinity of a factory discharg-
 ing epichlorohydrin  into  the atmosphere,  but these  were  quickly despersed,
with no  detection of the  chemical  at distances greater than 600  M  (Fomin,
 1966).
 III.  PHARMACOKINETICS
     A.  Absorption                               .  .     .  .
         Absorption  of epichlorohydrin  in man  and  animals  occurs  via  the
 respiratory and gastointestinal  tracts,  and by  percutaneous absorption (U.S.
EPA,   1979).  Blood  samples  obtained  from rats  after  6  hours  exposure"  to
 (1Z;C)epichlorohydrin at  doses of 1 and 100 ppm  in air  revealed  0.46+0.19
 and 27.8+4.7 ;jg epichlorohydrin  per ml  of plasma, respectively.   The rates


-------
epichlorohydrin  per ml  of  plasma,  respectively.   The rates  of uptake  at
these exposure levels were  determined  as  15.48 and 1394 ug per hour, and the
dose received was 0.37 and 33.0 mg/kg (Smith, et al. 1979).
     B.  Distribution
         The  distribution of  radioactivity  in various tissues of  rats fed
(1ZlC)-epichlorohydrin has been examined  (Weigel,  et al.  1978).   The  chemi-
cal was  rapidly  absorbed with tissue  saturation  occurring within two  hours
in males  and four  hours in females.  The kidney and  liver  accumulated the
greatest amounts  of radioactivity.  Major  routes of  excretion were in the
urine (38  to 40 percent), expired air (18 to 20  percent), and the  feces  (4
percent).   The  appearance  of large  amounts  of  14C02  in expired air  sug-
gests a rapid and extensive metabolism of (^C)-epichlorohydrin in rats.
     C.  Metabolism
         Limited  data  concerning mammalian  metabolism of  epichlorohydrin
suggest  in_  vivo  hydrolysis  of   the  compound,  yielding  alpha-chlorohydrin
(Jones, et al.  1969).   Upon exposure to radioactively-labeled epichlorohy-
drin a  small percentage of the  radioactivity was  expired  as intact  epi-
chlorohydrin, while a large percentage of the radioactivity  was  excreted  as
  C02>   indicating  a  rapid   and  extensive  metabolism  of   the  (  C)epi-
chlorohydrin.  Metabolites  in the urine have been  obtained  by  these  re-
searchers ,  but the final  analysis as to the  identity  of the compounds is not
yet complete.  Van Ouuren (1977)  has  suggested a  metabolite  pathway of  epi-
chlorohydrin to include glycidol,  glycidaldehyde and epoxy-propionic  acid.
     0.  Excretion
         The percentages  of total radioactivity  recovered in the urine and
expired air as 14C02  were 46 percent and   33 percent in  the  1 ppm group,
and 54  percent and  25  percent  in the  100  ppm  group, respectively.    Rats

-------
orally  treated  with 100 mg/kg  excreted  51 percent  of  the administered  epi-
chlorohydrin  in the urine and  38  percent  in  expired air, while 7 to 10  per-
cent  remained in the body 72 hours after exposure.   Tissue accumulation of
radioactivity was highest in kidneys and liver.
IV.  EFFECTS
     A.  Carcinogenicity
         Epichlorohydrin appears  to have low carcinogenic activity following
dermal  application.   In  two  studies, epichlorohydrin  applied  topically to
shaved  backs  of rats  or mice  did not induce any  significant  occurrence of
skin  tumors  (Weil,  1964;  Van Ouuren, et  al. 1974).   However,  subcutaneous
injection of  epichlorohydrin  at levels as  low as 0.5 mg have resulted in the
induction of tumors at the injection site.
         Extensive  inhalation studies have recently identified epichlorohy-
drin as  a  potent nasal  carcinogen in rats.   At concentrations of  100 ppm,
significant increases  in the occurrence of  squamous cell carcinomas  of the
nasal  turbinatss have  been   observed.   Such  tumors have  been reported in
lifetime exposure studies at 30 ppm but not at 10 ppm (Nelson, 1977, 1978).
         Several recent  epidemiological  studies  have  suggested the  risk of
cancer as  a  result  of occupational epichlorohydrin  exposure.   Both respira-
tory cancers  and leukemia are  in excess  among  some exposed worker  popula-
tions,  but this increase was not shown  to be statistically  significant
(Enterline and Henderson, 1978; Enterline, 1979).   The  data  suggest a laten-
cy period  of  roughly  15 years  before  the  onset  of- carcinogenic symptoms.  A
second  survey  has  failed to  substantiate  these   findings  (Shellenberger, et
al. 1979),  However,  this survey  used a younger  study  population with less
                                                                         »
exposure to epichlorohydrin.

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     B.  Mutagenicity
         Epichlorohydrin  has been  shown  to cause  reverse  mutations in sev-
eral organisms (SRC, 1979).
         Cytogenetic  studies  with  experimental  animals  have  revealed  in-
creased aberrations  in animals treated with epichlorohydrin.   Both mice and
rats have  displayed dose-dependent  increases  in  abnormal chromosome morpho-
logy at exposure  levels  ranging  from 1  to  50  mg/kg (Santodonato,  et  al.
1979) .
                                                           •
         In  humans,  the clastogenic properties of  epichlorohydrin have been
reported in  workers occupationally  exposed to the chemical and in cultured
"normal"  lymphocytes  exposed  to  epichlorohydrin  (SRC,  1979).   Cytogenetic
        i
evaluation of 'exposed  workers  has shown an increase  of somatic cell chromo-
some aberrations  associated  with concentrations ranging  from  0.5  to 5.0 ppm
(2.0 to 20  mg/m3)  (SRC,  1979).   Such  chromosomal  damage appears  to  be  re-
versible once exposure to the chemical ceases.
     C.  Teratogenicity
         Pregnant rats  and rabbits exposed to 2.5  to 25  ppm epichlorohydrin
during days  6  to  15 or days 6  to 18 of -gestation  showed a mild teratogenic
response (John, et al.  1979).   However  examinations of all  fetal tissue have
not been completed.   The incidence  of resorbed  fetuses  was not  altered  by
exposure to epichlorohydrin at the doses employed.
     D.  Other Reproductive Effects
         The antifertility properties  of epichlorohydrin have  been examined
by several investigators.  Administration of 15 mg/kg/day of epichlorohydrin
for 12  days  resulted in reduced fertility of  male  rats (Halen,  1970) .   Five
                                                                         »
repeated doses of 20 mg/kg were more effective in rendering male rats  infer-
tile than  was one  100 mg/kg  dose or  five  50 mg/kg  doses (Cooper, et  al.


-------
1974).  The  suggested mode of action  of epichlorohydrin is  via the in^  vivo
hydrolysis of the compound  which produces  alpha-chlorohydrin.  Altered  re-
productive function  has been reported for workers  occupationally exposed  to
epichlorohydrin at concentrations less than 5 ppm.
     E.  Chronic Effects
         Two  species of rats  and one specie  of mice  (both  sexes)  were  ex-
posed to 5  to 50 ppm epichlorohydrin for six  hours per day,  five  days  per
week for a total of  65 exposures.   All species and  sexes displayed inflamma-
tory and  degenerative  changes  in nasal  tissue,  moderate  to severe tubular
nephrosis,   and gross  liver  pathology  at  50  ppm  exposure  (Quast,  et  al.
1979a).  The  same research  group has also  examined the  effect of  100  ppm
exposure for  12 consecutive  days.   The toxicity to nasal tissues was similar
(Quast, et al. 1979b).
         Altered  blood  parameters  (e.g.  increased  neutrophilic megamyelo-
cytes,  decreased  hemoglobin,  hematocrit, and  erythrocytes)  have been   ob-
          i
served  in rats exposed to 0.00955 to  0.04774  ml epichlorohydrin per kg body
weight  administered  intraperitoneally (Lawrence,  et al. 1972).   Lesions  of
the lungs and reduced weight gains were also observed.
     Toxicity studies with various animal species have  established that epi-
chlorohydrin  is  moderately  toxic  by  systemic  absorption  (Lawrence,  et al.
1972).  Acute oral  LDjg values  in  experimental animals  have  ranged frcm
155 to 238 mg/kg for the mouse and from  90 to 260 mg/kg in the rat.   Inhala-
tion LC5Q  values range from  360  to  635 ppm  in rats,  to- 800  ppm  in mice
(SRC, 1979).   Single subcutaneous injections of epich-lorohydrin in  rats  at
doses  of  150  or 180 mg/kg have  resulted  in  severe  injury to  the kidney
                                                                         •
(Rotara and Pallade,  1966).
                                     -  ) /....>•» / ^
                                    ^^^^7

-------
         Accidental human  exposures  have been .reviewed  (NIOSH,  1976; Santo-
donato,  et  al.  1979).  Direct  exposure to epichlorohydrin  vapor  results in
severe irritation of  the eyes and respiratory membranes, followed by nausea,
vomiting, headache,  dyspnea,  and altered liver  function.   A significant de-
crease was reported in  pulmonary  function  among  workers exposed to epichlor-
ohydrin  in an epoxy-resin  manufacturing process.  Workers were simultaneous-
ly exposed to dimethyl amino propylamine.
V.   AQUATIC TOXICITY
     Pertinent data could not be located in the available literature.
VI.  EXISTING GUIDELINES AND STANDARDS-
     Existing occupational  standards  for exposure  to  epichlorohydrin are re-
viewed in the  NIOSH  (1976) criteria  document.   The NIOSH  recommended  envi-
ronmental exposure  limit is a  2 mg/m3  10-hour  time-weighted average and  a
19 mg/m3 15-minute  ceiling concentration.   The  current  Occupational Safety
and Health Administration  standard is  an  8-hour time-weighted average  con-
centration of 5 ppm (20 mg/m-5).
                                 I oo-

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         l-CHLORO-2,3-EPOXYPROPANE(EPICHLOROHYDRIN )

                         REFERENCES  '

Cooper, E.R.A., et al.   1974.   Effects  of  alhpa-chlorohydrin
and related compounds on the reproduction  and  fertility  of
the male rat.  Jour. Reprod. Pert.   38:  379.

Enterline, P.E.  1979.   Mortality experience of  workers  ex-
posed to epichlorohydrin.  In press: Jour.  Occup. Med.

Enterline, P.E., and V.L. Henderson.  1978.  Communication  to
Medical Director of the  Shell Oil Company:  Preliminary  find-
ing of the updated mortality study among workers exposed to •
epichlorohydrin.  Letter dated  July  31,  1978.   Distributed  to
Document Control Office, Office of Toxic Substances  (WH-557)
U.S. Environ. Prot. Agency.

Fomin, A.P.  '1966.  Biological  effects  of  epichlorohydrin and
its hygienic significance as an atmospheric pollutant.   Gig.
Sanit.  31: 7.

Halen, J.D.  1970.  Post-testicular  antifertility effects of
epichlorohydrin and 2,3-epoxypropanol.  Nature   226:  87.

John, J.A., et al.  1979.  Epichlorohydrin-subchronic
studies. IV.  Interim results of a study of the  effects  of
maternally inhaled epichlorohydrin on rats', and  rabbits' em-
bryonal and fetal development.  Jan. 12, 1979.   Unpublished
report from Dow Chemical Co.  Freeport,  TX.

Jones, A.R., et al.  1969.  Anti-fertility  effects and metab-
olism of of alpha- and epichlorohydrin  in  the  rat.  Nature
24: 83.

Lawrence, W.H., et al.   1972.   Toxicity profile  of epichloro-
hydrin.  Jour. Pharm. Sci.  61: 1712.

Nelson, N.  1977.  Communication to  the  regulatory agencies
of preliminary findings of a carcinogenic  effect in  the  nasal
cavity of rats exposed to epichlorohydrin.  New  York  Univer-
sity Medical Center.  Letter dated March 28, 1977.

Nelson, N.  1978.  Updated communication to the  regulatory
agencies of preliminary  findings of  a -carcinogenic effect in
the nasal cavity of rats exposed to  epickloroh-ydrin.  New
York University Medical  Center.  Letter dated June 23, 1978.
                                            *•
NIOSH.  1976.  NIOSH criteria for a  recommended  standard:
Occupational exposure to epichlorohydrin.   U.S.  DHEW.  Na-
tional Institute for Occupational Safety and Health.
                             ^T-U^^J^^^—
                              }) / &

-------
Oesterhof, D.  1975.  Epichlorohydrin-.  Chemical Economics
Handbook.  642.302/A-642.3022.  Stanford Research Corp.,
Menlo Park, Calif.

Quast, J.F., et al.  1979a.  Epichlorohydrin - subchronic
studies.. I. A 90-day inhalation study in laboratory rodents.
Jan. 12, 1979.  Unpublished report from Dow Chemical Co.
(Freeport, TX).

Quast, J.F., et al.  1979b.  Epichlorohydrin - subchronic
studies.  II. A 12-day study in laboratory rodents.  Jan. 12,
1979.  Unpublished report from Dow Chemical Co.  Freeport,
TX.

Rotara, G., and S. Pallade.  1966.  Experimental studies of
histopathological features in acute epichlorohydrin
(l-chloro-2,3-epoxypropane) toxicity.  Mortal Norm. Patol.
11: 155.

SantodonatOj J., et al.  1979.  Investigation of selected
potential environmental contaminants: Epichlorohydrin and
epibromohydrin.  Syracuse Research Corp.  Prepared for Office
of Toxic Substances, U.S. EPA.

Shellenberger, R.J., et al.  1979.  An evaluation of the
mortality experience of employees with potential exposure to
epichlorohydrin.  Departments of Industrial Medicine, Health
and Environmental Research and Environmental Health.  Dow
Chemical Co.  Freeport, TX.

Smith, F.A., et al.  1979.  Pharmacokinetics of epichlorohy-
drin (EPI) administered to rats by gavage or inhalation.
Toxicology Research Laboratory, Health and Environmental
Science.  Dow Chemical Co., Midland, MI.  Sponsored by the
Manufacturing Chemists Association.  First Report.

Syracuse Research Corporation.  1979.  Review and evaluation
of recent scientific literature relevant to an occupational
standard for epichlorohydrin: Report prepared by Syracuse
Research Corporation for NIOSH..

Van Duuren, B.L.  1977.  Chemical structure, reactivity, and
carcinogenicity of halohydrocarbons.  Environ. Health Persp..
21: 17.

Van Duuren, B.L., et al.  1974.  Carcinogenic action of alky-
lating agents.  Jour. Natl. Cancer Inst.  53: 695.

Weigel, W.W., et al.  1978.  Tissue distribution and excre-
tion of (-^c)-epichlorohydrin in male and female rats.     *
Res. Comm. Chem. Pathol. Pharmacol.  20: 275.
                            /06-/J

-------
Weil,  C.S.  1964.  Experimental carcinogenicity and acute
toxicity of representative epoxides.  Amer.  Ind. Hyg. Jour.
24: 305.
                            too-ff

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                                      No. 101
         Ethyl Methacrylate

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is  drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources, this  short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by the
subject chemical.   This document  has undergone scrutiny to
ensure its technical accuracy.

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                              ETHYL METHACRYLATE
                                    Summary

     Information on  the carcinogenic and mutagenic effects  of ethyl methac-
rylate was  not found in  the  available literature.  Ethyl  methacrylate has,
however,  been shown to cause teratogenic effects in rats.
     Chronic occupational exposure to  ethyl methacrylate  has not  been re-
ported in the available literature.
     Data concerning  the effects of ethyl methacrylate  on  aquatic organisms
were not found in the available literature.

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                              ETHYL METHACRVLATE
I.   INTRODUCTION
     Ethyl  methacrylate  (molecular  weight  114.15)  is the  ethyl  ester of
methacrylic 'acid.   It is a  crystalline solid that melts  at  less than 75°C,
has a  boiling point of 117°C,  a  density of 0.9135,  and an index of refrac-
tion of  1.4147.   It  is  insoluble in water  at 25°C  and is infinitely solu-
ble in alcohol  and ether (Weast,  1975).   It possesses' a  characteristic un-
pleasant odor (Austian, 1975).
      •
     Widely known  as "Plexiglass" (in  the polymer- form),  ethyl methacrylate
is used  to  make polymers, which  in  turn are used  for building, automotive,
aerospace,  and  furniture  industries.  It is also  used by  dentists as dental
plates, artificial teeth, and orthopedic cement (Austian,  1975).
II.  EXPOSURE
     Ethyl methacrylate is used in large quantities and therefore has poten-
tial for  industrial  release  and  environmental contamination.   Ethyl methac-
rylate in the polymerized form  is not  toxic; however, chemicals used to pro-
duce ethyl  methacrylate  are  extremely  toxic.  No  monitoring  data are avail-
able to indicate ambient air or water levels of the compound.
     Human exposure  to ethyl methacrylate from foods cannot  be assessed due
to a lack of monitoring data.
     Bioaccumulation data on ethyl methacrylate were  not  found in the avail-
able literature.
III. PHARMACOKINETICS
     Specific  information on  the metabolism,  distribution,  absorption,  or
elimination of ethyl methacrylate was not found in the available literature.
                                       y

-------
     No evidence has been  found  of the presence of ethyl methacrylate in the
human urine.   Therefore,  it is  hypothesized that it  is  rapidly metabolized
and undergoes complete oxidation (Austian, 1975).
IV.  EFFECTS
     A.   Carcinogenicity and Mutagenicity
          Information  on  the  carcinogenic  and mutagenic  effects of  ethyl
methacrylate was not found in the available literature.
     B.   Teratogenicity
          Ethyl methacrylate is  teratogenic  in rats.   Female rats were given
intraperitoneal  injections  of 0.12 mg/kg,  0.24 mg/kg,  and 0.41  mg/kg,  on
days 5,  10,  and 15 of gestation.   These doses were 10,  20,  and 33 percent,
respectively,  of the acute  intraperitoneal  LD5g dose.   Animals were sacri-
ficed one day before parturition (day 20).
     Deleterious effects,  were observed  in the developing  embryo and fetus.
Effects were  compound  and generally  dose-related.   A 0.1223  ml/kg injected
dose resulted  in unspecified gross abnormalities and skeletal abnormalities
in 6.3 percent  and  5.0 percent  of the test  animals,  respectively, when com-
pared  to  the untreated controls.   A dose of 0.476 ml/kg  resulted in gross
abnormalities  in 15.7 percent  of  the treated  animals  and  skeletal abnor-
malities in 11.7 percent of the treated animals (Singh, et al. 1972).
     C.   Other Reproductive Effects and Chronic Toxicity
          Information on  other  reproductive effects and  chronic toxicity of
ethyl methacrylate was not found in the available -literature.
     D.   Acute Toxicity
          Lower molecular weight acrylic monomers such as ethyl methacrylate
                                                                       •
cause  systemic  toxic effects.   Its administration  results in  an immediate

-------
increase in respiration rate, followed by a decrease  after  15-40 minutes.   A
prompt  fall  in  blood pressure  also occurs,  followed  by  recovery  in  4-5
minutes.   As  the  animal  approaches death, respiration  becomes  labored  and
irregular, lacrimation  may  occur,  defecation  and urination  increase,  and
finally reflex  activity  ceases,  and the  animal lapses into a coma and dies
(Austian,  1975).
          Acrylic  monomers  are  irritants to  the skin and  mucous membranes.
When placed in  the eyes of  animals,  they  elicit a very  severe  response and,
if not washed out, can cause permanent damage  (Austian, 1975).
          As  early  as 1941, Qeichmann  demonstrated  that injection  of 0.03
cc/kg  body weight  ethyl methacrylate  caused  a  prompt and  sudden  fall  in
blood pessure,  while  respiration  was stimulated immediately and  remained  at
this level for  30 minutes.  The  final  lethal dose (0.90-.12 cc/kg) brought
about  respiratory  failure,  although the hearts  of these animals were  still
beating (Qeichmann, 1941).
          Work  by  Mir,  et  al.  (1974)  demonstrated that respiratory  system
effects alone may not kill the animal, .but  that cardiac  effects may also
contribute to  the cause  of  death  (Austian,   1975).   Twelve  methacrylate
esters and methacrylic  acid were tested  on  isolated  perfused  rabbit  heart.
Concentrations  as  low as  1 part in  100,000  (v/v) produced significant  ef-
fects.   The effects were  divided into three  groups  according  to  the  rever-
sibility of the heart response.  Ethyl methacrylate was  placed  in "Group  1",
in  which  the   heart  response  is   irreversible  at  all  concentrations
                                                 •*     *'».
(1:100,000;  1:10,000;  1:1,000).   Five  percent (v/v)   caused  a   41.2 percent
decrease in the heart rate  of isolated  rabbit  heart.   The same  concentration
                                                                       •
reduced heart contraction by 64  percent and  coronary flow by  61.5 percent
(Austian,  1975).
                                         f
                                  -V7 $$-.
                                  ltt-6

-------
          The  findings  of Deichmann  (1941)  that ethyl  methacrylate  affects
blood  pressure  and  respiration  is  substantiated  by  studies  of  Austian
(1975).  Response  following  administration of ethyl  methacrylate was  charac-
terized by a biphenic response, an abrupt fall  in blood  pressure followed by
a more  sustained  rise.   Austian (1975)  also  found that  the respiration rate
is increased,  the duration of  effect being  approximately  20  minutes,  after
which time the respiration rate returned to normal.
          In the  available literature LO^Q  values were  found for  only rab-
bit and  rat;  these were  established  by Deichmanrr in  1941.  The oral  value
for the  rat is 15,000  mg/kg,  as opposed to 3,654-5,481 mg/kg  for the rab-
bit.   Inhalation values for the rat have been reported to be  3,300  ppm  for 8
hours  (Patty,  1962).   Deichmann also  established a  skin toxicity LD5Q  for
rabbit which was  greater than  10 ml/kg.  This  was substantiated by  another
test  which  showed  that moderate skin  irritation (in  rabbits)  does  result
from ethyl methacrylate exposure (Patty, 1962).
VI.   EXISTING GUIDELINES AND STANDARDS
     Information  on existing guidelines and  standards was  not  found in  the
available literature.
                                    11 r>7-
                                 * ff u i
                                        -7

-------
                              ETHYL METHACRYLATE
                                  References
Austian,  J.   1975.   Structure-toxicity  relationships  of  acrylic monomers.
Environ. Health Perspect.  19: 141.
Oeichmann, w.   1941.  Toxicity  of methyl, ethyl,  and  n-outyl methacrylate.
Jour. Ind. Hyg.  Toxicol.  23: 343.
Mir, G., et al.   1974.   Journal of toxicological and pharmacological actions
of  methacrylate  monomers.  III.   Effects  on  respiratory  and cardiovascular
functions of anesthetized dogs.-. Jour.-Pharm. Sci.  63: 376.
Patty,  F.A.   1962.   Industrial  Hygiene  and  Toxicology,  Vol.   II.    Inter-
science Publishers, New York.             .  ..
Singh, A.R., et al.  -1972.   Embryo-fetal toxicity and teratogenic effects of
a group of methacrylate esters in rats.  Tox. Appl. Pharm. 22: 314.
Weast,  R.  C.   1975.   Handbook  of Chemistry  and Physics.   56th  ed.  CRC
Press, Cleveland,  Ohio.

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                                      No. 102
           Ferric Cyanide

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is  drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources, this  short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                                 FERRIC CYANIDE
 I.    INTRODUCTION
      Ferric  cyanide is a  misnomer and is not  listed as a specific compound
 in  the comprehensive compendia of inorganic  compounds (Weast, 1978).  There
 are,  however,  a class of  compounds  known as "iron cyanide blues" consisting
 of  various  salts  where  the anions  are  the ferricyanide,  [FeCCN)^]3-,  or
 the   ferrocyanide,   [Fe(CN)6]4-,  and  the  cations  are  either  Fe(III)  or
 Fe(II)  and  sometimes  mixtures  of Fe(II)  and  potassium  (Kirk  and  Othmer,
 1967).   The  empirical  formula of  the  misnamed  ferric  cyanide,  Fe(CN),
 corresponds  actually to one  of  the ferricyanide compounds,  the ferric ferri-
 cyanide  with the  actual   formula  Fe[Fe(CN)6]}  also  known  as  Berlin  green.
 The  acid from  which these  salts  are derived  is  called  ferricyanic  acid,
 H3[Fe(CN)g]   (also  known  as  hexacyanoferric   acid),   molecular   weight
 214.98,  exists  as  green-blue deliquescent needles, decomposes upon  heating,
 and  is soluble in water  and alcohol.   In this EPA/ECAO  Hazard Profile only
 ferric     ferricyanide,      Fe[Fe(CN)6]?      and     ferric     ferrocyanide,
'^[^(CN)^]},   are   considered;  other   ferrocyanide  compounds  are   re-
 ported in a separate EPA/ECAO Hazard Profile  (U.S. EPA, 1980).
     These compounds are colored pigments, insoluble  in water or weak acids,
 although they  can  form colloidal  dispersions in aqueous media.   These  pig-
ments  are  generally  used  in  paint,  printing inks, carbon  paper  inks,  cray-
 ons,  linoleum,  paper pulp, writing inks and  laundry  blues.   These compounds
 are sensitive to alkaline decomposition (Kirk and Qthmer, 1967).
 II.  EXPOSURE
     Exposure to these  compounds may occur occupationally  or through  inges-
tion  of  processed  food  or contaminated water.   However,  the  extent of  Food
or water contamination  from  these compounds has  not been described in  the

-------
   available  literature.   Prussian  blue,  potassium  ferric  hexacyanoferrata
   (II),  has  been  reported as  an  antidote against  thallium  toxicity.   When
   administered at a  dose of 10 g twice  daily  by duodenal  intubation,  it pre-
   vents the intestinal reabsorption of thallium (Dreisbach, 1977).
   III.  PHARMACOKINETICS
        A.   Absorption and Distribution
             Pertinent data could not be located in the available literature.
        B.   Metabolism
             There is  no apparent metabolic alteration, of  these compounds.   As
   for the other  ferrocyanide  and ferricyanide  salts,  these compounds  are  not
   cyanogenic (Gosselin, et al.  1976).
        C.   Excretion
             No information  is  available  for  ferric hexacyanoferrates  (II)  or
   (III),  but information  is available for other related ferrocyanide  and fer-
   ricyanide salts  (U.S. EPA,  1980; Gosselin,  et al.  1976)  which seems  to  be
'   rapidly excreted in urine apparently without  metabolic alteration.
   IV.  EFFECTS
        A.   Carcinogenicity,  Mutagenicity,  Teratogenicity,  Chronic  Toxicity,
             and Other Reproductive Effects
             Pertinent data could not be located in the available literature.
        8.   Acute Toxicity
             No adequate toxicity data are  available.   All  ferrocyanide  and
   ferricyanide salts  are reported as  possibly moderately  toxic  (from 0.5  to
   5.0 mg/kg  as a  probable lethal dose  in  humans)  (Gosselin, 'et  al.  1976).
   V.    AQUATIC TOXICITY
        Pertinent  data could not be located in the available literature  regard-
                                                                            *
   ing the aquatic toxicity of ferric cyanide.


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VI.  EXISTING GUIDELINES AND STANDARDS


     Pertinent data could not be located  in the available literature.
                                   i >n ~> -
                                 "lit'-'

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                                  REFERENCES
Oreisbach, R.H.   1977.   Handbook  of Poisoning, 9th  edition.   Lange Medical
Publications, Los Altos,  CA.

Gosselin, R.E.:,  et al.  1976.  ,Clincial  Toxicology  of Commercial Products,
4th edition.   Williams and Wilkins,  Baltimore,  Maryland.

Kirk,R.E.  and O.F.  Othmer.   1967.  Kirk-Othmer  Encyclopedia of  Chemical
Technology,  II edition,  Vol.  12.   Intersciencs Publishers, div.  John  Wiley
and Sons, Inc., New York.

U.S. EPA.  1980.   Environmental Criteria and  Assessment Office.   Ferrocya-
nide: Hazard  Profile.  (Draft)

Weast, R.C.   1978.  Handbook of Chemistry and  Physics,  58th ed.  The Chemi-
cal Rubber Company, Cleveland,  Ohio.
                                           ,
                                  /OSL.-6

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                                      No. 103
         Fluoranthene
  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is  drawn chiefly
from secondary  sources  and  available reference documents.
Because of the limitations of such sources, this  short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.
                          — ) / o s—
                           ! I I *

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                         FLUORANTHENE


                           SUMMARY


     No  direct carcinogenic  effects  have  been  produced by


fluoranthene  after  administration   to  mice.    The  compound


has  also failed  to  show  activity   as  a tumor  initiator or


promoter.    However,  it  has  shown cocarcinogenic  effects


on  the  skin of mice when combined  with benzo(a)pyrene, in-
                          •

creasing tumor incidence and decreasing  tumor latency.


     Fluoranthene  has not  shown mutagenicr  teratogenic or ..


adverse reproductive effects.


     Daphnia magna appears to have low sensitivity to fluoran—


thene with  a  reported  48-hour  EC5Q of  325,000  ug/1.   The '•
                                                               »

bluegill, however,  is  considerably  more  sensitive  with an


observed  9-6-hour   LC5Q  value  of 3,980.    The  96-hour  LC5Q


for  mysid  shrimp  is 16  ug/1,. and  a  reported  chronic value


is  16  ;ug/l.   Observed  96-hour   EC^Q values  based on  cell


numbers for  fresh and saltwater algae are over 45,000 ug/1.
                         ^^^^^_^^^^^^^^
                         -Hi) -

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                         FLDORANTHENE
I.   INTRODUCTION
     This  profile  is  based  on  the  Ambient  Water Quality
Criteria Document for Fluoranthene  (U.S. EPA, 1979).
     Pluoranthene  (1,2-benzacenapthene,  M.W.  202)  is a  poly-
nuclear  aromatic  hydrocarbon  of  molecular  formula C]_gH^Q.
Its physical  properties include:  melting point, 111°C;  boil-
ing  point,  375°C;  water  solubility,  265 pg/1  (U.S.   EPA,
1978).
     Fluoranthene  is  chemically  stable,  but may  be removed
from  water by  biodegradation  processes  (U.S. EPA,  1979).
The  compound  is  relatively  insoluble  in aqueous  systems.
Fluoranthene  may  be adsorbed and concentrated  on  a variety
of particulate matter.   Micelle formation through the action
of  organic  solvents  or  detergents  may  occur.  (U.S.   EPA,
1979).                                       ,
     Flouranthene  is  produced  from  the  pyrolytic  processing
of coal  and petroleum  and  may  result  from  natural biosyn-
thesis (U.S.  EPA, 1979).
II.  EXPOSURE
     Fluoranthene  is  ubiquitous  in  the environment;  it has
been monitored  in food, water,  air, and  in  cigarette  smoke
(U.S.  EPA,  1979).   Sources  of contamination  include indus-
trial  effluents   and  emissions,  sewage,  soil   infiltration,
and  road  runoff  (U.S.  EPA,  1979).   Monitoring of  drinking
                                                          »
water  has   shown  an  average fluoranthene concentration of
27.5 ng/1  in positive  samples  (Basu,  et  al.   1978).    Food

-------
levels  of  the  compound  are  in  the ppb  range,  and will  in-

crease  in  smoked or  cooked foods  (pyrolysis  of fats)  (U.S.

EPA,  1979).   Borneff  (1977)  has estimated  that dietary  in-

take  of  fluoranthene occurs mainly  from fruits, vegetables,

and bread.

     An  estimated daily  exposure  to fluoranthene  has  been

prepared by EPA  (1979):
                                            •
               Source         Estimated  Exposure

               Water          0.017 ug/day

               Food           1.6 - 16 ug/day

               Air            0.040 - 0.080 ug/day

     Based  on the octanol/water partition  coefficient,  the

U.S.  EPA  (1979)  has  estimated  weighted  average bioconcen-

tration  factor  of 890 for  fluoranthene  for  the edible  por-

tion of fish and shellfish consumed by Americans.

III. PHARMACOKINETICS

     A.   Absorption

          Based  on  animal  toxici.ty  data  (Smythe,  et  al.

1962)  ,  fluoranthene  seems  well  absorbed following  oral or

dermal  administration.   The  related   polynuclear  aromatic

hydrocarbon (PAH), benzo(a)pyrene, is readily absorbed  across

the lungs  (Vainio, et al.  1976).

     B.   Distribution

          Pertinent  information  could   not  be  located  in
                                                           »
the available literature.   Experiments  with benzo(a)pyrene

indicate  localization in  a  wide variety  of  body  tissues,

primarily in body fats (U.S. EPA, 1979).

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     C.   Metabolism
          Pertinent  information  could   not   be   located   in
the  available  literature.   3y analogy  with  other  PAH com-
pounds,  fluoranthene  may  be  expected  to undergo metabolism
by  the  mixed function  oxidase enzyme  complex.    Transforma-
tion products  produced by this action  include ring hydroxy-
lated  products  (following epoxide  intermediate formation)
and  conjugated forms  of  these  hydroxylated  products  (U.S.
EPA, 1979).
     D.   Excretion
          Pertinent  information  could   not   be   located   in
the  available  literature.   Experiments  with  PAH compounds
indicate  excretion  through the hepatobiliary  system and  the -1
feces; urinary excretion  varies with  the degree of formation
of conjugated metabolites  (U.S. EPA,  1979).
                                                               i
IV.  EFFECTS
     A.   Carcinogenicity
          Testing of fluoranthene  in  a marine carcinogenesis
bioassay  failed  to  show  tumor  production  following  dermal
or  subcutaneous  administration  of  fluoranthene  (Barry,   et
al., 1935).
          Skin testing of fluoranthene  as a  tumor  promoter
or  initiator  in mice  has also  failed  to show  activity   of
the  compound  (Hoffman,  et al.,  1972;  Van Duuren and Gold-
schmidt, 1976).
          Fluoranthene  has been  demonstrated  to have  car-
cinogenic  activity  (Hoffmann  and   Wynder,  1963;  Van  Duuran

-------
and Goldschmidt, 1976).  The combination of  fluoranthene


and  benzo(a)pyrene produced  an  increased  number  of papil-


lomas  and  carcinomas,  with  shortened  latency  period   (Van


Duuren and Goldschmidt, 1976).


     B.   Mutagenicity


          Fluoranthene  failed  to  show  mutagenic   activity


in the Ames Salmonella assay in the  presence  of  enzyme activa-


tion mix  (Tokiwa, et al.  1977; La Voie, et  al.   1979).


     C.   Teratogenicity


          Pertinent  information  could  not be  located   in


the  available  literature.   Certain PAH  compounds  (7,12-di-


methylbenz(a)anthracene  and  derivatives)   have  been  shown


to produce  teratogenic effects  in  the rat  (Currie,  et al.


1970; Bird, et al.  1970).


     D.   Other Reproductive Effects


          Pertinent  information  could  not be  located   in


the available literature.


     E.   Chronic Toxicity


          Pertinent  information  could  not be  located   in


the available literature.


V.   AQUATIC TOXICITY


     A.   Acute Toxicity


          The  96-hour  LC5Q  value  for the. bluegill, Lepomis


macrochiruss is  reported  to be 3,980 ^ug/1 '(U.S.  EPA, 1978).

The  sheepshead  minnow?  Cyprinodon  variegatus^  was exposed
                                                           •

to concentrations  of  fluoranthene  as high  as  560,000  ug/1


with  no  observed LC5Q  value  (U.S.  EPA,  1978)  .   The fresh-

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water  invertebrate   Daphnia  magna  appears   to  have  a  low
sensitivity  to  fluoranthene  with  a  reported  48-hour  EC-g
value of 325,000  ug/1.   The 96-hour LC5Q value for  the salt-
water raysi'd shrimp, Mysidopsis bahia } is  16 ug/1.
     B.   Chronic. Toxicity
          There  are  no  chronic  toxicity data  presented  on
exposure of  fluoranthene to  freshwater  species.   A  chronic
value foe the mysid shrimp  is  16 pg/L.
     C.   Plant Effects
          The  freshwater   alga,  Selenastrum  capricornutum,
when  exposed  to  fluoranthene resulted  in  a  96-hour  ECeQ
value for cell number of 54,400 jag/1.   On the  same criterion,
the  96-hour   ECSO  value  for  the  marine  alga,  Skeletonema
costatum, is 45,600 ug/1 (U.S. SPA, 1979).
     D.   Residues
                         »
          No  measured  steady-state  bioconcentration  factor
(BCF)  is  available  for  fluoranthene.    A  3CF of  3,100  can
be  estimated  using  the  octanol/water  partition coefficient
of 79,000.
VI.  EXISTING GUIDELINES AND STANDARDS
     A.   Human
          The World Health Organization (1970)  has established
a  recommended standard  of   0.2  jug/1  for all--2AH  compounds
in drinking water.
          Based on the no-effect level  determined in  a  single
                                                           »
animal  study  (Hoffman,  et  al.  1972),  the U.S.  EPA  (1979)
has  estimated a  draft  ambient water criterion of  200 ;ag/l
for  fluoranthene.    However,  the  lower  level  derived  for
                             A

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total PAH compounds is expected to have precedence for  fluor-
anthene.
     B.   Aquatic
         'For  fluoranthene,  the  draft criterion  to protect
freshwater  aquatic  life  is  250 pg/1  as  a  24-hour  average,
not  to  exceed  560  ug/1  at  any  time.   For  saltwater  life,
the  criterion  is  0.30  ug/1  as  a  24-hour  average, not to
exceed 0.69 ug/1 at any time.

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                        FLUOROANTHENE

                         REFERENCES

Barry, G., et al.  1935.  The production of cancer  by  pure
hydrocarbons-Part III.  Proc. Royal Soc., London.   117:  318.

3asu, O.K., et al.  1978.  Analysis of water samples for
polynuclear aromatic hydrocarbons.  U.S. Environ. Prot.
Agency, P.O. Ca-8-2275B, Exposure Evaluation Branch, HERL,
Cincinnati, Ohio.

Bird, C.C., et al.  1970.  Protection from the embryopathic
effects of 7-hydroxymethyl-12-methylbenz(a)anthracene  by
2-methyl-I, 2-bis-{3 pyridyl)-l-propanone(metopirone ciba)
and£ -diethylaminoethyldiphenyl-n-propyl acetate  (SKR  525-A).
Br. Jour. Cancer^ 24: 348.

Borneff, J.  1977.  Fate of carcinogens in aquatic  environ-
ment.  Pre-publication copy received from author.

Currie, A.R., et al.  1970.  Embryopathic effects of 7,12-
dimethylbenz(a)anthracene and its hydroxymethyl derivatives
in the Sprague-Dawley rat.  Nature  226: 911.

Hoffmann, D., and E.L. Wynder.  1963.  Studies on gasoline
engine exhaust.  Jour. Air Pollut. Control Assoc.   13: 322.

Hoffmann, D., et al.  1972.  Fluoranthenes: Quantitative de-
termination in cigarette smoke, formation by pyrolysis,  and
tumor initiating activity.  Jour. Natl. Cancer Inst.   49:
1165.

La Voie, E., et al.  1979.  A comparison of the mutagenicity,
tumor initiating activity and complete carcinogenicity of
polynuclear aromatic hydrocarbons,  ^n:Polynuclear Aromatic
Hydrocarbons.  P.W. Jones and C. Leber (eds.).  Ann Arbor
Science Publishers, Inc.

Smythe, H.F., et al.  1962.  Range-finding toxicity data:
List VI. Am. Ind. Hyg. Assoc. Jour.  23: 95.

Tokiwa, H., et al.  1977.  Detection of mutagenic activity in
particullate air pollutants..  Mutat. Res.   48: 237.

U.S. EPA.  1978.  In-depth studies on health '&nd environmen-
tal impacts of selected water pollutants.  -U.S. Environ.
Prot. Agency.  Contract No. 68-01-4646.

U.S. EPA.  1979.  Fluoranthene: Ambient Water Quality Cri7
teria. (Draft).

-------
Vainio, H., et al.  1976.  The fate of  intratracheally  in-
stalled benzo(a)pyrene in the isolated  perfused  rat  lung  of
both control and 20-methylcholanthrene  pretreated  rats.   Res,
Commun. Chem. Path. Pharmacol.  13: 259.

Van Duuren, B.L., and B.M. Goldschmidt.  1976.   Cocarcino-
genic and tumor-promoting agents  in tobacco carcinogenesis.
Jour. Natl. Cancer Inst.  51: 1237.

World Health Organization.  1970.  European standards for
drinking water, 2nd ed., Revised^ Geneva.

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                                 No. 104
           Formaldehyde

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
      WASHINGTON, D.C.  20460

          APRIL 30, 1980
        soy-/

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                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is  drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources, this  short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

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                           FORMALDEHYDE




SUMMARY


     The ma.jor source of  formaldehyde  contamination in the envi-


ronment is combustion processes,  especially  automobile emissions.


Formaldehyde is a recognized  component of  photochemical smog.   A


recent source of concern  is the release of formaldehyde from


resins used in home construction  and insulation.


     Bioaccumulation of formaldehyde is considered  unlikely due


to its high chemical reactivity.  Formaldehyde  rapidly degrades


in the atmosphere by photochemical processes? however,  it can


also be formed by the photochemical oxidation of  atmospheric


hydrocarbons.


     Formaldehyde is rapidly  absorbed  via  the lungs or gut? fol-


lowing- absorption into the blood, however, formaldehyde dis-


appears rapidly due to reactions  with  tissue components and


because of its metabolism.


     The U.S. EPA1s Carcinogen Assessment  Group recently con-


cluded that "there is substantial evidence that formaldehyde is


likely to be a human carcinogen." This finding was based on pre-


liminary results from a chronic inhalation study  of formaldehyde


which reported carcinomas of  the  nasal cavity in  3  rats after  16


months of exposure.  This type of tumor is extremely rare is


unexposed rats of the strain  used in the study.


     There is an extensive data base showing that formaldehyde is


rautagenic in microorganisms,  plants, insects, cultured mammalian


cells, and mice.  It was  negative in a teratogenicity assay.


Formaldehyde is known to  be a mucous membrane irritant in humans?
                              -' ^ * &
                              / & * ^

-------
it is also known to be an allergen  in  sensitive  individuals.



I.   INTRODUCTION

     This profile  is based on  a U.S. EPA  report  entitled  "Inves-

tigation of Selected Potential Environmental  Contaminants:

Formaldehyde"  (1976) and other selected references.

     Formaldehyde  (HCHO; molecular  weight 30.03)  is  a  colorless
        •
gas having a pungent odor and  an  irritating effect on  mucous  mem-

branes.  It has the following  physical/chemical  properties  (U.S.

EPA, 1976; Windholz, 1976):

          Boiling  Point:          -19.2°C

          Melting  Point:          -92°C

          Density  in Air:         1.067

          Solubility:             soluble in  water and many

                                  organic solvents.

     A  review  of the production range  (includes  importation)

statistics for formaldehyde  (CAS  No. 50-00-0) which  is listed in

the initial TSCA Inventory (1979a)  has shown  that between 2 bil-

lion and 7 billion pounds of this chemical were  produced/imported

in 1977 JL/

     Formaldehyde  is usually sold as an aqueous  solution  contain-

ing 37% formaldehyde by weight; it  is  also available as a linear
—/ This production  range  information does  not  include  any    .
   production/importation data claimed as  confidential by  the
   person(s) reporting  for  the TSCA Inventory,  nor  does  it
   include any information  which would compromise Confidential
   Business Information.  The data submitted for the TSCA
   Inventory, including production range information,  are  subject
   to the limitations contained in the Inventory Reporting
   Regulations (40  CFR  710).

-------
polymer known as parafonnaldehyde and a cyclic  trimer  known  as


trioxane.  Formaldehyde is used in the production  of urea-formal-


dehyde resins, phenol-formaldehyde resins, polyacetal  resins,


various other resins, and as an intermediate in the production  of


a variety of chemicals.  Manufacture of resins  consumes  over 50%


of annual domestic formaldehyde production.  Urea-formaldehyde


and phenol-formaldehyde resins are used as adhesives for particle


board and plywood, and in making foam insulation.  Polyacetal


resins are used to mold a large variety of plastic parts for


automobiles, appliances, hardware, and so on (U.S. EPA,  1976).




II.  EXPOSURE


     HIOSH (1976) estimates that 1,750,000 workers are poten-


tially exposed to formaldehyde in the workplace.


     A.   Environmental Fate

     Formaldehyde and nascent forms of formaldehyde can  undergo


several types of reactions in the environment including  depoly-


merization, oxidation-reduction, and reactions  with other


atmospheric and aquatic pollutants.  Formaldehyde  can  react


photochemically in the atmosphere to form H and HCO radicals;

once formed, these radicals can undergo a wide  variety of


atmospheric reactions (U.S. EPA, 1976).  Hyj3rogen  peroxide can


also be formed during photodecomposition of formaldehyde (Purcell


and Cohen, 1967; Bufalini jt_ al., 1972).  The atmospheric half-
                                                             »
life of formaldehyde is less than one hour in sunlight (Bufalini


et al., 1972).

-------
     Even though formaldehyde is often used as a bacteriocide,



there are some microorganisms which can degrade the chemical



(U.S. EPA, 1976).  Kamata (1966) studied biological degradation



of formaldehyde in lake water.  Under aerobic conditions in the



laboratory, known quantities of formaldehyde were decomposed  in



about 30 hours at 20"C; anaerobic decomposition took about 48



hours.  No decomposition was noted in sterilized lake water.



     Paraformaldehyde slowly hydrolyzes and depolymerizes as  it



dissolves in water to yield aqueous formaldehyde.  Trioxane has



more chemical and thermal stability? it is inert under aqueous



neutral or alkaline conditions.  In dilute acid solutions, it



slowly depolymerizes (U.S. EPA, 1976).



     B.   Bioconcentration



     Formaldehyde is a natural metabolic product and does not



bioconcentrate (U.S. EPA, 1976).



     C.   Environmental Occurrence



     Environmental contamination from formaldehyde manufacture



and industrial use is small and localized compared with other



sources.  Combustion and incineration processes comprise the



major sources of formaldehyde emissions.  Stationary sources of



formaldehyde emissions include power plants, manufacturing facil-



ities, home consumption of fuels, incinerators, and petroleum



refineries.  Mobile sources of formaldehyde emissions include



automobiles, diesels, and aircraft.  The automobile, however,, is



the largest source of formaldehyde pollution.  It is estimated



that over 800 million pounds of formaldehyde were released to the



air in the United States in 1975; of this amount, over 600
                              -ifTtf •—
                              I Wl

-------
million pounds are estimated to result  from  the  use of automo-


biles.  In addition to formaldehyde, automobile  exhaust also
             ,                                        .-•-*•
contains large quantities of hydrocarbons.   Through photochemical

processes in the atmosphere, these hydrocarbons  are oxidized to

formaldehyde, among other things, further  adding to the environ-

mental load of formaldehyde (U.S. EPA,  1976).

     Urea-formaldehyde foam insulation  has been  implicated as a


source of formaldehyde fumes in homes insulated  with this

material.  Wood laminates (plywood, chip board,'  and particle

board) commonly used in the construction of  mobile  homes are also

known to release formaldehyde vapors into  the home  atmosphere

(U.S. EPA, 1979b).




III. PHARMACOKINETICS

     A.   Absorption

     Under normal conditions formaldehyde  can enter the "body

through dermal and occular contact, inhalation and  ingestion.   On

dermal contact, formaldehyde reacts with proteins of the skin

resulting in crosslinking and precipitation  of the  proteins.

Inhalation of formaldehyde vapors produces irritation and

inflammation of the bronchi and lungs;  once  in the  lungs,

formaldehyde can be absorbed into the bloodi  Ingestion of

formaldehyde is followed immediately by inflammation of the

mucosa of the mouth, throat, and gastrointestinal tract (U.S.
                                                              •
EPA, 1976).  Absorption appears to occur in  the  intestines

(Malorny _et_ _al_., 1965).

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     B.   Distribution


     Following absorption into the blood  stream,  formaldehyde


disappears rapidly due to condensation reactions  with  tissue


components and oxidation to formic acid  (U.S.  EPA,  1976).


     C.   Metabolism


     The main metabolic pathway for  formaldehyde  appears  to


involve initial oxidation to formic  acid,  followed  by  further


oxidation to CO2 and I^O.  In rats fed radiolabeled formaldehyde,


40% of the radiolabel was recovered  as respiratory  COj  (Buss et


al., 1964).  Liver and red blood cells appear  to  be the major


sites for the oxidation of formaldehyde to  formic acid  (U.S.  EPA,


1976; Malorny et_ _al_., 1965).


     D.   Excretion


     Some of the formic acid metabolite is  excreted in  the  urine


as the sodium salt; most, however, is oxidized to CO- and


eliminated via the lungs (U.S. EPA,  1976).




IV.  HEALTH EFFECTS


     A.   Carcinogenicity


     Watanabe et^ al. (1954) observed sarcomas  at  the site of


injection in 4 of 10 rats given weekly subcutaneous  injections of


formaldehyde over 15 months (total dose 260 mg per  rat).  Tumors


of the liver and omentum were reported in two "'other  rats.   The


authors do not mention any controls.
                                                             »

     Groups of mice were exposed to  formaldehyde  by  inhalation at


41 ppm and 81 ppm for one hour a day thrice weekly  for  35 weeks.


After the initial 35-week exposure to 41 ppm, the mice  were

-------
exposed for an additional 29 weeks at  122 ppm.   No  tumors  or


metaplasias, were found, although numerous changes were  observed


in respiratory tissues  (Horton et_ ^1^.,  1963).   The  study is


considered flawed for several reasons:  mice were not observed


for a lifetime; survival was poor; many tissues were not examined


histologically (U.S. EPA, 1976; U.S. EPA, 1979b).


     In a lifetime inhalation study of the  combination  of  hydro-


chloric acid (10.6 ppm) and formaldehyde  (14.6  ppm) vapors in


rats, 25/100 animals developed squamous cell carcinomas of the


nasal cavity (Nelson, 1979).  Nelson also reported  that bis-


chloromethyl ether, a known carcinogen, was detected in the


exposure atmosphere; however, concentrations were not reported.


     In a report of interim results (after  16 months of a  2-year


study) from a chronic inhalation study of formaldehyde  in  rats


and mice, the Chemical  industry Institute of Toxicology (1979)


reported that squamous  cell carcinomas  of the nasal cavity were


observed in three male  rats exposed to IS ppm of formaldehyde


(highest dose tested).  This type of tumor  is extremely rare in


unexposed rat of the strain used in this  study.


     Following receipt  of the CUT (1979) study, the U.S.  EPA's


Carcinogen Assessment Group (1979c) concluded that  "there  is


substantial evidence that formaldehyde  is likely to be  a human


carcinogen."  The unit  risk calculation (the lifetime cancer risk


associated with continuous exposure to 1 ug/m   of formaldehyde)


based on the preliminary results from  CUT  is estimated to be 3.4

   — 5
xJ-°  .  This estimate may change when  the final results of "the

CUT study become available.

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     B.   Mutagenicity


     There is an extensive data base showing that  formaldehyde  is


mutagenic in several species including mice, Drosophila, plants,


Saccharomyces cerevisiae, Neurospora Crassa, and several species


of bacteria.  Formaldehyde also produced unscheduled DNA syn-


thesis in a human cell line.  These and other early reports  of


mutagenic activity have been reviewed by Auerbach  et al. (1977)


and U.S. EPA (1976).


     Reports in the recent literature have  supported the finding


that formaldehyde is a mutagen:  Magana-Schwencke £t_ ^1_. (1978)


in a study with S. cerevisiae; Wilkens and  MacLeod  (1976)  in


E. coli; Martin ^t^ _al_. (1978) in an unscheduled DNA synthesis


test in human HeLa cells? Obe and Beek (1979) in sister chromatid


exchange assays in a Chinese hamster ovary  (CHO) cell line and  in


cultured human lymphocytes.


     C.   Teratogenicity

     Formaldehyde has been found negative in teratogenicity

assays in beagle dogs (Hurni and Ohden, 1973) and  rats  (Gofmekler


and Bonashevskaya, 1969).


     D.   Other Reproductive Effects


     No changes were observed in the testes of male rats exposed


to air concentrations of 1 mg/m  of formaldehyde.for 10 days


(Gofmekler and Bonashevskaya, 1969).

     E.   Other Chronic Toxicity
                                                             •
     Groups of rats, guinea pigs, rabbits, monkeys, and dogs were

continuously exposed to approximately 4.6 mg/m3 of formaldehyde


for 90 days.  Hematologic values were normal, however, some

-------
interstitial inflammation occurred  in  the  lungs  of all species



(Coon _et_ ai^., 1970).



     P.   Other Relevant Information



     Formaldehyde vapor is quite irritating  and  is a  major cause



of the mucous membrane irritation experienced  by people exposed



to smog.  Dermatitis from exposure  to  formaldehyde is a common



problem in workers and consumers who contact the chemical



regularly.  Formaldehyde is also known to  be an .allergen in



sensitive individuals (U.S. EPA, 1976).







V.   AQUATIC EFFECTS



     The use of formalin (aqueous formaldehyde)  as a  chemothera-



peutant for control of fungus on fish  eggs and ectoparasites  on



fish is a widely accepted and successful technique.   However,



unless certain criteria are met formalin may cause acute toxic



effects in fish (U.S. EPA, 1976).   The acute toxicity of formalin



to fish has been reviewed by the U.S.  Department of Interior



(1973).  Analysis of toxicity levels indicates that a wide range



of tolerances exist for different species; striped bass appear to



be the most sensitive with an LCg0  of  15 to  35 ppm.



     The LCjQ of formaldehyde for Daphnia  magna  is reported to



range between 100 to 1000 ppm (Dowden  and  Bennett,  1965).   The



48-hour median threshold limit  (TLm) for Daphnia"was  about 2  ppm



(McKee and Wolf, 1971).



     No effects were observed in crayfish  (Procambarus blandingi)



exposed to 100 ul/1 of formalin (concentration unspecified) for



12 to 72 hours (Helm, 1964).

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VI.  EXISTING GUIDELINES



     The OSHA standard for formaldehyde in workplace air is a



time weighted average (TWA) of 3 ppm with a ceiling concentration



of 5 ppm (39 CFR 23540).  The NIOSH recommended standard is a



ceiling concentration of 1.2 mg/m  (about 0.8 ppm) (NIOSH, 1976).



The ACGIH (1977) recommends a ceiling value of 2 ppm (3 mg/m3).

-------
                            REFERENCES

American Conference of Governmental  Industrial Hygenists (ACGIH).
1977.  TLVs:  Threshold  limit values for  chemical substances in
workroom air adopted.  Cinninnati, Ohio.

Auerbach, C./ M. Moutschen-Dahen,  and J.  Moutschen.   1977.
Genetic and cytogenetical effects  of formaldehyde and relative
compound.  Mutat. Res. 39:317-361  (as cited in U.S.  EPA, 1979c).

Bufalini, J.J., Gay, Jr., B.W. and Brubaker,  K.L.  1972.  Hydro-
gen Peroxide Formation from Formaldehyde  Photoxidation and Its
Presence in Urban Atmospheres.  Environ.  Sci.  Technol. ^(9), 816
(as cited in U-.S. EPA 1976).

Buss, J., Kuschinsky, K., Kewitz,  H.  and  Koransky,  W.  1964.
Enterale Resorption von  Formaldehyde.   Arch.  Exp. Path. Pharmak.,
247, 380 (as cited in U.S. EPA, 1976).

Chemical Industry Institute of Toxicology.   Statement Concerning
Research Findings, October, 1979.

Coon, R.S., Jones, R.A., Jenkins,  L.J.  and  Siegel,  J.  1970.
Animal Inhalation Studies on Ammonia, Ethylene Glycol, Formalde-
hyde, Dimethylamine, and Ethanol.  Tox. Appl.  Pharmacol, 16, 646
(as cited in U.S. EPA, 1976).

Dowden, B.F. and Bennett, H.J.  1965.   Toxicity of Selected Chem-
icals to Certain Animals.  J. Water  Pollut.  Cont. Fed., 37(9),
1308 (as cited in U.S. EPA, 1976).

Gofmekler, V.A. and Bonashevskaya, T.I.   1969.   Experimental
Studies of Teratogenic Properties  of Formaldehyde,  Based on
Pathological Investigations.  Gig. Sanit.,  _3_4J5), 266 (as cited
in U.S. EPA, 1976).

Helms, D.R.  1964.  The  Use of Formalin to  Control Tadpoles in
Hatchery Ponds.  M.S. Thesis, Southern  Illinois University,
Carbondale, 111. (as cited in U.S. EPA, 1976).

Horton, A.W., Tye, R. and Stemmer, K.L.   1963.   Experimental
Carcinogenesis of the Lung.  Inhalation of'Gaseous Formaldehyde
on an Aerosol Tar by C3H Mice.  J. Nat. Cancer Inst., .3_p_(l), 30
(as cited in U.S. EPA, 1976 and U.S.  EPA, 1979c).

Hurni, H. and Ohder, H.  1973.  Reproduction Study with
Formaldehyde and Hexamethylenetetramine in  Beagle Dogs.  Food
Cosmet. Toxicol., _11_(3), 459 (as cited  in U.S.  EPA,  1976).

Kamata, E.  1966.  Aldehyde in Lake  and Sea Water.   Bull. Chem.
Soc. Japan, _3JL(6)' I227  
-------
Induction of single strand breaks in DNA and their  repair.
Mutat. Res. 50; 181-193 (as cited by U.S. EPA in 1979a).

Malorny, G./ Rietbrock, N. and Schneider, M.  1965.  Die Oxyda-
tion des Forraaldeshyds zu Ameiscansaure im Blat. ein Beitrag  Zum
Stoffwechsel des Formaldehyds.  Arch. Exp. Path. Pharmak.,  250>
419 (as cited in U.S. EPA, 1976).

Martin, C.N., A.C. McDermid, and R.A. Garner.   1978.  Testing of
known carcinogens and non-carcinogens for their ability to  induce
unscheduled DNA synthesis in HeLa cells.  Cancer Res. 38; 2621-
2627 (as cited on U.S. EPA, 1979c).

McKee, J.E. and Wolfe, H.W.  1971.  Water Quality Criteria, 2nd
Ed., California State Water Resources Control Board, Sacramento,
Publication 3-8 (as cited in U.S. EPA, 1976)

National Institute of Occupational Safety and Health (NIOSH).
1976.  Criteria for a recommend standard.  Occupational Exposure
to Formaldehyde.  NIOSH Publication No. 77-126.

Nelson, N. (New York University) Oct. 19, 1979.  Letter to
Federal Agencies.  A status report on formaldehyde  and HC1
inhalation study in rats.

Obe, G. and B. Seek.  1979.  Mutagenic Activity of  Aldehydes.
Drug Alcohol Depend., 4(1-2), 91-4 (abstract).

Purcell, T.C. and Cohen, I.R.  1967.  Photooxidation of Formal-
dehyde at Low Partial Pressure of Aldehyde.  Environ. Sci.
Technol., 1(10), 845 (as cited in U.S. EPA, 1976).

U.S. Department of the Interior.  1973.  Formalin as a Thera-
peutant in Fish Culture, Bureau of Sport Fisheries  and Wildlife,
PB-237 198 (as cited in U.S. EPA, 1976).

U.S. EPA.  1976.  Investigation of selected potential environ-
mental contaminants:  Formaldehyde.  EPA-560/2-76-009.

U.S. EPA. 1979a. Toxic Substances Control Act Chemical Substance
Inventory, Production Statistics for Chemicals on the Non-Confi-
dential Initial TSCA Inventory.

U.S. EPA.  1979b.  Chemical Hazard Information Profile on
Formaldehyde.  Office of Pesticides and Toxics Substances.

U.S. EPA.  1979c.  The Carcinogen Assessment Group's Preliminary
Risk Assessment on Formaldehyde.  Type I - Air Programs.  Office
of Research and Development.

Watanabe, F., Matsunaga, T., Soejima, T. and Iwata, Y.  1954.
Study on the carcinogenicity of aldehyde, 1st report.  Experi-
mentally produced rat sarcomas by repeated injections of aqueous
solution of formaldehyde.  Gann, 45, 451.  (as cited in U.S.  EPA,
1976 and U.S. EPA, 1979c)

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Wilkins, R.J., and H.D. MacLeod.  1976.  Formaldehyde induced DNA
protein crosslinks in E. coli.  Mutat. Res.  36:11-16.

Windholz, M,, ed. 1976.  The Merck  Index,  9th  ed.,  Merck and
Company, Inc.

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                                 No. 105
          Formic Acid

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
      WASHINGTON, B.C.   20460

         APRIL 30, 1980
        /OS"/

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                          DISCLAIMER
     This report represents  a  survey of the.potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

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                                 FORMIC ACID
                                   Summary

     There is  no information available on  the  possible  carcinogenic,  muta-
genic, teratogenic,  or adverse reproductive  effects of formic acid.
     Formic acid has  been  reported to produce  albuminuria  and hematuria  in
humans following chronic exposure.  Exposure to high  levels of the  compound
may  produce  circulatory  collapse, renal  failure,  and  secondary  ischemic
lesions in the liver and heart.
     Formic acid is toxic  to freshwater organisms  at  concentrations  ranging
from 120,000  to 2,500,000  ug/1.   Daphnia magna was  the most  sensitive fresh-
water species  tested.   Marine  crustaceans were  adversely affected by  expo-
sure to formic  acid  at concentrations from 80,000 to 90,000 ug/1.
                                     /f
                                /of-3

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                                  FORMIC ACID
I.   INTRODUCTION
     Formic acid (CAS registry number  64-18-6)  is a colorless, clear, fuming
liquid with a pungent.odor  (Hawley,  1571;  Windholz,  1576; Walker, 1566).  It
is  a naturally  formed product, produced by bees, wasps,  and ants (Casarett
and Ooull, 1575).  Formic  acid has widespread occurrence  in  a large variety
of  plants,  including pine  needles,  stinging nettles,  and foods  (Furia and
Bellanca, 1971;  Walker,  1966).  Industrially,  it is made by  heating carbon
monoxide with sodium  hydroxide under heat and pressure,  or  it may be formed
as  a coproduct  from  butane oxidation  (Walker,  1966).   It has  the following
physical and chemical constants (Windholz, 1976; Walker, 1966):
    Property
    Formula:
    Molecular Weight:
    Melting Point:
    Boiling Point:
    Density:
    Vapor Pressure:
    Solubility:

    Demand (1979):
Pure
90%
8535
 46.02
  8.4°C
100. 5°C
  1.220?°
       4
-4°C
1.202^
-12°C
1.154;
25
'25
      33.1 torr ® 20°C
      Miscible in water, alcohol,
      and ether; soluble  in
      acetone,benzene, and toluene
      67.5 million lias.   (CMR,  1579)


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Formic  acid  is marketed  industrially  in 85,  90,  and 98 percent aqueous solu-
tions.   It  is  also  available  at 99+  percent purity  on  a semicommercial
scale.   Formic acid is  used  primarily  as  a volatile acidulating  agent;  in
textile  dyeing'  and  finishing,  including carpet  printing; in chemical syn-
thesis  and Pharmaceuticals; and  in tanning  and leather treatment (CMR, 1579;
Walker,  1966).
II.  EXPOSURE
     A.   Water
          Formic  acid has  been  detected in xaw sewage,  in effluents  from
sewage treatment plants,  and  in  river water (Mueller, et  al.  1958).   It has
also been identified  in  effluents  from chemical plants and paper mills (U.S.
EPA, 1576).
     B.   Food
          A  large  variety of plants  contain  free formic  acid;  it has  been
detected in  pine needles,  stinging nettle, and  fruits  (Walker,  1966).   It
has  been identified  in  a  number  of essential  oils, including  petitgrain
lemon and bitter orange  (Furia and  Bellanca, 1571).   Formic  acid  is  also re-
ported to be a constitutent of strawberry aroma  (Furia  and  Bellanca,  1971).
In the U.S.  this chemical may be used  as  a food  additive;  allowable  limits
in  food range  from  1  ppm in  non-alcoholic beverages  to 18 ppm in  candy
(Furia and Bellanca, 1571).   It may also occur in  food as  a result of  migra-
tion from packaging materials  (Sax, 1975).
     C.    Inhalation
          Ambient air concentrations  of formic acid  range from  4 to 72 ppb
(Graedel, 1578).   Emission sources include forest  fires,  plants,  tobacco
smoke,  lacquer manufacture, and  combustion  of  plastics (Graedel,  1978). '  It

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has  also been  identified in  the liquid  condensate from  the pyrolysis  of
solid municipal waste (Orphey and Jerman, 1970).
     0.   Dermal
          Pertinent data were not found in the available literature.
III. PHARMACOKINETICS
     A.   Absorption
          Acute toxicity  studies in  animals  and poisoning incidents  in  man
indicate  that  formic acid is  absorbed from the  respiratory  tract  and  from
the gastrointestinal tract (Patty, 1563; NIQSH, 1977)'.
     B.   Distribution
          Pertinent data were not found in the available literature.
     C.   Metabolism
          Formate may be  oxidized to produce carbon dioxide by the activity
of  a catalase-peroxide complex,  or  it  may enter  the  folate-dependent  one
carbon  pool following  activation and proceed  to carbon  dioxide via these
reactions (Palese  and Tephly,  1975).  Species differences in the  relative
balance of these two pathways for the metabolism  of  formate have  been  postu-
lated in  order  to  explain the greater accumulation  of  formate in the blood
of monkeys administered methanol, compared  to rats similarly treated (Palese
and Tephly,  1975).
     D.   Excretion
          Following intraperitoneal  administration  of  ^C  formate to rats,
significant  amounts  of  14CCL were  detected  in  these  samples  (Palese   and
                                                   - *     • •.
Tephly, J975).
IV.  EFFECTS
     A.   Pertinent data could not be located  in the  available  literature.'

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      8.    Chronic Toxicity
           Chronic human exposure to formic acid has been reported to  produce
 albuminuria and hematuria (Windholz,  1976).
      C.    Other Pertinent Information
           Formic acid is severely  irritating  to  th skin,  eyes, and respira-
 tory tract  (NIOSH,  1977).   Gleason  (1969) has  indicated  that  exposure  to
 high levels of compound may produce circulatory collapse, renal failure,  and
 secondary  ischemic lesions  in the liver  and heart.
 V.    AQUATIC TOXICITY
      A.    Acute Toxicity
           Dowden  and  Bennett  (1965)  demonstrated  a  24-hour LC_0  value  of
 175,000  ^ig/1  for bluegill sunfish (Lepomis  macrochirus)  exposed  to formic
 acid.  Bringmann  and  Kuhn (1959)  observed  a 48-hour LC5Q  vaiue  of  120,000
/jg/1 for waterfleas  (Daphnia  maqna) exposed to formic acid.-
           Verschueren  (1979)   reported  that a formic  acid  concentration  of
 2,500,000 >jg/l was lethal to  freshwater scuds (Gammarus pulex) and 1,000,000
jug/1 was a perturbation threshold value  for the fish Trutta  iridea.
           Portmann and Wilson (1971) determined  48-hour  LC5Q  values rang-
 ing  from 80,000 to 90,000 xig/1 for the marine shore crab  (Carcinus maenas)
 exposed  to formic acid in static  renewal bioassays.
      B.    Chronic Toxicity
           Pertinent data  were not found  in the available literature.
      C.    Plant Effects
           McKee and Wolf  (1963) reported that  formic acid at a concentration
 of  100,000xig/l was toxic to  the  freshwater algae, Scenedesmus sp.
      0.    Residue
           Pertinent data  were not found  in the available literature.
                                 -  \S\ 1 "7^
                                 *V oL*"^'

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VI.  EXISTING GUIDELINES AND STANDARDS
     A.   Humsn
          The  eight-hour,  TWA  exposure  limit  for  occupational exposure  to
formic acid is 5 pprn (ACGIH, 1977).
     B.   Aquatic
          Hahn  and Jensen (1977)  have suggested  an aquatic toxicity  rating
range  of  100,000  to  1,000,000 /jg/1 based on 96-hour LC_Q  values for  aqua-
tic organisms exposed to formic acid.

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                                  FORMIC ACID

                                  References
 American Conference  of  Government Industrial  Hygienists.   1977.   Threshold
 limit 'values  for chemical  substances  and physical  agents in  the workroom
 environment with intended changes  for  1977.   American Conference of Govern-
 mental Industrial Hygienists, Cincinnati,  OH.

 Bringmann,  G.  and R. Kuhn.   1959.  The toxic effects of wastewater on aqua-
 tic bacteria,  algae  and  small crustaceans.  Gesundheits-Ing 80:  115.

 Casarett,  L.J.  and  I.  Doull.   1975.    Toxicology:   The  Basic Science -of
 Poisons.  Macmillian Publishing Co., New  York.

 CMR.   1979.  Chemical Profile.   Formic  acid.   Chemical Marketing  Reporter,
 December  17, p. 9.

 Dowden,  B.F.  and  H.J.   Bennett.   1965.  Tgxicity  of selected  chemicals  to
 certain  animals.  Jour.  Water Poll. Contr. Fed.  37:  1308.

 Furia, T.E.  and N.  Bellanca (eds.)   1971.  Fenaroli's Handbook of Flavor In-
 gredients.   The Chemical Rubber Company, Cleveland, 0.

 Gleason,  M.   1569.   Clinical  Toxicology  of  Commercial  Products,  3rd  ed.
 Williams  and Wilkins, Baltimore, MO.

 Graedel,  T.E.   1978. Chemical  Compounds  in the Atmopshere.   Academic  Press,
 New York.

 Hahn, R.W.  and P.A.  Jensen.   1977.  Water Quality  Characteristics of Hazard-
 ous Materials.  Texas A  &  M  University.  Prepared  for National Oceanographic
 and Atmospheric Administration Special Sea Grant Report.   NTIS PB-285 946.

 Hawley, G.G.  (ed.)   1971.   The Condensed Chemical Dictionary, 8th ed.   Van
 Nostrand Reinhold Co, New York.

 McKee, J.E.  and H.W. Wolf.   1963.  Water Quality  Criteria Resources  Board,
 California Water Quality Agency, Publication No. 3-A.

 Mueller,  H.F.,  et al.   1958.  Chromatographic  identification  and determina-
 tion of organic acids in water.   Anal.  Chem.  30: 41.

 National  Institute  for  Occupational Safety and  Health.   1977.   Occupational
 Diseases:  A Guide to Their  Recognition.   Washington",  DC:  U.S.  DHEW,  Publi-
 cation No. 77-181.

Orphey, R.D. and R.I. Jerman.   1960.  Gas  chrpmatographic  analysis of  liquid
condensates  from  the pyrolysis of  solid municipal waste.   Jour. Chroma,to-
 graphic Science.  8: 672.

Palese, M.  and  T.  Tephly.    1975.   Metabolism of formate in the rat.  Jour.
Toxicol. Environ.  Health.  1: 13.
                                        - f

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Patty,  F.    1563.   Industrial Hygiene  and  Toxicology,  Vol.  II.   2nd  ed.
Intersciencs, New York.

Portmsnn, J.E.  and K.W. Wilson.   1971.   The toxicity of  140 substances to
the brown shrimp arid other marine animals.  Ministry of Agriculture, Fisher-
ies and  Food, Fisheries Laboratory, Burnham-on-Crouch, Essex, Eng. Shellfish
Leaflet No.  22,  AMIC-7701.

Sax,  N.I.    1975.   Dangerous  Properties  of  Industrial Materials.   4th  ed.
Van Nostrand Reinhold, Co,  New  York.

U.S. EPA.   1976.   Frequency of organic  compounds  identified in water.   U.S.
Environ. Prot. Agency, EPA-600/4-76-062.

Verschueren,  K.    1979.    Handbook of  Environmental Data  on  Organic  Chem-
icals.  Van  Nostrand Reinhold,  Co, New York.

Walker, J.F.  1966.  Formic acid and derivatives.  In:  Kirk-Othmer Encyclo-
pedia  of Chemical Technologyt  2nd ed.   A. Standen,  (ed).   John  Wiley  and
Sons,  New York.   Vol. 10,  p. 99.

Windholz, M. (ed.)  1976.   The Merck Index.  9th ed.  Merck and Co., Rahway,
NJ.
                               IOS-16

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                                      No.  106
           Futnaronitrile

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to  the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference documents.
Because of the limitations of such sources,  this short profile
may not reflect  a-11 available  information  including all the
adverse health  and  environmental impacts  presented  by the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

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                                 FUMARONITRILE
                                    Summary

     Information  on the  carinogenic,  mutagenic,  or teratogenic  effects  of
fumaronitrile was not found  in the available  literature.  LD5Q  values  for
injected mice and orally dosed  rats were  38  and 50 mg/kg,  respectively.  Re-
ports of chronic toxicity studies were not found in the available literature.

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                                 RJMARONITRILE

 I.    INTRODUCTION
          This'profile is based  upon relevant literature identified  through
 mechanized   bibliographic  searches   such  as   TOXLINE,   BIOSIS,  Chemical
 Abstracts,  AGRICOLA  and  MEDLARS,  as  well  as  manual  searches.    Despite
 approximately  70 citations  for  fumaronitrile,  approximately  95  percent  of
 these  concerned  the chemistry  of fumaronitrile or  its  reactions with  other
 chemicals.  Apparently,  the chief use of  fumaronitrile  is as a chemical  in-
 termediate  in  the manufacture  of other  chemicals,  rather than  end .uses  as
 fumaronitrile per se.  Undoubtedly,  this  accounts for its low profile in  the
 toxicological literature.
          Fumaronitrile   or   trans-l,2-dicyanoethylene   (molecular  weight
 78.07)  is  a solid  that  melts at  96.8°c  (Weast,  1975), has  a boiling  point
 of  186°c,   and  a  specific gravity  of 0.9416 at  25°C.  It is  soluble  in
water,  alcohol,  ether,  acetone,  chloroform,  and benzene.   Fumarcnitrile  is
used  as a bactericide  (Law,  1968),  and  as an antiseptic  for metal cutting
 fluids  (Wantanabe, et al., 1975).   It is used  to make polymers with styrene
numerous other  compounds.  This  compound is  easily isomerized to  the  cis-
 form, maleonitrile, which is  a  bactericide and fungicide (Ono, 1979).    It  is
conveniently  synthesized  from  primary amides under  mild conditions  (Cam-
pagna, et al., 1977).
 II.  EXPOSURE
          Human exposure  to fumaronitrile  from foods cannot  be assessed, due
to a lack of monitoring data.
          3ioaccumulation data on  fumaronitrile were not found in the avail-
able literature.

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 III. PHARMACOKINETICS
          Specific  information  on the metabolism,  distribution,  absorption,
or elimination of fumaronitrile was not found in the available literature.
IV.  EFFECTS  •
     A.   Carcinogenicity,  Mutagenicity,  Teratogenicity,  Reproductive
          Effects, and Chronic Toxicity
          Pertinent data could not be located in the available literature.
     8.   Acute Toxicity
          LD50 values  for  injected  mice  and orally  dosed  rats were 38 and
5Q mg/kg, respectively (Zeller,  et al., 1969).
V.   AQUATIC TOXICITY
          Data concerning the effects of fumaronitrile to  aquatic  organisms
were not found in the available  literature.
VI.  EXISTING GUIDELINES AND STANDARDS
          Data concerning  existing  guidelines  and  standards for  fumaroni-
trile were not found in the  available literature.
                                /0%'S

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                                  REFERENCES
Campagna,  F.,  et  al.    1977.   A  convenient  synthesis  of  nitriles   from
primary amides under mild conditions.  Tetrahendron Letters.   21: 1313.

Law, A.  1968.  Fumaronitrile as a bactericide.  Chen, Abst.   68: 1135.

Ono,  T.    1979.   Maleonitrile,  a  bactericide and  fungicide.   Chem.   Abst.
32: 126.

Wantanabe, M.,  et al.  1975.   Antiseptic for a metal cutting fluid.   Chem.
Abst.  .82: 208.

Weast, R.  1975.   Handbook of Chemistry  and  Physics^  56th ed.  Chem. Rubber
Publ. Co.  p. 2294.
                                                  •.

Zeller,  H.,  et  al    1969.    Toxicity   of  nitriles:   Results  of  animal
experiments  and  industrial  experiences  during   15  years.   Chem.   Abst.
71: 326.
                                     0

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                                      No. 107
            Halomethanes

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents a  survey  of  the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the report is drawn chiefly
from secondary  sources  and available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including  all the
adverse health  and  environmental  impacts  presented  by  the
subject chemical.   This  document  has undergone scrutiny  to
ensure its technical accuracy.
                          107-

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                                 HALOMETHANES'
                                    Summary
     The halomethanes  are a subcategory  of halogenated hydrocarbons.  There
is little  known concerning the  chronic toxicity of  these compounds.  Acute
toxicity results in central nervous system depression and liver damage.  The
fluorohalomethanes are  the  least toxic.  None of  the halomethanes have been
demonstrated to  be carcinogenic; however, chloro-, bromo-, dichloro-, bromo-
dichloro-, and  tribromomethane  have been  shown  to be mutagenic  in the Ames
assay.   There  are  no  available  data on  the  teratogenicity  of  the halo-
methanes,  although  both  dichloromethane  and bromodichloromethane  have been
shown to affect the fetus.
     Brominated methanes  appear  to be  more toxic to aquatic life than chlor-
inated methanes.   Acute toxicity  data  is  rather  limited in scope,  but re-
veals toxic concentrations in the  range of  11,000 to 550,000 jug/1.

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I.   INTRODUCTION
     This  profile  is based  on the  Ambient Water  Quality  Criteria Document
for Halomethanes (U.S. EPA, 1979).
     The  halomethanes  are a  sub-category  of halogenated hydrocarbons.  This
document   summarizes  the   following  halomethanes:   chloromethane   (methyl
chloride);  bromomethane  (methyl  bromide,  monobromomethane,  embafume);  di-
chloromethane  (methylene  chloride,  methylene dichloride,  methylene bichlor-
ide);  tribromomethane  (bromoform);  trichlorofluoromethane   (trichloromono-
fluoromethane,  fluorotrichloromethane,  Frigen 11,  freon-11,  Arcton  9);  and
dichlbrodifluoromethane  (difluorodichloromethane,  Freon 12,  Frigen 12, Arc-
ton 6, Genetron' 12,  Halon,  Isotron  2) and bromodichloromethane.  These  halo-
methanes  are  either  colorless gases or liquids at environmental  temperatures
and  are   soluble  in  water  at  concentrations from 13  x   10   to  2.5  x  10°
jug/1, except  for tribromomethane  which is only slightly soluble  and bromodi-
chloromethane  which  is insoluble.  Halomethanes are  used  as  fumigants, sol-
vents, refrigerants,  and in  fire extinguishers.   Additional  information  on
the  physical/chemical properties of chloromethane,  dichloromethane,  bromo-
methane,  and  bromodichloromethane,  can be  found in the ECAO/EPA (Dec.  1979)
hazard profile on these chemicals.
II.  EXPOSURE
     A.   water
          The  U.S. EPA (1975)  has  identified chloromethane, bromomethane,  di-
chloromethane,  tribromomethane,  and bromodichloromethane -in  finished drink-
ing waters in the United  States.  Halogenated hydrocarbons  have been  found
in finished waters at greater concentrations than  in  raw  waters (Symons,  et
                                                                         »
al. 1975), with  the  concentrations  related to the  organic content  of the  raw •
water.   The  concentrations of  halomethanes detected  in one  survey  of U.S.
drinking  waters are:

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                     Halomethanes in the U.S.  EPA Region V
                          Organics Survey (83 Sites)
Compound
Bromodichloromethane
Tribromome thane
Oichloromethane
Percent of
Locations with
Positive Results
78
14
8
Concentrations (mq/1)
Median
0.006
0.001
0.001
Maximum
0.031
0.007
0.007
Source:  U.S. EPA, 1975
Symons, et  al.  (1975) concluded that  trihalomethanes  resulting from chlori-
nation are  widespread in  chlorinated  drinking  waters.   An  unexplained in-
crease in the  halomethane concentration  of water  samples occurred  in the
distribution system water as compared to the treatment plant water.
     B.  Food
         Bromomethane  residues from  fumigation  decrease  rapidly  from both
atmospheric  transfer  and  reaction  with proteins  to form  inorganic bromide
residues.    With  proper  aeration  and  product  processing,  most  residual
bromomethane  will  disappear  rapidly   due   to  methylation  reactions  and
volatilization  (Natl.  Acad.  Sci.,   1978; Davis,  et al.  1977).   The U.S. EPA
(1979) has  estimated  the weighted   average  bioconcentration  factors  for
dichloromethane and tribromomethane to be 1.5 and  14,  respectively, for the
edible portions of  fish and shellfish consumed  by  Americans.   This estimate
is based  on  the octanol/water partition coefficient of  these two compounds.
Bioconcentration factors for the other halomethanes have not been determined.
     C.  Inhalation
         Saltwater  atmospheric background  concentrations  of  chloromethane
and  bromomethane,  averaging  about 0.0025 mg/m   and  0.00036  mg/m   respec-
tively, have been  reported  (Grimsrud  and  Rasmussen,   1975;  Singh, et  al.
1977).  These values  are higher than  reported average  continental background

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and urban  levels  suggesting that the oceans may  be  a major source of global
chloromethane and bromomethane.  Outdoor bromomethane concentrations as high
as O.QOQ85 mg/m  may occur  near light traffic.  This  results  from the com-
bustion of ethylene dibromide, a component  of leaded  gasoline  (Natl.  Acad.
Sci.,  1978).   Reported  background concentrations of dichloromethane in both
continental  and saltwater atmospheres  are  about 0.00012 mg/m  ,  while  urban
air  concentrations  ranged  from  less  than  0.00007  to 0.0005  mg/m .   Local
high  indoor concentrations can  be  caused by  the use  of  aerosol  sprays  or
solvents  (Natl.  Acad.  Sci., 1978).   Concentrations  of dichlorodifluorometh-
ane and  trichlorofluoromethane in the  atmosphere over urban areas are sev-
eral  times those  over  rural or oceanic areas.   This probably indicates that
the primary  modes of entry into the  environment, i.e., use  of refrigerants
and aerosols,  are greater in industrialized and  populated  areas  (Howard,  et
al.  1974).  Average concentrations  of  trichlorofluoromethane reported  for
urban  atmospheres  have  ranged  from  nil  to  3  x   IQ'^ mg/m5,  and concen-
                                                           -3               2
trations  for dichlorofluoromethane  ranged   from  3.5  x 10    to  2.9 x  10
mg/m  .
III. PHARMACOKINETICS
     A.  Absorption
         Absorption  via  inhalation  is of primary  importance and  is  fairly
efficient  for the halomethanes.  Absorption  can also occur via the skin and
gastrointestinal tract,  although this is generally  more significant for the
nonfluorinated  halomethanes  than for the fluortJCarbons- (Natl. Acad.  Sci.,
1978; Oavis, et al.  1977; U.S. EPA,  1976;  Howard, et al. 1974).
                                 107-6

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     8.  Distribution
         Halomethanes are  distributed rapidly to  various tissues  after  ab-
sorption  into  the  blood.   Preferential  distribution  usually  occurs   to
tissues witlVhigh lipid content (U.S.  EPA, 1979).
     C.  Metabolism
         Chloromethane  and bromomethane  undergo  reactions  with  sulfhydryl
groups  in  intracellular enzymes  and  proteins,  while bromochloromethane  in
the body  is hydrolyzed  in significant amounts  to yield  inorganic bromide.
Dichloromethane is  metabolized to carbon  monoxide which  increases carboxy-
hemoglobin  in  the  blood and  interferes with  oxygen transport  (Natl.  Acad.
Sci.,   1978).   Tribromomethane  is apparently  metabolized to carbon monoxide
by the cytochrome  P-450-dependent mixed  function  oxidase system  (Ahmed,  et
al. 1977).  The fluorinated halomethanes  form  metabolites  which  bind to cell
constituents,   particularly when  exposures are long-term  (Blake  and Mergner,
1974).  Metabolic data  for bromodichloromethane  could not be  located  in  the
available literature.
     D.  Excretion
         Elimination of  the halomethanes  and  their metabolites occurs mainly
through expired breath and urine (U.S. EPA, 1979).
IV.  EFFECTS
     A.  Carcinogenicity
         None  of  the halomethanes summarized  in  this document  are considered
to be  carcinogenic.  Theiss and  coworkers (1977). examined  the  tumorigenic
activity  of tribromomethane,  bromodichloromethane, --and dichloromethane  in
strain A mice.  Although increased  tumor  responses were noted with each,  in
                                                                        »
no case were  all  the requirements met  for a  positive carcinogenic response,
as defined  by  Shimkin and  Stoner  (1975).   Several  epidemiologic  studies have
                                /07-7

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 established an association between trihalomethane levels  in  municipal  drink-
 ing water supplies in the United States and certain cancer death  rates (var-
 ious sites) (Natl.  Acad.  Sci.,  1978; Cantor  and McCabe, 1977).  Cantor,  et
 al. (1978) cautioned that these  studies  have  not  been  controlled for  all
 confounding variables,  and  the limited monitoring  data that were  available
 may not have been an accurate reflection of past exposures*
     •B.  Mutagenicity
          Simmon,   et  al.  (1977)  reported  that chloromethane,  bromomethane,
_and dichloromethane  were  all mutagenic  to  Salmonella  typhimurium  strain
 TA1QQ  when assayed in a dessicator whose  atmosphere contained  the  test com-
 pound.   Metabolic activation  was  not required.  Only  marginal positive  re-
 sults  were obtained with  bromoform  and  bromodichloromethane.   Andrews,  et
 al. (1976) and  Jongen,  et al. (1978)  have confirmed  the- positive  Ames  re-
 sults  for  chloromethane and dichloromethane,   respectively.   Oichloronethane
 was negative in  tnitotic recombination in  S^  cerevisiae  03  (Simmon,  et  al.
 1977)  and  in mutagenicity  tests  in Drosophila  (~ilippova, et  al.  1967K
 Trichlorofluorcmethane and dichlorofluoromethane  were negative in  the  Ames
 assay  (Uehleke,   at  al.  1977), and dichlorodifluoromethane  in a rat feeding
 study  (Sherman,  1974) caused no increase in mutation rates over controls.
     C.   Teratogenicity
          Pertinent information could not be located in the available litera-
 ture.
     0.  Other'Reproductive Effects
          Gynecologic problems have  been  reported in  female workers exposed
 to dichloromethane and  gasoline vapors (Vozovaya,  1974).   Evidence  of  fete-
                                                                      »
 embryotoxicity has  been noted in  rats  and mice  exposed to  dichlorcmethane
                                  (07-?

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vapor on  gestation  days 6 to  15  (Schwetz,  et al. 1975).   Some  fetal  anoma-
lies were  reported  in  experiments  in which  mice were  exposed  to vapor  of
bromodichloromethane at  8375  mg/m ,  7 hours/day  during gestation days  6  to
15 (Schwetz, et al.  1975).
     E.  Chronic Toxicity
         Schuller, et al. (1978) have observed a. suppression  of  cellular and
humoral immune response  indices in female  ICR mice exposed by gavage  for  90
days to bromodichloromethane at 125  mg/kg  daily.   Tribromomethane suppressed
reticuloendothelial system function  (liver and  spleen phagocytic uptake  of
Listeria monocytoqenes)  in mice exposed  90 days at daily  doses  of 125 mg/kg
or less  (Munson,  et al. 1977,1978).  Information pertinent  to  the  chronic
toxicity  of the  other  halomethanes  could  not be located in the  available
literature.
     F.  Other Relevant Information
         In general,  acute  intoxication by  halomethanes  appears  to  involve
the central nervous system and liver  function (U.S. EPA, 1979).
V.   AQUATIC TOXICITY
     A.  Acute Toxicity
         Acute toxicity  studies  for halomethanes have  obtained acute  LC5Q
values for the bluegill sunfish  (Lepomis  machrochirus) of 11,000 ug/1  for
methylbromide, 29,300 ug/1  for bromoform,  224,000 ug/1  for methylene  chlor-
ide and 550,000 for methyl  chloride.  A static  bioassay  produced a  96-hour
LC5Q  value  of 310,000 ug/1  methylene  chloride  for. the  fathead  minnow
(Pimephales promelas) while  a flow-through assay  produced an LC5Q value  of
193,000 /jg/1.   In  freshwater  invertebrates  two  acute  studies  with  Daphnia
                                /07-J

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maqna  resulted in  LC^ values  of  46,500  .ug/1  for  bromoform,  and  224,000
pg/1  for  methylene chloride.   In  marine fish,  LC_Q  values for the  sheeps-
head  minnow  (Cyorinodon  varieqatus)  were  17,900 pg/1  for  bromoform  and
180,958 ug/1  for methylene chloride.   For  the tidewater silversides (Menidia
bervllina)  LC5Q values of  12,000 pg/1  for  methylbromide  and 147,610  ug/1
for methylene chloride were obtained.   Adjusted LC^ values  for  the  marine
mysid  shrimp  (Mysidopsis  bahia) were  24,400  ug/1 for bromoform and  256,000
pg/1 for methylene chloride (U.S. EPA, 1979).
     8.  Chronic Toxicity
         The  only  chronic value  for an aquatic species  was  9,165 jug/1 for
the sheepshead minnow.
     C.  Plant Effects
         Effective  concentrations  for  chlorophyll  a  and  cell  numbers  in
freshwater  algae  Selenastrum  capricomutum  ranged from  112,000 to  116,000
ug/1  for  bromoform  and 662,000 pg/1 for methylene chloride, while effective
concentrations  for  the marine  algae (Sketonema  costatum)  were  reported  as
11,500 to  12,300 pg/1 for bromoform and   662,000 ug/1  for methylene  chlor-
ide (U.S. EPA, 1979).
VI.  EXISTING GUIDELINES AND STANDARDS
     Neither  the human health nor  the aquatic criteria derived by U.S. EPA
(1979), which  are  summarized  below,  have gone through the  process  of  public
review;  therefore,  there  is  a  possibility  that these  criteria  will  be
changed.
     A.  Human
         Positive associations between human  cancer mortality  rates and  tri-
halomethanes  (chloroform, bromodichloromethane,  tribromomethane) in drinking
                                 107-10

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water have been reported.  There have also been  positive  results  for tribro-
momethane using  strain A/St. male  mice in  the  pulmonary adenoma  bioassay.
Bromomethane, chloromethane,  dichloromethane,  bromodichloromethane and  tri-'
bromomethane have been reported as  mutagenic in  the Ames test without meta-
bolic activation.   Dichlorodifluoromethane caused a significant  increase  in
mutant frequency in Neurospora  crassa  (mold), but  was  negative  in the  Ames
test.  No  data  implicating  trichlorofluoromethane  as  a  possible  carcinogen
have been published.
         Because positive  results  for  the mutagenic endpoint correlate  with
positive results  in in  vivo bioassays  for  oncogenicity, mutagenicity  data
for  the  halomethanes suggests  that several  of  the compounds might also  be
carcinogenic.  Since carcinogenicity data currently available for  the halo-
methanes were  not  adequate  for the development of water  quality  criteria
levels,  the  draft  criteria recommended  for  chloromethane, bromomethane, di-
chloromethane,  tribromomethane and  bromodichloromethane are the same  as  that
for chloroform,  2 /jg/1.
         Chloromethane:  OSHA (1976) has  established the maximum acceptable
time-weighted average air  concentration for  daily  8-hour occupational expo-
sure at 219 mg/m .
         Bromomethane:   OSHA  (1976)  has  a threshold  limit  value  of  80
mg/m  for  bromomethane,  and  the American  Conference of  Governmental  Indus-
trial Hygienists (ACGIH,  1971) has  a threshold limit value of  78 mg/m3. .
         Dichloromethane:  OSHA  (1976a,b)  has  established  an  8-hour time-
weighted average  for dichloromethane of 1,737  mg/m ,  however,  NIOSH  (1976)
has  recommended  a  10-hour   time-weighted   average  exposure   limit  of  261
mg/m  of dichloromethane  in the presence of no more  carbon monoxide  than
9.9 mg/m3.
                                  txrr
                                    7-//

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         Tribromomethane:   QSHA  (1976a,b)  has  established an  3-hour time-
weighted average for tribromomethane of 5 mg/m .
         Bromodichloromethane:   There is  no  currently  established  occupa-
tional exposure standard for bromodichloromethane.
         Trichlorofluoromethane  and  dichlorodifluoromethane:   The  current
OSHA  (1976)  8-hour  time-weighted  average occupational  standards  for tri-
chlorofluoromethane  and dichlorodifluoromethane  are 5,600  and  4,950 mg/m  ,
respectively.   The U.S.  EPA (1979)  draft'water  quality  criteria  for tri-
chlorofluoromethane  and dichlorodifluoromethane -are 32,000 and  3,000 /jg/1,
respectively.
     8.  Aquatic
         Draft  criteria  for the  protection  of  freshwater life  have been
derived  as 24-hour  average concentrations  for the  following halomethanes:
methylbromide - 140 ug/1 not to exceed  320 ug/1; bromoform - 840 jjg/1 not to
exceed 1,900 ug/1; methylene chloride -  4,000 ug/1 not to exceed 9,000 ug/1;
and methyl chloride - 7,000 jug/1 not to exceed 16,000 ug/1.
         Draft criteria  for the protection of marine  life have been derived
as 24-  hour average concentrations for  the  following halomethanes:  methyl-
bromide 170 ug/1 not  to exceed  380 ug/1;  bromoform  -  180 pg/1 not to exceed
420  ug/1;   methylene  chloride - 1,900 jug/1  not  to exceed 4,400  pg/1;  and
methyl chloride -  3,700 ug/1 not to exceed 3,400 ug/1.


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                        HALOMETHANES

                         REFERENCES

Ahmed, A.E., et al.  1977.  Metabolism of haloforms  to carbon
monoxide, I. Ijn vitro studies.  Drug. Metab. Dispos.  5: 198.
(Abstract).

American Conference of Governmental and Industrial Hygienists
1971.  Documentation of the threshold limit value for sub-
stances in workroom air.  Cincinnati, Ohio.

Andrews, A.W., et al.  1976.  A comparison of the mutagenic
properties of vinyl chloride and methyl chloride.  Mutat.
Res.  40: 273.

Blake, D.A., and G.W. Mergner.  1974.  Inhalation studies on
the biotransf ormation and elimination of '(^C)-trichloro-
fluoromethane and (14C)-diphlorodifluoromethane  in beagles.
Toxicol. Appl.  Pharmacol.  30: 396.

Cantor, K.P., and L.J. McCabe.  1977.  The epidemiologic
approach to the evaluation of organics in drinking water.
Proc. Conf. Water Chlorination: Environ. Impact  and  Health
Effects.  Gatlinburg, Tenn.  Oct. 31-Nov. 4.

Cantor, K.P. et al.  1978.  Associations of halomethanes in
drinking water with cancer mortality.  Jour. Natl. Cancer
Inst. (In press).

Davis, L.N., et al.  1977.  Investigation of selected poten-
tial environmental contaminants: monohalomethanes.   EPA 560/
2-77-007; TR 77-535.  Final rep. June, 1977, on  Contract No.
68-01-4315.  Off. Toxic Subst. U.S.' Environ. Prot. Agency,
Washington, D.C.

Filippova, L.M., et al.  1967.  Chemical mutagens.   IV.
Mutagenic activity of geminal system.  Genetika   8:  134.

Grimsrud, E.P., and R.A. Rasmussen.  1975.  Survey and analy-
sis of halocarbons in the atmosphere by gas chromatography-
mass spectrometry.  Atmos. Environ.  9: 1014.

Howard, P.H., et al.  1974.  Environmental hazard assessment
of one and two carbon fluorocarbons.  EPA 560/2-75-003.  TR-
74-572-1.  Off. Toxic Subst.  U.S. Environ. Prot. Agency,
Washington, D.C.

Jongen, W.M.F., et al.  1978.  Mutagenic effect  of dichloro-
methane on Salmonella typhimurium. Mutat. Res. 56: 246.
                             -Tj? V—-
                             jf IA. I ^

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Munson, A.E.,.et al.  1977.  Functional  activity  of  the  re-
ticuloendothelial system  in mice exposed  to  haloalkanes  for
ninety days.  Abstract.   14th Natl. Reticuloendothelial  Soc.
Meet. Tucson, Ariz.  Dec. 6-9.

Munson, A.E., et al.  1978.  Reticuloendothelial  system  func-
tion in mice exposed to four haloalkanes:  Drinking water con-
taminants.-  Submitted: Soc. Toxicol.  (Abstract).

National Academy of Sciences.  1978.  Nonfluorinated  halo-
methanes in the environment.  Washington,  D.C.

National Institute for Occupational Safety and  Health.   1976.
Criteria for a recommended standard:  Occupational exposure to
methylene chloride.  HEW  Pub. No.  76-138.  U.S. Dep.  Health
Edu. Welfare, Cincinnati, Ohio.

Occupational Safety and Health Administration.  1976.  Gener-
al industry standards.  OSHA 2206, revised January,  1976.
U.S. Dept. Labor, Washington, D.C.

Schuller, G.B., et al.  1978.  Effect of  four haloalkanes on
humoral and cell mediated immunity in mice.  Presented Soc.
Toxicol. Meet.

Schwetz, B.A., et al.  1975.  The  effect  of  maternally in-
haled trichloroethylene,  perchloroethylene,  methyl chloro-
form, and methylene chloride on embryonal  and fetal  develop-
ment in-mice and rats.  Toxicol. Appl. Pharmacol.  32:   84.

Sherman, H.  1974.  Long-term feeding studies in  rats and
dogs with dichlorodifluoromethane  (Freon  12  Food  Freezant).
Unpubl. rep. Haskell Lab.

Shimkin, M.B., and G.D. Stoner.  1975.   Lung tumors  in mice:
application to carcinogenesis bioassay.   Adv. Cancer Res.
21: 1.

Simmon, V.F., et al.  1977.  Mutagenic activity of chemicals
identified in drinking water.  S.  Scott,  et  al.,  eds.  In
Progress in genetic toxicology.

Singh, H.B., et al.  1977.  Urban-non-urban  relationships of
halocarbons, SFg, N20 and other atmospheric  constituents.
Atmos. Environ.  11: 819.

Symons, J.M., et al.  1975.  National organics "reconnaissance
survey for halogenated organics.   Jour.  Am.-• Water Works
Assoc.  67: 634.

Theiss, J.C., et al.  1977.  Test  for carcinogenicity of or,-
ganic contaminants of United States drinking waters  by pul-
monarv tumor response in  strain A  mice.   Cancer Res.  37:
2717."

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Uehleke, H., et al.  1977.  Metabolic activation of haloal-
kanes and tests _in vitro for mutagenicity.  Xenobiotica  7:
393.

U.S. EPA.  1975.  Preliminary assessment of suspected carcin-
ogens in drinking water, and appendices.  A report to Con-
gress, Washington, D.C.

U.S. EPA.  1976.  Environmental hazard assessment report,
major one- and two- carbon saturated fluorocarbons, review of
data.  EPA  560/8-76-003.  Off. Toxic Subst. Washington,
D.C.

U.S. EPA.  1979a.  Halomethanes: Ambient Water Quality Cri-
teria. (Draft).

U.S. EPA.  1979b.  Environmental Criteria and Assessment Of-
fice.  Halomethanes: Hazard Profile (Draft).

Vozovaya, M.A.  1974.  Gynecological illnesses in workers of
major industrial rubber products plants occupations.  Gig.
Tr.  Sostoyanie Spetsificheskikh Funkts.  Rab. Neftekhim.
Khim. Prom-sti. (Russian) 56. (Abstract).

Wilkness, P.E., et al.  1975.  Trichlorofluoromethane in the
troposphere, distribution and increase, 1971 to 1974.
Science  187: 832.

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                                       No.  108
             Heptachlor

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health*
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

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                     SPECIAL NOTATION
U.S;  EPA1s Carcinogen Assessment Group  (CAG) has  evaluated

heptachlor and has found sufficient evidence to indicate

that  this compound is carcinogenic.
                          -IAS*-
                          /Of" 3

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                                  HEPTACHLOR
                                    Summary
     Heptachlor  is  an organochlorinated cyclodiene insecticide, and has been
used mostly  in its technical, and  hence,  impure form, in  most bioassays up
to  the present.   Nevertheless,  it  has been  found that heptachlor  and its
metabolite,  heptachlor epoxide,  induce liver cancer in mice and rats.  Hep-
tachlor was  mutagenic in two mammalian assays  but not in the Ames test.  In
long-term  reproductive studies  in rats, heptachlor caused  reduction  in lit-
ter size,  decreased lifespan  in  suckling  rats,  and cataracts in both parents
and offspring.  Little  is  known about other  chronic effects  of heptachlor
except that  it induces  alterations  in glucose homeostasis.   It causes con-
vulsions  in  humans.   Heptachlor  epoxid_e,  its major  metabolite, accumulates
in adipose tissue and  is more acutely toxic than the parent compound.       ,
     Numerous  studies  indicate that  heptachlor  is  highly  toxic, both acutely
and chronically,  to  aquatic  life.   Ninety-six  hour  LC-Q  values  for fresh-
water  fish range from 7.0 ug/1  to  320 pg/1 and  24 to  96-hour LC5Q values
for invertebrates  from 0.9 ug/1  to 80 pg/1.   The 96-hour values  for salt-
water  fish range from 0.8  to 194 ug/1.   In  a 40-week life  cycle  test with
fathead  minnows, the determined  no-adverse-effect concentration  was  0.86
pg/1.   All fish  exposed at 1.84 ug/1  to heptachlor were  dead after 60 days.
The fathead  minnow bioconcentrated  heptachlor  and its  biodegradation  pro-
duct,   heptachlor  epoxide,   20,000-fold over  ambient  water  concentrations
after 276  days exposure.  The saltwater sheepshead minnow  accumulated these
two compounds  37,000-fold after  126 days  exposure."  Heptachlor  epoxide has
approximately the same toxicity values as  heptachlor.
                               /6 8"',

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I.   INTRODUCTION
     This  profile  is based  on the  Ambient Water  Quality Criteria  Document
for Heptachlor (U.S. EPA, 1979).
     Heptachlor  is  a broad  spectrum insecticide of  the group of  polycyclic
chlorinated hydrocarbons  called cyclodiene insecticides.   From 1971 to  1975
the  most important  use  of  heptachlor  was to  control agricultural  soil  in-
sects (U.S. EPA, 1979).
     Pure, heptachlor   (dnemical  name  l,4,5,6,7,8,8-heptachloro-3a,4,7,7a-
tetrahydro-4,7-fnethanoindene;   C,J-lcCl7;   molecular   weight  373.35)  is   a
white crystalline  solid  with a camphor-like  odor.    It  has a vapor  pressure
of  3 x  10~4  mm Hg  at  25°C, a solubility in  water  of 0.056  mg/1 at 25  to
29°C,  and is  readily soluble  in  relatively  nonpolar solvents  (U.S.  EPA,
1979).                                                                        i
     Technical  grade heptachlor  (approximately  73  percent  heptachlor;  21
percent  trans  chlordane,  5  percent  heptachlor epoxide  and 2 percent  chlor-
dene  isomers)  is  a tan,  soft, waxy  solid with  a  melting  range of  46  to
74°C and a vapor pressure of 4 x 10"4 mm Hg at  25°C (U.S.  EPA,  1979).
     Since 1975,  insecticidal uses  and  production  volume  have declined  ex-
tensively because of the  sole producer's  voluntary restriction and  the  sub-
sequent issuance of  a registration suspension  notice  by the U.S. EPA,  August
2, 1976,  for all  food crop  and  home  use of heptachlor.  However, significant
commercial use of  heptachlor for termite control and non-food  crop pests
continues.                                         -       \-
     Heptachlor  persists  for prolonged  periods  in  the environment.   It  is
converted  to  the  more  toxic  metabolite,  heptachlor  epoxide,  in  the  soil
                                                                        »
(Lichtenstein,  1960; Lichtenstein,  et  al.  1970,   1971;   Nash and  Harris,
1972),  in plants  (Gannon and  Decker,  1958),   and  in  mammals  (Oavidow  and

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Radomski, 1953a).  Heptachlor,  in  solution or thin films, undergoes photode-
composition  to photoheptachlor  (Benson,   et  al.  1971)  which  is  more  toxic
than  the  parent compound to  insects (Khan,  et al.  1969),  aquatic inverte-
brates  (Georgacakis  and Khan,  1971; Khan,  et al. 1973)  and rats, bluegill
(Lepomis  machrochirus) and  goldfish (Carassius  auratus)  (Podowski,  et  al.
1979).  Photoheptachlor. epoxide  is also  formed in sunlight., and is  more  toxic
than the parent compound (Ivie, et al. 1972).
     Heptachlor and  its epoxide will bioconcentrate  in numerous species  and
will accumulate in the food chain  (U.S. EPA, 1979).
II.  EXPOSURE
     A.  Water
         Various  investigators  have detected heptachlor  and/or  heptachlor
epoxide in  the major  river  basins of the U.S. at a  mean concentration  for
both of 0.0063 pg/1  (U.S.  EPA,  1976).  Levels of  heptachlor  ranged from .001
jjg/1  to 0.035  ug/L and heptachlor/heptachlor  apoxide were  found  in  25 per-
cent  of all  river samples  (Breidenbach,  et  al.  1967).   Average  levels  in
cotton sediments are around 0.8 ug/kg (U.S. EPA, 1979).
     B.  Food
         In  their  market basket  study  (1974-1975) for  20 different cities,
the FDA  showed that 3 of  12 food  classes contained  residues  of  heptachlor
epoxide ranging from 0.0006  to  0.003 ppm (Johnson and Manske, 1977).  Hepta-
chlor epoxide  residues greater than 0.03  mg/kg have  been found in 14 to  19
percent  of   red  meat,   poultry,  and  dairy products "sampled  from  1964-1974
(Nisbet, 1977).  Heptachlor  and/or heptachlor epoxide were  found  in  32 per-
cent of 590  fish  samples obtained nationally,  with  whole fish residues from
                                                                           »
0.01 to 8.33 mg/kg (Henderson, et al. 1969).


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         The  U.S.  EPA  (1979)  has estimated the  weighted average bioconcen-
tration  factor  for heptachlor in the  edible portions of  fish and shellfish
consumed  by  Americans  to  be  5,200.   This  estimate is  based  on  measured
steady-state  bioconcentration factors- for  sheepshead minnows,  fathead min-
nows, and spot  (Leiostomus xanthuru).
         Human  milk can be contaminated with  heptachlor epoxide.  A nation-
wide  survey indicated that  63.1 percent of  1,936 mothers'  milk samples con-
tained heptachlor  epoxide residues  ranging  from 1 to 2,050 ug/1  (fat adjust-
ed)  (Savage,  1976).  Levels of 5 ug/1 of  the epoxide have  been reported in
evaporated milk (Ritcey, et al. *1972).
      C.  Inhalation
         Heptachlor  volatilizes  from  treated  surfaces,   plants,  and  soil
(Nisbet, 1977).  Heptachlor,  and to a  lesser  extent heptachlor epoxide,. are
widespread  in ambient  air  with typical mean  concentratons of approximately
0.5  ng/m .   On the basis of  this  data, typical human exposure  was  calcu-
lated to be 0.01 ug/person/day (Nisbet, 1977).   Thus, it  appears that inha-
lation is not a major route for  human exposure to heptachlor.  Air downward
from  treated  fields may contain  concentrations as high  as  600 ng/m  .   Even
after three weeks, the  air from  these fields may contain up  to 15.4  ng/m3.
Thus, sprayers,  farmers and nearby  residents  of sprayed  fields  may  receive
significant exposures (Nisbet, 1977).
     0.   Dermal
         Gaines (1960)  found  rat  dermal LD^g "values • af  195  and 250  mg/kg
for males  and females,  respectively,  compared with  oral l-Dc0's of 100  and
162 mg/kg,  respectively,  for  technical  heptachlor.   Thus, dermal  exposures
                                                                      »
roay be important in humans under the right  exposure conditions.
                                   (07-7

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III. PHARMACOKINETICS
     A.  Absorption
         Heptachlor  is  readily  absorbed  from  the  gastrointestinal  tract
(Radomski  and  Davidow, 1953;  Mizyukova and  Kurchatov, 1970;  Matsumura  and
Nelson, 1971).  The  degree- to  which heptachlor is absorbed by  inhalation  has	
not been reported  (Nisbet,  1977).   Percutaneous absorption is  less  efficient
than through  the  gastrointestinal  tract,  as indicated  by  comparison of  the
acute toxicity resulting from dermal vs. oral exposures  (Gaines,  1960).
     8.  Distribution and Metabolism
         Heptachlor  reaches all  tissues of the rat within one  hour  of a sin-
gle oral dose and  is metabolized to heptachlor epoxide.  Heptachlor  has been
found to bind to hepatic  cytochrome P-450, an enzyme of the liver hydroxyla-
tion system (Donovan,  et  al.  1978).  By the end of  one month  traces  of heq-
tachlor epoxide were detectable only in fat  and liver.  Levels of  the  epox-
ide in fatty tissues stabilized  3 to 6 months after  a single  dose of hepta-
chlor  (Mizyukova and Kurchatov,  1970).   Human fat  samples  may also  contain
nonachlor  residues  derived from  technical heptachlor  or  chlordane  exposure
(Sovocool  and  Lewis,  1975).   When experimental animals were   fed heptachlor
for two months, the  highest levels of heptachlor  epoxide  were found  in fat,
with lower levels  in  liver,  kidney and  muscle and  none  in brain  (Radomski
and Davidow, 1953).   There is evidence to show that  the  efficiency  of con-
version to the epoxide  in  humans is less than in the rat (Tashiro and Matsu-
mura, 1978).  Various  researchers  have  found that heptachlor epoxide  is more
toxic to mammals than  the  parent compound (U.S. EPA', 1979).  There  is an  ap-
proximate  ten to fifteen-fold  increase in heptachlor residues found  in body
fat, milk  butterfat,  and  in the  fat of poultry,  eggs, and livestock  as com-
pared to residue levels found  in  their  normal food rations (U.S. EPA, 1976).

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Heptachlor  and its  epoxide  pass readily  through  the placenta  (U.S.  EPA,
1979).  The  epoxide can be  found in over 90  percent of the U.S. population
at approximate mean levels of 0.08 to 0.09 mg/kg  (Kutz,  at  al.  1977).
     C.  Excretion
         Elimination  of  non-stored  heptachlor   and  its  metabolites  occurs
within the  first  five days, chiefly in  the  feces and  to  a lesser extent  in
the urine (Mizyukova  and  Kurchatov,  1970).   In addition, a primary route for
excretion in females  is  through lactation,  mostly as  the epoxide.   Levels
can be as high as 2.05 mg/1  (Jonsson, et al. 1977).'
IV.  EFFECTS
     A.  Carcinogenicity
         The studies  on rats have generated, much controversy,  especially for
doses around 10 mg/kg/day.   However,  heptachlor and/or heptachlor epoxide  (1
to 18  mg/kg/day  of unspecified  purities) have induced hepatocellular  carci-
nomas  in  mice  during  three chronic  feeding  studies.   Heptachlor  epoxide
(also of  unspecified purity) has produced  the same  response  in rats  in one
study (Epstein,  1976; U.S.  EPA,  1977).   Clearly,  studies  with chemicals  of
specified purity  still need to be performed  to  establish if contaminants  or
species differences are responsible for the observed effects.
     8.  Mutagenicity
         Heptachlor  has  been  reported  to be  mutagenic in mammalian  assays
but  not  in  bacterial assays.   Heptachlor  (1  to  5  mg/kg)  caused  dominant
lethal changes in male rats  as  demonstrated  by the number 'of resorbed  fetus-
es in intact pregnant rats  (Carey, et al. 1973).  Bone marrow cells of  the
treated animals showed increases  in  the  incidence of abnormal  mitoses, chro-
                                                                          »
matid abnormalities,  pulverization,  and translocation.  9oth  heptachlor and
heptachlor  epoxide  induced  unscheduled  DNA  synthesis in  SV-AO transformed

                                   tlt^j /n -
                                  * I ai V "

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human  cells  (VA-4)  in  culture  with metabolic  activation  (Ahmed,  et al.
1977).  Neither  heptachlor nor  heptachlor epoxide was  mutagenic for Salmo-
nella tvphimurium in the Ames test (Marshall, et al. 1976).
     C.  Teratogenicity
         In  long-term feeding  studies with  heptachlor, cataracts  developed
in  the parent  rats and  in the  offspring shortly  after  their  eyes opened
(Mestitzova,  1967).
     D.  Other Reproductive Effects
         In  long-term feeding  studies  in rats, heptachlor  caused a marked
decrease in litter size and  a  decreased  lifespan  in  suckling rats (Mestit-
zova,  1967).   However,  newborn  rats were less susceptible  to  heptachlor than
adults  (Harbison, 1975).
     E.  Chronic Toxicity
         Little  information on  chronic effects is  available.   When  admini-
stered  to  rats  in  small daily doses  over  a  prolonged  period of  time, hepta-
chlor  induced alterations  in  glucose  homeostasis  which were  thought  to  be
related to an initial stimulation of the cyclic AMP-adenylate  cyclase system
in  liver  and kidney  cortex  (Kacew  and  Singhal,  1973,   1974;  Singhal  and
Kacew, 1976).
     F.  Other Relevant Information
         Heptachlor  is  a convulsant  (St.  Omer,  1971).   Rats fed protein-de-
ficient diets are  less susceptible  to heptachlor and  have lower heptachlor
epoxidase  activities  than pair-fed controls  (Webb  and  Miranda, 1973; Miran-
da,  et al.  1973;  Miranda  and  Webb,  1974).   Pnenobarbital  potentiates the
toxicity of heptachlor  in newborn  rats  (Harbison,  1975).  Many  liver and
brain enzymes are affected  by heptachlor down to 2 mg/kg doses in pigs  (U.S.
EPA, 1979).

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V.   AQUATIC TOXICITY
     A.  Acute Toxicity
        •Numerous  studies on the acute  toxicity of heptachlor  to  freshwater
fish and invertebrate  species  have  been conducted.  Many of these  studies on
heptachlor  have  used technical grade  material.   Available data  suggest  that
toxicity of the  technical material  is attributable to the heptachlor  and its
degradation product,  heptachlor  epoxide,  and that  toxicities of  these  com-
pounds  are  similar (Schimmel,  at  al.  1976).   In  addition,  during  toxicity
testing with  heptachlor, there is  apparently an  appreciable  loss of hepta-
chlor by volatilization  due  to aeration or mixing, leading to variability of
static  and flow-through results  (Schimmel,  et al.  1976;  Goodman,  at  al.
1978).
         Fish  are less  sensitive  to  heptachlor than  are invertebrate  spe.-
cies.   Ninety-six hour  l_C50 values  for  fish  range  from  7.0  ug/1   for  the
rainbow trout,  Salmo  qairdneri,  (Macek,  et  al. 1969)  to 320  ug/1   for  the
.goldfish  (Carassius auratus).   Ten  days  after a dose  of 0.863  ug/g    C-
heptachlor  to  goldfish,  91.2  percent  was  unchanged,  5.4 percent  was hepta-
chlor  epoxide,   1  percent was  hydroxychlordene, 1.1  percent  was  1-hydroxy-
2,3-epoxychlordene and 1.2 percent was  a  conjugate  (Feroz  and Khan,  1979).
Reported values  for  invertebrate  species  range  from  0.9 pg/1 for  the stone-
fly, Pteronarcella badia,  (Sanders  and Cope,  1968)  to 80 ug/1 for  the clado-
ceran  (Simoceohalus serrulatis).   These  data  indicate  that  heptachlor  is
generally highly toxic in acute exposures.
         The relative  toxicity of heptachlor to its'  common degradation  pro-
duct, heptachlor epoxide,  is 52 ug/1  to 120 ug/1-as  determined  in a  26-hour
                                                                       »
LCs  Oaohnia maona bioassay (Frear and Soyd, 1967).

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         Heptachlor  has  been shown to be  acutely toxic to a number  of salt-
water  fish  and invertebrate  species.   The 96-hour  LC5Q values derived  from
flow-through tests on  four  fish species  range from 0.85 to 10.5 jjg/1 (Hansen
and Parrish,  1977;  Korn and.Earnest,  1974;  Schimmel, et al. 1976).   Results
of static exposures  of eight fish species are  from  0.8 to 194 ug/1  (Eisler,
1970;  Kutz,  1961).   The commercially valuable pink  shrimp  (Penaeus  duorarum)
is especially  sensitive, with  reported  96-hour  values  as low  as 0.03  jjg/1
(Schimmel,  et  al. 1976).   Other species  such  as the blue crab,  Callinectes
sapidus, and American  oyster, Crassostrea virginica, are 2,100 and  950 times
less sensitive, respectively, than the pink shrimp  (Butler, 1963).
     B.  Chronic  Toxicity
         In a  40-week  life  cycle test with fathead minnows (Pimephales prom-
elas),  the  determined  no-adverse-effect  concentration  was 0.86  jjg/1.   All
•••IBM**                                                                        ^
fish  exposed  to  1.84 ug/1  were dead  after 60  days (Macek,  et  al.  1976).
Valid  chronic  test  data are not available for  any aquatic invertebrate  spe-
cies.
         In a  28-day exposure starting with sheepshead minnow embryo (Cypri-
nodon  varieqatus) growth of fry was significantly  reduced  at  2.04 jug/l,  the
safe dose being  at  1.22 jug/1 (Goodman,  et al.  1978).  In an 18-week partial
life cycle exposure  with this same  species,  egg production was significantly
decreased at 0.71 jug/1 (Hansen and Parrish, 1977).
     C.  Plant Effects
         In the only study  available,  a  concentration of 1,000 jug/1  caused a
94.4 percent  decrease  in productivity of a natural- saltwater  phytoplankton
community after a 4-hour exposure to heptachlor (Butler, 1963).
     0.  Residues
         The amount  of  total residues,   heptachlor  and heptachlor   epoxide,
accumulated by fathead minnows after  276 days  of exposure  was found to be

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20,000 times  the concentration  in water  (Macek,  et  al.  1976).  Heptachlor
epoxide constituted  10-24 percent  of the  total residue.   Adult sheepshead
minnows exposed  to  technical grade material  for 126 days accumulated  hepta-
chlor and  heptachlor epoxide 37,000 times  over  the concentration of ambient
water  (Hansen and  Parrish,  1977).   Juvenile sheepshead  minnows exposed  in
two  separate  experiments  for  28 days  bioconcentrated heptachlor  5,700  and
7,513 times the  concentration in the water (Hansen and Parrish, 1977;  Good-
man, et al. 1976).
VI.  EXISTING GUIDELINES AND STANDARDS
     The issue  of the carcinogenicity  of heptachlor  in  humans is being  re-
viewed; thus, it is possible that the human health criterion will be changed.
     A.  Human
         Based on the data for the carcinogenicity  of heptachlor epoxide  in
mice (Davis,  1965),  and using the  "one-hit"  model,  the U.S.  EPA (1979)  has
estimated  levels of heptachlor/heptachlor  epoxide  in ambient  water  which
will result in risk levels of human cancer as specified in the  table below. ,
Exposure Assumptions            Risk Levels and  Corresponding Draft Criteria
     (per day)
                                0         10-7           10-6        iQ-5
2 liters of drinking water      0       0.0023 ng/1     0.023 ng/1  0.23 ng/1
and consumption of 13.7
grams fish and shellfish.
Consumption of fish and         0       0.0023 ng/1     0.023 ng/1  0.23 ng/1
shellfish only.
                       Existing Guidelines and Standard? ._
Agency                     Published Standard       '     Reference
Occup.  Safety           500 ug/m^* on skin from  air    Natl. Inst. Occyp.
  Health Admin.                                          Safety Health,  1977
Am. Conf. Gov.          500 ug/rn-^ inhaled              Am. Conf. Gov. Ind.
  Ind.  Hyg. (TLV)                                        Hyg.,  1971
world Health Org.       0.5 ug/kg/day acceptable       Natl. Acad. Sci., 1977
                          daily intake in diet
                                    / o r-LL
                                  *T LA f I   .

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U.S. Publ. Health       Recommended drinking water     Natl. Acad.  Sci.,  1977
  Serv. Adv. Comm.        standard (1968) 18 jjg/1 of
                          heptachlor and 18 )jg/l of
                          heptachlor epoxide

*Time-weighted average


     B.  Aquatic

         For  heptachlor the.  draft criterion  to protect  freshwater aquatic

life is  0.0015 jjg/1 as  a  24-hour average,-  not to  exceed 0.45  ug/1 at  any

time.  To protect  saltwater aquatic  life,  the draft criterion is 0.0036  ug/1

as a 24-hour average, not to exceed 0.05 ug/1 at any time  (U.S. EPA,  1979).

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                          HEPTACHLOR
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Epstein, S.S.  1976.  Carcinogenic!ty of heptachlor and
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                                                   14
Feroz, M., and M.A.Q. Khan.  1979.  Metabolism of   C-hepta-
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Toxicol. 3: 519.

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Frear, D.E.H., and J.E. Boyd.  1967.  Use of Daphnia magna
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Goodman, L.R.-, et al.  1978.  Effects of heptachlor and
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Hansen, D.J., and P.R. Parrish.  1977.  Suitability of sheeps-
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Harbison, R.D.  1975.  Comparative toxicity of selected
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Johnson, R.D., and D.D. Manske.  1977.  Pesticide and other
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Jonsson, V., et al.  1977.  Chlorohydrocarbon pesticide
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Kacew, S., and R.L. Singhal.  1973.  The influence of p,p -
DDT, and chlordane, heptachlor and endrin on hepatic and
renal carbohydrate metabolism and cyclic AMP-adenyl cyclas«e
system.  Life Sci. 13: 1363.

-------
Kacew, S-., and R.L. Singhal.  1974.  Effect  of  certain,halo-
genated hydrocarbon insecticides on cyclic adenosine  3  ,5i-
monophosphate- H formation by rat kidney cortex.   Jour.
Pharmacol. Exp.  Ther. 183: 265.

Khan, M.H., et al.  1969.  Insect metabolism of photoaldrin
and photodieldrin.  Science 164:  318.

Khan, M.A.Q., et al.  1973.  Toxicity-metabolism  relation-
ship of the photoisomers of certain chlorinated cyclodien
insecticide chemicals.  Arch. Environ. Contain.  Toxicol.
1: 159.

Korn, S., and R. Earnest.  1974.  Acute toxicity  of twenty
insecticides to the striped bass, Morone saxtilis.  Calif.
Fish Game 60: 128.

Kutz, F.W., et al.  1977.  Survey of pesticide  residues
and their metabolites in humans.  In: Pesticide management
and insecticide resistance.  Academic. Press, New  York.

Kutz, M.  1961.  Acute toxicity of some organic insecticides
to three species of salmonids and to the threespine stickle-
back.  Trans. Am. Fish. Soc. 90:"264.

Lichtenstein, E.P.  1960.  Insecticidal residues  in various
crops grown in soils treated with abnormal rates  of aldrin
and heptachlor.  Agric. Food Chera. 8: 448.

Lichtenstein, E.P., et al.  1970.  Degradation  of aldrin
and heptachlor in field soils.  Agric. Food  Chem.  18:   100.

Lichtenstein, E.P., et al.  1971.  Effects of a cover crop
versus soil cultivation on the fate of vertical distribution
of insecticide residues in soil 7 to 11 years after soil
treatment.  Pestic. Monitor. Jour. 5: 218.

Macek, K.J., et al.  1969.  The effects of temperature on
the susceptibility of bluegills and rainbow  trout to  selected
pesticides.  Bull. Environ. Contam. Toxicol. 4:174.

Macek, K.J., et al.  1976.  Toxicity of four pesticides
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Marshall, T.C., et al.  1976.  Screening" *of  pesticides for
tnutagenic potential using Salmonella typhimurium  mutants.
Jour. Agric. Food Chem. 241 TSTT      '"

Matsumura, F., and J.O. Nelson.  1971.  Identification of
the major metabolite product of heptachlor epoxide in rat '
feces.  Bull. Environ. Contam. Toxicol. 5: 489.
                           I

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Mestitzova, M.  1967.  On reproduction studies on the occur-
rence of cataracts in rats after long-term feeding of the
insecticide heptachlor.  Experientia 23: 42.

Miranda, C.L., and R.E. Webb.  1974.  Effect of diet and
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Miranda, C.L., et al.  1973.  Effect of dietary protein
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                              i 1 >-*x
                             / c* / (/ -

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                                      No. 109
         Heptachlor Epoxide

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented  by the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

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                      HEPTACHLOR  EPOXIDE
                           SUMMARY
     Heptachlor epoxide is the principal metabolite of hepta-
chlor in microorganisms,  soil, plants,  animals,  and probably
man,  and  is  more acutely  toxic  than  the parent  compound.
Its  intrinsic  effects  are  difficult  to gauge  since  most
of  the  relevant  data in the  literature  is  a side  product
of  the  effects  of technical heptachlor.   Heptachlor  epoxide
(mostly  of  unspecified  purity)   has  induced  liver  cancer
in  mice and  rats and  was  mutagenic  in  a  mammalian  assay
system, but. not in a bacterial system.   Pertinent information
on  teratogenicity  and chronic toxicity could not  be  located
in  the  available  literature.   Heptachlor epoxide  accumulates
in adipose tissue..
     The chronic  value  for  the  compound derived from  a  26-
hour exposure of  Daphnia  magna  is  reported  to be  120  ug/1,
approximately the same value obtained for heptachlor.
     Fathead  minnows  bioconcentrated   heptachlor  and  its
biodegradation  product,   heptachlor  expoxide,  20,000  times
after 276 days  of exposure.   Heptachlor  epoxide  constituted
between 10 and 24 percent of the  total residue.
                         /09-J

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                      HEPTACHLOR EPOXIDE
I.   INTRODUCTION
     This  profile  is  based  on  the  Ambient  Water  Quality
Criteria Document for Heptachlor  (U.S. EPA,  1979a).
     Heptachlor epoxide is the principal metabolite  of hepta-
chlor in microorganisms,  soil,  plants, and mammals,  although
the  conversion in  man may  be  less  efficient  (Tashiro  and
Matsumura, 1978) .   Since  much of the  data has been  obtained
as  a side-product  of  the effects  of  technical heptachlor
and  the  purity of  the epoxide  is  often  unspecified,  there
is  a paucity  of  reliable literature  on  its  biological  ef-
fects (U.S. EPA, 1979a).
     Heptachlor  epoxide  is  relatively   persistent   in   the
environment  but has  been shown  to undergo photodecomposi-
tion  to  photoheptachlor  epoxide  (Graham,  et  al.   1973).
Photoheptachlor epoxide has been  reported  to exhibit  greater
toxicity than heptachlor epoxide  (Ivie, et al. 1972).   Hepta-
chlor epoxide  will  bioconcentrate   in numerous  species  and
will accumulate in the food chain (U.S. EPA, 1979a).
II.  EXPOSURE
     A.    Water
          Heptachlor  epoxide  has  been  detected  by  various
investigators in the major river  basins  of the United  States
(U.S. EPA,   1979a)  at  levels ranging from  0.001  to 0.020
ug/1 (Breidenbach, et al.  1967).
     B.    Food
          The PDA showed in their market basket survey  (1974-
1975) of  20  different  cities  that 3 of  12 food  classes con-

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tained  residues of  heptachlor  epoxide ranging  from  0.0006
to 0.003 ppm  (Johnson  and  Manske,  1977).  Heptachlor  epoxide
residues  greater  than 0.03  mg/kg were found  in 14  to  19
percent  of  red meat,  poultry,  and  dairy  products  during
the  period 1964-1974.   Average  daily  intake  was estimated
to be between  0.3  to 3 ug  from  1965  to 1974 (Nisbet,  1977).
Heptachlor  and/or  heptachlor  epoxide  were  found  in  32  per-
cent  of 590  fish  samples obtained  nationally,  with whole
fish  residues  containing  0.01  to  8.33  mg/kg   (Henderson,
et al.  1969) .    Human  milk  can  be contaminated  with  hepta-
chlor epoxi.de;  63  percent of  samples  in 1975-1976 contained
1  to 2,050 ug/1  (fat  adjusted)  (Savage,  1976).   Levels  of
5  ng/1  have been  reported in evaporated milk.   Cooking did
not  reduce the residue  level in  poultry  meat  by more  than
one-half (Ritcey, et al. 1972).
          The  U.S.  EPA  (1979a)   has  estimated  the  weighted
average  bioconcentration  factor  for  heptachlor to  be 5,200
for  the edible  portions  of  fish  and  shellfish  consumed  by
Americans..   This estimate  is based on  the  measured  steady-
state  bioconcentration  studies   in  three  species of fish.
Since heptachlor epoxide is  the  primary metabolite of  hepta-
chlor and  shows greater persistence  in body fat  (U.S.  EPA,
1976) , it may be assumed that  heptachlor  epoxide  is bioconcen-
trated to at least the same extent as heptachlor.
     C.    Inhalation
          Heptachlor  epoxide  is  present  in ambient  air ,to
a  lesser  extent than heptachlor  and  is not  thought  to  con-

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tribute  substantially  to  human  exposure  except   in  areas
near sprayed fields, where  concentrations  of up to 9.3 pg/ra
may be encountered  (Nisbet, 1977).
     0.   Dermal
          Gaines  (1960)   found  rat  dermal  LDen  values  of
195 and  200  mg/kg  for  males and  females,  respectively, com-
pared  with  oral LDcQ' s of  100 and  162  mg/kg,  respectively,
for technical  heptachlor.   Thus,   it  is likely  that  dermal
exposure in humans can be important under certain conditions.
III. PHAEMACOKINETICS
     A.   Absorption
          Heptachlor epoxide  is  readily  absorbed   from  the
gastrointestinal tract (U.S. EPA, 1979a).
     B.   Distribution
          Studies dealing directly  with exposure to  hepta-
chlor  epoxide could  not  be located  in the  available litera-
ture.    After oral  administration  of  heptachlor to experi-
mental  animals,  high  concentrations  of heptachlor  epoxide
have been  found in  fat,  with much  lower  levels in  liver,
kidney, and muscle, and none in  brain  (Radomski and  Davidow,
1953).    Another study  (Mizyukova  and Kurchatav,  1970)  also
demonstrated the  persistence  of  heptachlor epoxide in  fat.
Levels in  fatty  tissues  stabilize  after three  to  six  months
after   a  single  dose.   The  U.S.   EPA  (1979a)  states  that
there   is approximately 10-  to  15-fold increase  in heptachlor
                                                          »
residues found  in  body  fat, milk  butter fat,  and in the  fat
of poultry eggs  and livestock  as  compared  to residue  levels
found   in  their  normal  food  rations.   "Heptachlor  residues"
                          /of-/

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probably  refers  primarily to heptachlor epoxide.   Heptachlor
epoxide passes readily through  the placenta  (U.S.  EPA,  1979a)
and could  be  found  in over 90 percent of the U.S.  population
at average levels of  around  90  ng/kg  (Kutz,  et  al.  1977).
     C.   Metabolism  and  Elimination
          Heptachlor  epoxide accumulates  in adipose  tissue,
as  discussed  in  the  "Distribution"  section.    The  primary
route for excretion is fecal  (Mizyukova and  Kurchatav,  1970).
When  heptachlor  epoxide  was fed  to  rats  over a  period of
30  days,  approximately  20  percent of  the  administered  dose
(approximately 5  mg  heptachlor epoxide/rat/30 day)  was ex-
creted  in the feces,  primarily  as  1-exo-hydroxyheptachlor
epoxide   and   1,2-dihydroxydihydrochlordene  (Matsumura  and
Nelson,  1971; Tashiro  and  Matsuraura,  1978).    In  females,
a  primary route for excretion  is  via   lactation,   usually
as the  epoxide.   Levels can be as  high  as 2.05 mg/1  (Jonas-
son, et al. 1977).
IV.  EFFECTS
     A.   Carcinogenicity
          Heptachlor  epoxide of  unspecified  purity  induced
hepatocellular carcinoma  in a  chronic  feeding   study  with
mice  and  in   one  study with  rats  (Epstein,  1976;  U.S.  EPA,
1977) .
     B.   Mutagenicity
          Heptachlor  epoxide  induced  unscheduled  DNA  syn-
thesis  in SV-40   transformed human cells   (VA-4)   in culture
when  metabolically  activated  (Ahmed,  et  al.  1977),  but was

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not  mutagenic for  Salmonella typhimurium  in the  Ames test
(Marshall, et al. 1976) .
     C.   Teratogenicity,  Other   Reproductive  Effects  and
          Chronic Toxicity
          Pertinent data  could not be  located in the avail-
able literature.
     0.   Other Relevant Information
          Heptachlor  epoxide  is  more  acutely  toxic  than
heptachlor  (U.S.  EPA,  1979a) .  It  inhibits synaptic calcium
magnesium dependent ATPases in rats (Yamaguchi, et al. 1979).
V .   AQUATIC TOXICITY
     A.   Acute Toxicity
          Acute  toxicity  data could  not  be  located  in the
available literature  relative to  the effects of heptachlor
epoxide on fish or invertebrates.
     3.   Chronic Toxicity
          In  the only  reported  chronic  study,   the 26-hour
LCcn for • heptachlor epoxide  in  Daphnia magna was  120  ug/1
(Frear and Boyd,  1967) .   In  the  same  test, the corresponding
value for heptachlor was 52 ug/1.
     C.   Plant Effects
          Data  on  the  toxicity  of  heptachlor  epoxide  to
plants could not be located in the available literature.
     D.   Residues          *
          Macek,  et al.  (1976) determined 'the bioconcentra-
tion factor of  20,000  for heptachlor and  heptachlor epoxide
                                                          »
in  fathead  minnows  after 276 days'   exposure.    Heptachlor
epoxide  residues were  reported  as  constituting  10  to  24
percent of the  total residue.  The  geometric mean bioconcen-
                          /or-3

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    tration  factor  for heptachloc  in  all species of  fish  tested

    is 11,400  (U.S.  EPA, 1979a).   As explained  in the  "Distri-

    bution"  section of  this  text, the  bioconcentration  factor

    for  heptachlor  epoxide would  be as  least as  great as  that

    for heptachlor.

    VI.  EXISTING GUIDELINES AND STANDARDS

         A.   Human

              The  existing  guidelines  and standards  for  hepta-

    chlor and heptachlor epoxide are:
 AGENCY/ORG.

Occup. Safety
 Health Admin.

Am. Conf. Gov.
 Ind. Hyg. (TLV)

Fed. Republic
 Germany

Soviet Union
World Health
 Organ.**

U.S. Pub. Health
 Serv. Adv. Comm.
          STANDARD
500 ug/m * on skin from air
500 ug/m  inhaled


500 ug/ra3 inhaled
10 ug/m  ceiling value
 inhaled

0.5 ug/kg/day acceptable
 daily intake in diet

Recommended drinking water
 standard (1968) 18 pg/1 of
 heptachlor and 18 ug/1
 heptachlor epoxide
    REFERENCE

Natl. Inst. Occup.
 Safety Health, 1977

Am. Conf. Gov. Ind.
 Hyg., 1971

Winell, 1975
Winell, 1975
Natl. Acad. Sci.,
 1977

Natl. Acad. Sci.,
 1977
*   Time weighted average

** Maximum residue limits in certain foods can be found in Food Agric,
   Organ./World Health Organ. 1977, 1978


              The U.S. EPA (1979a)  is in the  process of establish-

    ing ambient water quality criteria  for  heptachlor and hepta-

    chlor epoxide.   Based on potential carcinogenicity of hepta-

    chlor epoxide, the draft criterion  is calculated on the esti-

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mate that 0.47  ng/raan/day would result in an increased addi-
tional  lifetime cancer  risk  of  no  more  than  1/100,000.
Based  on  this  lifetime  carcinogenicity  study  of heptachlor
epoxide at  10  ppm in  the diet  of C3Heb/Pe/J  strain mice,
the  recommended draft  criterion  is  calculated  to  be 0.233
ng/1.
     B.   AQUATIC
          No  existing guidelines  are  available  for  hepta-
chlor epoxide.  However, since heptachlor epoxide  is a biode-
gradation product of heptachlor, the hazard profile on hepta-
chlor should be consulted  *U.S. EPA, 1979b).
                             *

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                              HEPTACHLOR EPOXIDE

                                  REFERENCES
Ahmed, F.E.,  et al.  1977.   Pesticide-induced  DNA damage and  its repair in
cultured human cells.  Mutat. Res.  42: 1612.

American Conference  of  Governmental Industrial  Hygienists.   1971.  Documen-
tation of  the threshold limit values  for substances in workroom  air.   3rd.
«U

Breidenbach,  A.W.,  et   al.   1967.   Chlorinated  hydrocarbon  pesticides  in
major river basins, 1957-65.  Pub. Health Rep.  82: 139.

Epstein,  S.S.   1976.   Carcinogenicity  of heptachlor  and chlordane.   Sci.
Total Environ.  6: 103.

Frear, D.E.H.  and  J.E.  Boyd.  1967.  Use  of  Daphnia magna for the microbio-
assay and  pesticides.   I. Development  of standardized techniques for rearing
Daohnia  and preparation  of dosage-mortality curves for  pesticides.   Jour.
Econ. Entomol.  60: 1228.

Gaines,  T.B.   1960.  The acute  toxicity of pesticides  to  rats.   Toxicol..
Appl. Pharmacol.  2:88.                                                     i

Graham, R.E.,  et al.  1973.   Photochemical decomposition of heptachlor epox-
ide.  Jour. Agric. Food Chem.  21: 284.

Henderson,  C., et  al.   1969.  Organochlorine  insecticide residues  in  fish
(National Pesticide Monitoring Program).  Pestic. Monitor.  Jour.  3: 145.

Ivie, G.W., et al.   1972.   Novel  photoproducts  of heptachlor  epoxide,  trans-
chlordane, and trans-nonachlor.   Bull.  Environ.  Contain. Toxicol.  7: 376.

Johnson, R.D.  and D.D.  Manske.  1977.   Pesticide and other chemical residues
in total diet samples (XI).   Pestic. Monitor. Jour.  11: 116.

Jonasson,  V.,  et al.   1977.  Chlorohydrocarbon pesticide  residues  in  human
milk in greater St. Louis, Missouri, 1977.  Am.  Jour. Clin. Nutr.  30: 1106.

Kutz, F.w.,  et al.   1977.   Survey of  pesticide residues  and  their metabo-
lites  in  humans.   In:    Pesticide management   and  insecticide  resistance.
Academic Press, New York.

Macek, K.J.,  et  al.  1976.    Toxicity  of four pesticides to  water  fleas  and
fathead minnows.  U.S. Environ.  Prot. Agency,  EPA-600/3-76-099.

Marshall,  T.C., et  al.   1976.  Screening of pesticides for mutagenic  poten-
tial using Salmonella typhimurium  mutants.  Jour. Agric. Food  Chem.   24:  560.

Matsumura, F.  and J.O.  Nelson.  1971.   Identification of the  major metabolic
product of heptachlor epoxide in  rat feces.   Bull. Environ. Contam.  Toxicol.
5: 489.
                                /&?•//

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 Mizyukova,  I.G.  and  G.V.  Kurchatav.   1970.   Metabolism  of  heptachlor.
 Russian Pharmacol.  Toxicol.  33: 212.

 National   Academy  of   Sciences.    1977.    Drinking   water   and  health.
 Washington, O.C.

 National  Institute  for  Occupational Safety and  Health.   1977.   Agricultural
 chemicals  and  pesticides: a  subfield of the  registry  of toxic  effects of
 chemical substances.

 Nisbet,  I.C.T.   1977.   Human exposure  to chlordane,  heptachlor  and  their
 metabolites.   Unpubl.  rev.  prepared,  for  Cancer  Assessment  Group,  U.S.
 Environ. Prat.  Agency, Washington, O.C.

 Radmoski,  J.L.  and 8.  Oavidow.    1953.   The metabolite of  heptachlor,  its
 estimation, storage, and toxicity.  Jour. Phaimacol. Exp. Ther.   107: 266.

 Ritcey,  W.R.,  et al.   1972.  Organochlorine  insecticide residues  in  human
 milk, evaporated milk and some milk substitutes  in  Canada.  Can. Jour. Publ.
 Health.  63: 125.

 Savage,  E.P.    1976.   National  study  to  determine  levels  of  chlorinated
 hydrocarbon insecticides  in  human milk.   Unpubl.  rep.  submitted to  U.S.
 Environ. Prat.  Agency.
                                                                             4
 Tashiro, S. and  F.  Matsumura.  1978.   Metabolism of  trans-nonachlor and  re-
 lated chlortane  components  in rat and man.  Arch.  Environ.  Contain. Toxicol.
 7: 113

 U.S. EPA.   1977.  Risk assessment  of chlordane and  heptachlor.   Carcinogen
 Assessment Group.  U.S.  Environ.  Prat. Agency,  Washington, O.C.   Unpubl.  rep.

"U.S. EPA.  1979a.  Heptachlor:  Ambient Water Quality Criteria (Draft).

 U.S. EPA.  1979b.  Environmental Criteria and Assessment Office.   Heptachlor
 Epoxide: Hazard Profile.  (Draft)

 Winell, M.A.   1975.   An international comparison  of hygienic standards  for
 chemicals in the work  environment.   Ambio.   4:  34.

 Yamaguchi, I.,  et al.   1979.  Inhibition of  synaptic atpases by  heptachlor
 epoxide in rat  brain.  Pest. Biochem.  Physiol.   11:  285.
                                J

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                                   No. HO
        Hexachlorobenzene

  Health and Environmental Effects
U.S.  ENVIRONMENTAL PROTECTION AGENCY
      WASHINGTON,  D.C.  20A60

          APRIL 30, 1980
        /to--/

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone scrutiny  to
ensure its technical accuracy.

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                      SPECIAL NOTATION
U.S.  EPA13  Carcinogen Assessment Group  (CAG) has evaluated




hexachlorobenzene and has found sufficient  evidence to




indicate  that  this compound is carcinogenic.
                             //0-3

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                               HEXACHLOROBENZENE
                                    Summary

     Hexachlorobenzene is ubiquitous  in the environment and has an extremely
slow  rate of  degradation.   Ingested  hexachlorobenzene is  absorbed  readily
when  associated  with lipid material  and,  once absorbed,  is stored for  long
periods  of time  in the  body  fat.   Chronic exposures  can cause  liver  and
spleen damage and can induce the hepatic microsomal mixed functional  oxidase
enzyme.  Hexachlorobenzene can pass the placenta! barrier and produce toxic
or lethal  effects on the  fetus.  Hexachlorobenzene appears  to be neither a
teratogen nor a  mutagen;  however, this compound has produced tumors in  both
rats and mice.
     In  the  only  steady-state  study  with  hexachlorobenzene,  the  pinfish,
Laqodon.  rhoimboides, bioconcentrated  this compound  23,000 times in  42  days
of exposure.  The concentration  of  HC8 in muscle of pinfish was reduced  only
16 percent  after 28 days of depuration,  a  rate  similar to that  for  DOT in
fish.

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                               HEXACHLOROBENZENE
I.   INTRODUCTION
     This  profile  is based  on the Ambient  Water Quality Criteria  Document
for Chlorinated Benzenes (U.S. EPA, 1979).
     Hexachlorobenzene  (HCB;  CgCl^;  molecular weight  284.79)  is  a  color-
less solid with  a  pleasant aroma.   Hexachlorobenzene has a melting point  of
230°C,  a  boiling  point  of 322°C,  a density  of 2.044- g/ml,  and  is  vir-
tually  insoluble  in  water.   Hexachlorobenzene  is  used  in  the control  of
fungal  diseases  in cereal seeds intended  solely  for planting,  as a  plasti-
cizer for polyvinyl chloride, and as a flame retardant (U.S.  EPA, 1979).
     Commercial production  of  hexachlorobenzene in the  U.S.  was  discontinued
in 1976 (Chem. Econ.  Hdbk.,  1977).   However,  even prior to 1976, most, hexa-
chlorobenzene was  produced as a waste by-product during the  manufacture 4'of
perchloroethylene, carbon tetrachloride,  trichloroethylene, and  other  chlor-
inated  hydrocarbons.  This is  still the major  source of hexachlorobenzene  in
the  U.S.,  with  2,200 kg  being produced  by  these   industries  during  1972
(Mumma  and Lawless, 1975).
II.  EXPOSURE
     A.  Water
         Very little  is known regarding  potential  exposure  to  hexachloro-
benzene  as a result of ingestion  of  contaminated water.  Hexachlorobenzene
has been detected  in specific bodies of  water,  particularly near points  of
industrial discharge  (U.S.  EPA,  1979).  Hexachlorobenzene  has been detected
in the  polluted  waters  of the Mississippi River  (usually below  2 ng/kg) and
in  the   clean waters of  Lake  Superior  (concentrations  not  quantitatively
measured).    Hexachlorobenzene  was  detected  in drinking water  supdlies  at
                                 \io-s

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three  locations,  with  concentrations  ranging  from  6 to  10  ng/kg,  and  in
finished drinking water at two locations, with concentrations  ranging  from 4
to 6 ng/kg (U.S. EPA, 1975).
     B.  Food
         Ingestion of excessive amounts of hexachlorobenzene has been  a con-
sequence  of  carelessness,  usually  from  feeding  seed grains  to  livestock.
Foods high in  animal  fat (e.g., meat, eggs, butter,  and  milk)  have the high-
est concentrations of hexachlorobenzene.   The daily intake of  hexachloroben-
zene by  infants from human breast  milk in part of Australia was 39.5 jjg per
day per 4 kg baby.  This exceeded the acceptable daily intake  recommended by
the FAO/WHO of 2.4 jjg/kg/day  (1974).  The dietary intake by young  adults (15
to 18-year old  males.) was estimated to be 35 jug hexachlorobenzene  per  person
per  day (Miller  and Fox,  1973).   The  U.S.  EPA  (1979) has  estimated  the
weighted average bioconcentration factor for hexachlorobenzene to be  12,000
for the  edible portions  of  fish and  shellfish  consumed by Americans.   This
estimate is based  on the octanol/water partition  coefficient  of hexachloro-
benzene.
     C-  Inhalation
         Hexachlorobenzene  enters  the air  by   various  mechanisms,  such  as
release  from   stacks  and  vents  of industrial  plants,   volatilization  from
waste dumps and impoundments,  intentional spraying and dusting, and uninten-
tional dispersion  of hexachlorobenzene-laden  dust from  manufacturing  sites
(U.S. EPA  1979).   No  data  is  given on  the concentrations  of hexachloro-
benzene  in  ambient air.   Significant occupational"  exposure  can  occur  par-
ticularly to pest control operators  (Simpson and Chandar, 1972).

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     0.  Dermal
         Hexachlorobenzene may  enter the body by absorption through the  skin
as a result of skin contamination (U.S. EPA,  1979).
III. PHARMACOKINETICS
     A.  Absorption
         To date, only  absorption of hexachlorobenzene from the gut has  been
examined  in detail.   Hexachlorobenzene in  aqueous suspensions  is absorbed
poorly in  the  intestines of rats (Koss and  Koransky,  1975); however, cotton
seed oil  (Albro and  Thomas,  1974)  or olive  oil  (Koss and  Koransky, 1975)
facilitated  the  absorption.   Between  70  and 80  percent  of doses of hexa-
chlorobenzene ranging from 12  mg/kg to 180 mg/kg were absorbed.  Hexachloro-
benzene in  food  products will selectively partition  into  the lipid portion,
and hexachlorobenzene in  lipids  will be absorbed  far  better  than  that in an
aqueous milieu (U.S. EPA, 1979).
     8.  Distribution
         The highest  concentrations  of hexachlorobenzene  are  found  in  fat
tissue (Lu  and Metcalf,  1975).   In rats receiving a  single intraperitoneal
(i.p.)  injection or  oral dose  of  hexachlorobenzene  in olive  oil,  adipose
tissue contained  about 120-fold more hexachlorobenzene than muscle tissue;
liver,  4-fold;  brain,  2.5-fold; and kidney,  1.5-fold  (Koss  and  Koransky,
1975).   Adipose  tissue  serves  as a reservoir for  hexachlorobenzene,  and de-
pletion of  fat deposits  results  in  mobilization and redistribution of stored
hexachlorobenzene.  However, excretion is  not increased, and the  total body
                                                    s
burden is not lowered (Villeneuve,  1975).

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     C.  Metabolism
         Hexachlorobenzene  is metabolized  after i.p.  administration in  the
rat  to pentachlorophenol,  tetrachlorohydroquinone  and  pentachlorothicphenol
(Koss,  et al.  1976).   In  another study using  rats in which the  metabolic
products were  slightly  different,  only a small percentage of  the  metabolites
were present as glucuronide conjugates (Engst, et al.  1976).   Hexachloroben-
zene  appears  to  be  an inducer  of the  hepatic  microsomal  enzyme system  in
rats (Carlson,  1978).   It has been proposed that both  the phenobarbital type
and  the 3-methylcholanthrene type  microsomal enzymes  are  induced  (Stonard,
1975; Stonard and Greig,  1976).
     0.  Excretion
         Hexachlorobenzene  is excreted' mainly in the  feces  and,  to some  ex-
tent,  in  the  urine in the  form  of several metabolites  which  are more  polar
than • the  parent compound  (U.S.  EPA,  1979).   In  the rat, 34  percent of  the
administered hexachlorobenzene was excreted in the feces, mostly as unalter-
ed  hexachlorobenzene.    Fecal excretion  of  unaltered  hexachlorofaenzene  is
presumed to be due to biliary  secretion.   Five percent, of the administered
HCB was excreted in the urine (Koss and Koransky, 1975).
IV.  EFFECTS
     A..  Carcinogenic!ty
         Carcinogenic activity of  hexachlorobenzene was assessed in hamsters
fed 4.3 or 16  mg/kg/day  for life (Cabral,  et  al.  1977).  Whereas 10 percent
of the  unexposed hamsters  developed  tumors,  92 percent of  the hamsters  fed
16 mg/kg/day,  75 percent  fed 8 mg/kg/day,  and 56  percent  fed  4 mg/kg/day
developed  tumors.   The   tumors  were  hepatomas,  haemangioendotheliomas  and
thyroid adenomas.  In a study on mice fed  6.5, 13 or  26 mg/kg/day for  life,
£he only  increase in  tumors was  in  hepatomas (Cabral, et  al.  1978).   How-

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ever,  the incidence of  lung tumors in  strain A mice  treated three times  a
week  for  a total  of 24  injections of  40  mg/kg each  was not  significantly
greater than the incidence in control mice (Theiss, et al.  1977).   Also,  ICR
mice  fed  hexachlorobenzene at  1.5  or 7.0  mg/kg/day  for  24  weeks showed no
induced hepatocellular carcinomas (Shirai,  et  al.. 1978).
     B.  Mutageriicity
         Hexachlorobenzene  was assayed  for mutagenic activity  in the  domi-
nant  lethal assay.   Rats were  administered  60  mg/kg/day  hexachlorobenzene
orally for  ten  days;  there was no significant difference  in  the incidence of
pregnancies (Khera, 1974).
     C.  Teratogenicity
         Hexachlorobenzene  does  not  appear  to be  teratogenic  for  the  rat
(Khera, 1974).   CO-1 mice  receiving  100 mg/kg/day  hexachlorobenzene oraily
on gestational  days 7 to 11  showed a  small increase  in the incidence of  ab-
normal fetuses  per litter (Courtney,  et al. 1976).   However,  the  statistical
significance was  not mentioned,  and  the abnormalities appeared in both  the
exposed and unexposed groups.
     D.  Other Reproductive Effects
         Hexachlorobenzene  can  pass  through  the placenta  and  cause   fetal
toxicity  in rats  (Grant,  et al.  1977).   The  distribution  of hexachloro-
benzene  in the  fetus appears to  be the  same  as  in  the  adult,  with  the
highest concentration in  fatty tissue.
     E.  Chronic Toxicity
         In one long-term study  where  rats were given  50 mg/kg hexachloro-
benzene every  other  day for  53 weeks,  an equilibrium  between  intake  and
elimination was achieved after nine weeks.  Changes in  the  histology of  the

-------
liver and  spleen were noted  (Koss,  et al. 1578).  On  human exposure  for  an
undefined  time  period,  porphyrinuria  has  been shown  to  occur  (Cam  and
Nigogosyan, 1963).
     F.  Other Relevant Information
         At  doses far  below  those  causing mortality,  hexachlorobenzene  en-
hances  the capability  of animals to  metabolize  foreign  organic  compounds.
This type  of interaction may be  of  importance in determining the  effects  of
other concurrently encountered xenobiotics (U.S.  EPA, 1979).
V.   AQUATIC TOXICITY
     A.  NO  pertinent  information is available on acute and chronic  toxicity
or plant effects.
     B-  Residues
         Hexachlorobenzene  (HCS)  is bioconcentrated  from  water into  tissues
of  saltwater  fish  and  invertebrates.   Bioconcentration  factors  (BCF)   in
short 96-hour exposures are as  follow (Parrish, et al. 1974):  grass  shrimp,
Palaeomonetes  puqio,  - 4,116 jjg/1;  pink  shrimp, Penaeus  duorarum,   - 1,964
ug/1; sheepshead minnow,  Cyprinodon variegatus,  - 2,254 yg/1.   In a  42-day
exposure,  the pinfish,  Lagodon  rhomboides,   BCF was  23,000.   The   concen-
tration of HCS in pinfish  muscle was reduced only 16  percent after  28 days
of depuration; this slow  rate is similar to that  for DOT in  fish.
VI.  EXISTING GUIDELINES AND STANDARDS
     Neither  the human  health  nor  aquatic  criteria  derived  by U.S. EPA
(1979),  which are summarized  below,  have gone through  the  process of  public
review;   therefore,   there  is  a  possibility  that 'these   criteria will   be
changed.


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     A.  Human
         The  value  of 0.6  pg/kg/day  hexachlorobenzene  was  suggested  by
FAO/WHO as a reasonable upper limit for  residues  in  food for human consump-
tion (FAO/WHO, 1974).  The Louisiana State Department of Agriculture has set
the tolerated level of hexachlorobenzene  in meat fat  at 0.3 mg/kg (U.S. EPA,
1976).  The FAO/WHO recommendations for residues in foodstuffs are 0.5 mg/kg
in fat for milk and eggs, and 1 mg/kg in  fat  for  meat and poultry (FAO/WHO,
1974).   Based on  bioassay data,  and using  the  "one-hit"  model,  the  EPA
(1979) has estimated levels of hexachlorobenzene in ambient water which will
result in specified risk levels of  human cancer:

Exposure Assumption           Risk Levels  and Corresponding Draft Criteria
   (per day)
                              0       " .  10-7          10-6       io-5
2 liters of drinking water    0       0.0125 ng/1   0.125 ng/1  1.25 ng/1
and consumption of 18.7
grams fish and shellfish.
Consumption of fish and       0       0.0126 ng/1   0.126 ng/1  1.26 ng/1
shellfish only..

     B.  Aquatic
         Pertinent  information  concerning  aquatic  criteria could  not  be
located in the available literature.
                                +f A r 3'
                                  11*6-11

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                      HEXACHLOROBENZENE

                         REFERENCES

Albro, P.W., and R. Thomas.   1974.  Intestinal  absorption of
hexachlorobenzene and hexachlorocyclohexane  isomers in  rats.
Bull. Environ. Contam. Toxicol.   12:  289.

Cabral, J.R.P., et al.  1977.  Carcinogenic  activity of hexa-
chlorobenzene  in hamsters.  Nature  (London).  269:  510.

Cabral, J.R.P., et al.  1978.  Carcinogenesis study in  mice
with hexachlorobenzene.  Toxicol. Appl.  Pharmacol.   45: 323.

Cam, C., and G. Nigogosyan.   1963.  Acquired toxic  porphyria
cutanea tarda  due to hexachlorobenzene.  Jour.  Am.  Med.
Assoc.  183; 88.

Carlson, G.-P.  1978.  Induction of cytochrome P-450 by  halo-
genated benzenes.  Biochem. Pharmacol.   27:  361.

Chemical Economic Handbook.   1977.  Chlorobenzenes-Salient
statistics.  In: Chemical Economic Handbook, Stanford Res..
Inst. Int., Menlo Parkr Calif.. .

Courtney, K.D., et al.  1976.  The effects of pentachloro-
nitrobenzene,  hexachlorobenzene,  and  related compounds  on
fetal development.  Toxicol.  Appl. Pharmacol.   35:  239.

Engst, R., et  al.  1976.  The metabolism of hexachlorobenzene
(HCB) in rats.  Bull. Environ. Contam. Toxicol.  16:  248.

Pood and Agriculture Organization.  1974.  1973 evaluations
of some pesticide residues in food.   FAO/AGP/1973/M/9/1; WHO
Pestle- Residue Ser. 3.  World Health Org., Rome, Italy p.
291.

Grant, D.L., et al.  1977.  Effect of hexachlorobenzene  on
reproduction in the rat.  Arch. Environ. Contam. Toxicol.   5:
207.

Khera, K.S.  1974.  Teratogenicity and dominant lethal
studies on hexachlorobenzene  in rats.  Food Cosmet.  Toxicol.
12: 471.

Koss, R., and W. Koransky.  1975.  Studies on the toxicology
of hexachlorobenzene.  I*  Pharraacokinetics.  Arch  Toxicol.
34: 203.

Koss, G., et al.  1976.  Studies on the toxicology  of hexa-
chlorobenzene.  II. Identification and determination of
metabolites.  Arch. Toxicol.  35: 107.

-------
Koss, G., et al.  1978.  Studies on  the  toxicology  of  hexa-
chlorobenzene.  III. Observations  in a long-term  experiment.
Arch. Toxicol.  40: 285.

Lu, P.Y., and R.L. Metcalf.  1975.   Environmental fate  and
biodegradability of benzene derivatives  as studied  in a model
aquatic  ecosystem.  Environ. Health  Perspect.   10:  269.

Miller,  G.J., and J.A. Fox.  1973.   Chlorinated hydrocarbon
pesticide residues in Queensland human milks.   Med. Jour.
Australia  2: 261.

Mumma, C.E., and E.W. Lawless.  1975.  "Task I  -  Hexachloro-
benzene  and hexachlorobutadiene pollution from  chlorocarbon
processes".  EPA 530-3-75-003, U.S.  Environ. Prot.  Agency,
Washington, D.C.

Parrish, P.R., et al.  1974.  Hexachlorobenzene:  effects on
several  estuarine animals.  Pages  179-187 in Proc.  28th Annu.
Conf. S.E. Assoc. Game Fish Comm.

Shirai,  T., et al.  1978.  Hepatocarcinogenicity  of poly-
chlorinated terphenyl (PCT) in ICR mice  and its enhancement
by hexachlorobenzene (HCB).  Cancer  Lett.  4: 271.

Simpson, G.R., and A. Shandar.  1972.  Exposure to  chlori-    .
nated hydrocarbon pesticides by pest control operators.  Med..
Jour. Australia.  2: 1060.

Stonard, M.D.  1975.  Mixed type hepatic microsomal enzyme
induction by hexachlorobenzene.  Biochem. Pharmacol.  24:
1959.

Stonard, M.D., and J.B. Greig.  1976.  Different  patterns of
hepatic microsomal enzyme activity produced by  administration
of pure  hexachlorobiphenyl isomers and hexachlorobenzene.
Chem.-Biol. Interact.  15: 365.

Theiss, J.C., et al.  1977.  Test  for carcinogenicity of or-
ganic contaminants of United States drinking waters by pul-
monary tumor response in strain A mice.  Cancer Res.  37:
2717.

U.S. EPA.  1975.  Preliminary assessment of suspected carcin-
ogens in drinking water.  Report to Congress.  EPA 560/4-75-
003.  Environ. Prot.. Agency, Washington,  D.C.

U.S. EPA.  1976.  Environmental contamination from hexachloro-
benzene.  EPA 560/6-76-014.  Off. Tox.  Subst.  1-27.

U.S. EPA.  1979.  Chlorinated Benzenes: Ambient Water Quality
Criteria. (Draft).

-------
villeneuve, D.C.  1975.  The effect  of  food  restriction on
the redistribution of hexachlorofaenzene  in  the  rat.   Toxicol.
Appl. Pharmacol.  31: 313.

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                                      No. Ill
        Hexachlorobutadiene




  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY

       WASHINGTON, D.C.  20A60


           APRIL 30, 1980
             - / "i n *»
              / <*. >  >


             111-1

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all. available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                       SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (GAG) has evaluated



hexachlorobutadiene and has found sufficient evidence to




indicate that this compound is carcinogenic.

-------
                     HEXACHLOROBUTADI EH E



                           SUMMARY



     Hexachlorobutadiene (HCBD) is a significant by-product



of the manufacture of chlorinated hydrocarbons.  HCBD has



been found to induce renal neoplasms in rats (Kociba, et al.,



1971).  The mutagenicity of HCBD has not been proven conclu-



sively, but a bacterial assay by Taylor (1978)  suggests a



positive result.  Two studies on the possible teratogenic



effects of HCBD produced conflicting results.



     Ninety-six hour LC50 values for the goldfish, snail,



and sowbug varied between 90 and 210 ug/1 in static renewal



tests.  Measured bioconcentration factors after varying per-



iods of exposure are as follows: crayfish, 60;  goldfish, 920-



2,300; Scuyemouth bass, 29; and an alga, 160.
                         ///-y

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                     HEXACHLOROBUTADIEN E
I.   INTRODUCTION
     Hexachlorobutadiene (HCBD) is produced in the United
States as a significant by-product in the manufacture of
chlorinated hydrocarbons such as tetrachloroethylene, tri-
chloroethylene, and carbon tetrachloride..  This secondary
production in the U.S. ranges from 7.3 to 14.5 million pounds
per year, with an additional 0.5 million pounds being import-
ed (U.S. EPA, 1975).
     HCBD is used as an organic solvent, the major domestic
users being- chlorine producers.  Other applications include
its use as an intermediate in the production of rubber com-
pounds and lubricants.  HCBD is a colorless liquid with a
faint turpentine-like odor.  Its physical properties include:
boiling point, 210-220°C vapor pressure, 0.15 mm Hg; and
water solubility of .5 ug/1 at 20°C (U.S. EPA, 1979).
     Environmental contamination by HCBD results primarily
during the disposal of wastes containing HCBD from chlori-
nated hydrocarbon industries (U.S. EPA, 1976).  It has been
detected in a limited number of water samples.  HBCD appears
to be rapidly adsorbed to soil and sediment from contaminated
water, and concentrates in sediment from water by a factor of
100 (Leeuwangh, et al., 1975).
II.  EXPOSURE
                                           v
     A.   Water
          HCBD contamination of U.S. finished drinking water
supplies does not appear to be widespread.  The problem is
localized in areas with raw water sources near industrial
                          111-5'

-------
plants discharging HBCD.  From  its physical  and  chemical  pro-
perties, HBCD removal from water by adsorption  into  sediment
should be rapid (Laseter, et al., 1976).  Effluents  from
various industrial plants were  found  to contain  HCBD levels
ranging from 0.04 to 240 ug/1 (Li, et al., 1976).  An EPA
study of the drinking water supply of ten U.S. cities re-
vealed that HCBD was detected in one  of the  water  supplies,
but the concentration was less  than 0.01 ug/1  (U.S.  EPA,
1975).
     B.   Food
          Since the air, soil and water surrounding  certain
chlorohydrocarbon plants have been shown to  be contaminated
with HCBD (Li, et al., 1976), food produced  in the vicinity
of these plants might contain residual levels of HCBD.  A
survey of foodstuffs produced within  25 miles of tetrachloro-
ethylene and trichloroethylene  plants did not detect measur-
able levels of HCBD.  Freshwater fish caught  in  the  lower
Mississippi contained HBCD residues in a range from  0.01  to
1.2 mgAg«  Studies on HCBD contamination of  food  in several
European countries have measured levels as high  as 42 ug/kg
in certain foodstuffs (Kotzias,'et al., 1975).
          The U.S. EPA (1979) has estimated  a HCBD bioconcen-
tration factor of 870 for the edible  portions of fish and
shellfish consumed by Americans.  This estimate  is based  on
measured steady-state bioconcentration studies  in  goldfish.
     C.   Inhalation
          The levels of HCBD detected in the  air surrounding
chlorohydrocarbon plants are generally less  than 5 uc

-------
although values as high as 460 ug/ni  have been measured

(Li, et al.  1976).

III. PHARMACOKINETICS

     A.   Absorption

          Pertinent data were not found on the absorption of

HCBD in the available literature.

     B.   Distribution

          HCBD did not have a strong tendency to accumulate

in fatty tissue when administered orally with other chlori-

nated hydrocarbons.  Some of the chlorinated hydrocarbons

were aromatic compounds and accumulated significantly  in fat

(Jacobs, et al.  1.974) .

     C.   Metabolism

          Pertinent data were not found in the available

literature.

     D.   Excretion                    .-;

          Pertinent data were not found in the available

literature.

IV.  EFFECTS ON MAMMALS

     A.   Carcinogenicity

          Kociba, et al. (1977) administered dietary levels

of HCBD ranging from 0.2 mg/kg/day to 20.0 mg/kg/day for two

years to rats.  In males receiving 20 mg/kg/day, 18 percent

(7/39) had renal tubular neoplasms which were classified as

adenocarcinomas; 7.5 percent (3/40) of the females on  the
                                                          »
high dose developed renal carcinomas.  Metastasis to the lung

was observed in one case each for both male and female rats.
                              *
                          1/1-7

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No carcinomas were observed  in  controls,  however,  a nephro-
blastoma developed in one male  and one  female.
          A significant  increase  in  the  frequency  of lung
tumors was observed  in mice  receiving  intraperitoneal injec-
tions of 4 mgAg or  8 mgAg  of  HCBD, three  times per week un-
til totals of 52 mg  and  96 mg,  respectively,  were  admin-
istered (Theiss, et  al... 1977).
     B.   Mutagenicity
          Taylor (1978)  tested  the mutagehicity of HCBD on _S.
typhimurium TA100.   A dose dependent increase in reversion
rate was noted, but  the  usual criterion  for mutagenicity of
double the background rate was  not reached.
     C.   Teratogenicity
          Poteryaeva (1966)  administered  HCBD to nonpregnant
rats by a single subcutaneous injection of  20 mg/kg.   After
mating, the pregnancy rate for  the dosed  rats was  the same as
that of controls.  The weights  of the young rats from the
dosed mothers were markedly  lower than  the  controls.   Autop-
sies at 2-1/2 months revealed gross pathological changes in
internal organs including g-lomerulonephritis  of the  kidneys.
Degenerative changes were also  observed  in  the red  blood
cells.
     D.   Other Reproductive Effects
          Schwetz, et al. (1977) studied  the  effects  of  di-
etary doses of HCBD  on reproduction,  in  rats.   Males  and  fe-
                                                           »
males were fed dose  levels of 0.2 to 20 mg/kg/day  HCBD start-
ing 90 days prior to mating  and continuing  through  lactation.
At the two highest doses, adult rats suffered  weight  loss,
                                  ,
                             > *lAlS-
                            i J  I "-
                           III-7

-------
decreased food consumption and alterations  of  the  kidney cor-



tex, while the only effect on weanlings consisted  of  a  slight



increase in body weight at 21 days of age at the 20 mg/kg



dose level.  Effect on survival of the young was not  effected.



     E.   Chronic Toxicity



          The kidney appears to be the organ most  sensitive



to HCBD.  Possible chronic effects are observed at doses  as



low as 2 to 3 mg/kg/day (Kociba, et al., 1971, 1977;  Schwetz,



et al., 1977).  Single oral doses as low as 8.4 mg/kg have
                                         \


been observed to have  deleterious effects  on  the  kidney



(Schroit, et al. 1972).  Neurotoxic effects in rats have  been



reported at a dose of 7 mg/kg and effects may  occur at  even



lower dose levels (Poteryaeva, 1973? Murzakaev, 1967).   HCBD



at 0.004 mg/kg gave no indication of neurotoxicity.   Acute



HCBD intoxication affects acid-base equilibrium in blood  and



urine (Popovich, 1975; Poteryaeva, 1971).   Some investigators



report a cumulative effect for HCBD during  chronic dosing by



dermal (Chernokan, 1970) or oral Poteryaeva, 1973) routes.



An increase in urinary coproporphyrin was observed in rats



receiving 2 mg/kg/day and 20 mk/kg/day HCBD for up to 24



months (Kociba, 1977).



     F.   Other Relevant Information



          The possible antagonistic effect of  compounds  con-



taining mercapto (-SH) groups on HCBD have  been suggested by

                                            *•

two studies.  Murzokaev (1967) demonstrated a  reduction  in



free -SH groups in cerebral cortex homogenate  and  blood  serum



following HCBD injection in rats.  Mizyukova,  et al.  (1973)



found thiols (-SH compounds) and amines to  be  effective  anti-

-------
dotes against the toxic effects of HCBD when  administered
prior to or after HCBD exposure.
V.   AQUATIC TOXICITY
     A.   Acute Toxic ity
          Goldfish,  (Carassius auratus), had  an  observed 96-
hour LC50 of 90 ug/1 in a static renewal test  (Leeuwangh,  et
al. 1975).  A snail, (Lymnaea stagnalis), and  a  sowbug,
(Asellus aquaicus), were both exposed for 96-hours  to HCBD
resulting in EC5Q values of 210 and 130 v.g/1,  respective-
ly (Leeuwangh, et al., 1975).  No acute studies  with marine
species have been conducted.
     B.   Chronic Toxic ity
          Pertinent  information was not found  in the avail-
able literature.
     C.   Plant'Effects
          Pertinent data was not found in the  available
literature.
     D.   Residues
          Measured bioconcentration factors are  as  follows:
crayfish, Procambaeus clarhi, 60 times after 10 days expo-
sure; goldfish, Caressius auretus, 920-2,300 times  after 49
days exposure; large mouth bass, Microptorus salmoides, 29
times after 10 days exposure; and a freshwater alga, Oedogon-
ium cardiacum, .160 times after 7 days exposure (Laseter, et
al., 1976).  Residue data on saltwater organisms are not
available.

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VI.  EXISTING GUIDELINES AND STANDARDS



     Neither the human health nor aquatic criteria derived  by



U.S. EPA (1979), which are summarized below,  have gone



through the process of public review; therefore, there  is a



possibility that these criteria may be changed.



     A.   Human



          Standards or guidelines for exposure  to HCBD  are



not available.



          The draft ambient water quality, criteria for  HCBD



have been calculated to reduce the human carcinogenic risk



levels to ID"5, 10-6, and lO"7 (U.S. EPA, 1979).



The corresponding criteria are 0.77 ug/1, 0.077 ug/lf 0.0077



u.g/1, respectively.



     B.   Aquatic



          Draft freshwater or saltwater criterion for hexa-



chlorobutadiene have not been developed because of insuffi-



cient data (U.S. EPA, 1979).
                           ll-Ht

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                              HEXACHLOROBUTAOIENE

                                  REFERENCES
Chernokan,  V.F.   1970.  Some  data of the  toxicology  of hexachlorobutadiene
when ingested into  the organism through the skin. •  Vop.  Gig. Toksikol. Pes-
tits.  Tr.  Nauch.  Tr.  Sess. Akad. med.  Nauk.  SSSR.   (no vol.): 169.  CA:74:
97218r. (Translation)

Jacobs, A.,  et  al.    1974.  Accumulation  of noxious  chlorinated substances
from Rhine River water in the fatty tissue of rats.  Vom. Wasser  43: 259.

Kociba, R.J.,  et al.  1971.   Toxicologic  study of  female  rats administered
hexachlorobutadiene  or hexachlorobenzene  for  30  days.   Dow  Chemical  Co.,
Midland, Mich.
                                                   •.
Kociba, R.J.,  et al.   1977.   Results of  a two-year  chronic toxicity study
with hexachlorobutadiene in rats.  Am. Ind. Hyg. Assoc.  38:  589.

Kotzias, 0., et  al.  1975.  Ecological chemistry.   CIV.  Residue analysis of
hexachlorobutadiene in food and poultry  feed.  Chemosphere  4:  247.

Laseter,  J.L.,   et  al»  1976.   An  ecological  study  of hexachlorobutadiene
(HCSD).  U.S. Environ.  Prot. Agency, EPA-560/6-76-010.

Leeuwangh, P., et al.   1975.   Toxicity of  hexachlorobutadiene in aquatic or-
ganisms.   In:  Sublethal  effects  of  toxic  chemicals  on  aquatic  animals.
Proc.  Swedish-Netherlands  Symp.,  Sept.  2-5.  Elsevier Scientific Publ. Co.,
Inc., New York.

Li,  R.T.,  et  al.   1976.   Sampling  and  analysis  of  selected toxic  sub-
stances.  Task  IB - hexachlorobutadiene.   EPA-560/6-76-015.   U.S.. Environ.
Prot. Agency, Washington, O.C.

Mizyukova, I.G., et al.  1973.  Relation between  the structure and detoxify-
ing  action  of  several  thiols  and  amines during  hexachlorobutadiene poison-
ing.  Fiziol.  Aktive.  Veshchestva.  5: 22.  CA:81:22018M. (Translation)

Murzakaev,  F.G.   1967.   Effect of  small  doses  of  hexachlorobutadiene  on
activity  of the  central nervous  system  and morphological  changes  in  the
organisms of animals  intoxicated  with it.  Gig.  Tr.   Prog.  Zabol.   11: 23.
CA:67:31040a. (Translation)

Popovich,  M.I.   1975.   Acid-base equilibrium and  mineral metabolism follow-
ing  acute  hexachlorobutadiene  poisoning.    Issled.  Abl.  Farm.  Khim.   (no
vol.): 120.  CA:86:26706K.  (Translation)
                                                     ••

Poteryaeva,  G.E.   1966.   Effect of  hexachlorobutadiene on the  offspring  of
albino rats.  Gig Sanit.  31:  33.   ETIC:76:8965. (Translation)
                                                                        »
Poteryaeva, G.E.  1971.  Sanitary  and  toxicological  characteristics  of hexa-
chlorobutadiene.  Vrach. Oelo.   4:  130.  HAPAB:72:820.   (Translation)
                                  III'II

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Poteryaeva, G.E.   1973.   Toxicity  of hexachlorobutadiene during  entry into
the organisms through  the gastorintestinal tract.  Gig. Tr.   9:  98.   CA:85:
29271E. (Translation)

Schroit, I.G.,  et al.   1972.  Kidney  lesions  under experimental  hexachloro-
butadiene  poisoning.   Aktual.  Vop. gig.  Epidemiol.  (no  vol.):  73.   CA:81:
73128E. (Translation)

Schwetz, 8.A.,  et al.   1977.   Results of  a  reproduction study in  rats fed
diets containing hexachlorobutadiene.   Toxicol. Appl. Pharmacol.  42: 387.

Taylor, G.   1978.  Personal communication.  Natl. Inst.  Occup. Safety Health.

Theiss, J.C., et. al.   1977.  Test for carcinogenicity of  organic  contami-
nants of United States drinking  waters by pulmonary tumor response in strain
A mice.  Cancer Res.  37: 2717.

U.S. EPA.   1975.   Preliminary assessment of suspected  carcinogens  in drink-
ing water.   Rep. to Congress.  U.S. Environ. Prot. Agency.

U.S. EPA.  1976.   Sampling  and analysis of selected  toxic substances.   Task
IB  - Hexachlorobutadiene.  EPA-560/6-76-015.   Off.  Tox.  Subst.  U.S.  Envi-
ron. Prot.  Agency, Washington, D.C.

U.S. EPA.  1978.   Contract No. 6803-2624.   U.S.  Environ.  Prot.  Agency,  Wash-
ington, D.C.

U.S.  EPA.   1979.   Hexachlorobutadiene:   Ambient  Water  Quality  Criteria
(Draft).
                                //  /'-

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                                   No. 112
         ichlorocyclohexane


  Health and Environmental Effects
U.S.  ENVIRONMENTAL PROTECTION AGENCY
      WASHINGTON, D.C.   20460

          APRIL 30, 1980
            //a-/

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                          DISCLAIMER
     This report represents a survey of  the  potential health
and environmental hazards from exposure  to the  subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all  available information including all the
adverse health  and  environmental  impacts presented by the
subject chemical.  This  document has undergone scrutiny  to
ensure its technical accuracy.
                          ;//-*•

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                       SPECIAL NOTATION
U.S. EPA13 Carcinogen. Assessment Group (GAG) has evaluated



hexachlorocyclohexane and has found sufficient evidence to



indicate that this compound is carcinogenic.

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                             HEXACHLOROCYCLOHEXANE
                                    Summary

     Hexachlorocyclohexane (HCH),  a broad spectrum insecticide, is a mixture
of  five configurational  isomers.   HCH  is no  longer  used  in  the  United
States; however,  its gamma-isomer,  commonly known as  lindane, continues to
have  significant  commercial  use.   Technical  HCH,  alpha-HCH,  _beta-HCH,  and
lindane  (gamma-HCH)  have  all  been shown  to  induce  liver tumors  in mice.
Most of the  studies  on hexachlorocyclohexanes deal only with  lindane.  Evi-
dence for mutagenicity of  lindane  is equivocal.   Lindane was not  teratogenic
for  rats,  although it reduced  reproductive capacity in rats   in  a  study of
four generations.   Chronic exposure of  animals  to lindane  caused liver en-
largement  and,  at  higher  doses,  some liver  damage  and  nephritic  changes.
Humans chronically exposed to HCH  suffered  liver damage. Chronic  exposure of
humans  to  lindane: produced  irritation of  the central  nervous system.   HCH
and  lindane  are  convulsants.  The U.S. EPA (1979)  has estimated  the ambient
water  concentrations  of hexachlorocyclohexanes  corresponding  to  a  lifetime
cancer  risk  for humans of  10   as  follows:   21 ng/1  for  technical  HCH, 16
ng/1 for alpha-HCH, 28 ng/1  for beta-HCH, and  54 ng/1 for lindane  (gammaHCH).
     Lindane has  been  studied in a  fairly  extensive  series of acute studies
for both freshwater and marine  organisms.   Acute toxic levels as  low as 0.17
ng/1 have been reported for marine invertebrate species.

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                             HEXACHLOROCYCLOHEXANE
I.   INTRODUCTION
     This profile  is based  on the  Ambient  Water Quality  Criteria Document
for   Hexachlorocyclohexane   (U.S.    EPA,   1979).    1,2,3,4,5,6-Hexachloro-
cyclohexane   (CJ-LClj   molecular  weight  290.0)   is   a  brownish-to-white
crystalline  solid with  a  melting point  of 65°C  and a solubility  in water
of 10  to  32 mg/1.   It  is  a mixture  of five configurational  isomers and is
commonly referred  to as BHC or benzene hexachloride.   Lindane is the common
name  for  the  gamma isomer  of 1,2,3,4,5,6-hexachlorocyclohexane  (U.S.  EPA,
1979).
     Technical •  grade   hexachlorobenzene   (HCH)   contains  the  hexachloro-
cyclohexane  isomers in  the  following ranges:   alpha-isomer,  55 to  70 per-
cent;  beta-isomer,  6 to  8 percent;  gamma-isomer ,  10 to  18 percent; delta-
isomer, 3 to 4 percent; epsilon-isomer, trace  amounts.   Technical  grade HCH
may  also  contain 3  to  5 percent of  other  chlorinated  derivatives  of cyclo-
hexane,  primarily  heptachlorocyclohexane  and  octachlorocyclohexane  (U.S.
EPA,  1979).
     Hexachlorocyclohexane  (HCH)  is   a  broad  spectrum  insecticide  of  the
group of cyclic  chlorinated  hydrocarbons  called organochlorine insecticides.
Since  the  gamma-isomer  (lindane)  has  been shown  to be  the  insecticidally
active  ingredient  in   technical  grade  HCH,  technical  grade HCH   has  had
limited commercial use  except as  the raw material  for  production  of lin-
dane.  Use  of technical HCH has been banned in  the U.S., but  significant
commercial  use of  lindane  continues.  Lindane  is used in a  wide  range  of
applications including  treatment  of  animals,  buildings,  man  (for  ectopara-
sites), clothes, water  (for  mosquitoes), plants,  seeds,  and soils (U.S. 'EPA,
1979).

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     NO  technical grade  HCH  or  lindane  is currently  manufactured  in the
U.S.; all lindane used in the U.S. is imported (U.S. EPA, 1979).
     Lindane  has  a  low  residence  time in  the  aquatic environment.   It is
removed  by  sedimentation,  metabolism,  and volatilization.   Lindane contri-
butes less to  aquatic  pollution than the other hexachlorocyclohexane isomers
(Henderson, et al. 1971).
     Lindane  is  slowly  degraded by soil  microorganisms  (Mathur  and  Saha,
1975; Tu,  1975,  1-976)  and is  reported  to  be isomerized to  the alpha and/or
delta isomers  in  microorganisms and plants  (U.S.  EPA,  1979), though this is
controversial  (Tu,  1975, 1976;  Copeland and Chadwick,  1979;  Engst,  et al.
1977).   It  is-not isomerized  in  adipose tissues  of rats,  however  (Copeland
and Chadwick, 1979).
II.  EXPOSURE
     A.  Water
         The  contamination  of water  has  occurred  principally  from direct
application of technical hexachlorocyclohexane  (HCH) or lindane to water for
control of mosquitoes, from  the use of HCH in agriculture and  forestry, and,
to a  lesser  extent,  from  occasional contamination  of  wastewater from manu-
                                                 *
facturing plants  (U.S.  EPA, 1979).
         In the  finished- water of.  Streator,  Illinois,  lindane  has  been de-
tected at a concentration of 4 pg/1  (U.S. EPA, 1975).
     B.  Food
         The daily  intake  of  lindane  has  been reported to  be  1 to 5 ug/kg
body weight and the daily intake of  all other HCH  isoraers to  be 1 to 3 ug/kg
body weight (Duggan and Ouggan, 1973).   The  chief  sources  of HCH residues in
the human  diet are milk, eggs,  and other dairy  products  (U.S.  EPA,  1979),
and carrots  and  potatoes (Lichtenstein, 1959).   Seafood is  usually a minor

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source of  HCH,  probably because  of the relatively  high rate of dissipation
of HCH in the aquatic environment (U.S. EPA, 1979).
         The  U.S.   EPA  (1979)  has  estimated   the  weighted  average biocon-
centration factor  for lindane  to  be 780 for the  edible  portions of fish and
shellfish consumed  by Americans.   This estimate is based on measured steady-
state bioconcentration in bluegills.
     C.  Inhalation
         Traces of  HCH have  been  detected  in the air of central and suburban
London (U.S. EPA,  1979).  No further pertinent information could be found  in
the available literature.
     0.  Dermal
         Lindane has been  used to eradicate human  ectoparasites and few ad-
verse reactions have been reported (U.S. EPA, 1979).
III. PHARMACOKINETICS
     A.  Absorption
         The  rapidity  of  lindane absorption  is  enhanced  by  lipid mediated
carriers.  Compared to  other  organochlorine  insecticides, HCH  and lindane
are unusually  soluble  in  water,  which  contributes to  rapid  absorption and
excretion  (Herbst  and Bodenstein,  1972;  U.S.  EPA,  1979).  Intraperitoneal
injection  of lindane resulted in  35 percent  absorption   (Koransky,  et al.
1963).  Lindane is absorbed after oral.and dermal exposure  (U.S. EPA, 1979).
     8.  Distribution
         After administration  to  experimental  animals,  lindane  was detected
in  the  brain at  higher concentrations  than  in  other  organs  (Laug,  1948;
                                                     ^
Davidow and  Frawley,  1951;  Koransky,  et  al.  1963; Huntingdon  Res.  Center,

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1972).  At  least 75 percent  of an intraperitonial  dose of 14C-labeled lin-
dane was consistently found in  the skin,  muscle,  and fatty tissue (Koransky,
et al.  1963).   Lindane enters  the human fetus through the placenta; higher
concentrations were  found  in  the skin than  in the  brain  and  never exceeded
the  corresponding   values  for  adult  organs  (Poradovsky,   et  al.  1977;
Nishimura, et al. 1977).
     C.  Metabolism
         Lindane  is metabolized  to  gamma-3,4,5,6-tetrachlorocyclohexene  in
rat  adipose tissue,  but  is  not  isomerized  (Copeiand  and Chadwick,  1979);
other  metabolites are  2,3,4,5,6-pentachloro-2-cyclohexene-l-ol,  two  tetra-
chlorophenols, and  three  trichlorophenols (Chadwick, et al. 1975;  Engst,  et
al.  1977).   These are. commonly  found  in the urine as conjugates (Chadwick
and Freal,  1972).  Lindane metabolic  pathways are still matters of some con-
troversy  (Engst, et al.   1977;  Copeiand  and Chadwick,  1979).   Hexachloro-
cyclohexane  isomers  other than lindane  are  metabolized to trichlorophenols
and  mercapturic  acid conjugates  (Kurihara,  1979).   Both free  and conjugated
chlorophenols  are far  less  toxic than  the  parent compounds  (Natl.  Acad.
Sci., 1977).                                      ;
     D.  Excretion
         HCH and  lindane  appear to be eliminated  primarily as conjugates  in
the  urine.   Elimination of lindane appears  to be  rapid after  administration
ceases.  Elimination of beta-HCH is  much slower  (U.S. EPA,   1979).   In  fe-
males, HCH is excreted in the milk  as well as in  the urine.   The beta-isomer
usually  accounts for  above  90 percent  of   the  HCH -present  in human  milk
(Herbst and Bodenstein, 1972).

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IV.  EFFECTS
     A.  Carcinogenicity
         An increased  incidence  of liver tumors was  reported  in male and/or
female mice of various  strains  fed technical hexachlorocyclohexane (Goto, et
al. 1972;  Hanada,  et al.  1973;  Nagasaki, et al.  1972),  alpha-HCH (Goto, et
al. 1972;  Hanada,  et al. 1973;  Ito,  et  al.  1973,  1975),  beta-HCH (Goto, et
al.  1972;  Thorpe  and Walker,  1973)  and  lindane  (gamma-HCH)  (Goto,  et al.
1972:  Hanada,  et  al.  1973;  Natl.  Cancer Inst.,  1977a;  Thorpe  and Walker,
1973).   Male  rats  fed alpha-HCH  also  developed  liver tumors  (Ito,  et al.
1975).   A  mixture containing  68.7 percent  alpha-HCH,  6.5  percent beta-HCH
and 13.5 percent lindane  in  addition to other  impurities (hepta- and octa-
chlorocyclohexanes),  administered  orally (100 ppm  in the diet,  or 10 mg/kg
body  weight  by  intubation),  caused tumors  in  liver and  in lymph-reticular
tissues  in male  and  female mice after 45 weeks.   Application  by skin paint-
ing  had  no  effect   (Kashyap,  et  al.   1979).   A  review  by   Reuber  (1979)
suggests that lindane is carcinogenic on uncertain evidence.
     B.  Mutagenicity
         Evidence  for  the  mutagenicity of lindane  is equivocal.  Some alter-
ations in  mitotic  activity and  the karyotype of  human lymphocytes cultured
with  lindane at  0.1 to  10 ug/ml have been  reported  (Tsoneva-Maneva,  et al.
1971).   Lindane  was  not  mutagenic in  a dominant-lethal  assay  (U.S.  EPA,
1973) or a host-mediated assay (Buselmair, et al. 1973).
     Gamma-HCH  was   found  to   be  mutagenic  in  microbial   assays   using
Salmonella typhimurium with metabolic  activation,  the host-mediated  assay,
and the  dominant lethal test in rats.   Other  reports indicate that it  does
not have significant mutagenic activity (U.S. EPA,  1979).
                                 II-9

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     C.  Teratogenicity
         Lindane given  in the  diet during pregnancy "at levels of  12 or 25
mg/kg  body   weight/day  did  not   produce   teratogenic  effects   in  rats
(Mametkuliev, 1978; Khera, et al. 1979).
     0.  Other Reproductive Effects
         Chronic lindane  feeding in a study of  four  generations of rats in-
creased  the  average duration of pregnancy,  decreased the number  of  births,
increased  the proportion  of stillbirths, and  delayed  sexual  maturation in
F7  and  F,  females.    In  addition,  some of  the  F.  and   F_ animals  ex-
hibited  spastic paraplegia (Petrescu, et al. 1974).
         In  rats and  rabbits, lindane given  in  the diet during pregnancy in-
creased  postimplanation death of embryos  (Mametkuliev,  1978;  Palmer,  et al.
1978).   Testicular atrophy  has  been observed  for lindane in  rats  and mice
(National Cancer Institute, 1977b; Nigam,  et al. 1979).
     E.  .Chronic Toxicity
         Irritation of  the central  nervous system, with other toxic side ef-
fects  (nausea,  vomiting,  spasms,  weak respiration with cyanosis  and blood
dyscrasia),  was  reported  after  prolonged  or improper  use  of Hexicid  (1 per-
cent  lindane) for  the  treatment of  scabies  on humans  (Lee, et  al.  1976).
Production  workers exposed  to  technical HCH  exhibited  symptoms  including
headache,  vertigo,  irritation  of the  skin,  eyes,  and  respiratory  tract mu-
cosa.  In  some  instances,  there were apparent disturbances  of carbohydrate
and  lipid  metabolism  and dysfunction  of the  hypothalamo-pituitary-adrenal
system (Kazahevich, 1974;  Besuglyi,  et al. 1973).  A' study  of persons occu-
pationally exposed  to HCH for 11 to 23 years  revealed  biochemical manifes-
                                                                        »
tations  of toxic hepatitis (Sasinovich, et al. 1974).

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         In  chronic  studies with  rats  given  lindane  in  oil,  liver  cell;
hypertrophy,(fat degeneration and necrosis)  and nephritic changes were  noted
at higher  doses (Fitzhugh,  et  al. 1950; Lehman,  1952).  Rats  inhaling lin-
dane  (0.78  mg/m-3)  for seven  hours,   five  days a  week  for  180 days  showed
liver cell enlargement,  but showed no toxic symptoms or other  abnormalities
(Heyroth,  1952).   The addition  of 10 ppm lindane to  the diet  of rats for one
or  two  years   decreased  body  weight  after  five  months  of  treatment  and
altered  ascorbic  acid levels in  urine,  blood, and tissues (Petrescu,  et al.
1974).   Dogs  chronically  exposed  to  lindane  in  the  diet  had   slightly
enlarged livers (Rivett, et  al. 1978).
     F.  Other-Relevant Information
         Hexachlorocyclohexane  is a convulsant.
         Lindane is  the most acutely  toxic isomer of HCH.  The  toxic effects
of lindane are antagonized  by  pretreatment  with phenobarbital  (Litterst  and
Miller,  1975)  and by treatment  with  silymarin (Szpunar,  et al.  1976)  and
various tranquilizers  (Ulmann,  1972).
V.   AQUATIC TOXICITY
     A.  Acute Toxicity
         Among  16 species  of  freshwater  fish, LC_Q values   from  one  flow-
through  and  24  static  bioassays  for  the  gamma  isomer   of hexachloro-
cyclohexane  ranged from 2 jug/1  for the- brown trout (Salmo trutta) (Macek  and
McAllister,  1970)   to  152  jjg/1  for  the  goldfish   (Carassius   auratus)
(Henderson,  et  al.  1959).    In  general,, the salmon tended  to  be more sensi-
tive  to the action  of lindane  than did  warm  water  species.   Zebra fish
(Brachydanio rerio)  showed  a  lindane  LC5Q  value  of 120 ng/1,  but rainbow
trout (Salmo qairdnerl)  evidenced respiratory  distress at 40 ng/1   (Slooff,
1979).   Technical  grade HCH was much  less toxic than pure  lindane;  LC50

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values obtained  for  lindane  in 96-hour  studies of  the freshwater  goldfish
(Carassius  auratus)  ranged from 152 jjg/1  for  100 percent  lindane to  8,200
jjg/1  for  BCH (15.5  percent  gamma  isomer)  (Henderson,  et al.  1959).   Static
tests  on  freshwater  invertebrates revealed  a  range of LC5Q values of from
4.5  jug/1   (96-hour  test)   (Sanders   and   Cope,   1968)  for   the  stonefly
(Pteronarcys  californica)  to • 880 jug/1  (48-hour  test)  (Sanders  and  Cope,
1968)  for the clado- ceran  (Simocephalus  serralatus)  for  lindane.   Canton
and  Slooff  (1977)   re-  ported  an LC5(,  value  for  the  pond snail  (Lymnaea
staqnalis)  of l,200;jg/l  for  alpha-HCH in a 48-hour  static test.
         Among  seven species of marine fish tested for  the  acute  effects of
lindane,  static  test  LC5Q   values  ranged  from 9.0 ^g/1  for the  Atlantic
silversides (Menidia  menidia)  to  66.0 ug/1  for  the  striped  mullet  (Mugil
cephalus)  (Eisler,   1970).   The results  of six flow-through assays on five
species  of marine  fish produced  LC5Q values from  7.3 jjg/l for the  striped
bass  (Morone saxatilis)  (Korn  and Earnest,  1974)  to  240 jug/1 for the long  .
nose  killifish  (Fundulus similis)  (Butler,  1963).    A single species,  the
pinfish  (Laqodon rhomboides),  tested  with  technical  grade  hexachlorocyclo-
hexane, produced a  96-hour  flow-through LC_Q value  of  86.4 jjg/1  (Schimmel,
et al. 1977).   Acute tests on  marine  invertebrates  showed six  species  to be
quite  sensitive  to  lindane,  with  LC5Q values from  both  static  and  flow-
through assays ranging from  0.17 jug/1-for the pink shrimp  (Panaeus duorarum)
(Schimmel,  et al.   1977)  to 10.0 /jg/1  for the grass  shrimp  (Palaemonetas
vulqaris)  (U.S.  EPA,  1979).    An  LC5Q value of 0.34  jug/1  was obtained  for
technical grade  hexachlorocyclohexane  for the pink  shrimp (Schimmel, et  al.
1977).  The American  oyster  had an EC5Q  of 450 jjg/1  based on shell 'decom-
position (Butler, 1963).

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     8.  Chronic
         A chronic  value of  14.6 jjg/1 for  lindane  was obtained in  a  life-
cycle  assay  of  the  freshwater  fathead  minnow  (Pimephales promelas).   For
three  species  of  freshwater  invertebrates  tested  with  lindane,  chronic
values  of 3.3,  6.1,  and 14.5 pg/l  were  obtained  for Chironomus  tentans,
Gammarus  fasciatus,  and  Daphnia magna  (Macek,  et  al.  1976)..   No- chronic
marine data for any of the hexachlorobenzenes were available.
     C.  Plant Effects
         Concentrations  causing growth  inhibition of  the  freshwater  alga,
Scenedesmus acutus were reported to be 500,  1,000, 1,000,  and  5,000 jug/1 for
alpha-HCH,   technical   grade   HCH,   lindane,   and   beta-HCH,   respectively
(Krishnakumari,  1977).   In  marine  phytoplankton  communities,  an  effective
concentration  value  of 1,000 fig/1 (resulting  in  decreased productivity) was
reported  for  lindane;  and for the alga, Acetabularia mediterranea  an effec-
tive  concentration  of  10,000 jug/1 was  obtained for lindane-induced  growth
inhibition.  No effect  in 48 hours  was observed  for  the  algae  Chlamydomonas
so. exposed  to lindane  at the maximum solubility  limit.   Irreparable damage
to Chlorella  sp.  occurred at  lindane concentrations of  more  than  300  ;jg/l
(Hansen, 1979).
     0.  Residues
         Bioconcentration factors  for- lindane  ranging  from 35  to  938  were
reported  for  six species of  freshwater  organisms (U.S. EPA, 1979;  Sugiura,
et  al. 1979a). . In marine  organisms,  bioconcentration  factors  (after  28
days)  for  39 percent  lindane of 130,  218,  and  617..were  obtained for  the
edible  portion of  the  pinfish (Lagodon  rhomboides),  the American oyster
                                  111-13

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(Crassostrea virqinica),  and offal tissue  of the pinfish  (Schimmel,  et al.
1977). Sugiura, et al.  (1979a)  found alpha-, beta-, and  gamma-HCH had accu-
mulation  factors  of  1,216,  973  and   765  in  golden  orfe  (Leuciscusidus
melanotus); 330,  273,  and 281 in carp  (Cyprinus  carpio); 605, 658,  and 442
in  brown  trout   (Salmo  trutta  fario);  and  588,  1,485,  and  938 in  guppy
(Poecila reticula),  respectively.   Further, these accumulation  factors were
proportional to the  lipid content  of the fish.   Accumulation occurred in the
adipose tissues and  the gall bladder, with  the  alpha and  beta-HCH being more
persistent (Sugiura, et al.  1979b).
         Equilibrium  accumulation  factors  of  429  to  602  were  observed  at
days 2 to  6 after exposure of Chlorella sp. to 10 to 400 ;jg/l of lindane in
aqueous solution  (Hansen,  1979).
VI.  EXISTING STANDARDS AND  GUIDELINES
     Neither the  human health nor  the  aquatic criteria  derived  by U.S. EPA
(1979), which are summarized below, 'have gone through  the  process of public
review;  therefore,  there  is a  possibility  that   these criteria will  be
changed.
     A.  Human
                                                 i
         Based on the  induction  of liver tumors in  male  mice, and using the
"one-hit" model,  the U.S. EPA (1979) has  estimated  the  following  levels  of
technical  hexachlorocyclohexane  and its isomers in ambient  water which will
result in specified risk  levels of human cancer.
         The water  concentrations of technical HCH corresponding to  a  life-
        icer risk   for  I"
Nagasaki,  et al. (1972).
time  cancer risk  for  humans of  10~   is  21  ng/1, "'based  on  the  data  of
                                  ^^^^.^u.
                                  1 / JeL J

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         The water  concentrations of  alpha-HCH corresponding  to a lifetime
cancer  risk  for humans  of 10    is  16 ng/1,  based on  the data  of Ito, et
al. (1975).
         The water  concentrations of beta-HCH  corresponding  to  a lifetime
cancer  risk  for humans  of 10    is  28 ng/1, based  on the data  of Goto, et
al. (1972).
         The water  concentrations of  lindane  (gamma-HCH)  corresponding to a
lifetime cancer risk  for humans of  10    is 54 ng/1, based on  the data of
Thorpe and Walker (1973).
         Data  for  the  delta  and  epsilon isomers  are insufficient  for the
estimation of cancer risk levels  (U.S. EPA, 1979).
         An ADI of  1 jjg/kg  for HCH has been set by the Food and Agricultural
Organization and the World Health Organization (U.S. EPA, 1979).
         Tolerance  levels set by  the  EPA are as follows:  7 ppm for animal
fat, 0.3 ppm  for milk,  1 ppm for most  fruits  and vegetables, 0.004  pm for
finished drinking water, and 0.5 ug/m3 (skin) for air  (U.S. EPA, 1979).
     8.  Aquatic
         For  lindane,  freshwater criteria  have been  drafted  as  0.21 ug/1
with 24-hour average concentration not to  exceed 2.9 ug/1.  For marine or-
ganisms, criteria  for  lindane have  not  been drafted.  No  criteria for mix-
tures of isomers of hexachlorocyclohexane (benzene  hexachloride)  were draft-
ed for freshwater or marine organisms because of the lack of data.

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                    HEXACHLOROCYCLOHEXANE

                          REFERENCES

Besuglyi, V.P.,  et al.   1973.   State of  health  of persons
having prolonged  occupational contact  with hexachlorocyclo-
hexane.  Idrabookhr Beloruss.  19: 49.

Buselmair,  W.,   et al.    1973.    Comparative  investigation
on the mutagenicity of  pesticides in mammalian test systems.
Mutat. Res.  21: 25.

Butler,  P.A.    1963.    Commercial  fisheries investigations,
pesticide-wildlife  studies,  a  review  of  fish and  wildlife
service  investigations  during 1961-1962.   U.S.  Dept.  Inter.
Fish Wildl. Circ. 167: 11.
                                          ».
Canton, J.H., and W. Sloof.  1977.  The usefulness of Lymnaea
stagnalis  L.  as  a   biological   indicator   in  toxicological
bioassays (model substance cA-HCH).  Water Res. 11: 117.

Chadwick, R.W.,  and J.J.  Freal.   1972.   The identification
of  five  unreported  lindane  metabolites  recovered  from rat
urine.  Bull. Environ. Contam. Toxicol.  7:  137.

Chadwick, R.W.,  et al.   1975.  Dehydrogenation, a previously
unreported pathway of lindane metabolism in  mammals.  Pestic.
Biochem.  Physiol.  6:  575.

Copeland, M.F.,  and  R.W. Chadwick.   1979.   Bioisomerization
of  lindane  in  rats.     Jour.  Environ.  Pathol.  Toxicol.  2:
737.

Davidow,  B.  and J.P.  Frawley.    1951.    Tissue distribution
accumulation and elimination  of  the  isomers of benzene  hexa-
chloride (18631).  Proc. Soc. Exp. Biol.  Med.  76: 780.

Duggan,  R.E.,  and M.B.  Duggan.    1973.   Residues  of  pesti-
cides  in milk,  meat 'and foods.   Page  334  In:   L.A.  Edwards,
ed.  Environ. Pollut.  Pestic. London.

Eisler,  R.    1970.    Acute  toxicities  of  organochlorine and
organophosphorus  insecticides  to  estuarine  fishes.    Bur.
Sport Fish Wildl. Pap. No. 46.

Engst, R., et al.  1977.  Recent  state of lindane metabolism.
Residue Rev. 68: 59.

Fitzhugh, O.G.,  et al.   1950.  Chronic toxicities of benzene
hexachloride, and  its alpha,  beta,  and gamma isomers.    Jour.
Pharmacol. Exp. Therap.  100: 59.

Goto,  M.,  et al.   1972.   Ecological  chemistry.   Toxizitat
von a-KCH in mausen. Chemosphere  1: 153.

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Hanada,  M.,  et  al.    1973.    Induction  of  hepatoma  in mice
by benzene hexachloride.  Gann. 64: 511.

Hansen,  P.D.    1979.    Experiments  on  the  accumulation  of
lindane  (gamma BHC)  by  the  primary producers Chlorella spec.
and Chlorella pyrenoidosa.   Arch.  Environ.  Contam. Toxicol.
8: 721.

Henderson, C.,  et al.  1959.   Relative  toxicity  of  ten chlori-
nated  hydrocarbon- insecticides  to  four   species of  fish.
Trans. Am. Fish Soc. 88: 23.

Henderson, C., et  al.   1971.  Organochlorine pesticide resi-
dues in fish-fall 1969: Natl. Pestic. Monitor.  Progr. Pestic.
Monitor. Jour.   5: A.

Herbst,  M.,  and G.  Bodenstein.   1972.   tToxicology  of lin-
dane.  Page 23 In; E. Ulmann,  (ed.) Lindane. Verlag K. Schil-
linger Publishers, Freiburg.

Heyroth,  F.F.   1952.   In;  Leland, S.J.,  Chem. Spec. Manuf.
Ass. Proc. 6:110.

Huntingdon Research  Center.   1972.    In;  Lindane: Monograph
of  an  insecticide E.  Illmon  (ed.).    Lube  Verlag K.  Schil-
linger p. 97.

Itp, N.,  et  al.    1973.  Histologic' and  ultrastructur.al stu-
dies  on  the  hepatocarcinogenicity  of  benzene hexachloride
in mice.  Jour. Natl. Cancer Inst.  51: 817.

Ito, N. ,  et  al.    1975.   Development  of  hepatocellular car-
cinomas  in  rats  treated with  benzene hexachloride.   Jour.
Natl.  Cancer Inst.  54: 801.

Kashyap,  S.K., et  al.   1979.  Carcinogenicity of  hexachloro-
cyclohexane (BHC)  in pure inbred  Swiss mice.  Jour. Environ.
Sci. Health B14: 305.

Kazahevich,  R.L.    1974.   'State  of  the  nervous  system  in
persons with a prolonged professional contact with hexachlor-
ocyclohexane and  products  of  its synthesis.    Vrach.  Delo.
2: 129.

Khera, K.S., et al.   1979.   Teratogenicity studies on pesti-
cidal  formulations  of  dimethoate,  diuron and   lindane  in
rats.  Bull. Environ. Contam. Toxicol.  22:  5.22.

Kocansky, w.,  et al.   1963.   Absorption,  distribution,  and
elimination of alpha-  and  beta- benzene  hexachloride.  Arch.
Exp. Pathol. Pharmacol.  244: 564.

Korn,  S.,  and  R.  Earnest.   1974.   Acute  toxicity of twenty
insecticides  to  striped bass, Marone  saxatilis.    Calif.
Fish Game 60:  128.
                           -/ ^ ~i /^
                            I ^) *.u

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Krishnakumari, M.K.   1977.   Sensitivity of  the alga Scene-
desmus acutus to some pesticides.  Life Sci.  20: 1525.

Kurihara,  H.,  et  al.    1979.    Mercapturic acid  formation
from lindane in rats.  Pest. Biochem. Physiol.   10: 137.

Laug, E.P.  1948.   Tissue distribution of a toxicant follow-
ing oral  ingestion of the  gamma-isomer  of  benzene hexachlo-
ride by rats.   Jour. Pharmacol. Exp. Therap.  93: 277.

Lee, B., et al.   1976.   Suspected reactions to gamma benzene
hexachloride.   Jour. Am. Med. Assoc.  236: 2846:

Lehman, A.J.   1952a.   Chemicals in  food:   A  report  to the
Assoc.  of  Food  and  Drug  officials.    Assoc.  Food  and  Drug
Off., U.S. Quart. Bull.  16:  85.
                                          ".
Lehman, A.J.   1952b.     Chemicals  in foods:   A  report  to
the Association of Food  and Drug officials on current develop-
ments.   Part  II.   Pesticides Section  V.    Pathology.   U.S.
Assoc. Food Drug Off., Quart.  Bull.  16: 126.

Lichtenstein,  E.P..   1959.    Absorption of  some chlorinated
hydrocarbon  insecticides   from  soils  into  various  crops.
Jour. Agric. Food Chem. 7:  430.

Litterst, C.L.,  and  E.  Miller.   1975.   Distribution of lin-
dane  in brains of control  and phenobarbital pretreated dogs
at the  onset of  lindane induced  convulsions.  Bull. Environ.
Contam. Toxicol.   13: 619.

Macek,  K.J.,  and W.A. McAllister.   1970.   Insecticide  sus-
ceptibility  of  some  common  fish  family  representatives.
Trans. Am. Fish Soc. 99: 20.

Macek,  K.J.,  et  al.   1976.   Chronic  toxicity of  lindane
to  selected  aquatic  invertebrates  and  fishes.   EPA-600/3-
76-046.  U.S.  Environ. Prot. Agency.

Mametkuliev, C.H.   1978.   Study of  embryotoxic and terato-
genic properties  of  the gamma  isomer of HCH  in experiments
with rats.  Zdravookhr.  Turkm.  20: 28.

Mathur, S.P.,  and J.G. Saha.   1975.   Microbial degradation
of  lindane-C-14  in  a flooded sandy  loam  soil.    Soil  sci.
120: 301.
                                            .-
Nagasaki,  H. ,  et  al.   1972.    Carcinogenicity of  benzene
hexachloride (BHC).  Top. Chem.  Carcinog.,  Proc. Int.  Symp.,
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National  Academy of  Sciences -  National  Research  Council.
1977.   Safe Drinking  Water Committee.   Drinking  Water  and
Health,  p. 939.

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National  Cancer  Institute.   1977a.   A bioassay  for  possible
carcinogenicity of  lindane.   Fed.  Reg^  Vol.  42  No.  218.

National  Cancer  Institute.    1977b.    Bioassay  of  lindane
for  possible carcinogenicity.   NCI Carcinogenesis  Technical
Report, Series No.  14.

Nigam,  S.K.,  et al.   1979.   Effect of hexachlorocyclohexane
feeding  on  testicular  tissue  on pure  inbred  Swiss  mice.
Bull. Environ. Contain. Toxicol.  23: 431.

Nishiraura,  H.,  et  al.    1977.   Levels  of polychlorinated
biphenyls  and organochlorine insecticides  in  human  embryos
and  fetuses.  Pediatrician 5:  45.

Palmer, A.K.,  et al.  1978.   Effect  of lindane  on  pregnancy
in the  rabbit and rat.   Toxicology  9: 239.

Petrescu, S.,  et  al.   1974.  Studies on the effects  of  long-
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78:  831.

PoradovsJcy,  R.,  et al.    1977.   Transplacental  permeation
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405.

Reuber, M.D.   1979.   Carcinogenicity  of  lindane.    Environ.
Res.  19:  460.

Rivett,  K.P.,  et  al.    1978.    Effects  of  feeding  lindane
to dogs for periods of up  to  2  years.   Toxicology 9:  237.

Sanders,  H.O., and  O.B.  Cope.   1963.  The relative  toxicities
of serveral  pesticides to naiads  of  three  species of stone-
.flies.  Limnol. Oceanogr.  13:  112.

Sasinovich,  L.M.,  et  al.    1974.   Toxic  hepatitis  due  to
prolonged exposure  to  BHC.  Vrach. Delo.  10: 133.

Schimmel, S.E., et  al.   1977.  Toxicity and bioconcentration
of  BHC and  lindane  in  selected  estuarine animals.    Arch.
Environ.  Contain. Toxicol.   5:  355.

Sloof,  W.   1979.    Detection  limits  of a biological  monitor-
ing  system based on fish respiration.   Bull. Environ.  Contam.
Toxicol.  23:  517.

Sugiura,  K.,  et al.   1979a.   Accumulation of  organochlorine
compounds  in  fishes.    Difference  of accumulation  factors
of fishes.  Chemosphere  5:  359,

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Sugiura, K.,  et  al.   19795.   Accumulation of organochlorine
compounds  in  fishes.   Distribution  of 2,4,5-T, 
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                                    No.  113
    ganma-flexachlorocyclohexane


  Health and Environmental Effects
U.S.  ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.   20460

          APRIL 30,  1980
            1/3-1

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not  reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.
                      7

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                          Disclaimer Notice
Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.

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                     GAMMA-HEXACHLOROCYCLOHEXANE (Lindane)
                                    Summary

     Gamma-l,2,3,4,5,6-hexachlorocyclohexane, commonly  known as lindane, can
induce liver tumors  in  mice.   Evidence for mutagenicity  of lindane is equi-
vocal.  Lindane  was  not teratogenic for rats,  although it reduced reproduc-
tive capacity over four generations.   Chronic exposure of animals to lindane
caused liver enlargement and,  at  higher doses,  some liver damage and nephri-
tic  changes.   Humans   chronically  exposed  to  HCH  suffered  liver  damage.
Chronic exposure of humans to  lindane produced  irritation of  the  central
nervous system.  Lindane is a convulsant.
     Lindane  has been  extensively studied  in  a  number  of  freshwater  and
marine acute studies.  Levels as  low as 0.17 jjg/1 are toxic to marine inver-
tebrate species.

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                     GAMMA-HEXACHLOROCYCLOHEXANE (Lindane)
I.   INTRODUCTION
     This  profile is based  on the  Ambient Water  Quality Criteria  Document
for Hexachlorocyclohexane  (U.S. EPA, 1979).
     Gamma-l,2,3,4,5,6-hexachlorocyclohexane     or     lindane     (C^H^Cl^;
molecular  weight  290.0)  is  a crystalline  solid  with  a melting  point  of
112.8°C,  a  vapor pressure  of  0.003  mm Hg  at  20°C  (U.S.   EPA,  1979),  a
solubility  in water  at 25°C  of  7.8 rag/1  (Hansen, 1979),  and a  solubility
in ether of 20.8  g/100 g at  20°C  (U.S.  EPA,  1979).   Other trade names  in-
clude  Jacutin,  Lindfor 90,  Lindamul 20, Nexit-Staub, Prodactin,  gamma-HCH,
gamma-SHC,  and purified BHC  (U.S. EPA,  1979).  Technical  grade hexachlorocy-
clohexane contains 10 to 18 percent lindane.
     Lindane  is  a broad spectrum insecticide,  and  is a member  of  the  cyclic
                                                                             *
organo-chlorinated hydrocarbons.  It  is used  in a wide range of applications
including  treatment  of animals,  buildings, man  (for ectoparasites),  cloth-
ing, water  (for  mosquitoes),  plants, seeds,  and soil.   Lindane is not cur-
rently  manufactured  in the  U.S.; all  lindane  used in  the  U.S. is  imported
(U.S. EPA, 1979).
     Lindane has a low  residence  time in the  aquatic  environment.  It  is  re-
moved by sedimentation, metabolism, and volatilization.  Lindane contributes
less to  aquatic  pollution  than the  other hexachlorocyclohexane  isomers (Hen-
derson, et al. 1971).
     Lindane  is  slowly  degraded by  soil  microorganisms (Mathur  and  Saha,
1975; Tu, 1975,  1976) and  is  reported to be  isomerized to the  alpha- and/or
delta- isomers in  microorganisms and  plants  (U.S.  EPA,  1979), but not  in
rats (Copeland and Chadwick, 1979).   The  metabolic  pathway in microorganisms  '
is still controversial (Tu, 1975,  1976;  Copeland and Chadwick, 1979).

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II.  EXPOSURE
     A.  Water
         The  contamination of  water  has  occurred  principally  from  direct
application of technical hexachlorocyclohexane  (HCH) or  lindane  to  water for
control  of  mosquitoes or  from  the use of  HCH in agriculture and  forestry;
and  to  a  lesser  extent  from  occasional  contamination  of wastewater  from
manufacturing plants (U.S. EPA,  1979).
         Lindane has been  detected in the finished water of  Streator,  Illi-
nois, at a concentration of 4 ug/1 (U.S. EPA, 1975).
     B.  Food
         The daily  intake  of  lindane has been  reported  at  1  to  5 ug/kg body
weight and  the  daily  intake of all  other  HCH isomers at 1 to 3 ug/kg  body
weight (Duggan and  Duggan,  1973).   The chief sources of HCH  residues  in the
human  diet  are milk,  eggs,  and other  dairy products (U.S.  EPA,  1979)  and
carrots  and  potatoes   (Lichtenstein,  1959).   Seafood   is  usually  a  minor
source of HCH,  probably because of  the  relatively  high  rate of dissipation
of HCH in the aquatic environment (U.S. EPA, 1979).
         The U.S.  EPA  (1979)  has  estimated  the weighted average bioconcen-
tration  factor  for lindane to  be 780  for  the edible portions  of fish  and
shellfish consumed by Americans.   This estimate is based on measured steady-
state bioconcentration studies in bluegills.
     C.  Inhalation
         Traces of HCH have been detected in the air of  central  and suburban
London (Abbott, et al.  1966).   Uptake of lindane by , inhalation is estimated
at 0.002 jug/kg/day (Barney,  1969).
     0.  Dermal                                                       '
         Lindane has been  used  to eradicate human ectoparasites, -a few  ad-
verse reactions have been  reported  (U.S. EPA, 1979).
                                113-6

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III. PHARMACOKINETICS
     A.  Absorption
         The  rapidity  of  lindane absorption  is enhanced  by lipid-mediated
carriers.  Compared  to  other organochlorine insecticides,  lindane is unusu-
ally soluble  in water which  contributes to its  rapid  absorption and excre-
tion (Herbst  and  Bodenstein,  1972; U.S. EPA,  1979).   Intraperitoneal injec-
tions of  lindane  resulted in 35  percent absorption  (Koransky,  et al. 1963).
Lindane is also absorbed after oral and dermal exposure (U.S. EPA, 1979).
     B.  Distribution
         After  administration  to experimental animals,  lindane  was detected
in  the  brain  at  higher  concentrations than  in other  organs  (Laug,  1948;
Davidow and  Frawley, 1951; Koransky,_ et al. 1963; Huntingdon Research Cen-
ter, 1971).   At least 75  percent of an intraperitoneal  dose of   C-labeJ,ed
lindane was consistently found in the  skin,  muscle,  and fatty tissue (Koran-
sky, et  al.  1963).   Lindane  enters  the human  fetus through the placenta;
higher concentrations were found  in the skin than in  the brain,  but  never
exceeded the corresponding values for  adult  organs  (Poradovsky,  et al.  1977;
Nishimura, et al.  1977).
     C.  Metabolism
         Copeland and  Chadwick  (1979)  found that lindane  did not isomerize
in  adipose  tissues   in  rats,  but  noted dechlorination to  T*-3,4,5,6-tetra-
chlorocyclohexene.   Some other metabolites  reported  have  been 2,3,4,5,6-pen-
tachloro-2-cyclohexene-l-ol,    pentachlorophenol,   tetrachlorophenols,    and
three trichlorophenols  (Chadwick, et al. 1975;  Engst,  et  al. 1977),  all  of
which were  found in the  urine  as conjugates  (Chadwick  and Freal,  1972).
                                                                     »
Lindane metabolic pathways are  still  matters of some controversy  (Engst,  et
                                 . isii / -
                                ' I JJU
                                 1/3-7

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al. 1977; Copeland and  Chadwick,  1979).   Both free and conjugated chlorophe-
nols with  the possible exception' of pentachlorophenol (Engst,  et al. 1977)
are far less toxic than lindane (Natl. Acad. Sci., 1977).
     0.  Excretion
         Metabolites of lindane appear to be  eliminated  primarily as conju-
gates in the urine.  Very  little  unaltered lindane is excreted  (Laug, 1948).
Elimination of lindane  appears  to be rapid after administration ceases (U.S.
EPA, 1979).
IV.  EFFECTS
     A.  Carcinogenicity
         Nagasaki, et al.  (1972b) fed *{,/&, T~,  and 0 isomers separately
in the diet  to  mice at levels  of 100,  250,  and  500  ppm.   At termination of
the experiment  after 24 weeks, multiple  liver tumors, some  as  large as 2.0
centimeters in diameter were observed in  all  animals  given ^-HCH at the 500
ppm level.   The  250 ppmV -HCH level resulted in  smaller nodules,  while no
lesions were  found  at  levels of  100 ppm.   The various dosages  did  not pro-
duce any  tumors  with respect to  the other isomers.   Pathomorphological in-
vestigations by  Didenko,  et al.  (1973)  established  that the  IT isomer did
not induce  tumors in mice  given  intragastric administration  at doses of 25
mg/kg twice a week for five weeks.
         Hanada,   et  al. (1973)  fed  six-week-old mice a  basal  diet  of 100,
300, and 600 ppm  t-HCH  and  the^(, $,• IT isomers for a  period  of 32 weeks.
After  38  weeks,   liver  tumors  were  found  in   76.5  percent of the males and
43.5 percent  of  the  females fed  t-HCH,  indicating  males were  more highly
susceptible to HCH-induced  tumors than  females.   Multiple nodules were found
in  the  liver, although no  peritoneal  invasion  or  distinct  metastasis  was
found.   Thep -isomer-treated animals had no tumors.

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            Goto,  et al. (1972) essentially confirmed  the  findings of the above
   study using diets containing 600 ppm levels over a  26 week period.  The com-
   bination of/?-, T-,  or 0 -HCH  with  the highly carcinogenic  action  of °(-
   HCH  revealed  no  synergistic  or  antagonistic effect on  the  production of
   tumors by °( -HCH for dd strains of mice  (Ito,  et al. 1973).  Kashyap,  et al.
   (1979) found that 2T-HCH (14 percent lindane)  at 100 ppm  levels  in the diet
   or at  10 mg/kg/day  caused  liver and  lymphoreticular tissue tumors in both
   male and  female  mice after 45  weeks.   Application by  skin painting  had no
   effect.
            The National Cancer Institute conducted a  bioassay for the possible
   carcinogencity  of 0  -HCH  to Osbome-Mendel rats and 86C3F1  mice.   Adminis-
   tration continued  for  80 weeks  at  two  dose  levels:   time-weighted  average
   dose for male rats  was  236 and 472  ppm; for  female  rats, 135 and 275 ppm;
   and for all mice, 80 and 160 ppm.  NO  statistically significant incidence of
   tumor occurrence was noted  in any of  the experimental  rats as  compared to
   the controls.  At  the  lower dose concentration  in male mice,  the incidence
•'  of hepatocellular carcinoma was significant when compared to  the  controls,
   but not significant in the higher dose males.  "Thus, the  incidence of hepa-
   tocellular carcinoma in male mice cannot clearly be  related  to treatment."
   The incidence of hepatocellular carcinoma among female mice  was not signifi-
   cant.   Consequently, the  carcinogenic  activity  of  T'-HCH  in mice is  ques-
   tionable (Natl. Cancer Inst.,  1977).
        B.   Mutagenicity
            Some alterations  in mitqtic  activity and the karyotype of  human  ly-
   phocytes cultured with lindane  at 0.1 to 10 mg/ml have been  reported (Tsone-
   va-Maneva,  et al.  1971).  %" -HCH was  mutagenic  in  assays using  Salmonella
   typhimurium with  metabolic activation,  the  host-mediated  assay, and  the

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dominant lethal assay in rats.  Other reports  indicate  that  it  does  not  have
significant mutagenic activity (U.S.  EPA,  1979; Buselmair,  et al.  1973).
     C.  Teratogenicity
         Lindane given  in  the diet  during pregnancy at levels of 12 or 25
mg/kg body weight/day did  not produce teratogenic effects in rats (Mametku-
liev, 1978; Khera,  1979).
     0.  Other Reproductive Effects
         Chronic lindane feeding  in  a study of four generations  of  rats in-
creased the  average  duration of  pregnancy,  decreased  the  number of births,
increased the proportion of  stillbirths, and delayed sexual  maturation in F2
and F3 females.  In  addition, some of the  Fl and F2 animals  exhibited spas-
tic paraplegia (Petrescu,  et al.  1974 )_.
         In rats and rabbits, lindane given  in  the diet during  pregnancy in-
creased postimplantatlon death of embryos  (Mametkuliev, 1978; Palmer, et al.
1978).  Testicular atrophy has been observed in rats and mice (National Can-
cer Institute, 1977;  Nigam, et al. 1979).
     E.  Chronic Toxicity
         Irritation of the central nervous system with other toxic side ef-
fects  (nausea,  vomiting,  spasms,  weak  respiration  with .cyanosis and blood
dyscrasia) have been reported after  prolonged  or improper  use  of Hexicid (1
percent lindane) for the treatment of scabies on humans  (Lee,  et al.  1976).
         In chronic  studies  with  rats given lindane in oil, liver cell hy-
pertrophy  (fat  degeneration  and  necrosis)  and  nephritic changes  were noted
at  higher  doses (Fitzhugh,  et  al.  1950;  Lehman," 1952a,b).  Rats  inhaling
lindane (0.78 mg/m )  for  7  hours, 5  days  a week  for  180  days  showed liver
                                                                    »
cell  enlargement  but showed no  clinical   symptoms  or  other  abnormalities
(Heyroth,  1952).  The addition of 10 ppm lindane  to the diet  of rats for one
                               in-/*

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or  two  years decreased  body weight  after five months  of treatment and  al-
tered ascorbic  acid levels  in  urine, blood,  and tissues  (Petrescu,  et  al.
1974).   Dogs chronically  exposed  to lindane  in the  diet  had  friable  and
slightly enlarged livers (Rivett, et  al. 1978).
     F.  Other Relevant Information
         Lindane  is a  convulsant and  is the  most acutely  toxic isomer  of
hexachlorocyclohexane.  The  toxic effects of lindane are antagonized by  pre->.
treatment  with phenobarbitol  (LLtterst and  Miller, 1975)  and by  treatment
with  silymarin (Szpunar,  et al. 1976),  and various  tranouilizers  (Ulmann,
1972).
V.   AQUATIC TOXICITY
     A.  Acute Toxicity
         The  range  of adjusted LC5Q  values  for one  flow-through  and'24
static  bioassays  for lindane in  freshwater  fish ranged from  1 pg/l for  the
brown  trout Salmo  trutta  (Macek,- et al.  1970)  to 83 jug/1 for the goldfish
(Carassius  auratus),  and  represents  the  results  of tests  on 16  freshwater
fish  species  (U.S. EPA,  1979).   Zebrafish  (Brachydanio rerio)  showed  an
LC5Q  value of  120  ;jg/l  but rainbow  trout  (Salmo qairdneri)  exhibited  re-
spiratory distress  at 40 jug/1  (Slooff,  1979).   Among eight species of  fresh-
water invertebrates studied  with lindane,  stone flies (Pteronarcys  califomi-
ca) and three species of crustaceans: scuds  (Gammarus lacustris and G^ faci-
atus) and  sowbugs (Ascellus brevicaudus)  were most  sensitive, with adjusted
LC5Q  values  ranging  from  4  to  41 jug/1.    Three  species  of  cladocerans
(Daohnia pulex,  D^ maqna  and  Simocephalus  serralatus)  were  most resistant
with  LC5Q  values of  390  to 745 jjg/1.  The  midge (Chironomus  tentans)  was
intermediate in sensitivity with LC5Q values of 175 pg/1 (U.S. EPA, 1979).

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         Among eight  species  of marine fish  tested  in static bioassays  with
lindane, the Atlantic  silversides  (Menidia menidia)  was most sensitive,  with
an  acute  LC5Q  of 9  jjg/1 (Eisler,  1970), while  the  striped  mullet  (Muqil
cephalus)  was  reported  as having  an  acute  static  LC5Q of  66.0 ug/1  (U.S.
EPA,  1979).   The  results of six  flow-through  assays on  five  species  of
marine  fish  revealed  that the striped  bass (Morone  saxatilis) was most  sen-
sitive  with  an  acute LC_0  of 7.3  jug/1  (Korn  and  Earnest,  1974);  and  the
longnose  killifish (Fundulus  similis)  was  most  resistant  with  a  reported
LC-.,  of 240 jjg/1.   Acute studies  with six  species of marine  invertebrates
showed  these organisms   to  be  extremely  sensitive to  lindane,  with  LC5Q
values  ranging from 0.17 jjg/1 for the  pink shrimp,  Panaeus duorarum  (Schim-
mel, et al. 1977), to 8.5 ug/1 for the grass shrimp (Palaemonetes  vulqaris).
     B.  Chronic
         A chronic value  of  14.6 ug/1 was obtained for lindane in  a life-
cycle assay of the freshwater fathead  minnow (Pimephales promelas).  Chronic
values  of  3.3,  6.1,  and  14.5 ug/1  were obtained for three freshwater  inver-
tebrates, Chironomus  tentans,  Gammarus fasciatus, and  Daphnia maqna  (Macek,
et al.  1976).  No marine chronic studies were available.
     C.  Plant Effects
         For freshwater  algae,  Scenedesmus acutus,  the effective concentra-
tion  for  growth  inhibition  was  1,000 ug/1.   Effective concentrations   for
marine  phytoplankton  communities and  the  algae,  Acetabularia mediterranea,
were  1,000  and  10,000 pg/1, respectively.   Irreparable damage  to Chlorella
spec, occurred at concentrations greater than 300 ug/1  (Hansen, 1979).
     0.  Residues
                                                                     »
         Bioconcentration  factors  for  lindane  ranging  from  35  to  938 have
been obtained for  six  species of  freshwater fish and invertebrates.  No bio-
concentration factors  for lindane have been  determined for  marine organisms
                                  / -7 {('
                                X /'  ->' >" '^

-------
(U.S. EPA,  1979;  Sugiura,  et al. 1979).  Equilibrium accumulation  factors  of
429 to 602 were observed at  days 2  to 6 after exposure of Chlorella  spec,  to
10 to 400 ug/1 -of lindane  in aqueous  solution  (Hansen,  1979).
VI.  EXISTING GUIDELINES AND STANDARDS
     Neither  the  human health  nor  the aquatic  criteria  derived by  U.S. EPA
(1979), which are  summarized below, have gone through  the process of  public
review;  therefore,  there  is  a possibility  that  these  criteria  will  be
changed.
     A.  Human
         Using  the  "one-hit" model,  the  U.S.  EPA  (1979)  has estimated  that
the water  concentration of  lindane (gamma-HCH) corresponding  to a  lifetime
cancer risk for humans of 10    is  54 ng/1, based on the  data of Thorpe and
Walker (1973) for the induction  of  liver tumors  in male mice.
         Tolerance  levels set  by the  U.S.  EPA are  as follows:   7 ppm for
animal fat;  0.3 ppm for  milk;  1 ppm for most  fruits  and vegetables; 0.004
ppm  for  finished  drinking water;  and  0.5 mg/m  (skin)  for  air (U.S. EPA,
1979).  It  is not clear  whether these levels are for hexachlorocyclohexane
or for lindane.
     3,  Aquatic
         The criterion has been  drafted to  protect freshwater organisms as a
0.21 ug/1  24-hour average concentration  not  to exceed 2.9 ug/1.   Data are
insufficient to draft  criterion for  the protection of  marine life from gam-
ma-hexachlorocyclohexane (lindane).
                                I > 3

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           .  GAMMA-HEXACHLOROCYCLOHEXANE(LINDANE)


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-------
Hansen,  P.O.   1979   Experiments  on the accumulation of  lin-
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Ito, N-, et al.   1973.  Histologic and ultrastructural studies
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Laug, E.P.    1948.   Tissue  distribution  of  a  toxicant  fol-
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Lee, B., et al.   1976.  Suspected reactions to gamma benzene
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Lehman, A.J.   1952a.   Chemicals in food:''  A  report  to the
Assoc. of  Food  and  Drug  officials.  Assoc.  Food  and   Drug
Office, U.S. Quant. Bull.  16:  85.
                                                          »
Lehman,  A.J.    1952b.   U.S.  Assoc. Food  Drug Off.  Quant.
Bull. 16: 126.

-------
Hansen,  P.D.   1979  Experiments  on  the accumulation of  lin-
dane  (  -BHC)   by  the • primary  producers  Chlorella  spec.
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8: ITT.

Henderson, C.,  et al.   1971.  Organochlorine pesticide resi-
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Herbst, M., and G. Bodenstein.  1972.  Toxicology of lindane.
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Publ., Freiburg.

Heyroth, F.F.   1952.   In; Leland,  S.J.,  Chera.  Spec.  Manuf.
Assoc. Proc.  6:  110.

Huntingdon Research  Center.    1971.   In;  Lindane:  Monograph
of an  insecticide.   E.  Ullman  (ed.), Verlag K.  Schellenger,
(Pub.), p. 97,  1972.

Ito, N., et al.   1973.  Histologic and ultrastructural studies
on  the  hepato  carcinogenicity  of  benzene  hexachloride in
mice.  Jour. Natl. Cancer  Inst. 51:  817.

Kashyap, S.K.,  et al.  1979-.   Carcinogenicity of hexachloro-
cyclohexane  (BHC).  Jour.  Environ. Sci.  Health B14: 305.

Khera, K.S., et al.  1979.  Teratogenicity studies on pesti-
cides  formulations  of   dimethoate,   diuron  and lindane in
rats.  Bull. Environ. Contam. Toxicol. 22: 522.

Koransky, W.,  et al.   1963.   Absorption,  distribution,  and
elimination of  alpha- and  beta-  benzene hexachloride.   Arch.
Exp. Pathol. Pharmacol.  244: 564.

Korn,  S.,  and  R. Earnest.    1974.   The  acute   toxicity of
twenty  insecticides  to  striped  bass,  Marone  saxatilis>
Calif. Fish Game  60: 128.

Laug,  E.P.   1948.   Tissue  distribution  of  a  toxicant  fol-
lowing oral  ingestion  of  the  gamma-isomer  of benzene  hexa-
chloride by rats.  Jour.  Pharmacol. Exp. Therap.  93: 277.

Lee, B., et  al.   1976.   Suspected reactions to gamma benzene
hexachloride.  Jour. Am.  Med. Assoc.  236: 2846.

Lehman,  A.J.    1952a.  Chemicals in food:   A report  to the
Assoc.  of  Food  and  Drug  officials.    Assoc.  Food  and   Drug
Office, U.S. Quant. Bull.  16: 85.
                                                          »

Lehman,  A.J.   1952b.    U.S.  Assoc. Food  Drug   Off.  Quant.
Bull. 16: 126.

-------
Lichtenstein,  E.P.   195r.   Absorption  of  some chlotinted
hydrocarbon  insecticides  from  soils  into  various  crops.
Jour. Agric. Pood Chem. 7: 430.

Litterst, C.L.,  and E. Miller.  1975.   Distribution of lin-
dane  in  brains of  control and  phenobarbital  pretreated dogs
at the onset of lindane induced convulsions.   Bull. Environ.
Contarn. Toxicol. 13: 619.

Macek, K.J.,  and W.A^  McAllister.   1970.   Insecticide sus-
ceptibility  of  some   ccsnon  fish  family  representatives.
Trans. Am.  Fish. Soc.  99: 20.

Macek,  K.J.,  et al.   Ir76.   Chronic  toxicity  of lindane
to  selected aquatic  invertebrates  and  fishes.    EPA  600/3-
76-046.  U.S. Environ.  Prct. Agency.

Mametkuliev, C.H.   1978..   Study of embryotoxic  and terato-
genic  properties of the gamma  isomer  of HCH  in experiments
with rats.  Zdravookhr. T~ckm. 20:  28.

Mather,  S.P.,  and  J.G. Saba.   1975.   Microbial degradation
of  lindane-C-14 in a  flsoded  sand loam soil.   Soil  Sci.
120: 301.

Nagasaki,  H.,   et   al.   1972.   Carcinogenicity of benzene
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National  Academy of  Sciences -  National Research  Council.
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carcinogenicity of  lindans*  Fed. Reg.  Vol.  42. No.  218.

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feeding  on  testicular  -issue  on  pure  inbred  Swiss  mice.
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and fetuses.  Pediatrician 6: 45.

Palmer, A.K.,  et al.    1S73.   Effect of  lindane  on pregnancy
in the rabbit and rat.  Toxicology 9: 239. ,

Petrescu, S., et al.   19~4.   Studies on  the  effects  of long-
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Poradovsky,  R.,  et al.   1977.   Transplacental  permeation
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405.
                        IIZ-I7

-------
Lichtenstein,  E.P.   1959.    Absorption  of  some chlorinted
hydrocarbon  insecticides  from  soils  into  various  crops.
Jour. Agric. Food Chem. 7: 430.

Litterst, C.L.,  and E.  Miller.   1975.   Distribution of  lin-
dane  in  brains of control and phenobarbital pretreated  dogs
at the onset  of lindane induced convulsions.  Bull. Environ.
Contain. Toxicol. 13: 619.

Macek, K.J.,  and W.A.  McAllister.   1970.   Insecticide  sus-
ceptibility  of  some  common  fish  family  representatives.
Trans. Am.  Fish. Soc. 99: 20.

Macek,  K.J.,   et al.   1976.    Chronic  toxicity  of lindane
to  selected aquatic  invertebrates  and  fishes.    EPA 600/3-
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Mametkuliev, C.H.   1978.   Study of embryotoxic  and terato-
genic properties of the  gamma isomer  of HCH  in experiments
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Mather,  S.P.,  and J.G. Sana.    1975.   Microbial degradation
of  lindane-C-14  in  a  flooded  sand  loam soil.    Soil   Sci.
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Health p. 939.

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carcinogenicity of lindane.  Fed. Reg. Vol. 42. No.  2l8.

Nigam, S.K.,  et al.  1979.   Effect of hexachlorocyclohexane
feeding  on  testicular  tissue   on  pure  inbred  Swiss  mice.
Bull. Environ.. Contam. Toxicol.  23:  431.

Nishimura,  H.,  et  al.   1977.   Levels  of  polychlorinated
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Palmer, A.K.,  et al.   1978.   Effect of  lindane  on pregnancy
in the rabbit and rat.  Toxicology 9: 239.

Petrescu, S., et al.  1974.   Studies  on  the  effects of long-
term administration  of  chlorinated organic  pesticides  (lin-
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78: 831.

Poradovsky,  R.,   et  al.   1977.    Transplacental  permeation
of pesticides  during normal pregnancy.    Cesk Gynekol.  42:
405.

-------
Reuber,  M.D.    1979.   Carcinogenic!ty  of Lindane.   Environ.
Res. 19: 460.

Rivett,  K.F.,  et al.    1978.    Effects of  feeding  lindane
to dogs  for periods of up to  2  years.   Toxicology 9:  237.

Schimmel,  S.E.,  et al.    1977.  Toxicity  and  bioconcentration
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Environ. Contain.  Toxicol. 6:  355.

Sloof, W.   1979.  Detection  limits  of a biological  monitor-
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Toxicol. 23:  517.

Sugiura,  R.,  et  al.   1979.   Accumulation of  organochlorine
compounds  in  fishes.    Difference  of  accumulation  factors
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Szpunar, K.,  et al.   1976.  Effect of  silymarin  on  hepatoxic
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Thorpe,  E., and A.I.  Walker.   1973.  The toxicology  of  diel-
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Tsoneva-Maneva,  M.T., et  al.   1971.   Influence  of  Diazinon
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38: 344.

Tu,  C.M.   1975.   Interaction between  lindane and  microbes
'in soil.   Arch. Microbiol.  105: 131.

Tu,  C.M.   1976.   Utilization and  degradation  of  lindane
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U.S SPA.   1979.  Hexachlorocyclohexane:  Ambient Water Quality
Critera  (Draft).

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                                      No.  114
     Hexachlorocyclopentadiene

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.   20460

           APRIL 30,  1980
           /  '^^  f  I

          111-I

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

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                           HEXACHLOROCYCLOPENTADIENE
                                    Summary

     Hexachlorocyclopentadiene  (HEX) is  used as a  chemical intermediate  in
the  manufacture of  chlorinated pesticides.  Evidence  is  not  sufficient  to
categorize  this compound  as  a carcinogen or  non-carcinogen;  HEX  was  not
mutagenic  in either short-term  in  vitro  assays  or a  mouse dominant  lethal
study.   Teratogenic effects were not  observed in rats receiving oral  doses
of HEX during gestation.
     The  reported   96-hour  LC5Q value  for the  fathead minnow  under  static
and  flow-through conditions  using larval and  adult fish ranges from 7.0 pg/1
to  104 jug/1.   The  chronic  value for  fish in  an embryo-larval  test is  2.6
pg/1.

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                           HEXACHLOROCYCLOPENTADIENE
I.   INTRODUCTION
     Hexachlorocyclopentadiene  (HEX;   C5Clg)  is  a  pale  to  greenish-yellow
liquid.  Other physical properties include:  molecular  weight,  272.77;  solu-
bility  in  water,  0.805 mg/1;  and  vapor pressure, 1 mm Hg at  78-79°C.   HEX
is a highly reactive compound  and  is  used as a chemical  intermediate  in the
manufacture of  chlorinated pesticides (Kirk-Othmer, 1964).  Recent  govern-
ment bans  on  the use  of  chlorinated  pesticides  have  restricted the use  of
HEX  as  an   intermediate   to   the endosulfan  and   decachlorobi-2,4-cyclo-
pentadiene-1-yl industries.  Currently, the major use of  HEX  is as an inter-
mediate  in the  synthesis  of  flame retardants (Sanders, 1978;  Kirk-Othmer,
1964).   Production  levels  of HEX  approximate 50  million, pounds per  year
(Bell, et al.  1978).
   .  Environmental  monitoring  data for  HEX are  lacking, except  for  levels
measured in  the  vicinity  of  industrial  sites.   The  most likely route  of
entry of HEX  into the environment  arises from its  manufacture  or  the  manu-
facture  of HEX-containing  products.   Small amounts of  HEX  are present  as
impurities in pesticides made  from it; some HEX  has undoubtedly entered the
environment via this route.
     HEX appears  to be  strongly adsorbed  to soil  or  soil  components,  al-
though  others have  reported   its  volatilization  from  soil  (Rieck,   1977a,
1977b).    HEX  degrades   rapidly  by   photolysis,   giving   water-soluble
degradation products  (Natl. Cancer  Inst.,  1977).   Tests  on its  stability
towards hydrolysis at  ambient  temperature indicated  ''a half-life of about  11
days at pH3-6, which was  reduced to 6  days at pH  9.
                                  lit-

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II.  EXPOSURE
     A.  Water
         HEX has  been  detected in water near  points  of industrial discharge
at levels ranging  from 0.156  to  18 mg/1 (U.S. EPA, 1979).   Other than this,
there  is little   information  concerning  HEX concentrations  in  surface  or
drinking waters.   Due  to its low  solubility,  photolability,  and tendency to
volatize, one would not expect HEX to remain in flowing water.
     8.  Food
         HEX has  been  identified in a  few  samples-- of fish taken from waters
near  the Hooker   Chemical  Plant  in Michigan . (Spehar,  et  al.  1977).   No
reports concerning HEX contamination of other foods could be located.
         The U.S.  EPA. (1979) has  estimated the weighted  average bioconcen-
tration factor of  HEX  for  the edible portions of fish and shellfish consumed
by Americans  to be  3.2.   This  estimate is  based  on  measured  steady-state
bioconcentration studies in fathead minnows.
     C.  Inhalation
         The most  significant chronic  exposure  to HEX occurs  among persons
engaged directly in  its  manufacture  and among production  workers fabricating
HEX-containing products.   Inhalation is the primary mode  of  exposure  to HEX
in the event of accidental  spills, illegal discharges,  or occupational situ-
ations.
III.  PHARMACOKINETICS
     A.  Absorption
         Kommineni  (1978)  found  in  rats  that HEX is  absorbed  through  the
squamous  epithelium  of  the  nonglandular  part  of  the  stomach,  causing
                                                                       »
necrotic changes,  and  that the major route  of elimination of HEX is through
the  lungs.   This  information  is based on  morphological  changes in  rats


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administered HEX  by  gayage.   Further study-with  guinea pigs showed that HEX
was  absorbed  through the  skin;  but,  unlike the  rat  stomach,  the squamous
epithelium of these animals did not undergo necrotic changes.
     B.  Distribution
         The tissues  of four rats administered single  oral doses of HEX re-
tained only  trace amounts of  the compound  after 7 days  (Mehendale,  1977).
For example, approximately 0.5 percent of the total dose was retained in the
kidney and less  than 0.5 percent in  the liver.   Other organs  and tissues -
fat,  lung,  muscle, blood,  etc.  - contained even less HEX.   Tissue homoge-
nates  from  rats  receiving  injections of    C-HEX showed that  93 percent of
the radioactivity  in  the kidney  and  68 percent in  the  liver were associated
with the cytosol  fraction (Mehendale, 1977).
     C.  Metabolism
         At least  four  metabolites were  present  in the urine of rats admini-
stered HEX  (Mehendale,  1977).  Approximately  70  percent of  the metabolites
were extractable using a hexane:isopropanol mixture.
     0.  Excretion
         Mehendale (1977)  found that  approximately 33 percent  of the total
dose  of  HEX administered  to rats via oral intubation was excreted  in  the
urine  after  7 days.   About  87 percent  of  that  (28.7 percent  of the total
dose)  was eliminated during the  first 24 hours.   Fecal  excretion accounted
for  10 percent  of the  total  dose;   nearly 60 percent of  the 7  day  fecal
excretion occurred during  the  first  day.  These  findings  suggest that elim-
ination of  HEX may occur  by routes  other  than  urine  and  feces,  and  it  has
been postulated that a major route of excretion may be  the respiratory.tract.

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         Whitacre  (1978)  did not agree  with-the study  by  Mehendale (1977).
This recent study  of HEX  excretion  from mice and  rats  showed that excretion
was mainly by the fecal route with no more than 15 percent in the urine.
         Approximately nine  percent  of an injected dose  of  HEX was excreted
in the  bile in one hour  (Mehendale, 1977).  Because this  quantity is equi-
valent to that  excreted  in  the  feces  over seven days,  enterohepatic circu-
lation of this compound is probable.
IV.  EFFECTS
     A.  Carcinogenicity
         Only .one  in  vitro  test of HEX for  carcinogenic activity could be
located.  Litton Bionetics  (1977)  reported the  results of a test  to deter-
mine whether  HEX  could  induce  malignant  transformation in  BALB/3T3 cells.
HEX was  found  to  be relatively  toxic  to cells,  but no significant carcino-
genic activity was reported with this assay.
         The National  Cancer Institute  (1977)  concluded  that  toxicological
studies conducted  thus far have  not been adequate for evaluation of the car-
cinogenicity of HEX.  Because of this  paucity of information and HEX's high
potential for  exposure,  HEX has been  selected for the NCI's carcinogenesis
testing program.
     B.  Mutagenicity
         HEX has  been reported  to  be non-mutagenic  in  short-term in vitro
mutagenic  assays   (Natl.  Cancer Inst.,   1977;   Industrial  Bio-Test  Labora-
tories,  1977; Litton  Bionetics,  1978a) and in a  mouse  dominant lethal assay
                                                     .-
(Litton Bionetics, 1978b).

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     C.  Teratogenicity
         International  Research and  Development Corporation  (1978)  studied
the effect of oral  doses  of up to 300 mg/kg/day of  HEX administered to rats
on days 6 through 15 of gestation.   Teratogenic effects were not reported at
doses up  to  100 mg/kg/day;  the highest  dosage (300 mgAg/day)  resulted in
the death of all rats  by day ten of'gestation.   In this study, elimination
via the respiratory tract did not appear to be  significant.
     0.  Other Reproductive Effects
         Pertinent  information  could not be located in the available liter-
ature.
     E.  Chronic Toxicity
         There are  very few  studies  concerning the chronic  toxicity  of HEX
in laboratory  animals.   Naishstein  and Lisovskaya  (1965)  found  that  daily
administration of  1/30  the median  lethal dose (20  mg/kg)  for  6 months res-
ulted in the death of two of  ten  animals.   The investigators judged the cum-
ulative effects  of  HEX  to be weak;  no  neoplasms or ether abnormalities were
reported.  Naishstein and Lisovskaya  (1965) applied  0.5 to  0.6  ml of a solu-
tion  of  20  pom  HEX daily to the  skin  of rabbits  for  10 days  and  found no
significant   adverse effects  from exposure.   Treon, et  al. (1955)  applied
430-6130  mg/kg   HEX  to the  skin of rabbits.   Degenerative changes of the
brain, liver, kidneys,  and adrenal  glands of  these animals were  noted,  in
addition to  chronic skin  inflammation,  acanthosis,  hyperkeratosis,  and epil-
ation.  Further  study by Treon,  et al. (1955)  revealed  slight degenerative
changes in the liver  and  kidney of  guinea pigs, rabbits, and rats exposed to
0.15  ppm  HEX for daily seven-hour  periods over approximately  seven  months.
Four of five mice receiving the same dosage died within this period.

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         There  is  virtually  no information,  regarding the human  health ef-



fects  of  chronic  exposure  to  HEX.   According  to Hooker's  material safety



data sheet for  HEX,  (1972)  acute exposure to the compound results in irrita-



tion of  the eyes  and mucous membranes,  causing  lacrimation,  sneezing, and



salivation.  Repeated contact with the skin  can  cause blistering  and burns,



and inhalation  can cause pulmonary edema.  Ingestion  can  cause nausea, vom-



iting,  diarrhea, lethargy, and retarded respiration.



V.   AQUATIC TOXICITY



     A.  Acute Toxicity



         The .  reported   96-hour  LC5Q   values   for   the   fathead  minnow



(Pimephales promelas)  under static and  flow-through  conditions with larval



and adult  fish range from  7.0  fjg/1 to 104 jjg/1.  The effect of water hard-



ness is minimal (Henderson  1956; U.S. EPA,  1978).   There are  no  reports of



studies of the acute toxicity of HEX on saltwater organisms.



     B.  Chronic Toxicity



         In the only chronic  study  reported,  the lowest chronic  value for



the fat- head minnow  (embryo-larval) is 2.6 jjg/1 (U.S. EPA, 1978).



     C.  Plant Effects



         Pertinent  information  could  not be  located  in the available liter-



ature.



VI.  EXISTING GUIDELINES AND STANDARDS



     Neither the human health nor the aquatic criteria derived by U.S. EPA



(1979a),  which are summarized below, have  gone  through the process of public
                                                    .•


review;  therefore,  there  is a  possibility  that  these  criteria  will  be



changed.                                                               »


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     A.  Human
         The  Occupational Safety  and  Health Administration  has not  set a
standard  for  occupational  exposure  to  HEX.   The  American  Conference  of
Governmental  Industrial Hygienists  has  adopted a threshold limit value (TLV)
of 0.01  ppm (0.11 mg/m ) and  a  short term exposure  limit  of  0.03 ppm (0.33
mg/m3) (ACGIH, 1977).
         The draft ambient water quality  criterion for HEX is 1.0 ug/1 (U.S.
EPA, 1979).
     8.  Aquatic
         For HEX, the  draft  criterion to protect  freshwater  aquatic life is
0.39 ;jg/l  as a 24-hour average,  not to  exceed 7.0  jug/1  at any  time (U.S.
EPA, 1979).   Criteria., have not  been  proposed for  saltwater  species because
of insufficient data.
                                \ I if-in

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                           HEXACHLOROCYCLOPENTADIENE

                                  REFERENCES
American  Conference  of Governmental  Industrial  Hygienists.    1977,  TLV's:
threshold  limit  values for  chemical substances  and  physical agents  in the
workroom environment with intended changes for 1977.  Cincinnati, Ohio.

Bell, M.A., et al.   1978.  Review  of the environmental effects of pollutants
XI.   Hexachlorocyclopentadiene.   Report by  Battelle  Columbus Lab.  for U.S.
EPA Health Res. Lab., Cincinnati, Ohio.

Henderson, D.   1956.   Bioassay  investigations for International  Joint Com-
mission.   Hooker Electrochemical  Co.,  Niagara   Falls,  N.Y.   U.S.  Dep.  of
Health  Educ.  Welfare,  Robert  A.  Taft  Sanitary  Eng.   Center,  Cincinnati,
Ohio.  12 p.

Hooker  Industrial Chemicals  Division.   1972.   Material safety data  sheet:
Hexachlorocyclopentadiene.-  Unpublished internal memo dated April, 1972.

Industrial  Bio-Test  Laboratories,  Inc.   1977.   Mutagenicity  of  PCL-HEX
incorporated  in  the  test  medium tested against  five strains  of  Salmonella
typhimurium  and  as  a volatilate against tester  strain  TA-100.   Unpublished
report submitted to Velsicol Chemical Corp.

International  Research  and Development Corp.  1978.   Pilot teratology study
in rats.  Unpublished report submitted to Velsicol Chemical Corp.

Kirk-Othmer  Encyclopedia  of  chemical  technology.   2nd  ed.   1964.   Intersci-
ence Publishers, New  York.

Kommineni,  C.   1978.  Internal memo  dated  February  14, 1978,  entitled:
Pathology  report  on  rats  exposed to hexachlorocyclopentadiene.   U.S.  Dep. of
Health  Ed.  Welfare,   Pub.  Health  Serv.  Center for  Dis.  Control, Natl. Inst.
for Occup. Safety and Health.-

Litton  Bionetics, Inc.   1977.    Evaluation  of  hexachlorocyclopentadiene in
vitro malignant  transformation  in  BALB/3T3  cells:  Final  rep.   Unpublished
report submitted  to Velsicol Chemical Corp.

Litton  Bionetics,  Inc.   1978a.   Mutagenicity  evaluation of hexachlorocyclo-
pentadiene  in the mouse  lymphoma forward mutation assay.   Unpublished rep.
submitted  to Velsicol Chemical Corp.

Litton  Bionetics,  Inc.   1978b.   Mutagenicity  evaluation of hexachloropenta-
diene in  the mouse  dominant  lethal assay:  Final  report.   Unpublished rep.
submitted  to Velsicol Chemical Corp.

Mehendale,  H.M.    1977.  The chemical reactivity  - absorption,  retention,
metabolism,  and  elimination  of  hexachlorocyclopentadiene.   Environ.  Health,
Perspect.  21: 275.
                                     11

-------
Naishstein,  S.Y.,  and E.V.  Lisovskaya.   1965.   Maximum permissible concen-
tration of  hexachlorocyclopentadiene  in  water bodies.   Gigiena i Sanitariya
(Translation) Hyg. Sanit.  30: 177.

National  Cancer  Institute.  1977.  Summary of data  for chemical selection.
Unpublished  internal  working paper,  Chemical  Selection Working  Group, U.S.
Dep. of Health Edu. Welfare, Pub. Health Serv., Washington, O.C.

Rieck,  C.E.  1977a.   Effect of hexachlorocyclopentadiene  on soil  microbe
populations.   Unpublished  report  submitted  to  Velsicol  Chemical  Corp.,
Chicago, 111.

Rieck,  C.E.   1977b.    Soil  metabolism  of  14C-hexachlorocyclopentadiene.
Unpublished report submitted to Velsicol Chemical Corp., Chicago,  111.

Sanders,  H.J.   1978.   Flame  retardants.   Chem.  Eng.  News:   April  24,
1978: 22.

Spehar,  R.L.,  et  al.   1977.  A  rapid assessment of the toxicity  of three
chlorinated  cyclodiene  insecticide intermediates  to  fathead  minnows.   Off.
Res. Oev. Environ. Res. Lab., Ouluth,  Minn.  U.S. Environ. Prot. Agency.

Treon,  J.F.,  et  al.  '1955.    The   toxicity  of  hexachlorocyclopentadiene.
Arch. Ind. Health.  11: 459.

Whitacre,  O.M.   1978.   Letter  to  R. A.  Ewing,  Battelle Columbus  Labora-
tories,  dated  August  9,  1978.   Comments  on  documpnt  entitled:   Review  of
Environmental Effects of Pollutants  XI.   Hexachlorocyclopentadiene.

U.S.  EPA.   1978.  In-depth  studies  on  health  and environmental  impacts  of
selected  water  pollutants.  Contract No.   68-01-4646.   U.S.  Environ.  Prot.
Agency,.Washington, 0.C.

U.S. EPA.   1979.   Hexachlorocyclopentadiene:   Ambient Water Quality  Criteria
(Draft).

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                                      No. 115
          Hexachloroethane

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone scrutiny to
ensure its technical accuracy.

-------
                               HEXACHLOROETHANE

                                   SUMMARY

     Results of  a National  Cancer Institute  (NCI)  carcinogenesis  bioassay
showed  that hexachloroethane  produced an  increase  in  hepatocellular car-
cinoma incidence in mice.
     Testing of  hexachloroethane  in  the  Ames  Salmonella assay  showed no
mutagenic effects.   No  teratogenic effects were observed  following oral or
inhalation  exposure  of  rats to hexachloroethane, but  some toxic effects on
fetal development were observed.
     Toxic  symptoms  produced in  humans  following  hexachloroethane  exposure
include  central  nervous  system  depression and  liver,  kidney,  and  heart
degeneration.
     Hexachloroethane is  one of  the  more  toxic of  the chlorinated  ethanes
reviewed for aquatic organisms with marine invertebrates appearing  to  be the
most  sensitive organisms  studied.   This  chlorinated  ethane also  had the
greatest bioconcentration factor,  139  for bluegill sunfish, observed in this
class of compounds.
                                   //S--3

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                               HEXACHLOROETHANE
I.  INTRODUCTION
     This  profile  is based  on the  Ambient Water  Quality  Criteria Document
for Chlorinated Ethanes (U.S. EPA, 1979a).
     The chloroethanes are hydrocarbons  in which one or more of the hydrogen
atoms are  replaced by chlorine  atoms.   Water solubility  and vapor pressure
decrease with  increasing chlorination,  while density and  melting point  in-
crease.  Hexachloroethane  (Perchloroethane; M.w. 236.7)  is a  solid  at  room
temperature  with  a boiling  point of 186°C,  specific gravity  of 2.091;  and
is insoluble in water (U.S. EPA, 1979a).
     The chloroethanes are used  as solvents,  cleaning and degreasing agents,
and  in  the  chemical  synthesis of  a number  of  compounds.   Hexachloroethane
does not appear to  be commercially produced in the U.S., but 730,000 kg  were
imported in 1976.   (U.S. EPA, 1979a).
     The chlorinated  ethanes  form azeotropes with water  (Kirk  and Othmer,
1963).  All  are very soluble  in organic  solvents  (Lange,  1956).  Microbial
degradation  of  the  chlorinated -ethanes has not  been, demonstrated (U.S.  EPA,
1979a).
     The reader is  referred  to the Chlorinated  Ethanes  Hazard Profile for a
more general discussion of chlorinated ethanes (U.S. EPA, 1979b).
II.  EXPOSURE
     The chloroethanes are present in raw  and finished  waters due primarily
to industrial discharges.  Small amounts of  the  chloroethanes  may be formed
by chlorination  of drinking  water or treatment of sewage.   Air  levels  are
produced by evaporation of volatile chloroethanes.
                                                                        »
     Sources of  human  exposure to chloroethanes  include  water, air,  contam-
inated foods and fish, and dermal  absorption.  Fish and  shellfish have shown

-------
levels of  chloroethanes  in  the  nanogram range  (Dickson and  Riley,  1976).
Information on the levels of hexachloroethane in foods is not available.
     U.S.  EPA (1979a)  has  estimated  the weighted  average  bioconcentration
factor for hexachloroethane to  be 320  for  the edible  portion of  fish and
shellfish  consumed  by  Americans.  This  estimate  is  based  on  the octanol/
water partition coefficient.
III. PHARMOKINETICS
     Pertinent  data could  not  be located  in the  available  literature on
hexachloroethane  for absorption,  distribution, metabolism,  and  excretion.
However,  the  reader  is  referred  to a more general treatment of chloroethanes
(U.S. EPA,  1979b) which indicates rapid  absorption  of chloroethanes follow-
ing  oral  or  inhalation  exposure;  widespread distribution  of the chloro-
ethanes  through the  body;   enzymatic dechlorination  and oxidation to the
alcohol and ester forms;  and excretion of the chloroethanes  primarily in the
urine and in expired air.
IV.  EFFECTS
     A.  Carcinogencitiy
         Results  of  an  NCI  carcinogenensis  bioassay  for  hexachloroethane
showed that oral administration  of the compound produced an  increase  in the
incidence  of  hepatocellular  carcinoma in mice.  No statistically significant
tumor increase was seen in rats.
     8.  Mutagenicity
         The  testing of hexachloroethane in the Ames  Salmonella assay  or in
a yeast mutagenesis  system  failed to show any mutagenic activity  (Weeks, et
al. 1979).

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     C.  Teratogenicity
         Teratogenic  effects  were not  observed in pregnant  rats exposed  to
hexachloroethane by inhalation or intubation (Weeks, et al. 1979).
     0.  Other Reproductive Effects
         Hexachloroethane administered  orally  to pregnant rats decreased the
number of  live fetuses per  litter and  increased  the fetal  resorption rate
(Weeks, et al. 1979).
     E.  Chronic Toxicity
         Toxic symptoms  produced in humans  following hexachloroethane expo-
sure  include liver,  kidney,   and heart  degeneration,  and  central nervous
system depression (U.S. EPA, 1979a).
         Animal studies  have  shown that chronic exposure to hexachloroethane
produces both hepatotoxicity and nephrotoxicity  (U.S. EPA, 1979a).
V.   AQUATIC TOXICITY
     A.  Acute Toxicity
         Among   freshwater   organisms,   the    bluegill   sunfish   (Lepomis
macrochirus)  was  reported  to  have  the  lowest  sensitivity   to  hexachloro-
ethane, with a 96-hour  static LC5Q value  of  980 pg/1.   The 48-hour static
LC5Q  value  of  the  freshwater Cladoceran  (Daphnia maqna)  was  reported  as
8,070 ug/1 (U.S.  EPA,  1978).   For  the marine  fish,  the  sheepshead  minnow
(Cyprinodon  varieqatus).  a 96-hour  LC-0  value of  2,400 /jg/1 was  reported
from  a  static assay.   The  marine mysid shrimp  (Mysidopsis  bahia) was  the
most  sensitive aquatic  organism tested,  with a  96-hour static LC5Q  value
of 940pg/l  (U.S.  EPA, 1978).
     B.  Chronic Toxicity
                                                                        »
         Pertinent data could not be located in the available literature.
                                          //«

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     C.  Plant Effects

         For  the  freshwater  algae,  Selenastrum capricornutum,  the 96-hour

ECcn  effective  concentrations  based  on  chlorophyll and  cell  number were
  j\j

87,000  and  93,200  ug/1  for  chlorophyll  a  production  and  cell  growth,

respectively.    The   marine  algae,  Skeletonema  costatum,   was  much  more

sensitive,  with effective  concentrations  from  7,750  to  8,570  ug/1  being

reported.

     0.  Residues

         A  bioconcentration  factor  of 139  was dbtained for  the freshwater

bluegill sunfish (U.S. EPA, 1979a).

VI.  EXISTING GUIDELINES AND  STANDARDS

     Neither  the human health  nor  the aquatic criteria derived  by U.S. EPA

(1979a), which are summarized below,  have gone through  the  process of public

review;  therefore,  there  is  a  possibility  that  these  criteria  will  be

changed.

     A.  Human

         By applying  a linear,  non-threshold model to the  data from the NCI

bioassay for carcinogenesis,  the U.S.  EPA (1979a) has estimated the level of

hexachloroethane in  ambient water  that will  result in an  additional risk of

10"5 to be 5.9 ug/1.

     The  eight-hour  TWA exposure  standard  established  by  OSHA  for  hexa-

chloroethane is 1 ppm.

     8.  Aquatic Toxicity
                                                    »•
         The  proposed criterion  to  protect freshwater aquatic  life  is  62

ug/1 as  a  24-hour average and  should  not exceed 140 /jg/1 at any  time.   The

drafted  criterion  for saltwater aquatic  life  is a  24-hour  average concen-

tration of 7 ug/1 not to exceed 16 pg/1 at any time.

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                       HEXACHLOROETHANE

                          REFERENCES

Dickson, A.G., and J.P. Riley.  1976.  The distribution
of short-chain halogenated aliphatic hydrocarbons in some
marine organisms.  Mar. Follut. Bull. 79: 167.

Kirk, R., and D. Othxner.  1963.  Encyclopedia of Chemical
Technology.  2nd ed.  John Wiley and Sons, Inc. New York.

Lange, N.  (ed.)  1956.  Handbook of Chemistry.  9th ed.
Handbook Publishers, Inc. Sandusky, Ohio.

National Cancer Institute.  1978.  Bioassay of hexachloro-
ethane for possible carcinogenicity.  Natl. Inst. Health,
Natl.  Cancer Inst. DHEW Publ. No. (NIH)  78-1318.  Pub.
Health Serv. U.S. Oept. Health Edu. Welfare.

U.S. EPA.  1978.  In-depth studies on health and environ-
mental impacts of selected water pollutants.  U.S. Environ.
Prot.  Agency.  Contract No. 68-01-4646.

a.S. EPA.  1979a.  Chlorinated Ethanes:   Ambient Water Qual-
ity Criteria (Draft).

U.S. EPA.  1979b.  Environmental Criteria and Assessment
Office.  Chlorinated Ethanes:  Hazard Profile (Draft).

Van Dyke, R.A., and C.G. Wineman.  1971.   Enzymatic dechlori-
nation:  Dechlorination of chloroethane  and propanes in-
vitro.  Biochem.  Pharmacol. 20: 463.

Weeks, M.H.,•et al.  1979.  The toxicity of hexachloroethane
in laboratory animals.  Am. Ind. Hyg. Assoc. Jour. 40: 137.
                         //r-r

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                                      No. 116
          Hexachlorophene

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                                HEXACHLOROPHENE
                                    Summary

     Oral,  dermal,  and  subcutaneous  administration  of  hexachlorophene  in
animal studies has failed to show significant carcinogenic effects.
     Mutagenic effects  of hexachlorophene  exposure have been reported in one
study which  indicated increased chromosome aberrations in rats.   Testing  of
hexachlorophene in the  host mediated assay or  the dominant  lethal assay did
not produce positive effects.
     Several  reports have  indicated that hexachlorophene  may produce  some
teratogenic  and  embryotoxic effects.  A three generation  feeding study  in
rats  failed  to  show any  teratogenic activity.   Hexachlorophene has  shown
some adverse effects on male reproductive performance.
     Chronic administration  of hexachlorophene has produced  central  nervous
system effects and muscular paralysis.

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 I.   INTRODUCTION
     Hexachlorophene   (C13H6a2CL6,   molecular  weight   406.9)   is  a  white
 powder  which melts  between 166°c  and 167°C.   The compound  is practically
 insoluble in water but is soluble in ethanol,  ether,  and other organic sol-
 vents.  Under alkaline conditions,  hexachlorophene forms water-soluble salts
 (IARC,- 1979).
     The principle uses of hexachlorophene have been  for the  manufacture of
 germicidal  soaps,  as a  topical  anti-infective agent  for humans, as  a vet-
 erinary anti-helminthic, for  disinfection of  hospital  equipment,  and  as a
 broad-spectrum  soil  fungicide (IARC,  1979).   Limitation  of drugs and cos-
 metics containing hexachlorophene was instituted by the FDA in 1972.
     Commercial hexachlorophene  produced  from 2,4,5-trichlorophenol contains
 less than 15 ug/kg of 2,3,7,8-tetrachlorodibenzo-para-dioxin (IARC,  1979).
 II.  EXPOSURE
     There are  no  available estimates on daily exposure  levels  of  humans to
hexachlorophene from air, water,  or food.   .
     Water monitoring  studies  have detected  hexachlorophene in  two finished
drinking water  samples  (Shackelford  and  Keith,  1976) and  in  effluents  of
sewage treatment  plants at  levels  of 3.2 to 44.3 ug/1  (Sims.and  Pfaender,
1975),  as  well as in creek sediments (9.3 to 377 jjg/kg).
     Data on  hexachlorophene levels  in aquatic organisms indicate that  the
compound is bioaccumulated (Sims and Pfaender, 1975).
     Hexachlorophene has been  detected  in human milk at  levels  up  to  9 ;jg/l
 (West,  et al.  1975).   Blood  levels of the  compound  in  users  of soap  con-
                                                                         »
taining  hexachlorophene  have  been  reported  (0.02   to  0.14  mg/1   blood)
 (Butcher,  et al. 1973); blood levels fall  after use is discontinued.
     A 1974  survey  by NIOSH indicated  that  exposure  to  hexachlorophene  was
primarily  in hospitals, sanitariums, and convalescent  homes (IARC, 1979).

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III. PHARMACOKINETICS
     A.   Absorption
          Systemic  toxicity  following dermal  application  or  ingestion  of
hexachlorophene indicates that  the  compound  is  absorbed through the skin and
the gastrointestinal tract (AMA Drug Evaluations, 1977).
     B.   Distribution
          Whole-body autoradioigraphs  of the murine fetus  during  late ges-
tation following administration of  labelled  hexachlorophene  indicate an even
distribution pattern of the compound  .   The compound  crosses  the  placenta;
fetal  retention  increases during  the course  of pregnancy  (Brandt,  et  al.
                                                     \
1979).  Hexachlorophene has been detected in human  adipose samples  at levels
of 0.80 jug/kg (Shafik,  1973).
     C.   Metabolism
          Hexachlorophene is  metabolized by the liver,  producing   a  glucu-
ronide conjugate.   The clearance of blood  hexachlorophene  is  dependent  on
this hepatic activity (Klaassen, 1979).
     0.   Excretion
          Within three  hours of hexachlorophene  administration to  rats,  50
percent of the initial dose was excreted in  the bile (Klaassen,  1979).  Oral
administration of  the  compound to  a cow  resulted in excretion  of  63.8 per-
cent of  the  initial dose in  the  feces and  0.24 percent  in the urine (St.
John and Lisk,  1972).
IV.  EFFECTS
     A.   Carcinogenicity
          The lifetime  dermal application of  25-percept and 50-percent  so-
lutions  of  hexachlorophene   to mice  failed  to  produce  significant  car-
cinogenic effects  (Stenback,  1975);  the  levels  of compound  used caused nigh
toxicity.   Rudali   and   Assa   (1978)  were  unable  to  produce  carcinogenic
effects  in  mice  by  lifetime  feeding or  subcutaneous injection at  birth  of
hexachlorophene.    Oral  lifetime feeding of  hexachlorophene  to  rats  (17  to
150   ppm)   also   failed   to  show  carcinogenic   effects  (NCI,   1978).

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     B.   Mutagenicity
          Single    intraperitoneal    injections    of   2.5    or    5.0   mg/kg
hexachlorophene  failed  to induce dominant  lethal  mutations in mice (Arnold,
et al. 1975).
          Desi,  et  al.  (1975)   have reported  that  hexachlorophene   admin-
istered  to  rats   produced   chromosome   aberrations  (dose  and  route  not
specified).
     C.   Teratogenicity
          Kennedy, et al.  (1975a)  reported  that  the fetuses of pregnant rats
                                                     «.
exposed to hexachlorophene at 30 mg/kg on  days  6 to 15 of gestation  show a
low  frequency  of  eye defects and  skeletal abnormalities  (angulated   ribs).
Fetuses of rabbits exposed to this compound  at  6 mg/kg on days  6  to 18 of
gestation showed a  low incidence  of skeletal  irregularities,  but no soft
tissue anomalies  (Kennedy,  et al. 1975a).  A  three-generation feeding study
of hexachlorophene  to rats at levels of 12.5 to  50 ppm did not show tera-
togenic effects (Kennedy, et al.  1975b).
          A  single retrospective  Swedish  study  on  infants born  to  nurses
regularly exposed  to antiseptic  soaps containing hexachlorophene  has sug-
gested that  the  incidence of malformations in this  infant  population  is in-
creased (Hailing, 1979).
     D.   Other Reporductive Effects
          Gellert,  et  al.  (1978)  have  reported  that male neonatal  rats
washed for eight days with three  percent  hexachlorophene solutions showed as
adults a decreased fertility due  to inhibited reflex ejaculation.
          Oral administration of  hexachlorophene  to  rats  has been  reported
to.produce  degeneration  of  spermatogenic  cells  (Casaret  and Doull,  1975).
Subcutaneous  injection of hexachlorophene to mice  at  various periods of ges-
tation produced increased fetal resorptions  (Majundar, et al. 1975).

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     E.   Chronic Toxicity
          Administration  of hexachlorophene  by" gavage  (40  mg/kg)  produced
hind leg  paralysis  and growth impairment  after two to  three weeks (Kennedy
and  Gordon,  1976).   Histological  examination  showed  generalized edema  or
status spongiosus of  the  white matter of  the  entire  central nervous  system.
These  gross effects and  histopathological lesions  have been  reported  to  be
reversible  (Kennedy, et al. 1976).
          Central  nervous  system effects  in  humans  following chronic  ex-
posure to hexachlorophene include diplopia, irritability, weakness of  lower
extremities, and convulsions (Sax, 1975).
V.   AQUATIC TOXICITY
     A.   Acute and Chronic Toxicity and Plant Effects
          Pertinent data were not found in the available literature.
     8.   Residues
          Sims  and  Pfaender  (1975)   found levels  of  hexachlorophenol  in
aquatic organisms  ranging  from  335  ppb  in sludge  worms to  27,800  ppb  in
water boatman (Sigara spp.).
VI.  EXISTING GUIDELINES
     A.   Human
          Hexachlorophene  is  permitted as a  preservative  in drug and  cos-
metic products at levels up to 0.1 percent (USFDA, 1972).
     B.   Aquatic
          Pertinent data were not found in the available literature.

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American Medical Association.  1977.  AMA Council on Drugs, Chicago.

Arnold,  0.,   et  al.   1975.   Mutagenic   evaluation  of  hexachlorophene.
Toxicol. Appl. Phaimacol.  33: 185.

Brandt,  I.,  et  al.   1979.   Transplacental  passage  and  embryonic-fetal
accumulation of hexachlorophene in mice.  Toxicol. Appl. Pharmacol.  49: 393.

Butcher,  H./  et al.   1973.   Hexachlorophene  concentrations  in blood  of
operating room personnel.  Arch. Surg.  107: 70.

Casaret,  L.  and  J.  Doull.   1975.   Toxicology:   The  Basic  Science  of
Poisons.  MacMillan, New York.
                                                  \
Oesi,   I.,   et   al.    1975.   Animal   experiments   on  the   toxicity   of
hexachlorophene. Egeszsegtudomany  19: 340.

Gellert,   R.J.,  et  al.    1978.    Topical   exposure   of   neonates   to
hexachlorophene:  Long-standing effects  on  mating behavior and  prostatic
development in rats.  Toxicol. Appl. Pharmacol.  43: 339.

Hailing, H.   1979.   Suspected link  between exposure to  hexachlorophene  and
malformed infants.  Ann.  NY. Acad.  Sci.   320: 426.

International Agency  for Research  on Cancer.  1979.   IARC  monographs  on  the
evaluation of  the carcinogenic risk  of chemicals to  humans.  Vol.  20,  Some
Halogenated Hydrocarbons, p. 241.   IARC, Lyon.

Kennedy, G.L.,  Jr.  and D.E.  Gordon.   1976.   Histopathologic changes produced
by hexachlorophene in the rat as a function of both magnitude and number of
doses.  Bull. Environ. Contain. Toxicol.   16: 464.

Kennedy,  G.L.,   Jr.,  et  al.   1975a.   Evaluation  of  the  teratological
potential of hexachlorophene in rabbits and rats.  Teratology.  12: 83.

Kennedy, G.L. Jr., et al.  1975b.   Effect of hexachlorophene on  reproduction
in rats.  J. Agric.  Food Chem. 23:  866.

Kennedy, G.L. Jr.,  et al.   1976.   Effects  of  hexachlorophene in the rat  and
their reversibility.  Toxicol. Appl. Pharmacol.  35: 137.

Klaassen,  C.D.   1979.   Importance  of  hepatic  function  on   the  plasma
disappearance  and  biliary   excretion of   hexachloroghene.   Toxicol.   Appl.
Pharmacol.  49: 113.
                                  It 6'

-------
Majundar,  S.,  et  al.   1975.  Teratologic  evaluation of hexachlorophene  in
mice.  Proc. Pennsylvania Acad. Sci.  49: 110.
National Cancer  Institute.   1978.   Bioassay of Hexachlorophene  for Possible
Carcinogenicity  (Tech.  Rep.  Ser.  #40).   DHEW,   Publication  No.  78-840,
Washington.
Rudali,  G.  and  R.   Assa.    1978.   Lifespan  carcinogenicity  studies  with
hexachlorophene in mice and rats.  Cancer Lett.  5: 325.
Sax, N.   1975.   Dangerous Properties of Industrial Materials..  4th  ed.  Van
Nostrand Reinhold, New York.
Shafik,    T.     1973.     The   determination   of   pentachlorophenol    and
hexachlorophene in human  adipose tissue.  Bull.  Environ. Contamin.  Toxicol.
10: 57.
                                                  v
Shackelford,  W.  and  L.  Keith.   1976.   Frequency  of  organic  compounds
identified in water.   U.S. EPA, 600/4-76-062, p.  142.
Sims,  J.  and  F.  Pfaender.    1975.  Distribution  and  biomagnification  of
hexachlorophene in urban  drainage  areas.  Bull.  Environ. Contamin.  Toxicol.
14: 214.
St.  John,  L. and  0.  Lisk.   1972.  The  excretion  of hexachlorophene in  the
dairy cow.  J.  Agr. Food Chem.  20: 389.
Stenback,  F.   1975.    Hexachlorophene.   in   mice.   Effects  after  long-term
percutaneous applications.  Arch. Environ. Health,   30:  32.
West,  R.,   et  al.   1975.   Hexachlorophene concentrations  in  human  milk.
Bull. Environ.  Contamin. Toxicol.  13:  167.
                                    ±4 *7 7
                                    ) J r }
                                    \\t~1

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                                       No.  117
          Hydrofluoric Acid

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental  impacts  presented by  the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

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                             , HYDROFLUORIC ACID
                                   Summary

     Hydrofluoric acid (HF) has  produced mutagenic  effects in plants and
Drosophila, and lymphocyte chromosome aberrations in  rats. _Chromosome ef-
fects were not observed in mice  following sub-chronic inhalation exposure to
the compound.
     No data are avilable on the possible carcinogenic or teratogenic ef-
fects of HF.
     Chronic exposure to the compound has produced skeletal, fluorosis,  den-
tal mottling and pulmonary function impairment.
     One short-term bioassay test demonstrated that a concentration of
50,000 ug/1 HF was lethal to bluegill sunfish in one hour.
                                  It 7-J

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                              • HYDROFLUORIC ACID

 I.    INTRODUCTION
      Hydrofluoric acid (CAS registry number  7664-39-3)  (HF)  is a colorless,
 clear,  fuming  corrosive liquid made  by  treating  fluorspar fGaF-x)-Kith  sul-
 furic acid.  An unusual property  of  HF  is  that it will  dissolve glass  or  any
 other silica-containing material.  It has  the following physical and chem-
 ical  properties (Windholz,  1976;  Hawley, 1971; Weast, 1972):
                                   Pure                   Constant Boiling
      Formula:                     HF                          HF/H^
      Molecular Weight:           20.01                          —
      Melting Point:              -83.55QC                        —
      Boiling Point:              19.51QC                        —
      Density:                     0.987                    1.15 - 1.18
      Vapor Pressure:             1 atm i 19.5loc
      Solubility:                 Very soluble in water;
                                  soluble in many organic
                                  solvents, e.g., benzene,
                                 toluene, xylene, etc.
     HF is used in the aluminum industry, for the production of  fluoro-
carbons, for uranium processing,   for petroleum alkylation, for the produc-
tion of fluoride salts, and as a pickling agent for stainless steel.  It has
many other minor uses (CMR, 1978).
II.  EXPOSURE
     A.   Water
          Other than occasional leaks and spills, very small amounts of HF
are released into water from manufacturing and production facilities (Union
Carbide, 1977;  U.S. EPA, 1977a).   HF is released into the air from coal

-------
 fires  (U.S.  EPA,  1977b)  and from manufacturing and production facilities
 (Union Carbide, 1977).   HF released into the air has a high affinity for
 water,  and it is  expected  that  it will  rain out (Fisher,  1976).   The amounts
 of HF  in water and  the extent of its presence could not be determined from
 the  available literature.   Under alkaline conditions,  HF  will form aqueous
 salts.
     8.   Food  •
          Pertinent data were not found  in the available  literature.
     C.   Inhalation
          HF  occurs in the atmosphere from coal fires  and  from manufacturing
 and production facilities  (see above), as  well as  from the photochemical re-
 action of &Lf2 with NO and humid air (Saburo,  et al.  (1977).  It  is
 present in the stratosphere (Zander, et al. 1977; Orayson,  et al. 1977;
 Farmer and Paper, 1977).  The extent and amounts of HF in  the atmosphere
 could not be  determined from the available  literature.
     D.   Dermal
          Pertinent data were not found in the available literature.
 III. PHARMACOKINETICS
     A.   Absorption
          The major route of HF absorption is by the respiratory system;
penetration of liquefied anhydrous HF through the skin has been reported
 (Burke, et al. 1973).   Fatal inhalation of HF fumes resulted in a blood
 fluoride level of 0.4 mg/100 ml (Greendyke and Hodge, 1964),. while skin
penetration of anhydrous HF produced a maximum blood fluoride concentration
of 0.3 mg/100 ml (Burke,  et al.  1973).   These levels are 100-fold higher

-------
 than normal serum fluoride levels (Hall et al. 1972).  Forty-five percent of
 fluoride present in the air in gaseous or particulate form is absorbed on
 inhalation (Dinman,  et al.  1976).
      B.    Distribution
           Absorbed fluoride is deposited mainly in the skeleton and teeth;
 it  is also found in  soft tissues  and body fluids (NAS,  1971;  NIOSH,  1975;
 NIOSH, 1976)>  Fluoride reaches fetal circulation via the placenta and is
 deposited in the fetal skeleton (NAS,  1971).
           Fluoride deposition  in  bone is not  irreversible (NAS,  1971).   How-
 ever,  laboratory animals chronically exposed  to HF gas  retained abnormally
 high  levels of fluoride in  the skeleton  for up to 14  months after  exposure
 (Machle  and Scott, 1935).
      C.    Metabolism
           The physiological or biochemical basis of fluoride  toxicity  has
 not been  established,  although it appears that enzymes  involved  in vital
 functions are inhibited by  fluoride  (NAS, 1971).   Examination of the data of
 Collins,  et al.  (1951)  indicates  that metabolism of absorbed  fluoride  is the
 same whether it  is inhaled  as  a particulate inorganic or  gas  (as HF) (NIOSH,
 1976).
     0.    Excretion
           Fluoride is excreted in the urine,  feces  and  sweat, and  in trace
 amounts in milk, saliva, hair  and probably tears.   Data are lacking regard-
 ing loss of fluoride by expired breath (NAS,  1971).
                                                     *•
          The primary route of  fluoride elimination is  through the urine.
The urinary fluroide concentration is influenced by factors such as total
absorption, the  form of  fluoride absorbed, frequency of exposure and general

-------
 health (MAS,  1971).   It is recognized that urinary fluoride levels are di-
 rectly related to the concetration of absorbed fluoride (NAS, 1571).
           In  a relatively unexposed person,  about one-half of an acute dose
 of fluoride is excreted within 24 hours in the urine,  and about one-half is
 deposited in  the skeleton (NAS,  1571).
 IV.   EFFECTS
      A.    Carcinogenic!ty
           Pertinent  data were  not found in the available literature.
      B.    Mutagenicity
           Mohamed (1563)  has reported various  aberrations in  second  genera-
 tion  tomato plants following parenteral treatment with HF at  3 ^g/m^.
 These results  could  not be duplicated by Temple and Weinstein (1576).
           Rats inhaling 0.1 mg HF/m5  chronically  for two months were re-
 ported to  develop lymphocyte chromosomal aberrations;  aberrations could not
 be.detected in sperm cells of  mice administered the same levels of HF-
 (Voroshilin, et al.  1573).
           Weak mutagenic  effects  in the  offspring of Drosophila exposed to
 air bubbled through  2.5 percent HF have  been reported  (Mohamed, 1571).
      C.    Teratogenicity
           Pertinent  data  were  not found  in the  available literature.
     0.    Other Reproductive Effects
           Reduced  fertility in Drosophila and decreased  egg hatch have been
 reported following exposure to 2.5 ppm HF  (Gerdes, et al. 1571).
     E.    Chronic Toxicity
          Among the adverse physiologic effects of long-term exposure to HF
are skeletal fluorosis, dental mottling and pulmonary impairment (NAS,  1571;
NIOSH, 1575; NIOSH, 1576).  Skeletal fluorosis is characterized by increased

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bone  density,  especially  in  the  pelvis  and  spinal  column,  restricted spinal
motion,  and ossification  of  ligaments.   Nasal irritation,  asthma  or  short-
ness  of  breath, and in some  cases pulmonary fibrosis  are associated  with
HF-induced pulmonary distress  (NIOSH, 1976).   Digestive disturbances have
also  been noted (NIOSH, 1976}.   Fluoride-induced renal pathology  has not
been  firmly established in man (Adler,  et al.  1970).  Causal relationships
in industrial  exposures are  difficult to determine because exposure  often
involves other compounds  in  addition to  fluorides  (NIOSH, 1976).
          Laboratory animals chronically exposed to 15.2 mg HF/m3 devel-
oped  pulmonaryr kidney and hepatic pathology  (Machle  and Kitzmiller,  1935;
Machle, et al. 1934), while  animals exposed to 24.5 mg HF/m3 developed
lung  edema (Stokinger, 1949).  Testicular pathology was also observed in
dogs  at 24.5 mg HF/m3 (Stokinger, 1949).  Several animal studies have
demonstrated that inhalation of HF increased fluoride deposition in the
bones (NIOSH,.  1976).
      F.   Other Relevant  Information
          Fluoride has anticholinesterase character which,  in conjunction
with  the reduction in plasma calcium observed in fluoride intoxication, may
be responsible for acute nervous system effects (NAS, 1971).  The severe
pain  accompanying skin injury from contact with 10 percent HF has been at-
tributed to immobilization of calcium,  resulting in potassium nerve stimula-
tion  (Klauder, et al 1955).
          Inhibition of enolase,  oxygen uptake, and tetrazolium reductase
                                                      .•
activity has been demonstrated in vitro from application of HF to excised
guinea pig ear skin (Carney,  et al.  1974).

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 V.    AQUATIC TOXICITY
      A.    Acute  Toxicity
           McKee  and Wolf  (1963) reported that HF was toxic to. fish
 (unspecified at  concentrations ranging from 40,000 to 60,000 ;jg/l.  Bonner
 and Morgan (1576) observed that 50,000 ^ig/1 HF was lethal to bluegill sun-
 fish  (Lepomis macrochirus) in one hour.
      B.    Chronic Toxicity, Plant Effects, and Residue
           Pertinent data were not found in the available literature.
      C.    Other Relevant Information
           Bonner and Morgan (1976) observed a marked increase in the oper-
cular "breathing" rate of bluegill sunfish exposed to a concentration of
25,000 ug/1  for four hours.  The fish recovered within three days.
VI.  EXISTING GUIDELINES AND STANDARDS
      A.    Human
           In 1976, NIOSH proposed a workplace environmental  limit for HF of
2.5 mg/nh5  (3 pom) as a time-weighted average to provide protection  from
the effects of HF over a working lifetime  (NIOSH,  1976).  A  ceiling limit of
5 mg HF/nv5 based on 15-minute exposures was also recommended to  prevent
acute irritation from HS (NIOSH,  1976).
     B.   Aquatic
          Pertinent data were not found in  the  available  literature.
                                    . ^JS^
                                 •*; j J u
                                  117-f

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                               HYDROFLUORIC ACID

                                   References
 Adler,  P., et al.  1970.   Fluorides  and Human Health.  World Health Organi-
 zation,  Monograph 59,  Geneva.

 Sonner,  W.P.  and E.L.  Morgan.   1976.  On-line surveillance of industrial ef-
 fluents  employing chemical-physical methods  of  fish as sensorsa.   Dept. of
 Civil    Engineering,    Tennessee   Technological   University,   Cookeviller-
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 Available  from NTIS:   PB261-253.

 Burke,  W.J.,  et  al.   1973.   Systemic  fluoride poisoning resulting  from  a
 fluoride skin burn.  Jour. Occup. Med.   15: 39.

 Carney,  S.A., et  al.   1974.  Rationale of the treatment of hydrofluoric acid
 burns.   Br. Jour.  Ind. Med.  31: 317.

 Chemical  Marketing Reporter.   1978.   Chemical Profile  -  Hydrofluoric acid.
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 Collins,  G.H.,   Jr..,   et  al.   1951.   Absorption  and  excretion  of  inhaled
 fluorides.  Arch.  Ind. Hyg. Occup. Med.  4: 585.

 Dinman,  D.B.,  et  al.   1976.   Absorption  and excretion of  fluoride  immedi-
 ately after exposure.  Pt. 1.  Jour. Occup. Med.   18: 7.

 Drayson,  S.R.,  et al.   1977.   Satellite  sensing  of  stratospheric  halogen
 compounds  by  solar  occulation.   Part  1.   Low  resolution  spectroscopy.
 Radiat. Atmos. Pap. Int. Symp.   p. 248.

 Farmer,  C.B.  and  O.F.  Raper.    1977.   The hydrofluoric acid:   Hydrochloric
 acid ratio in the 14-38 km region of the  stratosphere.  Geophys. Res.  Lett.
 4: 527.

 Fisher, R.W.  1976.  An air pollution  assessment of  hydrogen fluoride.   U.S.
 NTIS.  AD Rep. AS-AS027458, 37 pp.

 Gerdes, R., et al.  1971.  The  effects of atmospheric hydrogen fluoride upon
 Drosophila  melanogaster.    I.    Differential  genotypic   response.    Atmos.
 Environ.  5: 113.

Greendyke, R.M.  and H.C. Hodge.  1964.   Accidental death due  to hydrofluoric
 acid.  Jour. Forensic Sci.  9:  383.

Hall, L.L., et al.  1972.  Direct potentiometric deterination  of total  ionic
 fluoride in biological fluids.   Clin.  Chem.  18:  1455.
                                                                       »
Hawley,  G.G.    1971.   The Condensed  Chemical  Dictionary.    8th   ed.    Van
Nostrand Reinhold Co.,  New York.
                                   / / /-/O

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 Klauder, J.V., et  al.   1955.  Industrial  uses of compounds  of fluorine and
 oxalic acid.   Arch. Ind.  Health.   12: 412  •

 Machle,  W. and K.  Kitzmiller.  1935.   The effects of the  inhalation of hy-
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 tion.   Jour.  Ind. Hyg.  Toxicol.   17:  223.

 Machle,  W. and  E.W. Scott.   1935.   The  effects of  inhalation of  hydrogen
 fluoride —  III.   Fluorine  storage  following exposure to sub-lethal concen-
 trations.   Jour.  Ind. Hyg. Toxicol.   17:  230.

 Machle,  W., et al.  1934.   The  effects of the inhalation of  hydrogen fluor-
 ide  —  I.  The   response  following  exposure to  high concentrations.   Jour.
 Ind. Hyg.   16: 129.

 McKee,  J.E. and H.W. Wolf.  1963.  Water Quality  Criteria.  California State
 Water  Quality Control Board  Resources Agency  Publication  No. 3-A.

 Mohamed,  A.   1968.   Cytogenetic  effects of hydrogen  fluoride treatment  in
 tomato plants.  Jour. Air Pollut.  Cant. Assoc.  18: 395.

 Mohamed,  A.   1971.   Induced recessive  lethals in  second  chromosomes  of
 Drosoghila  melanogaster. by hydrogen  fluoride.  In:   Englung,  H.,  Berry, W.,
 eos. Proc.  2nd internet. Clean Air Cong.- New YorT<:  Academic  Press.

 National  Academy  of Sciences.  1971.  Fluorides.  U.S.  National  Academy  of
 Sciences, Washington, DC.

 National  Institute  for Occupational  Safety and Health.   1975.  Criteria for
 a recommended standard  -  occupational exposure to inorganic fluorides.  U.S.
 OHEW,  National Institute for Occupational Safety and Health.

 National  Institute  for  Occupational  Safety and Health.   1976.  Criteria for
 a  recommended standard  - occupational  exposure  to hydrogen  fluoride, U.S.
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 Pub. No. 76-43.

 Saburo, K., et al.   1977.  Studies on  the  photochemistry of aliphatic  halo-
 genated  hydrocarbons.    I.    Formation of  hydrogen   fluoride  and  hydrogen
 chloride  by the   photochemical reaction of dichlorodifluoromethane  with ni-
 trogen oxides in  air.  Chemosphere p. 503.

 Stokinger,  H.E.   1949.  Toxicity  following  inhalation of fluorine and hydro-
gen fluoride.   In:   Voegtlin, Hodge,  H.C.,  eds.   Pharmacology  and Toxicology
of Uranium Compounds.  McGraw-Hill Book Co., Inc., New York.  p. 1021.

Temple,  P.  and L.   Weinstein.   1976.   Personal   communication.  Cited in:
Drinking Water and  Health.   Washington,  DC:   National Research Council,  p.
486.
                                                                        »
Union Carbide.  1977.   Environmental  monitoring report,  United States Energy
Research  and  Development  Administration,  Paducah  gaseous diffusion  plant.
NTIS Y/UB-7.

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U.S.  EPA.   1977a.   Industrial  process  profiles  for  environmental  use:
chapter  16.   The fluorocarbon-hydrogen  fluoride  industry.   U.S.  Environ.
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U.S. EPA.   1977b.  A survey  of sulfate,  nitrate  and  acid aerosol emissions
and their control.  U.S. Environ. Prot. Agency.  U.S. DHEW PB276-558.

Voroshilin,  S.I.,  et al.   1973.   Cytological effect  of  inorganic compounds
of fluorine on human and animal cells in vivo and in vitro.  Genetika 9: 115.

Weast, R.C..  1972.  Handbook  of Chemistry  and Physics.  53rd ed.  Cleveland,
OH:  Chemical Rubber Co.

Windholz, M.  1976.  The Merck  Index.   9th ed.   Merck  and Co., Inc., Rahway-,
N.J.

Zander, R.,  et  al.  1977.  Confirming  the presence of hydrofluoric acid in
the upper stratosphere.   Geophys. Res.  Lett.  4: 117.

-------
                                       No.  118
          Hydrogen Sulfide

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including  all the
adverse health  and   environmental  impacts  presented  by  the
subject chemical.   This  document  has  undergone scrutiny  to
ensure its technical accuracy.
                            11 *-3L

-------
                       Hydrogen Sulfide



                           Summary
     Pertinent information could not be located on the



carcinogenicity,  mutagenicity,  or teratogenicity of H£S.



     Hydrogen, sulfide is very toxic to humans via inhalation



and has been reported to cause death at concentrations of



800 to 1000 ppm.



     Hydrogen sulfide is reported to be very toxic to fish




with toxic effects resulting from 1 to 100 ppm.

-------
I.   INTRODUCTION

     Hydrogen sulfide (^S; CAS No. 7783064) is a colorless

flammable gas with a rotten egg odor.  It has the following

physical properties:

          Formula                  t^S

          Molecular Weight         34.08

          Melting Point            -85.5°C

          Boiling Point            -60.4°C

          Density                  1.539 gram per liter at 0°C

          Vapor Pressure           20 atm. at 25.5°C



     Hydrogen sulfide is soluble in water, alcohol, and

glycerol (ITII, 1976).   Hydrogen sulfide is a flammable gas

and the vapor may travel considerable distance to a source of

ignition and flash back.

     Hydrogen sulfide and other sulfur compounds occur to some

extent in most petroleum and natural-gas deposits.  Very

substantial quantities  of this gas are liberated in coking

operations or in the production of manufactured gases from

coal (Standen, 1969).  Hydrogen sulfide is used to produce

substantial tonnages of elemental sulfur, sulfuric acid, and

a variety of other chemicals.  Completely dry hydrogen sulfide,

whether gaseous or liquid, has no acidic properties.  Aqueous

solutions, however, are weakly acidic (Standen, 1969).  In

1965,  some 5.2 million  metric tons of H2S was recovered from
                                                            »
fossil fuels (Standen,  1969).

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II.  EXPOSURE



     A.   Water



          Bacterial reduction of sulfates accounts for the



occurrence of I^S in numerous bodies of water, such as the



lakes near El Agheila, Libya.  Hydrogen sulfide is familiarly



formed as a bacterial decomposition product of protein



matter, particularly of animal origin (Standen, 1969) and this



gas can be found in most sewage treatment plant and their



piping systems.



     B.   Food



          H2S may be formed within the gastrointestinal tract



after the ingestlon of inorganic sulfide salts or elemental



sulfur due to the actions of gastric acid and of colonic



bacteria. (Division of Industrial Hygiene, 1941).



     C.   Inhalation



          Wherever sulfur is deposited, pockets of hydrogen



sulfide may be encountered, thus it is found at coal, lead,



gypsum, and sulfur mines.  Crude oil from Texas and Mexico



contain toxic quantities of H2S (Yont and Fowler, 1926).   The



decay of organic matter gives rise to the production of I^S



in sewers and waste from industrial plants where animals



products are handled.  Thus, there has been accidental poisoning



from H2§ in tanneries, glue factories, fur-dressing and



felt-making plants, abattoirs, and beet-sugar factories;  for



example, in Lowell, Massachusetts five men were poisoned
                                                            »


(three died)  when sent to repair a street sewer which drained



waste from a tannery (Hamilton and Hardy, 1974).

-------
     Hydrogen sulfide is formed in certain industrial processes




such as the production of sulfur dyes, the heating of rubber




containing sulfur compounds, the making of artificial silk or




rayon by viscose process (Hamilton and Hardy, 1974).




     D.   Dermal




          Pertinent information could not be found in the




available literature.




III. PHARMACOKINENTICS




     A.   Absorption




          By far the greatest danger presented by hydrogen




sulfide is through inhalation, although absorption through




Ithe skin has been reported (Patty, 1967).




     B.   Distribution




          Pertinent information could not be. found in the




available literature.




     C.   Metabolism and Excretion




          Evidence has been obtained for the presence of a




sulfide oxidase in mammalian liver (Baxter and Van Reen,




1958; Sorbo, 1960), but important nonenxymatic mechanisms for




sulfide detoxication are also recognized.  Sulfide tends to




undergo spontaneous oxidation to non-toxic products such as




polysulfides, thiosulfates or sulfates (Gosselin, 1976).




     When free sulfide exists in the circulating blood a




certain amount of hydrogen sulfide is excreted in the exhaled




breath, this is sufficient to be detected by odor, but the




greater portion, however, is excreted in the urine, chiefly as




sulfate, but some as sulfide (Patty, 1967).
                             11    -

-------
IV.  EFFECTS



     A..   Carcinogenic! ty



          Pertinent information could noc be found in the



available literature.



     B.   Mutagenicity



          Pertinent information could not be found in the



available literature.



     C.   Teratogenicity



          Pertinent information could not be found in the



available literature.



     D.   Other Reproductive Efforts



          Pertinent information could not be found in the



available literature.



     E.   Chronic Toxicity



          At low concentrations of hydrogen sulfide (e.g., 50



to 200 ppm) the toxic symptoms are due to local tissue



irritation rather than to systemic actions.  The most



characteristic effect is on the eye, where superficial injury



to the conjunctiva and cornea is known to workers in tunnels,



caissons, and sewers as "gas eye" (Grant, 1972).  More



prolonged or intensive exposures may lead to involvement of



the respiratory tract with cough, dyspnea and perhaps pulmonary



edema.  Evidence of severe pulmonary edema has been found at



autopsy and in survivors of massive respiratory exposures
                                                            »


(Gosselin, 1976).  The irritating action has been explained



on the basis that I^S combines with alkali present in moist



tissues to form sodium sulfide, a caustic (Sax, 1979).  Chronic

-------
poisoning results in headache, inflammation of the conjunctivas



and eyelids, digestive disturbances, loss of weight, and



general debility (Sax, 1979).



     F.   Other Relevant Information



          Hydrogen sulfide is reported with a maximum safe



concentration of 13 ppm (Standen, 1969), although at first



this concentration can be readily recognized by its odor, H2S



may partially paralyze the olfactory nerve to the point at



which the presence of the gas is no longer sensed.  Hamilton



and Hardy (1974) report that at a concentration of 150 ppm,



the olfactory nerve is paralyzed.



     Exposures of 800-1000 ppm may be fatal in 30 minutes,



and high concentrations are instantly fatal (Sax, 1979).



There are reports of exceptional cases of lasting injury,



after recovery from acute poisoning, which point to an



irreversible damage to certain cells of the body resulting



from prolonged oxygen starvation (Hamilton and Hardy, 1974).



Hydrogen sulfide has killed at concentrations as low as



800 ppm (Verschueren, 1974).



V.   AQUATIC TOXICITY



     A.   Acute Toxicity



          Verschueren (1974) has reviewed the effects of H2S



on several aquatic organisms.  Goldfish have been reported to



die at a concentration of 1 ppm after long time exposure in
                                                            »


hard water.   Verschueren (1974) reports a 96-hour LC50 value of



10 ppm for goldfish.  Verschueren also reports on a large number



of fresh water fish with toxic effects resulting from exposure

-------
Co H2§ at concentrations ranging from 1 to 100 ppm.

     Verschueren (1974) reports median threshold limit values

for Arthropoda: Asellus, 96-hour at 0.111 mg/1;  Crangonyx,

96 hour at 1.07 mg/1;  and Gammarus, 96-hour at 0.84 mg/1.

     B.   Chronic Toxicity, Plant Effects and Residues

          Pertinent information could not be located in the

available literature.

     C.   Other Relevant Information

          Verschueren (1974) reports that sludge digestion is

inhibited at 70-200 mg/1 of ^S in wastewater treatment plants

VI.  EXISTING GUIDELINES AND STANDARDS

     A.   Human

          The 8-hour,  time-weighted average occupational

exposure limit for H£S has been set in a number  of countries

and are tabled below (Verschueren, 1974):


           T.L.V.:  Russia            7 ppm
                    U.S.A.            20 ppm "peak"
                    Federal German    10 ppm
                      Republic


     #2$ is a Department of Transportation flammable and

poisonous gas and must be labelled prior to shipment.

     B.   Aquatic

          Maximum allowable concentration of .0.1 mg/1 for

Class I and Class II waters has been established in Romania

and Bulgaria for I^S (Verschueren, 1974).
                             11 Co
                            ) -J/ U

-------
                          References
Baxter, C. F. and R. Van Reen.  19S8a.  Some Aspects of
Sulfide Oxidation by Rat Liver Preparations.  Biochem.
Biophys.  Acta 28: 567-572.  The Oxidation of Sulfide
to Thiosulfate by Metalloprotein Complexes and by
Ferritin.  Loc. cit. 573-578.  1958b.

Division of Industrial Hygiene.  1941.  Hydrogen Sulfide,
its Toxicity and Potential Dangers.  National Institute
of Health, U.S. Public Health Service.  Public Health
Rep. (U.S.) 56: 684-692.

Gosselin, R. £., et al.  1976.  Clinical Toxicology of
Commercial Products.  The Williams and Wilkins Company,
Baltimore.

Grant, W. M.  1972.  Toxiciology of the Eye.  2nd ed.
Charles C. Thomas, Springfield, Illinois.

Hamilton, A. and Harriet Hardy.  1974.  Industrial
Toxicology.  Third edition.  Publishing Science Group, Inc.

ITII.  1976.  Toxic and Hazardous Industrial Chemicals
Safety Manual for Handling and Disposal with Toxicity
and Hazard Data.  The International Technical Information
Institute.  Toranomon-Tachikawa Building, 6-5, 1 Chome,
Nishi-Shimbashi, Minato-ku, Tokyo, Japan.

Patty, F.  1967.  Industrial Hygiene and Toxicology.
Interscience Publishers.  New York.

Sax, N. Irving.  1979.  Dangerous Properties of Industrial
Materials.  Van Nostrand Reinhold Company, New York.

Sorbo, B.  On the Mechanism of Sulfide Oxidation in Bio-
logical Systems.  Biochem.  Biophys.  Acta 38: 349-351.

Standen, A. et. al. (editors).  1969.  Kirk-Othmer
Encyclopedia of Chemical Technology.  Interscience
Publishers.  New York.

Verschueren, K.  1977.  Handbook of Environmental Data
on Organic Chemicals.  Van Nostrand Reinhold Company,  New
York.

Yant, W. P. and H. C. Fowler.  1926.  Hydrogen Sulfide
Poisoning in the Texas Panhandle.  Rep. Invest. U.S. Bureau
of Mines.  Number 2776.

-------
                                     No. 119
      Indeno (1,2,3-aOpyTene

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL  30, 1980
          n  j-l

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                       SPECIAL NOTATION
U.S. EPA1s Carcinogen Assessment Group (CAG) has evaluated
indeno(l,2,3-c,d)pyrene and has found sufficient evidence to
indicate that this compound is carcinogenic.

-------
                          INDENO[1,2,3-cd]PYRENE




                                 Summary






     IndenoC1,2,3-cd]pyrene (IP) is a member of the polycyclic aromatic




hydrocarbon (PAH) class.  Several compounds in the PAH class are well




known to be potent animal carcinogens.  However, IP is generally regarded




as only a weak carcinogen to animals or man.  There are no reports




available concerning the chronic toxicity of IP.  Exposure to IP in the




environment occurs in conjunction with exposure to other PAH; it is not




known how these compounds may interact in human systems.




     There are no reports available concerning standard acute or chronic




toxicity tests of this chemical in aquatic organisms..

-------
I.   INTRODUCTION



     This profile is based primarily on  the  Ambient Water  Quality Criteria



Document for Polynuclear Aromatic Hydrocarbons  (U.S. EPA,  1979a)  and  the




Multimedia Health Assessment Document for Polycyclic Organic  Matter (U.S.




EPA, 1979b).




     IndenoCl,2,3-cd]pyrene (IP; 022^12) is  one of fcne  family of  polycyclic



aromatic hydrocarbons (PAH) formed as a  result of incomplete  combustion



of organic material.  Its physical and chemical properties have not been



well-characterized.



     PAH, including IP, are ubiquitous in the environment.  They  have



been identified in ambient air, food, water, soils, and sediments  (U.S.



EPA, 1979b).  The PAH class contains several potent carcinogens (e.g.,



benzCblfluoranthene), weak carcinogens (benzCa]anthracene), and cocarcinogens



(e.g.,  fluoranthene), as well as numerous non-carcinogens  (U.S. EPA,



1979b).



     PAH  which contain more than three  rings (such as IP) are relatively



stable in the environment,  and may be transported in air and  water  by



adsorption to particulate matter.  However,  biodegradation and chemical



treatment are effective in eliminating most  PAH in the environment.  The



reader is referred to the PAH Hazard Profile for a more general discussion




of PAH (U.S. EPA, 1979o).



II.  EXPOSURE



     A.   Water



          Basu and Saxena (1977, 1978) have conducted monitoring  surveys



of U.S. drinking water for the presence of six representative PAH,  including
                                                                           »


IP.  They found the average total level of the six PAH (fluoranthene,



benzolk]fluoranthene,  benzoCjlfluoranthene,   benzoCajpyrene, benzoCg,h,i]-



perylene, and indenoC1,2,3-cd]pyrene) to be  13.5 ng/1.

-------
     B.   Food




          Levels of  IP are not  routinely  monitored  in food,  but it has




been detected in foods such as  butter  and smoked  fish (U.S.  EPA,  1979a).




However, the total intake of  all  types of PAH  through the  diet has been




estimated at 1.6 to  16 ug/day (U.S.  EPA,  1979b).  The U.S. EPA (1979a)




has estimated the bioconcentration  factor of IP to  be 15,000 for  the




edible portion of fish and shellfish consumed  by  Americans.   This estimate




is based upon the octanol/water partition coefficient for  IP.



     C.   Inhalation




          There are  several studies  in which IP has been detected in




ambient air (U.S. EPA, 1979a).  Measured  concentrations ranged from 0.03




to L.34 ng/m^ (Gordon, 1976;  Gordon  and Bryan, 1973).  Thus,  the  human



daily intake of IP   by inhalation of ambient air  may be in the range  of




0.57 to 25.5 ng, assuming that  a human breathes 19  m3 of air per  day.




III. PHARMACOKINETICS




     There are no data available concerning the pharmacokinetics  of IP,




or other PAH, in humans.  Nevertheless, some experimental animal  results




were published on several other PAH, particularly benzo[a]pyrene.



     A.    Absorption




          The absorption rate of IP  in humans  or  other animals has  not



been studied.  However, it is known  (U.S.  EPA, 1979a)  that,  as a  class,




PAH are well-absorbed across  the respiratory and  gastrointestinal  epithelia




membranes.   The high lipid solubility  of  compounds  in  the PAH  class supports



this observation.

-------
     B.   Distribution



          Based on an extensive literature review, data on the distribution




of IP in mammals were not found.  However, it is known (U.S. EPA, 1979a)




that other PAH are widely distributed throughout the body following their-



absorption in experimental rodents.  Relative to other tissues, PAH tend



to localize in body fat and fatty tissues (e.g., breast).



     C.   Metabolism



          The metabolism of IP in animals or man has not been directly




studied.  However, IP, like other PAH, is most likely metabolized by the



microsomal mixed-function oxidase enzyme system in mammals (U.S.  EPA,



1979b).  Metabolic attack on one or more of the aromatic rings leads to




the formation of phenols and isomeric dihydrodiols by the intermediate



formation of reactive epoxides.  Dihydrodiols are further metabolized  by



microsomal mixed-function oxidases to yield diol epoxides, compounds



which are known to be biologically reactive intermediates for certain



PAH.   Removal of activated intermediates by conjugation with glutathione



or glucuronic acid, or by further metabolism to tetrahydrotetrols,  is  a



key step in protecting the organism from toxic interaction with cell



macromolecules.



     D.   Excretion



          The excretion of IP by mammals has not been studied.   However,



the excretion of closely related PAH is rapid, and occurs mainly  via the



feces (U.S.  EPA, 1979a).  Elimination in the bile may account for a



significant percentage of administered PAH.   It is unlikely that  PAH will



accumulate in the body as a result of chronic low-level exposures.
                                 119-7

-------
IV.  EFFECTS




     A.   Carcinogenicity




          IP is regarded as only a weak carcinogen  (U.S.  EPA,  1979b).   LaVoie




and coworkers  (1979) reported that IP had  slight activity as a tumor  initiator




and no activity as a complete carcinogen on the skin of mice which  is  known




to be highly sensitive to the effects of carcinogenic  PAH.




     B.   Mutagenicity




          LaVoie and coworkers  (1979) reported that IP gave positive results



in the Ames Salmonella assay.




     C.   Teratogenicity and Other Reproductive Effects



          There are no data available concerning the possible  teratogenicity




or other reproductive effects as a result  of exposure  to  IP.   Other related



PAH are apparently not significantly teratogenic in mammals (U.S. EPA,  l979aX.




V.   AQUATIC TOXICITY




     Pertinent information could not be located in the available literature.




VI.  EXISTING GUIDELINES AND STANDARDS




     Neither the human health nor aquatic  criteria derived by  U.S. EPA  (1979a),




which are summarized below, have not gone  through the  process  of public




review; therefore, there is a possibility  that these criteria  may be changed.




     A.   Human




          There are no established exposure criteria for  IP.   However,  PAH,



as a class, are regulated by several authorities.  The World Health Organization




(1970) has recommended that the concentration of PAH in drinking



water (measured as the total of fluoranthene, benz[g,h,i]perylene, benzCb]-



fluoranthene, benz[h]fluoranthene,  indeno[1,2,3-cd]pyrene, and benz[a]p,yrene)



not exceed 0.2 ug/1.  Occupational exposure criteria have been established
                                     X

-------
for coke oven emissions, coal tar products, and coal tar pitch volatiles,



all of which contain large amounts of PAH, including IP (U.S. EPA, 1979a).



     The U.S. EPA O979a) draft recommended criteria for PAH in water are



based upon the extrapolation of animal carcinogenicity data for benzCa]-



pyrene and dibenz[a,h]anthracene.



     B..   Aquatic



          There are no standards or guidelines concerning allowable concen-



trations of IP in aquatic environments.
                               //

-------
                           INDENO[1,2,3-cd]PYRENE

                                 REFERENCES
 Basu,  O.K.,  and J.  Saxena.   1977.  Analysis of raw and drinking water
 samples for  polynuclear aromatic hydrocarbons.  EPA P.O. No. CA-7-2999-A,
 and CA-8-2275-B.   Exposure  Evaluation Branch, HERL, Cincinnati, Ohio.

 Basu,  O.K. and J.  Saxena.   1978.  Polynuclear aromatic hydrocarbons in
 selected U.S.  drinking waters and their raw water sources.  Environ.  Sci.
 Technol., 12:   795.

 LaVoie, et al. 1979.   A comparison of the mutagenicity,  tumor initiating
 activity, and  complete carcinogenicity of polynuclear aromatic hydrocarbons
 In: "Polynuclear Aromatic  Hydrocarbons".   P.W. Jones and P.  Leber (eds.).
 Ann Arbor Science Publishers, Inc.
 Gordon,  R.J.   1976.   Distribution of airborne polycyclic aromatic hydro-
 carbons  throughout Los Angeles,  Environ.  Sci. Technol.  10:  370.

 Gordon,  R.J.  and R.J.  Bryan.   1973.   Patterns of airborne polynuclear
 hydrocarbon concentrations at four Los Angeles sites.   Environ.  Sci.  7:
 T050.

 U.S.  EPA.   1979a.   Polynuclear aromatic hydrocarbons.   Ambient water
 quality  criteria.   (Draft).

.U.S.  EPA.   1979.   Multimedia health assessment document for polycylic
 organic  matter.   Prepared under  contract  by J. Santodonato, et al., Syracuse
 Research Corp.

 U.S.  EPA.   1979.   Environmental  Criteria  and Assessment Office.   Poly-
 chlorinated Aromatic Hydrocarbon:  Hazard Profile.  (Draft).

 World  Health Organization.  1970.  European standards  for drinking water,
 Ind ed.   Revised,  Geneva.

-------
                                      No.  120
          Isobutyl Alcohol

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental  impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny  to
ensure its technical accuracy.
                           -777V

-------
                                                                      If 7


                            Isobutyl Alcohol



I.   Introduction

     Isobutyl alcohol (2-methyl-l-propanol, C,ILQ0; molecular weight

74.12) is a flaaanabla, colorless, refractive liquid with an odor  like  of

amyl alcohol,, but weaker.  Isobutyl alcohol is used in  Che manufacture of

esters for fruit flavoring essences, and as a solvent in paint and varnish

removers.  This compound is soluble in approximately 20 parts water, and is

miscible with alcohol and ether.

II.  Exposure-

     No data were readily available.

III. Pharmacokinetics

     A.   Absorption

          Isobutyl alcohol is absorbed through che intestinal cract and

the lungs.

     B.   Distribution

          No data were readily available.

     C.   Metabolism

          Isobutyl alcohol is oxidized Co isobutyraldehyde and isofaucyric

acid in che rabbit, with further metabolism proceeding  Co acetone and  carbon

dioxide.  Some conjugation with glucuronic acid occurs  in che rabbic and dog.

     D.   E
          Approximately 141 of isobutyl alcohol is excreted as urinary

conjugates in che rabbic.

IV.  Effects

     A.   Carcinogenic!£7

          Rats receiving isobutyl alcohol, either orally or subcucaneously,

one co cvo times a week for 495 co 643 days showed liver carcinomas and
                                ' // "->
                               / /'-c.

                              /ifl-  ?

-------
sarcomas, spleen sarcomas and myeloid leukemia  (Gibel, £C_ al_.,  Z.  Exp.

Chir. Chir. Forsch. 7: 235  (1974).

     B.   Teratogenicity

          No data were readily available.

     C.   Other Reproductive Effects

          No data were readily available.

     D.   Chronic Toxicity

          Ingestion of one molar solution of isobutyl alcohol in water by

rats for 4 months did not produce any inflammatory reaction of  the liver.

On ingestion of two molar solution for two months rats developed Mallory's

alcoholic hyaline bodies in the liver, and were observed  to have decreases

in fat, glycogen, and RNA in the liver.

     E.   Other Relevent Information

          Acute exposure to isobutyl alcohol causes narcotic effects, and

irritation to the eyes and throat in humans exposed to 100 ppm  for repeated

8 hour periods.  Formation of facuoles in the superficial layers of  the

cornea, and loss of appetite and weight were reported among workers  subjected,

to an. undetermined, but apparently high concentration of  isobutyl alcohol and

butyl acetate.  The oral LD_Q of isofautyl alcohol for rates if  2.46  g/kg

(Smith et. al., Arch.. Ind. Hyg. Occup. Med. 10_: 61, 1954).

V.   Aquatic Toxicity

     A.   Acute Toxicity

          The LC_- of isobutyl alcohol for 24-hour-old Daphnia  magna is

between 10-1000 mg/1.

VI.  Existing Guidelines and Standards

     OSHA   -  100 ppm
     NIOSH  -  None
     ACGIH  -   50 ppm

-------
711. Information Sources

     1.   NQi Toxicology Data Bank.
     2.   March. Index, 9th ad.
     3.   NIOSH Registry of Toxic Effects of Chemical Substances, 1978.
     4.   NCM Toxline.
     3.   Sax, I. "Dangerous Properties of Industrial Materials."
     6.   Proctor, N. and Hughes, J. " Chemical Hazards of ehe Workplace"
          Lippincott Co., 1978.
     7.   Occupational Diseases.  A Guide to Their Recognition, NIOSH
          publication Ho. 77-181, 1977.
     3.   Hunter, D.  "The Diseases of Occupations" 5th ad., Hodder and
          Stoughton, 1975.
                                 ,- >'((('
                                 ''   i r '/**

-------
                                     No.  121
               Lead


  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D,C.  20460


           APRIL 30, 1980
                       ,
            JU-I

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources/ this short profile
may not reflect  all - available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                             LEAD
                           SUiMMARY
     The hazards of human exposure to lead have been well-
recognized for centuries.  The hematopoietic system is  the
most sensitive target organ for  lead in humans, although
subtle neurobehavioral effects are suspected in children
at similar levels of exposure.   The more serious health
effects of chronic lead exposure, however,, involve neuro-
logical damage, irreversible renal damage, and adverse  repro-
ductive effects observed only at higher levels of lead  expo-
sures.  Although certain inorganic lead compounds are car-
cinogenic to some species of experimental animals, a clear
association between lead exposure and cancer development
has not been shown in human populations.
     The effects of lead on aquatic organisms have been
extensively studied, particularly in freshwater species.
As with other heavy metals, the  toxicity is strongly depen-
dent on the water hardness.  Unadjusted 96-hour LC^Q values
with the common fathead minnow,  Pimephales promelas, ranged
from 2,400-7,480 pg/l in. soft water to 487,000 pg/1 in  hard
water.  Toxicity is also dependent on the life stage of
the organism being tested.  Chronic values ranged from  32
jug/1 to 87 /jg/1 for six species  of freshwater fish.  Lead
at 500 jug/1 can reduce the rate  of photosynthesis by 50
percent in freshwater algae.  Lead is bioconcentrated by   •
all species tested - both marine and freshwater - including

-------
fish, invertebrates, and algae.  The mussel, Mytilus edulis,
concentrated lead 2,568 times that found in ambient water.
Two species of algae concentrated lead 900-1000-fold.

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                             LEAD
I.    INTRODUCTION
     This hazard profile is based primarily upon  the Ambient
Water Quality Criteria Document  for Lead  (U.S. EPA, 1979).
A number of excellent comprehensive reviews on the health
hazards of lead have also been recently published.  These
include the U.S. EPA Ambient Air Quality Criteria Document
for Lead and the lead criteria document of the National
Institute for Occupational Safety and Hearth  (1978).
     Lead (Pb, At. No. 82)  is a  soft gray acid-soluble metal
used in electroplating, metallurgy, and the manufacture
of construction materials, radiation protection devices,
plastics, electronics equipment, storage batteries, gasoline
antiknock additives, and pigments  (NIOSH, 1978).  The solu-
bility of lead compounds in water depends heavily on pH
and ranges from about 10  pg/1 at pH 5.5 to 1 /ag/1 at pH
9.0 (U.S.  EPA, 1979).  Inorganic lead compounds are most
stable in the +2 valence state,  while organolead compounds
are more stable in the -t-4 valence state (Standen, 1967) .
     Lead consumption in the United States has been fairly
stable from year to year at about 1.3 x 10  metric tons
annually.  Consumption of lead as an antiknock additive
to gasoline (20 percent annual production) is expected to
decrease steadily.  Since lead is an element, it will remain
indefinitely once released to the environment (U.S. EPA,
1979).

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II.  EXPOSURE
     A.   Water
          Lead is ubiquitous in nature, being a natural
constituent of the earth's crust.  Most natural groundwaters
have concentrations ranging from 1 to 10 ug/1.
          Lead does not move readily through stream beds
because it easily forms insoluble lead sulfate and carbonate.
Moreover, it binds tightly to organic ligands of  the dead
and living flora and fauna of stream beds.-- However, lead
has been found at high concentrations in drinking water
(i.e., as high as 1000 ug/1), due primarily to conditions
of water softness, storage, and transport  (Beattie, et al.
1972).
          The magnitude of"the problem of  excessive lead
in drinking water is not adequately known.  In one recent
survey of 969 water systems, 1.4 percent of all tap water
samples exceeded the 50 pg/1 standard (McCabe, 1970).  The
U.S. EPA (1979) has not estimated a bioconcentration factor
for lead in aquatic organisms.
     B.   Food
          It is generally believed that food constitutes
the major source of lead absorption in humans.  The daily
dietary intake of lead has been estimated  by numerous investi-
gators, and the results are generally consistent  with one
another.  This dietary intake is approximately 241 jig/day
for adults  (Nordman, 1975; Kehoe, 1961).   For children  (ages'
3 months to 3.5 years) the dietary intake  is 40 to 210 ug
of  lead per day  (Alexander, et al. 1973).

-------
     C.   Inhalation
          A great deal of controversy has been generated
regarding the contribution of air to total daily  lead  absorp-
tion.  Unlike the situation with food and water,  ambient
air lead concentrations vary greatly.   In metropolitan areas,
average air lead concentrations of 2 jjg/m , with  excursions
of 10 ug/m  in areas of heavy traffic or industrial point
sources, are not uncommon (U.S. EPA, 1979).   In non-urban
areas average air lead concentrations are ..usually on the
order of 0.1 pg/m2  (U.S. EPA, 1979).
III. PHARMACOKINETICS.
     A.   Absorption
          The classic studies of Kehoe  (1961) on  lead  metabo-
lism in man indicate that on the average and  with consider-
able day-to-day excursions, approximately eight percent
of the normal dietary lead  (including beverages)  is absorbed.
More recent studies have confirmed this conclusion  (Rabino-
witz, et al. 1974).  The gastrointestinal absorption of
lead is considerably greater in children than in  adults
(Alexander, et al.  1973; Ziegler, et al. 1978).
          It has not been possible to accurately  estimate
the extent of absorption of inhaled lead aerosols.  To vary-
ing degrees,, depending on their solubility and particle
size, lead aerosols will be absorbed across the respiratory
epithelium or cleared from the lung by mucociliary action
and subsequently swallowed.
          Very few  studies concerning dermal  absorption
of lead in man or experimental animals  are available.  A
                             I ! ! ""I j —
                          <* '/ V1 BL I

-------
recent study by Rastogi and Clausen  (1976)  indicates  that
lead is absorbed through intact skin when applied  at  high
concentrations in the form of lead acetate  or  naphthenate.
     B.   Distribution
          The general features of lead distribution  in  the
body are well known/ both from animal studies  and  from  human
autopsy data.  Under circumstances of long-term  exposure,
approximately 95 percent of the total amount of  lead  in
the body (body burden) is localized  in the  skeleton  after
attainment of maturity (U.S. EPA, 1979).  By contrast,  in
children only 72 percent is in bone  (Barry, 1975).  The
amount in bone increases with age but the amount in  soft
tissues, including blood, attains a  steady  state early  in
adulthood (Barry, 1975; Horiuchi and Takada, 1954).
          The distribution of lead at the organ  and cellular
level has been studied extensively.  In blood, lead  is  pri-
marily localized in the erythrocytes (U.S.  EPA,  1979).
The ratio of the concentration of lead in the  cell to lead
in the plasma is approximately 16:1.  Lead  crosses the  pla-
centa readily, and its concentration in the blood  of  the
newborn is quite similar to maternal blood  concentration.
     C.   Excretion
          There are wide interspecies differences  concerning
routes of excretion for lead.  In most speci.es biliary  ex-
cretion predominates in comparison to urinary  excretion,
except  in the baboon  (Eisenbud and Wrenn, 1970).   It  also
appears that urinary excretion predominates in man (Rabino-
                              x
                          - j ,Y i o ^
                          ^^^F^^^C^

-------
witz, et al. 1973).  This conclusion, however, is based
on very limited data.
IV.  EFFECTS
     A.   Carcinogenicity
          At least three studies have been published which
report dose-response data for lead-induced malignancies
in experimental animals  (Roe, et al. 1965; Van Esch, et
al. 1962; Zollinger, 1953; Azar, et al. 1973).  These studies
established that lead caused renal tumors in rats.
          Several epidemiologic studies have been conducted
on persons occupationally exposed to leaa (Dingwall-Fordyce
and Lane, 1963; Nelson, et al. 1973; Cooper and Gaffey,
1975; Cooper, 1978).  These reports do not provide a con-
sistent relationship between lead exposure and cancer develop-
ment.
     B.   Mutagenicity
          Pertinent information could not be located in
the available literature concerning mutagenicity of lead.
However, there have been conflicting reports concerning
the occurrence of chromosomal aberrations in lymphocytes
of lead-exposed workers  (O'Rioraan and Evans, ly74; Forni,
et al. 1976).
     C.   Teratogenicity
          In human populations exposed to high concentra-
tions of lead, there is evidence of embryotoxic effects
                                                             »
although no reports of teratogenesis have oeen published
(U.S. £PA, 197y).  In experimental animals,  on the other
hana, leaa has repeatedly produced teratogenic effects  (Cat-

-------
zione ana Gray, 1941; Karnofsky ana Ridgway, 195
-------
(Kline, 1960), electrocardiographic abnormalities  (Kosmider
and Pentelenz, 1962), impaired liver  function  (Dodic,  et
al. 1971), impaired thyroid function  (Sandstead, et  al.
1969) , and intestinal colic (Beritic, 1971).
V.   AQUATIC TOXICITY
     A.   Acute Toxicity
          The available data base on  the  toxic  effects of
lead to freshwater organisms is quite large and clearly
demonstrates the relative sensitivity of  freshwater  orga-
nisms to lead.  The data base shows that  the different lead
salts have similar LC^Q values, and that  LCcQ values for
lead are greatly different in hard and soft water.   Between
soft and hard water; the LC   values  varied by  a factor
of 433 times for rainbow trout, 64 times  for fathead min-
nows,  and 19 times for bluegills  (Davies, et al. 1976; Picker-
ing and Henderson, 1966).
          Some 96-hour LC^Q values for freshwater  fish are
2,400 to 7,480 pg/1 for fathead minnows in  soft water  (Tarz-
well and Henderson, 1960; Pickering and Henderson, 1966),
482,000 for  fathead minnows in hard water  (Pickering and
Henderson, 1966), 23,800 ug/1 for bluegills in  soft  water
(Pickering and Henderson, 1966), and  442,000 ug/1  for  blue-
gills in hard water  (Pickering and Henderson, 1966).
          For invertebrate species, Whitely  (1968) reported
24-hour LC5Q values of 49,000 and 27,500  ug/1 for  sludge
worms  (Tubifex sp.) obtained from tests conducted  at pH


                             /r
                             hi* «£-
                           *•/ 7 9-J

-------
levels of 6.5 and 8.5, respectively.  The effects of water
hardness on toxicity of lead to invertebrates could not
be located in the available literature.
          The acute toxicity data base for saltwater orga-
nisms is limited to static tests with invertebrate species.
The LC5Q values ranged from 2,200 to 3,600 ug/1 for oyster
larvae in a 48-hour test  (Calabrese, et al. 1973) to 27,000
ug/1 for adult soft shell clams (Eisler, 1977) in a 96-hour
test.
     B.   Chronic Toxicity
          Chronic tests in soft water have been conducted
with lead on six species of fish.  The chronic values ranged
from 32 ug/1 for la'ke trout (Sauter, et al. 1976) to 87
ug/1 for the white sucker (Sauter, et al. 1976), both being
embryo-larval tests.
          Only one invertebrate chronic test result was
found in the literature.  This test was with Daphnia magna
in soft water, and the resulting chronic value was 55 jug/1,
about one-eighth the acute value of 450 ug/1  (Biesinger
and Christensen, 1972).
          Life cycle or embryo-larval tests conducted with
lead on saltwater organisms could not be located in the
available literature.
     C.   Plant Effects
          Fifteen tests on eight different species of aqua-
tic algae are found in the literature.  Most studies mea-
                                           14
sured the lead concentration which reduced   CO- fixation
by 50 percent.  These values range from 500 ug/1 for Chlorella

-------
sp. (Monahan, 1976) to 28,000 for a diatom, Navicula  (Malan-
chuk and Gruendling, 1973).
          Pertinent data could not be located  in  the  avail-
able literature on the effects of lead on marine  algae.
     D.   Residue
          The mayfly (Ephemerella grandis) and  the  stonefly.
(Pteronarcys californica)  have been studied for their ability
to bioconcentrate lead (Nehring, 1976).  The bioconcentra-
tion factor for lead in the mayfly is 2,366 and in  the stone-
fly 86, both after 14 days of exposure.
          Schulz-Baldes  (1972) reported that mussels  (Mytilus
edulis) could bioconcentrate lead 2,568-fold.   Two  species
of algae bioconcentrate lead 933 and 1,050-fold  (Schulz-
Baldes, 1976).
VI   EXISTING GUIDELINES AND STANDARDS
     A.   Human
          As of February 1979, the U.S. Occupational  Safety
and Health Administration  has set the permissible occupa-
tional exposure limit for  lead and inorganic lead compounds
at 0.05 mg/m  of air as an 8-hour time-weighted average.
The U.S. EPA  (1979) has also established an ambient airborne
lead standard of 1.5 pg/m  .
          The U.S. EPA (1979) has derived a draft criterion
for lead of 50 jug/1 for ambient water.  This draft  criterion
is based on empirical observation of blood lead in  human
population groups consuming their normal amount of  food     ,
and water daily.
                             A

-------
     B.   Aquatic
          For lead, the draft criterion  to protect  fresh-
water aquatic life is:
               e(1.51 In  (hardness) - 3.37
as a 24-hour average, where e is the natural logarithm;
the concentration should not exceed:
               e(1.51 In  (hardness) - 1.39)
at any time  (U.S. EPA, 1979).
          For saltwater aquatic life, no draft criterion
for lead was derived.

-------
                             LEAD

                          REFERENCES
Alexander, F.W., et al. 1973.  The uptake and excretion
by children of lead and other contaminants.  Page  319  in
Proc. Int.  Symp. Environ. Health.  Aspects of Lead.   ATnster-
dam, 2-6 Oct., 1972.  Comm. Eur. Commun.  Luxembourg.

Azar, A., et al. 1973.  Review of lead studies in  animals
carried out at Haskell Laboratory - two-year feeding study
and  response to hemorrhage study.  Page 199 iji Proc. int.
Symp. Environ. Health, Aspects of Lead.  Amsterdam, 2-6
Oct., 1972.  Comm.  Eur. Commun. Luxembourg.

Barry, P.S.I. 1975.  A comparison of concentrations of lead
in human tissues.  Br.  Jour. Ind. Med. 32: 119.

Beattie, A.D., et al. 1972.  Environmental lead pollution
in an urban soft-water area. Br. Med. Jour. 2: 4901.

Beritic, T. 1971.  Lead concentration found in human blood
in association with lead colic. Arch. Environ. Health. 23:
289.

Biesinger, K.E., and G.M. Christensen.  1972.  Effect  of
various metals on survival, growth, reproduction and metabo-
lism of Daphnia magna.  Jour. Fish. Res. Board Can.  29:
1691.

Calabrese, A., et. al.  1973.  The toxicity of heavy metals
to embryos of the American oyster Crassostrea virginica.
Mar. Biol. 18: 162.

Carpenter, S.J., and V.H.. Ferm. 1977.  Embryopathic effects
of lead in the hamster.  Lab. Invest. 37: 369.

Catzione, 0., and P. Gray..1941.  Experiments on chemical
interference with the early morphogenesis of the chick.
II.  The effects of lead on the central nervous system. Jour.
Exp. Zool.  87: 71.

Chisolm, J.J. 1968.  The use of chelating agents in the
treatment of acute and chronic lead, intoxication in child-
hood.  Jour. Pediatr. 73: 1.

Chisolm, J.J., et. al. 1975.  Dose-effect and dose-response
relationships for lead in children.  Jour. Pediatr. 87:
1152.
                                                          »
Clarkson, T.W., and J.E. Kench.  1956.  Urinary excretion
of amino acids by men absorbing heavy metals. Biochem. Jour.
62:  361.

-------
Cooper, W.C. 1978.  Mortality in workers  in  lead  production
facilities and lead battery plants during  the  period  1971-
1975.  A report to International Lead  Zinc Research Organiza-
tion, Inc.

Cooper, W.C., and W.R. Gaffey. 1975.   Mortality of lead
workers. Jour. Occup. Med.  17: 100.

Cramer, K., et al. 1974.  Renal ujtrastructure renal  func-
tion and parameters of lead toxicity in workers with  dif-
ferent periods of lead exposure.  Br.  Jour.  Ind.  Med 31:.
113.

Davies, P.H., et al.  1976.  Acute and chronic toxicity
of lead to rainbow trout  (Salmo gairdneri) in  hard and soft
water.  Water Res. 10: 199.

Dingwall-Fordyce, J., and R.E. Lane. 1963.   A  follow-up
study of lead workers.  Br. Jour. Ind. Mech. 30:  313.

Dodic, S., et al. 1971.  Stanjc jetre  w pojedinih profesion-
alnih intosksikaiija In:  III Jugoslavanski  Kongres Medicine
Dela, Ljubljana, 1971.

Eisenbud, M., and M.E. Wrenn. 1970.  Radioactivity studies.
Annual Rep. NYO-30896-10. Natl. Tech.  Inf.  Serv. 1:  235.
Springfield, Va.

Eisler, R.  1977.  Acute toxicities of selected heavy metals
to the softshell clam, Mya arenaria.   Bull.  Environ.   Contain.
Toxicol.  17: 137.

Forni, A., et al.  1976.  Initial occupational exposure
to lead.  Arch. Environ. Health 31: 73.

Horiuchi, K., and I. Takada.  1954.  Studies on the indus-
trial lead poisoning.  I.  Absorption, transportation, deposi-
tion and excretion of lead.  1.  Normal limits of lead in
the blood, urine and feces among healthy Japanese urban
inhabitants.  Osaka City Med. Jour. 1: 117.

Jacquet, P., et al.  1975.  Progress report  on studies into
the  toxic action of lead in biochemistry of  the developing
brain and on cytogenetics of post-meiotic germ cells.  Eco-
nomic Community of Europe, Contract No. 080-74-7, Brussels,
Belgium.

Jacquet,  P., et al.  1977.  Cytogenetic investigations on
mice  treated with lead.  Jour. Toxicol. Environ.  Health
2:  619.

Karnofsky,  D.A.,  and L.P. Ridgway. 1952.  Production  of
 injury  to the  central  nervous system of the  chick embryo
by  lead salts.  Jour.  Pharmacol. Exp.  Therap.  104: 176.

-------
  Kehoe,  R.A.  1961.   The metabolism of lead in man in health
  and  disease.   The  Harben Lectures,  1960.   Jour.  R.  Inst.
  Publ.  Health  Hyg.   34: 1.

.  Kimmel,  C.A.,  et al.  1976.   Chronic lead  exposure:   Assess-
  ment of  developmental toxicity.   Teratology 13:  27  A (ab-
  stract) .

  Kline,  T.S.  1960.   Myocardial changes in  lead poisoning.
  AMA  Jour.  Dis.  Child. 99:  48.

  Kosmider,  S.,  and  T.  Pentelenz.  1962.  Zmiany elektro kardio-
•  grayficzne u.  starszychosol, 2.  prezwleklym zauo-dowym zatru-
  ciem olowiem.   Pol. Arch.  Med.  Wein 32:  437.

•  Lancranjan,  I., et al.  1975.  Reproductive ability of work-
  men  occupationally exposed to lead.  Arc'h.  Environ.  Health
  30:  396.

•  Lane,  R.E. 1949.  The care of the lead worker.   Br. Jour.
  Ind. Med.  6:  1243.

  Malanchuk, J.L., and  G.K.  Gruendling.  1973.   Toxicity of
  lead nitrate  to algae.  Water Air and Soil. Pollut.   2: 181.

.  McCabe,  L.J.  1970. Metal levels found in distribution sam-
  ples.   AWWA Seminar on Corrosion by Soft  Water.  Washing-
  ton, D.C.

  McClain,  R.M.,  and B.A. Becker. 1975.  Teratogenicity,
  fetal  toxicity and placental transfer of  lead nitrate in
  rats.   Toxicol. Appl. Pharmacol. 31: 72.

•  Monahan,  T.J.   1976.   Lead inhibition of  chlorophycean micro-
  algae.   Jour.  Psycol.  12:  358.

  Morgan,  B.B.,  and  J.D. Repko- 1974.  In  C. Xintaras, et
  al.  eds.  Behavioral  toxicology.  Early detection of occu-
  pational hazards.   U.S.Dep- Health Edu.  Welfare.  Washington,
  D.C.

  Nehring,  R.B.   1976.   Aquatic insects as  biological monitors
  of heavy metal pollution.   Bull. Environ. Contain. Toxicol.
  15:  147.

  Nelson,  W.C. ,  et al.  1973..  Mortality among orchard workers
  exposed  to lead arsenate spray:   a cohort study.  Jour.
  Chron.   Dis.  26: 105.

  NIOSH.   1978.   Criteria for a recommended standard.  Occupa-
  tional exposure to inorganic lead.   Revised criteria 1978.
  National Institute for Occupational Safety and  Health.
  DHEW (NIOSH)  Publication No. 78-158.

-------
Nogaki, K. 1958.  On action of lead on body of lead  refinery
workers:  Particularly conception, pregnancy and parturition
in case of females and their newborn.  Excerp.  Med. XVII.
4: 2176.

Nordman, C.N. 1975.  Environment lead exposure in Finland.
A study on selected population groups.  Ph.D. thesis.  Univer-
sity of Helsinki.

O'Riordan, M.L., and H.J. Evans.  1974.  Absence of  signifi-
cant chromosome damage in males occupationally exposed to-
lead.  Nature (Lond.) 247: 50.

Pickering, Q.H., and C. Henderson.  1966.  The acute toxicity
of some heavy metals to different species of freshwater
fishes.  Air. Water Pollut. Int. Jour. 10: 453.
                                        >.

Rabinowitz, M.B., et al. 1974.  Studies of human lead metabo-
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Perspect. Exp. Issue 7: 145.

Rastogi, S.C., and J. Clausen.  1976.  Absorption of lead
through the skin.-  Toxicol. 6: 371.

Roe, F.J.C., et al. 1965.  Failure of testosterone or xanthop-
terin to influence the induction of renal neoplasms  by lead
in rats.  Br. Jour. Cancer 19: 860.

Sandstead, H.H., et al. 1969.  Lead intoxication and the
thyroid.  Arch. Int. Med.  123: 632.

Sauter, S., et al.  1976.  Effects of exposure to heavy
metals on selected freshwater fish.  Ecol. Res. Ser. EPA
600/3-76-105.

Schulz-Baldes, M.  1972.  Toxizitat und anreicherung von
Blei bei der Miesmuschel Mytilis edulis im Laborexperiment.
Mar. Biol.  16: 266.

Schulz-Baldes, M.  1976.  Lead uptake in two marine  phyto-
plankton organisms.  Biol. Bull.  150: 118.

Standen, A., ed.  1967.  Kirk-Othmer encyclopedia of chemi-
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Stowe, H.D., and R.A. Goyer. 1971.  The reproductive ability
and progeny of F, lead-toxic rats.  Fertil. Steril.  22:
755.            i

Tarzwell, C.M., and C. Henderson.  1960.  Toxicity of less
common metals to fishes.  Ind. Wastes  5: 12.
                            \n^ o _
                           ) i J<^-

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U.S. EPA.  1979.  Lead:  Ambient Water Quality Criteria.
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Van Esch, G.J., et al. 1962.  The induction of renal tumors
by feeding basic lead acetate to rats.  Br. Jour. Cancer
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Wedeen, R.P., et al. 1975.  Occupational lead nephropathy,
Am. Jour. Hed.  59: 630.

Whitley, L.S.  1968.  The resistance of tubificid worms
to three common pollutants.  Hydrobiologia  32: 193.

Ziegler, E.E., et al. 1978.  Absorption and retension of
lead by infants.  Pediatr. Res.  12: 29.

Zollinger, H.U. 1953.  Durch Chronische Bleivergiftung Er-
zeugte Nierenadenome und Carcinoma bei Ratten und Ihre Bezie-
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poisoning and their relationship to corresponding human
neoplasma).  Virchow Arch. Pathol. Anat.  323: 694.
                               -IT

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                                      No. 122
          Maleic Anhydride

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not  reflect  all available  information  including all the
adverse health  and  environmental  impacts  presented by the
subject chemical..   This document  has undergone  scrutiny  to
ensure its technical accuracy..
                             )/li£-
                          <^ I } JJ

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                                                       151
                         MALEIC  ANHYDRIDE
SUMMARY



     Maleic anhydride is readily soluble in water where it



hydrolyzes to form maleic acid.   It is readily biodegraded by



microorganisms and is not expected to bioconcentrate.



     Maleic anhydride induced local tumors in rats following



repeated subcutaneous injections.   Maleic anhydride is an acute



irritant and can be an allergen in sensitive individuals.







I.   INTRODUCTION



     A.   Chemical Characteristics



     Maleic anhydride (C4H203; 2,5-furandione; CAS No.  108-31-6)



is a white, crystalline solid with an acrid odor.   The chemical



has the following physical/chemical properties (Windholz,  1976):







              Molecular Weight:    98.06



              Boiling Point:       202. O'C



              Melting Point:       52. 80°C



              Solubility:         Soluble in water and many



                                  organic solvents







     A review of the production range (includes importation)



statistics for maleic anhydride (CAS No.  108-31-6) which is



listed in the initial TSCA Inventory (1979a) has  shown thai
                         VJUJt-3

-------
between 200 million and 300 million pounds  of  this  chemical  were

produced/imported in 1977. _V

     Maleic anhydride is used as a chemical intermediate  in  the

production of unsaturated polyester resins,  fumaric  acid,

pesticides, and alkyd resins (Hawley, 1977).



II.  EXPOSURE

     A.   Environmental Fate

     Maleic anhydride is readily soluble  in water where it

hydrolyzes to form maleic acid  (Hawley, 1977;  Windholz, 1976).

Matsui e t_ al. (1975) reported that maleic anhydride  in wastewater

is easily biodegraded by activated sludge.

     B.   Bioconcentration

     Maleic anhydride is not expected to bioaccumulate  (U.S.  EPA,

1979b).

     C.   Environmental Occurrence

     The major source of maleic anhydride emissions  is associated

with release of the chemical as a byproduct of phthalic anhydride

manufacture.  Emissions can also occur during  the production  and

handling of maleic anhydride and its derivatives (U.S. EPA,

1976).
*/This production range information does not include any
production/importation data claimed as confidential by the
person(s) reporting for the TSCA Inventory, nor does it include
any information which would compromise Confidential Business
Information.  The data submitted for the TSCA Inventory,
including production range information, are subject to the
limitations contained in the Inventory Reporting Regulations  (40
CFR 710).

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III. PHABMACOKINETICS



     No data were found.  Nonetheless,  it  is  expected  that any



maleic anhydride that is absorbed would be hydrolyzed  to  maleic



acid and then neutralized to a maleate  salt.  Maleate  should be



readily metabolized to COj and HjO.








IV.  HEALTH EFFECTS



     A. Carcinogenicity



     Dickens (1963) reported that local fibrosarcomas  developed



in rats after repeated subcutaneous injections of maleic



anhydride suspended in arachis oil.  Multiple injections  of



arachis oil alone or a hydrolysis product derived from maleic



anhydride (sodium maleate) did not produce any tumors  at  the



injection site.



     A long term dietary study of maleic anhydride  in  rats  for



possible carcinogenicity is now in progress.  Terminal necropsies



are schedules for January, 1980 (CUT,  1979).



     B.   Other Toxicity



     Maleic anhydride vapors and dusts  are acute irritants  of the



eyes,  skin,  and upper respiratory tract (ACGIH, 1971).  Repeated



exposures to maleic anhydride concentrations above  1.25 ppm in



air have caused asthmatic responses in workers.  Allergies  have



developed in which workers have become  sensitive to even  lower



concentrations of the compound.  An increased incidence of  bron-



chitis and dermatitis has also been noted among workers with*



long-term exposure to maleic anhydride.  One case of pulmonary



edema in a worker has been reported (U.S. EPA, 1976).

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V.    AQUATIC EFFECTS




     The 24 to 96-hr median threshold limit  (TLm) for maleic




anhydride in mosquito fish is 230-240 mg/1.  The 24-hr TLm  for




bluegill sunfish is 150 mg/1  (Verschueren, 1977).








VI.   EXISTING GUIDELINES




     The existing OSHA standard for maleic anydride is an 8-hour



time weighted average (TWA) of 0.25 ppm in air  (39CFR23540).

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                            REFERENCES
American Conference of Governmental  Industrial Hygienist (1971).
Documentation of Threshold Limit  Values  for Substances in Work-
room Air, 3rd ed. , 263.

Chemical Industry  Institute  of  Toxicology (1979).   Research
Triangle Park, N. C. , Monthly Activities  Report (Nov-Dec 1979).

Dickens, F.  (1963).  Further Studies on  the Carcinogenic and
Growth-Inhibiting  Activity of Lactones and Related Substances.
3r. J. Cancer. 17(1);100.

Hawley, G. G.  (1977).  Condensed Chemical Dictionary,  9th ed.  Van
Nostrand Reinhold  Co.

Matsui, S. _et_ _al_.  (1975).  Activated sludge degradability of
organic substances in the waste water of the Kashima  petroleum
and petro chemical industrial complex in Japan.  Prog.  Water
Technol. 2:645-659

U.S. EPA  (1976).   Assessment of Maleic Anhydride as a Potential
Air Pollution Problem Vol. XI.  PB 258 363.

U.S. EPA  (1979a).  Toxic Substances  Control Act Chemical Sub-
stances Inventory, Production Statistics for Chemicals Listed on
the Non-Confidential Initial TSCA Inventory.

U.S. EPA  (1979b).  Oil and Hazardous Materials.  Technical
•Assistance Data System  (OHMTADS DATA BASE).

Verschueren,  K (1978).  Handbook  of  Environmental Data on Organic
Chemicals.   Van Nostrand Reinhold Co.

Windholz, M.  (1976).  The Merck Index, 9th Edition.   Merck and
Company,  Inc.

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                                      No. 123
           Malononitrlle

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental  impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                                  DISCLAIMER
     This  report  represents a brief  assessment of the  potential health and
environmental  hazards  from exposure  to the subject  chemical.   The informa-
tion contained in the  report is  drawn chiefly  from secondary  sources and
available  reference  documents.   Because of the limitations  of such sources,
this short profile may  not reflect all available  information  on the subject
chemical.  This document  has  undergone scrutiny to ensure  its technical ac-
curacy .

-------
                                 MALONONITRILE
                                    Summary

     Nitriles, as a group, are sources of the cyanide ion, which interferes
with basic cellular oxidative mechanisms.  Malononitrile has effects on the
cardiovascular, renal, hepatic and central nervous systems.  This compound
can take effect after inhalation, dermal contact or ingestion.  No carcino-
genic, mutagenic or teratogenic effects have been reported.
     Malononitrile has been used in the treatment of various forms of mental
illness.  A thorough documentation of the side effects of this compound
exists.  The only human toxicity data on malononitrile found in the avail-
able literature are those reported during clinical psychiatric use.

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                                 MALONONITRILE

 I.   INTRODUCTION
      Malononitrile (NCCH2CH),  CAS  registry number 109-77-3, is an odor-
 less, yellow crystalline chemical  with a molecular weight of 66.06 and a
 specific gravity  of 1.049.   Its melting point, is between 30°c and 31°C.
 Malononitrile is  soluble in  water, acetone, alcohol and ether, but is insol-
 uble in ethanol (Weast,  1974).  When heated to decomposition, nitriles emit
 toxic fumes  containing cyanides (Sax, 1968).
      Malononitrile is used in  the  following applications:  as a lubricating
 oil  additive,  for thiamine synthesis, for pteridine-type anti-cancer agent
 synthesis, and in the synthesis of photosensitizers, acrylic fibres,  and
 dyestuffs  (Eur. Chem. News,  1975; Lonza Inc., 1978).
      Imports of malononitrile, which currently is not manufactured in the
 United  States,  were 60,000 pounds for 1976 (NIOSH, 1978).
.II.   EXPOSURE
      A.  Water and Food
          Pertinent data were not found in the available literature.
      8.   Inhalation
          Research by Panov  (1969)  indicates that malononitrile was readily
 absorbed by  the lungs of animals.   As test chamber temperatures increased,
 the  mortality  rate also increased,  presumably due to higher  absorption.
          The major occupational exposure to nitriles occurs principally  by
 inhalation of  vapor or aerosols and by skin absorption.   The likelihood of
 such  exposure  increases during the handling,  transferring and quality con-
                                                                       »
 trol  sampling of these compounds.
                                 IJL3-5

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      C.   Dermal
           Panov (1969) reported that malononitrile was readily absorbed
 through the eyes of rabbits.   He also reported that mice and rabbits absorb
 the compound through the skin.  Extreme irritation resulted from both modes
 of application.
 III.  PHARMACOKINEJICS
      A.   Absorption
           Animal studies indicated that malononitrile is absorbed through
 the lungs  and by the skin (Panov,  1969).
      8.   Distribution
           Hicks (1950)  determined  that,  to  some extent,  malononitrile exerts
 tissue  specificity  (brain,  liver,  kidney, lung and thyroid)  in  its action.
           The formation of  thiocyanate  in vitro from  malononitrile and thio-
 sulfate was .highest in  the  presence of  liver tissue,  lowest  with brain, and
 intermediate with kidney  (Stem et al.,  1952).
      C-   Metabolism
           The dinitrile compounds  (such as malononitrile) presumably can ex-
 ert a greater toxic effect  than the mononitriles due  to  the  more rapid re-
 lease of cyanide from the parent compound.  Malononitrile released cyanide
j£ vivo and was ultimately excreted as thiocyanate after oxidation
 (Ghiringhelli, 1955).
          The CSN group may be converted to a carboxylic acid derivative and
ammonia, or may be  incorporated into cyanocobalamine.  Ionic cyanide also
reacts with carboxyl groups and with disulfides (McKee et al., 1962).
                                      Jt
                                   < n,ii
                                  I ) > 9*

-------
           Stern et al.  (1952)  found that in vitro respiration of brain,  kid-
 ney,  and liver slices was inhibited by  0.01 M  malononitrile.   The same in-
 vestigators  also demonstrated  the  formation of thiocyanate  from  malononi-
 trile and thiosulfate in  liver and kidney tissues in  vitro.   The release of
 cyanide  from dinitriles suggests that their mechanism of acute toxicity  may
 be similar to  that of the mononitriles.
           The  enzyme  rhodanase,. which catalyzed the formation  of thiocyanate
 from  cyanide and thiosulfate,  was  ineffective  in  the  catalysis of thiocya-
 nate  from malononitrile.  In vivo  thiocyanate  formation apparently came  from
 an intermediate  metabolite and not the malononitrile  molecule.
      D.    Excretion
           After  absorption, malononitrile may be metaboilized  to an organic
cyanide, which is  oxidized to  thiocyanate and excreted in the urine (McKee
et al, 1962).  No  evidence of  respiratory excretion was found in the avail-
able  literature.
IV.   EFFECTS
      A.    Carcinogenic!ty, Mutagenicity, Teratogenicity and Reproductive
           Effects
           Pertinent data were not  found in the available literature.
      B.    Chronic  Toxicity
           The only available human toxicity data on malononitrile are those
reported during  the clinical use of the compound in the treatment of mental
illness.
           Hyden and Hartelius  (1948) reported on the clinical use of malo-
nonitrile during psychiatric treatment.   Its intended purpose was to stimu-
late  the production of proteins and nucleic acids in the pyramidal cells, of
the frontal cortices of psychiatric patients, particularly  those who were
depressed or schizophrenic.   All patients experienced tachycardia 10 to 20

-------
 minutes after the infusion of malononitrile (1-6 mg/kg).  Facial redness,
 headache, nausea, vomiting, shivering, cold hands and feet, muscle spasms
 and numbness were also reported with varying frequency.  Similar results
 were also submitted by MacKinnon et al.  (1949),  Hartelius (1950), and Meyers
 et al.  (1950) in the treatment of mental patients.
           Hicks (1950) reported that malononitrile  poisoning induced brain
 lesions in rats.  The compound produced  demyelinating lesions of the optic
 tract and nerve, the cerebral cortex,  the olfactory bulb and the substantia
 nigra.
           Panov (1969)  found  the repeated exposure  to malononitrile (36
 mg/m5 for 2 hours per day  for 35 days) was slightly toxic to rats-   The
 exposure  caused slight anaplasia of bone.marrow,  i.e.  a  lower hemoglobin
 level and elevated reticulocyte  count.
      F.    Acute Toxicity
           Panov (1969)  subjected mice to  a single,  2-hour inhalation expo-
 sure  to malononitrile.  The mice showed signs of restlessness  and increased
 respiration  rate in the early post-treatment period  followed by Lassitude,
decreased  respiration rate, cyanosis, noncoordination of movement, tremb-
ling, convulsions and eventual death of some animals.  The exposure concen-
tration was not  noted.
          Panov  (1969) reported that liquified malononitrile applied to the
eyes of rabbits caused tearing, blepharospasm. hyperemia of the conjunctiva,
and swelling of  the eyelids.  Panov also applied malononitrile solution
(concentration not stated) to the tails of mice.   The -animals showed signs
of restlessness, rapid respiration and slight cyanosis of the extremities
                                                                        »
and the mucosa of the lips.  He also observed trembling and skin irritation
following dermal application of malononitrile to  a rabbit.

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          Nuclear changes in neurons and  satellite  spiral  ganglia were  seen
in rats administered single doses  (6-8 mg/kg) of malononitrile  (Van Breeman
and Hiraoka, 1961).
V.   AQUATIC TOXICITY
     Pertinent data were not found in the available literature.
VI.  EXISTING GUIDELINES AND STANDARDS
     A.   Human
          Because malononitrile is about three times as toxic as  isobutyro-
nitrile, NIOSH recommends that employee exposure to malononitrile not exceed
3 ppm (8 mg/nv5) as a TWA limit for up to 10-hour workshift in a 40-hour
work week (NIOSH, 1978).
     B,   Aquatic
          Pertinent data were not found in the available literature.

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                                 MALONONITRILE

                                   References


Eur.  Chem.  News.  1975.   Lonza develops malononitrile  process  for wide  ap-
plication.  March 15, 1975.

Ghiringhelli,  L.  1955.   Toxicity of  adipic nitrile—Clinical  picture  and
mechanism of poisoning.  Med. Lav.  46: 221.

Hartelius, H.   1950.  Further experiences in  the use of malononitrile in  the
treatment of mental illnesses.  Am. Jour. Psychiatry.  107: 95.

Hicks,  S.P.   1950.   Brain metabolism  in vivo—II.  The  distribution of  le-
sions  caused  by azide  malononitrile,  plasmocid and  dinitrophenol poisoning
in rats.  Arch. Pathol.  50: 545.

Hyden,   H.,   and   H.   Hartelius.    1948.    Stimulation  of   the  nucleo-
protein-production in the nerve  cells  by  malononitrile  and its  effect  on
psychic  functions in mental  disorders.   Acta.   Psychiatr.  Neurol.  Suppl.
48: 1-

Lanza, Inc.  1978.  Malononitrile—Production Information.  Fairlawn, NJ.

MacKinnon, I.H., et al_  1949.  The use  of malononitrile in the treatment of
mental illness.  Am.  Jour. Psychiatry.  105: 686.

McKee, H.C., et  al.   1962.  Acetonitrile in body  fluids  related to smoking.
Public Health Rep.  77: 553.

Meyers,  0.,  et  al.   1950.  Effect of  malononitrile on  physical  and mental
status of schizophrenic patients.   Arch. Neurol. Psychiatry.  63: 586.

National Institute for  Occupational  Safety and Health.   1978.   Criteria for
a  recommended  standard...occupational  exposure  to  nitriles.    U.S.  DHEW
(NIOSH) Report No. 78-212.

Panov,  I.K.    1969.   Study  of  acute  dicyanomethane toxicity  in  animals.
Jour. Eur. Toxicol.   2:  292.

Sax, N.I.  1968.  Dangerous Properties of Industrial  Materials,  3rd ed.   Van
Nostrand Reinhold Co., New York.

Stem, J., et  al.  1952.   The effects and the  fate  of malononitrile and re-
lated compounds in animal tissues.  Biochem.  Jour.   52:  114.

Van Breeman, V.L. and J. Hiraoka.  1961(abst.)  Ultra structure  of nerve and
satellite cells  in spinal ganglia of  rats  treated with  malononitrile.   Am.
Zool.  1: 473.

Weast, R.C.  (ed.)  1974.   CRC Handbook  of  Chemistry and Physics  —  A Ready
Reference    Book   of     Chemical    and    Physical    Data,    54th    ed.
                                           -I H

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                                       No.  124
              Mercury

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, B.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy. .

-------
                             MERCURY



                             SUMMARY





      Short chain alkyl mercurials represent a toxic species



that distributes widely and accumulates in the liver, kidneys



and other organs.  These compounds are eliminated from the body



at a slow rate.  In humans, mercurials have been associated with



neurological disorders, sensory impairment and tremors.  Prenatal



exposure has produced psychomotor disorders.v  Brain development



is impaired by accumulation of mercurials, and lesions in the



cerebral and cerebellar areas have been observed.



      Methylmercury crosses the placental barrier and is secreted



in milk.  Methylmercury and mercuric chloride have been shown



to produce teratogenic effects in animals.  Reproductive effects



in animals of alkyl mercury compounds involve reversible inhibi-



tion of spermatogonia and damage to unfertilized gametes.  A



high infant mortality rate has been reported in a study of mothers



exposed to high levels of mercurials.



      Mercurials have induced chromosome breakage in plant cells



and point mutations in Drosophila.  Mercurials have not been



shown to produce carcinogenic effects other than non-specific



injection site sarcomas.  The U.S. EPA (1979)  has calculated



an Acceptable Daily Intake (ADI)  for mercury of 200 pg/day.



      Mercury can be bioconcentrated many-fold in fish and other



aquatic organisms because of rapid uptake and the excretion of

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mercury from their tissues.  In general, the methylmercury com-
pounds are more toxic than the inorganic forms of mercury.  Toxi-
city varies widely among species.  Concentrations as low as 0.1
ug/1 have been shown to be toxic to freshwater crayfish.

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                             MERCURY
I.     INTRODUCTION
      This profile is based on the Ambient Water Quality Criteria
Document for Mercury  (U.S. EPA, 1979).
      Mercury  (Hg; atomic weight 200.59) is a silver-white metal,
which is a liquid at room temperature.  It has the following
physical properties:  melting point, -38.87°C; boiling point, •
356-358°C; specific gravity, 13.546; and vapor pressure at 20°C,
0.0012 mm Hg (Stecher, 1968).
      Mercury exists in three oxidation states:  elemental (0),
mercurous (+1)  , and mercuric (+2).  The solubilities of some
common mercuric salts are as follows: HgCl2 (1 g/13.5 ml water),
Hg(N03)2 (soluble in a "small amount" of water), Hg(CH3COO)2
(1 g/2.5 ml water) (Stecher, 1968).  Mercurous salts are much
less soluble in water; Hg2Cl2 is practically insoluble in water
(Stecher, 1968).
      Major usage of mercury include the following:  as a cathode
in the electrolytic preparation of chlorine and caustic soda,
in electrical apparatus, in industrial and control instruments,
in general laboratory applications, in dental amalgams, in anti-
fouling and mildew-proofing paints, and as a fungicide in treat-
ing seeds, bulbs, and plants.  However, mercury is no longer
registered by the U.S.. EPA for this last application.
      Elemental mercury can be oxidized to the mercuric form
in water in the presence of oxygen (Stock and Cucuel, 1934);
this transformation in water is facilitated by the presence of
organic substances (Jensen and Jernelov, 1972).  The mercuric

-------
ion is a substrate for biomethylation reactions; both dimethyl
and monomethyl mercury may be formed by bacteria present  in sedi-
ments (Wood, 1976 and Cotton and Wilkinson, 1966).  Considerable
bacterial demethylation of methylmercury occurs in the environ-
ment, limiting the buildup of methylmercury (Tonomura and Konzaki,
1969).  The degree of oxygenation, pH, and the presence of inor-
ganic and organic ligands are determining factors regulating
which state of mercury is present in water.  On thermodynamic
grounds, one would expect inorganic mercury to be present mainly
as mercuric compounds in well-oxygenated water and, in an increas-
ing fraction of total mercury, as the elemental form or the sul-
fide form under reducing conditions (NAS, 1978).
II.   EXPOSURE
      Mercury undergoes a global cycle of emission and deposi-
tion.  Total entry of mercury into the atmosphere is approximately
40,000 to 50,000 metric tons per year, mainly from natural sources
(NAS, 1978 and Korringa and Hagel, 1974).  Deposition from the
atmosphere into the ocean is estimated at about 11,000 tons per
year (NAS, 1978).  These waters represent a relatively large
mercury pool that maintains a stable concentration (U.S. EPA,
1979).
      Industrial release of mercury involves both organic and
inorganic forms.  These emissions are from the burning of fossil
fuels, discharges of waste from the chloralkali industries, dis-
charges of methylmercury from chemical manufacturers, and runoff
from the use of ethyl and methylmercury fungicides (U.S. EPA, •
1979) .

-------
      Based on available monitoring data, the U.S. EPA  (1979)

has estimated the uptake of mercury by adult humans  from air,

water, and food:
                       Adult - ug/day
Source
Air
Water
Food
Minimum
0.3
0.1
3.0
Maximum
0.8
0.4
5.0
                        Predominant form
                                                  elemental
                                                  mercuric
                                               methylmercury
          Total
3.4
6.2
      Fish and shellfish represent a source of high methylmercury

intake.  The U.S. EPA  (1979) has estimated average bioconcen-

tration factors of 1,700 for mercuric chloride and 6,200 for

methylmercury in the edible portions of fish and shellfish con-

sumed by Americans.  This estimate is based on bioconcentration

studies in several species, and on other factors.

III.  PHARMACOKINETICS

      A.   Absorption

           Inorganic mercury salts are absorbed poorly by the

human gastrointestinal tract; less than 15 percent absorption

was reported (Rahola, et al., 1971).  Inhalation of mercuric

oxide has been shown to produce pulmonary deposition and absorp-

tion of the compound, with 45 percent of the administered dose

cleared within 24 hours (Morrow, et al., 1964).  Dermal absorp-

tion of mercuric chloride has been reported in studies with guinea

pigs (Friberg,  et al., 1961; Skog and Vahlberg", 1964).

           Metallic mercury is not absorbed significantly from

the gastrointestinal tract.  Friberg and Nordberg  (1973)  calculate

that less than 0.01 percent of an orally administered- dose is

absorbed.  Studies with human subjects reveal approximately 80


                                X

-------
percent of inhaled mercury vapor is retained  (Hursh, et  al.,
1976}, with alveolar regions indicated as the probable site of
absorption into the bloodstream  (Berlin, et al., 1969).  Animal
studies indicate dermal absorption of metallic mercury can occur
(Juliusberg,  1901; Schamberg, et al., 1918).
           Methylraercury shows virtually complete absorption
from the gastrointestinal tract  (Aberg, et al., 1969; Miet-
tinen, 1973).  Inhalation of alkyl mercurials leads to high
retention, perhaps as high as 80 percent (Task Group on
Metal Accumulation, 1973).  Severe poisoning of humans follow-
ing topical methylmercury applications indicates some dermal
absorption of the compound (U.S. EPA, 1979).
      B.   Distribution
           Methylmercury, after absorption from the gastrointes-
tinal tract,  distributes readily to all tissues in the body (WHO
Expert Committee, 1976), with the highest concentrations being
found in the kidney cortex and red blood cells.  Approximately
five percent of an ingested dose is found in the blood compart-
ment following tissue distribution.  Human studies with  a radio-
actively labeled compound have indicated that approximately ten
percent of the body burden may be transferred to the head region
following complete tissue distribution (Aberg, et al., 1969).
The ratio of methylmercury in the brain to levels in the blood
may be as high as 10:1 (U.S.  EPA, 1979).   In muscle tissue, an-
alysis of the mercury present indicates that it is almost entirely
methylmercury, while liver and kidney contain a substantial amount
of demethylated, inorganic forms (Magos,  et al., 1976).

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           Determination of methylmercury  in cord  blood  and  fetal
red cells indicates that the compound  is transported  across  pla-
cental membranes  (Tejing, 1970; Suzuki, et al.,  1971).   Methyl-
mercury is secreted in mother's milk and may average  as  much
as five percent of the maternal blood  level  (Bakir, et al.,  1973).
           Mercury in the mercuric form concentrates  in  the  kid-
neys following inhalation of mercury vapor.  Animal studies  show
that up to 90 percent of an administered dose  may  localize at
this site (Rothstein and Hayes, 1964).  Experiments using  radio-
labeled mercury in human volunteers have shown approximately
seven percent accumulation of the inhaled compound in the  head
region (Hursch, et al., 1976).  Oxidation of absorbed elemental
mercury to the mercuric form takes place _iri vivo,  probably largely
through the enzymatic activities of red blood  cells (Clarkson,
et al., 1978).
           Mer.cury has been shown to be transferred into the
fetus after maternal exposure.  The rate of transfer  of  elemental
mercury appears to be greater than ionic forms of  mercury  (Clark-
son, et al., 1972).
           Animal studies with inorganic mercury salts indicate
the distribution pattern is similar to the pattern observed  after
exposure to mercury vapors (Friberg and Vostal,  1972); however,
the ratio of mercuric ion in red cells to plasma levels  is lower
(Rahola, et al., 1971)..  The major site of mercuric ion  accumula-
tion is the kidney (U.S. EPA, 1979).
      C.   Metabolism
                                                              »
           Methylmercury undergoes cleavage of the carbon mercury
bond, resulting in the production of inorganic mercury in vivo.

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Plasma, liver, and kidney all contain substantial  amounts  of
inorganic mercury following administration of  the  organic  form
of the compound  (Bakir, et al.,  1973).  Norseth and Clarkson
(1971) have suggested that gut microflora may  aid  in this  bio-
transformation.  Bakir, et al.  (1973) have determined  a mean
half-life value of 65 days for 16 hospital cases.  However, a
wide range of blood half-lives have been determined in human
studies (U.S. EPA, 1979).  Whole body half-life values for methyl-
mercury appear to be in the same range  (-<-52^93 days)  as blood
clearance half-lives (Miettinen, 1973).
           Elemental mercury can undergo oxidation in  the  body
to the mercuric form, which is then capable of interacting with
many tissue ligands  CClarkson, et al.,  1978).  Limited experi-
ments with subjects exposed to mercury  vapor indicate  a two com-
ponent loss of mercury from the bloodstream.  Clarkson (1978)
has estimated half-lives of 2.4 days for the fast component and
14.9 days for the slow component following a brief exposure to
mercury vapor.  Hursh,  et al. (1976) have estimated that the
whole body half-life of elemental mercury is comparable with
that of methyl mercury.
      D.   Excretion
           The excretion of methylmercury occurs predominantly
by the fecal route in humans.  Less than ten percent of excretion
occurs in the urine  (U.S. EPA, 1979).   Norseth and Clarkson (1971)
have determined significant biliary secretion of methylmercury
in animals, raising the possibility that biotransforraation to»
the inorganic form might be affected by microflora in  the gut.

-------
           Elemental mercury exposure has been shown  to  lead
to mercury excretion predominantly through the feces  and  urine
(Lovejoy, et al., 1974).  As kidney levels of mercury  increase,
a greater urinary excretion of the compound occurs  (Rothstein
and Hayes, 1964).  Urinary excretion values from 13 percent to
58 percent have been determined.  Elimination of inhaled  mercury
has been observed in expired air  (7 percent)  (Cherian, et al.,
1978)  and in sweat (Lovejoy, et al., 1974).
           Human studies with small ingested*, doses of mercuric
salts have indicated that following excretion of the  unabsorbed
compound, urinary and fecal excretion of inorganic mercury were
approximately equal (Rahola, et al., 1971).
IV.   EFFECTS
      A..   Carcinogenicity
           Intraperitoneal injection of metallic mercury  into
rats produced injection site sarcomas (Druckrey, et al.,  1957).
           Pertinent data could not be located in the available
literature indicating that mercury is carcinogenic.
      B.   Mutagenicity
           Methylmercury has been shown to block mitosis  in plant
cells and in human leukocytes treated in vivo, and human  cells
in vitro, as well as to induce chromosome breakage in plant cells
and point mutations in Drosophila (Swedish Expert Group,  1971;
Ramel, 1972).
           No evidence for the mutagenic effects of elemental
or inorganic mercury could be located in the available literature.

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      C.   Teratogenicity
           Oharazawa  (1968) reported increased frequency of cleft
palate in mice treated with an alkyl mercury compound.  Embryo-
toxic effects without gross teratological effects were reported
by Fujita (1969)  in mice.  Prenatal exposure to methylmercury
has produced histological evidence of brain damage in several
species (Matsumoto, et al., 1967; Nonaka, 1969; Morikawa, 1961).
Spyker and Smithburg  (1972) and Olson and Massaro (1977) have
also reported anatomical malformations in animals exposed pre-
natally to methylmercury.
           Teratological effects of mercuric chloride have been
reported in animals (Gale and Perm, 1971).  However, data are
not available on the  teratogenicity of inorganic mercury in human
populations.
          • Exposure of rats prenatally to mercury vapor produced
fetal toxicity without evidence of teratological effects (Baranski
and Szymczyk, 1973).
      D.   Other Reproductive Effects
           A high mortality rate in infants born to women suffer-
ing mercury poisoning has been reported  (Baranski and Szymczyk,
1973) .
           Methylmercury has been reported to interfere with
reproductive capability in adult animals treated with this com-
pound (Ramel, 1972; Suter, 1975).  Khera (1973) has observed
that administration of alkyl mercury compounds to rats may damage
gametes prior to fertilization.  Reversible inhibition of spesma-
togonial cells in mice treated with mercuric chloride has been
reported (Lee and Dixon, 1975).

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      E.   Chronic Toxicity



           Chronic exposure to methylmercury has produced several



outbreaks of poisoning, characterized by neurological symptoms



following central nervous system damage (Nordberg, 1976; NAS,



1978).   Adult exposure to methylmercury has produced symptoms



of paresthesia of the extremities, impaired peripheral vision,



slurred speech, and unsteadiness of gait and of limbs (U.S.  EPA,



1979).   Neuropathological investigation showed cerebellar atrophy



and focal atrophy of the calcarine cortex (Hunter and Russell,



1954) .



           Prenatal exposure to methylmercury produced psycho-



motor brain abnormalities (Engleson and Herner, 1952; Harada,



1968) .   Brain development was shown to be disturbed, and both



cerebral and cerebellar lesions were observed (U.S. EPA, 1979).



An epidemiological study on school children in the Minamata  Bay



area has reported a higher incidence of neurological deficits,



learning difficulties, neurological symptoms, and poor performance



on intelligence tests for these residents of a high methylmercury



exposure region (Med.  Tribune, 1978).



           An ethylmercury poisoning outbreak indicated renal



and cardiac damage following this exposure (Jalili and Abbasi,



1961).



           Mercury vapor poisoning may produce signs of mental .



disturbances, tremors, and gingivitis (U.S.  EPA,  1979).   Exposure



to extremely high concentrations can damage lung  tissue causing



acute mercurial pneumonitis.  Kidney dysfunction  (proteinuria)
                                                             »


in workers exposed to mercury vapor has also been reported  (Kazantzis,



et al., 1962; Joselow and Goldwater, 1967).
                             ^ . >i / 4 -
                            **) } v -> *

                            /a/-/3

-------
V..    AQUATIC TOXICITY
      A.   Acute Toxicity
           Observed LC^Q values for three  flow-through  and  two
static-renewal assays for mercuric chloride with  the  rainbow
trout as the test species ranged from 155  to 903  ug/1.   The re-
sults of two flow-through and three static-renewal  assays on
rainbow and brook trout provide an LC50 range  for methylmercuric
compounds from 24 to 84 ug/1, with the rainbow trout  being  from
three to five times as sensitive as the brook  trout.  For five
other mercury compounds, LC5Q values ranged from  5.1  for phenyl-
mercuric acetate to 39,910 ug/1 for merthiolate.  Ethyl- and
phenylmercury compounds generally were more toxic while merthio-
late and pyridylmercuric acetate were less toxic.   A  total  of
14 freshwater invertebrate species have been tested in  static
and static-renewal bioassays for acute toxicity to  mercuric chloride
and mercuric nitrate.  LCCQ values ranged  from 0.02 to  2,100
ug/1 (U.S. EPA, 1979).  Heit and Pingerman (1977) and Beisinger
and Christensen (1972) reported the more sensitive  species  to
be the crayfish Faxonella clypeata and the daphnid, Daphnia magna,
respectively.  Warnick and Bell (1969) reported that  the mayfly
(Ephemerella subvaria), the stonefly  (Acroneuria  lycorius),  and
the caddisfly (Hydropsyche betteni) were among the  most resistant
freshwater invertebrates to mercuric chloride.  Two static  tests
have produced 96-hour LC5Q values of 800 and 2,000  ug/1 for  mer-
curic chloride to the marine fish, the mummichog  (Fundulus  heter-
clitus).  Among marine invertebrates exposed to mercuric chloride,
LCgg values ranged from 3.6 to 32,000 pg/1 for 21 species.   Embryo
                                itf

-------
stages of the oyster (Crassostrea virginica),  the hard-shell
clam (Mercenaria mercenaria),  and the mysid shrimp (Mysidopsis
bahia),  the latter in the only acute flow-through test reported,
were the more sensitive species reported.  Lockwood and Inman
(1975)  provide the only acute  study for methylmercuric chloride
with a adjusted 96-hour LC5Q value of 150 ug/1.
      B.   Chronic Toxicity
           McKim, et al. (1976)  offered the single source reported
for chronic effects to freshwater fish.  Examining the long-term
effects of methylmercury chloride on three generations of the
brook trout (Salvelinus fontinalis), adverse effects  were reported
at 0.93 ug/1,.but not at 0.29  pg/1.   Brook trout were from three
to four times more resistant than rainbow trout (Salmo gairderi).
Sosnowski, et al. (1979) have  examined the effects of mercuric
chloride by a flow-through, life-cycle bioassay on the mysid
shrimp,  Mysidppsis bahia.  The highest concentration  producing
no-observed-effect was 0.82 ug/1.
      C.   Plant Effects
           A number of different parameters have been used to
determine the toxic effects of mercury compounds on freshwater
plants.   Effective concentrations .of mercuric  chloride ranged
from 60 to 2,590 pg/1.  Blinn, et al.  (1977)  demonstrated altered
photosynthetic activity in a summer assemblage of algal species
at 60 pg/1.  Two of these studies on the effects of methylmercury
chloride to freshwater algae revealed enzyme inhibition at 1,598
pg/1 in Anklstrodesmus braunii' and 50 percent  growth  inhibition
to Coelastrum microporum at concentrations of  2.4 to  4.8 pg/1.
For other organomercury compounds, effective concentrations ranged

                                if
                             ^tii tS-
                            ^^^^^^^^^
                            /ay-AT

-------
 from  less  than  0.6  to  200.6  ug/1.   Using 18  marine species,  Ber-


 land, et al.  (1976) measured growth inhibition at mercuric chloride

 concentrations  from 5  to  15  pg/1 and lethalities from 10 to 50

 ug/1.  Effective  concentrations  for the alga Isochrysis galbana

 ranged up  to  2,000,000 jig/1, at  which no growth was observed

 (Davies, 1976).   For other organomercury compounds, effective

 concentrations  ranged  from 0.1  to less than  2,000 ug/1.  Harriss,

 et  al.  (1970) reported reduced  photosynthetic activity to methyl-

 mercury hexachlorophthalimine in the diatom,. Nitzchia delictissima,
       .                              ,   *
 at  the level  of 0.1 ug/1.  Methylmercury chloride was reported

 by  Overnell  (1975)  to  reduce photosynthetic  activity at concen-

 trations of less  than  2,000  ug/1.

      D.   Residues

           Bioconcentratlon  data for freshwater species for  various

 mercury compounds can  be  summarized by the following.bioconcen-

 tration factors':   33,800  for the algae Synedra ulna (Fujita  and

 Hashizuma, 1972)  exposed  to  mercuric chloride; 4,532 to 8,049

 for juvenile  rainbow trout exposed to methylmercury chloride

 (Reinert,  et  al.,  1974);  12,000  to 20,000 for brook trout exposed

 to  methylmercury  chloride  (McKin,  et al., 1976); and 62,898  for

 the fathead minnow exposed to methylmercury  chloride (Olson,

 et  al., 1975).  It should be noted that for  the high bioconcen-

 tration value for the  fathead minnow, the fish were allowed  to .

 forage on  aquatic organisms  growing within the mercury enriched

 exposure chambers;  therefore, this measurement may more closely

 reflect actual  field data.   The  trout were fed a pelleted diet.

 A variety  of  marine organisms have been used to demonstrate  the

'rapid accumulation of  inorganic  and organic  mercury compounds.

-------
Bioconcentration values for marine algae ranged from 853 to 7,400,

with exposure periods of two to eight days for mercuric chloride.

A 30-day bioconcentration factor of 129 for the lobster, Homarus

americanus, has been reported by Thurberg, et al.  (1977), and

a range of 2,800 to 10,000 reported for adult oysters, Crassostrea

virginica, (both species for mercuric chloride).  Kopfler  (1974)

reports a biomagnification value of 40,000 for the oyster C.

virginica to methylmercury and phenyl-mercury chloride.  The

biological half-lives of rapidly accumulated., mercuric compounds

indicate that clearance is not rapid even after several months.

VI.   EXISTING GUIDELINES

      A.   Human

           The U.S. EPA has recommended a drinking water standard

of 2 ug Hg/1 to protect human health (U.S. EPA, 1973).

           Calculation of an acceptable daily intake (ADI) of

mercury by the U.S. EPA (1979)  has produced a tentative criterion

of 0.2 pg/1 (with an uncertainty factor applied) for ambient

water.

      B.   Aquatic

           The criteria for mercury are divided into tentative

recommendations for inorganic and organic mercury.  Freshwater

criteria have been drafted as follows:  for inorganic mercury,

the draft criterion is 0.064 pg/1 for a 24-hour average exposure,

not to exceed 3.2 pg/1 at any time.  For methylmercury, the draft

criterion is 0.016 pg/1 for a 24-hour average, not to exceed
                                                             •
8.8 pg/1 at any time.  To protect marine life from inorganic

mercury, the draft criterion is 0.19 pg/1 for a 24-hour average,

not to exceed 1.0 pg/1 at any time.  For methylmercury, the tenta-

-------
tive criterion is 0.025 ug/1 as a 24-hour average not  to  exceed



2.6 pg/1 at any time (U.S. EPA, 1979).



           The above criteria have not yet gone through the  pro-



cess of public review;  therefore, there is a possibility  that '



the criteria may be changed.

-------
                           MERCURY

                          REFERENCES

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Bakir, F., et al. 1973.  Methyl mercury poisoning in  Iraq.
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Baranski, B., and I. Szymczyk.  1973,  Effects of mercury
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Beisinger, K.E., and G.M. Christensen.  1972.  Effects of
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Berland, 3.R., et al.  1976.  Action toxique de quarte metaux
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Berlin, M.H. , et al.  1969..  On the site and mechanism of
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Blinn, D.W., et al.  1977.  Mercury inhibition on primary
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Clarkson, T.W.  1978.  Unpublished data.  Environ. Health
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Clarkson, T.W., et al. 1972.  The transport of elemental
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Clarkson, T.W., et al. 1978.  The metabolism of inhaled
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Cotton, F.A., and G. Wilkinson.  1966.  Advanced inorganic
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Davies, A.G.  1976.  An assessment of the basis of mercury
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Assn. U.K. 56": J9.

-------
Druckrey, H./ et al. 1957.  Carcinogenic action of metallic
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Engleson, G., and T. Herner.  1952.  Alkyl mercury poison-
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Friberg, L.', and F. Nordberg.  1973.  Inorganic mercury-
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Friberg, L., and J. Vostal, eds.  1972.  Mercury  in  the
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Friberg, L., et al. 1961.  Resorption of mercuric chloride
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normal skin and through skin pre-tested with  acetone,  alkyl-
arylsulphonate and soap.  Acta. Derm. Venerol.  41:  40.

Fujita, E.  1969.  Experimental studies on organic mercury
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either placenta or breast milk.  Jour.  Kumamoro Med.  Soc.
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Fujita, M., and K. Hashizuma.  1972.  The accumulation of
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Gale, T., and V. Ferm. 1971. Embryopathic effects of mer-
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Harada, 7..C. 1968.  Clinical investigations on Minamata
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Harriss, R.C., et al.  1970.  Mercury compounds reduce photo-
synthesis by plankton.  Science  170: 736.

Heit, M., and M. Fingerman.  1977.  The influence of size,
sex and temperature on the toxicity of mercury to two  species
of crayfishes.  Bull. Environ. Contain. Toxicol.  13: 572.

Hunter, D., and D.S. Russell. 1954. Focal cerebral and cere-
bellar atrophy in a human subject due to organic mercury
compounds. Jour. Neurol. Neurosurg. Psychiatry  17:  253.

Hursh, J.3., et al.  1976.  Clearance of mercury  (197Hg,
203rtg) vapor inhaled by human subjects.  Arch. Environ.
Health  4: 302.
                              16

-------
Jalili, M.A., and A.H. Abbasi.  1961.  Poisoning  by  ethyl
mercury toluene sulphonanilide.  Br. Jour.  Ind. Med.   18:
303.

Jensen, S., and A. Jernelov.  1972.  Behavior of  mercury
in the environment. Page 43.  iri Mercury contamination  in
man and his environment. Vienna Int. Atomic Energy Agency.
Tech. Rep. Ser. 137.

Joselow, M.M., and L.J. Goldwater.  1967.   Absorption  and
excretion of mercury"  in man.  XII.  Relationship  between
urinary mercury and proteinuria.  Arch. Environ.  Health
15: 155.

Juliusberg, F.  1901.  Experimentelle untersuchungem uber
quicksilber-resorption bei der schmierkur.  Arch. Derm.
Syph.  56: 5.

Kazantzis, G., et al.  1962.  Albuminuria and the nephrotic
syndrome following exposure to mercury and  its compound.
Q. Jour. Med. 31: 403.

Khera, K.S. 1973. Reproductive capability of male rats and
mice treated with methyl mercury.  Toxicol. Appl. Pharmacol.
24: 167.

Kopfler, F.C.  1974.  The accumulation of organic and  inor-
ganic mercury compounds by the eastern oyster  (Crassgstrea
virginica) .  Bull. Environ. Contain. Toxicol.  111 275.

Korringa, P., and P.  Hagel.   1974.  Ln.  Proc.  International
symposium on problems of contamination of man and his  environ-
ment by mercury and cadmium.  Comm. Eur.  Commun., Luxembourg
July 3-5, 1973.

Lee, I.D., and R.L. Dixon.  1975.  Effects  of mercury  on
spermatogenesis studied by velocity sedimentation, cell
separation and serial mating. Jour. Pharmacol Exp. Ther.
194: 171.

Lockwood, A.P.M., and C.B.E.  Inman.  1975.  Diuresis in
the amphipod, Gammarus duebeni induced by methylmercury,
D.D.T., lindane and fenithcothien.  Comp. Biochem. Physiol.
52C: 75.

Lovejoy, H.B. , et al.  1974.  Mercury expos.ure evaluations
and their correlation with urine mercury excretion.  Jour.
Occup. Med.  15: 590.

Magos, L., et al. 1976. Tissue levels of mercury  in  autopsy
specimens of liver and kidney. Page 93. in  WHO Conf.  on   '
intoxication due to alkyl mercury treatea~seed.   Baghdad
Nov. 9-13, 1974. Geneva, WHO  11  (Suppl. to  Bull.  WHO 53).
                            - i(Ml -
                            I }  /I1'1*

-------
Matsumoto, H., et al. 1967.  Preventative effect of penicil-
lamine on the brain defect of fetal rat poisoned transpla-
centally with methyl mercury.  Life Sci.  6:  2221.

McKim, J.M., et al.  1976.  Long-term effects of merthylmer-
curic chloride on three generations of brook trout  (Salv-e-
linus fontinalis):  Toxicity, accumulation, distribution,
and elimination.  Jour. Fish. Res. Board Can.  33:  2726.

Medical Tribune. 1978. Methyl mercury affects Japanese school-
children.  13 September, 1978.

Miettinen, J.K. 1973.  Absorption and elimination of dietary
(Hg "*") and methyl mercury in man.  Page 233. in  M.W.  Miller,
and T.W., Clarkson, eds. Mercury, mercurials and mercaptans.
Charles C. Thomas, Springfield, 111.

Morikawa, N. 1961.'  Pathological studies in organic mercury
poisoning.  Kumamota Med. Jour.  14:  71.

Morrow, P.2., et al. 1964.  Clearance of insoluble dust
from the lower respiratory tract.  Health Phys. 10: 543.

NAS.  1978. An assessment of mercury in the environment.
Panel on Mercury. Washington, D.C.

Nonaka, I. 1969. An electron microscopic study of the experi-
mental congenital Minamata Disease in rat.  Kumamoto Med.
Jour.  22:  27.

Nordberg, G.F., ed. 1976.  Effects and dose-response of
toxic metals.  Elsevier-Amsterdam.

Norseth, T. and T.W. Clarkson.  1971. Intestinal transport
of    Hg-labelled methyl mercury chloride; role of biotrans-
formation in rats.  Arch. Environ. Health 22: 258.

Oharazawa.  1968.  Chromosomal abnormalities and teratogenesis
induced by ethyl mercuric phosphate in pregnant mice.  Nippon
Sanka-Fujinka Gakka:  Zasshi  20: 1479.

Olson, F.C., and E.J. Massaro.  1977.  Pharmacodynamics
of methyl mercury in the marine maternal/embryo fetal unit.
Toxicol. Appl. Pharmacol.  39:  263.

Olson, G.F., et al.  1975.  Mercury residue's in fathead
minnows, Pimephales promelas Rafinesque, chronically exposed
to merthyimercury in water.  Bull. Environ. Contain.  Toxicol.
14: 129.
                                                            •
Overnell, J.  1975.  The effect of heavy metals on photo-
synthesis and loss of cell potassium in two species of marine
algae, Dunaliella tertiolecta and Phaeodactvlum tricornutum.
Mar .  BToTT"29 : 957

-------
Rahola, T., et-al- 1971.  The biological half-time of inorgan-
ic mercury (Hg  )  in man.  Scand. Jour. Clin. Invest. Abst.
27: 77 (Suppl. 116).

Ramel, C. 1972.  Genetic effects. Page 9 jji  L. Friberg,
and J. Vostal, eds.   Mercury in the environment -a toxico-
logical and epidemiological appraisal.  Chemical Rubber
Co.,  Cleveland.

Reinert,  R.E. et al.;  1974.  Effect of temperature on accu-
mulation of methylmercuric chloride and p,p'DDT by rainbow
trout  (Salmo gairdneri).  Jour. 'Fish. Res. Board Can.  31:
649.

Rothstein, A., and A.D.  Hayes.  1964.  The turnover of mer-
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Health Phys.  10: 1099.

Schamberg, J., et al. 1918.  Experimental studies of the
mode of absorption of mercury when applied by inunction.
Jour. Am. Med. Assoc. 70: 142.

Skog, E.  and J.E. Wahlberg.  1964.  A comparative investi-
gation of the percutaneous absorption of metal compounds,
in the,guinearpig by,means of,the radioactive isotopes   Cr,
5SCo, 63Zn, llSAq, ll3mCd, 2()3Hg.  Jour. Invest. Derm. 43:
187.

Sosnowski, S.L., et al.   1979.  The effects of chronic mer-
cury exposure on the mysid shrimp.  Mysidopsis bahia Abst.
N.E.  Fish & Wildlife Conf.  April 1-4.  Providence R.I.

Spyker, J.M., and M.  Smithberg.  1972.  Effects of methyl
mercury on prenatal development in mice.  Teratology  5: 181.

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

Stock, A., and F. Cucuel.  1934.  Die Verbreitung des Quick-
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Suter, K.E.  1975.  Studies on the dominant lethal and fertil-
ity effects of the heavy metal compounds methyl mercuric
hydroxide, mercuric chloride, and cadmium chloride in male
and female mice.  Mutat. Res.  30:  365.

Suzuki, T., et al. 1971.  Comparison of mercury contents
in maternal blood, umbilical cord blood and placental tissue.
Bull. Environ. Contam. Toxicol. 5.

Swedish Expert Group.   1971.  Methyl mercury in fish - toxi-
cological-epidemiological evaluation of risks.  Report from
an expert group.  Nordisk. Hyg. Tidsknift. Suppl. 4.

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Task Group on Metal Accumulation.  1973.  Accumulation of
toxic metals, excretion and biological half-times.  Environ.
Phys. Biochem.  3: 65.

Tejning, S. 1970. Mercury contents in blood corpuscles and
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Thurberg, F.P., et al.  1977.  Response of the lobster,
Homarus americanus,- to sublethal levels of cadmium and mer-
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Tonomura, K., and F. Kanzani.  1969.  The reductive decom-
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184: 227.         '

U.S. EPA.  1973.  Water, quality criteria, 1972.  Ecol. Res.
Ser. Rep. Comm. of Water Quality Criteria.  Natl. Acad.
Sci.  EPA/R3/73/033.  U.S. Government Printing Office.
Washington, D.C.

U.S. EPA.  1979.  Mercury:  Ambient Water Quality Criteria
(Draft).

Warnick, S.L., and ELL. Bell.  1969.  The acute toxicity
of some heavy metals to different species of aquatic insects.
Jour. Water Pollut.  Control Fed.  41: 280.

Wood, J.M.  1976.  Les mataux toxiques dans I1environment.
La Recherche 7: 711.

World Health Organization.  1976.  Environmental health
criteria,  Mercury.  Geneva.

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                                      No.  125
              Me thorny 1

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a survey of the  potential  health
and environmental hazards from exposure to  the  subject  chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources,  this short profile
may not reflect all available  information  including all  the
adverse health  and  environmental  impacts  presented  by  the
subject chemical.   This document has undergone scrutiny  to
ensure its technical, accuracy.
                               i If "T Ctr
                            •* /"J1 ) O

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                         Disclaimer Notice
Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.

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                                    METHOMYL
                                    Summary
     Methomyl  is  a  toxic  carbamate  insecticide  used on  field  crops and
 fruit.   It  is  readily absorbed through inhalation  or dermal exposure ana  is
 almost  completely eliminated  from  the body  within 24 hours.   Chronic tox-
 icity studies  in rats  and  dogs show  that no  effects occur below  100 ppm.
 The  threshold  limit  value  for methomyl in  air is 2.5 fig/m^.   Methomyl in-
 hibits the  activity  of cholinesterase  in the boay.   Studies have shown that
 methomyl is  not  carcinogenic  in rats and dogs  or mutagenic  in the Ames oio-
 assay.  However,  a different type of bioassay showed  mutagenic activity at a
 methomyl concentration  of 50  ppm.   A potential  product  of the  reaction  of
methomyl with certain nitrogen compounds  in the  environment or in mammalian
 systems is nitrosomethomyl,  which is a potent mutagen, carcinogen,  ana tera-
togen.
     Methomyl is  toxic to many  aquatic organisms  with 96-hour  LC5Q  levels
 ranging from 0.1 to 3.4 ppm.

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                                   METHOMYL
 I.    INTRODUCTION
      Methomyl  is a broad-spectrum insecticide used on many vegetables, field
crops, certain fruit  crops,  and ornamentals  (3erg, et al. 1977).  Introduced
by  DuPont  in   1966  as  an  experimental  insecticide-nematacide  (Martin  and
Worthing  1974),  methomyl -is now  manufactured by  DuPont  and  Shell (Stanford
Research  Institute 197A)  and used commercially as a foliar treatment to con-
trol  aphids, army  worms,  cabbage looper, tobacco_budworm,  tomato fruitworm,
cotton  leaf perforator,  and  ballworm  (Martin and  Worthing  1974).   About
three million  pounds  (1360 "tonnes) of  methomyl were produced" in'the united
States in 1974 under the  trade name Lannate® (Pest  Control,  1975).   Wastes
                                           *
associated with methomyl  production  may contain  methylene  chloride.   Metho-
myl  formulations  may contain  pyridine  as  a  contaminant  (Sittig,  1977).
Methomyl  is  highly soluble  in  water.   Its  bioconcentration  factor  is  1.0;
octanol/water coefficient, 2.0 (see Table 1).
II.  EXPOSURE
     A.   Water
          Methomyl is considered  stable in  ground water  ana  decomposes at  a
rate of less than  10 percent in  5 days in  a river environment.  In  a  lake
environment, methomyl decomposes  at  a rate  of  less than  85 percent  per  year
(U.S. EPA 1980).
     B.   Food
          After the application of methomyl  from 0.25 to 0.50  kilograms  per
hectare (kg/ha) on tomatoes, plant residues  were  below 0.2  ppm..  Application
of 1 kg/ha left residues of  0.3,  0.13,  and 0.06 ppm at 1, 2,  and 3 days,  re-
spectively, after  spraying  (Love and Steven,  1974).  Methomyl applied at  a
rate of 3 oz/acre (0.2 kg/ha) left a  17 ppm residue on rape  plants immediate-

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            TABLE 1.  PHYSICAL AND CHEMICAL PROPERTIES OF METHOMYL
    Synonyms:  S-methyl N-(methylcarbamoyl)oxy)thioac3tamidate;
           l-(methylthio)ethylideneamino methylcarbamate;
           l-(methylthio)acetaldehyde 0-methylcarbamoyloxime;
           methyl N-(((methylamino)carbonyl)oxy)ethanimidothioate;
           CAS Registry No. (16752-77-5); OuPont 1179; Lannate;
           Mesomile; Nudrin
    Chemical Formula:  (CH3S)(CH3)C=N-0(C=0)NHCH3

    Molecular Weight:  162.2

    Description:  White crystal solid
             Slight sulfurous odor
             Soluble in organic solvents

                                   24
    Specific Gravity and/or DensityJ  d   = 1.2946

    Melting and/or Boiling Points:  nip 78 to 79QC  -

    Stability: Stable in aqueous solution .
               Subject to decomposition in moist soil
               Overall degradation rate constant (0.01/day)

               Half-life approximately 50 days

Solubility (water):  5.3 g/ioo'ml at 25Qc

                   sediment .  .5
                     H20    '   1
    Vapor Pressure:  5 x 10-5 mm Hg at 25°C

    Bioconcentration Factor (BCF) and/or
    Octanol/water partition coefficient (Kow):   KQW =2.0
                                            BCF = 1.0
    Source:   Martin and worthing,  1974;  Fairchild,  1977;
    Windholz,  1976, U.S. EPA,  1980.
                                      X

                                   - >U.^^
                                  S*) I u u

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 ly  after application.   This concentration declined rapidly to 1.5, 1.0, 0.4,
 and 0.2  ppm,  1,  2,  5,  and 9  days later,  respectively.   Methomyl resioues
 were  not detected  (less than  0.02  ppm)  in  seed harvested  22  days  after
 application.   Rape  plant leaves collected after  the  application of methomyl
 at  3  to 4 oz/acre  (0.2-0.3  kg/ha)  had 2.5  to  16 ppm residues  (Lee,  et al.
 1972).
          Methomyl  has  a half-life in plants of  3 to 7 days.   Harvey (1975)
 detected  methomyl  residue,   its  oxime, and  small polar fractions  one  month
 after  application.   Methomyl residue  standards for  crops  are noted in the
 Existing  Guidelines and Standards Section of this report.
     C.   Inhalation and Dermal
          Data are  not  available indicating  the number  of  people  exposed  to
methomyl  by  inhalation  or dermal contact.  Most  human exposure  would appear
to  occur during  production and  application.  The  U.S. EPA  (1976)  listed the
 frequency of illness among  occupational  groups  exposed to pesticides.   In
1157  reported  cases,  most illnesses occurred among ground  applicators  (229)
and mixer/  loaders  (142).  The  lack  of or  refusal to  use safety  equipment
was a  major factor of  this  contamination,   other groups affected  were  gar-
deners  (101), field workers  exposed to pesticide  residues  (117), nursery and
greenhouse  workers   (75),  soil  fumigators  in  agriculture  (29),   equipment
cleaners  and mechanics  (28), tractor  drivers  and  irrigators  (23),  workers
exposed  to  pesticide drift  (22),  pilots (crop dusters) (17), and  flaggers
for aerial application  (6).  Most illnesses  resulted  from carelessness,  lack
of  knowledge of  the  hazards,  and/or   lack of safety  equipment.  Under  dry,
hot conditions workers  tended not to  wear protective clothing.  Such condi-
tions  also  tended  generally  to increase  pesticide levels and  dust on the
                                                                         »
workers.

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 III.  PHARMACOKINETICS
      A.   Absorption and Distribution
          Methomyl  is  a highly water-soluble carbamate insecticide which can
 be absorbed  readily by moist mucous membranes or through the skin  (Guerzoni,
 et al.  1976).   Methomyl applied to  the  skin is less toxic than methomyl ao-
 ministered  orally  (Kaplan" and Sherman,  1977).   Kaplan  and  Sherman  (1977)
 noted that  there was  no  buildup  of methomyl in  fish after a 30-day  feeding
 study,  indicating that methomyl was not distributed or  retained  in any one
 specific  organ  of the body.   In another study, there was ho cumulative oral
 toxicity  in  rats  (Harvey,  et al.  1975).   The  investigators measured a total
 clearance rate  of  less than 24 hours after  oral administration of methomyl
                                            %
 to rats.
      B.   Metabolism
          Harvey,   et   al.   (1973)  administered  l4C-labeled  methomyl  to
 rats.   The  radioactive methomyl  was eliminated  in  the  form of  carbon ai-?
 oxide,  acetonitrile,   and  urinary  metabolites.   They  noted the  absence  of
 methomyl, S-methyl  N-hydroxythioacetimioate,  methyl  S,S-aioxide,.  and conju-
 gates  of the  former   two  compounds.   Radiolabeled methomyl  administereo  in
 the rat by Huhtanen and Oorough (1976) also was metabolized to carbon dioxide
 and  acetonitrile.  Carbon dioxide  was  also   found  in  soils treated  with
methomyl  (Heywood 1975), without the presence  of sulfoxide  or sulfone (Baron
 1978).
          Han  (1975)   investigated  the   formation of  nitrosomechomyl  from
cured meats  containing methomyl  and  residual  sodium  nitrite.  The  samples
were  incubated  under  simulated   stomach  conditions  (pH^)  for  1  and   3
hours.   Nitrosomethyl  was  not founo  in  the  test  material;  the  detection
limit was less than 1 ppb.

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      C.    Excretion
           Methomyl   is   eliminated  primarily  through  the  urinary  system
 (Harvey, et al. 1975).
 IV.   EFFECTS
      A.    Carcinogenicity
           No  evidence of methomyl carcinogenicity  was observed in tests with
 rats  and dogs  (Kaplan and Sherman,  1977).   LLjinsky  and Schmaehl (1978) con-
 cluded that if nitrosomethyl  carbamates  (nitrosomethomyl)  were formed by the
 reaction of  the parent  insecticide  (methomyl)  with nitrite in  the environ-
 ment  or  in the stomach,  the  carcinogenic risk of the parent  compound could
 increase.
           In  pesticide  workers,  two cases  of embryonal cell  carcinoma have
 been  associated  with  exposure,  to  methomyl  and-  three  other  pesticides
 (carbaryl, paration,  and dimethoate).   One  of the pesticide  workers under-
 went  surgery for a testicular mass; the  second  worker died of  metastatic em-
 bryonal cell  carcinoma.   These  cases led the authors  to  suggest that testi-
 cular cancer may be  related to  agricultural  chemical  exposure  (Prabhakar and
 Fraumeni, 1978).
      B.   Mutagenicity
          Blevins, et  al. (1977)  screened  methomyl  and its nitroso  deriva-
 tive  for mutagenic  activity.   Using histidine  auxotrophs  of  S^_ typhimurium
 derived by  Ames,  they  noted  that methomyl,  unlike  its nitroso  derivative,
 did not cause  a significant increase  in  the number of  revertant  colonies  in
 any of the strains used.  Thus,  while nitrosomethomyl appeared to oe  a po-
 tent mutagen,  they considered methomyl to be non-mutagenic.
          Guerzoni,  et  al.  (1976) tested methomyl for mutagenic  activity  on
Saccharomyces  cerevisiae.  Methomyl was  considered  mutagenic at  50 ppm.  The

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 authors  noted,  however,  that  the  mutagenic  effect  depended  on  the  S^
 cerevisiae  strain.
      C.   Teratogenicity and Other Reproductive  Effects
          Methomyl  was fed to pregnant  New Zealand White  raobits on days  3
 to  16 of gestation.  Teratogenic effects were not found at any  of three  die-
 tary  levels,  0,  50,  and  100  ppm  (Kaplan  and Sherman,  1977).   The  same
 authors  also  reported on  a  3-generation,  6-litter  reproduction study  with
 rats  with the  same  dietary levels.   Methomyl did not have adverse  effects  on
 reproduction and lactation  performance;  in addition,  pathological changes
 were  not observed in the third-generation weanling pups,  using a  model  eco-
 system,  Howe (1978) did not see effects  on quail egg production or egg  fer-
 tility from  a  diet  of  10,  40, and 30 ppm methomyl.
          Blevins,  et al.  (1977)  treated  normal  human  skin cells with six
 insecticidal  esters  of N-
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     0.   Chronic Toxicity
          Rats  of  both sexes  were  fed nutritionally  complete diets  con-
taining  0,  10,-50,  125, and  250 ppm of methomyl in a  90-day  feeding study
and  0,  50,  100, 200.,  and  400 ppm  of methomyl in a  22-month feeding study.
The weight gain for the high-dose males  was  significantly  lower than that of
controls.   No clinical, hematological,  biochemical,  urinary,  or  pathologic
evidence  of toxicity  was  observed  at  90 days.   However,  in  the  22-month
study,  decreased  Hb  values  were noted  in  the  two higher-dose female  test
groups.   A  higher  testis/body  weight ratio was observed  in  the  high-dose
males.   Histopathologic alterations  were observed  in  kidneys of male  and
female  rats  receiving 400 ppm and  in spleens of  the female rats  receiving
200  and  400 ppm of methomyl.   Beagles  of both sexes  fed  nutritionally  com- •
plete diets  containing 0,  50,  100, and  400 ppm of  methomyi in 90-day  and
2-year  feeding  studies  showed  no  nutritional,  clinical,  urinary,  or  bio-
chemical  evidence of  toxicity.   In  the  2-year study, an  additional  dietary
level of 1000  ppm  caused  some clinical  signs of  toxicity and  mortality.
Similar  to  findings in the 22-month feeding study in rats,  histopathologic
changes were  observed after  2 years in the  kidney, spleen,  and liver  at the
two higher feeding levels.  Dogs receiving the high-level  diet  showed  a  com-
pound-related anemia.   Results  of  the long-term studies indicated that  the
no-effect level for rats and dogs was 100 ppm (Kaplan and Sherman,  1977).
     E.   Other Relevant Information
          Several incidents  of  acute occupational  exposure have  been  re-
ported  in the  literature.    In  the  first  incident,  four  crews  of fiela
workers  harvesting  vegetables and  fruits  treated with  pesticides  including
methomyl were studied.  One crew had  depressed blood  cholinesterase  activity

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  after harvesting corn treated with methomyl.   Forty-eight percent of another
  crew  had  significant  cholinesterase  depression  after  harvesting  treated
  citrus, tomatoes, and gladiolas (Owens, et al. 1978).
            A second  incident involved  120  grape pickers where  IQb displayed
  symptoms  suggesting  pesticide   poisoning.    Methomyl  and  other   cnolin-
  esterase-inhibiting pesticides, such as dimethorate  and torak,  were  named in
  a legal complaint against the grower.   The major symptoms  claimed by the ex-
  posed workers were  headache,  dermatitis,  vomiting, nausea, fatigue,  and eye
'  pain (McClure,  1976).
            Kumagaya,  et  al.  (1978) reported  on two  cases  of poisoning  from
  swallowing methomyl.  The general symptoms were loss  of consciousness,  re-
  spiratory failure,  miosis,  myofibrillary  twitching,  increase in airway  se-
  cretions, and reduced serum cholinesterase  activity.  Complications  of  pul-
  monary edema, hepatitis, and polyneuritis were also observed.
            The  oral  LD^ values for  rats,  mice, ducks, and wild biros  have
  been reported as  17,  10, 15,  and 10  mg  methomyl per  kilogram body weight
  (mg/kg),   respectively.   The  oral  LD5Q values for  dogs, monkeys, guinea
  pigs,  and chickens  are  reportedly 30,  40,  15, and  15 mg/kg,  respectively.
  Inhalation i_C5Q values  for  rats, quails,  and ducks are  77,  3680,  and  Ib90
  ppm,  respectively.   The  dermal LD5Q  for rabbits is  5000  mg/kg.  NO  adverse
  effects were noted  when bootail  quail  and  aloino  rabbits  were sprayed  six
  times  (at  5-day  intervals)  with  1.1 kg/ha of a 90 percent  formulation of
  methomyl. Methomyl  is  relatively  non-toxic to bees, once the spray has dried
  (Fairchild,  1977; Martin and worthing,  1974).
            Carbamate  pesticides,  such  as methomyl,  have  cholinergic   proper-
  ties similar to  those  of the  organic phosphates,  out  of shorter duration.
  Methomyl  inhibits both  RBC  and  plasma cnolinesterase  activity.   The period

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 of inhibition  of the  cholinesterases  is  approximately  1-2  hours,  and  re-
covery  usually  occurs between 24  and  48 hours after  contact.  Atropine  ad-
ministration is the treatment of choice  (Simpson and Bermingham, 1977).
V.   AQUATIC TOXICITY
     A.   Acute and Chronic Toxicity
          Methomyl  24-hour   TLm  (median  toxic   limit)   values  for  carp
(Cyprinus carpio) and tilapia  fish range from 1.054 to  3.16 mg/1 (El-Refai,
et  al.  1976).   The  LC5Q  (96-hour  exposure)  for  rainbow  trout  (Salmo
gainneri) was 3.4 ppm;  for bluegill  (Lepomis  macrochirus),  0.87 ppm; and for
goldfish  (Carrasius  auratus),  greater  than  0.1 ppm  (Martin  and  Worthing,
1974).   Following exposure  (4-48  hours) of  marine or  estuarine fishes  to
carbamate pesticide,  the acetylcholinesterase activity in  the brain was in-
hibited by 77 to 89 percent (Coppage, 1977).
     B.   Plant Effects and Residues
          Pertinent data could not be located in the available literature.
VI.  EXISTING GUIDELINES AND STANDARDS
     A.   Human
          The  threshold limit  value  for  air  is  established at  2.5 mg/m-3
(Fairchild, 1977).  The  Office of  Water  and Waste Management  is  in  the pro-
cess  of  conducting  preregulatory  assessment  of  methomyl under  the  Safe
Drinking Water  Act.   The Office of Toxic Substances has  promulgated regula-
tions for methomyl under Section 3 of the Federal  Insecticide, Fungicide and
Rodenticide Act.
          Methomyl  residue   concentrations  in  crops   are   regulated   as
follows:  0.1  ppm  for  lentils and  pecans;  1  ppm  for  forage, hay,  barley
                                                                         »
(grain), and  oats (grain); 2  ppm  for strawberries and  avocados; 5  ppm  for
Chinese cabbage;  6  ppm  for  blueberries, beets,  collard,   danoelions,  kale,

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mustard greens,  parsley,  swiss chard, turnip greens, and  watercress;  10 ppm
for wheat,  rye,  barley,  and oats used as  hay,  straw, or  forage;  and  40 ppm
for bermuda grass  hay (Federal. Register [43(98):  21700,  1978;  43 and (112):
25120, 1978;  44(63):  18972; 44(83):  24846;  44(129): 38844;  44(160):  47934,
and 44(227): 67117, 1979]);
     B.  Aquatic  -
          Guidelines  or  standards  to protect aquatic life coulo  not  be lo-
cated in the available literature.

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                                   REFERENCES
 Baron,  R.L.  1978.  Terminal residues of carbamate  insecticides.   Pure Appl.
 Chem.   50:  503.

 Berg,  Gil., et al., (ed)   1977.   Farm Chemicals Handbook Meister  Publishing
 Company,  Willoughby, Ohio.

 Blevins,  R.O., et al.   1977.   Mutagenicity  screening of  five methyl  car-
 bamate   insecticides  and.  their   nitroso  derivatives   using  mutants   of
 Salmonella  typhimurium LT2.   Mutat.  Res.  56:  1.

 Coppage,  O.L.   1977. - Anticholinesterase  action of pesticide carbamates  in
 the  central nervous system  of  poisoned fishes.  Physiological  Response  Mar.
 Biota.  Pollut. Proc. Symp.   pg.  93.

 El-Refai,  A.,  et  al.  1976.  Toxicity of  three  insecticides to two  species
 of fish.  Int.-Pest  Control,  18: 4.

 Fairchild,  E.J.  (ed.)   1977.  Agricultural Chemicals and Pesticides:   A  sub-
 file  of the NIOSH Registry  of  toxic effects of chemical  substances,  U.S.
 Dept. of  HEW,  July.

 Guerzoni, M.E.,  et al.   1976.   Mutagenic activity-of  pesticides.  Riv.  Sci.
 Tecnol. Alimenti.  Nutr. Urn.,  6:  161.

 Han,  J. C-Y,  1975.   Absence of  nitroso  formation  from  (14-C)  methomyl and
 sodium  nitrite under simulated  stomach  conditions.  Jour. Agric.  Food Chem.
 23: 892.

 Harvey, J.J.,  et  al.  1973.  Metabolism of methomyl in the rat.   Jour.  Agr.
 Food Chem., 21: 769.

 Harvey,  J.J.  1975.   Metabolism  of  aldicarb and  methomyl.   Environmental
 Quality Saf., Suppl.  Vol. 3, ISS Pesticides,  389.

 Heywood,  O.L.    1975.    Degradation  of  carbamate  insecticides  in  soil.
 Environ. Qual.  Saf., 4: 128.

 Howe, G.J.   1978.   The effects of various  insecticides applied to a terres-
 trial model  ecosystem  or  fed in the  diet  on  the serum cholinesterase level
 and  reproductive  potential of coturnix   quail.    Oiss.   Abstr.   Int.  3.
 38: 4785.

 Huhtanen,  K.  and   H.W. Dorough   1976.  Isomerization.ana  Beckman rearrange-
ment  reactions in  the  metabolism  of  methomyl  in.rats.   Pest.  Biochem.
 Ftiysiol.  6: 571.

Kaplan,    A.M.   and  H.   Sherman   1977.    Toxicity   studies   with   methyl
N-(((methylamino)carbonyl)oxy)ethanimidothioate.  Toxicol.  Appl.  Pharmacol.
40: 1.
                                     XL

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 Kumagaya,  S.,  et  al.   1978.   Pesticide  Intoxication.   Yamaguchi  Igaku
 27:  211.

 Lee,  Y.M.,  et  al.   1972.   Residues  of  methomyl in  rape  plant ana  seed
 following  its  application  for  the  control of  bertha  army  worm,  Mamestra
 configurata  Lepidoptera Noctuidae.

 Lijinsky,  w.  and 0.  Schmaehl  1978.  Carcinogenicity  of N-nitroso  deriva-
 tives of N-methylcarbamate insecticides in rats.   Ecotoxicol.  Environ.  Saf.,
 2: 413.

 Love,  J.L.  and  0.  Steven   1974.   Methomyl residues on  tomatoes.  N ana  J.
 Exp. Agric.,  2:  201.  ••••-•

 Martin and worthing  (ed.), 1974.  Pesticide Manual, 4th edition.

 McClure,  C.O.   1976.   Public  health concerns  in  the  exposure  of  grape
 pickers to high  pesticide residues  in  Madera County, Calif. Public Health
 Report-.93:  421,  September.

 Owens,  C.O.,  et  al.   1978.   The  extent of exposure of  migrant workers  to
 pesticide  and  pesticide  residues.   Int. Jour. Chronobiol.   5:  428.

 Pest Control  1975.   pg. 314.

 Prabhakar,  J.M.  and J.F. Fraumeni   1978.   Possible relationship of insecti-
 cide exposure  to  embryonal cell carcinoma.   Jour. Am. Med.  Assoc.   240: 288.

 Simpson, G.R.  and S.  Birmingham  1977.  Poisoning by  carbamate pesticides.   .
Med. Jour.  Aust.  2: 148.

 Sittig,   M.   1977.    Pesticides  Process  Encyclopedia,  Chemical   Technology
Review no.  81.  Noyes Data Corporation, Park Ridge, N.J.

 Stanford Research Institute.  1977.   Directory of  Chemical Producers. Menio
Park, California.

 U.S. Environmental  Protection Agency.  1976.   Organophosphate Exposure from
 Agricultural Usage,  EPA 600/1-76-025.

U.S. Environmental  Protection Agency.   1980.   Aquatic  Fate  and Transport
Estimates   for  Hazardous   Chemical   Exposure   Assessments.    Environmental
Research Laboratory,. Athens,  Georgia.

windholz,   M.   1976.   The  Merck Index, Ninth  Edition,  Merck  and  Co., Inc.,
Rahvvay,  N.J., USA.

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                                     No.  126
          Methyl Alcohol

 Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in tha report is  drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources, this  short profile
may not reflect  all available  information  including all the
adverse health  and  environmental  inpacts  presented by  the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

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                          BIOLOGIC EFFECTS OF EXPOSURE





Extent of Exposure


      Methyl alcohol, CH30H, also called methanol, is the first member of a


homologous series of monohydric aliphatic alcohols.  At  room  temperature,
                                            -'  .
methyl alcohol is a colorless, neutral liquid possessing a mild distinctive


odor. [1]  Additional chemical and physical properties  of  methyl  alcohol

are presented in Table XIII-1. [2,3,4]

      The greater part of methyl alcohol manufactured in the US is produced

synthetically. [5]  One  widely  used  synthetic  process  is  the  "medium

pressure   process"   which  involves  the  reduction  of  carbon  monoxide

(containing small amounts of carbon dioxide) with hydrogen.  The  reduction

step  is carried out at 250-400 C and at 100-600 atmospheres pressure using

a catalyst. [1]

;  "   During  the years 1968-»73, synthetic methyl alcohol production in  the


US increased at an average  annual  rate  of  over   13.2%.   In   1973,   the

production  of  synthetic  methyl  alcohol  amounted to slightly over seven

billion pounds, around one billion gallons.  In addition, an   estimated   10

million   pounds    (1.5  million  gallons)  of  "natural"   (eg,  from  wood

distillation) methyl alcohol were produced.  [5]

      Methyl  alcohol  is  used  in a variety of industrial processes.   The

major use is in 'the production of formaldehyde which amounted  to  39% of  the
                            •  r                    •      '
methyl  'alcohol  consumed in the US in  1973..   [5] "Other commercial uses of

methyl alcohol are  in the  production  of  chemical  derivatives,  such   as

dimethyl  terephthalate,  methyl halides, methyl methacrylate, acetic a'cid,

and  methylamines, and because of its solvent properties, methyl alcohol   is

-------
 also  used  in  paints,  varnishes, cements, and  other  formulations such as

 inks and dyes. [1,5]  Table XIII-2 lists the consumption  of methyl  alcohol

 by product and quantity produced in the US for the year 1973.  [5]

       A. number of occupations with potential exposure to  methyl alcohol are

 listed in Table SLII-3. [6]

       NIOSE  estimates, .that  approximately  175,000  workers  in the US.,ara
 ipotentially  exposed to methyl alcohol:

                              EFFECTS  ON HUMANS
              Burk  [26]  attributed  the toxic effects of  methyl alcohol to

 formaldehyde and formic acid, indicating that both compounds were oxidation

 products of  methyl alcohol.  The author stated that the diagnosis of methyl

 alcohol poisoning is sometimes very difficult,  and  would   be  more  easily

 verified  by  quantitative  determinations  of  formic acid in the urine of

 persons suspected of being poisoned with methyl alcohol.

                         ^rcutaneous absorption of methyl alcohol can lead

to serious consequences, including death..  In  1968,  Gimenez   et  al [27]

reported  an  analysis  of  19  cases  of children, ranging in  age from  1.5

months to 4 years, who were poisoned as a result of having cloths soaked in

methyl  alcohol  applied  to  their,  abdomens  to  relieve gastrointestinal

troubles or other unspecified complaints.  There were  2   additional cases

reviewed  in which both methyl and ethyl alcohols  had been employed  in this

way, making a total of  21 cases.  Although absorption of methyl alcohol  via

 the resoiratory tract was possible in  these cases, the fact  that the cloths

 were held in place by  rubber baby pants would favor percutaneous absorption

 of  the  alcohol  as  the significant route of exposure.  The length of time

 between  application   and onset of symptoms of intoxication was 1-13 hours

 (7  1/4  hours average) .  The  aarly signs of  intoxication  were  described* 'oy

 the authors  as central nervous  system  depression with   13  children  having

-------
exhibited severe respiratory depression and U. of these having convulsions.

Slood pH in the 21 patients ranged from 6.4  to  7.38  (normal:   7.36-7.41

[23])>  indicating  acidosis in most cases.  Twelve of the 21 children died

of cardiac or respiratory arrest 2-10 days after hospital  admission.   The

survivors  recovered  without  apparent  permanent damage.  Papilledema and
                                           .  >  .
ocular fundus bleeding were observed in 2 of the infants  who  subsequently

died.   Abdominal  skin  lesions  were  present  in  5  patients,  3 of the

erythematous type and '2 of the scaling type.^  The  authors  [27]  commented

that while there was no relationship between methyl alcohol blood levels as

tested  in  11  children   (57-1,130  mg%)  and  prognosis,  there   was   a

relationship  between the initial blood pH and the subsequent course of the

illness.   In  general,  treatment  consisted   of   administering   sodium

bicarbonate, glucose, ethyl alcohol, fluids, and electrolytes.  Other forms

of treatment included peritoneal dialysis, exchange transfusion, mechanical

raspiration,.  and  the  administration of anticonvulsant drugs.  It must be

pointed out that the absorptive properties  of  the  skin  of  infants  are

probably   different   from   those   of  adults  and  consequently  infant

susceptibility to, and manifestations of, methyl alcohol  intoxication  may

not  parallel those seen, in adults.                                          :

      In 1952, Leaf and Zatman [30] reported on experiments in which 5 male

voluntaers ingested 2.5-7.0 ml of methyl alcohol diluted  to  100  ml  with

water.   These  amounts  of  methyl  alcohol corresponded to doses of 29-84

ing/kg.  Two blood samples were taken from 3 subjects, 2-5 hours  after  the

ingestion.  Urine was collected frequently for 11-16 hours following methyl

alcohol administration.  Both the blood and urine samples were analyzed for

methyl  alcohol  by  a colorimetric method based on the oxidation of methyl

alcohol to formaldehyde and formation of a colored complex with a  modified
                                                                       »
Schiff's  reagent.   The  results  of  this experiment indicated that under

these  conditions  methyl   alcohol   was   rapidly   absorbed   from   the

gastrointestinal  tract.    The  maximum methyl alcohol concentration in the

urine  was  achieved  approximately  one  hour  after  ingestion  and  then

decreased  exponentially. .. The  ratio  of  blood  to  urine methyl alcohol

-------
concentrations remained almost constant for the 3 subjects in which it  was


determined,  and  the  authors  [20]  concluded  that  che  change  in  the


concentration of methyl alcohol in the urine was an accurate  .indicator  of


the change in aethyl alcohol concentration in the body.  At the levels used


in this experiment, the  concentration  of  methyl  alcohol  in  the  urine


declined  to  control  values within 13-16 hours after ingestion.  Leaf and


Zatmaa [30] also stated'that only 0.4-1.2Z of the ingested  methyl  alcohol

                                     I        >  •
was eliminated unchanged in the urine.


       In  another  experiment  in  the  same  study, [30]  2 male volunteers


 ingested 15  ml of ethyl alcohol and 4 ml of methyl alcohol  simultaneously.


 They  then ingested 10 ml of ethyl alcohol every hour for the next 7 hours.

                                                 j  .
 The  same individuals served as their own controls in a previous  experiment


 in  which  they  ingested only 4 ml of methyl alcohol.  Urine was collected


 hourly and analyzed for methyl alcohol.  The maximum urinary methyl alcohol


 concentrations  for  those individuals who ingested both methyl alcohol and


 ethyl alcohol were 8.82 and 9.20 mg/100 ml, compared to values of 6.05  and


 5.50 mg/100 ml when methyl alcohol alone was ingested.  Moreover, the total


 amount of methyl alcohol excreted unchanged in the urine  in  the  first  7


 hours  after  ingestion  was  107.1  tag  and 125.5 mg (3.7 and 3.96% of the


 administered dose respectively) when both methyl alcohol and ethyl  alcohol


 were  ingested,  whereas  only  from  18.2  to  30.8  mg (0.57-0.97% of the


 administered dose) was excreted unchanged in a similar  time  period  after


 ingestion of 4 ml methyl alcohol alone.  The authors  [30] concluded that  in


 humans ethyl  alcohol  interfered,  with  the  normal  oxidation  of  methyl


 alcohol,  causing  more  of  it  to  be  excreted 'unchanged  in the urine.


 Moreover, according to the authors' conclusion,  higher  concentrations   of


 methyl alcohol in the blood are maintained in the presence of ethyl alcohol


 at any given time after absorption, as compared to concentrations  achieved


 In  che absence of ethvl alcohol.

-------
                                 Ethyl alcohol may inhibit the oxidation of


methyl alcohol in  vivo  by  competing  (competitive  inhibition)  for  the


alcohol  dehydrogenase  system.  It is conceivable, therefore, that chronic


alcoholics might exhibit measurable concentrations of methyl alcohol in the


blood or urine even though they have not been exposed to methyl alcohol. L.3?J


      In  summary,  an  integration of in vitro  [33-35] and in vivo studies


[29-31,37]-  indicates that in humans methyl alcohol is  oxidized  primarily


by  alcohol  dehydrogenase.  The results discussed in the section on Animal


Toxicity, however, suggest that in nonprimates methyl alcohol  is  oxidized


primarily by the catalase-peroxidase system.



                          . ANIMAL TOXICITY



         Gilger   and   Potts  [42]  concluded  from their studies  that the results.


   of  oral administration   of  methyl  alcohol  to*  rats,   rabbits,   and  dogs


   differed  from  those reported on  humans  in 4 important areas,  namely, lethal


   dose,  time course  of development and  signs of intoxication,   eye  effects,


   and acidosis.  The  authors also  concluded that, following intoxication with


   methyl alcohol,  the responses  of  primates more closely   approximated  human


   responses  than, did  those   of   nonprimates.    An  extensive review of the


   literature dealing  with the  oral  toxicity of methyl alcohol in  humans  and


   nonprimates was  supportive of  their  conclusion.  The authors  concluded that


   the approximate  lethal  oral  dose  of  methyl  alcohol in humans   (0.85-1.4


   g/kg) was  1/3  the equivalent dose in monkeys and 1/9  the equivalent dose in


   rats.  Moreover,  nonprimates  exhibited   severe  early  intoxication  with


   narcosis   lasting   until   death   whereas   primates  showed  much  less early


   intoxication followed by a symptomless  latent  period, then  by sickness  and


   death.    The  only   eye changes observed  with  certainty in  nonprimatea were


   early pupillary  changes and  corneal  opacities  following exposure  keratitis.


   Some  monkeys,   however,   and  many  humans  developed   partial or  complete
                                      i >'n->
                                   .-• ) i  > 7 "

-------
    blindness accompanied by eyeground changes such as hyperemia of   the  optic



    discs  and venous engorgement.  Finally, humans and monkeys often developed



    severe acidosis (COZ-combining capacity  less  than  20  volumes  7,}  after



    methyl  alcohol  ingestion;  this  condition  was  rare  in nonprimates and



    occurred only at near lethal or lethal doses.
Correlation of Exposure and Effect




      Well-documented studies that correlate environmental, levels of methyl




alcohol with observed toxic effects have not been found in the  literature,




nor   have   any  long-term  epidemiologic  studies  of  chronic  low-level




occupational exposure been found.



      Effects  seen from either of the 2 most common routes of  occupational




exposure  (inhalation  and  percutaneous  absorption)  include:    headache




[14,16,17,39];   dizziness   [13,19];  nausea   [16,17,26];  vomiting   [17];




weakness  (unspecified) [16]; vertigo  [17,26]; chills  [13];  shooting   pains




in   the   lower  extremities  [13];  unsteady  gait   [17];  dermatitis  [U];



multiple  neuritis characterized by paresthesia,  numbness,  prickling,  and




shooting  pain  in  the back of  the hands and forearms, as well as  edema of




the  arms  [15]; nervousness  [19];  gastric pain  [19];  insomnia  [19];  acidosis




[19];  and  formic  acid  in   the urine.  [26]   Eye  effects,'such  as blurred




vision,  [16,17] constricted visual fields,   [17,19,25]  blindness,   [13,25]




changes   in   color  perception,  [17]  double vision,  [19]  and  general  visual

-------
disturbances [17] have been reported.  Eye examinations have shown sluggish


pupils,  [13,17] pallid optic discs,  [13] retinal edema,   [17]  papilledema,


[26] ' hyperemia  of  the  optic discs with blurred edges  and dilated veins.


[17]


      The  study  by  Bennett  et al  [40] showed similar  symptoms resulting
                                              >  •
from  ingestion.   These  are  acidosis,  headache,  visual   disturbances,


dizziness,  nausea  and  vomiting, severe upper abdominal pain, dilated  and


nonreactive pupils.  Eyeground examinations showed hyperemia of   the  optic


discs and retinal edema.  The eyeground changes were almost always found in


acidotic patients.  This finding is  suggestive  of  a  correlation  between


acidosis  and visual disturbances.   However,  a number of  patients, with  and


without  acidosis, complained of visual disturbances.   Additionally,  blood


tests  showed  elevated  serum  amylase  levels in 14 of  21 patients.  This


finding  in  conjunction  with  complaints  of  upper  abdominal  pain   and


pancreatic  necrosis  seen at autopsy led the authors  [40] to conclude that


hemorrhagic pancreatitis resulted from acute  methyl  alcohol  intoxication.


However,  reports  of  acute  hemorrhagic pancreatitis by parenteral routes


have-not been found.


      Direct  skin  contact  with  methyl  alcohol  has   been said to cause


dermatitis,  [14] erythema, and scaling.   [27] The reported variability   in


susceptibility   [14]  is  probably largely becaus.e of variations  in  tine of


contact  with methyl alcohol; it is evident that sufficient  -dermal  contact


.,with  any lip id solvent such as methyl alcohol has  the potential  for causing


skin  irritation.


 Basis for the Recommended Environmental Standard


       Epidemiologic   studies   incorporating  comprehensive   environmental


 surveys, well-planned surveillance,   a  sufficient  study  population,   and


 statistical  analysis  have  not  been  found  in  the  literature.    It is


 therefore  difficult  to  recommend  an  environmental  limit  based    upon


 unequivocal scientific data... .

                                      r '/rffl—
                                  **?)//

-------
                                TABLES AND FIGURE
                               TABLE XIII-1

            PHYSICAL AND CHEMICAL PROPERTIES OF METHYL ALCOHOL
Molecular formula
j
Formula weight
Apparent specific gravity at 20 C
Boiling point at 760 mmHg
Vapor pressure at 20 C
Melting point
Solubility in water
Solubility in alcohols, ketones, esters,
and halogenated hydrocarbons
Flash point, Tag open cup
Flash point, Tag closed cup
Flammable limits
(% in air)
Vaoor density
CH30H
32.04
0.7910
64.5 C
96 nmHg
-97.6 C
Miscible
Miscible
16 C
12 C
6.72-36.50
1.11
(air-1)

Corrosivity
Conversion factors
(760 tnmHg and 25 C)
Noncorrosive at
normal atmospheric
temperatures.
Exceptions: lead and
aluminum

1 ppm-"l.310 mg/cu m
1 mg/cu o».763 ppm
Adapted from ANSI 237  [2],  Che Manufacturing  Chemists  Association [3],
and the Handbook of Chemistry and Physics  [4]
                            ISL

-------
 f
i."
 I
*
                                             TABLE XIII-2

                                  US METHYL ALCOHOL CONSUMPTION, 1973
-
Formaldehyde
Dimethyl terephthalate
Solvent usage
Methyl halides
Methylamines
Methyl methacrylate
Inhibitor for formaldehyde
Exports
Glycol methyl ethers
Acetic acid
Miscellaneous
Total
Million Pounds
2,778
435
565
435
232
265
66
824
81
240
1,207
7,128
Million Gallons
420
66
85
66
35
40
10
124
12
36
181
1,075
               From Blackford  [5]
                                             I-It

-------
                              TABLE XIXI-3
           POTENTIAL OCCUPATIONAL EXPOSURES TO METHYL ALCOHOL
Acetic acid makers
Adhesive workers
Alcohol distillery workers
Alcohol lamp users
Aldehyde pumpmen
Antifreeze workers
Art glass workers
Automobile painters
Aviation fuel handlers
Bookbinders
Bronzers
Brushmakers
Denatured alcohol workers
Dimethyl sulfate makers
Drug makers
Drycleaners
Dye makers
Dyers
Ester makers
Explosives workers
Feather workers
Felt-hat makers
Flower makers, artificial
Formaldehyde makers
Foundry workers
Furniture polishers
Gilders
Glassmakers, safety
Hectograph operators
Incandescent lamp makers
Inkmakers
Japan makers
Japanners
Jet fuel workers
Lacquerers
Lacquer makers
Lasters
Leather workers
Linoleum makers
Lithographers
Metal polishers
Methyl acrylate makers
                                Methyl alcohol workers
                                Methyl amine makers
                                Methylation workers
                                Methyl bromide makers
                                Methyl chloride makers
                                Methyl methacrylate makers
                                Millinery workers
                                Motor fuel blenders
                                Organic  chemical synthesizers
                                Painters
                                Paintmakers
                                Paint remover workers
                                Patent leather makers
                                Perfume  makers
                                Photoengravers
                                Photographic film makers
                                Polish makers
                                Printers
                                Rayon makers
                                Resin makers
                                Rocket fuel handlers
                                Rocket fuel makers
                                Rubber shoe cementers
                                Rubber workers
                                Shellackers
                                Shellac  makers
                                Shea factory workers
                                Shoe finishers
                                Shoe heel  coverers, wood
                                Shoe stitchers
                                Soapmakers
                                Straw-hat  makers
                                Sugar  refiners
                                Textile  printers
                                Type cleaners
                                Vacuum tube makers
                                Varnish  workers
                                Vulcanizers
                                Wood alcohol distillers
                                Wood stainers
                                Wood stain makers
From Gafafer [6]

-------
                               TABLE XIII-4

                      ANIMAL EXPERIMENTATION RESULTS
                        OF METHYL ALCOHOL EXPOSURE
Species
 Route of
 Exposure
  Dose
         Effect
 Ref-. .
erence
Monkeys
Inhalation
5,000 ppm
duration
unknown
The monkey survived for
an unstated period of time.
  47
                           1,000 ppm
                           duration
                           unknown
                              The monkey died promptly      47
                              upon exposure at this level.
Dogs
               450-500 ppm
               8 hr/day
               7 days/week
               for 379 days
               Blood levels of methyl
               alcohol were found to range
               from 10 to 15 mg/100 ml
               of blood and on occassion
               went as high as 52 rag/100 ml.
               No abnormal eye findings
               were reported.
                              41
            Oral
               2.5 to 9.0     Of the 9 treated dogs, 2
                  g/kg        died at doses of 4 and
               body weight    9 g/kg.  C02 combining
                              capacities dropped below
                              normal in 2 dogs, and no
                              ophthalmoscopic changes
                              were noted.
                                             42

-------
                         TABLE XIII-4 (CONTINUED)

                      ANIMAL EXPERIMENTATION RESULTS
                        OF METHH. ALCOHOL EXPOSURE
Species
 Route of
 Exposure
  Dose
Effect
 Ref-
erence
Monkeys
Oral
1.0 to 3.0     Acidosis developed in
   g/kg        monkeys receiving doses
               ranging from 3.0 to 6.0
               g/kg.  The animal receiving
               1.0 g/kg did not develop
               acidosis.  Definite eye-
               ground change occurred to
               2 of the acidotic monkeys.
                     42
Rats
               4.75 g/kg      70% mortality                  42


               4.5 g/kg       None of  the 9  tested rats      42
                              developed acidosis.
Rabbits
               3.5 g/kg       One animal receiving  this
                              dose died in  less  than  24
                              hours.  No eye  fundus
                              changes were  reported.
                                             42
Rabbits
               2.1 g/kg       Of  the  3 animals  tested  at     42
                              this dose, all died between
                              24  hours, and  3 days after
                              dosing.
             Intra-          10  ing  and       At 10  nig,  there was  no skin   49
             cutaneous       35  mg           reaction, .whereas  at 35
                                           mg,  a  9-sq  mm skin reaction
                                           occurred.

-------
                         TABLE XIII-4 (CONTINUED)

                      ANIMAL EXPERIMENTATION RESULTS
                        OF METHYL ALCOHOL EXPOSURE
Species
 Route of
 Exposure
  Dose
         Effect
 Ref-
erence
Monkeys
i.p.  inj
0.5 g/kg of
14 C-methyl
alcohol with
an equimolar
amount of
ethyl al-
cohol
The ethyl alcohol reduced
the oxidation of methyl
alcohol 90%.
  52
                           1.0 g/kg       The methyl alcohol was
                           14 C-methyl    oxidized at a rate of
                           alcohol and    37 mg/kg/hour between the
                           6.0 g/kg       first and fourth hour.  The
                           14C-raethyl     C02 formation was linear at
                           alcohol        the high dose; the oxidation
                                          rate was 47 mg/kg/hour which
                                          is a significant difference.
                                                            52
Rats
               1.0/kg 14C-    The oxidation rate of the     51
               methyl         methyl alcohol was 24 mg/kg/hr
               alcohol        for the first 28 hours.  At
                              the end of 36 hours 77% of
                              the methyl alcohol had been
                              oxidized to 14C-labled C02
                              and 24% was excreted unchanged
                              in approximately equal amounts
                              by the pulmonary' and combined
                              urinary and fecal routes.

-------
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47.   McCord  CP:   Toxicity  of  methyl  alcohol (methanol) following skin
      absorption and inhalation—A progress report.  Ind Eng  Chem  23:931-
      36, 1931

48.   Cooper  JR,  Felig  P:   The  biochemistry of methanol poisoning—II.
      Metabolic acidosis in the monkey.  Toxicol Appl  Pharmacol  3:202-09,
      1961.

49.   Renkonen  KO,  Teir H:  Studies on the local reactions of the skin to
      chemical compounds.  Ann Med Exp Biol Fenn 35:67-69, 1957

50.   Carpenter  CP, Smyth HF Jr:  Chemical burns of the rabbit cornea.  Am
      J Ophthalmol 29:1363-72, 1946

51.   Tephly  TR,  Parks  RE  Jr, Mannering GJ:  Methanol metabolism in the
      rat. J Pharmacol Exp Ther 143:292-300, 1964

52.   Makar  AB,  Tephly  TR,  Mannering  GJ:   Methanol  metabolism in the
      monkey.  Mol Pharmacol 4:471-83, 1968

53.   Clay  KL,  Murphy  RC, Watkins WD:  Experimental raethanol toxicity in
      the primate—Analysis of metabolite acidosis.  Toxicol Appl Pharmacol
      34:49-61, 1975

54.   Saha  AK,  Khudabaksh AR:  Chromosome aberrations induced by methanol
      in germinal cells of grasshopper, Oxya velox Fabricius.  J  Exp  Biol
      12:72-75, 1974

55.   Technology  Committee  (GA  Hedgecock, chtnn), Working Party (SJ Silk,
      chmn):  Chemical indicator tubes for measurement of the concentration
      of  toxic  substances  in air—First report Of a working party of the
      Technology Committee of  the  Britisn  Occupational  Hygiene  Society.
      Ann Occup Hyg 16:51-62,  1973
                                                                     »
56.   Smith  BS,  Pierce  JO:   The  use of plastic bags for industrial air
      sampling.  Am Ind Hyg Assoc J 31:343-48, 1970

-------
57.   Rogers GW:  Sampling and determination of methanol in air.  J Ind Hyg
      Toxicol 27:224-30, 1945

58.   Documentation  of  NIOSH Validation Tests, NIOSH contract No. CDC 99-
      74-45. .US Qapt of Health,  Education,  and  Welfare,  Public  Health
      Service',-'   Center   for   Disease  Control,  National  Institute  for
      Occupational Safety and Health, 1975, pp S59-1 to S59-9

59.   Feldstein  M,  Balestrieri  S,  Levaggi DA:  The use of silica gel in
      source testing.  Am Ind Hyg Assoc J 28:381-85, 1967

60.   Methyl alcohol Class 3, NIOSH Sampling Data Sheet #36.01.  US Dept of
      Health, Education, and Welfare, Public  Health  Service,  Canter  for
      Disease  Control,  National.  Institute  for  Occupational  Safety and
      Health, December 15, 1975, December 16, 1975, January 26, 1976

61.   Skoog  DA, West DM:  Fundamentals of Analytical Chemistry.  New York,
      Hblt, Rinehart and Winston, 1963, pp 667-69

62.   Deniges MG:  [Analytical chemistry-Study of methyl alcohol in general
      and especially in the presence  of  ethyl  alcohol.]  C  R  Acad  Sci
      (Paris) 150:832-34, 1910 (Fr)

63.   Ellvove  E:'  A note on the detection and estimation of small amounts
      of methyl alcohol.  J Ind Eng Chem 9:295-97, 1917

64.   Wright  LO:  Comparison of sensitivity of various tests for methanol.
   •   In'd Eng Chem 19:750-52, 1927

65.   Chapin RM:  Improved Deniges test for the detection and determination
      of methanol in Che presence of ethyl alcohol.  J Ind Eng Chem 13:543-
      45, 1921

66.   Jephcott  CH:   Determination  of methyl alcohol in the air.  Analyst
      60": 588-92,  1935

67.   Jansson BO, Larson BT:  Analysis of organic compounds in human breath
      by gas chromatography-mass spectrometry.  J Lab Clin  Med  74:961-66,
      1969

68.   Matsumura   Y:   The  adsorption  properties  of  active  carbon—II.
      Preliminary study on adsorption of various organic vapors  on  active
      carbon by gas chromatography.  Ind Health 3:121-25, 1965

69.   Baker RN, Alenty LA, Zack JF Jr:  Simultaneous determination of lower
      alcohols, acetone and acetaldehyde in blood by gas chromacography.  J
      Chromatogr  Sci 7:312-14,  1969

70.   Hurst  RE:   A  method  of  collecting  and  concentrating head space
      volatiles for gas-chromatographic analysis.  Analyst 99:302-05, 1974

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                 71.    Occupational  Health   Hazards   in Massachusetts  Industries—IV.  Wood
                       heel covering,  WPA  No.  65-14-6060.   Boston, Massachusetts  Department
                       of Labor and Industries,  Division of Occupational Hygiene, 1937

                 72.    Goss  AE, Vance GH:  Methanol vapors from duplicating machines may be
                       health hazard.   Ind Hyg Newsletter  8:15, 1948

                 73.    McAllister   RG:    Exposure  to  methanol  from spirit  duplicating
                       machines.  Am Ind Hyg  Assoc Q 15:26-28,  1954

                 74.    Dutkiewicz  T,   Blockowicz A:   [Evaluation of exposure  to methanol in
                       view of field studies.] Med Pr  18:132-41, 1967  (Pol)

                 75.    Methyl  alcohol  (methanol), AIHA Hygienic Guide Series.  Southfield,
                       Michigan, American  Industrial Hygiene Association,  1957

                 76.    Methanol,  Data  Sheet 407,  Revision   A.   Chicago, National Safety
                       Council, 1967,  pp 1-5

                 77.    American  Conference of Governmental Industrial  Hygienists, Committee
                       on  Industrial  Ventilation:    Industrial  Ventilation—A  Manual  of
                       Recommended Practice,  ed  13.  Lansing, Michigan, ACGIH, 1974

                 78.    American  National  Standards   Institute:  Fundamentals Governing the
                       Design and Operation of Local Exhaust Systems,  Z9.2-1971.  Mew  York,
                       American National Standards Institute Inc, 1971

                 79.    Bowditch  M, Drinker CK,  Drinker P,  Haggard HH,  Hamilton A:  Code for
                       safe concentrations of  certain  common toxic  substances  used  in
                       industry.  J Ind Hyg Toxicol 22:251, 1940

                 80.    Cook  WA:  Maximum  allowable concentrations of  industrial atmospheric
                       contaminants.  Ind  Med 14:936-46, 1945

                 81.    Methyl  alcohol  (methanol), AIHA Hygienic Guide Series.  Southfield,
                       Michigan, American  Industrial Hygiene Association,  1964

                 82.    American  Conference of Governmental Industrial  Hygienists, Committee
                       on Threshold Limit  Values:  Documentation of Threshold  Limit  Values
                       for  Substances  in Workroom   Air,  ed 3. Cincinnati, ACGIH, 1971, pp
                       155-56

                 83.    American  Conference   of  Governmental Industrial Hygienists:  .TLVs—
J Ji                     Threshold Limit Values for Chemical Substances  and  Physical Agents in
                       the Workroom Environment  with Intended Changes  for  1974.  Cincinnati,
                       ACGIH, 1974
                 84.   Winell  MA:    An  international  comparison  of  hygienic  standards  for
                       chemicals in the work environment.   Ambio  4:34-36,  1975,  p  2J
                                          /3.6-3-f

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85.   Smelyanskiy ZB, Ulanova IP:  [New standards for peraissible levels of
     coxic gases, fumes, dust in che air of  work  areas.]  Gig  Tru  Prof
     Zabol 5:7-15, 1959 (Rus)

86.   Czechoslovakia  Committee  of HAG (J Teisinger, chmn):  Documentation
     of MAC in Czechoslovakia.  Prague, June 1969, pp 114-15

87.   Elkins, HB, in Patty FA (ed):  Industrial Hygiene and Toxicology, rev
     ed 2; Toxicology (Fassett DW, Irish DD, eds).  New York, Interscience
     Publishers, 1963, vol 2, pp 1409-22,

88.   Methyl  alcohol  (methanol), AIHA Hygienic Guide Series.  Southfield,
     Michigan, American Industrial Hygiene Association, 1964
                                          •
89.   Methanol—Storage  and  Handling.   Wilmington,  Delaware, du Pont de
     Nemours Ca, W74, 10 pp

90.   National   Electric   Code   1975,   NFPA   No.   70-1975.    Boston,
     Massachusetts, National Fire Protection Association, 1975

91.   American  National  Standards—Occupational  and  Educational Eye and
     Face  Protection,  Z87.1.   New  York,  American  National  Standards
     Institute Inc, 1968
                               ' ~s ' ^
                           *•"'/_!)  /<*.

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                                                No. 127
S,S'-Methylene - 0,0,0',o'-Tetraethyl Phosphorodithioate

            Health and Environmental Effects
          U.S. ENVIRONMENTAL PROTECTION AGENCY
                 WASHINGTON, D.C.  20460

                     APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a survey of the  potential health
and environmental hazards from exposure  to  the  subject chemi-
cal..  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources,  this short profile
may not reflect  all available  information including  all  the
adverse health  and  environmental  impacts  presented  by  the
subject chemical.   This document has undergone scrutiny  to
ensure its technical accuracv.
                           i i~
                        *«*Tj /  /'

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                            PHOSPHORODITHOIC ACID,
              S,S'-METHYLENE,0,0,0',0'-TETRAETHYL ESTER  (ETHIQN)
                                   Summary

     The S,S'-methylene,0,0,0',0'-tetraethyl ester  of phosphorodithoic acid,
ethion, has  not shown mutagenic  effects  in mice or  teratogenic  effects  in
fowl.  Subcutaneous injection of the  compound into  atropinized chickens pro-
duced neurotoxic effects.  There is no available information on the possible
carcinogenic effects of ethion.
     Ethion has  shown  acute toxicity in  stonefly  naiads  at  a  96-hour LC5Q
range from 1.8 to 4.2 jjg/1.
                              J3.7-J

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 I.   INTROOUCTON
      0,0,0',0'-Tetraethyl-S,S'-methylene bisphosphorodithioate (CAS registry
 number 563-12-2),  also called  ethion,  is an  insecticide and  miticide made
 from phosphorus pentasulfide (SRI, 1976).  Ethion  has the following physical
 and chemical properties -(Windholz, 1576; FAQ, 1569):
               Formula:                       C9H22°4P2S4
               Molecular Weight:              384.48
               Melting Point:                 -12°C to -13°C
               Density:.                      1.22020
               Vapor Pressure:                Practically non-volatile at
                                             ordinary temperatures
               Solubility:                    Insoluble in water, soluble in
                                             organic solvents
               Consumption:                   0.7 million Ibs/year (SRI,  1576)
      Ethion is a  pre-harvest topical insecticide  used  primarily on  citrus
 fruits, deciduous  fruits,  nuts and cotton (SRI, 1976).   It  is  also used as a
 cattle  dip for ticks and  'as a  back-line treatment  for  buffalo  flies  (FAO,
 1969):
'll.   EXPOSURE
      A.    Water
           Pertinent data  could  not be  located  in  the available  literature.
 Water contamination from ethion  manufacturing may be minimal due  to  the com-
 mon use of industrial wastewater treatment plants (U.S. EPA, 1977).
     3.    Food
           Residues  on a variety of foods have  been reported (FAO,  1969).  A
 sampling shows the  residues on  fruits and vegetables .range from 10.4 ppm for
 raisins to less than 0.1 ppm for almonds.  The majority  are less  than  1 ppm.
 Treated cotton  showed no  residue  in  the seed.   Tea  at harvest showed 'resi-
 dues  of up to 7 ppm; since tea  is blended  prior to sale, residues are  lower


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when  consumed.   Lactating  cows  fed up to 20 ppm radioactive ethion  showed  no
residues  in  their  milk.   In  meat,  the  highest  radioactivity was  in the
liver;  however,  chemical analysis showed these  residues were not ethion but
metabolites.  , When  animals were  dipped,  residues  from' skin  absorption  of
ethion  were found  in the body fat.
      C.   Inhalation and Dermal
          Pertinent  data could not be  located in the available literature.
III.  PHARMACOKINETICS
      A.   Absorption
          Results  of acute toxicity studies in  animals indicate that ethion
is absorbed following oral  and dermal  exposure (Gaines,  1969).
      3.   Distribution
          Following  feeding of  dairy cattle with  ethion,  small  amounts of
the  unchanged compound  were  found  in milk and fatty  tissues  (Vettorazzi,
1976) .           '
      C.   Metabolism
          Rao and  McKinley (1969)  have reported that _in_ vitro  metabolism of
ethion  occurs through  oxidative desulfuration  of  the  compound by chicken
liver homogenates.
     D.   Excretion
          Pertinent  data could  not  be located  in the  available literature.
Based on studies  of  other organophosphorous insecticides, it may  be antici-
pated that  ethion metabolites  would  be  eliminated primarily  in the  urine
(Matsumura,  1975).
IV .   EFFECTS
     A.    Carcinogenicity
          Pertinent data could not be located in  the available literature.


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      8.   Mutagenicity
          Vettorazzi  (1976)  has  cited an  unpublished study which  found no
dominant lethal effects in mice administered ethion.
      C.   Teratogenicity
          Oral administration  of ethion (100 ppm)  to  chickens,  chukars,  and
quail failed  to  produce teratogenic or adverse reproductive  effects (Abbott
and Walker, 1972).
      0.   Other Reproductive Effects
          Oral feeding  of ethion  to  chickens,  chukar,  and quail  failed to
affect egg hatch (Abbott and Walker, 1972).
      E.   Chronic Toxicity
          Subcutaneous injection of atropinized chickens with 400  mg/kg eth-
ion produced  neurotoxic effects '(flaccid  paralysis) (Gaines, 1969).   Ethion
will produce anti-cholinesterase effects in  mammals (Vettorazzi,  1976),
V.   AQUATIC TOXICITY
     A.   Acute Toxicity
          Sanders  and  Cope  (1968)  observed  96-hour  LCCQ  values  ranging
from  1.3 to  4.2  ug/1 for stonefly naiads  (Pteronarcys califomica)  exposed
to ethion.
     3.   Chronic Toxicity,  Plant Effects and Residues
          Pertinent data could  not be  located in the available literature.
VI.  EXISTING GUIDELINES AND  STANDARDS
     A.   Human
          The World  Health  Organization (FAO, 1969) 'has established an  ADI
level of 0.005 mg/kg for ethion based  on cholinesterase inhibition  studies.
  '   9.   Aquatic
          Pertinent data could  not be  located in the available literature.
                                     / S\a
                                    / _> I Q  *
                                     7
                                 I2.T-B

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                                   REFERENCES
 Abbott,  U. and  N.  Walker.   1972.   Effects of  pesticides  and  related  com-
 pounds  on  several  avian  species,  chemistry and  toxicology of agricultural
 chemicals.   Summary  Report 1971.   Food Protection  and  Toxicology Center.,..
 University  of California at Davis,  p. 9.

 Food  and  Agriculture  Organization/World  Health Organization.  1969.  Evalua-
 tions of  some pesticide residues  in  food.   The monographs  FAO/WHO/Pl:1968/-
 M/9/1.

 Gaines, T.   1969.  Acute  toxicity  of  pesticides.   Toxicol. Appl. Pharmacol.
 14: 515.

 Matsumura,  F.   1975.   Toxicology  of Insecticides.  Plenum  Press,  New York.
 p. 223.

 Rao,  S. and w.  McKinley.   1969.   Metabolism of organophosphorus insecticides
 by liver homogenates  from  different species.  Can. Jour. Biochem.  47: 1155.

 Sanders, H.O. and 0.8. Cope.   1968.   The relative toxicities of several pes-
 ticides  to  naiads   of  three  species  of  stoneflies.   Limnol.  Oceanogr.
 13: 112.

 Stanford  Research Institute.  1976.   Chemical Economics  Handbook,  Insecti-
 cides..

 U.S.  EPA.   1977.  Industrial  process  profile for  environmental use:  chapter
 8, pesticides industry.  U.S. Environ. Prot. Agency, U.S. NTIS PB 266 225.

 Vettorazzi, G.  1976.  State  of the art  of  the toxicological evaluation car-
 ried out by the joint FAO/WHO  meeting  on pesticides residues.   II.  Carbamate<
 and  organophosphorus  pesticides   used  in   agriculture  and public  health.'
Residue Reviews.  63:  1.

Windholz,  M.  (ed.)    1976.  The Merck Index, 9th  ed.   Merck and Co.,  Inc.,
Rahway, New Jersey.

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                                     No. 128
        Methyl Ethyl Ketone

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980
              -J *s* A -
               J  y ^^
          /a ?-

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                         . DISCLAIMER
     This report represents  a survey of the potential  health
and environmental hazards from exposure  to  the  subject  chemi-
cal..  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources,  this short profile
may not reflect  all available  information  including all  the
adverse health  and  environmental  impacts  presented  by  the
subject chemical.   This document has undergone scrutiny  to
ensure its technical accuracy.
                              > r*>'i.it--
                             y j ^i

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                                                    SJ-27-10








                     Methyl Ethyl Ketone






I.   INTRODUCTION




     Methyl 'ethyl ketone or (MEK) as it is commonly referred




to i3 a clear, colorless, volatile liquid (VP of 100 mm at 25°C)




with a molecular weight of 72.12.  It has a melting point of




-86.35°C and a boiling point of 76.6°C.  It is very soluble




in water (25.5 g/loo at 2 percent) and soluble in all




proportions in alcohol, ether, acetone and benzene.2  it is




also highly flammable (22"? - open cup).3




     MEK is produced and used as a solvent in nitrocellulose




coatings and vinyl films; in the synthesis of colorless




resins; in the manufacture of smokeless powder;  in paint




removers, cements, adhesives, and cleaning fluids; in printing




industry; as a catalyst carrier; in lube oil dewaxing and in



acrylic coatings.^




II.  ROUTES OP EXPOSURE




     MEK is rapidly absorbed through Che skin by inhalation.




III. PHA&MACOKINETICS




     MEK occurs in trace amounts in normal human urine and




may have a dietary origin.^  Most probable precursor is




  - methylacetoacetic acid.^-




     Urine of rabbits exposed to MEK reported to contain




glucuronide of 2-butanol.3-




IV .  EFFECTS ON MAMMALS



     The chief effect of MEK is narcosis but is  also a strong*



irritant of the mucous membranes of the eyes and nose.  The




oral LD50 for rats is 3.3 g/kg and the inhalation LC50 is
                            /2.S--.3

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around 700 ppm.l




     Repeated exposure of guinea pigs for 12 weeks to 235 ppm




caused no symptoms.^




     Lethal doses in animals caused marked congestion of




internal organs and slight congestion of brain.  Lungs showed




emphysema (see Table 1) .




     Slight throat irritation in humans occured at 100 ppm




and in eyes at 200 ppm.




     Dermatoses among workers having direct contact and



exposed to vapors are not uncommon.  Some workers complained




of numbness of fingers and arms.-'-

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                           Table 1

          Effects of Methyl Ethyl Ketone on Animals
           Concentration/
              Duration
                              Animal
Methyl
 Ethyl
  Ketone
33,000-100,000 ppm/200 tain.   Guinea Pigs
           3,300 ppm/810 min.

           1,125 ppm/24 hr/3,55d
           1,126 or 2,618 ppm/7 hr/d
             on d 6-15 of gestation
                              Guinea Pigs

                              Rats
                              Pregnant
                                Rats
Effects
Gasping, death,
emphysema, slight
congestion of the
brain, marked
congestion of the
systemic organs
especially the
lungs and corneal
opacities

No abnormal signs

No evidence of
peripheral neuro-
pathy

Embryo toxicicy,
facotoxicity and
possible terato-
genicity

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                          References
1.   Toxicity and Metabolism of Industrial Solvents.

2.   Ketonic Solvents, Open File Report, Working Draft prepared
     by Clement Associates, Inc., September 19, 1978.

3.   Sax, N. Irving, Dangerous Properties of Industrial Materials,
     Fourth Edition, '1975, Van Nostrand Reinhold, New York,
     New York  10001

4.   Patty, F. A., Schrenk, H. H., Yant W. P.: Acute Respone of
     guinea Pigs to Vapors of Some New Commercial Organic
     Compounds—VIII.  Butanone.  U.S. Public Health  Pep 50:
     1217-28, 1935.

5.   Spencer, P. S., Schaumburg, H. H. : Feline Nervous System
     Response to Chronic Intoxication With Commercial Grades
     of Methyl n-Butyl Ketone, Methyl Isobutyl Ketone, and
     Methyl Ethyl Ketone.  Toxicol. Appl. Pharmacol. 37:301-11,
     1976.

6.   Griggs, J. H. , Waller, E. M. , Palmisano, P. A., Niedermeier,
     W.: The Effect of Noxious Vapors on Embryonic Chick Develop-
     ment, Ala. J Med. Sci.  8:342-45, 1971.

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                                      No. 130
       Methyl Methacrylate

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

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                             METHYL METHACRYLATE
                                   Summary

     Oral or skin  painting studies in  rats  have failed to show carcinogenic
effects of administration of methyl methacrylate.  Implantation of  the  com-
pound in mice also failed to produce tumors.
     Exposure of  rats  to a mixture  of chloroprene and methyl  methacrylate
produced an increase in lymphocyte chromosome aberrations.   Increased  chrom-
atid breaks and chromosome breaks have been reported in workers exposed to
this same chemical mixture.
     Teratogenic effects  (hemangiomas)  have been reported  following  intra-
peritoneal administration of methyl methacrylate to pregnant rats.  Inhala-
tion exposure of pregnant rats  to an  acrylic cement containing  methyl  metha-
crylate failed  to  produce  significant teratogenic effects.
     Ninety-six hour LC5Q values for  four  species of  fish range  from 159
to 368 ppm.   Inhibition of cell multiplication of an alga begins at 120 ppm.
                                      s
                               730-3

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 I.    INTRODUCTION
      Methyl methacrylate,  CAS  registry  number  80-62-6,  is  a  colorless,
 clear,  volatile  liquid.   It is made from acetone cyanohydrin which  is  hydro-
 lyzed in  sulfuric acid to yield methacrylamide sulfate, which is  then  treat-
 ed with methanol to yield methyl methacrylate.   It has the following  physi-
 cal  and  chemical  properties  (Windholz,  1976;  Hawley,  1971;  Weast,   1972;
 Verschueren, 1977):
              Formula:                  C5H802
              Molecular Weight:         100.12
              Melting Point:.            -48.2°C
              Boiling Point:            101°c
              Density:                  0.944020
              Vapor Pressure:           28 torr 1 20°C
              Solubilityr               Sparingly soluble in water, miscible
                                        in alcohol, benzene, ether, etc.
              Production:               706 million Ibs (1973) (Gruber,  1975)

     virtually all  the methyl  methacrylate produced in this  country is used
 for  polymers,  e.g.,  surface  coating  resins and  plastics (plexiglass, lu-
 cite), ion exchange resins, dentures, etc.
 II.  EXPOSURE
     A.   Water
          According to Gruber  (1975),  about  1.8  g of methyl methacrylate per
kilogram final product  (methyl methacrylate) is present in wastewater.  The
amount  of  methyl methacrylate entering  domestic, water supplies  is probably
small.
     B.   Food
          Polymethyl methacrylate  is  used  for  food  storage.   A  very  .small
amount of  residual monomer may  migrate into food from the  polymer.
                                      If

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     C.    Inhalation
           Fugitive  emissions  from  production,  storage,  and  transportation
probably  constitute the  only major  sources  of methyl  methacrylate in  the
air.   The concentration would most  likely  be highest in production facili-
ties.  Production was estimated  to be 7.9 million pounds in 1974  (U.S.  EPA,
1976).   A 550-million  pound-per-year production facility  with 0.5  percent
loss emits 39.6 grams of  methyl  methacrylate  per second.   If this  is consi-
dered  to be a  virtual  point  source,  the downwind concentration 500 meters
avray would be 1.5 ppm one-hour average (U.S. EPA,  1976).
     0.    Dermal
           Pertinent data could not be located in the available literature.
III. PHARMACQKINETICS
     A.    Absorption and Distribution
          Pertinent data could not be located  in the available literature.
     3.   Metabolism
          Bratt and Hathway  (1977)  found that  up to 88  percent of a single
methyl(^C) methacrylate  dose  of  5.7  mg/kg  body weight  was expired  as C07
within 10  days.  Neither- the route of administration nor the specific label-
ling of  the propylene residue changed this  value.   Small amounts of several
metabolites  were  excreted   in  the  urine,   including  i4C-methylmalonate,
^•4C-succinate,  2-formylpropionate,  and possibly ^C-/-hydroxybutyrate.
          Corkill,  et  al.  (1976)  found  that  the  disappearance of methyl
methacrylate in  human  blood  _!£ vitro  showed  a  first order dependence on
methyl methacrylate concentration.   The calculated  half-life was  20 to 40
minutes,   irrespective of  the  sex or  age  of the blood donor.   More than 40
percent of the initial  dose  of methyl methacrylate  was converted  to metha-
crylic acid within  90  minutes.
                                no-s

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      C.    Excretion
           Pertinent data could not be located in the available literature.
 IV.   EFFECTS
      A.    Carcinogenicity
           The  International Agency  for Research on Cancer (IARC, 1979) has
 evaluated  the  available data and concluded that there is not  enough  informa-
 tion  to determine  the  potential  carcinogenicity of  methyl methacrylate  to
 humans.  Borzelleca, et  al.  (1964)  observed no  treatment-related tumors  in
 male  and female Wistar rats administered  6,  60, or 2,000 mg/1 methyl metha-
 crylate  in their drinking water  for two years.  Oppenheimer,  et al.. (1955)
 found  no  local  tumors  in ten  Wistar rats painted  with methyl methacrylate
 three  times  per week for  four months and"observed for their entire life span.
           Another  study,  by Spealman,  et  al.  (1945),  in  which male  and
 female mice  received implants consisting of  0.075  gm of methyl methacrylate
 in a gelatin capsule also yielded negative results.
     8.    Mutagenicity                                                   '
           The  only  data available on the mutagenic  effects of methyl metha-
 crylate  are  two studies involving exposure  to  a mixture of  chloroprene  and
 methyl methacrylate (Bagramjan, et  al.  1976; Bagramjan and  Babajan, 1974).
 In  both  studies,  an increased frequency  of  chromosomal  aberrations  were
 found  in rats   exposed  to the mixture.  Bagramjan,  et al.   (1976)  also  mea-
 sured  a  significant increase  in  chromatid breaks  and chromosome breaks  in
 the lymphocytes of  workers  exposed  to  a  mixture of chloroprene  and methyl
methacrylate.
     C.   Teratogenicity
          Singh, et al.  (1972a,b)  and  Autian  (1975) injected  intraperito-
neally three groups  of  pregnant Sprague-Oawley  rats  with methyl methacrylate
at doses of  0.1, 0.2, or 0.4 g/kg body weight on days 5, 10,  and  15  of ges-

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 cation.   In  animals  administered  the  two  higher  doses,  a significantly
 greater  number  of hemangiomas were seen at various  sites.   All three groups
 exhibited  reduced  fetal, weights,  but no  significant increase  in skeletal
 defects  was observed in any group.
          McLaughlin, et'al.  (1978)  exposed pregnant mice to a vapor concen-
 tration  of 1,330  ppm  methyl methacrylate (as acrylic  cement,  Simplex  p) for
 two  hours  two times per  day  for days 6  through 15 of  gestation.   No feto-
 toxic'or teratogenic effects  were noted  other than a  slight  decrease  in the
 average  fetal weight.
     0.   Other Reproductive Effects
          Pertinent data could not be located in the available literature.  .
     E.   Chronic Toxicity
          Spealman,  et   al.   (1945)  conducted  a   series   of   sub-chronic
 inhalation  ex-  periments  involving guinea  pigs, and  dogs.   Guinea  pigs
 exposed  to  39.0 mg/1  methyl  methacrylate  for  two hours per day  for  three
 days exhibited  significant liver  degeneration,  while dogs  exposed to  46.3
 mg/1 methyl  methacrylate  for two  hours  per day for  8 to 15  days  exhibited
 liver degeneration and tubular degeneration of the kidneys.
          Sorzelleca,  et  al.  (1964)  administered 6,  60,  and  2,000 ppm of
 methyl methacrylate in drinking  water to male and  female  rats for  a  period
 of two years.   Weight  gain was decreased for the first  few weeks in animals
 given the highest dose.  No changes  in hematological values or urine concen-
 trations of  protein and  reducing agents were  noted..  Females  receiving the
 highest dose level exhibited an increase  in kidney to 'body weight  ratios.
          Blagodatin,  et  al.  (1970) reported symptoms of  headache, pain in
 the- extremities,  fatigue,  sleep  disturbance,  loss of  memory, and  irritabi-
 lity in  152 workers exposed to concentrations of 0.5 to  50  ppm methyl metha-
crylate.   Most of the  workers  had been employee  for  longer than 10 years.
                                    ]^^£r
                                  "f J J U "'
                                 S30-7

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      F.    Acute Toxicity
           No  detectable  acute  effects were  noted  in workers  employed  in
 manufacturing  polymethyl methacrylate  sheets  (Cromer and Kronoveter,  1976).
 The  airborne concentrations of methyl methacrylate  varied  from  4  to  49  ppm.
 V.    AQUATIC TOXICITY   .
      A.    Acute Toxicity
           Pickering  and  Henderson  (1966)  observed  the  fallowing 96-hour
 ^50 va^ues  f°r  fish  exposed  to  methyl  methacrylate:   fathead   minnow
 (Pimephales. promelas)  - 159  ppm  in  soft water  (20  mg/1);  fathead  minnow -
 311  ppm  in hard water (360  mg/1); bluegill  (Lepomis macrochirus) - 357 ppm
 in  soft water  (20  mg/1);  goldfish  (Carassius auratus)  - 277 ppm  in soft
 water (20  mg/1);  guppies (Lebistes retieulatus) - 368  ppm in soft water (20
 mg/1).
     B.   Chronic Toxicity
          Pertinent data could not be located in the  available  literature.
     C.   Plant Effects
          Inhibition of  cell multiplication  of the  alga, Microcystis aerugi-
 nosa, by methyl methacrylate begins at 120 ppm (Bringmann and Kuhn 1976).
 D.   Residues
          Pertinent data could not be located in the available literature.
 VI.  EXISTING GUIDELINES AND STANDARDS
     A.    Human
          Guidelines have been  established for exposure to methyl methacry-
 late by  the American  Conference  of Governmental  Industrial  Hygienists  and
QSHA.  Both the TLV and the  federal standard have  been set  at  100 ppm (or
410 mg/rn3)  (ACGIH,  1977; 29 CFR  1910).

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     B.   Aquatic
          No guidelines  have been established for the  protection  of aquatic
organisms from acute  or  chronic methyl methacrylate toxicity because  of the
lack of pertinent data.

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                                   REFERENCES
 American Conference of Governmental  Industrial  Hygienists.   1977.  Threshold
 limit  values  for  chemical  substances  and physical  agents in  the workroom
 environment.  Cincinnati, Ohio.

 Autian,  J.'   1975.   Structure-toxicity  relationships  of  acrylic  monomers.
 Environ. Health Perspect.  11: 141.

 Bagramjan, S.B.. and E'.A. Babajan.   1974.   Cytogenetic study of the mutagenic
 activity of chemical  substances  isolated  from Nairit  latexes  MKH and LNT-1.
 (Russ.) Biol.  Zh. Arm.  17: 102.

"Bagramjan, S.B., et al.   1976.  Mutagenic effect  of  small  concentrations'of
 volatile substances emitted from polychloroprene  latexes LNT-1 and MKH,  dur-
 ing their combined uptake by the animal.   (Russ.) Biol. Zh.  Arm.  19: 98.

 Blagodatin, V.M.,  et  al.   1970.   Issues  of  industrial hygiene  and occupa-
 tional  pathology  in  the manufacture  of  organic glass.   (Russ.)  Gig.  Tr.
 Prof. Zabol.  14: 11.

 Borzelleca, J.F.,  et al.   1964.    Studies on the chronic oral  toxicity  of
 monomeric ethyl aerylate and methyl  methacrylate.   Toxicol. Appl. Pharmacol.
 6: 29.

 Bratt,  H.  and  O.E.  Hathway.   1977.   Fate of  methyl methacrylate  in  rats.
 Br. Jour. Cancer.  36: 114.

 Bringmann,  G.  and R.  Kuhn.   1976.   Vergleichende Befunde  der Schadwirkung
 wassergefahrdender Stoffe genen Bakterien  (Speudomonas putida) und Blaualgen
 (Microcystis aeruginosa).  Nwf-Wasser/Abwasser,  (117)  H.9.

 Corkill, J.A., et al.   1976.   Toxicology of methyl methacrylate:  The rate of
 disappearance of methyl methacrylate in  human blood i.n vitro.   Clinica Chim-
 ica Acta.  68: 141.

 Cromer, J.  and  K.  Kronoveter.  1976,   A study of methyl methacrylate  expo-
 sures and  employee health.   National  Institute  for Occupational  Safety and
 Health, Cincinnati, Ohio.  DHEW 77-119.

 Gruber,  G.I.    1975.   Assessment   of industrial  hazardous  waste  practices,
 organic chemicals,  pesticides, and explosive industries.  TRW  Systems Group,
 NTIS PB-251 307.

 Hawley, G.G.  (ed.)  1971.   The Condensed Chemical Dictionary.   8th ed., Van
 Nostrand Reinhold Co.,  New York.

 International  Agency for Research  on Cancer.   1979.   IARC  monographs on the
 evaluation of  the carcinogenic risk of chemicals  to humans.   Vol.  19,.Methyl
 methacrylate:  187.

 McLaughlin, R.E., et al.  1978.  Methyl  methacrylate:  a study  of teratogeni-
 city and  fetal  toxicity  of  the vapor in the mouse.   Jour.  Bone  Jt.  Surgery
 Am.  Vol. 60A: 355.

-------
Oppenheimer,  3.S.,  at al.   1955.   Further  studies  of polymers  as carcino-
genic agents in animals.  Cancer Res.  15: 333.

Pickering,  Q.H.  and C. Henderson.   1966.   Acute toxicity  of some important
petrochemicals to fish.  Jour. Water Poll. Con. Fed.  38: 1419.

Singh,  A.R.',  et  al.   1972a.  Embryonic-fetal toxicity and  teratogenic ef-
fects of a group of methacrylate esters in rats.  Jour. Dent. Res.  51: 1532.

Singh,  A.R.,  et al.   1972b.  Embryo-fetal toxicity  and  teratogenic effects
of  a group of  methacrylate  esters  in  rats  (Abstract  No.  106).   Toxicol.
Appl. Pharmacol.  22: 314.  .

Spealman,  C.R.,  et  al.    1945.   Monomeric  methyl  methacrylate:  Studies • on
toxicity.  Ind.  Med.  14: 292.

U.S. EPA.   1976.   Assessment of methyl methacrylate  as  a  potential air pol-
lution problem.   U.S. Environ. Prat.  Agency, NTIS P8-258 361.

verschueren, K.   1977.   Handbook of Environmenal Data, on  Organic Chemicals.
Van Nostrand Reinhold Co., New York.

Weast,  R.C.   1972.   Handbook of Chemistry  and Physics.   53rd  ed., Chemical
Rubber Company,  Cleveland, Ohio.

Windholz, M.  (ed.)   1976.   Merck Index.  9th  ed.,  Merck  and Co., Inc., Rah-
way, New Jersey.
                             I  3 o-lt

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                                      No. 131
            Naphthalene

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

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                               NAPHTHALENE
                                 Summary


     Naphthalene is present in ambient water as well  as  drinking water.
Naphthalene can be absorbed by any route, although  the efficiency of
absorption has not been determined.  The  toxicological properties are due
to the formation of highly reactive metabolites.  Chronic  exposure produces
cataracts, hemolytic anemia, and kidney disease.  Naphthalene  can cross
the placenta and produce these effects on newborns.   Naphthalene has been
found to be nonmutagenic in several microsomal/bacterial assay systems.
Chronic toxicity  studies of naphthalene  have shown it to  be noncarcinogenic.
     Naphthalene has been shown to be acutely toxic in freshwater fish
                                                                          *
with LC5Q values of 150,000 ug/1 being reported in  one static  bioassay.
Freshwater invertebrates were more sensitive with LC5Q values  of 8,570
jig/1,  as were marine fish with LC5Q values ranging from 2,350 to 2,600  .

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INTRODUCTION



     This profile is based on the Ambient Water Quality Criteria Document



for Naphthalene (U.S. EPA, 1979).



     Maphthalene (CigHs; molecular weight 128.16) is a bicyclic, aromatic



hydrocarbon which in a pure grade, forms a white crystalline solid at



room temperature  (Windholz, 1976). Pure naphthalene has a melting point



of 80.2°C, a boiling point of 217.96°C (Manufacturing Chemists Assoc.,



1956) and a  vapor pressure of 0.0492 mm Hg at 19.8°C (Gil'denblat, et



al. 1960).  Naphthalene is water soluble, with solubility ranging from



30,000 ug/1 (Mitchell, 1926) to 40,000 ug/1 (Josephy and Radt, 1948).



Naphthalene vapor and dust can form explosive mixtures with air (Windholz,



1976).  Naphthalene is used as an intermediary in the production of dye



compounds, in the formulation of solvents, lubricants and motor fuels,



and as a feedstock in the synthesis of phthalic anhydride.   Naphthalene



is also used directly as a moth repellant, insecticide,  antihelminthic ,



vermicide, and an intestinal antiseptic (U.S. SPA, 1979).   In. 1974,



production of naphthalene was approximately 2.9 x 10^ metric tons (U.S.



SPA, 1976).








II.  EXPOSURE



     A.   Water



          The two major sources of naphthalene in the aquatic environment



are from industrial effluents and from oil spills.  The  final effluents



of sewage treatment plants receiving discharges from these  industrial



facilities have been noted to have up to 22 ug/1 naphthalene, while. natural



waters have up to 2.0 ug/1, and drinking water supplies  have up to 1.4



     naphthalene (U.S. EFA, Region IV, unpublished data).
                                            131-1

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     B.    Food




          The U.S. EPA (1979) has estimated the weighted average  bio-




concentration factor for naphthalene to be 60 for the edible  portions of



fish and shellfish consumed by Americans.  This estimate was  based on



octanol/water partition coefficients.



     C.    Inhalation




          In the ambient air, inhalation of naphthalene is negligible



with vapor concentrations ranging from 0.00005 to 0.0001 ug/m3 and




particulate concentrations ranging from 0.000003 to 0.00025 ug/m^



(Krstulovic, et al. 1977).  Industrial exposure can range from 0.72 yig/m^



to 1.1 x 10^ ug/m3  in the vapor phase (Bjrseth, et al., 1978b; Robbins,



1951) and from 0.09 ug/m3 to 4.40 pg/m^ in particulates (Bjrseth, 1978a,



1978b).   Naphthalene has also been found in cigarette smoke condensate



(Akin, et al. 1976).



III.  PHARMACOKINETICS



     A.    Absorption



          Little detailed information is available on the absorption of



naphthalene in man or animals. Adequate amounts of naphthalene can be



absorbed when ingested as a solid, or by inhalation, to cause significant



toxicity (U.S. EPA, 1979).  Absorption seems to be facilitated if naphthalene



is dissolved in oil (Solomon, 1957), and hindered if naphthalene, is bound




to protein (Sanborn and Malins, 1977).



     B.    Distribution



          Naphthalene distributes widely after absorption.  In mallards,



the relative distribution of naphthalene was as follows:  greatest* in



skin, followed by liver,  brain, blood, muscle, and heart (Lawler, et al.



1978).

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     C.   Metabolism



          Naphthalene is first metabolised by hepatic mixed-function




oxidases to an epoxide, which is an obligatory step  in  the metabolism  of




naphthalene.  Further metabolism can occur leading to the formation of a




variety of compounds.  Most of these compounds are enzymatically conjugated




with glucuronic acid or sulfate.  During metabolism  a number of highly




reactive compounds are formed such as  1,2-dihydroxynaphthalene and 1,2-




naphthoquinone (U.S. EPA, 1979).



     Naphthalene metabolites undergo further conversions in the eye.




This multi-step pathway can lead to the formation of 1,2-naphthaquinone




which can irreversibly bind to lens protein and amino acids (Van Heyningen



and Pirie, 1966).




     D.   Excretion




          Naphthalene has not been identified in urine  after absorption. •




With sufficient absorption of naphthalene to result  in  toxicity to an  18




month old infant, Mackell, et al. (1951) noted metabolites of naphthalene




in the urine that were still identifiable two weeks  after exposure but




which had disappeared 18 days after exposure.



     1-Napthol is the predominant spontaneous decomposition product of



the epoxide of napohthalene.  1-Napthol is excreted  unchanged as well  as



congugated with glucuronic acid or sulfate prior to  excretion.  The finding



of 1,4-nathoquinone in the urine of a child poisoned with naphthalene



(Mackell, at al. 1951) suggests that 1-napthol can also be further oxidized



in mammals (Cerniglia and Gibson, 1977).



IV.  EFFECTS



     A. Carcinogenicity




          In attempts to demonstrate its carcinogenicity, naphthalene  has



Seen given orally, subcutaneously, implanted in the  bladder, and painted

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on the backs of a number of animal species  (U.S.  EPA,  1979).   In  these




experiments naphthalene caused no increase  in tumor  formation.  Two




experiments have produced increases in lymphosarcoma and  lymphatic




leukemia after treatment with coal tar derived naphthalene.  The  first of




these studies (Knake, 1956) was complicated by the presence of  10 percent




impurities in the naphthalene and the painting of the  injection site with




carbolfuchsin, a known experimental carcinogen, prior  to  injection.  In




the second study (Knake, 1956) where excess leukemia was  noted, naphthalene




was dissolved ,in benzene, a known human leukemogenic agent,  and painted on




the backs of mice. Benzene treatment resulted in  no  leukemia.  Skin




papillomas have been produced on mice following painting  with 1,4— naththa-




quinone, a metabolite of naphthalene (Takizawa, 19^0).  Also, Pirie (1968)




noted abnormal mitotic figures in metaphase and cell overgrowth in the
                                                                        *



epithelial cells of the lens of rabbits given 1 g/kg/day  of naphthalene




by gavage.




     8.   Mutagenicity




          Naphthalene has been found to be nonmutagenic in several microsomal/




bacterial assay systems (McCann, et al. 1975; Kraemer, et al. 1974).




     C.   Teratogenicity




          Pertinent data could not be located in  the available literature.




     D.   Other Reproductive Effects




          Naphthalene or its metabolites can cross the placenta in sufficient




amounts to cause fetal toxicity (Zinkham and Childs, 1958; Anziulewicz,




et al.  1959)-   When a metabolite of naphthalene," 2-naphthol, was admin-




istered to pregnant rabbits, their offspring were born with cataracts and




evidence of retinal damage (Van der Hoeve, 1913).

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     E.   Toxicity

          Oral administration of two percent naphthalene or 2-napthol  to

rats for at Least 60 days resulted in the development of cataracts

(Fitzhugh adn Busckke, 19^9).  Van Heyningen and Pirie  (1976) dosed  rabbits

daily by gavage with 1000 mg/lcg of naphthalene  for a maximum of 28 days.

Lens changes developed after the first dose, and retinal changes developed

after the second dose.  Rabbits fed 1000 mg/kg/day developed cataracts

between day 3 and 46.  Topical application of a 10 percent solution  in oil

to the eyes of rabbits did not produce cataracts after  a period of 50  days.

Intraperitoneal injection of 500 mg/kg of naphthalene in an oily solution

produced weight loss over a period of 50 days (Ghetti and Mariani, 1956).

Hemolytic anemia- with associated jaundice and occasionally renal disease

from precipitated hemoglobin has been described in newborn infants,     4

children and adults after exposure to naphthalene by ingestion, inhalation,

or possibly by skin contact (U.S. EPA, 1979).   The extent or duration  of

exposure was not given.  Mahvi, et al. (1977) noted a dose related damage

to bronchiolar epithelial cells in mice given intraperintoneal injections

of naphthalene in corn oil.  Bronchiolar epithelial changes were not

noted in two control groups.  The authors noted minor bronchiolar epithelial

changes in the treated group receiving 67.4. mg/kg of naphthalene.

Those mice receiving higher doses (128 and 256  mg/kg of naphthalene)

developed reversible necrosis of bronchiolar cells.

     F.   Other Relevent Information

          Alexandrov and Frayssinet (1973) demonstrated that naphthalene

administered intraperitoneally to rats could inhibit the mixed-function
                                                                  »
tnicrosomal oxidase enzyme system, and could also inhibit the induction of

these enzymes by 3-
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V.   AQUATIC TOXICITY




     A.   Acute




          For the freshwater mosquitofish  (Gambusia affinis)  a  96-hour




ststic bioas'say provided an LCcg value of  150,000 ug/1  (Wallen,  et  al.




1957), while the freshwater cladoceran (Daphnia magna)  was  shown to have




an US-hour LC5Q value of 8,570 ug/1 (U.S.  EPA, 1978).   Marine organisms




tended, to be somewhat more sensitive to naphthalene with an 24-hour static




^050 value of 2,400 ug/1 for the sheepshead minnow (Cyprinodon  variegatus).




Two 24-hour static LC5Q values of 2,500,' 2,600 were obtained  for two




species of marine shrimp, (Penaeus aztecus) and (Palaemonetes pugio),




respectively (Anderson, et al. 1974).  A 96-hour LC5Q value of  2,350 ug/1




was obtained for grass shrimp (Palaemonetes pugio) (Tatem,  1976).




     B.   Chronic Toxicity




          A single embryo-larval test on the fathead minnow (Piaephales




promelas) stated that no effects were observed at concentrations as high




as 440 ug/1 (U.S. EPA, 1978).




          Data pertaining to the chronic toxicity of naphthalene for any




marine species could not be located in the available literature.




     C.   Plant Effects




          A 48-hour EC5Q value of 33,000 ug/1 for reduced cell  numbers




has been reported for the freshwater algae (Chlorella vulgaris)  exposed




to naphthalene.  Data pertaining to the effects of naphthalene  to marine




plants could not be located in the available literature.




     D.   Residues




          Using the octanol/water partition coefficient of  2,300  for




naphthalene, a bioconcentration factor for aquatic organisms  with an 8




percent lipid content has been estimated as 210.   Bioconcentration

-------
factors determined for niarine invertebrates ranged from 50 to 60 in the



marine copepod Calanus helgolandicus after one day (Harris, et al. I977a,



I977b) .to 5,000 in the copepod Eurytemcra affinis, after nine days,



(Harris, et al. I977b) indicating that equilibrium may not occur rapidly.



Bioconcentration factors of 32 to 77 after 1 to 24 hours were reported



for these 3 species of marine fish and one species of mussel (Lee, at al.



1972a; 1972b).




VI.  EXISTING GUIDELINES AND STANDARDS



          Neither the human health nor aquatic criteria derived by U.S.



SPA (1979'), which are summarized below, have gone through the process of



public review; therefore, there is a possibiliity that these criteria



will be changed.




     A.   Human



          The Occupational Safety and Health Administration standard for



exposure to vapor for a time-weighted industrial exposure is 50 tng/m3.



The American Conference of Governmental Industrial Hygienists (ACGIH,



1971) threshold limit value is 75 tng/m3, while at present the ACGIH also



suggests a maximum 15 minute exposure value of 75 mg/m3 (ACGIH, 1978).



The acceptable daily intake for naphthalene is 448 ^ug/day for a 70 kg



person.  The U.S.  EPA (1979) draft ambient water criterion  for



naphthalene is 143 ug/1.



     B.   Aquatic



          Criterion can not be derived for naphthalene for either fresh-



water or marine organisms, because of the lack of sufficient toxicological



data.
                               l-3/r/O

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                                  NAPHTHALENE

                                  REFERENCES
Akin,  F.J.,   at  al.   1976.   Identification  of polynuclear  aromatic hydro-
carbons in- cigarrette  smoke  and their importance as tumorigens.  Jour. Natl.
Cancer Inst.  57: 191.

Alexandrov,   K.   and-  C.  Frayssinet.   1973.   In   vitro  effect  of  some
naphthalene-related  compounds  on  aryl hyrocarbcn  (benzo(a)pyrene) hydroxy-
lase.  Jour.  Natl. Cancer Inst.  51: 1067.

American  Conference of Governmental  Industrial -Hygienists.    1971.   Docu-
mentation  of  the  threshold  limit values for substances in workroom air.  3rd
ed.  Cincinnati, Ohio.

American Conference  of Governmental Industrial Hygienists.  1978.  Threshold
limit  values  for  chemical  substances  and physical  agents in  the workroom
environment with intended changes for 1978.  Cincinnati, Ohio.

Anderson,  J.w.  et  al.   1974.   The  effects  of  oil  on  estuarine animals:
Toxicity uptake  and depuration,  respiration.   In;  pollution  and physiology
of marine  orgasnisms.  Adademic Press.' New York
                                                                           •<«
Anziulewicz,  J.A.,  et  al.   1959.  Transplacental naphthalene poisoning.  Am.
Jour. Obstet. Gynecol.   78: 519.

Bjorseth,  A.  et al.   1978a.    Polycyclic  aromatic  hydrocarbons  in  the work
atmosphere. II.  Determination  in  a coke  plant. Scand. Jour.  Work. Environ.
Health.  4: 212.

Bjorseth,  A.  et al.   1978b.    Polycyclic  aromatic  hydrocarbons  in  the work
atmosphere. I.  Determination in an aluminum  reduction  plant.   Scand. Jour.
Work Environ. Health.  4: 212.

Cerniglia,   C.E.  and  D.T.  Gibson.   1977.   Metabolism  of  napthalene  by
Cunninqhamella eleoans.  Appl. Environ. Microbiol.   34: 363.

Fitzhugh,  O.G.  and  W.H. Buschke.   1949.   Production of cataract in  rats by
beta-tetralol and other derivatives of naphthalene.   Arch.  Ophthal.  41: 572.

Ghetti, G. and L.  Mariani.  1956.   Eye  changes  due to  naphthalene.  Med.
Lavoro.  47:  524.

Gil'denblat,  I.A.,  et  al.   1960.   Vapor  pressure over  crystalline naphtha-
lene.  Jour. Appl. Chem. USSR.  33: 245.

Harris, R.P.  et al.   1977a.  Factors  affecting the  retention  of a petroleum
hydrocarbon   by  marine  planktonic  copepods.   In:   Fate  and  Effetts   of
petroleum hydrocarbons in marine ecosystems and organisms.   Proc. Symp.  286.

Harris, R.P.  et  al.   1977b.   Accumulation of carbon-14-l-napthalene by  an
oceanic and an estuarine copepod during long-term exposure to  low-level con-
centrations.   Mar.  Biol.  42: 187.
                                   /3/-V

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Josephy,  E.  and F.  Radt,  (eds.)  1948.   Encyclopedia of organic  chemistry:
Series  III.  Elsevier Publishing Co.,  Inc.,  New  York.

Xnake,  E.   1956.  Uber  schwache geschwulsterzengende Wirkung  von  Naphthalin
und Benzol.  Virchows Archiv.  Pathol.  Anat.  Physiol.   329:  141.

Kraemer,  M.,' et al.   1974.   S^ typhimurium and §_._  coli to detect  chemical
mutagens.  Arch. Pnarmacol.  284: B46.

Xrstulovic,  A.M.,   et al.   1977.   Distribution  of  some  atmospheric  poly-
nuclear aromatic hydrocarbons.   Am. Lab.   9(7):  11.

Lawier,  G.C.,  et  al.   1978.   Accumulation  of  aromatic  hydrocarbons  in
tissues of petroleum-exposed  mallard  ducks  (Anas  olatyrhynchos).   Environ.
Sci. Technol.   12: 51.

Lee,. R.F..  et al.  1972a.   Uptake, metabolism and discharge of polycyclic aro-
matic hydrocarbons by marine fish.  Mar. Biol.   17:  201.

Lee, R.F.  et al.  1972b.   Petroleum hydrocarbons:   uptake  and discharge  by
the marine mussel Mytilus edulis.  Science.   177: 344.

Mackell,  J.V.,  et  al.   1951.   Acute hemolytic anemia  due  to ingestion  of
napthalene moth balls.   Pediatrics.  7: 722.
                                                                              A
Mahvi,  D.,  et  al.   1977.   Morphology of  a naphthalene-induced bronchiolar
lesion.  Am. Jour. Pathol.  86:  559.
                                                                      •.
Manufacturing  Chemists  Assoc.   1956.   Chemical  safety  data  sheets  SD-58:
Napthalene.  Washington, O.C.

McCann,  J.,   et al.   1975.   Detection  of  carcinogens  as  mutagen  in the
Salmonella/microsome test.   Assay of 300 chemicals.   Proc.  Natl.  Acad.  Sci.
72: 5135.

Mitchell,  S.  1926.   A  method  for  determining  the  solubility of sparingly
soluble substances.  Jour. Chem. Soc.  129:  1333.

Pirie,   A.  1968.   Pathology in  the eye of the naphthalene-fed rabbit.   Exp.
Eye Tes.  7: 354.

Robbins, M.C.   1951.  Determination of Napthalene in air.   Arch.  Ind.  Hyg.
Occuo.   Med.  4: 85.

Sanborn, H.R.  and  D.C.  Malins.   1977.   Toxicity and  metaoolism  of  naphtha-
lene: a study with marine  larval invertebrates.  Proc- Soc.  Exp.  3iol.  Med.
154: 151.

Solomon, T.   1957.   A  manual of pharmacology  and  its applications to  th^ra-
peutics and toxicology.  3th ed.  W.3. Saunders Co., Philadelphia.

Takizawa,  N.    1940.   Carcinogenic  action  of certain quinones.   Proc.  Imp.
Acad. (Tokyo)  16:  309.
                                    13 l-

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Tatem, H.E.   1976.   Toxicity and physiological affects  of oil and petroleum
hydrocarbons  on  astuarine  grass shrimp, Palaeminetes  pugio.   Holthuis Ph.D.
dissertation.  Texas A and M University.  133 pp.

U.S. EPA..  1971-1977.  Unpublished data from Region IV, Atlanta, Ga.

U.S.  EPA.   1976.  Organic  chemical producer's data  base program.  Chemical
No. 2701.  Radian Corporation.

U.S.  EPA.   1978.  In-depth  studies on  health  and environmental  impacts of
selected  water  pollutants.   Contract  NO.  68-01-4646,  U.S.   Environ.  Prot.
Agency.

U.S. EPA.  1979.  Naphthalene: Ambient Water Quality Criteria.  (Draft)

Van Heyningen,  R. and  A.  Pirie.   1966.   Naphthalene cataract.  In:  M.U.S.
Dardenne,  ed.   Symposium  on  the  biochemistry  of the  eye.    Karger,  Asel,
Switzerland.

Van Heyningen,  R. and  A.  Pirie.   1976.   Naphthalene cataract  in pigmented
and albino rabbits.  Exp. Eye Res.  22: 393.

Van der  Hoeve,  J.   1913.   Wirkung  von napbthol auf  die  augen von menschen,
tieren, und auf fatale augen.  Graele Arch. Ophthal.  85: 305.

Wallen,  I.E.,  et al.   1957.   Toxicity of  Gambusia affinis  of  certain pure
chemicals in turbid waters.  Sewage Ind. Wastes.  29: 695.

Windholz, M., ed.  1976.  The Muck Index, 9th ed.  Muck and Co. Rahway, N.J.

Zinkham,  W.J. and 8. Childs.  1958.   A defect  of  glutathione metabolism in
erythrocytes  from patients  with   a   naphthalene-induced hemolytic  anemia".
Pediatrics  22: 461.
                                          I 3  H3

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                                       No.  132
         1,4-Naphthoquinone

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENC7
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to  the subject chemi-
cal..  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources,  this short profile
may'not- reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

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                                                                                itA
I                                      1,4-NAPHTHOQUTNONE                      '

|             SUMMARY.

                   1,4-Naphthoquinone is used as a polymerization regulator and an
              intermediate.  Some data are available which indicate that 1,4-naphtho-
              quinona is biodegradable.
                   The most consistent findings reported in the literature for health
              effects of 1,4-naphthoquinone involve hematological changes, irritant and
              allergenic activity, and inhibition of biochemical oxidation processes.
              One study found 1,4-naphthoquinone to be oncogenic.  Some evidence of
              inhibition of in vitro endocrine function and of nerve activity was re-
              ported.

              I.  INTRODUCTION.

                   1,4-Naphthoquinone (1,4-aaphthalenedione; CiQH,0 • molecular weight 158.15)
              is a solid at room temperature.  It occurs as a greenish yellow powder or
              as yellow triclinlc needles.  It has a melting point of 123-126 C and begins
              to sublime at 100 C; its density is 1.422.  1,4-Naphthoquinone is only
              slightly soluble in water; it is soluble in a variety of organic solvents
              (Windholz 1976; Hawley 1971).
                   Current production (including importation) statistics for 1,4-naphtho-
              quinone (CAS Mo. 130-15-4) listed in the initial TSCA Inventory (U.S. EPA 1979)
              show that between 1,000,000 and 9,000,000 pounds of this chemical were
                                         *
              produced/imported in 1977.
                   1,4-Naphthoquinone is used as a polymerization regulator for rubber
              and polyester resins, in the synthesis o'f dyes and Pharmaceuticals, and as
              a fungicide and algicide (Hawley 1971).

              II.  EXPOSURE
                   A.  Environmental Fate
                   No specific information on the biological, chemical or photochemical
              transformation of 1,4-aaphthoquinone under environmental conditions was
              identified in the literature.  Napthoquinones undergo few substitution
              * This production range information does not include any production/importation
1             data claimed as confidential by the person(s) reporting for the TSCA Inventory,
              nor does it include any information which would compromise Confidential
              Business Information.  The data submitted for the TSCA Inventory, including
              production range information, are subject to the limitations contained in
              the Inventory Reporting Regulations (40 CFR 710).

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reactions (Thirtle 1965).  Like other quinones, 1,4-naphthoquinone can
interconvert with its corresponding hydroquinone iny an oxidation-reduction
                                                   X-
system.
     Talakin (1964) reported that 1,4-naphthoquinone in river water apparently
undergoes slow biochemical oxidation, based on an observed increase in
BOD.  Verschueren (1977) reports that the BOD- is 0.81, using the standard
dilution technique with normal sewage as seed material,'and that the theoretical
oxidation demand is 2.1.                               .'..

     B.  Bioconcentration

     No information was found on the bioconcentration potential of 1,4-naph-
thoquinone.  Based on its low water solubility and its solubility in organic
solvents, 1,4-naphthoquinone could be expected to bioconcentrate to some
extent.

     C.  Environmental Occurrence
     No information was found on the presence of 1,4-naphthoquinone in
environmental media.
     In addition to its potential entry into the environment from its
manufacture, processing and uses, 1,4-naphthoquinone may also enter the.
environment as a degradation product of certain naphthalene derivatives.
For example, the U.S. EPA (1975) reported studies showing that the pesticide
carbaryl (1-naphthyl-n-methyl-carbamate) undergoes hydrolysis to 1-naphthol,
which is then converted by bacteria to 1,4-naphthoquinone and other products.

III.  PHAKMACOKINETICS

     No information was obtained.

IV.  HEALTH EFFECTS

     A.  Carcinogenicity

     1,4-Naphthoquinone was found to induce neoplasm when applied dermally
to mice for 28 weeks.  The total dose applied was 2000 mg/kg.  (Proceedings
of the Imperial Academy of Tokyo 16:309, 1940, as cited in NIOSH 1975).

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     3.  Reproductive Effects

     I,4-Naphthoquinone completely  inhibited  the  gametokinetic  effect
of human chorionic gonadotropin in  toads  (Pakrashi  1963) .

     C.  Other Toxicity
     The oral LD5Q for rats was reported  as 190 mg/kg  (NIOSH, 1975) .  The
LD.QQ of 1,4-naphthoquinone in rats was 0.5 g/kg, 0.25 g/kg, and 0.5 g/kg
for intraventricular, subcutaneous, and intraperitoneal' administrations,
respectively.  The LC.nn in air was 0.45  mg/L for a one-hour exposure.
Acute (0.5 g/kg) and subchronic (0.3 g/kg for 4 days) exposure  of rats re-
sulted in the formation of 39 and 13% methemoglobin, respectively,  followed
by the appearance of Heinz bodies and development of hemolytic  anemia.
A decrease in total respiration and hypothermia due to disturbances in
oxidation-reduction processes was also observed.  According to  the authors,
"threshold concentrations of 1,4-oaphthoquinone detected for rats and rabbits
in single-exposure and chronic experiments were 0.0004 and 0.0007 mg/L with
respect to their irritant and toxic, effects"  (Slyusar  et al. 1964).   !

     D.  Other Relevant Information
     1,4-Naphthoquinone exerted an allergenic effect in guinea  pigs
(Kryzhanovskaya et al. 1966).  A possible role for  1,4-naphthaquinone
in drug-induced thrombocytopenia was suggested by Niewig et al.  (1973)
as 1,4-naphthoquinone was found to be involved in the destruction of normal
blood platelets by serum antibodies in vitro.  1,4-Naphthoquinone blocks
the biosynthesis of adrenal steroids by bovine adrenal cortex in vitro
(Kahnt and Neher 1966), and has an inhibitory effect on mixtures of cytochrome
_c and dehydrated succinate oxidase from beef heart  (Heymann and Falser 1966) .
1,4-Naphthoquinone inhibited ATPase and nerve activity in the (American)
             •j
cockroach.   (Baker and Norris 1971, Baker 1972).

V.  AQUATIC  TOXICITY
     Very little information was available.  For  1,4-naphthoquiaone, a
median threshold limit value (TLM:24-28 hr) of 0.3-0.6 mg/L was  listed for
an unspecified species of fish (Verschueren 1977).

VI.  GUIDELINES
     No guidelines for exposure to  1,4-naphthoquinone were located.

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                               References


Baker JE. 1972.  Effects of feeding-inhibitory quiripnes on the nervous
system of Periplaneta.  Experientia.  28(l):31-32.

Baker JE, Norris DM.  1971.  Neurophysiological and biochemical effects
of naphthoquinones on the central nervous system, of Periplaneta.
J. Insect Physiol. 17:2383-2394.

Hawley GG.  1971.  Condensed Chemical Dictionary, 8th edition.  Van Nostrand
Reinhold Co.

Heymaan H, Feiser LF. 1948.  Naphthoquinone antimalarials.  XXI.  Anti-
succinate oxidase activity.  Jour. Biol. Chem. 176(3):1359-1369.

Kahnt FW, Neher R. 1966.  Biosynthesis of adrenal steroids in vitro.  II.
Importance of steroids as inhibitors.  Helv. Chim. Acta 49(1):123-133. (Ger.)

Kotsifopoulos PN.  1975.  In vitro effect of oxidizing and analgesic agents
on the erythrocyte membrane protein electrophoretic pattern.  Nouv. Rev.
Fr. Hematol. 15(1) -.141-146.  (Abstract in Chemical Abstracts, 83,72709Z).

Kryzhanovskaya MV, et al. 1966.  Allergenic activity of some atmospheric
pollutants of a chemical nature.  Gig. Sanit. 31(3):8-11.

National Institute of Occupational Safety and Health.  Registry of Toxic
Effects of Chemical Substances. 1975.                     .     -

Nieweg HO, et al. 1973.  Drugs and thrombocytes.  Proc. Eur. Soc. Study
Drug Toxic.  14:101-109.

Pakrashi A.  1963.  Endocrinological studies of plant products.  IV.  Effect
of certain coumarins upon the biological potency of human chorionic gonado-
tropin.  Ann. Biochem. Exptl. Med. (Calcutta) 23:357-370.

Slyusar MP, et al. 1964.  Data on the toxicology of alpha-naphthoquinones:
and its permissible concentration in a working area.  Gigiena 95-100.
(Abstract in Zh. Farmakol. Toksikol.  11.54.373, 1965).

Talakin YN.  1965.  The experimental determination of the maximum permissible
concentration of alpha-naphthoquinone in water resources.  Hyg. and Sanit.
30:184.

Thirtle JR.  1965.  Quinones.  In: Kirk-Othmer Encyclopedia of Chemical
Technology.  2nd Edition.  John Wiley and Sons, -Inc., New York.

U.S. EPA 1975.  Microbial degradation and accumulation of pesticides in
aquatic systems.  EPA 660/3-75-007, PB 241 293.
                                                                     »
U.S. EPA 1979.  Toxic Substances Control Act Chemical Substance Inventory,
Production. Statistics for Chemicals on the Non-Confidential TSCA Inventory.

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Verschueren, K.  1977.  Handbook of Environmental Data on Organic Chemicals, -
Van Nostrand Reinhold Co.
        , M. .1. 1976.  The Merc, Index, Here. . Co., lac.,  Rahway, «e» Jersey.
                                   m-7

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                                  No.  133
             Nickel

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
      WASHINGTON, B.C.  20460

          APRIL 30,  1980
           /33-/

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                       SPECIAL NOTATION










U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated



nickel and has found sufficient evidence to indicate that



this compound is carcinogenic.

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                                    NICKEL
                                    Summary

     Nickel  is  a  ubiquitous  multi-media  environmental  contaminant.   Al-
though nickel  is toxic and  appears to be a carcinogen to  man,  there  is an
increasingly  strong  indication that  nickel  is  an  essential element.   The
route of  exposure to nickel is  very  important, since  oral intake of nickel
metal is  comparatively nontoxic.   However,  exposure to nickel by  inhalation
or parenteral  administration as well as  cutaneous contact can produce toxic
         •                                           •
affects.   In  terms of human health effects,  probably the most acutely toxic
nickel compound is  nickel  carbonyl.   Nickel  in  several  chemical forms  has
been  associated  with  lung cancer in_ man  and  experimental  animals  upon
inhalation;  carcinogenic effects,  however, are  not  indicated by  the  oral
route.  The  acceptable daily intake (ADI)  of  nickel is 254 ug per day  for a
70 kg man.
     The  toxicity  of nickel to  aquatic life  is  affected by water hardness.
In the aquatic environment nickel  is  acutely toxic to freshwater fishes at a
concentration of 2,480 jug/1 (26 mg/1 hardness).   Chronic toxicity to fishes
has  been   reported  at  527  jug/1  (210 mg/1 hardness).   Nickel  toxicity  is
affected  by  water hardness.   Algae appears  to be  more  sensitive to nickel
than  fish.   Based on  the  limited  number  of studies  performed,  the  biocon-
centration  factor for  fish is 61,  for   algae  the  factor  is 9.8,  and  the
weighted average bioconcentration  factor is 11 for  fish and shellfish.

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                                    NICKEL
I.   INTRODUCTION
         This  profile  is based on  the Ambient Water  Quality Criteria Docu-
ment for Nickel (U.S. EPA, 1979).
     Nickel  (Ni;  atomic weight 58.71),  a bright,  silver  metal of the iron-
cobalt-nickel  triad, is  a  hard  and  malleable  metal with  a  high  tensile
strength used  in  virtually  all areas of metallurgy.  Nickel does not  readily
form chloro-complexes  under environmental  conditions and  would not  be  ex-
pected to form significant amounts of  sulfate complexes (U.S. EPA,  1979).
     In 1972,  U.S.  consumption of nickel,  exclusive  of  scrap,  was estimated
to total about 160,000 tons (Reno,  1974).   The estimate consisted mainly of
commercially  pure nickel  (about  110,000  tons)  which  is used  in stainless
steel, electroplating, and various other alloys.                            *
II.  EXPOSURE
     The route by which  most  people  in the general  population receive  the
largest portion  of daily nickel  intake is  through foods.   Total  daily  di-
etary intake  values may  range up  to 900 ;jg nickel,  depending  on the nature
of the diet,  with average  values  of 300 to  500  ug daily  (NAS,  1975).    The
U.S. EPA (1979) has  estimated  a weighted average  bioconcentration  factor  for
nickel to  be 11  for  the edible portions of  fish and shellfish  consumed by
Americans.   This  estimate is based on measured steady-state bioconcentration
studies in  fathead minnow larvae  (Pimephales  promelas)  (Lind,  et  al. Manu-
script).   The  values for  nickel levels in 969 U.S., public water supplies  for
1969-1970 was 4.8 pg/1, with only 11  systems  of this" total exceeding 25 pg/1
(NAS, 1975).   The  levels of  nickel  in the  air  are  also  low, with  a  1974
                                                                       »
arithmetic  mean level for urban air of 9 ng/m3 (U.S. EPA,  1976).

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III. PHARMACOKINETICS
     A.  Absorption
         The major  routes of nickel absorption  are inhalation and  ingestion
via  the diet.   Percutaneous  absorption is  a  less significant  factor  for
nickel's systemic effects but  important in the allergenic responses to  nick-
el.  Collectively the data  of  Tedeschi and Sunderman (1957), Perry  and  Perry
(1959), Nomoto  and  Sunderman (1970), Nodiya  (1972),  and Horak and  Sunderman
(1973)  indicate that 1 to  10  percent  of  dietary nickel  is absorbed.   Skin
penetration of  nickel has  been- demonstrated with, nickel  entering at sweat-
duct °and hair-follicle ostia (Wells, 1956).  The  extent to which nickel en-
ters the bloodstream  by way of the skin cannot be  stated  at the present time
(U.S. EPA, 1979).
         Respiratory  absorption of  various forms of nickel is probably  the
                                                                            «
major route of  nickel entry into man  under conditions  of occupational  expo-
sure.   Pulmonary absorption into the  bloodstream is  probably greatest  for
nickel  carbonyl vapor, with animal  studies suggesting  that  as much as  half
of  the  inhaled .amount  is  absorbed  (Sunderman  and  Selin,  1963).   Nickel  in
particulate matter  is absorbed from the pulmonary tract  to a lesser degree
than nickel carbonyl  (Leslie,  et  al. 1976).  Based on  animal studies, nickel
appears to have a half-life cf  several  days in  the body, yet there  is little
evidence for tissue accumulation.
     B.  Distribution
         Blood  is  the main vehicle  for  transport of  absorbed nickel,   with
serum albumin being the main carrier protein, although  a  specific nickelrich
metalloprotein  has  been  identified in  man  (NAS,  1975).  Tissue distribution
                                                                       »
of absorbed nickel appears  to  be dependent on the  route of intake.   Inhaled
nickel  carbonyl leads to highest  levels in  the  lung,  brain, kidney,  liver,

-------
and adrenals (Armit,  1908;  Sunderman  and Selin, 1968; Mikheyev, 1971).   Par--
enteral administration  of  nickel salts usually  results  in highest levels  in
the kidney,  with  significant uptake  shown by  endocrine glands,  liver and
lung (Wase, -et al. 1954; Smith and Hackley, 1968).
     C.  Metabolism
         A  number of disease  states and  other physiological  stresses are
reported to  alter the movement and tissue  distribution of  nickel in man  as
well as experimental  animals.  In man,  increased  levels of serum nickel are
seen in cases  of acute myocardial infarction  (D'Alonzo  and Pell,  1963;  Sun-
derman, et al.  1972),  acute stroke  and extensive burn injury (McNeely,  et
al. 1971).   Reduction is  seen in hepatic cirrhosis or uremia, possibly  sec-
ondary to hypoalbuminemia.
         Nickel  appears to  be an essential  element,  at  least  in experimental
animals.   Nickel deficient  diets  have produced  decreased  growth  rates and
impaired  reproduction in  swine (Anke,  et  al.  1974) and  rats  (Schnegg and
Kirchgessner, 1975).                               •
     D.  Excretion
         The routes  of elimination for  nickel in man  and animals depend  in
part on the chemical  forms  of nickel  and the mode of nickel intake.  Dietary
nickel, due  to the  low  extent of gastrointestinal absorption, is simply  lost
in the feces (U.S. EPA, 1979).   Urinary  excretion  in  man and animals is usu-
ally the major clearance route for absorbed nickel.   In some  instances sweat
can constitute a major  route of nickel elimination (Hohnadel, et  al. 1973).
Nodiya  (1972)  reported a fecal  excretion average of 258  jjg in Russian stu-
dents.  Horak  and Sunderman  (1973) determined fecal  excretion  of nickel  in
                                                                       *
10 healthy  subjects and arrived  at a value  identical  to that  found in the
Russian study.

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IV.  EFFECTS
     A.  Carcinogenicity
         A  carcinogenic  response to"'various nickel compounds upon  injection
has deen observed in a number  of animal studies (Lau,  et al. 1972;  Stoner,
et  al.  1976;  IARC,  1976).   In nickel  refinery  workers, an  excess risk  of
nasal and lung cancer  has been demonstrated (IARC, 1976).  However,  there  is
no  evidence at present  to indicate  that orally  ingested nickel is  tumori-
genic.
     The qualitative and ..quantitative  character  of the carcinogenic  effects
of nickel as  seen in experimental animal models has  been shown to vary  with
the chemical  form of the  nickel, the route of exposure,  the animal model em-
ployed, and the amounts of the  substance administered (U.S. EPA,  1979).
3.   Mutagenicity
         Pertinent information  could  not be located in the available  litera-
ture.
     C.  Teratogenicity
         While Ferm  (1972) has  claimed unspecified malformations in  surviv-
ing hamster embryos  when mothers were  exposed to  parenteral  nickel (0.7  to
10.0 mg/kg),  Sunderman, et al.  (1978) found no teratogenic effects  from  oral
administration of either  nickel chloride (16 mg/kg) or nickel subsulfide (80
mg/kg) in rats.   Exposure of pregnant  rats  by inhalation to nickel carbonyl
on days 7 or  8 of gestation  frequently  caused the  progeny to develop ocular
anomalies,   including anophthalmia and  microphthaimia.   The  incidence of ex-
traocular anomalies  is very low.  The specificity  of nickel carbonyl for in-
duction of ocular anomalies in  rats  appears to be unique among known terato-
genic agents (Sunderman,  et al. 1979).

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     D.  Other Reproductive Effects
         Schroeder and  Mitchner (1971) have  demonstrated adverse affects in
a three generation  study with rats  at a level of  5 mg/1  (5  ppm) nickel in
drinking water.   In  each of the generations,  increased numbers of runts and
enhanced neonatal mortality were  seen.  A  significant  reduction in  litter
size  and  a .reduced  proportion of  males in  the  third  generation were also
observed.  Nickel sulfate  (25  mg/kg) has been demonstrated to be  gametotoxic
in  rats,  with  complete obliteration  of spermatozoa  following  exposure for
120 days (Hoey, 1966; Waltschewa, et al. 1972).
     E.  Chronic  Toxicity
         Chronic  exposure  to nickel  has resulted  in injury to both the upper
and lower  respiratory tract in  man  (Tolot, et al.  1956;  McConnell,  et al.
1973).  Inhalation of nickel particulate matter  is  likely to play a role in
chronic, respiratory  infections by  effects on  alveolar macrophages.  Contact
dermatitis  in man with nickel  sulfate has  been observed  (Fregert,  et al.
1969;   Brun,  1975).    Also,  dietary nickel  can elicit  a dermatitic response
(Kaaber, et al. 1978).
     F.  Other Relevant Information
         There are experimental data that  demonstrate that nickel has a syn-
ergistic effect  on  the  carcinogenicities  of  polycyclic hydrocarbons  (Toda,
1962;   Maenza,  et al.  1971;  Kasprzak,  et  al.  1973).   Nickel  and  other ele-
ments  are known  to  be  present in  asbestos and may  possibly be  a factor in
asbestos carcinogenicity   (Cralley,  1971).   Also,., a synergistic  action be-
tween nickel and  viruses has been suggested (Treagon'and Furst, 1970).
V.   AQUATIC TOXICITY
                                                                       »
     A.  Acute Toxicity
         Water hardness  significantly  influences  the acute toxicity of nick-
el  to freshwater fish.   For  fish,   observed  LC5Q  values range  from  2,480


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/jg/1  for  the  rock  bass  (Ambloohites  ruoestris)  (hardness  =  26  mg/1)  to
110,385  jug/1 for  the  bluegill  (Lsoomis macrochirus)  (hardness  = 42  mg/1).
At  a hardness  of 20-29  mg/1,  six  freshwater  species .have  LC_Q values  of
between  2,916 and 5,360 ug/1 (Pickering  and  Henderson,   1966;  Lind et  al.,
manuscript).  At  a hardness of  360  ug/1,  values range from 39,600  to  44,500
/jg/1.  In  comparison,  acute tests with  freshwater invertebrate  species  have
a  greater  range  of  LC--  values at a  fixed hardness.   The  stonefly  (Aero-
neuria  lycorias)   exhibited the  highest  LC5Q  of  33,500 jug/1  (Warnick  and
Bell, 1969)  and Daphnia maqna gave  the  lowest value  of  510  ug/1 (Biesinger
and Christensen,  1972).   LLnd, et al.  (1979)  provide the only data obtained
under  relatively  high hardness  conditions  (244  mg/1),  an  LC5Q  value  of
2409 jug/1  for Daphnia  pulicaria.
         Data on  the acute toxicity of  nickel  to saltwater fishes is  limit-
ed.   The  LC.Q  values  range  from  29,000 jug/1  for  the  Atlantic Silverside
(Menidia menidia)  to 350,000 ug/1 for the mummichog (Fundulus heteroclitus)
(Eisler and  Hennekey,  1977).   The invertebrate acute toxicity data base  con-
sists of  14 results,  with a range  of LC-- values from- 310 ug/1 for  larvae
                                          >u
of  the  hard  clam (Mercenaria  mercenaria)  (Calabrese  and Nelson,  1974)  to
500,000 ug/1 for adults of  the cockle Cardium edule (Portmann, 1963).
     B.  Chronic Toxicity
         A  life  cycle test  (Pickering,   1974)  and  an  embryo-larval  test
(Lind,  at   al.,  manuscript)  have been  conducted  with  the  fathead   minnow
(Pimeohales  oromelas).  The chronic  values are 527 ug/L (210 mg/1 hardness)
and  109  jug/1 (44 mg/1 hardness) respectively,   aiesinger and  Christensen
(1972) conducted a life cycle  test with  Daohnia manna resulting  in a chronic
                                                                       »
value of 53  ug/1  at  a hardness  of 45  mg/1.  There are no chronic saltwater
data available (U.S.  SPA,  1979).

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     C.  Plant Effects
         Hutchinson  (1973)  and Hutchinson and Stokes (1975) observed  reduced
growth of  several algae  species  at concentrations  ranging from  100  to  700
jjg/1.  A decrease  in diatom diversity was observed by Patrick, et  al.  (1975)
to occur at concentrations as  low as 2 ug/1.
     D.  Residues
         Bioconcentration  data is limited  to the fathead minnow,  Pimephales
promelas,  (Lind,  et  al., manuscript)  and  the  alga,  Scenedesmes acuminata
(Hutchinson  and Stokes,  1975).  The  bioconcentration factor  for  the whole
body of the fathead minnow is  61 and for  the  alga the factor is 9.8.
VI.  EXISTING GUIDELINES AND STANDARDS
     Neither  the  human health  nor  the_ aquatic criteria  derived  by U.S.  EPA
(1979a), which  are summarized  below,  have gone through the process of  public
review;  therefore,  there  is   a  possibility  that  these  criteria will be
changed.'
     A.  Human
         The  American   Conference   of  Governmental  Industrial   Hygienists
(ACGIH, 1971) has  adopted a threshold limit  value  (TLV)  for a workday expo-
sure of 1  ppb.  The  acceptable daily intake (ADI) for man has been determin-
ed to be 294 jjg/day  (U.S.  EPA,  1979).  The U.S.  EPA (1979) draft water qual-
ity criterion for nickel is 133 ug/1.
     B.  Aquatic
         For  nickel,  the draft criterion  (U.S. EPA,  1979)  to protect  fresh-
water aquatic life is:
                (1.01  . In  (hardness) - 1.02)
              e
                                                                       »<
as a 24-hour average, and the concentration should not exceed at any time:
                (0.47  . In  (hardness) + 4.19)
              Q                             *
         The  draft criterion to protect  saltwater  aquatic life  is 220 ug/1
as a 24-hour average, not to exceed 510 ug/1 at any time (U.S. EPA, 1979).

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                           NICKEL

                         REFERENCES

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Threshold- limit values for chemical  substances  and  physical
agents  in the workroom environment with  intended changes for
1978.   94 pp.

Anke, M. , et al.  1974.  Low  nickel  rations  for growth and
reproduction in pigs.  In; Trace Element  Metabolism in Ani-
mals- 2.  W.G. Hoekstra, J.W.  Suttie,  H.S.  Ganther and W.
Mertz (eds.).  University Park Press,  Baltimore,  MD.,  pp.
715.'

Armit,  H.W_  1908.  The toxicology of  nickel carbonyl.   Part
II.  Jour. Hygiene  8: 565.
                                                        0
Biesinger, K.E., and G.M. Christensen.   1972.   Effects of
various metals on survival, growth,  reproduction,  and  metabo-
lism of Daphnia magna.  Jour. Fish.  Res.  Board  Can..  29:
1691.

Brun, R.  1975.  Epidemiology of contact  dermatitis in Geneva-
(1,000  cases).  Dermatol.  150: 193.  (French)

Calabrese, A., and D.A. Nelson.  1974.   Inhibition  of  embry-
onic development of the hard  shell clam,  Mercenaria mercen-
aria, bv heavy metals.  Bull. Environ. Contain.  Toxicol.  2:
Cralley, L.J.  1971.  Electromotive phenomenon  in  metal  and
mineral particulate exposures.  Relevance  to  exposure  to as-
bestos and occurrence of cancer.  Am. Ind.  Hyg.  Assoc. Jour.
32: 653.

D'Alonzo, C.A., and S. Pell.  1963.  A  study  of  trace  metals
in myocardial  infarction.  Arch. Environ.  Health  6:  381.

Sisler, R, , and R.J. Hennekey.  1977.   Acute  toxicities  of
Cd2"*", Cr2*, Ni2"*". amd Zn2* to estaurine macro fauna.
Arch. Environ. Contam. Toxicol.  6: 315.

Perm, v.H.  1972.  The teratogenic effects  of metals  on  mam-
malian embryos.  In: Advances in Teratology,  Vol.  5.   D.H.M.
Wollam (ed. )   Academic Press, New York.  pp.  51-75.

Fregert, S.,  et al.  1969.  Epidemiology of contact dermati-
tis.  Trans.  St. Johns Hosp. Derm. Soc. 55: 71.

Hoey, M.J.  1966.  The effects of metallic  salts on the  his-
tologv and functioning of the rat testes.   Jour. Reprod.
Fertil.  12:  461.

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Hohnadel, D.C., et al.   1973.  Atomic  absorption  spectrometry
of nickel, copper, zinc, and  lead  in sweat  collected  from
health subjects during sauna  bathing.   Clin.  Chem.   19:
1288.

Horak, E., and F.W. Sunderman.   1973.   Fecal  nickel  excretion
by healthy adults.  Clin. Chem.  29: 429.

Hutchinson, T.C.  1973.  Comparative studies  of the  toxicity
of heavy metals to phytoplankton and their  synergistic  inter-
actions.  Water .Pollut.  Res.  (Canada)  8:  68.

Hutchinson, T.C., and P.M. Stokes.  1975.   Heavy  metal  toxi-
city and algal bioassays.  ASTM  STP 573,  Am.  Soc. Test.
Mater.  pp. 320-343.

International Agency for Research  on Cancer.   1976.   Nickel
and nickel compounds.  In; Evaluation  of  Carcinogenic Risk of
Chemicals to Man  (International Agency for  Research  on  Cancer
Monographs, 11) IARC, Lyon, p. 111.

Kaaber, K., et al.  1978.  Low nickel  diet  in the treatment
of patients with chronic nickel-dermatitis.   Brit.. Jour.
Derm..  98: 197.

Kasprzak, K.S., et al.   1973.  Pathological reactions in  rat
lungs following intratracheal injection of  nickel subsulfide
and 3,4-benzpyrene.  Res. Comm. Chem.  Pathol.  Pharmacol.  6:
237.

Lau, T.J., et al.  1972.  The carcinogenicity of  intravenous
nickel carbonyl in rats.  Cancer Res.   32:  2253.

Leslie, A.C.D., et al.   1976.  Prediction of  health  effect of
pollution aerosols.  In; Trace Substances in  Environmental
Health - X.  D.D. Hemphill (ed.),  University  of Missouri,
Columbia, Mo.  pp. 497-504.

Lind, D., et al.  Regional copper-nickel  study, Aquatic Tox-
icology Study, Minnesota Environmental Quality Board, State
of Minnesota (Manuscript).

Maenza, R.M. et al.  1971.  Rapid  induction of sarcomas in
cats by combination of nickel sulfide  and 3,4-benzypyrene.
Cancer Res.  31: 2067.

M.cConnell, L.H., et al.  1973.  Asthma caused  by  nickel sen-
sitivity.  Ann. Ind. Med.  78: 888.

McNeely, M.D., et al.  1971.  Abnormal concentrations of
nickel in serum in cases of myocardial  infarction, stroke,
burns, hepatic cirrhosis, and uremia.  Clin.  Chem.   17:
1123.
                          J

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Mikheyev, M.I.   1971.   Distribution  and  excretion of nickel
carbonyl.  Gig.  Tr.  Prof.  Zabol.   15:  35.

National Academy of  Sciences.   1975.   Nickel.   National Acad-
emy of Sciences  Committee  of Medical  and Biological Effects
of Environmental Pollutants.  Washington,  DC.

Nodiya, P.I.  1972.  Cobalt and  nickel balance  in students of
an occupational  technical  school.  Gig.  Sanit.   37:  108.

Nomoto, S., and  F.W. Sunderman,  Jr.   1970.   Atomic absorption
spectrometry of  nickel  in  serum,  urine,  and  other biological
materials.  Clim. Chem.  16: 477.

Patrick, R., st  al.  1975.  The  role  of  trace elements  in
management of nuisance  growths.   U.S.  Environ.  Prot.  Agency,
EPA 660/2-75-008, 250 p.

Perry, H.M., Jr.,. and E.P. Perry.  1959.  Normal concentra-
tions of some trace  metals in human urine:   Changes  produced
by ethylenediametetracetate.  Jour. Clin.  Invest.   38:  1452.

Pickering, Q.H.  1974.  Chronic  toxicity of-  nickel to the
fathead minnow.  Jour.  Water Pollut.  Control Fed.   46:  760.

Pickering, Q.H., and C. Henderson.  1966.  The  acute  toxicity
of some heavy metals to different  species  of warmwater
fishes.  Air Water Pollut. Int.  Jour.  10: 453.

Portmann, J.E.   1968.   Progress  report on  a  program  of
insecticide analysis and toxicity  testing  in relation to the
marine environment.  Helgolander wiss. Meeresunters   17:
247.

Reno, H.T.  1974.  Nickel.  In:  Minerals Yearbook 1972,  Vol.
I.  Metals, Minerals and Fuels.  Washington, DC,  U.S.   Gov-
ernment Printing Office, pp. 871.

Schnegg, A.,  and M.  Kirchgessner.  1975.  The essentiality of
nickel for the growth of animals.  Z.  Tierphysiol.,  Tierer
naehr.  Futtermittelkd.  36: 63.

Schroeder, H.A., and M. Mitchner.  1971.  Toxic  effects  of
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Smith, J.C. ,  and 3.  Hackley.  1968.   Distribution  and excre-
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Nutr.  95: 541.
                                                          »
Stoner, G.D., at al.  1976.  Test  for  carcinogenicity of me-
tallic compounds by  the pulmonary  tumor  response  in  strain A
mice.  Cancer Res.   36: 1744.

-------
Sunderraan, F.W., et  al.   1978.   Embryotoxicity and  fetal
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                             s
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after exposure  to nickel  carbonyl.  Arch.  Ind.   Health   16:
486.

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cancer.  Bull. Tokoya  Med.. Dentr Univ.   9: 441.

Tolot, F. , et al..  1956..  Asthmatic forms  of  lung disease  in
workers exposed to chromium, nickel and  aniline  inhalation.
Arch. Mol. Prof. Med..  Tran.. Secur. Soc.  18:  288.

Treagon, L., and A.  Furst.  1970.  Inhibition  of  interferon
synthesis in mammalian cell cultures after nickel treatment.
Res. Comm.- Chem. Pathol.  Pharmacol.  1:  395.

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through 1974 from the  national  air surveillance  network.
EPA-600/4-76-041, U.S. Environ.  Prot.  Agency,  Research
Triangle Park, NC.

U.S. EPA,  1979.  Nickel:  Ambient Water Quality Criteria.

Waltschewa, V.W. et  al.   1972.   Hodenveranderungen  bei
weissen Ratten durch chronische  Verabreichung  von Nickel sul-
fate.  (Testicular changes due  to long-term administration of
nickel sulphate in rats.)  Exp..  Pathol.  6: 116.  In German
with Engl. abstr.

Warnick, S.L., and H.L. Bell.   1969.   The  acute  toxicity of
some heavy metals to different  species of aquatic insects.
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                                                          »
Wase, A.W., et al.   1954.  The metabolism of nickel.  I.
Spatial and temporal distribution of Ni^3  ^n the mouse.
Arch. Biochem. Biophys.   51: 1.

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Wells, G.C.  1956.  Effects of nickel on the skin.  Brit.
Jour. Dernatol.  68: 237.

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                                      No. 134
            Nitrobenzene

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy»

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                                 NITROB5NZENE
                                    Summary

     Nitrobenzene  is a.  pale yellow  oily liquid  with an  almond-like odor.
There is little or no  information  available  on its teratogenic, mutagenic or
carcinogenic  effects.    Nitrobenzene  yielded  negative results  in  the Ames
assay for  mutagenicity.   Gross abnormalities  were observed  in  4 fetuses of
30 rats administered nitrobenzene.
     Chronic  exposure  to nitrobenzene produces cyanosis, methemoglobinemia,
jaundice, anemia, and sulfhemoglobinemia in man.
     Static  tests  with  the  bluegill,  sunfish,  Daphnia maqna,  and  an alga,
Selenestrum  capricomutum,  indicates little  difference in  sensitivity with
no  50  percent effective  concentration  lower  than 27,000 ug/1.   An embryo-
                                                                            ^
larval test  with  the fathead minnow demonstrated  no  adverse chronic effects
at the highest  concentration tested (32,000 ug/1).   Static  tests with salt-
water  fish,   shrimp,  and  alga gave  repeated  96-hour LC_Q  or  EC5Q  values
of 58,538 jjg/1, 6,676 ug/1  and 9,600 jug/1, respectively.

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I.  INTRODUCTION
     This  profile is based  on the  Ambient Water  Quality Criteria  Document
for Nitrobenzene  (U.S.  EPA,  1979).   The principal  uses  of nitrobenzene  are
for reduction  to aniline (97  percent),  solvent for Friedel-Crafts  reaction,
metal polishes,  shoe black,  perfume,  dye  intermediates,  crystallizing  sol-
vent  for  some substances,  and  as a  combustible  propellant  (Dorigan  and
Hushon, 1976).
     Nitrobenzene   (CJHJOJ   is   a  pale   yellow  oily   liquid  with   an
almond-like  odor.   Its  physical  properties  include:   melting  point,  6°C;
vapor pressure,  0.340 mm Hg  at 25°C; and  solubility in  water  of 1000  mg/1
at 20°C  (U.S.  EPA,  1979).   Nitrobenzene is miscible  with most organic  sol-
vents, a  fairly  strong  oxidizing agent,  and undergoes  photoreduction  when
irradiated with  ultraviolet light in organic  solvents that contain  abstraci-
table hydrogen atoms.
II.  EXPOSURE
                                                     i
     A.  Water
         Levels of nitrobenzene in  wastewater are  monitored  by plants  pro-
ducing and using  the  chemical, but nitrobenzene levels in city water  systems
are usually too low to measure (Pierce, 1979).
     3.  Food
         Nitrobenzene is  not an approved food  additive  (Oorigan and  Hushon,
1976).  There have been  reports of nitrobenzene poisoning resulting  from  its
contamination of alcoholic drinks and food (Nabarfo, 1948).
     The U.S. EPA  (1979) has  estimated  the  weighted average bioconcentration
factor for nitrobenzene to be  4.3  for  the edible portions of fish and.shell-
fish consumed  by  Americans.   This  estimate  was based on  octanol/water par-
tition coefficients.

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     C.  Inhalation
         Atmospheric  nitrobenzene levels  outside a  plant  are not  monitored
by industry.   Since  inner plant  levels  are below  the Threshold Limit  Value
(TLV)  of  5 mg/m   and  nitrobenzene vapors  accumulate at the floor  level  due
to their  high density, the  external  concentrations  are  expected to be  very
low (Dorigan and Hushon, 1976).
III. PHARMACOKINETICS
     A.  -Absorption
       •  Nitrobenzene  absorption can  occur by  all  possible  routes,  but  it
takes  place mainly through the respiratory tract and skin.  On  the average,
80 percent  of the nitrobenzene vapors are retained in the human respiratory
tract  (Piotrowski, 1977).
         Nitrobenzene,  as  liquid  and  vapor,  will  pass  directly through  the
skin.  The  rate  of vapor absorption depends on  the air concentration,  rang-
ing  from  1 mg/hr  at  5  mg/m  concentration to  9 mg/hr  at 20 mg/m .   Maxi-
                                                                     2
mal cutaneous  absorption  of liquid nitrobenzene is 0.2 to  3 mg/cm /hr  de-
pending on skin temperature.
     3.  Distribution
         Upon'  entry   into   the body,  nitrobenzene  enters  the  bloodstream.
Nitrobenzene  is  a very  lipid soluble with an  oil  to water  coefficient  of
800.   In a  rat study,  the ratio  of  concentration of nitrobenzene in adipose
tissue versus blood in internal organs and muscle was approximately 10:1  one
hour after  an intravenous injection (Piotrowski, -1977).   Oorigan and HusHon
(1976)   found  that 50  percent of the nitrobenzene  administered to rabbits
accumulated unchanged  in tissues within two days  after intubation.

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     C.  Metabolism
         There are two  main  pathways for the metabolism of  nitrobenzene:   1)
reduction  to  aniline   followed  by  hydroxylation  to  aminophenols,  and  2)
direct  hydroxylation  of nitrobenzene  to form  nitrophenols.  Further  reduc-
tion of nitrophenols to- aminophenols may also occur (Piotrowski,  1977).   The
first  pathway proceeds  via  the  unstable intermediates,  nitrosobenzene  and
phenylhydroxylamine, both  of which  are  toxic  and have  pronounced  methemo-
globinemic  capacity.   These  reactions occur  in the  cytoplasmic and  micro-
somal  fractions  of liver cells  by the nitro-reductase  enzyme system  (Fouts
and Brodie,  1957).  The  aniline  is then excreted as an acetyl derivative,  or
hydroxylated  and excreted as  an aminophenol.   The second  pathway  does  not
occur  in the microsomal fraction.  This .reaction  proceeds via peroxidase  in
the presence of oxygen (Piotrowski, 1977).                                    ;
         Robinson, at al. (1951)  found p-aminophenol  to be the main  metabol-
ic product  of nitrobenzene metabolism in rabbits.  Little  unchanged  nitro-
benzene was  excreted  in the urine and only 1  percent was expired as  carbon
dioxide.  Together  with nitrophenols  and nitrocatechol,  p-aminophenol con-
stituted 55  percent of  the  urinary metabolites.   Metabolites were  detected
in the urine up to one week after dosing.
     D.  Excretion
         In  man,  the primary  known  excretion  products of  nitrobenzene  are
p-aminophenol  and  p-nitrophenol  which  appear  in the  urine  after chronic  or
acute  exposure.   In   experimental   inhalation  exposure -to  nitrobenzene,
p-nitrophenol  was formed  with  the  efficiency of  6  to  21  percent.    The
efficiency  of  p-aminophenol  formation  is  estimated  from  acute poisorjing

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cases  where  the molar  ratio of  excreted p-nitrophenol  to  p-aminophenol  is'
two to one, since p-aminophenol is  not formed at a detectable level  in  short
subacute exposure (Piotrowski, 1977).
         Ikeda  and  Kita' (1964)  found the  rate of  excretion of  these  two
metabolites to parallel the level of methemoglobin in the blood.
         Nitrobenzene  remains in the  human  body  for a  prolonged  period  of
time.   The excretion  coefficient  of  urinary  p-nitrophenol  (followed  for
three  weeks)  in man is  about 0.008 per  hour.   The  extended systemic reten-
tion and slow excretion of metabolites in man is determined by the low  rates
of metabolic  transformation "(reduction and  hydroxylation)  of the nitroben-
zene itself.  The conjugation and excretion  of  the  metabolites,  p.-nitrophe-
nol and p-aminophenol,  is  rapid (Piotrowski,  1977).   The urinary metabolites
in man account  for  only  20 to 30 percent of  the nitrobenzene dose; the  fate
of the rest of the metabolites is not  known (Piotrowski, 1977).
IV.  EFFECTS
     A.  Carcinogenicity
         The  available  literature  does not  demonstrate  the Carcinogenicity
of nitrobenzene, although it is suspect (Dorigan and Hushon,  1976).
         Some nitrobenzene  derivatives have  demonstrated carcinogenic  capa-
cities.  Pentachloronitrobenzene  (PCN8)'induced hepatomas and papillomas  in
mice (Courtney,  et al. 1976).
         l-Fluoro-2,4-dinitrobenzene  (ONFB)  was  found  to be a  promoter  of
skin tumors  in  mice,  although  it  does  not  induce  them when  administered
alone (Bock,  et al 1969).
     B.  Teratogenicity
                                                                      »
         There  is  a paucity  of  information  on  the  teratogenic effects  of
nitrobenzene.    In  one  study,  125  mg/kg  was  administered  to  pregnant   rats
                                      J*

-------
during preimplantation  and placentation periods  (Kazanina,  1963).  Delay  of
embryogensis,  alteration  of  normal placentation,  and  abnormalities  in  the
fetuses  were  observed.   Gross  morphogenic defects  were  seen  in  4  of  30
fetuses examined.
     C,  Mutagenicity .
         Nitrobenzene was  not found to  be mutagenic  in the Ames Salmonella
assay  (Chiu,  et al.,  1978).  Trinitrobenzene  and 'other nitrobenzene  deriva-
tives  have  demonstrated mutagenicity  in  the  Ames Salmonella microsome  assay
and  the  mitotic  recombination  assay  in  yeast (Simmon,  et  al.  1977),  thus
raising questions concerning the mutagenicity  of  nitrobenzene.
     0.  Other Reproductive Effects
         Changes  in the tissues  of  the  chorion  and placenta  of pregnant
women  who  worked in  the  production of  a  rubber catalyst  that  used  nitro-
benzene were  observed.   No mention was made of the effects  on  fetal develop-
ment or  viability (Dorigan and Hushon, 1976).  Menstrual disturbances  after
chronic nitrobenzene exposure have been reported.
         Garg,  at-al. (1976)  tested substituted nitrobenzene derivatives  for
their  ability  to  inhibit  pregnancy in  albino rats.   Two  of  the compounds
tested (p-methoxy and p-ethoxy derivatives)  inhibited implantation and  preg-
nancy 100 percent when administered on days 1  through  7 after impregnation.
     E.  Chronic Toxicity
         Symptoms of chronic occupational nitrobenzene  absorption are  cyan-
osis,  methemoglobinemia, jaundice,  anemia,  sulfhemoglobinemia,  presence  of
Heinz  bodies  in  the erythrocytes,  dark  colored  urine,  and the  presence  of
nitrobenzene  metabolites  (e.g.,   nitrophenol) in  the  urine  (Pacseri and
                                                                      9
Magos,  1958;  Hamilton,   1919;  Wuertz,  et  al.  1964; Browning,  1950;  Maiden,
1907; Piotrowski, 1967).


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         Chronic exposure of  laboratory  animals to nitrobenzene (via  inhala-


tion,  ingestion  or  subcutaneous   injection)  produced  symptoms  similar  to


those  mentioned above  for humans as  well as  tissue  degeneration  of  the


heart,  liver,  and  kidney,  and  reductions in  erythrocytes  and   hemoglobin


levels in the blood (U.S. EPA, 1979).


     F.  Other Relevant Information


         Alcohol ingestion  has been found to act synergistically with  nitro-


benzene in man and animals  (Dorigan and Hushon, 1976; Smyth, et al., 1969).


         Kaplan, et al.  (1974) showed that caffeine, an inducer of microsom-


al enzymes,  increases the  rate  of metabolism and' excretion of nitrobenzene


thus causing a rapid decline in nitrobenzene induced methemoglobin  levels-


         Metabolism and  excretion  of nitrobenzene in humans is slower by  an


order of magnitude than in  rats or rabbits  (Piotrowski, 1977).              .


V.   AQUATIC TOXICITY


     A. • Acute Toxicity


         The  96-hour  LC5Q  reported  value  for the  bluegill (Lepomis  macro-


chirus)  is  42,600 ug/1  and the observed  48-hour LC^n for  Daohnia magna  is
—•«-^•  ^           '                                  J\J      —•••••••••••_ i^HHMBMBM-


27,000 ug/1.  Saltwater  species  tested  are the sheepshead minnow,   Cyprinodon


varieqatus, which  has a reported  96-hour LC5Q of  58,539 jug/1 and the mysid


shrimp, Mysidopsis  bahia,  with a  reported 96-hour  LC5_.  of  6',676  jug/1  (U.S.


EPA, 1979).


     8.  Chronic Toxicity


         In  the  only  chronic  data  available, .,no  adverse  effects  were


observed during  an embryo-larval  test  with the  fathead minnow  (Pimephales


promelas) at  nitrobenzene  test concentrations  as  high as 32,000  pg/1  (U.S.
                                                                       »

EPA, 1978).


-------
     C.  Plant Effect
         Based  on  cell numbers  and  chlorophyll  a  concentration,  reported
EC5Q  values for  the  freshwater alga,  Selenastrum capricornutum.  are  42,000
and  44,100 ug/1; and  for  the marine  alga, Skeletonema  costatum,  there  are
reported £C.Q values of-9,600 and  10,300 ug/1  (U.S. EPA,  1979).
     0.  Residues
        _A  bioconcentration  factor of 15 was estimated for  aquatic organisms
that contain 8 percent lipids.
VI.  EXISTING GUIDELINES'AND STANDARDS
     Neither the human health nor the aquatic  criteria  derived by  U.S.  EPA
(1979), which are  summarized below, have gone  through the process  of  public
review;  therefore,  there  is  a   possibility  that these  criteria  will  be
changed.                                                                      ^
     A.  Human.
         The  TLV  for  nitrobenzene  is 5  mg/m .   This  is  the  OSHA Federal
standard, the value set by  the  ILO/WHO committee on Occupational  Health,  and
the TLV suggested  by  the American  Conference  of Governmental and  Industrial
Hygienists  (Goldstein, 1975, ACGIH, 1977).
         The draft  water  quality  criteria  for nitrobenzene  is 30 ug/1  (U.S.
EPA, 1979).  This  value is  based  on the TLV  and organoleptic level  (minimum
detectable odor limit in water) of  nitrobenzene.
     S.  Aquatic
         For nitrobenzene the drafted criterion to.-protect  freshwater  aquat-
ic life is  480  ug/1 as a 24-hour  average  concentration  not to  exceed  1,100
ug/1 at any time.   To  protect  saltwater aquatic life, the 24-hour average is
                                                                         •
53 ug/1 and this concentration should not  exceed 120 ug/1 at any time  (U.S.
EPA, 1979).

-------
                                 NITROBENZENE

                                  REFERENCES
American  Conference  of  Governmental  Industrial Hygiensts.   1977.   Docu-
mentation  of  the  threshold  limit  value  for  substances  in  workroom  air.
Cincinnati, Ohio.

Bock, A.G.,  et al.  1969.  Tumor  promotion  by  1-fluoro-2, 4-dinitrobenzene,
a potent skin sensitizer.  Cancer-Res.  29: 179.

Browning,  E.    1950.    Occupational  jaundice   and   anemia.   Practitioner
164: 397.

Chiu, C.W., et al.   1978.  Mutagenicity of some commercially available nitro
compounds for Salmonella typhinurium.  Mut. Res.  58: 11.

Courtney, K.D.,  et al.   1976.   The effects of pentachloronitrobenzene, hexa-
chlorobenzene, and  related compounds on  fetal  development.   Toxicol.  Appl.
Pharmacol.  35: 239-

Dorigan,  J.,   and  J.  Hushon.    1976.   Air  pollution  assessment  of  nitro-
benzene.  U.S. Environ. Prot. Agency.

Fouts,  J.R.,  and  B.B.  Brodie,   1957.   The  enzymatic  reduction of  cloram-
phenicol,  p-nitrobenzoic  acid  and  other  aromatic   nitro  compounds   in
mamma]*.  Jour. Pharaiacol. Exp. Ther.  119: 197.

Garg, S.X., et al.   1976.  Potent  female  antifertility  agents.  Indian Jour.
Med. Res.  64: 244.

Goldstein, I.   1975.   Studies   on  MAC  values of  nitro  and amino-derivatives
of  aromatic  hydrocarbons.   Adverse  Effects  Environ.  Chem.  Psychotropic
Drugs  1: 153.

Hamilton,  A.    1919.   Industrial  poisoning  by  compounds  of the  aromatic
series.  Jour. Industr. Hyg.   1: 200.

Ikeda, M., and A. Kita.   1964.   Excretion  of p-nitrophenol and p-aminophenol
in  the  urine  of a  patient  exposed  to nitrobenzene.   Br.  Jour.  Ind.  Med.
21: 210.

Kaplan, A.M.,  et al.   1974.   Methemoglobinemia  and metabolism of  nitro  com-
pounds.  Toxicol. Appl. Pharmacol.  29:  113-

Kazanina, S.S.   1968.   Morphology and  histochemistry -of hemochorial  placen-
tas  of white  rats during  poisoning of  the maternal   organisms  by  nitro-
benzene.  Bull. Exp. Biol. Med.  (U.S.S.R.)  65: 93-
                                                                        *
Maiden, W.   1907.   Some observations on  the condition  of  the blood  in  men
engaged  in aniline  dyeing and  the manufacture  of nitrobenzene and  its  com-
pounds.  Jour. Hyg. 7:  672.
                                     _/  ^£a
                                     1J U >

-------
Nabarro,  J.D.N.   1948.   A  case  of  acute mononitrobenzene  poisoning.   3r.
Med. Jour.  1: 929.

Pacseri, I.,  and  L.  Magos.  1953.  Determination  of the measure of exposure
to  aromatic  nitro and  amino  compounds.   Jour.  Hyg.  Epidemiol.  Microbiol.
Lnmunol.  2: 92.

Pierce,   M.     1979.    Personal   communication.    Quality  Control   Dep.,
Philadelphia Water Treatment Div., Philadelphia, ?a.

Piotrowski, J.   1967.   Further investigations on  the  evaluation of exposure
to nitrobenzene.  3r. Jour. Ind. Med.  24: 60.

Piotrowksi, J.   1977.   Exposure  tests for  organic compounds  in industrial
toxicology.  NIOSH 77-144.  U.S. Dep. Health, Edu. Welfare.

Robinson, D.,  et  al.   1951.   Studies in detoxication.   40.   The metabolism
of  nitrobenzene  in  the rabbit,   o-,  o-, and p-aitrophenols,  o-,  m-,  and
p-aminophenols   and   4-nitrocatechol  as   metabolites  of   nitrobenzene.
Biochem.  Jour.  50:  228.

Simmon, V.F.r  et  al.   1977.  Munitions wastewater treatments:   Does chlori-
nation  or ozonation  of  individual  components  produce  microbial  mutagens?
Toxicol. Appl. Pharmacol.   41: 197.

Smyth,  H.F.,   Jr.,  et  al.   1969.   An exploration of  joint  toxic  action:
Twenty-seven  industrial chemicals intubated  in  rats in  all  possible  pairs.
Toxicol. Appl. Phannacol.   14: 340.

U.S.  EPA.   1978.   In-depth  studies  on health  and  environmental  impacts  of
selected water pollutants.  Contract No. 68-01-4646.

U.S. SPA.  1979.  Mitrobenzenes.  Ambient Water Quality Criteria   (Draft).

Wuertz,   R.L.,  et al.  1964.   Chemical  cyanosis  - anemia  syndrome.   Diag-
nosis, treatment,  and recovery.  Arch. Environ. Health  9: 478.

-------
                                      No.  135
           •4-Nitrophenol

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.   20460

           APRIL 30,  1980
    / -3 ff-

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference documents..
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                          4-NITROPHSNOL



                             SUMMARY



     There is-no evidence to indicate that 4-nitrophenol  is carcin-



ogenic.



     Weak mutagenic  effects in Saccharomyces  and  in Proteus have



been  observed.  Results  from  the  Ames  assay, the  E.   coli,  and



the dominant  lethal assay  failed  to show  mutagenic effects from



4-nitrophenol.



     No  information on  the  teratogenic  or  adverse reproductive



effects of 4-nitrophenol is available.



     A single animal study  indicates cumulative chronic .toxicity;



the methodology of this study was not available for  review.



     For freshwater  organisms,  acute values for the toxic effects



of  4-nitrophenol  ranged  from  8,280  to  60,500  pg/1,   and  7,170



to  27,100  ug/1  for marine  organisms.    Effective  concentrations



for aquatic plants fall within these ranges of concentrations.

-------
                          4-NITROPHENOL
I.    INTRODUCTION
     This profile  is  based on the  Ambient Water Quality Criteria
Document for Nitrophenols (U.S. EPA, 1979).
     The  raononitrophenols  are  a  family  of  compounds  composed
of  the  isomers  resulting  from  nitro group  substitution  at the
2,3, and 4 position of  phenol (the  ortho, meta,  and para isomers,
respectively).   The -para  isomer,  4-nitrophenol, has  a molecular
weight of  139.11,  a  boiling point of  279°C,  a' melting  point of
113-114°C,  a  density  of 1.479 g/ml; it  is  soluble  in  water  (U.S.
EPA, 1979).
     Uses of  the mononitrophenols  include  the  following: produc-
tion of dyes, pigments,  Pharmaceuticals, rubber  chemicals,  lumber
preservatives, photographic chemicals, and pesticidal and fungici-
dal  agents  (U.S. EPA,  1979).  'Production  was 17.5   x  10  tons
per year in 1975 (Chem.  Market.  Reporter, 1976).
     The nitrophenols  may  be  formed  via microbial   degradation
or  photodegradation   of pesticides  (e.g., parathion)  containing
the  nitrophenol  moiety.   4-Nitrophenol  may  be produced  in the
atmosphere  through  the photochemical   reaction between  benzene
and nitrogen monoxide  (U.S. EPA,  1979).   Partial  microbial degrada-
tion  of  certain nitrophenols  has  been  shown, particularly • by
acclimated microorganisms.    Mononitrophenols  appear to  be  effi-
ciently degraded by unacclimated microorganisms  (Haller, 1978).
II.  EXPOSURE
                                                              »
     The lack of  monitoring data  on the  mononitrophenols  makes
it difficult to assess  exposure from water, inhalation, and foods.

-------
Mononitrophenols  in  water  have  been  detected  in  the  effluents
of  chemical plants  (U.S.  EPA,  1976,  1979).    4-Nitrophenol  has
been shown  to  penetrate the skin and  to  produce damage  at  thres-
hold concentrations of  0.8  and  0.9 percent  (w/v) ,  respectively
(U.S.  EPA, 1979) .
     Exposure  to  nitrophenols appears, to  be  primarily through
occupational contact  (chemical  plants, pesticide  applications) .
Contaminated water may result  in isolated poisoning incidents.
     The U.S.  EPA  (1979)  has  estimated the  weighted  average bio-
concentration  factor  for 4-nitrophenol to  be 4.9  for  the  edible
portions of fish and  shellfish consumed by  Americans.  This  esti-
mate is based on the octanol/water partition coefficient.
III. PHARMACOKINETICS
     A.   Absorption and Distribution
          Pertinent  data could  not  be located  in  the  available
literature regarding absorption or distribution.
     B.   Metabolism
          Metabolism  of  the  mononitrophenols  occurs   primarily
by conjugation.  Other  possible  routes  are  reduction  of  the  nitro
group  to  an  amino  group or  oxidation  to  dihydric-nitrophenols
(U.S.   EPA, 1979).    These  reactions  are  mediated  primarily  by
liver enzyme systems, although other tissues show lower metaboliz-
ing activity (U.S. EPA, 1979).
     3.   Excretion
          An animal study has indicated  that- oral  or intraperi-
                                                              •
toneal administration of  4-nitrophenol  leads to rapid  elimination
in  all  species  tested,  and  that   the  total elimination  period
is not likely to exceed one week  (Lawford, et al.  1954) .


-------
IV.  EFFECTS
     A.   Carcinogenicity
          There is no evidence available regarding the carcinogeni-
city of raononitrophenols.
     B.   Mutagenicity_
          A  weak  rautagenic  effect was  detected  in  Saccharomyces
cerevisiae by  4-nitrophenol  (Fahrig,  1974);  this  was also  indi-
cated .by  testing  4-nitrophenol  for  growth  inhibition  in  a  DNA
repair  deficient  strain of  Proteus  mirabilis  (Adler,  et  al. ,
1976).  This compound has also induced chromosome breaks  in  plants
(U.S.   EPA,  1979).   4-Nitrophenol has  failed to  show  mutagenic
effects in the  Ames  assay,  in E.  coli,  or  in the dominant  lethal
assay (U.S. EPA, 1979).
     C.   Teratogenicity and Other Reproductive Effects
          Pertinent  data could  not  be  located  in  the  available
literature regarding teratogenicity and other  reproductive effects.
     D-.   Chronic Toxicity
          A  single Russian  study (Makhinya,  1969)  reported  that
chronic  administration  of   mononitrophenol  to  mammals  produced
hepatitis, splenic hyperplasia, and neurological symptoms. Method-
ology of this study was not available  for review.
V.   AQUATIC TOXICITY
     A.   Acute Toxicity
          L<~50 va^-ues have been obtained for  two  species  of  fresh-
water  fish:  8,280  ug/1  for  bluegills,  Lepomis  macrochirus,  in
                                                              »
a  96-hour static  assay  (U.S.  SPA,  1973),  and  60,510  ug/1  for
the fathead minnow, Pimephales oromelas, in a 96-hour  flow-through
assay  (Phipps,  et  al.  unpublished manuscript).    For the  fresh-

-------
water  invertebrate,  Daphnia magna,  determined LCen  values  range



from  8,396  to 21,900  ug/1 (U.S.  EPA,  1979).   The  marine  fish,



sheepshead  minnow,  Cyprinodon  variegatus,   has  produced  deter-



mined  LCjQ  va'lue of 27,100  816  ug/1 in  a 96-hour  static  assay,



while  the  marine mysid shrimp,  Mysidopis bahia,  was  more  sensi-



tive^ with a reported LCen value of 7,170 pg/1.



     B.   Chronic Toxicity



          No chronic studies on freshwater  organisms are available.



In an  embryo-larval test  of  the marine  fish,  sheepshead  minnow,



a chronic  value  of  6,325  ug/1 was obtained.    No  chronic  testing



for marine invertebrates was available.



     C.   Plant Effects



          Four  species  of  freshwater   plants have  been  tested



with  4-nitrophenol.    The  algae,   Selenastrum capricornutum  and



Chlorella  vulgar is,  and   the  duckweed,  Lemna minor ,  were  most



sensitive  with  effective  concentrations  of  4,190,  6,950,  and



9,452  pg/1,  respectively; while the  alga,  Chlorella pyrenoidosa,



was much more resistant, with an effective concentration of 25,000



ug/1.   The  marine alga,  Skeletonema  costatum,  provided effective



concentrations of 7,370 to 7,570 pg/1 (U.S. EPA, 1979).



     D.   Residues



          No  bioconcentration  factors  for either   freshwater • or



marine species were available.



VI.  EXISTING GUIDELINES AND STANDARDS



     Neither the  human health  nor  aquatic  criteria  derived  by
                                                             9


U.S.   EPA (1979), which  are  summarized below, have  gone  through



the  process  of  review;   therefore,  there is  a possibility  that



these criteria will be changed.
                             /3S-7

-------
     A.   Human



          Available data  pertaining to  4-nitrophenol is  insuffi-



cient for deriving a criterion to protect human health.



     B.   Aquatic



          A  criterion  for  protecting  freshwater   organisms  has



been drafted  as  240  ug/1, for  a 24-hour  average   concentration,



not  to  exceed 550 ug/1.   For marine  life, a  criterion has been



drafted as  53  ug/1 for a  24-hour average,  not  to exceed 120 ug/1
                                        •


(U.S. EPA, 1979).

-------
                        4-NITROPHENOL
                          REFERENCES

Adler, Bi", et al.  1976.  Repair-defective mutants of Pro-
teus mirab'ilis as a prescreening system for the detection
of potential carcinogens.  Biol. Zbl. 95: 463.

Chemical Marketing Reporter.  1976.  Chemical profile:
p-nitrophenol.  Chem. Market. Reporter p. 9.

Fahrig, R.  1974.  Comparative mutagenicity studies with
Pesticides.  Pages 161-181 In: R. Montesano and L. Tomatis
eds.  Chemical carcinogenesls" essays.  Proc. workshop on
approaches to assess the significance of experimental chemi-
cal carcinogenesis data for man organized by IARC and the
Catholic University of Louvain, Brussels, Belgium.  IARC
Sci. Publ.  No. 10.  Int. Agency Res. Cancer, World Health
Organization.

Haller, H.D.  1978.  Degradation of mono-substituted ben-
zoates and phenols by wastewater.  Jour. Water Pollut. Con-
trol Fed; 50: 2771.

Lawford, D.J., et al.  1954..  On the metabolism of some
aromatic nitro-compounds by different species of animals.
Jour. Pharm. Pharmacol. 6: 619.

Makhinya, A.P.  1969.  Comparative hygienic and sanitary-
toxicological studies of nitrophenol isomers in relation
to their normalization in reservoir waters.  Prom Zagryazneniya
Vodoemov. 9: 84.

Phipps, G.L., et al.  The acute toxicity of phenol and sub-
stituted phenols to the fathead minnow.  (Manuscript).

U.S. EPA.  1976.  Frequency of organic compounds identified
in water.  U.S. Environ. Prot. Agency.  Contract No. EPA
600/4-76-062.

U.S. EPA.  1978.  In-depth studies on health and environ-
mental impacts of selected water pollutants.  Contract No.
68-01-4646..

U.S. EPA.  1979.  Nitrophenols:  Ambient Water Quality Cri-
teria  (Draft) .

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                                  No.  136
           Nitrophenols

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
      WASHINGTON, D.C.  20460

          APRIL 30,  1980
       J36'l

-------
                          DISCLAIMER
     This report represents  a survey of- the potential health
and environmental hazards from exposure to  the  subject chemi-
cal..  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources,  this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by the
subject chemical.   This document  has undergone  scrutiny  to
ensure its technical accuracy.
                          .. ( f f\L
                         " f ]} V I

-------
                         NITROPHENOLS
                           SUMMARY
     None of the nitrophenols have shown carcinogenic  activity.
     Mutagenicity testing has indicated positive  effects
of:  2,4-dinitrophenol in mouse bone marrow cells and  E.
coli; 2,4,6-trinitrophenol in E. coli and Salmonella;  and
4,6-dinitro-ortho-cresol in Proteus.  Weak mutagenic effects
of 4-nitrophenol have been reported  in Saccharomyces and
in Proteus.  Other mutagenic test assays have shown negative
results for these compounds.
     Teratogenic effects have been reported in the develop-
ing chick embryo following administration of 2,4-dinitro-
phenol.  This compound did" not produce teratogenic effects
in mammalian studies.  Adverse reproductive effects (embryo
toxicity)  were seen- in rats exposed  to 2,4-dinitrophenol.
     {The chronic effects of 2,4-dinitrophenol ingestion
have included cases of agranulocytosis, neuritis, functional
heart damage, and cataract formation.  Ingestion  of 4,6-
dinitro-ortho-cresol has also produced cataracts  in humans.
     One Russian study has reported cumulative toxic effects
in animals produced by the mononitrophenols; methodology
of this study was not available for review.
     Freshwater fish appeared to be the most sensitive or- •
ganism to the action of nitrophenols, wi'th acute  values
ranging from 230 to 167,000 ug/1.  The reactivities of vari-
ous nitrophenols in order of decreasing toxicity  are,  in  .
general:  2,4-dinitro-6-methylphenol, 2,4-dinitrophenol,
2-nitrophenol, 4-nitrophenol, and 2,4,6-trinitrophenol.

-------
                         NITROPHENOLS



I.   INTRODUCTION



     This profile is based on the Ambient Water Quality



Criteria Document for Nitrophenols  (U.S. EPA, 1979).



     The nitrophenols are a family of compounds which, de-



pending on the extent and position of nitro group substituents,



include the mononitrophenols, dinitrophenols, and trinitro-



phenols.  Dinitrocresols are related compounds bearing an



additional 2-position methyl group.  The mononitrophenols



(molecular weight 139.11) show boiling points from 194-279°C



(depending on the isoraeric form) and melting points of 44—



114°C.  They have a density of 1.485 and are soluble in



water.  The dinitrophenols (molecular weight 184.11) have



melting points from 63.5-144°C and show a density of 1.67



to 1.70.  Water solubility is from 0.42 to 2.3 g/1.  Tri-



nitrophenols (molecular weight 229.11) have melting points



from 96-123°C; they are slightly soluble in water.  2,4,6-



Trinitrophenol, the most widely used isomer, has a density



of 1.763 g/ml and a solubility of 1.28 g/1.  Of the six



isomers of dinitrocresol, 4,6-dinitro-o-cresol is the only



one of any commercial importance.  The physical properties



of 4,6-dinitro-o-cresol, hereafter referred to as dinitro-



ortho-cresol, include a molecular weight of 198.13, a melt-



ing point of 85.8°C and a solubility of 100,mg/l in water



(U.S.  EPA, 1979).



     Uses of the mononitrophenols include the following:



production of dyes,  pigments, Pharmaceuticals, rubber chemi-

-------
cals, lumber preservatives, photographic chemicals, and
pesticidal and fungicidal agents.  The dinitrophenols are •
used as chemical intermediates for sulfur dyes, azo dyes,
photochemicals, pest control agents, wood preservatives,
and explosives.  2,4,6-Trinitrophenol  (picric, acid) is used
for dye intermediates, germicides, tanning agents, fungi-
cides, tissue fixative, photochemicals, Pharmaceuticals,
and for the etching of metal surfaces.  Dinitro-ortho-cresol
is used primarily as a blossom-thinning agent on fruit trees
and as a fungicide, insecticide, and miticide on fruit trees
during the dormant season (U.S. EPA, 1979).

Current Production:   2-nitrophehol      5-. 7.5x10  tons/year  .(1976)
                      4-nitrophenol        17.5x10  tons/year  (1976)
                                                  2
                      2,4-dinitrophenol     4.3x10  tons/year  (1968)

     The nitrophenols may be formed via microbial degrada-
tion or photodegradation of pesticides (e.g., parathion)
containing the nitrophenol moiety  (U.S. EPA,   1979).  Partial
microbial degradation of certain nitrophenols has been shown,
particularly by acclimated microorganisms.  Mononitrophenols
appear to be efficiently degraded by unacclimated microorgan-
isms (Haller, 1978).
II.  EXPOSURE
     The lack of monitoring data on the nitrophenols makes
it difficult to assess exposure from water, inhalation,
and foods.  Nitrophenols in water have been detected in

                              •a-
                           -V 6 W "

-------
effluents from chemical plants  (U.S. EPA,  1976;  1979) or


following dumping of explosives  (Harris, et al.  1946).


Dermal absorption of mononitrophenols, dinitrophenols,  tri-


nitrophenols  (picric acid), and dinitro-ortho-cresol  (DNOC)


has been detected .(U.S. EPA, 1979).


     Exposure to nitrophenols appears to be primarily through


occupational contact (chemical plants, pesticide applica-


tion) .  Contaminated water may result in isolated poisoning


incidents.
                                                    •

     The U.S. EPA (1979) has estimated weighted  average


bioconcentration factors for the following nitrophenols:


2-nitrophenol, 4.0; 4-nitrophenoJL, 4.9; 2,4-dinitrophenol,


2.4; 2,4,6-trinitrophenol, 6.0; and 4 ,6-dinitrocresol,  7.5


for fish and shellfish consumed by Americans.  This estimate


is based on octanol/water partition coefficients.


III. PHARMACOKINETICS


     A.   Absorption


          Specific data on the absorption  of the mononitro-


phenols is not available.  The dinitrophenols are readily


absorbed following oral, inhalation, or dermal administra-


tion.  Data on the absorption of trinitrophenols is not


available.  Animal studies with oral administration of  2,4,6-


trinitrophenol indicate that it is readil-y absorbed from


the gastrointestinal tract.  Dinitro-ortho-cresol is readily


absorbed through the skin, the respiratory tract, and the
                                                          »

gastrointestinal tract in humans (NIOSH, 1978).

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     B.   Distribution
          No information on the distribution of the raono-
nitrophenols is available.  Dinitrophenol blood levels rise
rapidly af-ter absorption, with little subsequent distribu-
tion or storage at tissue sites (U.S. 2PA, 1979) .  2,4,6-
Trinitrophenol and dinitro-ortho-cresol have been found
to stain several body tissues; however, the compounds may
be bound to serum proteins, thus producing non-specific
organ distribution (U.S. SPA, 1979).
     C.   Metabolism
          Metabolism of the nitrophenols occurs through
conjugation, reduction of nitro groups to amino groups,
or oxidation to dihydric-nitrophenols (U.S. EPA, 1979) .
These reactions are mediated primarily by liver enzyme systems,
although other tissues show lower metabolizing activity
(U.S. EPA, 1979) .   The metabolism of dinitro-ortho-cresol
is very slow in man as compared to that observed in animal
studies (King and Harvey, 1953) .
     Q.   Excretion
          Evidence from human poisoning with parathion indi-
cates that excretion of 4-nitrophenol in the urine is quite
rapid (Arteberry,  at al. 1961).   Experiments with urinary
clearance of dinitrophenols in several animal species indi-
cate rapid elimination of these compounds (Harvey,  1959) .
2,4,6-Trinitrophenol has been detected in the urine of ex-
posed human subjects indicating at least partial urinary
elimination (Harris,  et al. 1946).  The experiments of Parker
                           136-7

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and coworkers  (1951) in several animal species  indicate



that dinitro-ortho-cresol is rapidly excreted following



injection; however, Harvey, et al.  (1951) have  shown  slow



excretion 'of dinitro-ortho-cresol in human volunteers given



the compound orally.



IV..  EFFECTS



     A.   Carcinogenicity



          There are no available data to indicate that the



mononitrophenols are carcinogenic.  Both 2- and 4-nitrophenol



failed to show promoting activity for mouse skin tumors



(Boutwell and Bosch, 1959); this same study failed to show



promoting activity for 2,4-dinitrophenol.  No evidence is



available to indicate that dinitrophenols, trinitrophenols,



or dinitro-ortho-cresol produce any carcinogenic effects



(U.S. EPA, 1979).



     B.   Mutagenicity



          A weak mutagenic effect was detected  in Saccharo-



myces cerevisiae for 4-nitrophenol  (Fahrig, 1974); this



was also indicated by testing 4-nitrophenol for growth in-



hibition in a DNA repair deficient strain of Proteus mirabilis



(Adler, et al. 1976).  This compound has also induced chromo-



some breaks in plants (U.S. EPA, 1979).  4-Nitrophenol has



failed to show mutagenic effects in the Ames assay, in E.



coli, or in the dominant lethal assay (U.S.,EPA, 1979).



          Testing of 2,4-dinitrophenol has indicated muta-



genic effects in E. coli  (Demerec, et al. 1951)  and damage'



in murine bone marrow cells (chromatid breaks)  (Mitra and



iManna, 1971) .  Tn vitro assays of unscheduled DNA synthesis

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(Friedman and Staub, 1976) and DNA damage  induced  during
call culture (Swenberg, et ai. 1976)  failed  to  show  positive
results with this compound.
          '2,4,6-Trinitrophenol has produced  mutations  in
E. coli and Salmonella assays  (Demerec, et al.  1951; Yoshikawa,
et al. 1976) .  Testing in Drosop'nila  has failed to indicate
mutagenic activity.
          Adler, et al. (1976) have reported that  dinitro-
ortho-cresol shows some evidence of producing DNA  damage
in Proteus mirabilis.  Testing of this compound in the Ames
Salmonella system (Anderson, et al. 1972) or in E. coli
(Nagy, et al. 1975)  failed to show any mutagenic effects.
     C.   Teratogenicity
          No information is available to indicate  that mono-
nitrophenols, 2,4,6-trinitrophenol, or dinitro-ortho-cresol
produce teratogenic effects.
          2,4-Dinitrophenol has produced developmental abnor-
malities in the chick embryo  (Bowman, 1967;  Miyamoto,  et
al. 1975).  No teratogenic effects were observed following
intragastric administration to rats (Wulff,  et  al. 1935)
or intraperitoneal administration to mice  (Gibson, 1973).
     D.   Other Reproductive Effects
          Feeding of 2,4-dinitrophenol to. pregnant rats
produced an increased mortality in offspring  (Wulff, et
al. 1935); similarly, intraperitoneal administration of
the compound to mice induced embryotoxicity  (Gibson, 1973) '

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Influence of the compound on maternal health may have contri-
buted to these effects  (U.S. EPA, 1979).
     E.   Chronic Toxicity
         • Chronic administration of mononitrophenols to
mammals has been reported to produce hepatitis, splenic
hyperplasia, and neurological symptoms  in a single Russian
study (Makhinya, 1969).  Methodology of this study was not
available for review.
          Use of 2,4-dinitrophenol as a human dieting aid
has produced some cases of agranulocytosis, neuritis, func-
tional heart damage, and a large number of cases of cataracts
(Homer, 1942).  Cataracts have also been reported in patients
poisoned with dinitro-ortho-cresol (NIOSH, 1978).
          Human effects resulting from  2,4,6-trinitrophenol
exposure have been reported as temporary impairment of speech,
memory, walking, and reflexes (Dennie, et al. 1929).
     F.   Other Relevant Information
          A synergistic action in producing teratogenic
effects in the developing chick embryo has been reported
with a combination of 2,4-dinitrophenol and insulin (Landauer
and Clark, 1964).
          The combination of 2,4,6-trinitrophenol and opioids
or minor analgesics produced an increase in analgesia (Huidobro,
1971).
          2,4-Dinitrophenol is a classical uncoupler of
oxidative phosphorylation, which accounts for its marked
acute toxicity.  Dinitro-ortho-cresol is also well known
for its activity as an uncoupler.

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V.   AQUATIC TOXICITY
     A.   Acute Toxicity
          Freshwater fish LC   values reported  for  the  blue-
gill (Lepomis macrochinus) ranged from 230 to 167,000 ug/1
and for the juvenile fathead minnow  (Pimephales promelas),
from 2,040 to 60,510 ug/1.  The order of decreasing toxicity
for five nitrophenols examined was:  2,4-dinitro-6-methyl
phenol, 2,4-dinitrophenol, 2-nitrophenol, 4-nitrophenol,
    •
2,4,5-trinitrophenol (U.S. EPA, 1979).  For three of the
phenols tested with both the bluegill and fathead minnow,
                                              »
the bluegill appeared more sensitive.  In static bioassays
with the freshwater invertebrate, Daphnia magna, 48-hour
LC5Q values of 4,090 to 4,710; 8,396 to 21,900; and 84,700
ug/1 were reported for 2,4-dinitrophenol, 4-nitrophenol
and 2,4,6-trinitrophenol, respectively (U.S. ,SPA, 1979).
The marine fish, sheepshead minnow (Cyprinodon variegatus),
was the only fish species acutely tested for three nitro-
phenols, with reported LC5Q values of 29,400; 27,100 and
134,000 ug/1 being obtained for 2,4-dinitrophenol, 4-nitro-
phenol, and 2,4,6-trinitrophenol.  Observed LC.-Q values
of 4,350; 7,170 and 19,700 ug/1 were reported for the mysid
shrimp  (Mysidopsis bahia) for the same three formulations,
respectively.
     3.   Chronic Toxicity
          Pertinent information on the chronic effects  on
freshwater species could not be located in the available
literature searches.  The only chronic test on a marine

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species was an embryo-larval assay of the sheepshead minnow
that produced a chronic value of 6,325 ug/1  (U.S. EPA, 1978).
Pertinent information relative to chronic effects on marine
invertebrates could not be located in the available literature.
     C.   Plant Effects
          The effects of various nitrophenols vary widely
among species of freshwater plants and according to the
formulation of nitrophenol tested.  The duckweed, Lemna
minor, was the most sensitive plant tested with 2,4-dinitro-
phenol and was the most resistant with 2-nitrophenol, hav-
ing effective concentrations (50 percent growth reduction,
time unspecified)ranging from 1,472 to 62,550 ug/1 for the
two respective formulations.  The marine alga, Skeletonema
costatum, appeared to be slightly more resistant than fresh-
water species tested, with effective concentrations ranging
from 7,370 to 141,000 ug/1 for 4-nitrophenol and 2,4,6-tri-
nitrophenol, respectively.
     D.   Residues
          Bioconcentration factors were not determined for
any freshwater or marine species.  However, based on octanol/
water partition coefficients, bioconcentration factors were
estimated as 8.1, 21, and 26 for 2,4-dinitrophenol, 2,4,5-
trinitrophenol, and 2,4-dinitro-6-dimethylphenol, respectively.
VI.  EXISTING GUIDELINES AND STANDARDS
     The human health and aquatic criteria derived by U.S.
EPA (1979), which are summarized below, have not yet gone
through the process of public review; therefore, there is
a possibility that these criteria may be changed.
                          / 3 6-

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     A.   Human
          Eight-hour TWA exposures for 2,4,6-trinitrophenol
(0.1 mg/m )  and 4,6-dinitro-ortho-cresol  (0.2 rag/m ) have
been established by the ACGIH (1971).
          Draft, water quality criteria for the following
nitrophenols have been estimated, by U.S. SPA  (1979) based
on adverse effects data:  dinitrophenols - 63.6 pg/1; tri-
nitrophenols - 10 /ig/1; and dinitrocresols - 12.3 }ig/l.
     B.   Aquatic
          Criteria drafted to protect freshwater life from
nitrophenols follow:  57 ^g/1 as a 24-hour average concen-
tration, not to exceed 130 ug/1, for 2,4-dinitro-6-methyl-
phenol; 79 jug/1, not to exceed 180 ug/1, for 2,4-dinitro-
phenol; 240 ug/1, not to exceed 550 ug/1, for 4-nitrophenol;
2,700 pg/1, not to exceed 6,200 pg/1, for 2-nitrophenol;
and 1,508 ug/1, not to exceed 3,400 ug/1, for 2,4,6-trinitro-
phenol.  For marine life the following criteria have been
drafted as 24-hour average concentrations:  37 ug/1, not
to exceed 84 ug/1, for 2,4-dinitrophenol; 53 pg/1, not to
exceed 120 pg/1, for 4-nitrophenol; and 150 ug/1, not to
exceed 340  g/1, for 2,4,6-trinitrophenol.
                             if (11
                            J U I B^
                         I36-/3

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                        NITROPHENOLS

                         REFERENCES

Adler, B., et al.  1976.  Repair-defective mutants  of  Proteus
mirabilis as a prescreening system for  the detection of  po-
tential carcinogens.  Biol. Zbl.  95: 463.

American Conference of Governmental Industrial Hygienists.
1971.  Documentation of the threshold limit  values  for sub-
stances in workroom air.  Vol. 1. 3rd ed.  Cincinnati, Ohio.

Anderson, K.J., et al.  1972.  Evaluation of  herbicides  for
possible mutagenic properties.  Jour. Agric.  Food Chem.   20:
649.

Arterberry, J.D., et al.  1961.  Exposure to  parathion:   Mea-
surement by blood cholinesterase level  and urinary  p-nitro-
phenol excretion.  Arch. Environ. Health  3:  476.

Boutwell, R.K., and O.K. Bosch.  1959.  The  tumor-promoting
action of phenol and related compounds  for mouse skin.
Cancer Res.  19: 413.

Bowman, P.  1967.  The effect of 2,4-dinitrophenol  on  the
development of early chick embryos.  Jour. Embryol. Exp..
Morphol.  17: 425.

Demerec, M., et al.  1951.  A survey of chemicals for  muta-
genic action on E_. coli.  Am. Natur.  85: 119.

Dennie, C.C., et al.  1929.  Toxic reactions  produced  by  the
application of trinitrophenol (picric acid).  Arch. Dermatol.
Sypnilol.  20: 698.

Fahrig, R.  1974.  Comparative mutagenicity  studies with Pes-
ticides.  Pages 161-181 In; R. Montesano and  L. Tomatis.
(eds.)  Chemical carcinogenesis essays.  Proc. workshop on
approaches to assess the significance of experimental  chemi-
cal carcinogenesis data for man.  Organized by IARC and the
Catholic University of Louvain, Brussels, Belgium.  IARC Sci.
Publ. No.  10.  Int. Agency Res. Cancer, World Health  Organi-
zation.

Friedman, M.A., and J. Staub.  1976.  Inhibition of mouse
testicular DMA synthesis by mutagens and carcinogens as a po-
tential simple mammalian assay for mutagenesis.  Mutat. Res.
37: 67.

Gibson, J.E.  1973.  Teratology studies in mice with 2-sac-
butyl-4, 6-dinitrophenol (dinoseb).  Food Cosmet. Toxicol.'
11: 31.

-------
Haller, H.D,   1978.   Degradation of mono-substituted benzo-
ates and ohenols  by wastewater.   Jour.  Water Pollut. Control
Fed.   50:  2771.

Harris, A.H.,  et  al.   1946.   Hematuria  due to picric acid
poisoning  at a naval  anchorage  in Japan.   Am. Jour.  Pub.
Health  3.6: 727.

Harvey, D.G.   1959.   On  the  metabolism  of some aromatic nitro
compounds  by different species of animal.   Part III.  The
toxicity of the dinitrophenols,  with a  note on the effects of
high environmental temperatures.   Jour.  Pharm. Pharmacol.
11: 462.                                       »•

Harvey, D.G.,  et  al.   1951.   Poisoning  by dinitro-ortho-cre-
sol.  Some observations  on the effects  of  dinitro-ortho-cre-
sol administration by mouth  to human vplunteers.   3r.  Med.
Jour.  2:  13.

Horner, W.D.   1942.   Dinitrophenol and  its relation  to forma-
tion to cataracts.  Arch. Ophthal.   27:  1097.

Huidobro,  F.   1971.   Action  of picric acid on the  effects of
some drugs acting on  the  central  nervous  system, with  special
reference  to opiods.   Arch.  Int.  Pharmacodyn Ther.   192:.
362.

Xing, E.,  and  D.G. Harvey.   1353.   Some  observations on the
absorption and excretion  of  4,6-dinitro-o-creosol  .(DNOC). I.
Blood dinitro-o-cresol levels in  the cat  and rabbit  following
different methods of  absorption.   Biochem.  Jour.   53:  185.

Landauer, W. ,  and E.  Clark.   1964.   Uncouplers of  oxidative
ohosphorylation and teratogenic  activity  of  insulin.   Nature
204: 285.

Makhinya, A.P.  1969.  Comparative hygienic  and  sanitary
toxicological  studies  of  nitrophenol isomers in  relation to
their normalization in reservoir  waters.   Prom.  Zagryazneniya
Vodoemov.  9:  84. (Translation).

Mitra, A.B., and G.K.  Manna.  1971.   Effect  of some  phenolic
compounds on chromosomes  of  bone  marrow cells of mice.   In-
dian Jour. Med. Res.   59: 1442.

Miyamoto, K.,   et al.   1975.  Deficient myelination by  2,4-
dinitrophenol  administration  in  early stage  of development.
Teratology  12: 204.

Nagy, A., et al.  1975.   The correct mutagenic affect  of pes-
ticides on Escherichia coli WP2  strain.  Acta. Microbiol. '
Acad.  Sci. Hung.22: 309.

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National Institute for Occupational Safety and Health.   1978.
Criteria for a recommended standard: Occupational exposure  to
dinitro-ortho-creosol.  Dep.. Health Edu. Welfare, Washing-
ton, D.C.

Parker, V.H., et al.  1951.  Some observations on the  toxic
properties of 3,5-dinitro-ortho-cresol.  Br. Jour. Ind. .Med.
9: 226.

Swenberg,' J.A., et al.  1976.  In vitro DNA damage/akaline
elution assay for predicting carcinogenic potential.
Biochem. Biophys. Res. Cbiranun.  72: 732.

U.S. EPA.  1976.  Frequency of organic compounds identified
in water.  U.S. Environ. Prot. Agency.  Contract No. EPA
600/4-76-062.

U.S. EPA.  1978.  In-depth studies on health and environmen-
tal impacts of selected water pollutants.  Contract No.
6801-4646.

U.S. EPA.  1979.  Nitrophenols: Ambient Water Quality  Cri-
teria. (Draft).

Wulff, L.M.B., et al.  1935.  Some effects of alpha-dinitro-
phenol on pregnancy in the white rat.  Proc. Soc. Exp.  Biol.
Med.  32: 678.

Yoshikawa, K., et al.  1976.  Studies on the mutagenicity of
hair-dye.  Kokurltsu Eisei Shikenjo  94: 28.

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                                       No.  137
            NItrosamines

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

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                       SPECIAL NOTATION










U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated



nitrosamines and has found sufficient evidence to indicate



that this compound is carcinogenic.

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                                 NITROSAMINE5
                                    Summary

     Nitrosamines and nitrosamides  are  widespread in the environment and can
also be  produced end.ogenously by  nitrosation  of constituents  of food.  Ni-
trosamines  and  nitrosamides are  considered to  be  among the  most potent of
all carcinogenic,  mutagenic, and  teratogenic  agents  known.   The  livers of
rats chronically exposed to nitrosamines exhibit pathological changes.
     Toxicity data  examining the effects  of nitrosamines on  aquatic organ-
isms is  scant.   For  freshwater  life forms, acute toxicity levels of 5,850 to
7,760 ug/1  were reported,  while for marine fish an  acute  value  of nearly
3,300,000 ug/1  was  reported (both values for N-nitrosodiphenylamine).  N-ni-
trosodimethylamine  has  been shown  to   induce  hepatocellular carcinoma .in
rainbow trout.
                                    i / i &
                                *~) v> I

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                                 NITROSAMINES
I.   INTRODUCTION
     This  profile is  based on  the  Ambient  Water  Quality Criteria  Document
for Nitrosamines  (U.S. EPA,  1979).
     The  nitrosamines _(and  nitrosamides)  belong to a  large group of  chemi-
cals generally  called N-nitroso compounds.   Because they  frequently coexist
with N-nitrosamines  in the  environment and  are  structurally related  to  ni-
trosamines,  nitrosamides  .are also  included in the  U.S.  EPA (1979)  document
and in .this profile.
     The nitrosamines  vary  widely  in their physical properties and may exist
as solids, liquids or gases.  Nitrosamines of low molecular  weight are vola-
tile at room  temperature,   while  those of high molecular weight are  steam
                  *                                                         ' •
volatile.  Nitrosamines are soluble in water and organic  solvents (U.S. SPft,
1976) .
     Synthetic  production of  nitrosamines  is  limited" to small  quantities.
The only nitrosamine  produced  in quantities greater than  450 kg per year  is
N-nitrosodiphenylamine, which  is used in  rubber  processing and in the  manu-
facture of pesticides.   Other N-nitroso compounds  are  produced primarily  as
research chemicals (U.S. EPA,. 1976).
     Nitrosamines are  rapidly  decomposed by sunlight and thus do not persist
in ambient  air or water  illuminated by sunlight (U.S.  EPA, 1979;  Fine,   at
al. 1977a).  Some nitrosamines  have been found to persist  for extended peri-
ads of  time  in the aquatic  environment (Fine,  at. al.  1977a; Tate and  Alex-
ander,  1975).
II.  EXPOSURE
                                                                      »
     Nitrosamines  are  widespread   in  the environment.   The most  probable
source of environmental  nitrosamines is nitrosation of amine and amide ' pre-
cursors which are ubiquitous in the environment (Bogovsxi,  et.al. 1972).
                                      Y

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     It  has been estimated  that air, diet,  and smoking all  play  a roughly
equivalent  role  in  human exposure to preformed  nitrosamines,  contributing a
few  micrograms per  day; intake  from drinking  water  is probably  much less
than 1 ug per  day (U.S.  EPA, 1976).
     A.  Water
         Significant concentrations  of  nitrosamines have been reported for a
limited  number of  samples of ocean  water,  river water,  and  waste  treatment
plant  effluent (3  to  4 ug  dimethylnitrosamine/1)  adjacent to  or  receiving
wastewater  from industries using  nitrosamines or secondary  amines in produc-
tion operations  (Fine,   et al.  1977b).   Well water  with high nitrate levels
and coliform  counts had nitrosamine  concentrations of  less  than 0.015 ug/1
(U.S.  EPA,  1977).   Non-volatile  nitrpsamines have  been tentatively identi-
fied in  New Orleans drinking water  at  levels of 0.1  to 0.5  ug/1  (Fine,  .-.at
al. 1976).
         Contamination  of water  can occur  both from  industrial wastewater
and from agricultural runoff.
     3.  Food
         Nitrosamines have  been  found  in  foods, particularly in meats such
as sausages, ham, and bacon  which have been  cured  with nitrite.   N-nitroso-
dimethylamine  was present in a variety of  foods in the  1  to  10  ug/kg  range
and occasionally at levels up  to 100 ug/kg  (Montesano  and  Bartsch,  1976).
N-nitrosopyrrolidine  has been  consistently  found   in  cooked bacon in  the
range of 10-50 ug/kg (Fine,  et al. 1977a).
         Many  food constituents can  either  be converted directly  to N-nitro-
so compounds or give  rise  to nitrosatable products  after a metabolic inter-
                                                                     #
mediate step which can be involved directly  or  indirectly in  such reactions.
                                  ,. / / o^L^
                                  7 0 ^
                                      *
                                137-6

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  Constituents  include nitrate, nitrite, some amino acids,  choline,  phospholi- •
  pids,  purines,  pyrimidines,  some vitamins,  caffeine,  and some  pesticides
  (Walters,  1977;  Elsperu and Lijinsky,  1973).
           Nitrate and nitrite are well supplied  in the  diet.  Eighty-six per-
  cent  of the  nitrate ingested  comes  from  vegetables;  9  percent comes  from.
  cured  meats.   Only 2 percent of the  nitrite ingested  comes from vegetables,
  while  21 percent comes  from cured meat (White,  1975).
           The  U.S. EPA  (1979) has  estimated  the weighted  average  bioconcen-
  tration  factor to be  500 for N-nitrosodiphenylamine  in the edible  portions
  of  fish  and shellfish consumed by Americans.  This  estimate is based on mea-
  sured  steady-state  bioconcentration  studies with bluegills.   Based on  the
  octanol/water partition coefficient  for. each compound,  the U.S. EPA  (1979)
  has  estimated weighted average  bioconcentration factors  for the  following-.
  compounds  in the  edible portions  of  fish and  shellfish consumed by  Ameri-
  cans:  N-nitrosodimethylamine,  0.06;  N-nitrosodiethylamine,  0.39;  N-nitroso-
  di-n-butylamine,  4.9; and N-nitrosopyrrolidine,  0.12.
       C.  Inhalation
           Due  to  the  photolabile  nature  of nitrosamines,   concentrations  in
•  ambient air are  very  low,  except near sources of direct emissions of nitros-
  amines  (i.e.  chemical  plants)  (Fine,  at al. 1977a).   Nitrosamines  were  de-
  tected only  twice at 40 collection points  in New Jersey  and New York  City,
  and then only below the 0.01  ug/m  level.
           Tobacco  and  tobacco  smoke contain  both secondary amines and nitros-
  amines  (Hoffman,  st al.  1974).   The  intake of  nitrosamines from smoking  20
  cigarettes per  day has been  estimated  at approximately  6  pg/day (U.S. EPA,
  1979).

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III. PHARMACOKINETICS
     A.  Absorption
         Pertinent data could not be  located  in  the  available  literature.
     B.  Distribution
         Following intravenous  injection into rats, nitrosamides and  nitros-
amines  are  rapidly  and  fairly  uniformly  distributed  throughout  the  body
(Magee, 1972;  Stewart,  et al.  1974).   Both nitrosamides and nitiosamines  ap-
pear to cross the placenta since they  induce neoplasms  in offspring if  ad-
ministered maternally to  rats in  late pregnancy  (Magee, et  al.  1976).
     C.  Metabolism and Excretion
         Nitrosamides are rapidly metabolized in animals and excreted in  the
urine within  24 hours (Magee, et al.  1976)..
         Nitrosamines are metabolized  less  rapidly and persist  in the  body
unchanged  for a  longer  period.  The  rate  of  metabolism  depends  upon  the
chemical structure (U.S.  EPA, 1979).
         After  administration  of    C-labeled  dimethylnitrosamine,  diethyl-
nitrosamine,  or  nitrosomorpholine,  the  amount  of  isotope  appearing  as
14
  C02  within 12  hours  is  60,  45,   and 3  percent,  respectively,  while  the
corresponding urinary excretions  are  4, 14,  and 80 percent.  Urinary metabo-
lites  include  other  nitroso  compounds  formed by  oxidation  of  the  alkyl
groups to the alcohols  and carboxylic acids  (Magee,  et al. 1976).  Dimethyl-
nitrosamine is excreted in the milk of  female rats (Schoental,  et al. 1974).
         The  liver appears to be the major  site for' metabolism  of nitrosa-
mines;  kidney and lung  also  metabolize nitrosamines  ('Magee,   et  al.  1976).
The metabolites  of nitrosamines  are  thought  to be the  active teratogenic,
                                                                          »
mutagenic and carcinogenic forms (U.S. EPA, 1979).

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IV.  EFFECTS
     A.  Carcinogenicity
         The  epidemiological studies conducted  to date have been  inadequate
to  establish  any  correlation  between  exposure  to  N-nitroso  compounds  or
their precursors and human cancer  (U.S.  EPA,  1979).
         In  animals,  nitrosamines  and  nitrosamides  are  potent  carcinogens,
inducing  tumors in essentially  all vital  organs via  all routes of  admini-
stration (Montesano and Bartsch, 1976; Druckrey,  at al. 1967).
         Many  of the N-nitroso  compounds which have been  tested  are carcino-
genic.  There is a strong  relationship between  chemical  structure and  type
of  tumors  produced.   Symmetrically  substituted  dialkylnitrosamines and  some
cyclic nitrosamines produced carcinomas.of the liver.  Asymmetrical dialkyl-
nitrosamines  produced  carcinomas  of the esophagus (Druckrey,  et al.  1967)r<
Apparently  all  N,N-dialkylnitrosamines  containing  a  tert-butyl  group   are
noncarcinogenic (Heath and Magee,  1962).
         There  are large  differences  in   species  response  to  carcinogenic
nitrosamines  and nitrosamides, both in  type of tumor produced and in  suscep-
tibility, but all  animal  species tested are  vulnerable.   The late fetus  and
neonate  appear  to be  highly  susceptible  (U.S.  EPA,  1979).    Exposure  to
nitrosamides  during pregnancy may  result in a risk not only to the immediate
offspring, but also  for at  least  two more generations of animals (Montasano
and Sartsch,  1976).  There is no evidence  to indicate that nitrosamines pose
a similar threat (U.S.  EPA,  1979).
         Daily  oral  doses   of  N-nitroso  compounds 'of  2.5  percent  of   the
LDjQ  values  were  sufficient  to  induce cancer  in   rats  (Druckrey,  et   al.
1967).
                                3 7-

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     8.  Mutagenicity
         The N-nitroso  compounds  include some of the most  powerful  mutagens
known.  Nitrosamides are  mutagenic  in almost all  test  systems,  due to  non-
                  i
enzymic  formation  of   active  degradation  products.   Nitrosamines  must  be
metabolically activated to be mutagenic in microbial assays  (U.S.  EPA,  1979).
         Oimethylnitrosamine  and  diethylnitrosamine  have  been  reported  to
induce forward and reverse  mutations in §_._ typhimurium, E^ coli,  Neurospora
crassa and  other  organisms; gene  recombination  and conversion in  Saccharo-
myces cerevisiae;  "recessive lethal mutations" in Drosophila;  and  chromosome
aberrations in mammalian  cells (Montesano and Bartsch,  1976).  Negative  re-
sults were obtained in the mouse dominant lethal  test.
     C.  Teratogenicity
         N-nitroso  compounds  can  be  potent  teratogens  (U.S.  EPA,  1979.).
Nitrosamides are  teratogenic over an extended period  of gestation, whereas
nitrosamines are  active only when administered late in pregnancy  (Druckrey,
1973) probably because  of the inability of  the  embryonic tissue to metabo-
lize nitrosamines during early pregnancy (Magee,  1973).
     0.  Other Reproductive Effects
         Nitrosamines and  nitrosamides are embryotoxic (Druckrey,  1973).
     E.  Chronic  Toxicity
         The livers of  rats  and other species chronically exposed to nitros-
amines exhibit pathological changes including biliary hyperplasia,  fibrosis,
nodular parenchymal  hyperplasia,  and  the formation of enlarged hepatic par-
enchymal cells with large  nuclei (Magee,  et al. 1976).
                                137-1°

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     F.  Other Relevant Information
         Unlike  nitrosamines ,  nitrosamides cause  tissue injury  at  the site
of contact  (Magee, et  al.  1976).   This is thought to be  due  to the nonenzy-
matic  decomposition  of nitrosamides  into active products  upon contact with
tissues.
         Aminoacetonitrile , which  inhibits  the metabolism of dimethylnitros-
amine,  prevented the toxic and  carcinogenic  effects  of dimethylnitrosamine
in rat- liver (Magee,  et al. 1976).
         Ferric  oxide,  cigarette  smoke,  volatile  acids,  aldehydes,  methyl
nitrite, and  benzo(a)pyrene have been  suggested to act  in a cocarcinogenic
manner  with dimethylnitrosamine (Stenback, et al. 1973; Magee, et al. 1976).
V.   AQUATIC TOXICITY
     A.  Acute Toxicity
         The LC5Q  value  of 5,850  ug/1 for bluegill sunfish  (Lsoomis  macro-
chirus) exposed  to N-oitrosodiphenyiamine represents the  sole acute  toxicity
data  for  freshwater fish,  while an  LCCQ value of  7,760 ,ug/l  was  obtained j
for the freshwater invertebrate, Daohnia  maqna (U.S. EPA, 1978).  The  marine
mummichog  (Fundulus  heteroclitus) was  relatively  resistant  to N-nitrosodi-
methyiamine  in  a  96-hour  static  test,  where  an  adjusted  LC5Q  value  of
3,300,000 ug/i was reported (Ferraro,  et  al.  1977).  NO  additional  data con-
cerning marine organisms was presented  in the Ambient  Water Quality  Criteria
Document (U.S. EPA, 1979).
     3.  Chronic Toxicity
         The chronic effects  of M-nitrosodiphenyl  anrine have  been  examinee
in Daohnia maqna, with no  adverse  effects being reported at  a concentration
                                                                        •
of 48  ug/1.   NO chronic data  concerning  marine organisms were  found  in  the
available literature.
                                   -LL 1 / -
                                    ^^^^^^^^^^
                                      ?
                                 137-U

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     C.  Plant Effects
         Pertinent data could not be located in the  available  literature.
     D.  Residues
         A bioconcentration  factor  of  217  was reported,  as was.a  biological
half-life of less than .one day in the  freshwater bluegill  sunfish  (U.S. EPA,
1978).  No data  on  residues  in marine life were found in  the available lit-
erature.
     E.  Miscellaneous
         Shasta  strain  rainbow   trout (Salmo  gairdneri)  fed  N-nitrosodi-
methylamine in their  diet  for 52 weeks developed a dose-response  occurrence
of  hepatocellular  carcinoma  at  doses  of  200,  400,  and  800 mg N-nitrosodi-
methylamine per kg body weight (Grieco, et  al.  1978).
VI.  EXISTING GUIDELINES AND  STANDARDS                                      ^
     Neither the human health nor  the  aquatic  criteria  derived by U.S. EPA
(1979), which are summarized  below,  have gone through the process of public
review;  therefore,  there  is  a  possibility   that  these  criteria will  be
changed.
     A.  Human
         Using the  "one-hit" model, the U.S.  EPA  (1979)  has estimated the
following levels of nitrosamines in ambient water which  will  result in spe-
cified risk levels  of human cancer.
         The water  concentration of dimethylnitrosamine  corresponding  to  a
lifetime cancer  risk  for humans of  10~  is 0.026.-ug/l,  based on  the induc-
tion of liver tumors in rats  (Druckrey,  1967).
         The water  concentration of dibutylnitrosamine  corresponding  to  a
                                        c                              •
lifetime cancer  risk  for humans  of 10   is 0.013  ug/1,  based  on induction
of tumors of the bladder and  esophagus  in mice  (Bertram and Craig,  1970).
                                  -M&r-
                                13

-------
         The water concentration of  N-nitroso-pyrrolidine corresponding to a  •
lifetime cancer  risk for humans  of 10    is  0.11 ,ug/l,  based  on the  induc-
tion of hepatocellular carcinomas in rats  (Preussman, et  al. 1977).
         NO other guidelines or standards  are available.
     B.  Aquatic
         No criteria  for freshwater or  marine  life have been  drafted  (U.S.
EPA, 1979).

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                         NITROSAMINES

                          REFERENCES

Bertram, J-.S., and A.W. Craig.  1970.  Induction of bladder
tumours in mice with dibutylnitrosamine.  Br. Jour. Cancer
24: 352.

Bogovski, P., et al.  1972.  N-nitroso compounds, analysis
and formation.  IARC Sci. Pub. No.  3.  Int. Agency Res.
Cancer, Lyon, France.

Druckr.ey, H., et al.  1967.  Organotropic carcinogenic action.
of 65 different N-nitroso compounds  in BD rats.  Z. Krebs-
forsch.  69: 103.

Druckrey, H.  1973.* ,  Specific carcinogenic and teratogenic
effects of "indirect" alkylating methyl and ethyl compounds,
and their dependency on stages of oncogenic development.
Xenobiotica 3: 271.

Elsperu, R., and W. Lijinsky.  1973.  The formation of N-
nitroso compounds from nitrite and  some agricultural, chemi-
cals.  Food Cosmet. Toxicol. 11: 807.

Ferraro-, A.F., et al.  1977.  Acute  toxicity of water-borne
dimethylnitrosamine (DMN) to Fundulus heteroclitus (L).
Jour.- Fish. Biol. 10: 203.

Fine, D.H., et al.  1976.  N-Nitroso compounds in air and
water.  IARC Sci. Publ. No. 14.  Int. Agency Res. Cancer,
Lyon, France.
              • «•»
Fine, D.H., et al.  1977a.  Human exposure to N-nitroso
compounds in the environment.  In:   H.H. Hiatt, et al.,
eds.  Origins of human cancer.  Cold Spring Harbor Lab.,
Cold Spring Harbor, New York.

Fine, D.H., et al.  1977b.  Determination of dimethylnitro-
samine in air, water and soil by thermal energy analysis:
measurements in Baltimore, Md.  Environ. Sci. Technol. 11:
581.

Grieco, M.P., et al.  1978,  Carcinogenicity and acute toxi-
city of dimethylnitrosamine in rainbow trout (Salmo gaird-
neri).  Jour. Natl. Cancer Inst. 60: 1127.

Heath, D.F., and P.N. Magee.  1962.  Toxic properties of
dialkylnitrosamines and some related compounds.  Br.  Jour!
Ind. Med. 19: 276.

-------
                         N1.TSOSAM1NES

                          REFEBENC2S

Bertram, 3.S., and A.W. Craig.  1970.  Induction  of  bladder
tumours in mica with dibutylnitrosamine.  Br. Jour.  Cancer
24: 352..

Bogovski, P., et al.  1972.  N-nitroso compounds,  analysis
and formation.  IARC Sci. Pub. Ho.  3.  Int. Agency Res.
Cancer, Lyon, France.

Druckrey, H., et al.  1967-  Qrganotropic carcinogenic action.
of 65 different N-nitroso compounds  in 3D rats.   2.  Krebs-
forsch.  69: 103.'

Druckrey> H.  1973.*   Specific carcinogenic and teratogenic
effects of "indirect" alkylating methyl and ethyl compounds,
and their dependency on stages of oncogenic development..
Xenobiotica  3: 271.

Slsperu, R., and W. Lijinsky.  L973.  The formation  of N-
nitroso compounds from nitrite and  some agricultural, chemi-
cals.  Food Cosmet. Toxicol. 11: 807.

Ferrara,'A.F., at al.  1977.  Acute  toxicity of water-borne
dimethylnitrosamine (DMN) to Fundulus heteroclitus (L).
Jour.- Fish. Biol. 10: 203.

Fine, D.H.,  et al.  1976.  N-Nitroso compounds in air and
water.  IARC Sci. Publ. No. 14.  Int. Agency Res.  Cancer,
Lyon, France.
              > <»
Fine, D.H.,  et al.  1977a.  Human exposure to N-nitroso
compounds in the environment.  In:   H.H. Hiatt, et al.,
eds.  Origins of human cancer. ~"£old Spring Harbor Lab.,
Cold Spring Harbor, New York.

Fine, D.H., et al.  1977b.  Determination of dimethylnitro-
samine in air, water and soil by thermal energy analysis:
measurements in Baltimore, Md.  Environ. Sci. Technol. 11:
581.

Grieco, M.P., et al.  1978.  Carcinogenicity and  acute toxi-
city of dimethylnitrosamine in rainbow trout (Salmo  gaird-
neri).  Jour. Natl. Cancer Inst. 50: 1127.           "

Heath, D.F., and P.N. Magee.  1962.  Toxic properties of
dialkyinitrosamines and some related compounds.   Br. Jour'
Ind. Jled. 19: 276.

-------
Hoffman, D., et al.  1974.  Chemical studies on  tobacco
smoke.  XXVI.  On the isolation and identification of vola-
tile and non-volatile N-nitrosamines and hydrazines  in ciga-
rette smoke.  In; N-Nitroso compounds  in the environment.
IARC Sci. Pub. No. 9.  Int. Agency Res. Cancer,  Lyon, France.

Magee, P.N.  1972.  Possible mechanisms of carcinogenesis
and mutagenesis by nitrosamines.  In;  W. Nakahara,  et al.,
eds.  Topics in chemical carcinogenesis.  University of
Tokyo Press, Tokyo.

Magee, P.N.  1973.  Mechanisms of transplacental carcino-
genesis by nitroso compounds.  In;  L.. Tomatis and U. Mohr,
eds.  Transplacental carcinogenesis,   IARC Sci.  Pub. No. .
4.  Int.  Agency Res. Cancer, Lyon, France.

Magee, P.N., et al.  1976.  N-Nitroso  compounds  and  related
carcinogens.  In;  C.S. Searle, ed.  Chemical Carcinogens.
ACS Monograph No. 173.  Am. Chem. Soc., Washington,  D.C.

Montesano, R., and H. Bartsch.  1976.  Mutagenic and carcino-
genic N-nitroso compounds:  Possible environmental hazards.
Mutat.  Res. 32: 179.

Preussmann, R., et al.  1977.  Carcinogenicity of N-nitroso-
oyrrolidine:  Dose-response study in rats.  Z. Krebsforsch.
90: 161.              "  ..

Schoental, R., et al.  1974.  Carcinogens in milk:   Transfer
of ingested diethylnitrosamine into milk by lactating rats.
Br. Jour. Cancer 30: 238.

Stenback, ?., et al.  1973.  Synergistic effect  of ferric
oxide on dimethylnitrosamine carcinogenesis in the Syrian
golden hamster.  Z. Krebsforsch. 79: 31.

Stewart, 3.W., et al.  1974.  Cellular injury and carcino-
genesis..  Evidence for the alkylation  of rat liver nucleic
acids in vivo by N-nitrosomorpholine.  Z. Krebsforsch. 82:
1.

Tate, R.L., and M. Alexander.  1975.   Stability  of nitro-
samines in samples of lake water, soil and sewage.   Jour.
Natl.  Cancer Inst. 54: 327.

U.S. EPA.  1976.  Assessment of scientific information on
nitrosamines.  A report of an ad hoc study group of  the
U.S. Environ. Prot. Agency Sci. Advis. Board Executive Comm.
Washington, D.C.

U.S. EPA.  1977.  Scientific and assessment report on nitro-
samines.  EPA 600/6-77-001.  Off. Res. Dev. U.S. Environ.
Prot. Agency, Washington, D.C.
                          • • i / si i
                          **/ uj>

-------
U.S. EPA.  1978.  In-depth studies on health and  environ-
mental impacts of selected water pollutants.  Contract No.
68-01-4646.  U.S. Environ. Prot. Agency.

U.S. EPA.  1979.  Nitrosamines:  Ambient Water Quality Cri-
teria (Draft) .

Walters/ C.L.  1977.  Nitrosamines - environmental carcinogens?
Chem. Br. 13: 140. '

White, J.W,, Jr.  1975.  Relative significance of" dietary
sources of nitrate and nitrite.  Jour. Agric. Food Chem.
23: 886.            " ~"

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                                      No.  138
       N-Nitrosodiphenylamina

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report  is  drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources, this  short profile
may not reflect  all available  information  including all  the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny  to
ensure its technical accuracy.
                              - i /  '

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                      N-NITROSODIPHENYLAMINE




SUMMARY

     Formation of N-nitrosodiphenylamine  (NDPhA) has been  shown

experimentally in the stomachs of individuals receiving nitrite

and diphenylamine.  N-nitrosodiphenylamine  undergoes photochemi-

cal decomposition in -solution or in the atmosphere in the

presence of sunlight.  Bacterial degradation of NDPhA has  been

demonstrated in soil.

     Prior to the release of recent findings from the NCI  bio-

assay program, NDPhA was considered a non-carcinogenic nitros-

amine.  In the NCI lifetime rat feeding study, however, NDPhA was

found to induce a significant incidence of  urinary bladder tumors

in both males and females.  Few urinary bladder tumors were

observed in mice in a similar experiment, although there was a

high incidence of non-neoplastic bladder  lesions.

     N-nitrosodiphenylamine has consistently been found negative

in a variety of mutagenicity assays.



I.   INTRODUCTION

     This document is based on the Ambient  Water Quality Criteria

Document on Nitrosamines (U.S. EPA, 1979b),..the Scientific and

Technical Assessment Report on Nitrosamines (U.S. EPA, 1977), and

other selected references.  The term "N-nitrosodiphenylamine"
                                                            »
(NDPhA) in this report refers specifically  to that compound; the

term "nitrosamine" when used in this report refers to nitrosa-

mines in general.

-------
    • N-nitrosodiphenylamine  (NDPhA; molecular weight 198.23;

molecular formula C]_2H10N2°^ ^s a yellow  to brown  or orange

powder or flakes.  It has the  following physical/chemical

properties  (Hawley, 1977):

          Melting Point:               64-66°C

          Solubility:                  insoluble in  water;

                                       soluble in  organic ~
                          •
                                       solvents.

     NDPhA  is used as a vulcanization retarder in  the rubber

industry (Hawley, 1977).

     A review of the production range (includes importation)

statistics  for N-nitrosodiphenylamine (CAS No.  86-30-6)  which  is

listed in the initial TSCA Inventory (1979a)  has shown that

between 400,000 and 900,000 pounds of this chemical  were pro-

duced/imported in 1977.-!/



II.  EXPOSURE

     A.   Formation

     The chemistry of formation of nitrosamines is quite complex,

however,  they are in general formed by the combination of amines

(Rj_R2N-)  with some nitrosating agent.  Formation has  been shown

to occur with primary, secondary, and tertiary amines, as well as
   This production range information does not include any produc-
   tion/importation data claimed as confidential by the person(s)
   reporting for the TSCA Inventory, nor does it include any*
   information which would compromise Confidential Business
   Information.  The data submitted for the TSCA Inventory,
   including production range information, are subject to the
   limitations contained in the Inventory Reporting Regulations
   (40 CFR 710).
                              Tlf ' "7 ' —
                             ^ ) U -J u

-------
other amino compounds.  The nitrosating agent can be  derived  from




nitric oxides (NO, NO2, ^2°3' or N2°4^ or inor9anic nitrite  (U.S.



EPA, 1977)'.




     The in vivo formation of nitrosamines  following  ingestion  of



precursors has been demonstrated in human and animal  studies



(U.S. EPA, 1977)  Sander and Seif  (1969) showed the formation of




NDPhA in the stomachs of humans given- nitrite and diphenylamine.



     B.   Environmental Fate



     In the absence of light, nitrosamines  are quite  stable and



will decompose hydrolytically only following prolonged contact



with strong acid.  There is no evidence of  thermal instability  of



nitrosamines in the gas phase? however, they do undergo photo-



chemical decomposition in solution or in the atmosphere in the



presence of sunlight or ultra-violet light  (U.S. EPA, 1977).



     Transnitrosation reactions involving direct transfer of  the



nitroso group from NDPhA to other amines have been demonstrated



(Challis and Osborn, 1972).  Such a reaction yields-a new



N-nitroso compound and diphenylamine.



     In unattended soil, 70% of added NDPhA  was lost within 30



days.  In soil amended with bacteria, added NDPhA had disappeared



completely at the end of day 10 (Mallik, 1979).



     C.   Bioconcentration



     See Section V.C.

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III. PHARMACOKINETICS

     Intestinal bacteria  common in the gastrointestinal tract of

many animals and humans have been  shown capable of degrading

NDPhA  (Rowland and Grasso,  1975).




IV.  HEALTH EFFECTS  IN MAMMALS

     A.   Carcinogenicity

     In a one-year study, Argus and Hoch-Ligeti (1961)  adminis-

tered  NDPhA by gavage to  25 male rats  for 45 weeks (total dose
                                               • •
244 mg/rat).  So tumors were observed.  In other studies (Boyland

.et^jal.', 1968) no tumors were seen  when 20 rats were given the

test compound in the diet for 100  weeks at daily doses  of 120

mg/kg, or when 24 male rats were administered DNPhA by  intra-

peritoneal injection once per week for 6 months at a dose of 2.5

rag/week.  Both tests were terminated after 2 years.  When two

groups of mice (18 male and 18  female  per group) were admin-

istered NDPhA by gavage daily for  3 weeks at 1,000 mg/kg',  then in

diet at 3,769 ppm for 18  months, no significant incidences of

tumors were observed.  However,  in another assay reticulum cell

sarcomas were observed in the mice when the chemical was injected

subcutaneously (NCI, 1968;  Innes jst_ _al>.,  1969).  Druckrey et al.

(1967) reported a lack of tumorigenicity in rats administered 120

mg/kg/day of NDPhA for 700  days, for a tota"! dose of 65 g/kg.

Taken  together, these studies were viewed as a demonstration of

the non-carcinogenicity of  NDPhA.

     Recent results  from  the NCI bioassay program, however,  have

demonstrated that NDPhA is  a carcinogen in rats (NCI,  1979;  Cardy
                                 y

-------
^t_ jal_. / 1979).   In  these  studies  NDPhA was administered in the

diet to rats and mice  at  two  doses,  the "maximum tolerated dose"

for each species and one-half that amount.  Groups of 50 animals

of each sex were tested at  each dose for approximately 100

weeks.  The study  found that  dietary exposure to NDPhA gave rise

to a significant incidence  of urinary bladder tumors in both male

(40%)  and  female (90%) rats.   Pew urinary bladder tumors.were

observed in the mice,  although there was a high incidence of non-

neoplastic bladder  lesions.   The  authors (Cardy _et_ ^1_. , 1979)

ascribed the strong carcinogenic  effect seen in rats in this

study  to the higher doses used; they estimated that the maximum

daily  intake of NDPhA  was 320 mg/kg in females and 240 mg/kg in

males.  These  levels are  somewhat higher than those used by

Druckrey ^t_ ^1_.  (1967) in the only other known chronic feeding

study  done in  rats.

     B.   Mutagenicity

     NDPhA has consistently been  reported negative in a variety

of mutagenicity assays:   S. typhimurium (Ames test), with and

without activation  (Yahagi  e-t^ al_. ,  1977; Bartsch _et_ _al_. ,  1976;

Simmon, 1979a; Rosenkranz and Poirier,  1979); E. coli, with

activation- (Nakajima _et_ _al_.,  1974);  (Pol A~) E. coli (Rosenkranz

and Poirier, 1979); N. crassa (Marquardt £t^-_al_., 1963); Chinese

hamster V79 (lung)  cell line,  with and without activation (Kuroki

et al., 1977); Saccharomyces  cerevisiae D3,  with activation
                                                             •
(Simmon, 1979b); host  mediated assay (tester strains: S.

typhimurium and S.  cerevisiae D3) (Simmon et al.,  1979);  in vivo

mouse  testicular DNA synthesis assay (Friedman and Staub,  1976).

-------
      C.    Other Toxicity

      The oral LD50 in rats is 1650 rag/kg; in mice the oral LD50

 is  3,850 mg/kg (NIOSH,  1978).




 V.    AQUATIC EFFECTS

      A.    Acute

      The 96-hour LC^Q for NDPhA in bluegill sunfish under static

 test  conditions is 5.9  mg/1 (nominal concentration).  The-48-hour

 ECcQ  (static conditions) in Daphnia magna is 7.7 rag/1 (nominal

 concentration).  The adjusted 96-hour LCgg for the mummichog (a

 marine fish) under static conditions is 3,300 mg/1 (nominal con-

 centration)  (U.S.  EPA,  1979b).

      B.    Chronic

      No  adverse effects were reported at any test concentration

 in  a  chronic toxicity study in Daphnia magna at concentrations

 below 0.048  mg/1 (U.S.  EPA, 1979b).

      C.    Other

      Bioconcentration of NDPhA by bluegill sunfish reached equi-

 librium  within 14 days; the bioconcentration factor was 217.  The

 half-life of the compound in bluegill sunfish was less than one

'day (U.S.  EPA,  1979b).



 VI.   EXISTING GUIDELINES

      Criteria for the protection of aquatic species from excess
                                                             »
 NDPhA exposure have not been established (U.S. EPA, 1979b).

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                            REFERENCES

Argus, M.F., and Hoch-Ligeti, C.   1961.  Comparative  study of the
carcinogenic activity of nitrosamines.   J.  Natl.  Cancer Inst. 27,
695.

Bartsch, H., C. Mala-veille, and R. Montesano.  1976.   The  predic-
tive value of tissue-mediated mutagenicity  assays  to  assess the
carcinogenic risk of chemicals.   IARC  Scientific  Publications
(Lyon), No. 12, 467.

Boyland, E. , R..L. Carter, J.W. Gorrod, and  F.J.C.  Roe.   1968.
Carcinogenic properties of certain rubber additives.   Europ.  J.
Cancer 4_, 233.  (as cited in NCI,  1979).

Cardy, R.H., W. Lijinsky, P.K. Hilderbrandt.   1979.   Neoplastic
and non-plastic urinary bladder lesions  induced  in Fischer 344
rats and B6C3F, hybrid mice by N-nitrosodiphenylamine.   Ectotox-
icol. Env. Safety, J_(D, 29.

Challis, B.C. and M.R. Osborn.  1972.  Chemistry of nitroso
compounds.  The reaction of N-nitrosodiphenylamine with N-methyl-
aniline—a direct transnitrosation.  Chem.  Comm.  518

Druckrey, H., R. Preussmann, S. Ivankovic,  and D.  Schmahl.   1967.
Organotrope carcinogene Wirkungen bei 65 verschiedenen N-Nitroso-
Verbindungen an BD-Ratten.  Z. Krebsforsch.  69,  103.   (as  cited
in NCI, 1979 and Cardy _et_ _al_., 1-979).

Friedman, M.A., J. Staub.  1976.   Inhibition of  mouse  testicular
DNA synthesis by mutagens and carcinogens as a potential simple
mammalian assay for mutagenesis.  Mutat. Res., 37(1),  67-76.

Hawley, G.G.  1977.  The Condensed Chemical Dictionary,  9th ed.,
Van Nostrand Reinhold Co.

Innes, J.R.M., B.M. Ulland, M.G. Valerio, L. Petrucelli,
L. Fishbein, E.R. Hart, A.J. Pallotta. R.R. Bates,  H.L.  Falk,
J.J. Gart, M. Klein, I. Mitchell, and J. Peters.   1969.  Bioassay
of pesticides and industrial chemicals for  tumorigenicity  in
mice:  a preliminary note.  J. Natl. Cancer Inst.  42_(6), 1101-
1106.  (as cited in NCI, 1979).

Kuroki, T., C. Drevon, and R. Montesano.  1977.   Microsome-
mediated mutagenesis in V79 Chinese hamster cells  by  various
nitrosamines.  Cancer Res. 37, 1044-1050.

Mallik, M.A.  1979.  Microbial contribution to nitrosamine
formation in soil.  Smithsonian Scientific  Information Exchange
No. GY 70884 2.

Marquardt, H., R. Schwaier, 'and F. Zimmerman.  1963.   Nicht-
Mutagenitat von Nitrosamininen bei Neurospora  Crassa.   Natur-
wissenschaften _50_, 135.  (as cited in Cardy _et_ al_. , 1979).

-------
Nakajima, T., A. Tanaka, and K.I. Tojyo.  1974.  The  effect  of
metabolic activation with rat liver preparations on the mutagen-
icity of aeveral N-nitrosamines on a streptomycin-dependent
strain of Escherichia coli.  Mutat. Res. 26, 361-366.

National Cancer Institute.  1968.  Evaluation of Carcinogenic,
Teratogenic, and Mutagenic Activities of Selected Pesticides and
Industrial Chemicals'.  Vol. I.  Carcinogenic Study.   (as  cited  in
NCI, 1979).

National Cancer Institute.  1979.  Bioassay of N-Nitrosodiphenyl-
amine for Possible Carcinogenicity.  NIH Publication  No.  79-1720.

National Institute for Occupational Safety and Health.  1978.
Registry of Toxic Effects of Chemical Substances.

Rosenkranz, H.S. and L.A. Poirier.  1979.  Evaluation of  the
mutagenicity and DNA-modifying activity of carcinogens and non-
carcinogens in microbial systems.  J. Natl. Cancer Inst.  62,  873-
892.

Rowland, I.R. and P. Grasso.  1975.  Degradation of N-nitros-
amines by intestinal bacteria.  Appl. Microbiol. ^9.(1), 7-12.
(Abstract only).

Sander, J. and P. Seif.  1969.  Bakterielle reduction von nitrat
in Magen des Menschen als Ursoche einer nitrosaminbildung.
Arnz.-Forsch 19, 1091.-  (as cited in U.S. EPA, 1977).

Simmon, V.F.  1979a.  In vitro mutagenicity assays of chemical
carcinogens and related compounds with Salmonella typhimurium.
J. Natl. Cancer Inst. 62, 893-899.

Simmon, V.F.  1979b.  In vitro assays for recombinogenic  activity
of chemical carcinogens and related compounds with Saccharomyces
cerevisiae D3.  J. Natl. Cancer Inst. 62, 901-909.

Simmon, V.F., H.S. Rosenkranz, E. Zeiger _et_ _al_.  1979.  Mutagenic
activity of chemical carcinogens and related compounds in the
intraperitoneal host-mediated assay.  J. Natl. Cancer Inst.  62,
911-918.

U.S. EPA.  1977.  Scientific and Technical Assessment Report on
Nitrosamines, EPA-600/6-77-001.

U.S. EPA.  1979a.  Toxic Substances Control Act Chemical  Sub-
stances Inventory, Production Statistics for Chemicals on the
Non-Confidential Initial TSCA Inventory.
                                                             »
U.S. EPA.  1979b.  Ambient Water Quality Criteria:  Nitrosamines.
PB 292 438.'

Yahagi, T., M. Nagao, Y. Seino, T, Matsushima, T. Sugimura,  and
M. Okada.  1977.  Mutagenicities of N-nitrosamines on Salmonella.
Mutat. Res. 48, 121.
                             / 3

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                                      No. 139
     N-Nitrosodi-n-propylamine

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                      SPECIAL NOTATION
U.S. EPA1s  Carcinogen Assessment Group  (CAG) has evaluated
n-nitrosodi-n-p^opj^ylamine and has  found sufficient evidence
to indicate that this compound is carcinogenic.
                        /3?-3

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                    N-NITROSODI-n-PROPYLAMINE






SUMMARY



     The International Agency for Research on  Cancer has  con-



cluded that "N-nitrosodi-n-propylamine should  be  regarded for



practical purposes as if it were carcinogenic  in  humans."  The



conclusion is based on positive findings in several long-term



animal studies with the compound.  It has also been found muta-



genic in several test systems with activation.



     The chemistry of formation of nitrosamines is quite  complex,



however, they are formed in general by the combination of amines



with some nitrosating agent.  Nitrates, nitrites, and amines



(primary, secondary, and tertiary), the precursors in the



formation of nitrosamines, are ubiquitous in the  environment.



Significant quantities of the precursors are also produced'



through human activities.



     The in vivo formation of nitrosamines following ingestion of



precursors has been demonstrated in humans and animals.



     Nitrosamines degrade in the presence of sunlight; however,



in the dark they are quite stable.  Microorganisms can function



both in the formation and degradation of nitrosamines.  The half-



life of aliphatic nitrosamines in the environment ranges  from  one



hour in the atmosphere in sunlight to more'than 40 days in soils



and water (in the absence of light).







I.   INTRODUCTION



     This document is based on the Ambient Water Quality  Criteria



Document for Nitrosamines (U.S. EPA, 1979a), Volume 17 of the

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IARC Monographs on the Evaluation of  the  Carcinogenic  Risk  of

Chemicals to Humans  (IARC,  1978), the Scientific  and Technical

Assessment Report on Nitrosamines (U.S. EPA,  1977), and. other

selected references.. The term  "N-nitrosodi-n-propylamine"  (NDPA)

in this report refers specifically  to that  compound; the term

"nitrosamine" when used in  this  report refers  in  general to

'simple aliphatic nitrosamines.

     N-nitrosodi-n-propylamine  (NDPA; CgH^NjO; molecular weight

130.2) is a yellow liquid having the  following physical  chemical

properties (IARC, 1978).

          Boiling Point:          81°C

          Density:                d|° 0.9160


          Solubility:             soluble in  water, organic

                                  solvents, and lipids.

          Volatility:             can be  steam distilled

                                  quantitatively.

     A review of the production  range (includes importation)

statistics for NDPA  (CAS No. 621-64-7) which  is listed in the

initial TSCA Inventory (1979b) has  shown  that  between  zero  and

900 pounds of this chemical were intentionally produced/imported

in 1977.I/
*/
—' This production range  information does  not  include  any  pro-
   duction/importation data claimed as confidential by the per-
   son(s) reporting  for the TSCA  Inventory, nor does it include
   any information which  would compromise  Confidential Business
   Information.  The data submitted for  the TSCA  Inventory,
   including production range information, are subject to  the
   limitations contained  in the Inventory  Reporting Regulations
   (40 CFR 710).

-------
     No information on the commercial  uses  of  NDPA was  Located,
however, it appears likely that most,  if  not all,  of  that pro-
duced is used solely in the  laboratory.

II.  EXPOSURE
     Nitrates, nitrites, and amines  (in this case  the propyl-
amines), which are precursors in the formation of  nitrosamines,
are ubiquitous in the environment and  occur in food,  water,  soil,
and air.  The natural occurrence of nitrates,  nitrites,  and
secondary and tertiary amines results  from  their  formation during
the nitrogen cycle.  In addition to the naturally  formed precur-
sors , significant quantities are produced through  human activi-
ties (U.S. SPA, 1977).  Some of the major man-made sources of the
precursors are listed in Table 1.
     A.   Formation
     The chemistry of formation of nitrosamines is quite complex,
however, they are formed in general by the  combination  of amines
(R,R~N-) with some nitrosating agent.  Formation has  been shown
to occur with primary, secondary, and  tertiary amines,  as well as
other amino compounds.  The nitrosating agent  can  be  derived from
nitric oxides (NO, N02, ^2°3' or N2°4^ or inor
-------
 Table 1.   Man-Made Sources of Nitrosamine Precursors (U.S. EPA, 1977)

Nitric Oxides                               Amines
Transportation
  Motor vehicles
  Aircraft
  Railroads
Fuel combustion in stationary sources
 . .Coal
  Fuel Oil
  Natural gas
  Wood
Industrial processes
Solid waste disposal
Miscellaneous
  Forest fires
  Structural fires
  Coal refuse
  Agricultural
Feedlots
Rendering plants
Antioxidants
Vulcanization
  accelerators
Pharmaceuticals
Self-polishing waxes
Synthetic detergents
Pesticides
Solvents
Corrosion inhibitors
Animal glues
Photographic products
Leather tanning
Primary amine
  production
     The in vivo formation of nitrosamines  following the  inges-

tion of precursors has been demonstrated in human and animal

studies (U.S. EPA, 1976,').

     Nitrosamines can be formed in soil, water, and sewage  under

appropriate conditions (Ayanaba jt_ ^_1_., 1973a, b; Ayanaba and

Alexander, 1974? Kohl ^t_ ^1_. , 1971).   Microorganisms in soil and

water can participate in the formation of nitrosamines  (Ayanaba

.et_ .al_. , 1973b; Mills and Alexander, 1976),  although microbial

involvement in such formation reactions is  not essential  (Mills,

1976; Mills and Alexander, 1976).

     B.   Environmental Fate

     In the absence of light, nitrosamines  are quite stable and

will decompose hydrolytically only following prolonged contact

with strong acid.  There is no evidence of  thermal instability of

nitrosamines in the gas phase; however, they do undergo photo-
                             1.3.9-7

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chemical decomposition in solution or  in  the  atmosphere in the



presence of sunlight or ultra-violet light.   There  are very few



quantitative studies on the rate of photochemical degradation of



nitrosamines or on the rate effects of other  factors  (U.S.  EPA,



1977; IARC, 1978).  Nonetheless, it has been  shown  that N-nitro-



sodimethylamine has an atmospheric half-life  (during  ambient



atmospheric conditions) of between 30  minutes and one hour in



sunlight (Hanst _et_ _al_., 1977).  The atmospheric half-life of NDPA



should be similar (U.S. EPA, 1979a).



     N-nitrosodi-n-propylamine appears to be  fairly resistant to



microbial attack under environmental conditions.  The soil half-



life of NDPA under varying conditions  has been reported as rang-



ing between 10 and 40 days (Tate and Alexander, 1975; Saunders et



al., 1979; Oliver et_ _al_., 1978).  In lake water under laboratory



conditions, NDPA persisted for more than 4 months (Tate and



Alexander, 1975).



     A laboratory soil leaching study  (Saunders _et_  _al>.,  1979) has



indicated that NDPA (which is about 1% soluble in water)  will



leach under heavy simulated rainfall conditions.  In  a field



study, however, NDPA did not leach below a depth of 20 cm.   The



authors suggest that under field conditions,  NDPA is  dissipated



due to volatilization and degradation.



     C.   Bioconcentration



     No information on the bioaccumulation potential  of NDPA was



located, although it should be fairly  low.                   •



     D.   Environmental Occurrence



     NDPA has been detected in food, alcoholic beverages,and



s.everal pesticides (IARC, 1978).  It has also been  detected in




                              '"/ 13 J (!)'

-------
the waste-water from several chemical plants  (Cohen  and  Bachman,



1978).








III. PHARMACOKINETICS




     A.   Absorption'



     In goats, one hour after oral administration, NDPA  was  found



in milk and blood, indicating fairly rapid uptake.   Only traces



were found in the milk after 24 hours (Juszkiewicz and Kowalski,



1974).



     B.   Distribution



     No information was located on the distribution  of NDPA;



however, simple aliphatic nitrosamines tend to distribute rapidly



and fairly uniformly in the body  (U.S. EPA, 1979a).



     C.   Metabolism



     Available evidence suggests  that NDPA must be metabolically



activated to exert its toxic and  carcinogenic effects.   Urine



collected during the 43 hours after oral administration  of an



LD^g dose of NDPA to rats contained the following compounds:



N-nitroso-3-hydroxy-n-propyl-n-propylamine, N-nitroso-2-carboxy-



ethyl-n-propylamine, and to a lesser extent, N-nitrosocarboxy-



methyl-n-propylamine, and N-nitroso-2-hydroxy-n-propyl-n-



propylamine (Blattman and Preussmann, 1973).  The last named



metabolite, N-nitroso-2-hydroxy-n-propyl-n-propylamine,  has been



found carcinogenic in rats (Reznik et al., 1975) and hamsters



(Pour _et_ ^1_., 1974a,b) , thus it may be the active carcinogenic



metabolite (proximate and/or ultimate carcinogen) of NDPA.

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IV.  HUMAN HEALTH EFFECTS

     A.   Carcinogenicity

     Groups of rats were given NDPA in the drinking water  at

doses of 4, 8, 15, or 30 mg/kg day.  Of the 48 animals on  test,

45 developed liver carcinomas, 8 developed papillomas or car-

cinomas of the esophagus, and 6 showed carcinomas of the tongue

(Druckrey _et_ _al/ , 1967).

     Groups of rats were injected subcutaneously with 1/5,  1/10,

or 1/20 the LD5Q of NDPA (LD5Q:  487 mg/kg) once weekly for

life.  The average total dose of NDPA ranged between 0.93  and  2.7

g/kg.  A high incidence of neoplasms was observed in the nasal

cavities.  In addition, tumors of the liver, lung, kidney,  and

esophagus were observed (Althoff et al., 1973a; Reznik et  al.,

1975).

     Groups of Syrian golden hamsters were injected

subcutaneously with 1.2% NDPA in olive oil once weekly for life

at 5 dose levels (highest dose was 60 rag/kg).  Tumors were

observed in the nasal cavities, laryngobronchial tract, lungs,

and a variety of other organs (Althoff e_t_ £l_., 1973br Pour et

al., 1973).

     The International Agency for Research on Cancer  (1978) has

concluded:

     There is sufficient evidence of a carcinogenic effect of
     N-nitrosodi-n-propylamine in two experimental animal
     species.  Although no epidemiological data were avail-
     able. ..N-nitrosodi-n-propylamine should be regarded for
     practical purposes as if it were carcinogenic to humans,

     B.   Mutagenicity

     NDPA was positive in the Ames test (S. typhimurium strains

TA 1530, TA 1535, and TA 100) with activation (Barstch et  al.,

                              ~1 / ^ ^
                             * ] US 
-------
1976? Camus et_ &1_., 1976; Olajos  and  Cornish,  1976;  Sugimura et



al., 1976).  NDPA  was also mutagenic  in  E.  coli  (Nakajima et al.,



1974) and in Chinese hamster V79  cells  (Kuroki £t_ al_.,  1977),  in



both cases with activation.



     C.   Other Toxic Effects



     The acute oral LD5Q of NDPA  was  480 rag/kg in rats  (Druckrey



et_ _al_., 1967); the subcutaneous LD5Q  was 487 mg/kg in  rats and



600 mg/kg in hamsters (Pour _et_ ^1_. , 1973;  Reznik ^t_ _al_. ,  1975).








V. AQUATIC EFFECTS



     No data on the aquatic effects of NDPA were located.








VI.  EXISTING GUIDELINES



     The class of  compounds "nitrosamines"  was included  in.the



American Conference of Governmental Industrial Hygienists (1977)



list of "Industrial Substances Suspected of Carcinogenic  Poten-



tial for Man."  No threshold limit value (TLV) was given.



     As noted in Section IV.A, the International Agency  for



Research on Cancer (1978) has  concluded  that  "N-nitrosodi-n-



propylamine should be regarded for practical purposes  as  if it



were carcinogenic  to humans."

-------
                            REFERENCES

Althoff, J., F.W. Kruger, J. Hilfrich, D. Schmahl, and  U.  Mohr.
1973a.  Carcinogenicity of B-hydroxylated dipropylnitrosamine.
Naturwissenschaften, 60, 55 (as cited in IARC,  1978).

Althoff, J., F.W. Kruger, and U. Mohr.  1973b.   Carcinogenic
effect of dipropylnitroaamine and compounds related by  B-oxida-
tion.  J. Nat. Cancer Inst.., 51, 287-288 (as cited in IARC,
1978).

American Conference of Governmental Industrial  Hygienists,
Threshold Limit Values for Chemical Substances  and Physical
Agents in the Workroom Environment, 1977.

Ayanaba, A., W. Verstraete, and M. Alexander.   1973a.   Formation
of dimethylnitrosamine, a carcinogen and mutagen in soils  treated
with nitrogen compounds.  Soil Sci. Soc. Amer.  Proc. 37, 565-568.
(as cited in U.S. EPA, 1977).

Ayanaba, A., W. Verstraete, and M. Alexander.   1973b.   Possible
microbial contribution to nitrosamine formation in sewage  and
soils.  J. Nat. Cancer Inst. 50, 811-813.   (as  cited in U.S. EPA,
1977).

Ayanaba, A. and M. Alexander.  Transformation of methylamines  and
formation of a hazardous product, dimethylnitrosamine,  in  samples
of treated sewage and lake water.  J. Environ.  Qual. _3_, 83-89.
(as cited in U.S. EPA, 1979).                 t

Bartsch, H., C. Malaveille, and R. Montesano.   1976.  The  predic-
tive value of tissue-mediated mutagenicity assays to assess the
carcinogenic risk of chemicals.  In:  Montesano,  R., Bartsch,  H.
and Tomatis, L., eds.  Screening Tests in Chemical Carcinogen-
esis, Lyon (IARC Scientific Publications No.12), pp. 467-491.
(as cited in IARC, 1978).

Blattman, L. and R. Preussmann.  1973.  Struktur von metaboliten
carcinogener dialkylnitrosamine im rattenurin.   Z. Krebsforsch.,
79, 3-5.  (as cited in IARC, 1978).

Camus, A., B. Bertram, Kruger, F.W., C. Malaveille, and
H. Bartsch.  1976.  Mutagenicity of 3-oxidized  N,N-di-n-propyl-
nitrosamine derivatives in S. typhimurium.. mediated by rat  and
hamster tissues.  Z. Krebsforsch., 86, 293-302  (as cited in IARC,
1978).

Cohen, J.B. and J.D. Bachman.  1978.  Measurement of environ-
mental nitrosamines.  In:  Walker, E.A., Castegnaro, M.,
Gricuite, L. and Lyle, R.E., eds., Environmental Aspects of
ET-Mitroso Compounds, Lyon (IARC Scientific Publications No. 19) .
(as cited in IARC, 1978).

Druckrey, H., R. Preussmann, S. Ivankovic, D. Schmahl.  1967.
Organotrope carcinogene Wirkungen bei 65 verschiedenen  N-nitroso-
verbindungen an BD-ratten.  2. Krebsforsch., 69,  103-201.   (as
cited in IARC, 1978).

-------
Hanst, P.L., J.W. Spence, and M. Miller.   1977.  Atmospheric
chemistry of N-nitroso dimethylamine.  Env. Sci. Tech.,  11(4),
403.

Juskiewicz, T. and B. Kowalski, 1974. Passage of nitrosamines
from rumen into milk in goats.  In: Bogavski, P. and  E.A. Walker,
eds., N-Nitroso Compounds in the Environment, Lyon  (as  cited  in
IARC, 1978).

International Agency for Research on Cancer.  1978.   IARC Mono-
graphs on the Evaluation ..of the Carcinogenic Risk of  Chemicals to
Humans, Vol. 17.

Juskiewicz, T. and B. Kowalski, 1974. Passage of nitrosomines
from rumen into milk in goats.  In: bogaski, P. and E.  A. Walker,
eds., n-Nitroso Compounds in the Environment, Lyon  (as  cited  in
IARC, 1978).

Kuroki, T., C. Drevon, and R. Montesano.   1977.  Microsome-
mediated mutagenesis in V79 Chinese hamster cells by  various
nitrosamines.  Cancer Res., 37, 1044-1050.  (as cited in IARC,
1978).

Mills, A.L.  1976.  Nitrosation of secondary amines by  axenic
cultures of microorganisms and in samples  of natural  ecosystems.
Ph.D. Thesis. Cornell University, Ithaca,  New York. 95  pp.  (as
cited in U.S. EPA, 1977).

Mills, A.L. and M. Alexander.  1976.  Factors affecting dimethyl-
nitrosamine formation in samples of soil and water.   J.  Environ.
Qual. , _5_(4), 437.

Mirvish, S.S.  1977.  N-Nitroso compounds:  Their chemical and in
vivo formation and possible importance as  environmental carcino-
gens.  J. Toxicol. Env. Hlth. , 2_, 1267.

Nakajima, T., A. Tanaka, and K.I. Tojyo.   1974.  The  effect of
metabolic activation with rat liver preparations on the mutagen-
icity of several N-nitrosamines on a streptomycin-dependent
strain of Escherichia coli.  Mutat. Res.,  26, 361-366.   (as cited
in IARC, 1978).

Olajos, E.J. and H.H. Cornish.  1976.  Mutagenicity of
dialkylnitrosamines:  metabolites and derivatives (Abstract No.
43).  Toxicol. Appl. Pharmacol., 37, 109-110.   (as cited in IARC,
1978).

Oliver, J.E., P.C. Kearney, and A. Kontson.  1978.  Abstract
presented at the 175th National Meeting of the American Chemical
Society, Paper No. 80, Pesticide Division.  (as cited by Saunders
et_al_., 1979).

Pour, P., F.W. Kruger, A. Cardesa, J. Althoff, and U. Mohr.
1973.  Carcinogenic effect of di-n-propylnitrosamine  in Syrian
golden hamsters.  J. Nat.:Cancer Inst., 51, 1019-1027.   (as cited
in IARC, 1978).

-------
Pour, P., F.W. Kruger, A. Cardesa, J. Althoff, and U.  Mohr.
1974a.  Effect of beta-oxidized nitrosamines on  Syrian golden
hamsters.  I.  2-Hydroxypropyl-n-propylnitrosamine.  J.  Nat.
Cancer Inst., 52, 1245-1249.   (as cited in IARC,  1978).

Pour, P., J. Althoff, A. Cardesa, F.W. Kruger, and U.  Mohr.
1974b.  Effect of beta-oxidized nitrosamines on  Syrian golden
hamsters.  II.  2-Oxopropyl-n-propylnitrosamine.  J. Nat.  Cancer
Inst., 52, 1869-1874.  (as cited in IARC, 1978).

Reznik, G., U. Mohr, F.W. Kruger.  1975.  Carcinogenic effect  of
di-n-propylnitrosamine, B-hydroxypropyl-n-proylnitrosamine,  and
methyl-n-propylnitrosamine on  Sprague-Dawley rats.  J. Nat.
Cancer Inst., 54, 937-943.   (as cited in IARC, 1978).

Saunders, D.G., J.W. Mosier, J.E. Gray, and A. Loh.  1979.   Dis-
tribution and movement of N-nitrosodipropylamine  in soil.  J.
Agric. Fd. Chem., ^7J3), 584.

Sugimura, T., T. Yahagi, M. .Nagao, M. Takeuchi,  T,. Kawachi,
K. Hara, E. Yamakaki, T. Matsushima, Y. Hashimoto, and M.  Okada.
1976.  Validity of mutagenicity tests using microbes as  a  rapid
screening method for environmental carcinogens.   In:   Montesano,
R., Bartsch, H. and Tomatis, L., eds., Screening  Tests in  Chemi-
cal Carcinogenesis, Lyon (IARC Scientific Publication  No.  12),
pp. 81-101.  (as cited in IARC,. 1978).

Tate, R.L. and M. Alexander.   1975.  Stability of N-nitrosamines
in samples of lake water, soil, and sewage.  J.  Nat. Cancer  Inst.
54, 327-330.  (as cited in U.S. EPA, 1977).

U.S. EPA.  1977.  Scientific and Technical Assessment  Report on
Nitrosamines.  EPA-600/6-77-001.

U.S. EPA.  1979a.  Ambient Water Quality Criteria:  Nitrosamines.
PB 292 438.

U.S. EPA.  1979b.  Toxic Substances Control Act  Chemical Sub-
stances Inventory, Production  Statistics for Chemicals on  the
Non-Confidential Initial TSCA  Inventory.

-------
                                      No.  140
            Paraldehyde

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is  drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources, this  short profile
may not reflect .all available  information  including all the
adverse health  and  environmental impacts  presented by the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                                 PARALDEHYDE
                                   Summary

     There  is  no  evidence  in  the  available  literature  to indicate  that
paraldehyde, a  central nervous  system depressant,  is carcinogenic,  muta-
genic, or teratogenic.
     In low doses  (4-8 ml)  paraldehyde  has  a hypnotic effect on the central
nervous system.   Following chronic  and acute exposures  at higher  concen-
trations,  paraldehyde affects the respiratory and circulatory systems.
     Data concerning the  effects  of paraldehyde  on  aquatic organisms  were
not found in the available literature.
     Guidelines or standards concerning  air  or water  exposures were  not
found in the available  literature.
                                  If 0

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                                  PARALDEHYDE
 I.    INTRODUCTION
          ParaJdehyde,  2,4,6-trimethyl-l,3,5-trioxane,   also  known  as  para-
 acetaldehyde,  is  a colorless liquid with  a molecular weight of  132.2.  This
 compound  melts at  D°C  and boils  at  125°C.  It  has a specific  gravity  of
 0.994  at. 20°C,  and  its  solubility in  water is  120,000 mg/1 at  D°C  and
 58,000  mg/r'at 100°C  (Verschueren,  1577).  The  odor of paraldehyde is  not
 pungent  or  unpleasant,  but  it  is  characterized  by  a disagreeable  taste
                        „    r
 (Wilson,, .et  al. 1577).
          Paraldehyde  was  introduced  into medicine  by  Ceruello  in  1882  as
 the  second  synthetic  organic  compound  to be used   as  a sedative  hypnotic
 (Wilson,  et al.  1577).   It is  used frequently in  delirium tremens and  in
 treatment pf psychiatric states characterized  by excitement when  drugs must
 be given  over  a 'long period of  time (Wilson,  et al. 1577).   It also is  ad-
 ministered  for intractable  pain which does not respond  to  opiates  and  for
 basal  and  obstetrical  anaesthesia  (Goodman   and  Oilman,   1570).    It   is
 effective against  experimentally induced  convulsions and has  been  used  in
 emergency therapy  of tetanus,  eclampsia,  status epilepticus,  and poisoning
by convulsant drugs (Goodman and Oilman,  1970).
          It is  used  primarily  in medicine,  and therefore, the  chance  of
 accidental human exposure  or environmental contamination  is low.   However,
paraldehyde  decomposes to  acetaldehyde  and  acetic   acid  (Gosselin,  et  al.
 1576); these compounds have been  found to  be toxic-." In this sense,  occupa-
tional  exposure  or environmental  contamination  is  possible.   Since paral-
dehyde is prepared from acetaldehyde by polymerization in the presence of  an
                                                                         »
 acid catalyst,  there  exists a potential for adverse effects,  although none
have been reported in the  available literature.
                                      X
                                   JtJO'1/

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II.  EXPOSURE
          No monitoring data are  available  to  indicate  ambient  air  or water
levels of the compound.  Human exposure to par aldehyde from ingestion cannot
be  assessed, due  to a lack of monitoring data.  No  data on dermal exposure
of humans were found in the available  literature.
III. PHARMACOKINETICS
          Paraldehyde  is  rapidly  absorbed  from the  gastrointestinal tract
and  parenteral  sites.   Following  oral  administration to rats,  the  maximum
concentration  in  the  brain is  reached within  30 -minutes  (Figot,  et  al.
1953).   A  significant  percentage  is  excreted  unchanged through the  lungs.
Lang,  et al.  (1969)   reported that  human-subjects  given  unspecified  oral
doses  exhaled  7  percent  of the  administered  dose  within  4  hours.   Only
traces  are  observed in the urine;  the  rest is metabolized  by the  liver.
There is indirect evidence that paraldehyde  is depolymerized to  acetaldehyde
in the  liver, then  oxidized by  aldehyde  dehydrogenase to acetic acid  which,
in  turn,  is ultimately  metabolized   to  carbon dioxide  and water  in  mice
(Hitchcock and  Nelson,  1943).
          No data  on  bio accumulation  of  paraldehyde   were  found  in  the
available literature.  Based on the  evidence of metabolism  above,  however,
significant  bio accumulation would  appear unlikely.
IV.  EFFECTS
     A.   Carcinogenicity
          Paraldehyde   has  been designated  a  "suspect  carcinogen"   (NIOSH,
1978),   although  no increase  in   neoplasms  was  observed  in the mouse-skin
painting study  (Row and Salaman, 1955), which was cited by NIOSH.
     8.   Mutagenicity, Teratogenicity and Other Reproductive Effects.
          Pertinent data could not be  located in the available literature.
                                      s
                                  +}$t^~
                                 /V0-J

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   '  C.   Chronic Toxicity
          In  low  doses (4-8 ml), par aldehyde has  found use as a therapeutic
agent.   However,  if  used  for  a  prolonged  period  of  time,  intoxication
results  in  tolerance  and  dependence.   Paraldehyde addiction resembles alco-
holism;  withdrawal  may result in  delirium tremens  and vivid hallucinations
(Goodman and Gilman, 1570).
          Acidosis,  bleeding  gastritis,  muscular  irritability,  azotemia,
oliguria, albuminuria,  leukocytosis,  fatty changes  in the  liver  and kidney
with toxic  hepatitis and nephrosis, pulmonary  hemorrhages,  edema,  and dila-
tion of  the right heart  have  all been  observed  in cases of  chronic paral-
                                           *
dehyde poisoning.   Metabolic acidosis is  a manifestation  of paraldehyde in-
toxication  in  the paraldehyde  addict.   The etiology of the  acidosis is un-
certain (Beier, et al. 1963).
     0.   Acute Toxicity          , .
          Figot,  et  al.   (1953)  reported  an  oral LD50  Qf  1.55  g/t
-------
          Toxic  doses  of unspecified  amounts, given  intravenously,  cause
diffuse, massive pulmonary hemorrhages  and  edema,  as well as dilation of the
right heart.   Adverse  effects,  as seen in cases of  severe  acute paraldehyde
intoxication,  resemble those seen in chronically  exposed individuals,  e.g.,
addicts.
          Metabolic  acidosis. is., also found in the severe  acute  cases.   Hay-
ward and Boshell  (1957)  produced metabolic  acidosis  and other toxic effects,
including pulmonary  edema in dogs,  by administering unspecified  amounts  of
deteriorated  paraldehyde through gastric  tubes over a period of  18  hours.
In  this case  it  is  uncertain  whether the  paraldehyde or  the  deteriorated
product was the cause  of the observed  effects.   The sane is  true  in another
study where  a deteriorated  product (40 percent acetic  acid)  produced  sudden
death  with   intense  corrosion  of  buccal  mucosa  and  upper  air  passages.
Rectal  administration  (a common route in  therapeutic  settings) in  another
poisoning victim caused great pain and  sloughing of  rectal  mucosa (Gosselin,
                                                        i
et al.  1976).
          High concentrations  (unspecified)  depressed cholinergic  junctions
in frogs, apparently  by  reducing the amount of acetyleneline  liberated  from
nerve endings (Nicholls and Quillam,  1956;  Quillam,  1959).
          The lethal dose in humans  is disputable.   Less  than one ounce  by
mouth has been shown to be  lethal in some cases, while  others have  tolerated
four ounces.   Death  results from respiratory failure preceded by  prolonged
and profound coma (Goodman and Gilman,  1970).
          Paraldehyde  has been  used  in  obstetrics; "however,   it readily
crosses the  placental barrier and appears  in the  fetal circulation.  Unde-
                                                                         *
sirable  effects,   including  delay  in  respiratory  movements,   have  been
                                         -7

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   observed  in  neonates  following  administration  to  the mother  during  labor
   (Goodman  and  Gilman,  1970).   Consequently,  paraldehyde  finds  little or  no
   use in  obstetrics  today.
            The  lowest  dose  of  paraldehyde  reported  to  produce  any  toxic
   effect  (unspecified)  in  tiumsns  is  121 mg/kg.   Oral LD50  values have been
   reported  for the following species:  rats,  1530 mg/kg;  r^Dbits,  3304 mg/kg;
•   and dogs, 3500 mg/kg.  NIOSH (1978) has  reported  the lowest lethal  inhala-
   tion concentration to  be  2000 ppm.
   V.   AQUATIC TOXICITY
            Data concerning the effects of paraldehyde on aquatic organisms
   were not  found in  the  available  literature.
   VI.  EXISTING GUIDELINES  AND  STANDARDS
            NO exposure  limits  or standards  were found in the available liter-
   ature to  exist for air or water.  .

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                                  References
Beier,  L.*,  et  al.   1963.  Metabolic  acidosis occurring  during  paraldehyde
intoxication.   Chem. Abst.  59:  1022.

Figot,  P.,  et al.  1953.  Estimation  and  significance  of paraldehyde levels
in blood  and brain.  Chem. Abst.  47: 660.

Goodman,  L.  and A.  Oilman.   1970.   The Pharmacological  Basis of  Thera-
peutics.  4th ed.  MacMillan Co., New York.

Gosselin,  R.E., et  al.   1976.   Clinical  Toxicology of Commercial Products.
Williams  and Wilkins Co., Baltimore, Maryland.

Hayward,  J.  and B. Boshell.  1957.  Paraldehyde  intoxication with metabolic
acidosis.  Am.  Jour. Med.  23: 965.

Hitchcock,  P.   and  E.  Nelson.   1943.   The metabolism  of paraldehyde:   II.
Jour. Pharmac. Exp. Ther.  79: 286.

Kirk,  R.E.  and D.F. Othmer.    1979.  Encyclopedia of  Chemical  Technology.
John Wiley and  Sons, New York.

Lang, 0.,  et  al.  1969.  Data of pulmonary excretion of paraldehyde in man.
Chem. Abst.  71: 202.

Nicholls, J.  and J.  Quillam.   1956.  Mechanism of  action of paraldehyde and
methyIpentyno1  on  neuromuscular transmission  in  the  frog.   Chem.  Abst.
50: D295.                                                                   I

National  Institute  for  Occupational  Safety  and  Health.    1978.   Suspected
Carcinogens.  A Subfile on the  Registry  of Toxic  Effects of Chemical  Sub-
stance.  U.S.  Department of Health,  Education  and Welfare, Cincinnati,  Ohio.

Quillam,  J.   1959.   Paraldehyde and  methy Ipentyno 1 and  ganglionic  trans-
mission.  Chem.  Abst.  53: 20562.
                                                     f
Row, .F.J.C, and M.H. Salaman.   1955.  Further studies  on incomplete carcin-
ogenesis: Triethylene  melanine   (TEM),  1,2-benzanthracene,  and g-propiolac-
tone as initiators of skin tumor formation in  the mouse.  Brit. Jour.  Cancer
(London).  9:  177.

Verschueren, K.    1977.  Handbook of Environmental Data  on  Organic Chemicals.
Van Nostrand Reinhold Company, New  York.

Wilson,  C., et  al.  (ed.)  1977.  Textbook of Organic  Medicinal  and Pharma-
ceutical Chemistry.   J.B.  Lippincott Co.,  Philadelphia,  Pa.

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                                      No. 141
         Pentachlorobenzene

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report'represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.
                            **jout

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                               PENTACHLOROBENZEIC
                                     Summary

      Oral feeding of pentachlorobenzene  to  pregnant rats has produced devel-
 opmental effects and  decreased body weights  in fetuses.  No  adverse repro-
 ductive or developmental effects were  seen  in mice following maternal admin-
 istration of the compound orally.
      There is no  information available  on the mutagenic effects of penta-
 chlorobenzene.
         •
      A single study  has alluded to carcinogenic  effects  of pentachloroben-
 zene in mice and lack of carcinogenic effects  in dogs  and rats.   The details
 of this study were  not available for evaluation.
      Reported 96-hour  IC50  values for  the  bluegill,  mysid  shrimp,  and
 sheepshead minnow  range from 250 to 830  ,ug/l.  Oaphnia  is considerably less
 sensitive.   Studies with  algae, with  96-hour EC_Q  values  based on  chloro-
 phyll a_  concentration, have  reported values ranging  from 2,000  to  7,000
jug/1.   The steady-state bioconcentration  factor for the bluegill  is  1,800.

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 I.    INTRODUCTION
      Pentachlorobenzene,  CAS registry  number  608-93-5, is a colorless  crys-
 talline  solid with a pleasant  aroma.   It is produced  mainly as a  byproduct
 of  other chlorobenzenes and has  the  following physical and chemical  proper-
 ties  (Windnolz,  1976; We.ast, 1972; Hawley,  1971):
                   Formula:                 C6HC15
                   Molecular. Weight:        250.34
                   Melting Point:           86°c
                   Boiling Point:           277°c
                   Density:  .               1.S34216-5
                   Solubility:              Soluble  in carbon disulfide,
                                            chloroform,  and hot alcohol,
                                            insoluble in water
     Pentachlorobenzene is used primarily as a precursor in the synthesis of
 the fungicide pentachloronitrobenzene,  and  as a flame retardant.
 II.  EXPOSURE
     A.   Water
          Burlingame  (1977)  has identified  pentachlorobenzene  in the efflu-
 ent  from a  wastewater  treatment  plant  in  southern  California.   Access to
water  can occur  by  industrial discharge or  from  the degradation  of other
organochlorine compounds.
     8.   Food
          Pentachlorobenzene has  been  detected  in plants  (Balba  and Sana,
 1974;  Kohli,  et  al.  1976a)  and in animal  fat  (Stijve, 1971; Saha  and  Bur-
 rage,  1976;  Greve,  1973), and  was shown to arise  from the  metabolic break-
down of  lindane  or other  organochlorine compounds.   The U.S. EPA (1979) has
estimated  the  weighted average bioconcentration factor  for  pentachloroben-
 zene to  be  7,800 for the edible  portions of fish  and  shellfish  consumed by
Americans.  This  estimate is based on  steady-state bioconcentration studies
in bluegills.

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      C.    Inhalation
           The primary site for inhalation exposure could  be the workplace in
 industries utilizing or producing pentachlorobenzene.
      0.    Oential
           Pertinent data could not be located in the available literature.
 III.  PHARMACOKINETICS
     •A.    Absorption
           From studies with rabbits  it would appear  that pentachlorobenzene
                                                                     «
 is  very  poorly absorbed from the gastrointestinal tract (Parke  and  Williams,
 1960).
      3.    Distribution
           The  distribution of  pentachlorobenzene favors retention in the  fat
 (Parke  and Williams,  1560).   Khera  and  villeneuve (1575)  have found wide-
 spread  tissue distribution of  the  compound  following  oral  administration  to
 pregnant rats  and accumulation  in  fetal tissues.-
      C.    Metabolism
           There  appear to  be  some qualitative  and quantitative differences
 between  species in  the metabolism  of pentachlorobenzene.   In  the  rat  and
 rabbit, pentachlorobenzene was shown to be metabolized  to a variety of iso-
 mers  of  tetrachlorophenol, with  the  amount  of  unchanged  pentachlorobenzene
 excreted in the  urine of the rabbit being one percent (Kohli, et al. 1576b),
 and in the rat being nine percent (Koss and  Koransky, 1577).   Kohli and co-
workers  (1576b)  suggest  that  the  dechlorination hydroxylation  step  to the
tetrachlorophenol derivative proceeds through an arene oxide intermediate.
     0.   Excretion
           In rats and rabbits  urinary excretion  of metabolites  or  unchanged
pentachlorobenzene predominated.  Rozman, et  al.  (1578)  found  the biological
half-life of pentachlorobenzene to be two to  three  months  in rhesus monkeys.
                                      4

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After  40 days, ten percent of  the total dose was  secreted  in the urine; of
this,  58 percent  was pentachlorophenol.   After the  same period,  about 40
percent  of  the dose  was  excreted in the feces, 99  percent as pentachloroben-
zene.  The  authors suggest that biliary excretion was occurring.
IV.  EFFECTS
     A.   Carcinogenicity
          There  is  ane  report,  which  could not  be  critically  evaluated,
which  alludes  to pentachlorobenzene  being carcinogenic  in  mice but  not in
rats or  dogs (Preussman, 1975).
     9.   Mutagenicity
          Pertinent data could not be located in the available literature.
     C.   Teratogenicity
          Rats receiving 50,  100,  and  200  mg/kg  pentachlorobenzene on days 6
to 15 of gestation had pups with  increased suprauni ribs at  all doses (Khera
and Villeneuve,  1975).   The high  dose  also produced sternal  defects consist-
ing  of  unossified  or nonaligned sternabrae  with  cartilagenous  precursors
present.  The authors did not consider these defects to be teratogenic.
     0.   Other Reproductive Effects
          Oral  administration  of pentachlorobenzene  (50 or  100 mg/kg)  to
pregnant mice  on days 6 to  15 of gestation  produced no  teratogenic  or ad-
verse reproductive effects (Courtney, et al. 1977).
     E.   Chronic Toxicity
          Pertinent data could not be located in-the available literature.
V.   AQUATIC TOXICITY
     A.   Acute
                                                                     »
          The  U.S.  EPA  (1978)  reported 96-hour  LC5Q  values  for the blue-
gill  (Lepomis  macrochirus) exposed  to  pentachlorobenzene to  be  250 ug/1.
                                       y
                                         ^
                                           *

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The  48-hour  EC5Q value reported  for  Daphnia magna is 5,280 ug/1  (U.S.  EPA,
1978).  For  the  saltwater species, sheepshead minnow  (Cyprinodon  variegatus)
and  mysid shrimp  (Mysidoosis  bahia),. the  determined  96-hour  LC5Q  values
are 830 and 160/ug/l, respectively.
     9.   Chronic
          Pertinent data could not be located in the available  literature.
     C.   Plant Effects
          The  reported  96-hour  £C5Q vaj.ue  f0r  selenastrum  caoricornatum
based on  chlorophyll a_ concentration  is  6,780  jjg/1  (U.S.  EPA,  1978).-  For
the  marine  alga Skeletonema  costatum,  a  96-hour EC5Q  value on  the  same
basis is 1,980 ug/1 (U.S. EPA, 1978).
     0.   Residue
          After  a  28-day  exposure, the steady-state  bioconcentration  factor
for the bluegill for pentachlcrocenzene  is  1,800.  The half-life is greater
than seven days (U.S. EPA,  1973).
VI.  EXISTING GUIDELINES AND STANDARDS
     A.   Human
          The U.S. EPA (1979)  has drafted  a criterion of  O.S^ug/l for the
protection of human health.
     8.   Aquatic
          No  criteria  have  been  developed  or  proposed  to protect aquatic
organisms from pentachlorobenzsne  toxicity due to  the  lack of pertinent data.
                                    ift-7

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                                   REFERENCES


 Balba,  M.H.  and  J.G.  Saha.   1974.   Metabolism  of  l_indane~14c  by  wheat
 plants  grown  from treated seed.   Environ.  Let.   7:  181.

 Burlingame,  A.L.  1977.   Assessment  of the trace organic molecular  composi-
 tion of  industrial  and   municipal   wastewater  effluents  by  capillary  gas
 chromatography/real  time  high resolution  mass  spectrometry:  a  preliminary
 report.  Ecotoxicol.  Environ.  Saf_   1:  111.

 Courtney,  K.D.,  et al.   1977.   Teratology  study of  pentachlorobenzene  in
 mice: no  teratogenic effect at 50 or 100 mg/kg/day from day  6 to day  15  of
 gestation.  IRCS Med. Sci.   5: 587.

 Greve,  P.A.   1973.   Pentachlorobenzene as  a  contaminant  of  animal  feed.
 Meded.  Fac. lanbouwwet  Rijksuniv  Gent.   38: 775.

 Hawley, G.G.  (ed.)  1971.  The Condensed  Chemical  Dictionary.  8th ed.,  Van
 Nostrand Reinhold Co.,  New York.

 Khera,  K.S.  and  O.C.  Villeneuve.  1975.  Teratogenicity  studies on haloge-
 nated benzenes  (pentachloro-,  pentachloronitro-,  and  hexabromo-)  in  rats.
 Toxicology.  5:  117.

 Kohli,  J.,  et al.   1976a.   Balance  of  conversion  of  carbon-14 labeled lin-
 danes in lettuce  in hydroponic culture.  Pestic. Biochem. Physiol.  6: 91.

 Kohli,  J., et  al.  1976b.   The metabolism of higher chlorinated benzene iso-
 mers.  Can Jour.  Biochem.  54: 203.

 Koss, G.  and  W.  Koransky.  1977.  Pentachlorophenol in  different species of
 vertebrates  after administration of  hexachlorobenzene  and  pentachloroben-
 zene.  Pentachlorophenol, K.R. Rao, (ed.), Plenum Press,  New York.  p. 131.

 Parke,  D.V.  and  R.T.  Williams.   1960.   Studies  in  detoxification  LXXXI.
 Metabolism  of halobenzenes:  (a)  Penta- and  hexachlorobenzene:  (b)  Further
 observations of 1,3,5-trichlorobenzene.  Biochem. Jour.  74: 1.

 Preussman, R.   1975.   Chemical carcinogens in the  human environment.  Hand.
 Allg. Pathol.  6: 421.

 Rozman,  K., et al.  1978.  Metabolism and body  distribution  of pentachloro-
 benzene after  single oral  dose in rhesus monkeys.   Toxicol.  Appl. Pharmacol.
 45: 283.

 Saha, J.G. and R.H. Burrage.   1976.   Residues  of lindane and its metabolites
 in eggs, chicks and body tissues  of hen  pheasants after  ingestion of Lindane
 carbon-14 via  treated wheat seed or  gelatin capsules.  Jour.  Environ.  Sci.
Health Bull.  67.
                                                                        »

 Stijve,  T.   1971.  Determination and occurrence of hexachlorobenzene resi- .
 dues.  Mitt. Geb. Lebenmittelunters.  Hyg.  62:  406.

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U.S. EPA.   1978.  In-depth  studies on  health  and environmental  impacts of
selected water pollutants.   Contract No.  68-01-4646,  U.S.   Environ.  Prot.
Agency.

U.S.  EPfl.    1579.   Chlorinated  Benzenes:  Ambient  Water Quality  Criteria.
(Draft)

Weast, B.C.   1971.   Handbook of Chemistry  and Physics.   53rd ed., Chemical
Rubber Company, Cleveland, Ohio.

windnolz, M.  (ad.)   1976.  The Merck  Index.   9th ed.,  Merck  and  Co.,  Inc.,
Pahway, New Jersey.

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                                      No.  142
      Pentachloronitrobenzene

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980
            -7 II) 7f'

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                          DISCLAIMER
     This report represents  a survey of the potential  health
and environmental hazards from exposure to  the subject  chemi-
cal.  The information contained in the report is  drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources,  this  short profile
may not reflect  all available  information including all  the
adverse health  and  environmental impacts  presented  by  the
subject chemical.   This document  has undergone scrutiny  to
ensure its technical accuracy.
                          -^ ( "T '
                           I \y ,)T

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                       SPECIAL NOTATION










U.S. EPA's Carcinogen Assessment Group (GAG) has evaluated



pentachloronitrobenzene and has found sufficient evidence



to indicate that this compound is carcinogenic.
                               - 3

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                                DISCLAIMER
     The  mention of  company  trade names or  products  does not  constitute
endorsement by the U.S. EPA or  the Federal government.
                                 '-77/T-

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                           PENTACHLORONITROBENZENE
                                   Summary

     Increased'incidence  of  hepatoma formation was  reported in hybrid mice
treated with  pentachlorobenzene  (PCNB).   PCNB was found  to be mutagenic  in
the hcr-strain of Escherichia coli ochre,  but not  in  another §_._ coli strain.
     PCNB containing  a number of contaminants produced  renal agenesis and
cleft palate in C57B1/6 mice, cleft palate in  CD-I mice,  but was not terato-
genic in  CD  rats.  Purified PCNB (less  than  20  ppnv. hexachlorobenzene) re-
sulted  in  fewer cleft palates  in the fetuses.   No  significant teratogenic
effects in rats were  detected  at dosages as high as 1,563 ppm.  In  a  three
generation study using doses as high as 500  ppm,  PCNB  had no significant ef-
fects on the reproduction of rats.
     Acute toxicity  data for  fish were:   a 96-hour LCCg in bluegill from
0.29 to 0.38 ppm and a 96-hour LC5Q Of 0.31 pprn in rainbow trout.

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I.   INTRODUCTION
     This profile  is  based on the Initial Scientific  Review of Pentachloro-
nitrobenzene,  PCNB,   plus  relevant  scientific  research articles  published
subsequent to that document (U.S. EPA, 1976).
     Pentachloronitrobenzsne (molecular weight, 255.34)  is  a pale yellow-to-
white  solid,  depending  on purity,  that  melts between  142° and  146°C,  has
a  boiling point of  328°C  at  760 mm Hg, and a density  of 1.718  g/crn3 at
25°C.   Reported vapor  pressure values  for  PCNB are:   1.16  x  10~5  mm Hg
at  10°C,  5.0  x  lO-5  mm  Hg  at 20°C,  and 11.3  x  1Q-5  mm Hg  at 25°C
(U.S.  EPA,  1976).   PCNB has.  a relative vapor  density  (air  =  1)  of 10.2
(verschueren,  1977).   Water solubility of  PCNB  is  0.44 mg/1 at 20°C  and 2
mg PCNB will dissolve  in  one  liter ethanol at  25°C.   PCNB is  freely sol-
uble  in carbon  disulfide,  benzene,   chloroform,  ketones,  and  aromatic  and
chlorinated hydrocarbons, and slightly soluble in alkanols (U.S. EPA, 1976).
     PNC3 is primarily  registered  as a soil  fungicide for a wide  variety of
crops  and  is  also  used  as  a  seed-treatment fungicide.   It  is  effective
against  bunt  of wheat,  Botrytis,  Rhizoctonia,  and  Sclerotinia spp.   There
are no  current nonagricultural  uses  of PCNB  (U.S. EPA,  1976).  PGNB is manu-
factured  domestically  under  the  trade  name TerraclorS) with  an  estimated
annual  production  in  1971  of 3  million  pounds (U.S. EPA,  1972).   According
to the Olin Corporation (1974), 60  to 70 percent of the PCNB  produced will
be used in the United States.   The United States  has imported from 20,000 to
132,000  Ibs.  between 1966 and  1969  (U.S. EPA,  1976)..   PCNB manufactured in
Europe  is marketed under  the  common  name  Quintozene (Dejonckheere, et  al.

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1976).  It may be  worth noting that commercial  PCNB fungicides contain  im-
purities such as hexachlorobenzene, pentachlorobenzene and  tetrachloronitro-
benzene, which may  be more hazardous than  PCNB  itself  (Dunn,  et al. 1978;
Simon, et al. 1979).
     No data  are  available for  the disassociation'  of PCNB in aqueous sys-
tems.  Crosby and  Hamadmad (1971) studied  the photoreduction of PCNB.   The
compound remained unchanged in sunlight,  probably excluding photolysis as a
major  route  of  environmental  degradation.  At  temperatures  above  328°C,
some decomposition of PCNB has been noted  (U.S. EPA,  -1976).
     PCNB can be biodegraded  by  pure  cultures of actinomycetes and  filamen-
tous fungi during their active growth  phase  (Chacko,  et al.  1966).
II.  EXPOSURE
     PCNB  is  prepared by  either chlorination or  nitration reactions.   The
reaction temperature  for  the  chlorination process is  60 to 70°C.   Although
this reaction is well below the boiling point of PCNB, atmospheric emissions
are possible because of PCNB's relatively high vapor pressure.  Furthermore,
there exists  a potential  for  environmental  release via wastewater effluents
at  the  manufacturing sites.   No  monitoring data are  available for  ambient
air or water levels of the compound.  The major source of environmental con-
tamination is through its application as a  fungicide.  In the  United States,
PCNB is used primarily on cotton and peanut crops.   Geographic  use distribu-
tion is mainly concentrated west  of the Mississippi River  (U.S. EPA,  1976).
Carey,'et al. (1979) in their study of pesticide residues in the soil  detec-
ted PCNB in only three of  the 1,483 sample  sites.  The .detected residue con-
centration was  from 0.22  to  2.61 ppm.    It  should  be noted,  however, that
their study was primarily confined to  the  eastern  United  States.

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     Routes  of human  exposure  to  PCN8  include  water,  air,  contaminated
foods, and fish.   Casanova  and Oubroca  (1973)  studied  the residues of  PCNB
found in  lettuce  grown in soil  treated with the fungicide.   Residue  values
were 0.73  ppm '(15 kg PCNB/ha)  and 1.56 ppm  (45  kg PCN8/ha).  Gcursaud,  et
al. (1972) detected PGNB.contamination in endive roots.  Since the main ob-
jective  of their  study  was  the  uptake of  hexachlorobenzene,  actual  PCNB
concentrations  were  not  noted.    However,  in  a  subsequent   experiment
Goursaud,'-et  al.   (1972)  fed cows endive roots  containing  2.16  ppm  PCNB.
PCNB residues  found  in the cows'  milk were negligible.  Bioaccumulation  of
PCNB in  White Leghorn cockerels  (Dunn, et  al.  1978) was  also  found to  be
negligible   (accumulation  ratio   0.001   =  tissue   concentration/dietary
concentration).   Broiler  chickens  (Reed,  et al.  1977)  did  not  accumulate
PGNB  or  its  metabolites  to  any  appreciable  extent   (0.002   ppm).    NO
additional information on the levels of PCNB in foods is available.
     Bioaccumulation data on PCNB  were  not found in  the literature for aqua-
tic organisms,  Ko and  Lockwood (1963)  resorted  that the mycelium of  fungi
had accumulated a  concentration  of PCNB seven times  that of  the surrounding
soil.
III. PHARMACOKINETICS
     A.   Absorption
          Absorption data on PCNB  were  restricted to  oral administration in-
volving  three  test species.   Betts, et  al.  (1955)  reported that  60  percent
of  the oral  dosage was not absorbed from  the gastrointestinal tract  in  rab-
bits.  Two subsequent studies, however,  report  that PCNB  is readily absorbed
from the gastrointestinal tract and/or  metabolized by  gut flora  to  another
                                                                         »
compound and  then almost fully absorbed.   Kogel,  et al. (1979).  found  that
PGNB  was readily  and  almost completely  absorbed  from the gastrointestinal
                                     i  .- n ]  -"
                                   •*7 b I <*•

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tract of  Rhesus  monkeys.   After  a  single dose of  2 mg/kg  given  in methyl
cellulose suspension,  only  7.4 percent  of the administered  amount  was ex-
creted as unmetabolized FCNB in the feces.  When  91  mg/kg PCNB was given in
sesame oil,  only 4.3 percent of the dose was excreted unmetabolized.  Uptake
occurs mainly  by the  portal  venous route, with  little  involvement  of the
lymphatic system,  bringing  the  absorbed PCNB  directly  to  the  liver where
biotransformation can  begin.  Studies of  Comet Red  and White Leghorn chick-
ens yielded  similar  results.   Chickens  fed 300 ppm  PCNB in laying mash for
sixteen weeks excreted only  1.1 ppm  PCNB (Simon, et ai. 1979).
     B.   Distribution
          Several studies have been conducted  on  the distribution and stor-
age of ingested  PCNB.   Due to  rapid  metabolism  and elimination,  this com-
pound shows very little accumulation  in body  tissues.   Betts, et al. (1955)
used rabbits and Borzelleca,  et al.  (1971)  employed beagles  and  rats.   In
neither experiment was PCNB detected in  liver,  kidney,  muscle,  or adipose
tissue.   Othe'r  studies have  indicated  very low  concentrations of  PCNB in
various tissues.  Simon, et al.  (1979)  found  PCNB at concentrations of 0.85
ppm in fat and 0.005  ppm in  egg whites of chickens fed 300 ppm PCNB for six-
teen weeks.   Other tissues examined contained  no  detectable levels of PCNB.
Dunn, et  al.  (1978)  found  the highest tissue  residues  of  PCNB  in adipose
tissue (1.14 and 1.87 ppm)  and the gizzard (1.60 and  0.84 ppm) in chickens
given 100 ppm and 1,000 ppm PCNB in feed, respectively.  Leg and breast mus-
cles and  heart,  kidney,  and liver contained very low  (0.16 to 0.07 ppm) or
trace amounts of PCNB.
          Concentrations of PCNB  in various organs  of  Rhesus monkeys after
chronic feeding of  2  ppm PCNB in the daily diet were (in ppm):  blood, O.D7;
muscle, 0.01; brain, 0.03; liver, 0.19;  kidney,  0.14;  adrenal cortex, 0.08;


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thymus,  0.20;  lymph nodes  (large  intestine), 0.12;  bone marrow, 0.13;  and
omental  fat,  0.21  (Mueller,  et  al.  1978).   Kogel,  et al. (1979) found  the
highest concentration of PCNB and/or its metabolites occurring  in bile (7.73
> 0.2  ppm  in males and 3.72 +_ 0.05  in females)  after feeding  of 2  ppm PCNB
for 70 days.
     C.   Metabolism
          PCNB metabolism has been studied in rats,  dogs,  cows, and  rabbits.
Pentachloroaniline and methyl pentachlorophenyl sulfide are the major  metab-
olites.  Tissue retention of  these compounds is found primarily  in  body  fat
with minimal concentrations  found  in muscle (U.S.  EPA,  1976).  Two  major
                                   •
pathways for the. biotransformation of PCNB  in Rhesus monkeys are:   1)  the
reduction of the hitro-moeity to the corresponding aniline, and 2) the clea-
vage of the C-N  bond,  presumably  via conjugation with  sulfur-containing
amino acids  (Kogel, et al.  1979).
     0.   Excretion
          PCNB  and its  metabolites  ar.e excreted mainly  in  the urine  and
feces.   Mueller, et al.  (1978)  reported that Rhesus monkeys  excreted  almost
80  percent  of  the  ingested  PCNB  within   5 days;  of the  excreted  radio-
activity, 91.2 percent was  in  the form of metabolites.
IV.   EFFECTS
     A.   Carcinogenicity
          Very  little  information  on  possible  carcinogenic effects of PCNB
was found  in the  available literature.   Courtney,  et al.  (1976)  cite  one
study  which found  PCNB  to be  carcinogenic  in a hybrid  mouse with an  in-
creased incidence of hepatoma  formation.  Levels of exposure were not given.

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     8.   Mutagenicity
          PCNB was  found to be  mutagenic in the  hcr-strain of Escherichia
coli B/r ochre, but not in another £_._ coli strain.  In the host-mediated as-
say  in  mice,  no   significant  increase  in  mutation  rates  in  Salmonella
typhimurium and Serratia morcescens was  observed  after subcutaneous injec-
tion of  PCNB.   The compound also gave  negative  results in spot tests (U.S.
EPA, 1976).
     C.   Teratogenicity and Other Reproductive Effects
          PCNB was administered to pregnant rats by intubation on days 6 and
15 of gestation at dosages from 100 to  1,563  ppm.  Fetuses were examined for
gross malformations.  No significant effects on the number of corpora lutae,
the position and  numbers of dead or  resorbed fetuses,  or the fetal weights
and sex ratios were observed at any dose  level.   No significant skeletal or
soft tissue anomalies were reported  in the fetuses  (U.S. EPA, 1976).
          A three-generation study with groups of  rats  fed diets containing
0, 5,  50 or 500.ppm  (Olin  technical  PCNB)  showed  no significant effects on
fertility, gestation,  viability,  lactation,   rats  born  per  litter,  or rats
weaned per litter  or their average weaning weights  (U.S. EPA, 1976).
          PCNB containing a  number of contaminants,  however, produced renal
agenesis and cleft palate in C56B1/6 mice and cleft palate in CD-I mice, but
was not teratogenic in CD rats.  Purified PCNB (less than 20 ppm hexachloro-
benzene) resulted  in few cleft palates in  fetuses (Courtney,  et al. 1976).
     D.   Chronic  Toxicity
          PCNB does not  appear  to be  chronically  toxic,when administered in
feeding  studies.   Rhesus monkeys  given  2 ppm or  91 ppm  PCNB  in  their diet
for 70 days were  monitored for clinical chemistry and hematology parameters
throughout the study.   These parameters  remained  unchanged,  indicating that
                                 l/JL-lf

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organ function and hematopoiesis was not affected  by  PCNB or its metabolites
(Kogel, et al. 1979).
          White Leghorn chickens  fed  PCNB  at concentrations up  to  1,000 ppm
for the first 8 weeks of life did not develop tissue  lesions and hens fed up
to 1,000 ppm  for 35  weeks- failed  to develop histopathological  changes (Dunn,
et al. 1978).
          Kogel, et  al. (1979)  cite--studies which -report  that  the  toxic ef-
fect of  Terraclor^  in rats and dogs  is limited to liver  enlargement due to
              .                                  •*   s
hepatocellular hypertrophy.   Also,  cats had increased methemoglobin levels
after moderate  and  high doses  of Terraclor^ and  dogs  fed a very  high  dose
(5,000 ppm) of PCNB  of undetermined purity  for  two years were  found  to  have
reduced hematopoiesis.  These effects,  however, may  be due to  the presence
of hexachlorobenzene as a contaminant  (Kogel, et al.  1979).
     E.   Acute Toxicity
          Cholakis,   under  contract with  the U.S.  EPA, administered single
doses of  pentachloronitrobenzene  by  gavage  to  several species  of  microtine
rodents  (voles)  (U.S. EPA, 1978).  The acute  oral LD5Q  values in male and
female M^ montanus  were 4,194  mg/kg  and 5,717 mg/kg,  respectively.  In M^
ochrogaster,  M^  canicaudus,  and M^ pennsylvanicus, values  were greater  than
5,000 mg/kg  for  both sexes.   Toxicologic signs observed were  some  piloerec-
tion,  loss  of  righting  reflex and  lachrymation.   Most  signs  disappeared
after 24 hours.  Most deaths occurred  within two to six days of dosing.
V.   AQUATIC TOXICITY                               .:
     A.   Acute Toxicity
          In  static,  acute toxicity  bioassays  using  various   PCNB  formula-
                                                                          »
tions, bluegill  (Lspomis  macrochirus)  had  96-hour median  lethal  concentra-
tion  (LC5Q)   values  ranging   from  0.29 to  0.38 ppm.   Rainbow  trout (Salmo
                                    -i / P /.
                                  "7" u o

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gairdneri) had a 96-hour LC5Q value of 0.31 ppm  (U.S. EPA, 1976).
     B.   Chronic Toxicity,  Plant Effects  and Residues
          Pertinent data could not be  located in the available literature.
VI.  EXISTING GUIDELINES AND STANDARDS
     No guidelines or standards were located in the available literature for
humans or aquatic life.


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                                  REFERENCES
Betts, J.J., et  al.   1955.   The metabolism of PCN8  and 2,3,4,6-tetrachloro-
nitrobenzene and the formulation of  meriapturic acids in  rabbit.   Biochem.
Jour.  61: 611.

Borzelleca, J.F., et  al.   1971.   Toxicologic and metabolic studies  on PCNB.
Toxicol. Appl.  Pharmacol..  18: 522.

Carey, A.E., et  al.   1979.  Pesticide residue levels in soils and crops from
37 states, 1972.  Pest. Monit. Jour.  12: 209.

Casanova, M. and 3. Oubroca.   1973.   Etude  des residues de divers fongicides
utilises dans  le traitement des  cultures de laitures en serre.   Ann.  Phyto-
pathoi.  5: 65.
                                                     *.
Chacko, C.I.,  et al.   1966.   Chlorinated hydrocarbon pesticides:  degradation
by mirobes.  Science  154: 393.

Courtney, K.D.,  et al.  1976.  The  effects  of pentachloronitrobenzene, hexa-
chlorobenzene,  and related  compounds on  fetal  development.  Toxicol.  Appl.
Pharmacol.  35: 239.

Crosby, O.G. and N.  Hamadmad.  1971.  The  photoreduction  of  pentachloroben-
zenes.  Jour. Agr.  Food Chem.   19: 1171.

Oejonckheere,  w.f  et al.   1976.  Residues  of  Quinotozene.   Pest.  Monit.
Jour.  10: 68.    .                                      ,

Dunn, J.S., et al.   1978.  The accumulation and  elimination  of tissue resi-
dues after feeding PCNB to White Leghorn  cockerels.   Poultry Sci.   57:  1533.

Goursaud, J.,  et al.   1972.   Sur la  pollution du lait  par les residues HC3.
Industries Alimientoires et Agr.   89:  31.

Ko, W.H. and J.L. Lockwood.   1968.  Accumulation  and concentration of  chlor-
inated hydrocarbon pesticides  by  microorganisms  in  soil.  Can. Jour.  Micro-
biol.  14: 1075.

Kogel, W.,  et al.   1979.  Biotransformation of PCNB  - 14c  in  Rhesus  mon-
keys after single and chronic oral administration. Chemosphere  8: 97.

Mueller,  W.F.,  et al.    1978.   Comparative metabolism  of HC3 and  PCNB  in
plants, rats', and Rhesus monkeys.  Ecotoxicol. Environ.  Safety.  2: 437.

Olin Corporation.  1974.   Agricultural Products  Division,  Little Rock,  Ark.
Personal  communication  to Scon.  Branch, Criteria and Evaluation Division,
Office of Pesticide Programs,  U.S. EPA.

Reed,  E.L.,  et  al.   1977.   Tissue  residues from   feeding PCNB  to broiler
chickens.  Toxicol. Appl.  Pharmacol.  42: 433.

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 Simon,  G.S., et al.   1979.   Distribution  and clearance of pentachloronitro-
 benzene in  chickens.   Toxicol. Appl. Pharmacol.  50: 401.

 U.S.  EPA.  1972.  The pollution  potential in pesticide manufacturing.  Mid-
 west  Research  Institute.  EPA Rep. No. OWP-TS-00-72-04.  NTIS PB-213 782.

 U.S.  EPA.   1976.   Initial  Scientific Review  of PCNB.  Office  of Pesticide
 Programs, Washington,  D.C.  EPA-540/1-75-016.

 U.S.  EPA.   1978.   Study  of  the  chemical  and  behavioral  toxicology  of
 substitute   chemicals   in  microtine  rodents.   EPA-600/3-78-082.   Midwest
 Research Institute, Kansas City,  MO.

•Verschueren,  K.   1977.  Handbook of Environmental Data on Organic Chemicals.
 Van Nostrand Reinhold  Co., New York.

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                                       No.  143
         Pentachlorophenol


  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY

       WASHINGTON, D.C..  20460


           APRIL 30, 1980
            '{ / 0 ^
            ') V I  v

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                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

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                        PENTACHLOROPHENOL



                             SUMMARY



      Pentachlotophenol has  shown  no evidence of carcinogenicity.



Evidence  for mutagenicity   is  equivocal.    Pentachlorophenoi  is



teratogenic  in experimental  animals at levels which produce mater-



nal or  fetal toxicity.   Adverse health  effects  have been minimal



in workers  chronically exposed  to  pentachlorophenol.   Relatively
                                          k


high levels  of  continous exposure  produce  muscle weakness, head-



ache,  anorexia,  abdominal-  pain,  weight  loss,  and  irritation  of



skin, eyes,  and respiratory  tract.   Pentachlorophenol  is a strong



uncoupler of oxidative phosphorylation.



      Pentachlorophenol has  been demonstrated to be acutely toxic



to freshwater salmonids  at  levels as  low as  37  ug/1.   Comparable



levels  of  toxicity  were' observed  for  marine  fish. .  Freshwater



plants were  also  highly susceptible to  the action  of  this chemi-



cal- with effective concentrations_as low as 7.5 ug/1.

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                        PENTACHLOROPHENOL

I.     INTRODUCTION

      This profile  is  based on the Ambient Water Quality Criteria

Document for Pentachlorophenol  (U.S. EPA, 1979).

      Pentachlorophenol. (PCP;  CgCl5OH;  molecular  weight  266.35)

has  the  following  physical  and   chemical  properties   (Stecher,

1966;_-Natl.. Fire  .Prot.  Assoc., 1973;  Sax, 1975;  Spector, 1956;

Weast, 1975-76):

           Melting 'point Range      190 - 191° C
           Boiling Point Range      309 - 310  (decomposes)
          ' Vapor Pressure           0.12 mm Hg at 100  C
           Solubility  '             Water: 14 mg/1 at 20° C


           Commercial  preparations  of  pentachlorophenol  contain

"caustic  insolubles"  or   "nonphenolic.  neutral   impurities"  such

as  octachlorodibenzofurans  and tetra-,  penta-,   hexa-,   hepta-,

and  octachldrodibenzo-p-dioxins (Johnson,  et  al.  1973;  Schwetz,

et al.  1974).   In addition, commercial pentachlorophenol contains

three to  ten percent  tetrachlorophenol  (Goldstein, et  al. 1977;

Schwetz, et al. 1978).

      Pentachlorophenol  is  a  commercially  produced  bactericide,

fungicide,  and  slimicide  used primarily   for  the  preservation

of wood,  wood products,  and other  materials.    As a chlorinated

hydrocarbon,  PCP  is also  used as  a herbicide,  insecticide,  and

molluscicide  (U.S. EPA, 1979).

      Pentachlorophenol and its sodium  salt are  widely dissemi-

nated  in  the  environment  (U.S.  EPA,  1979).    Pentachlorophenol

undergoes photochemical degradation  in  solution in  the presence

of  sunlight   (Mitchell,  1961;  Hanadmad,  1967;  Wong  and Crosby,

1977) and is  reported  to  persist in warm moist soils  for a period


                                X
                              -, i J- Q 1 -
                             ^ U I U •

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of 12 months (Harvey and Crafts, 1952).  in laboratory experiments,
some microorganisms have  been reported to metabolize  pentachloro-
phenol  and  its  sodium  salt  (Watanabe,   1973;  Suzuki  and  Nose,
1971; Cserjesi, 1967; Reiner, et al.  1977).
II.  EXPOSURE
      Residues  of  pentachlorophenol  have  been  found  in  food,
water and  human tissues.   Pentachlorophenol  levels  of 0,06  ug/1
in finished  drinking  water prepared from untreated water  contain-
ing  0.17  ug/1  have been  reported  (Buhler, et al.  1973).    Penta-
chlorophenol has  been detected in 13  of 240  food composites  at
levels  of  0.01 to  0.04  mg/kg  (Johnson and  Manske,   1977).   The
calculated  daily  dietary  exposure  is one  to  six pg/person/day
(Duggan and Corneliusen, 1972).
      The  U.S. EPA  (1979)   has  estimated  the weighted   average
bioconcentration factor of pentachlorophenol  at  58 for the  edible
portion of  fish  and shellfish  consumed by Americans.  This  esti-
mate  is based on  measured steady-state  bioconcentration  studies
in goldfish  (Carassius  auratus) ,  bluegill (Lapomis macrochirus),
eastern  oyster  (Crassostrea  virginica) ,   and  sheepshead  minnow
(Cyprinodon variegatus).
      Inhalation and dermal  exposure  data for the general  popula-
tion are  not available  (U.S. EPA,  1979).   These  routes  of  expo-
sure are more likely to occur occupationally. :
      Total  body  exposures,  based on reported' urine  levels  of
pentachlorophenol,  appear  to  be in the  range of 10-17 /ig/person/
                                                               •
day  for the general  population  and  1500-4400  pg/person/day for
occupational exposures  (U.S.  EPA, 1979).    These values  may  be

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due  not  only  to direct  exposure  to pentachlorophenol,  but  also



to exposure  to hexachlorobenzene  (pesticide,  fungicide)  and  lin-



dane  (pesticide) , which  are degraded in part to pentachlorophenol



(Yang, et  al.'  1975;  Lui  and Sweeney,  1975;  Mehendale,  et al.



1975;  Koss and  Koransky,  1978;  Karapally,  et al.  1973;  Engst,



et al. 1976) .



III. PHARMACOKINETICS.



      A.    Absorption



           The  half-life  for  absorption  in  humans  after   oral



ingestion  of  pentachlorophenol  is  1.3  +  0.4  hr.    In  humans,  a

                                        ,'


peak  plasma  concentration of  0.248  mg/1  was observed  four  hours



after ingestion of a 0.1 mg/kg dose  (Braun, et  al. 1978).  Absorp-



tion  in  rats  is  similar  to that  found  in humans  (Braun,  et al.



1977).



      Pentachlorophenol  is  readily  absorbed  through  the   skin



as  indicated  by  its lethality after dermal  exposure  (Deichmann,
   *


et'al.  1942; Armstrong, et al. 1969).



      B.    Distribution



           In  humans  (fatal  pentachlorophenol  intoxication) and



in rats  (non-lethal  exposure),  the  highest levels of pentachloro-



phenol are found  in liver,  kidney,  and  blood,  with  the  lowest



levels in  brain, spleen,  and fat  (Cretney,  1976;  Armstrong, et



al. 1969; Braun,  et al. 1977; Larsen, et al.  1975).



      C.    Metabolism



           In  four  male volunteers  ingesting 0.1  mg pentachloro-



phenol/kg,   approximately  74  percent of  the  dose  was  elimina'ted



in  the  urine  as pentachlorophenol  (PCP)  and  12  percent  as PCP



glucuronide;  four percent  was  eliminated  in feces as pentachloro-
                             * fu r?

-------
phenol  and  PCP glucuronide  (Braun,  et  al.  1973) .   Rats excrete
75  percent  of  administered  pentachlorophenol  as   the  unchanged
PCP,  16  percent  as  tetrachlorohydroquinone,   and  nine  percent
as PCP glucuroriide  (Braun, et al. 1977).  In another study  (Ahlborg,
1978) , trichloro-£-hydroquinone was found as an additional metabo-
lite of  pentachlorophenol in  rats.   Mice also  metabolize penta-
chlorophenol  to  tetrachlorohydroquinone  (Jakobson  and  Yllner,
1971).
      D.   Excretion
           In  humans  and in  experimental  animals,  the  primary
mode of excretion for pentachlorophenol is  in the  urine (Diechmann,
et al.  1942; Braun, et al. 1977, 1978;  Larsen, et  al.  1975; Jakobson
and Yllner,  1971) .
           In  humans,   the   plasma   pentachlorophenol  half-life
is  30.2  + 4.0 hours.   -The half-lives for elimination  of penta-
chlorophenol  and  PCP glucuronide  from urine  are 33.1  +  4.5 and
12^7 + 5.4 hours,  respectively  (Braun,  et  al.  1978) .  Elimination
of  pentachlorophenol by  the  rat is  similar   to elimination  by
humans (Braun, et al.  1977) .
           The  available literature  indicates  that pentachloro-
phenol  does  not accumulate  in  body   tissues  to any  significant
extent  (U.S.  EPA,  1979).    Long  term,  low  level tissue  binding
has not been adequately studied.
IV.  EFFECTS
      A.   Carcinogenicity
           Pentachlorophenol  has  not  shown  evidence  of carci'no-
genicity.   Pentachlorophenol did not  promote  papillomas  or  car-
cinomas  when  applied  repeatedly  to the  skin  at high concentra-
                            //J-7

-------
tions  after   initiation  with  dimethylbenzanthracene   (Boutwell


and  Bosch,  1959).   Mice  receiving  commercial pentachlorophenol


in  the diet  throughout  their  lifespans   (about  18  months)  did


not  have  a  significant incidence of  tumors  (Innes,  et al.  1969) .


Pentachlorophenol,  with low  levels  of  nonphenolic contaminants,


was non-carcinogenic when fed to rats for 22 to  24 months  (Schwetz,


et al. 1978).                          -  -


      B.    Mutagenicity
                      •                             *   A
           Pentachlorophenol  has been  shown  to be  mutagenic  in


a  few test  systems.  Recrystallized  pentachlorophenol  increased


the  frequency  of  mutations  and mitotic gene  conversion in  Sac-


charomyces cerevisiae  when used at  a level  (400 mg/1)  which re-


sulted in a 59  percent survival rate of  test organisms  (Fahrig,


et al.  1978).   Four of the  473 offspring of- female mice  injected


with  a  single  high  dose of  pure pentachlorophenol  during  gesta-


tion  were reported  to  have  changes in  hair,  coat  color  (spots)


of genetic significance  (Fahrig, et  al.  1978).


           No mutagenic activity was detected  in  male germ cells


of Drosophila  (Vogel and Chandler, 1974) , in the mouse host-mediated


assay, in in  vitro  spot tests  (Buselmaier,  et al. 1973),  or  in


histidine-required  mutants  of  Salmonella  typhimurium  (Anderson,


et al. 1972).


      C.    Teratogenicity


           Information  suggesting  pentachlorophenol  is  a  human


teratogen  was  not  encountered.   Pentachlorophenol of  both  com-
                                                               »

mercial  and purified   grades   produced  fetal  anomalies  in  rats


at  levels considered to be  toxic either  to the maternal rat  or
                                •3"

-------
to  the  fetus (Larsen, et  al.  1975; Schwetz,  et al. 1974?  1978).
Abnormalities  included subcutaneous  edema,  dilated ureters,  de-
layed  ossification of  the  skull,  skeletal  anomalies, dwarfism,
exencephaly, macropthalmia, and taillessness.
      D.   Other Reproductive Effects
           In  a study in  which  male  and  female   rats  were  fed
3  or  30  mg/kg pentachlorophenol  continuously  starting  62 days
before  mating,  no adverse  effects were  observed at  the  3  mg/kg
level.  At 30 mg/kg, the following  indices were decreased: maternal
body  weight;  percent   liveborn  pups;  7,  14,  21  day survival;   1,
7, 14, 21 day body weight-pups; 7, 14,  21 day  litter size.  Selected
abnormalities were also seen at this dose.(Schwetz,  et  al. 1978).
      E. Chronic Toxicity
           Adverse  health  effects  have  been  minimal  in workers
chronically exposed to pentachlorophenol (JClemmer, 1972; Takahashi,
et  al.   1976).    Increased  levels  of  serum  enzymes  SCOT,   SGPT,
   »
and- LDH,  and elevated  levels of  total  bilirubin  and  creatine
phosphokinase were noted,  but  all  levels were still within  normal
limits.    A  significantly  higher  prevalence  of gamma  mobility
C-reactive protein  (CRP)  was detected  in the  sera  of  chronically
exposed  workers.   CUP levels  are  often  elevated  in acute  states
of  various  inflammatory  disorders or  tissue  damage  (Takahashi,
et  al.  1976).   A  chronic health effect  which  has been associated
with  human  exposure  to certain types of  commercial  PCP is  chlor-
acne  (Saader and  Bauer,  1951;  Nomura,  1953).    Chloracne   could
have resulted from impurities in the pentachlorophenol; commercial
PCP  containing  high  levels of  chlorodioxins  produced chloracne
in  the   rabbit  ear  test,  while  pure  pentachlorophenol or  penta-
                             d^^^^^^Q: -
                             ~ / t> / 0

-------
chlorophenol  with  reduced  dioxin  content  did  not  (Johnson,  et

al. 1973).

           Chronic  intoxication  in  humans results from  relatively

high  levels  of  continuous  exposure.    Symptoms  include muscle

weakness,  headache,  anorexia,   abdominal pain,  and  weight   loss

in addition  to  skin,  eye, and respiratory  tract irritation  (U.S.

EPA, 1979) .

           Rats  fed  pentachlorophenol  containing  low  levels  of

nonphenolic  contaminants at  daily  levels  of 1  to 30  mg/kg  for

eight months  (Goldstein, et al.  1977)  and  22  to 24 months (Schwetz,

et  al.  1978)  had decreased  body weight  gains  at  dosage levels

of 30  and  10 mg/kg,  respectively.   In the  22 to 24 month study,

the 30 mg/kg  dose resulted in increased  serum enzyme  SGPT levels

and increased specific gravity of the urine.

      F.   Other Relevant Information

           Pentachlorophenol  is  a  strong uncoupler  of  oxidative

phdsphorylation  (Weinbach and Garbus, 1965; Mitsuda, et  al. 1963).

V.    AQUATIC TOXICITY

      A.   Acute Toxicity

           The  results  of  38  freshwater  flow-through  bioassays

reveal a  range  of  96-hour LC5Q  values of  from 63 ug/1 for  the

sockeye  salmon  (Oncorhynchus  nerka)   (Webb  and Brett,  1973)  to

340 ug/1  for  the fathead  minnow  (Pimephalea- promelas)  (Ruesink

and Smith,  1975).  In  19 static assays, LCcn values  ranged  from

37  ug/1  for  the coho  salmon  (0.  kisutch)   to  600  pg/1 for  the
                                                                #
fathead minnow.    Five species  of  salmonids  were  more  sensitive

than 4  other species  of  minnows or  centrachids.   Freshwater  in-

vertebrates displayed  LC5Cj  values ranging from  310  ug/1  to  1,400

-------
ug/1  for  the tubificid  worm (Tubifex  tubifex)  and were  affected
by  increasing  the  PH from  7.5  to  9.5.   The  acute  toxicity  of
pentachlorophenol  to  saltwater  fish  ranged  from  38  ug/1  in  a
96-hour  static  pinfish  (prolarvae)   (Lagodon  rhomboides)   assay
(Borthwick  and  Schimme,!, 1978)  to 442  ug/1  for juvenile  sheeps-
head  minnows  (Cyprinodon  variegatus)   (Parrish,   et   al.   1978).
For  three  marine invertebrate  species  tested,  LC5Q values  ranged
from  40 to 5,600 ug/1, with the  eastern oyster (Crassostrea  vir-
ginica) being the most sensitive  marine  invertebrate.
      B.   Chronic Toxicity
           Freshwater  chronic   studies  for fish or invertebrates
were  not  available.   A  life-cycle  chronic  test  of 151  days  in
the marine  sheepshead minnow produced  a chronic value of 64  ug/1•
(Parrish,   et  al. 1973).   Data  for  marine invertebrates  was  not
available  (U.S. EPA, 1979).
      C.   Plant Effects       '
           For freshwater plants,_ the lowest affective concentra-
tion  was  7.5  ug/1,  which  resulted   in  the  total  destruction  of
chlorophyll  in  the  alga  Chlorella  pyrenoidosa after  72  hours.
A drastic  decrease  in  cell  numbers of the marine alga Monochrysis
lutheri was observed after 12 days of exposure to 293 ug/1  (Woelke,
1965) ,  and 50  percent  inactivation  of  photosynthesis  was  seen
in  kelp (Macrocystis  pyrifera)   exposed  for  .4  days to  300  ug/1
(Clendenning and North, 1960).
      D.   Residues
                                                               r
          . Equilibrium  levels  of PCP   in  water  and   tissues  of
aquatic organisms are  attainable  within four  days;  and when  pre-
viously  exposed  marine  eastern  oysters  (Crassostrea  virginica)

-------
or  freshwater  bluegills  (Lepomis  macrochirus)  were  held in PCP-



free water, a rapid loss- of PC? from the organism occurred  (Schim-



mel, et  al.  1978;  Pruitt,  et al.  1977).  Bioconcentration  factors



in marine organisms ranged from 0.26 for the juvenile brown shrimp



(Penaeus  aztecus)  to  78  for the  eastern  oyster.   In freshwater



fish,  bioconcentration factors  of 1,000  for   the  whole  body  of



the goldfish  (Carassius auratus)  and of 13  for the muscle tissue



of  the bluegill  have  been reported  (Kobayashi  and  Akitake, 1975;



Pruitt, et al. 1977) .



VI.   EXISTING GUIDELINES AND STANDARDS



      A.    Human.



           The U.S.  EPA  (19.79)  draft  criterion  for  pentachioro-



phenol in ambient water is 680 ug/1.



           The maximum air concentration established by the Ameri-



can Industrial Hygiene Association (1970)   is  0.5 rag  pentachloro-



phenol or  0.5 mg  sodium pentachlorophenate/mj  for an 8-hour expo-



sur'e (TLV) .   The  code of Federal Regulations  21, part  121, para-



graph  12±:2556  allows up  to  50  ppm pentachlorophenol  in  treated



wood which will come in contact with food.



           A NOEL in drinking water of 0.021 mg pentachlorophenol/1



is  suggested  by  the  National Research  Council  (1977) ,  based  on



a NOEL of  3 mg/kg  in  90 day  and  8 month rat studies and  an uncer-



tainty factor of 1,000.



      B.    Aquatic



           The  draft  criterion   to  protect  marine  life  is  6.2



ug/1 as  a 24-hour  average,  not  to exceed  14  ug/1 at  any time.



The  draft criterion  to protect  marine life  is 3.7  ug/1  for  a



24-hour  average,  not to exceed  8.5 ug/1 at  any time  (U.S.  EPA,



1979) .
                          If 3-

-------
                               PENTACHLOROPHENOL

                                  REFERENCES

Ahlborg,  U.O.    1978.   Dechlorination  of pentachiorophenol  in  vivo  and  in
vitro,   pp.  115-130.   In:  K.R.  Rao  (ad.),  Pentachlorophenol:   Chemistry,
pharmacology and environmental toxicology.  Plenum Press,  New York.

American  Industrial  Hygiene  Association.    1970.    Hygienic   Guide  Series:
Pentachlorophenol  and  sodium  pentachlorophenate.   Am.  Ind.   Assoc.  Jcur.
31: 521..

Anderson, K.J.,  et al.   1972.   Evaluation of  herbicides  for  possible  muta-
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Armstrong, R.W., et  al.   1969.   Pentachlorophenol poisoning in a  nursery  for
newborn   infants.    II.    Epidemiologic  and   toxicologic  studies.    Jour.
Pediatr.  75: 317.

Baader,  E.w.  and H.J. Bauer.   1951.    Industrial  intoxication  due  to  penta-
chlorophenol.  Industr. Med. 4 Surg.  20: 286.

Borthwick, P.W.,  and  S.C.  Schimmmel.  1978.   Toxicity of Pentachlorophenol
and  related  compounds to  early life  stages  of  selected  estuarine animals.
Pages  141-146  In:   K.R.  Rao  (ed.),  Pentachlorophenol:   Chemistry,  pharma-
cology and environmental toxicology.  Plenum Press, N,Y.

Boutwell, R.K. and K.K.  Bosch.  1959.  The tumor-promoting action  of  phenol
and related compounds  for mouse skin.   Cancer Res.  19: 413.

Braun,  W.H.,  et al.   1977.   The  pharmacokinetics and metabolism of  penta-
chlorophenol in rats.  Toxicol.  Appl.  Pharmacol.  41:  395.
s_.
Braun, W.H.,  et al.   1978.   The metabolism/pharmacokinetics of pentachloro-
phenol in  man, and  a comparison  with the  rat  and monkey  model.  Toxicol.
Appl. Pharmacol.  45: 135.

Buhler, D.R.,  et  al.  1973.  Occurrence  of hexachlorophene and Pentachloro-
phenol in sewage and water.  Environ.  Sci. Technol.  7: 929.

Buselmaier, et  al.   1973.   Comparative investigations of the mutagenicity of
pesticides in mammalian test systems.  Mutat. Res.  21: 25.

Clendenning,  K.A.  and  W.J.  North.   1960.   Effects  of wastes on the giant
kelp, Macrocystis pyrifera.   Pages 82-91  In:  Proc. 1st  Conf.  on waste dis-
posal in the marine environment.   Pergamon Press, New York•

Cretney,   M.J.  1976.   Psntachlorophenol  death.   Bull.   T.I.A.F.T.   12  10.
In:   T.J.   Haley,   1977.   Human  poisoning  with  Pentachlorophenol and  its
treatment.  Ecotoxicol. Environ. Safety (In press).

Cserjesi,  A.J.  1967.   The adaptation of fungi  to  pentachloroohenol and its
biodegradation.  Can. Jour. Microbiol.  13: 1234.

-------
Deichmann, W.,  et  al.   1942.   Acute and chronic effects of pentachlorophenol
and  sodium  pentachlorophenate upon experimental animals.   Jour. Pharm. Exp.
Therap.  76: 104.

Ouagan,  R.E.,   and P.E.  Corneliussen.   1972.   Dietary intake  of  pesticide
chemicals  in -  the United  States  (III),  June   1968-April   1970.   Pestic.
Monitor. Jour.  5: 331.

Engst,  R.,   et  al.   1976.    The  metabolism of  lindane and  its metabolites
gamma-2,3,4,5,6-pentachlorocyclohexene,  pentachlorobenzene  and   pentachloro-
phenol  in  rats  and the pathways  of lindane metabolism.  Jour.  Environ. Sci.
Health  2: 95.

Fahrig,  R.,  et al.   1978.   Genetic activity  of chlorophenols and chloro-
phenol  impurities.  Pages  325-338' In:   K.R.  Rao (ed.),  Pentachlorophenol:
Chemistry,  pharmacology  and  environmental toxicology.   Plenum  Press,  New
York.

Goldstein,  J.A.,   et  al.   1977.   Effects of  pentachlorophenol  on hepatic
drug-metabolizing  enzymes and porphyria related to contamination with  chlor-
inated  dibenxo-p-dioxins and  dibenzofurans.  Biochem. Pharmacol.  26: 1549.

Hanadmad,  N.   1967.   Photolysis of  pentachloronitrobenzene, 2,3,5,6-tetra-
chloronitrobenzene and pentachlorophenol.   Ph.D.  dissertation.  University
of California,  Davis.

Harvey,  W.A. and  A.S.  Crafts.  1952.   Toxicity  of pentachlorophenol and its
sodium  salt  in  three yolo soils.  Hilgardia  21: 487.

Innes,  J.R.M.,  et al.   1969.  Bioassay  of pesticides  and  industrial chem-
icals  for  tumorigenicity  in  mice.   A preliminary note.   Jour.  Natl.  Cancer
Inst.   42: 1101.
Jakobson,  I.  and  S.  Yllner.   1971.  Metabolism  of  *C -pentachlorophenol  in
the mouse.  Acta.  Pharmacol. Toxicol. 29: 513.

Johnson,  R.D.  and  D.D.  Manske.   1977.'   Pesticides in  food  and- feed:  Pes-
ticide  and other  chemical  residues in  total diet  samples  (XI).   Pestic.
Monitor. Jour.  11: 116.

Johnson,  R.L.,  et al.   1973.   Chlorinated dibenzodioxins  and pentachloro-
phenol.  Environ.  Health Perspec. Exp. Issue  5: 171.
Karapally,  J.C.,  et  al.   1973.   Metabolism of  lindane-1 ^C in  the rabbit:
ether-soluble urinary metabolites.  Jour. Agric. Food Chem. 21: 311

Klemmer,  H.w.   1972.   Human  Health  and Pesticides  •-  community pesticide
studies.  Residue Rev.  41: 55.

Kobayashi,  K.  and H. Akitake.   1975.   Studies on  the  metabolism of chlaro-
phenols  in  fish.   I.  Absorption  and  excretion  of PCP  by  goldfish.   Bull.
Jap. Soc. Sci.  Fish.  41: 87.

-------
Koss, G.  and w. Koransky.  1978.   Pentachlorophenol in different  species  of
vertebrates   after   administration   of  hexachlorobenzene  and   pentachloro-
benzene.   Pages 131-137  In:   K.R.  Rao (ed.), Pentachiorophenol:   Chemistry,
pharmacology  and environmental toxicology.  Plenum  Press,  New  York.

Larsen,  R.V.;-  et  al.   1975.   Placental transfer  and  teratology  of  penta-
chlorophenol  in rats.  Environ. Lett.   10:  121.

Lui,  H.,  and C.D.  Sweeney.  1975.   Hepatic  metabolism of hexachlorobenzene
in rats.  FE3S  Lett.' 51:'225.

Mehendale,  H.M.,  et al.   1975,   Metabolism and effects of hexachlorobenzene
on hepatic microsomal enzymes  in  the  rat.   Agric. Food  Chem.   23: 261.

Mitchell,  L.C.   1961.   Effect of ultraviolet light (2537A)  on  141  pesticide
chemicals by  paper chromatography.  Jour. Off. Anal. Chem.  44:  643.

Mitsuda,  W.,  et al.  1963.   Effect of  chlorophenol analogues on  the  oxida-
tive phosphorylation in rat liver mitochondria.  Agric.  Biol.  Chem.   27:  366.

National  Fire  Protection  Assoc.   1973.  Fire protection  guide on  hazardous
materials.   5th ed.  Natl. Fire Prot.  Assoc.  Int.,  Boston.

National  Research Council.   1977.   Drinking water  and health.   Natl. Acad..
of Sci.  Washington, O.C.

Nomura,  S.   1953.   Studies  on  chlorophenol poisoning.  Podo  Kaguku Jour.
Sci. Labor   29: 474.

Parrish,  P.R.,  et  al.    1978.   Chronic toxicity of chlordane,   trifluralin,
.and pentachloroohenol  to sheepshead  minnows, Cyprinoden variegatus.   Report
No. EPA 60013-78-010:1.

Pruitt,  G.W.,  et al.   1977.   Accumulation and  eliminatrion  of  pentachloro-
ohenol  by   the  bluegill,   Leoomis  macrochirus.    Trans.  Am.  Fish.   Soc.
106: 462.

Reiner,  E.A.,  et  al.    1977.    Microbial  metabolism  of  pentachlorophenol.
Proc.  Symp.  on Pentachlorophenol,  June 27-29.   U.S.   Environ.  Prot.  Agency
and Univ. West  Florida.

Ruesink,  R.G. and L.L.  Smith, Jr.   1975.    The relationship  of the 96-hour
LCso  to  the  lethal  threshold concentration  of  hexavalent chromium, phenol,
and  sodium  pentachlorophenate   for   fathead minnows,  Pimeohales   oromelas
rafinesoue.   Trans.  Am. Fish. Soc.  104: 567.

Sax,  N.I.   1975.   Dangerous  properties  of  industrial materials.    4th  ed.
Van Nostrand Reinhold Co., New York.

Schimmel,  S.C.,  et  al.    1978.    Effects  of  sodium   pentachlorophenoL  on
several estuarine animals:  toxicity,  uptake, and depuration.   Pages 147-155
In:   K.R.   Rao  (ed.),   Pentachlorophenol:   Chemistry,   pharmacology  and
environmental toxicology.  Plenum Press, N.Y.

-------
Schwetz,  B.A.,  at al.   1974.   The  effect of purified  and commercial grade
pentachlorophenol  on rat  embryonal and  fetal development.   Toxicol. Appl.
Pharmacol.  28: 151.

Schwetz,  B.A.,  et al.   1978.  Results of  two-year toxicity and reproduction
studies  on pentachlorophenol  in  rats.   In:  K.R.  Rao  (ed.),  Pentachloro-
phenol:    Chemistry,   pharmacology   and  environmental  toxicology.   Plenum
Press, New York.

Spector,   U.S.     1956. '  Handbook   of   toxicology.    W.B.   Saunders  Co.,
Philadelphia.

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

Suzuki,  T. and  K. Nose.   1971.   Decomposition of  PCP  in  farm  soil.  Part
II.   PCP metabolism by  a  microorganism  isolated  from  soil.   Moyaku Seisan
Gijutsi  (Japan)   26: 21.

Takahashi, W.,  et al.   1976.   Acute phase  proteins  and pesticide exposure.
Life Sci.  19: 1645.

U.S. EPA.  1979.   Pentachlorophenol:  Ambient Water Quality  Criteria  (Draft).

Vogel, E.  and  J.L.R.  Chandler.   1974.   Mutagenicity testing of cyclamate  and-
some pesticides in Drosoohila melanooaster.  Experientia  30:  621.

Watanabe,  I.    1973.   Decomposition  of pesticides  by  soil  microorganisms.
Jap. Agric. Res.  Q.  7: 15.

Weast, R.C.  (ed.O.  1975-1976.   Handbook of chemistry and  physics.   5th  ed.
CRC Press, Cleveland, Ohio.

Webb,  P.W. and  J.R.  Brett.  1973.  -Effects of sublethal  concentrations of
sodium  pentachlorophenate  on  growth rate,  food  conversion efficiency,   and
swimming  performance in  underyearling  sockeye  salmon  (Oncorhynchus  nerka).
Jour. Fish. Res.  Board Can.  30: 499.

Weinbach,  E.C.  and J.  Garbus.   1965.  The  interaction  of uncoupling  phenols
with  Mitochondria and   with   Mitochondrial  protein.    Jour.  Biol.  Chem.
240: 1811.

Woelke,  C.E.    1965.   Development  of  a  bioassay method  using  the marine
algae, Monochrysis lutheri.  Wash. Dep. Fish. Shellfish Progress Rep.  9p.

Wong, A.S.  and D.G. Crosby.  1977.   Photodecomposition  of  pentachlorophenol
(PCP).   Proc.  Symp.  on Pentachlorophenol,  June  27-29.,  U.S.  Environ. Prot.
Agency and Univ. West Florida.

Yang,  R.S.H.,  et  al.   1975.   Chromatographic methods  for the  analysis, of
hexachlorobenzene  and  possible  metabolites  in  monkey fecal samples.  Jour.
Assoc. of Anal. Chem.  58:  1197.
                                     r.' Yf f*
                                   S  ) / I) $

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                                    No.  144
             Phenol

 Health  and  Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980
      - /Y/7 /-
    <^V )  V U

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented  by the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

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                            PHENOL


                            SUMMARY


     Insufficient data  exist  to  indicate that phenol is a


carcinogenic agent.  In skin  painting  studies,  phenol appears


to function primarily as  a  nonspecific irritant.  Information


on the mutagenicity of  phenol is  equivocal.   Phenol does not


appear to be teratogenic.   Chronic exposure  to  phenol at rel-


atively high levels causes  liver  damage  in humans and ani-
                         •

mals, and kidney damage in  animals.  Exposure to acutely tox-


ic levels of phenol causes  CNS depression.


     The toxic effects  of phenol  have  been extensively exam-


ined in freshwater organisms  by acute  studies in 13 fish and -


13 invertebrate species.  Considerable interspecies and intra


species variation were  described,  with acute  values ranging


from 5,020 to 780,000 ug/1.   Only three  marine  species were


examined in acute tests,  and  LC5Q values ranged from


5,200 to 58,250 ug/1.
                             i «-t*,i 7-
                            ^ / U 
-------
                           PHENOL

I.   INTRODUCTION

     This profile is based on the Ambient Water Quality Cri-

teria Document for Phenol (U.S. EPA, 1979).

     Phenol (CgH^OH; molecular weight 94.11) is a clear,

colorless (light pink when impurities are present) hygro-

scopic, crystalline solid at 25° C with the following physi-

cal and chemical properties (Manufacturing Chemist Assoc. ,

1974; Kirk and Othmer, 1963; Weast, 1974).

          Melting Point   43° C
          Boiling Point   182° C at 760 mm Hg
          Flash Point     open cup 85° C
                          closed cup 79° C
          Vapor Pressure  0.35 mm Hg at 25° C
          Solubility      Water:- 6.7 g/100 ml at 16° C and
                          is soluble at all proportions at
                          66° C.  Also soluble in e.ther, al-
                          cohol, acetic acid, glycerol, liq-
                          uid sulfur dioxide, benzene, and
                          oils.

     Industrial capacity for production is 1.44 to 10*> tons

per year (Chem. Eng. News, 1975).  About 90 percent of the

phenol produced is used in the production of phenolic resins,

caprolactam, bisphenol-A, alkylphenols, and adipic acid

(Chemical Profiles, 1972).

     Phenol may be biochemically hydroxylated to ortho- and

para-dihydroxybenzenes and readily oxidized to the corres-

ponding benzoquinones.  These may in turn react with numer-

ous components of industrial waters or sewage such as mercap-

tans, amines,  or the -SH or -NH groups of proteins (Stom,

1975).  When ambient water containing phenols is chlorinaced,

various chlorinated phenols may be produced in sufficient
                         yyy-y

-------
quantities to produce an  objectionable  taste and odor (Aly,

1968; Barnhart and Campbell,  1972;  Jolley,  1973;  Jolley,  at

al. 1975).

II.  EXPOSURE

     A.   Water  -

          There have been no  market basket  surveys  of free

and conjugated phenols with which  to estimate the average

daily dietary intake of phenols.   The National Organic Moni-

toring Survey (U.S. EPA,  1977)  reported  finding unspecified
                                             •
concentrations of phenol  in 2 out  of 110  raw water  supplies.

The Survey found no phenol in any  finished  water supplies.

The National Commission on Water Quality  (1975)  reported  an

annual mean concentration of  1.5 ug phenol/1 in raw water

from the lower Mississippi River.

     B.   ?ood

          Phenol is produced  endogenously in the mammalian

intestinal tract through  microbial  metabolism (Harborrie,

1964) and free and conjugated phenol is  a normal constituent

of animal matter (U.S. SPA, 1979).   Phenol  concentrations of

7 mg/kg in smoked summer  sausage and 28.6 mg/kg  in  smoked

pork belly have been reported (Lustre and Issenberg,  1970).

Several mouthwashes and lozenges contain  phenol  in  amounts  of

up to 32.5 mg total phenol/lozenge.

     The U.S. EPA (1979)  has  estimated  the-weighted average

bioconcentration factor for phenol  to be  2.3  in  the edible
                                                          #
portions of fish and shellfish  consumed by  Americans.  This

estimate is based on the  octanol/water partition  coefficient

of phenol.


-------
      C.   Inhalation



           The inhalation of. phenol  vapor  appears to be large-



• ly restricted to the occupational environment (U.S. EPA,



 1979).  Dermal exposures, can be  from a  number of medicinal



 preparations for skin application (lotions,  powders,  oint-



 ments) containing up to 4.75 percent phenol,  or from certain



 feminine hygiene products,  and hemorrhoidal  products (U.S.



 EPA, 1979) .



 III. PHARMACOKINETICS



      A.   Absorption



           Phenol is readily  absorbed by all  routes.  This is



 illustrated by the fact that acutely toxic doses of phenol



 can produce symptoms within  minutes  of  administration regard-'



 less of the route of entry  (U.S. EPA, 1979).   Sixty to 80



 percent of inhaled phenol is retained in  the  lungs.  Piotrow-



 ski (1971) found that phenol vapor  could  be  readily absorbed



 by intact human skin.  The  rate  of  dermal absorption for



 phenol, vapor can be represented  by  the  formula A=(0.35)C,



 when A is the amount of phenol absorbed in mg/hour  and C is



 the phenol concentration in  mg/m3 (Piotrowski,  1971;  recal-



 culation of data of Ohtsuji  and  Ikeda,  1972  by U.S. EPA,



 1979).



      B.   Distribution



           Free and conjugated phenol  appear  to be normal



 trace  constituents in humans  and other  mammals (Harborne,



 1964).  Values reported for  free and  conjugated  phenol in'



 normal human blood vary greatly  due  in  part  to  the  specifi-



 city of the analytical methods used  in  and in  part  to the

-------
amount of the dietary protein  which  increases urinary phenol


excretion.  Recent values  in normal  human  blood are between


0.04 to 0.56 mg/1 for the  free phenol  and  1.06 to 5.18 rag/1


for conjugated phenols  (Dirmikis  and Darbre,  1974).  For the


total phenol (free and  conjugated) a range between 2 and 18


mg/1 has been reported  (Van Haaften  and  Sie,  1965).


          Upon absorption, phenol is rapidly  distributed to


all organ systems, followed by relatively  rapid metabolism


and excretion.  Within  15  minutes of an  oral  dose, the high-
                           •                                 ^

est concentrations are  found in the  liver,  followed by heart,


kidneys, lungs, brain and  blood (Deichmann, 1944).


     C.   Metabolism


          The major metabolites "of phenol  are sulfate and


glucuronic acid conjugates of  phenol and 1,4-dihydroxyben-


zene.  There are, however, species differences in'the excre-


tion pattern of these metabolites (Capel,  et  al.  1972).   The


cat, which is sensitive to phenol, in  addition to sulfate


conjugated phenols, excretes also, as  a major metabolite,


1,4-dihydroxybenzene (Miller,  et  al.   1976).   The metabolic


pattern is also dose dependent.   Other agents,  which  are nor-


mally metabolized to phenol, such as benzene  or phenylsalicy-


late, produce increased urinary excretion  of  phenol metabo-


lites (Kociba, et al.   1976).


     D.   Excretion


          In humans and in all mammals that have  been tested,


nearly all of the phenol and its  metabolites  are  excreted .in


the urine within 24 hours  (U.S. EPA, 1979; Piotrowski,  1971;


Deichmann and Keplinger, 1963).   Reported  normal  background
                            . I **, 1 ^

                         *•• } > I U~ >

-------
values for human urinary phenol  range  from 1.5  to 5  mg/1



(Fishbeck, et al. 1975? U.S. EPA,  1979).   Urinary excretion



levels of phenol metabolites in  workers  exposed to phenylsal-



icylate ranged from 150 to  1,371 mg/1.   Upon  ingestion  of



eight chloraseptic lozenges at the  recommended  dosing  sched-



ule, the total phenol and the free  phenol  concentrations in



the urine peaked at 270 and 10 mg/1, respectively.   When dogs



were fed 125 mg phenylsalicylate/kg/day  for 41  days,  the peak



urinary phenol concentration was 6,144 mg/1 and the  treatment



was not associated with ill effects (Kociba,  et al.  1976).



The half-life of phenol in  man is  approximately 3.5  hours



(U.S.  EPA, 1979).



IV.  EFFECTS



     A.   Carcinogenicity



          There is no convincing evidence  that  phenol  acts  as



a carcinogen, particularly  at concentrations  within  normal



physiologic limits.  Phenol appears to function primarily as



a nonspecific irritant (NIOSH, 1976).  Only one case of  human



cancer associated with exposure  to  phenol  was found  in  the



literature.  A 12-year old man who  had applied  a  salve of



phenol and ergot to his back daily  for 20  years developed an



invasive squammous cell epithelioma (Stevens  and  Callaway,



1940).



     Phenol produced papillomas  but not "carcinomas when  ap-



plied  to the skin of some strains of mice.  Phenol has car-



cinogenic activity when applied  repeatedly to the  skin of a



specially bred strain of Sutter  mice at concentrations which



produce repeated skin damage (Boutwell and Bosch,  1959;  Sala-

-------
man and Glendenning, 1956) .   Phenol  promotes  skin cancer in
mice when repeatedly applied  after  initiation with known car-
cinogens (Boutwell and Bosch,  1959;  Salaman and  Glendenning,
1956; Van' Duuren, et al. 1971).   Tumorigenesis is highest at
dose levels of phenol, which have  some sclerosing activity.
Phenol has no cocarcinogenic  activity when applied simultane-
ously and repeatedly with  benzo(a)pyrene  to mouse skin (Van
Duuren, et al. 1973).
     .8.   Mutagenicity
          Phenol was found to  be  mutagenic in Drosphila (Ha-
dorn and Niggli, 1946) and also reported  to be nonmutagenic
for Neurosoora (Dickey, et al. 1.949).   Phenol produced back
mutations in E. coli from  streptomycin  dependence to  non-de-
pendence at phenol concentrations high  enough that the
survival of bacteria was only '0.5 to 1.7  percent (Demerec,  et
al. 1951).
     C. Teratogenicity
          Studies dealing directly with teratogenicity were
not reported in the O.S. EPA  (1979) or  NIOSH  (1976) docu-
ments.  In a study, not designed  specifically as a teratogen-
icity study, rats were given phenol at  concentrations  of 100
to 12,000 mg/1 in their drinking water  over three to  five
generations.  Specific teratogenic effects were  not noted
(Heller and Pursell, 1938).
     D.   Other Reproductive Effects
                                                          #
          In the study mentioned under  teratogenicity,  higher
concentrations of phenol in the drinking  water (7,000  mg/1)
produced stunted growth in the young, death of the offspring

-------
at birth (10,000 mg/1), and  failure  to  reproduce  (12,000



mg/1) (Heller and Pursell, 1938).



     E.   Chronic Toxicity



          Repeated exposures  to phenol  at  high  concentrations



have resulted in chronic liver damage  in humans (Merliss,



1972).  In unpublished studies by Dow Chemical  Company. 	



(1976), rats received  135 doses of 100  mg  phenol/kg  or  50  mg



phenol/kg by gavage over a six month period.  The  growth of



the rats was comparable to that of controls»  Very slight



liver changes and slight to moderate kidney damage were  seen.



at the higher dose of phenol.  The lower dose of phenol pro-



duced only slight kidney damage;



          Rats given phenol  in their drinking water  at  300,



1,200, 1,600, 2,000, and 2,400 mg/1 had corresponding average



intakes of 21, 30, 49, 56, and 55 mg phenol per rat  per day



based on actual water consumption data.  The rats  at the



three lower dosage levels showed no overt  symptoms of toxic-



ity.  The weight gain of the  rats at the two highest dose



levels was depressed (Deichmann and Oesper, 1940).



     F.   Other Relevant Information



          The primary effect  of exposure to acutely  toxic



levels of phenol is CNS depression.  Significant evidence



could not be found to support the occurrence of synergistic



or antagonistic actions of phenol with  other compounds  in



mammals (U.S. EPA, 1979).

-------
V.   AQUATIC TOXICITY



     A.   Acute Toxicity



          Acute toxicity data  for  phenol  display a wide range



of interspecific variability and  intraspecific  sensitivity.



The range of LCgQ values for 13 species of  freshwater



fish is 5,020 ug/1"for the  rainbow trout  (Salmo gairdneri)  to



200,000 ug/1 for the goldfish  (Carassius  auratus)  (Cairns,  et



al. 1978).  Several studies have  indicated  an  inverse rela-



tionship between survival time and temperature  for rainbow  .
               r


trout, golden shiner (Notemigonius crysoleueus)  (U.S.  EPA,



1979).  Similar intraspecific  sensitivity and  interspecific



variability was demonstrated by bioassays with  freshwater in-



vertebrates as test, organisms.  The  cladbcerans,  Daphnia



magna and _D. lonqispina, displayed the greatest  sensitivity



to phenol with LC5Q values  as  low  as  7,000  ug/1  reported.



The freshwater clam, Sphaerium corneum, was  the most  resis-



tant species with an LC50 value of 780,000  ug/1  (U.S.



EPA, 1979).



          Data for the acute toxicity of  phenol  to marine or-



ganisms is not nearly as extensive as that  for  freshwater



species.  For marine fish,  LC$Q values of 5,200 and 6,014



ug/1 were obtained for rainbow trout  in saline waters and



mountain bass (Kuhlia sandvicensis), respectively  (U.S.  EPA,



1979).  Eastern oyster embryos (Crassostrea virginica)  and



hardclam embryos (Mercenaria mercenaria)  were much more  re-



sistant with LC5Q values of 58,250  and 52,630 ug/1,



respectively (Davis and Hidu, 1969).
                              jar

-------
     B.   Chronic



          Data for the chronic  effects of  phenol  on fresh-



water fish are not available.   In a  life cycle  chronic test,



a chronic•value of 3,074 ug/1 was obtained for  the  freshwater



cladoceran, Daphnia magna  (U.S. EPA,  1978).   Chronic  data



for marine organisms were  not available..



     C.   Plant Effects



          Plants are relatively insensitive  to  phenol  expo-



sure with effective concentrations ranging from 20,000 to



1/504,000 ug/1 for three species of  algae,  one  species of



diatom, and duckweed.  Marine plants  species  have not  been



examined for toxic effects of phenol.



     D.   Residues



          Measured bioconcentration  factors  of  1.2  to  2.3



have been determined for goldfish (Kobayashi, et  al. 1976;



Kobayashi and Akitake, 1975).   Bioconcentration factors have



not been determined for freshwater invertebrates  or plants,



or for any marine species.



VI.  EXISTING GUIDELINES AND STANDARDS



     Neither the human health nor the aquatic criteria de-



rived by U.S. EPA (1979), which are  summarized  below,  have



gone through the process of public review;  therefore,  there



is a possibility that these criteria will  be  changed.



     A.   Human



          On the basis of  chronic toxicity data for rats  and



an uncertainity factor of  500,  the U.S. EPA  (1979)  has de-



rived a draft criterion of 3.4  mg/1  for phenol  in ambient



water corresponding to the calculated acceptable daily intake

-------
of 0.7 rag.  The draft  criterion  for phenol is 1.0 ug/1 in
those instances where  chlorination  of  phenol may take place
during water purification  processes.
         :The 1974 Federal standard and  the ACGIH (1977)
recommendation foe phenol  in  air in the  workplace is 19
mg/m^ (5 ppm) as a time-weighted average.
          The NIOSH  (1976)  criterion for a recommended stand-
ard for occupational exposure to phenol  is 20 mg/m^  in air
as a time weighted average for.up,to a 10-hour work  day and a
40-hour work week, with .a  ceiling concentration of 60 mg/rn^
for any 15-rainute sampling period.
          The U.S. EPA interim drinking  water limit  for
phenol is 1 ug/1, which is largely  an  organoleptic standard
based on the objectionable taste and odor  produced by chlori-
nated phenols.  In response to a phenol  spill, in southern
Wisconsin, the "J.S_  EPA proposed on November 26,  1974 an
emergency standard of  0.1  rag  phenol/1  as being temporarily
acceptable for human consumption.
     B.    Aquatic
          The draft criterion for protecting freshwater or-
ganisms is 600 ug/1, not to exceed  3,400 ug/1.   No criterion
for marine organisms was derived (U.S. SPA,  1979).
                             I l-r "*-
                            •7l) i h

-------
                                    PHENOL

                                  REFERENCES

Aly,  O.M.  • 1968.   Separation  of phenols  in  waters  by  thin-layer chromato-
graphy.  Water Res.  2: 287.

American  Council for  Governmental  Industrial  Hygienists.   1977.  Threshold
limit  values  for chemical  substances  and physical agents  in workroom envi-
ronment with intended changes for 1977.

Barnhart,  E.L.   and G.R.  Campbell.   1972.   The  effect  of  chlorination on
selected organic chemicals.  U.S. Environ. Prot. Agency.

Boutwell,  R.K.  and  O.K.  Bosch.  1959.  The tumor-promoting action .of phenol
and related compounds.  Cancer Res.  19: 413.
                       a

Cairns, J.,  Jr.,  et al.   1978.  Effects  of  temperature on aquatic organisms
sensitivity to  selected  chemicals.   Project B-084-VA.  Bull. 106.  Virginia
Polytechnic Inst. State University.

Capel, I.D., et al.  1972.  Species variations in the metabolism of phenol.
Biochem. Jour.  127: 25.

Chemical and Engineering News.  July 28, 1975.

Chemical Profiles.  1972.  Phenol. Schnell Publishing Co., New York.

Davis, H.C.  and H. Hidu.   1969.   Effects of  pesticides  on embryonic devel-
opment of  clams  and oysters and on  survival and growth of  the larvae.  Fish
Wildl. Fish. Bull.  67: 393.  U.S.. Oep. Inter.

Deichmann,  W.B.   1944.   Phenol studies.  V.   The  distribution,  detoxifica-
tion, and excretion of phenol in the mammalian  body.  Arch. Biochem.  3: 345.

Deichmann, W.B.  and M.L.  Keplinger.   1963.  Phenols  and -phenolic compounds.
Page  1363  In:  F.A. Patty  (Ed.),  Industrial  hygiene  and  toxicology.   Inter-
science Publishers, New York.

Deichmann, W.B. and P. Oesper.   1940.   Ingestion of phenol — Effects on the
albino rat.  Ind. Med.  9: 296.

Demerec,  M., et al.  1951.   A  survey  of chemicals for mutagenic action on E.
coli.  Am. Natur.  85: 119.

Dickey, F.H., et  al.   1949.  The role  of  organic  peroxides in the induction
of mutations.  Proc. Natl. Acad. Sci.   35: 581.

Dirmikis,  S.M.  and A.  Darbre.  1974.   Gas-liquid chromatography  of "simple
phenols for urinalysis.  Jour.  Chromatogr.  94: 169.

-------
Oow  Chemical  Co.    1976.   References  and  literature review  pertaining  to
toxicological properties of phenol.  Toxicol. Res.  Lad.   Unpubl.  Manuscript.

Fishbeck, W.A.,  et  al.   1975.   Elevated urinary phenol levels  not  related to
benzene exposure.   Am. Ind. Hyg. Jour.  36:  820.

Hadom,  E.  and  H.   Niggli.   1946.  Mutations  in  Orosophila after  chemical
treatment of gonads in vitro.  Nature   157:  162.

Harborne, J.B.   1964.  Biochemistry of phenolic compounds.  Academic  Press,
New York.

Heller,  V.G.  and  L. Pursell.   1938.   Phenol'-contaminated  waters  and  their
physiological action.  Jour. Pharmacol.. Exp. Ther.   63: 99.

Jolley, R.L.   1973.  Chlorination effects  on  organic constituents in  efflu-
ents  from domestic sanitary sewage  treatment plants.   Ph.D. dissertation,
University of Tennessee, Knoxville.

Jolley,  R.L..,  et   al.   1975.   Chlorination of cooling  water:  a  source  of
envi-  ronmentally  significant chlorine-containing  organic compounds.   Proc.
4th Natl. Symp. Radioecology.  Corvallis, Ore.

Kirk,  R.E.  and  O.F. Othmer.   1963.    Kirk-Qthmer  encyclopedia  of  chemical
technology.  2nd ed.  John Wiley and Sons, Inc., New York.                   ;

Xobayashi, K.  and  H. Akitake.   1975.   Metabolism• of chlorophenols  in  fish.
IV.  Absorption   and  excretion  of  phenol  by   goldfish.   Nippon  Suisan
Gakkaishi.  41: 1271.

Kobayasni, K., at  al.  1976.  Studies  on  the  metabolism of  chlorophenols  in
fish.  VI.  Turnover of absorbed  phenol in  goldfish.  Bull.  Jap.  Soc. Sci.
Fish.  42: 45.

Xociba, R.J.,  et al.  1976.  Elevated  urinary phenol  levels in beagle dogs
treated with salol.  Am. Ind. Hyg. Jour.  37: 183.

Lustra, A.O.  and P.  Issenberg.   1970.  Phenolic  components  of  smoked meat
products.  Jour.  Agric. Food Chem.  18: 1056.

Manufacturing  Chemists  Assoc.    1974.    Chemical   safety  data  sheet   SD-4;
Phenol.  Washington, O.C..

Merliss,  R.R.  1972.  Phenol moras.  Mus. Jour. Occup. Med.   14:  55.

Miller, J.J., et al.  1976.  The  toxicity  of dimethylphenol and  related com-
pounds on the cat.  Toxicol. Appl. Pharmacol.  38:  47.

National  Commission on Water  Quality.   1975.   Water  quality and  environ-
mental assessment  and predictions  to   1985  for  the  lower  Mississippi River
and Barataria Bay.  Vol. 1.  Contract WQ5AC062.

National  Institute  for  Occuoational Safety  and Health.   1976.   Criteria for
a recommenced standard...Occupational exposure to phenol.  NIOSH  76-196.

-------
Ohtsuji,  J.  and M.  Ikeda.  '1972.  Quantitative  relationship  between atmos-
pheric  phenol vapor  and phenol  in the'  urine of  workers in  bakelite  fac-
tories.  Br. Jour. Ind. Med.  29: 70.

Piotrowski,, J.K.   1971.   Evaluation  of  exposure  to phenol:  absorption  of
phenol  vapour  in  the lungs and through the  skin and excretion  of phenol in
urine.  Br. Jour. Ind. Med.  28: 172.

Salaman, M.H.  and O.M. Glendenning.   1956.   Tumor promotion in mouse skin by
sclerosing agents.  Br. Jour. Cancer  11: 434.

Stevens,  J.B.  and  J.L.  Callaway.    1940.   Mixed  epithelioma  of  the  back
arising  from  daily  applicaton  of  a phenl  and  ergot  ointment.   Am.  Jour.
Cancer.  38: 364.

Stom, O.J.   1975.   Use of thin-layer and paper chromatography for detection
of  ortho- and para-  quinones  formed  in the  course  of  phenol  oxidation.
Acata Hydrochim. Hydrobiol.  3: 39.

U.S.  EPA.    1977.   National Organic  Monitoring  Survey.   General  review  of
results and methodology: phases I-III.  Water Supply Res. Oiv.

U.S.  EPA.   1978.   In-depth  studies on" health and environmental  impacts  of
selected water pollutants.  Contract No. 68-01-4646.                        ;

U.S. EPA.  1979.  Phenol: Ambient Water Quality Criteria. (Draft)

Van Duuren,  B.L., et  ai.   1971.  Cccarcinogenesis  studies  on  mouse skin and
inhibition of  tumor production.  Jour. Natl.  Cancer Inst.  46: 1039.

Van Duuren,  B.L.,  et al.   1973.  Cocarcinogenic agents  in  tobacco carcino-
genesis.  Jour. Natl. Cancer Inst.  51: 703.

Van Haaften,  A.B. and S.T.  Sie.  1965.   The measurement of phenol  in urine
by gas  chromatography as a check on  benzene exposure.   Am.  Ind. Hyg. Assoc.
Jour.  26: 52.

Weast,  R.C.  (Ed.)   1974.  Handbook  of  chemistry and physics.   55th ed.  CRC
Press, Cleveland, Ohio.
                                     -li

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                                     No. 145
              Phorate

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON,  D.C.   20460

           APRIL 30,  1980
           MS-1

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                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is  drawn chiefly
from secondary  sources  and  available reference documents.
Because of the limitations of such sources, this  short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented, by the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                         Disclaimer Notice
Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.

-------
                                    PHORATE
                                    Summary

     Phorate is an organophosphorous insecticide  used on a variety of crops,
mainly in south-central states.  Phorate  is  readily  absorbed through inhala-
tion and skin contact and'is  highly toxic to humans and other animals.  Pri-
marily, it  affects the  central and peripheral nervous  systems by inhibiting
cholinesterase activity.   Information  concerning carcinogenic and mutagenic
effects was not located  in the available literature.   The  threshold limit
value  for  phorate  is  50  ug/m3,  based  on  dermal  contact.   Additionally,
phorate has been classified for restrictive use by the U.S. EPA.
     Although phorate is  highly toxic to  certain aquatic organisms,  no  ap-
parent adverse effects have been observed in the aquatic environment.

-------
 I.    INTRODUCTION
      Phorate  is  a  highly  toxic organophosphorous insecticide used on a vari-
 ety of agricultural crops.   It was  introduced in 1954 by the American Cyana-
 mid  Co.  under the trade  name Thin.iiS) (Martin  and worthing, 1974).  Phorate
 is prepared by the reaction of phosphorous pentasulfide  vdth ethanol, for-
 maldehyde,  and ethyl  mercapton.  Production in  the  U.S.  totaled 3400 tonnes
 in 1977  (NAS,  1977).  Virtually all of the phorate is used on root and field
 cropsoils  to  control  sucking insects  and  nematodes  (NAS,  1975).  Phorate is
 slightly  soluble in water  and hydrolyzes  in moisture.   It has  an overall
 degradation rate constant of 0.02/d'ay and  a  bioconcentration factor of 5.2.
 Other properties are listed in Table 1.
 II.   EXPOSURE
      A.   Mater
          Phorate is produced in  the  United States  by  the American Cyanamid
 Co. at Hannibal, Mo.  (SRI,  1977).   Available information on an annual U.S.
 production  shows that  1900  tonnes  were  produced in  1971,  3600  tonnes  in
                                                           i
 1974,  and 3400 tonnes in  1977 (NAS,  1975,  1977).  Berg, et  ai.  (1972)  noted
 an application rate  of 1 pound of actual  material  per acre  (1.1  kg/ha;  in
 this case,  to control corn borers).   Application rates vary according'to use.
          Phorate has found  increasing use on croplands in the south-central
 states to   protect cotton,  hops,  alfalfa,  barley,  sorghum,  peanuts,  sugar
beets, sugar  cane, potatoes,  rice, and  tomatoes.   Only  small amounts  are
used  in  the southeastern and northeastern U.S.   American Cyanamid  Co.  re-
ported that phorate may fill  the  void  left by the removal from the market of
chlorinated hydrocarbons  and projected a  strong demand  for phorate in  the
corn rootworm market (Berg,  et al. 1977).

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              TABLE  1.   PHYSICAL AND CHEMICAL PROPERTIES OF PHORATE
Synonyms:  0,0-diethyl-S-(ethyIthiomethyl)phosphorodithioate;
           0,0-diethyl-S-ethylmercaptomethyl dithiophosphate;
           THIMET American Cyanamid (3911): timet (USSR);
           CAS Registry No. (298-02-2); Dranutox; Rampart; Vergfru

Structural Formula:  (C2H50)2(P=S)SCH2SC2H5

Molecular Weight:  260.4

Description:  Clear liquid
              Miscible  with:   CQ4,  dioxan,  vegetable  oils,  xylene,  alco-
              hols, ethers, esters

Soil Attenuation:  Kd approx.  5 x 1Q2; KOC = 3199

Specific Gravity and/or Density:  d25 - 1.167

Melting and/or Boiling Points:  bp 118 to 12QOC at 0.8 mm
                                mp less than -150C

Stability:  Stable at room temperature
            Hydrolyzed in the  presence of moisture
            Overall degradation rate constant (0.02/day)

            Soil half-life: 1-4 weeks
            Bacterial/Hydrolysis:  constant = 8 x

Solubility (water):  50 ppm at room temp.
                   sediment .  4.5
                     H20    *   1
Vapor Pressure:  8.4 x 10-4 mm Hg at 20°C

Bioconcentration Factor (BCF)  and/or
Octanol/water partition coefficient (KQW):   KOW  = 18
                                            BCF  =5.2
Source:  Martin and Worthing,  1974;  Fairchild,  1977; Windholz, 1976;
         U.S.  EPA, 1980

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          Little  information  was  found  on phorate  production  processes.
Lawless, et al.  (1977)  noted that in the production, crude phorate was wash-
ed  and filtered.  No  information was  given on the  treatment of  the waste
water  or  filter  cake associated with this process.   No  information on waste
sludge or landfill disposal was found in the available literature.
          Phorate  can enter  water  by  runoff  or  by  ground  water  drainage
after  application.   Phorate  is  relatively stable in  ground  water.  Only 10
percent decomposition was  estimated in a  river environment  in 5 days (50 to
250 mile transport; 80-400 km).   Also,  estimates show  that less than 90 per-
cent decomposition per  year  occurs  in  a lake environment- (U.S.  EPA, 1980).
There are no estimates on  the amount of phorate entering  the environment or
on the levels of phorate in  ambient water.  Menzie (1974)  noted that phorate
decomposes to  phorate sulfoxide  and phorate sulfone  and  the  sulfoxide  and
sulfone of the oxygen analog.
          Walter-Echols and  Lichtenstein  (1977) showed  that  some  oxidation
products of phorate (phorate  sulfoxide)  reduce  to  phorate in  lake mud under
certain conditions.   Using a flooded phorate  sulfoxide-treated loam  soil,
they noticed the  production of only small amounts  of phorate.  After  lake
mud was added, the reduction of phorate sulfoxide to  phorate increased dra-
matically and,  after two weeks'  incubation,  accounted for  44  percent  of  the
recovered residues.  They  related the reduction process to  the activity  of
microorganisms in an  environment of organic nutrients.
                    f
     B.   Food
          Information available in the open literature does' not quantify  the
amount of phorate detected on foods.   In a study  reported by  Menzie  (1974),
phorate was applied to  bermuda grass and  com  at the rate  of 2 pounds per

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acre  (2.2 kg/ha).  Fourteen  days  after treatment,  less than  1 ppm  phorate
residue  was noted on  the  corn;  after  21  days less than 1  ppm was found  on
bermuda  grass.
     C.   Inhalation and Dermal
          Data are  not available indicating the  number  of  people subject  to
inhalation  or  dermal  exposure to phorate.   The  primary  human exposure would
appear to occur  during production  and application.. The-.U.S. EPA (1976) list-
ed by occupational  group  the frequency of illness  caused by exposure to or-
ganophosphorous  pesticides.   Of  1157 reported cases, most illnesses occurred
among ground applicators  (229)  and mixer/loaders (142); the lack of,  or re-
fusal to use, safety  equipment was a major  factor of  this contamination.
Other groups affected  were gardeners (101), field  workers  exposed to pesti-
cide residues  (117), nursery  and greenhouse  workers (75),  soil  fumigators  in
agriculture (29), equipment cleaners and mechanics  (28),  tractor drivers and
irrigators  (23), workers exposed to  pesticide  drift (22),  pilots (crop dust-
ers) (17),  and  flaggers for  aerial  application  (6).  Most  illnesses  were a
result of carelessness, lack of  knowledge  of  the hazards, and/or  lack  of
safety equipment.   Under  dry,  hot  conditions,  workers tended  not to  wear
protective  clothing.    Such   conditions also  tended to increase  pesticide
levels and dust on the workers.
III. PHARMACOKINETICS
     A.    Absorption
          Newell and Dilley  (1978)  exposed four different groups  of  rats to
phorate  via four  routes   of  administration.  They compared  LD-Q an(j  \_£5Q
values and found that  inhalation was the  most toxic route,  followed,  in  de-
creasing  order, by  intravenous,  oral,  and dermal routes.  The phorate ae.ro-

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 sol generated  in  the  laboratory had a particle size range of  0.3-3.0 \m dia-
 meter,  a size  small enough to enter the  gas  exchange regions of the lung.
          Young,  et  al. (1979)  reported on two  occupational  exposure  inci-
 dents that  suggested  absorption in the lungs was the most effective -route of
 entry.   In both  cases,  the individuals wore protective  clothing, goggles,
 and respirators while working in the  dust house where technical grade  phor-
 ate was  produced.  Gas  chromatographic analyses  of air samples from the dust
 house  showed phorate  levels ranging  from  0.7  to 14.6 mg/nP.   NO estimate
 of  particle size was reported by the authors.
     B.   Distribution
          Phorate would be expected to  distribute in the body like organo-
 phosphorous pesticides  of  similar  solubility.   A report by Pugh and Forest
 (1975)  described  the  distribution  in  calves exposed to phorate in a manger
 containing 1200 ppm.  Phorate concentrations in  the liver ranged from 0.004-
 0.26 ppm; in the kidney, O.CQ2-Q.021. ppm; and in the brain, 0.025-0.19 ppm.
     C..  Metabolism
          The  major  phorate metabolites found  in  blood after  oral admini-
 stration to rats are phorate sulfoxide,  phorate  sulfone,  and phoratoxon sul-
 fone (NAS,  1977).   Bowman  and  Casida  (1958) showed that  phorate hydrolyzes
 in  rats  to  produce  urinary diethylphosphorodithioic acid,  diethylphosphoro-
 thioic  acid,   and diethylphosphoric  acid.   Oxidative metabolites are  not
 found as components  of  excretory  products  of animals treated with  phorate
 (NAS,  1977).   However,  DuBois,  et al.  (1950)  showed  that   in  rat  liver
 slices,  phorate was converted to its oxidative products.
     D.    Excretion
          The previous  section  notes  that  phorate  is eliminated  primarily
through the urinary system.

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IV.  EFFECTS
     A.   Carcinogenicity and Mutagenicity
          Pertinent data could not  be located in the available literature  on
the carcinogenicity  or mutagenicity  of phorate.   Formaldehyde,  a suspected
carcinogen, and other contaminants may be present in technical grade phorate.
     B.   Teratogenicity
          In  a  study  described  in  the absorption  section of  this  report,
Newell and Oilley  (1978) did  not find dose-related teratogenesis in rats ex-
posed  to  phorate via  inhalation,  intravenous,  dermal,  or oral  routes.   In
the chick embryo test, Richert and  Prahlad  (1972)  injected 1.5 or 2.0 ppm  in
a peanut oil  medium  into eggs on the tenth day of incubation.  Controls re-
ceived only peanut oil.  Hatchability of the eggs  decreased in a dose-depen-
dent manner.   Malformations  were  produced,  but  these  did not  seem  to be
dose-related.    The  relevance of these studies to  mammalian teratology is
unclear (NAS,  1977).
     C.   Other Reproductive Effects
          In a  study  in which CFI  mice were fed diets  containing  98.7 per-
cent phorate at 0.6,  1.5,  and 3.0  ppm, the no-adverse-effect level  for re-
productive performance was 1.5 ppm (NAS, 1977).
     0.   Chronic Toxicity and Other Relevant Information
          Pertinent data  on  chronic  toxicity  could  not  be  located  in the
available literature.   Several  subchronic  studies have been reported.  In
subchronic feeding studies  of 1, 5,  and  25 ppm phorate for 28  days,  choli-
nesterase in the 1 ppm group  was not decreased  (Tusing, 1955).   In  a second
rat study, Tusing (1956) fed  groups of 50  males and females 92 percent phor-
ate for 13 weeks  at  0.22,  0.66,  2.0,  6.0,  12.0, and  18.0 ppm.   He  noted  a
no-adverse-effect dosage at 0.66  ppm.

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           Tusing  (1956)  fed  three dogs  92 percent  phorate at  0.01, 0.05,
 0.25 and 1.24 mg/kg 6  days per  week  for 13-15 weeks.  The no-adverse-effect
 dosage  was judged to be  0.01 mg/kg;  even  at  this level, a  very slight de-
 crease  in plasma cholinesterase resulted.  Higher dosages, caused significant
 depression of  cholinesterase,  culminating  in  death at  the   two  highest
 dosages.
           Rat  feeding  studies showed  higher subchronic toxicities on phorate
 oxidative metabolites  than  on  phorate,  according   to  Rombunski,  et  al.
 (1958).   Others have  also noted  that  phorate metabolites  are  more  potent
 cholinesterase inhibitors than phorate (Curry, et al.  1961).
           Young, et al.  (1979) reported  on  acute  exposures to high levels of
 phorate  (up to  14.6  mg/m5)  in  a production  facility (see  absorption  sec-
tion).  The  symptoms  accompanying the exposures  were confusion,  dizziness,
nausea,  vomiting, pupil constriction, respiratory distress, cardiac arrhyth-
mia, and  unconsciousness.  Treatment  involved  a regime of RAM and  atropine.
According  to Gleason  (1969),  the symptoms  produced by a sub-lethal dose  are
typical of central  and peripheral nervous  system toxicity.   EPA's  accident
files contain  reports  of 21  episodes  of  poisoning   involving  phorate  for
1971-1973.  Eleven  were  agriculturally  related.   There  are  no controlled
studies  in humans from which no-adverse-effect  dosages could be derived.
          For humans,  the lowest published lethal (UD^) value is estimat-
ed to be  5 mgAg.  The  following studies list acute  phorate rtoxicity  levels
for human and nonhuman species, reported  by  Fairchild  (1977):
                                Hf-ll

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         Species              Exposure           LDgn  (mg/kg)
         Rat                   Oral.                    1.1
         Rat                   Skin                    2.5
         Rat                Intravenous                1.2
         Mouse                 Oral                   11
         Guinea pig            Oral                   20
         Guinea pig            Skin                   20
         Duck                  Oral                    2.55
         Duck                  Skin                   203
         Wild Bird             Oral                    1
V.   AQUATIC TOXICITY
     A.   Acute and Chronic Toxicity
          Phorate  is  highly toxic  to certain species of fish, crustaceans,
and terrestrial wildlife .(NAS, 1977).   NAS noted that there were no reported
killings of these species in the environment.
     B.   Plant Effects and Residues
          Pertinent data could not be located in the  available literature.
VI.  EXISTING GUIDELINES
     A.   Human
          The threshold  limit  value for  phorate is  50 jUg/nP,  based on skin
contact  (Fairchild,  1977).   An  8-hour  time-weighted  average  of  50  mg/m?
was adapted for phorate by the Tennessee  Department of Health (Young,  et al.
1979).  In  addition,,  phorate is classified for restrictive  use  by the U.S.
EPA for liquid  formulations  containing 65 percent  and greater active ingre-
dients.  The  restriction  was  influenced by  the  acute  dermal toxicity  of
phorate and by residue effects on avian species  (applicable  to foliar appj.i-
cations only).

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8.   Aquatic
     Pertinent data could not be located in the available literature.

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                                   REFERENCES
 Berg,  G.L.,  et al.  (ed.)   1977.   Farm Chemicals Handbook.   Meister Publish-
 ing  Company, willoughby, Ohio.

 Bowman,  J.s.  and  J.E.  Casida.   1958.  Further  studies on the  metabolism of
 Thimet by plants,  insects and mammals.  Jour.  Econ.  Entomol.   51:  838.

 Curry, A.M.,  et al.  1961.  Determination  of residue of phorate  and its in-
 secticidally active metabolites by  cholinesterase  inhibition.  Jour.  Agric.
 Food Chem.  9:  469.

 DuBois,  K.P.,  et al.  1950.  Studies on  toxicity and  pharmacological  action
 of octamethyl pyrophosphoramide.   Jour. Pharmacol.  Exp.  Ther.  99: 376.

 Fairchild, E.J. (ed.)  1977.  Agricultural Chemicals and Pesticides.  A Sub-
 file of  the NIOSH Registry of Toxic Effects of  Chemical Substances.   U.S.
 DHEW.

 Gleason,  M.N.,  et al.  1969.   Clinical  Toxicology  of Commercial  Products.
 Acute Poisoning.   3rd ed.

 Lawless, E.W.,  et al.  1972.   The Pollution Potential in Pesticide Manufac-
 turing.  U.S. EPA, Office  of Water Programs, Technical Studies  Report  TS-00-
 72-04.

 Martin,  H. and  C.R.  Worthing (eds.)  .1974.   Pesticide Manual.  4th ed.

 Menzie,  C.M.  1974.   Metabolism of Pesticides: An Update.  U.S. Dept.  of the
 Interior Special Scientific Report - Wildlife  No. 184,  Washington,  O.C.

 National Academy  of  Sciences.  1975.   Pest Control:   An  Assessment of Pre-
 sent and Alternative Technologies, Vol. I.

 National  Academy  of Sciences.   1977.   Drinking  Water  and  Health.   Natl.
 Acad. Sci., Washington, D.C.

 Newell,  G.W.  and  J,.V. Dilley.   1978.   Teratology  and Acute  Toxicology  of
 Selected Chemical  Pesticides Administered by Inhalation.   U.S. NTIS, PB Rep.
 PB-277077, 66 pp.

 Pugh, W.S. and  O.N.T. Forest.  1975.   Outbreak  of organophosphate  poisoning
 (Thimet) in cattle.   Can. Vet.  Jour.  16: 56.

 Richert, E.P.  and K.V.  Prahlad.   1972.   The effect- of  the organophosphate
0,0  diethyl  S-C(ethylthio)methyl]phosphorodithioate  on the  chick.  Poultry
 Sci.  51: 613..

Rombunski,  et al.  1958.  Cited in National Academy of Sciences, 1977.
                                                                          »
Stanford Research  Institute.  1977.   Directory of  Chemical  Producers.   Stan-
 ford Research Institute, Menlo Park, California.

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Tusing,  T.w.   1955.   Unpublished report  of American Cyanamid.   Cited in  U.S.
EPA Initial Scientific and Minieconomic Review of Phorate, 1974.

Tusing,  T.W.   1956.   Unpublished report  of American Cyanamid.   Cited in  U.S.
EPA Initial Scientific and Minieconomic Review of Phorate, 1974.

U.S.  Environmental  Protection  Agency.    1976.   Organophosphate  Exposure  from
Agricultural Usage,  EPA 600/1-76-025.

U.S. Environmental Protection Agency.   1980.   Aquatic Fate and Transport Esti-
mates  for Hazardous  Chemical  Exposure  Assessments.   Environmental  Research
Laboratory, Athens,  Georgia.

Walter-Echols, G. and  E.P.  Lichtenstein.  1977.  Microbial  reduction of phor-
ate sulfoxide to phorate in  a soil-lake mud-water  microcosm.    Jour.  Econ.
Entomol.  70: 505.

Windholz, M.  (ed.)  1976.   The Merck Index, 9th ed.  Merck  Co.,  Inc., Rahway,
New Jersey.

Young, R.J., et al.   1979.  Phorate  intoxication at an insecticide formulating
plant.  Am. Ind.  Hyg.  Assoc.  Jour.   40: 1013.

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                                      No. 146
          Phthalate Esters

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                           DISCLAIMER
     This report  represents a survey oz  the  potential health
and environmental hazards  from  exposure  to  the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources,  this short profile
may not reflect all available  information including  all the
adverse health  and  environmental  impacts  presented  by  the
subject chemical.   This document has  undergone scrutiny  to
ensure its technical accuracy.
                            ^^^^^^^^^^J^i
                           ~ / J 0 "

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                               PHTHALATE ESTERS
                                    Summary

     Certain' phthalates  (dimethyl  phthalate,  diethyl  phthalate,  mono-2-
ethyl-hexyl  phthalate  and  dimethoxyethyl phthalate),  have  shown  mutagenic
effects in both bacterial systems and the dominant lethal assay.
     All  eight phthalates  tested  by  injection  in  pregnant  rats  produced
teratogenic  effects.   These  effects were  not  noted when  DEHP or  dibutyl
phthalate were  administered orally  to  pregnant rats.   Additional  reproduc-
tive  effects  produced  include  impaired implantation,  parturition  and  de-
creased fertility  in  rats.   Testicular  damage has been  reported following
intraperitoneal  (i.p.)  or oral  administration of DEHP,  or  oral administra-
tion of dibutyl  phthalate.   No evidence of carcinogenic  effects produced, by"
                                                                             «
phthalates is available.
     Chronic toxicity includes  toxic polyneuritis  in workers exposed primar-
ily to dibutyl phthalate.  DEHP  animal  studies show  induced liver and kidney
changes while  dimethyl phthalate induced only kidney effects.  Following in-
jection dibutoxyethyl phthalate, di-(2-methoxyethyl) phthalate,  and  octyli-
sodecyl phthalate  have  caused  damage to the  developing  chick embryo nervous
system.
     Toxicity  of  the  phthalate  esters  to aquatic  organisms  varies within
this group  of  chemicals.  Freshwater organisms have appeared  somewhat  more
sensitive than  marine  species.   The data is insufficient to  allow  for  the
drafting of criteria to protect aquatic life  for any of the phthalates.
                                       -J

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                               PHTHALATE  ESTERS
I.   INTRODUCTION
     This profile  is  based primarily on the draft Ambient  Water Quality Cri-
teria Document for Phthalate Esters  (U.S. EPA,  1979).
     The  phthalate esters  are  esters of the  benzenedicarboxylic acid  ortho
form.  Esters of the  parent compound meta and para forms will  not be review-
ed in this profile.   The phthalate esters are  colorless  liquids of  low  vola-
tility, poorly  soluble  in  water  and soluble  in organic  solvents  and  oils.
Some physical and  chemical properties of the  phthalate esters  are  indicated
in Table.. 1.on the following page  (U.S. EPA,  1979).
     The phthalate  esters are widely used  as  placticizers, and  through this
application  are incorporated  into  wire and   cable  covering,  floor tiles,
swimming pool liners, upholstery and seat  covers,  footwear, and in  food and.^
medical  packaging  materials.    Non-plasticizer  uses   include  incorporation
into pesticide  carriers, cosmetics,  fragrances,  munitions, industrial  oils,
and insect repellants (U.S. Int. Trade Commission,  1978).   The most current
production figure is  6 xlO5 tons/year in  1977  (U.S. EPA,  1979).
     Phthalate  esters  are ubiquitous.   Monitoring   surveys   have   detected
phthalates in soil,  air, water,  animal and human tissues,  and  certain  vege-
tation.  Some plants  and animal, tissues  may synthesize phthalic acid esters
(Peakall, 1975).  From  in vitro  studies indications,  certain bacterial  flora
may be capable of metabolizing phthalates to the monoester  form (Englehardt,
at al.  1975).
II.  EXPOSURE
     Phthalate esters appear in  all areas of the environment.   Environmental
release of  the  phthalates  may  occur through  leaching of  plasticizers 'from
polyvinyl chloride  (PVC)  materials,  volatilization of phthalates  from  PVC

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materials, and the  incineration  of PVC items.   Human exposure  to phthalates



includes  contaminated  foods and fish,  dermal  application of phthalates  in



cosmetics and insect repellants, and  parenteral administration  by use of PVC



blood  bags, tubings, and infusion devices (U.S. EPA, 1979).



                                TABLE  1



            PHYSICAL AND CHEMICAL PROPERTIES OF FHTHALATE ESTERS

Phthalate
Compounds
Dimethyl
Diethyl
Oiallyl
Oiisobutyl
Dibutyl
Oimethoxyethyl
Oicyclohexyl
Butyl octyl
Oihexyl
Butylphthayl
butyl glycolate
Dibutoxyethyl
sthyl
Di-2-ethylhexyl
Diisooctyl
Di-n-octyl
Dinonyl
Molecular
Weight
194.18
222.23
246.27
278.30
278.34
282.00
330.00
334.00
334.00
336.37
366.00
391.00
391.00
391.00
419.00
Specific
Gravity
1.189 (25/25)
1.123 (25/4)
1.120 (20720)
1.040
1.047 (21)
1.171 (20)
1.200 (25/25)
—
0.990
1.097 (25/25)
1.063
0.985 (20/20)
0.981
0.978
0.965
Bp, Percent
OC Solubilitv in
H20, g/100 ml
282
296.1
290
327
340
190-210
. 220-228
340
—
219/5 mm
210
386.9/5 mm
239/5 mm
220/5 mm
413
0.5
Insoluble
0.01
Insoluble
0.45 (25°C)
0.85
Insoluble
—
Insoluble
0.012
0.03
Insoluble
Insoluble
Insoluble
Insoluble,
Source:  U.S.  EPA, 1979
                            rifts

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     Monitoring  studies have indicated  that most water phthalate  concentra-
tions  are  in the  ppm range, approximately  1-2 ug/1  (U.S.  EPA,  1979).   Air
levels of  pnthalates  in closed  PVC tiled rooms have been  reported  to be from
0.15  to  0.26  mg/m   (Peakall,  1975), while industrial monitoring has  mea-
sured  air  levels of  phthalates from 1.7 to  66 mg/m  (Milkov, et  al.  1973).
Phthalate  levels  in  various foods have  ranged  from non-detectable to  82 ppm
(Tomita, et  al.  1977).  Cheeses, milk,  fish and shellfish present potential
sources  of high  phthalate  intake  (U.S. EPA,  1979).   Estimates of  patient
parenteral  exposure   to  di-2-ethylhexyl phthalate  (DEHP)  during  use of  PVC
                           m
medical  appliances have indicated approximately  150'mg OEHP exposure  from  a
single hemodialysis   course.   Through application  of  certain cosmetics  and
insect repellants dermal exposure to  phthalates is  possible  (U.S. EPA,  1979).
     Using  average human  fish  and  shellfish consumption  data,  the U.S.  EHA
(1979) has derived  the following  bioconcentration  factors  for  the  edible
portions of fish  and shellfish consumed  by Americans  -  diethyl  phthalate,
270;  dibutylphthalate,  1500; OEHP,  95;  dimethyl  phthalate,  130.   OMP,.  DEP
and 8BP  are  based on the  steady-stata  bioconcentrations in bluegills  and in
fathead  minnows  for  OEHP.   A weighted average  bioconcsntration factor  of 26
was calculated  for dibutyl  phthalate utilizing  the  octanoi water partition
coefficient (U.S. EPA, 1979).
III. PHARMACOKINETICS
     A.  Absorption
         The phthalic acid  asters  and/or  their metabolites  are  readily  ab-
sorbed from the intestinal tract, the peritoneal cavity,  and the lungs  (U.S.
EPA, 1979).  Daniel  and Bratt  (1974) found  that seven days following  admin-
                                                                       »
istration of radiolabelled-OEHP, &2  percent  of the dose is  recovered in  the
urine and  57 percent recovered in  the  faces of  rats.   3iliarv excretion of
                                 /•ft.- 6

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orally administered  CEHP has been noted by  Wallin,  et al.   (1974).  Limited
human studies indicate  that  2 to 4.5 percent  of  orally administered OEHP  is
recovered in the urine  within 24 hours (Shaffer,  et al. 1945).  Lake,  et al.
(1975) suggest  orally administered  phthalates are  absorbed  after metabolic
conversion to the monoester form in the gut.
         Dermal absorption of OEHP  in rabbits has been  reported at 16 to  20
percent  of  the initial dose within  three  days  following administration
(Autian, 1973).
     B.  Distribution
         Studies in  rats  injected  with  radiolabelled-OEHP  have  shown that
from 60 to 70 percent of the  administered  dose was detected in the liver and
lungs within  2 hours after  injection (Daniel and  Bratt,  1974).  Wadell,  et
al.  (1977)  have reported  rapid  accumulation of radiolabelled-OEHP  in the
kidney  and  liver  of rats  after intravenous  (i.v.)  injection,  followed  by
                                                     *.
rapid excretion into the urine,  bile, and intestine.   Seven  days after i.v.
administration  of  radiolabelled-OEHP  to  mice, levels of  the  compound were
found preferentially in  the  lungs and to  a lesser extent  in  the brain, fat,
heart, and blood (Autian, 1973).
         An examination  of  tissue samples  from two deceased patients,  recip-
ients of  large volumes  of transfused  blood,  detected.  OEHP  in  the spleen,
liver, lungs, and  abdominal fat (Jaeger and  Rubin,  1970).  Daniel and Bratt
(1974) have suggested phthalates achieve a steady-state concentration, after
which the compounds or metabolites are rapidly eliminated by various routes.
         Injection of  radiolabelled-OEHP  and  diethyl phthalate  in pregnant
rats has shown the phthalates may cross the  placental barrier (Singh, et al.
1975).

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     C.  Metabolism
         Various  metabolites  of OEHP  have  been identified  following  oral
feeding of  rats (Albro, et al.  1973).   These results  indicate that OEHP  is
initially converted  from the diester to  the monoester, followed by  the  oxi-
dation of the  monoester'side chain forming  two  different alcohols.  The al-
cohols  are  then  oxidized to  the corresponding  carboxylic acid  or ketone.
Enzymatic clearance  of phthalates to  the monoester  form may  take place  in
the liver or  in the gut  (Lake,  et  al.  1977).  This enzymatic  conversion has
been observed' in stored  whole blood .indicating widespread  distribution ' of
                                                   a
this metabolic activity (Rock, et al. 1978).
     0.  Excretion
         Elimination  of orally  administratered  OEHP  is virtually completed
within four days in the rat  (Lake, et  al. 1975).   Major excretion  is through;
the urine and  faces, with  biliary  excretion  increasing the content of  DEHP
(or metabolites) in the  intestine (U.S.  EPA, 1979).   Schulz and Rubin  (1973)
have noted  a progressive increase in total  water soluble metabolites  in the
first  24  hours fallowing  injection  of  radiolabelled  OEHP  to  rats.    Within
one hour, eight percent of  the  OEHP was found  in the  liver,  intestine  and
urine.   After  24  hours, 54.6  percent  DEHP  was  recovered  in  the  intestinal
tract,  excreted feces and urine,  and only 20.5 percent OEHP was recovered  in
organic extractable form.
         The half-life  of phthalate elimination  from  the  tissues  and total
body is short (U.S. EPA, 1979).  Siphasic elimination  of OEHP from the blood
of rats  showed half-life  values of 9  minutes and 22-'minutes, respectively
(Schulz and Rubin, 1973).

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IV.  EFFECTS
     A.  Carcinogenicity
         Pertinent data could not be found in the available literature.
     8.  Mutagenicity
         Testing  of  several  phthalates  in  the Ames  Salmonella  assay has
shown  that diethyl  phthalate  has  some  mutagenic  activity  (Rubin,  et al.
1979).  Oibutyl,  mono-2-ethylhexyl,  di-(2-ethylhexyl) and butylbenzyl phtha-
late  all  produced  negative  effects  in  this  test  system.    Yagi,   et al.
(1978)  have  reported mutagenic  effects of mono-2-ethylhexyl  phthalate  in a
Bacillus subtillus recombinant assay system.
         Results  of a dominant  lethal assay  in  mice  have indicated DEHP and
dimethoxyethyl phthalate showed some mutagenic activity (Singh, et al. 1974).
     C.  Teratogenicity                                                      ;
         The teratogenic effects  of a number of  phthalate  esters (DEHP, di-
methyl, dimethoxyethyl, diethyl,  diisobutyl,  butylcarbobutoxymethyl,  and di-
octyl phthalates) have  been reported in  rats  (Singh,  et  al.  1972).  Terato-
genic effects were  not  seen following oral administration  of  DEHP and  dibu-
tyl phthalate to  rats (Nikonorow, et al. 1973).  Damage  to the nervous sys-
tem or  developing chick  embryos has been produced  by injection of dibutoxy-
ethyl  phthalate,  di-(2-methoxy-ethyl)  phthalate,  and  octyl-isodecyl  phtha-
late (Bower, et al.  1970).
     0.  Other Reproductive Effects
         Effects  on implantation  and parturition  have been  observed in preg-
nant rats  injected  intraparenteneally  with  OEHP, dibutyl phthalate,  and di-
methyl  phthalate  (Peters and  Cook,  1973).  A three generation rat reproduc-
                                                                       »
tion study has  indicated decreased fertility following maternal  OEHP  treat-
ment (Industrial Bio-Test,  1978).

-------
         Testicular  damage  has been  reported in  rats  administered OEHP  in-
traparenteneally or  orally.   Seth, at  al.  (1976)  found  degeneration of  the
seminiferous  tubules and changes in  spermatagonia; testicular  atrophy  and
morphological damage  was  noted in rats  fed OEHP  or dibutyl phthalate  (Car-
ter, et al. 1977).
     E.  Chronic Toxicity
         An increase  in toxic polyneuritis  has been reported  by Milkov,  et
al. (1973) in workers exposed  primarily  to dibutyl phthalate.  Lasser levels
   >
of  exposure  to  dioctyl,  diisooctyl,  benzylbutyl phthalates,  and  tricresyl
phosphate were also  noted.   Neurological symptoms have  been observed in  sev-
eral  phthalate  plasticizer  workers  (Gilioli,   1978).   Animal  studies   have
shown central nervous system degeneration and ancephalopathy  in rats admin-
istered  large  oral or  intiaperitoneal doses of  butylbenzyl phthalate  (Mai-
lette and Von Hamm, 1952).
         Oral OEHP feeding has produced  liver and kidney weight increases  in
           i
several  animal  studies (U.S.  EPfl,  1979).   Chronic exposure  to transfused
blood containing OEHP  has  produced  liver  damage  in monkeys  (Kevy,  et al.
1973).  Lake, at al.  (1975)  have produced  liver damage in  rats  by adminis-
tration of mono-2-ethylhexyl phthalate.
         Two-year  feeding  studies with  female  rats have shown  some kidney
effects produced by dimethyl phthalate (Draize, et al. 1948).
     F.   Other Relevant Information
         Several animal studies  have demonstrated- that OEHP pretreatment of
rats resulted  in increased  hexobaroital sleeping  times  (Daniel  and 9ratt,
1974;  Rubin and Jaeger,  1973; Swinyard, at al. 1976).
                                  _-^a/A/ —
                                 '  I / " 8"

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V.   AQUATIC TOXICITY
     A.  Acute Toxicity
         Acute  values  for  freshwater  fish were  derived  from eight  96-hour
bioassays for  four  phthalate esters.  LC5Q  values  ranged from 730 ug/1  for
di-n-butyl  phthalate  in  the  bluegill sunfish  (Lepomis macrochirus)  (Mayer
and  Sanders,  1973)  to  98,200 ug/1  in diethyl  phthalate  for the  bluegill,
Lepomis macrochirus.  Butylbenzyl  and dimethyl phthalates were  intermediate
in  their  toxicity in bluegill  assays with LC5Q  values of 43,300 to  49,500
ug/1 respectively (U.S.  EPA,  1978).   The scud, Gammarus pseudolimnaeus,  was
the most  sensitive  of  freshwater species tested, producing a  static  48-hour
adjusted  LC-Q  value  of  765 ug/1  (Mayer and  Sanders, 1973).   In  48-hour
static Daphnia  magna  assays,  the adjusted LC-Q  values for butylbenzyl,  di-
ethyl dimethyl, and di-n-ethylhexyl  phthalates  were 92,300,  52,100,  33,QOOn
and  11,100  ug/1,  respectively.   Among marine fish, juvenile  sheepshead min-
nows,  Cyorinodon  variegatus,  were most  susceptible   to  diethyl  phthalate,
producing a static  96-hour LC__ value of 29,600 jug/1.  In  similar  assays,
the  LC_Q  values  for  butylbenzyl and  dimethyl phthalate  were 445,000 ug/1
and  58,000  ug/1 respectively.   The  marine mysid  shrimp,  Mysidopsis  bahia,
was  tested  with  diethyl phthalate,  and produced  a   96-hour  LC_Q value  of
7,590 ug/1.   LC5Q  values of  9,630  and 73,700 ug/1 were reported  for  butyl-
benzyl and dimethyl phthalates,  respectively,  in  mysid  shrimp  assays.
     B.  Chronic Toxicity
         The only chronic studies available  are for  one  species of  fresh-
water  fish  and one  species  of  freshwater invertebrate  (Mehrle  and  Mayer,
1976; Mayer and Sanders,  1973).  A chronic value  of 4.2 ug/1  was  obtained  in
                                                                       »
a rainbow trout,  Salmo  gairdneri,  embryo-larval  study of di-(2-ethylhexyl)
                                 If I'll

-------
 phthalate.   In Oaphnia maqna significant reproductive  impairment was observ-
 ed  for di-2(-ethylhexyl)  phthalate  at  3.0  ug/1,  the lowest  concentration
 tested.   Chronic marine  data  was  not available.
     C.   Plant Effects
          In  the freshwater algae,  Selenastrum capricomutum, effective  con-
 centration  ranges  of 110  to  130 ug/1;  85,600 to 90,300  ug/1 and 39,300  to
 42,700 ug/1 were  obtained for butylbenzyl,  diethyl and  dimethyl  phthalates
 respectively.   Effective concentrations were  based  on  chlorophyll a content
 and cell  number.
     0.   Residues
          Bioconcentration  factors have been  obtained for five of  the phtha-
 lates.  In  the scud, bioconcentration factors of 1400  were reported  for  di-
 n-butyl  phthalate,  and  54,2680  for di-(2-ethylhexyl)  phthalate.   In  the.'
 bluegill, bioconcentration factors  of 57,  117, and 663 were obtained for  di-
 methyl, diethyl, and butylbenzyl phthalates,  respectively.  For  di-(2-ethyl-
 hexyi)  phthalate  bioconcentration  factors  were  reported  from 24 to  150  for
'the  sowbug,  Ascellus brevicaudus,  42 to 113  for the rainbow trout,  and  155
 to 386 for the fathead minnow.
 VI.  EXISTING  GUIDELINES AND  STANDARDS
     Neither  the  human  health  nor  the aquatic criteria  derived by  U.S.  EPA
 (1979), which  are  summarized  below, have gone through  the process of public
 review;  therefore,  there  is  a  possibility  that  these  criteria  will  be
 changed.
     A.  Human                                    '   ,
         3ased on  "no effect" levels observed in chronic feeding studies  of
 rats or dogs,  the  U.S.  EPA calculated acceptable daily  intake   (ADI)  levels
 for  several  phthalates,  and  established recommended water quality  criteria

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levels to  protect human  health for  dimethyl  phthalate,  diethyl phthalate,
dibutyl phthalate, and  OEHP.   These levels are  listed  in Table 2 (U.S. EPA,
1979)
     8.  Aquatic
         Data  are insufficient to  derive draft  criteria  for  any  of  the
phthalate  esters in  either  freshwater or  marine  environments  (U.S.  EPA,
1979).
                                    TABLE 2
                CALCULATED ALLOWABLE DAILY INTAKE IN WATER AND
              FISH FOR VARIOUS PHTHALATE ESTERS (U.S. EPA, 1979)

Ester NO Effect Species Days
Dose
(mg/kg/day)
Dimethyl
Diethyl
Dibutyl
Dicyclohexyl
Methyl phthayl
ethyl glycolate
Ethyl phthayl
ethyl glycolate
Butyl phthayl
ethyl glycolate
Oi-2-ethyhexyl
1000
625
18
14
750
250
140
60
Rat
Dog
Dog.
Dog
Rat
Rat
Dog
Dog
104
52
52
52
104
104
104
52
ADI**- F*** Recommended
(mg/day) Criteria
(mg/1)
700.
438
12.6
9.8
525
175
98
42
130
270
26
Not
Established
Not
Established
Not
Established
Not
Established
95
160
60
5




10
      **Allowable Daily Intake for 70 kg person (100 safety factor)
     ***F = Biomagnification factor
                                      XT

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                       PHTHALATE ESTERS

                          REFERENCES

Albro, P.W., et al.  1973.  Metabolism of diethhexyll phtha-
late by rats.  Isolation and characterization of the urinary
metabolies.  Jour. Chromatogr.  76: 321.

Autian, J.  1973. .Toxicity and health threats of phthalate
esters:  Review of the literature.  Environ. Health Perspect.
June 3.    ...

Bower, R.K., et al.  1970.  Teratogenic affects in the chick
embryo caused by esters of phthalic acid.  Jour. Pharmacol.
Exp. Therap.  171: 314.

Carter, B.R., et al.  1977.  Studies on dibutyl phthalate-
induced testicular atrophy in the rat:  Effect on zinc metabo-
lism.  Toxicol. Appl. Pharmacol.  41: 609.

Daniel, J.W., and H. Bratt.  1974.  The absorption, metabo-
lism and tissue distribution of di(2-ethylhexyl) phthalate
in rats.  Toxicology  2: 51.

Draize, J.H., et al.  1948.  Toxicological investigations
of compounds proposed for use as insect repellents.  Jour.
Pharmacol. Sxp. Ther.  93: 26.

Engelhardt, G. et al.  1975.  The microbial metabolism of  .
di-n-butyl phthalate and related dialkyl phthalates.  Bull.
Environ. Contain. Toxicol.  13: 342.

Gilioli, R. et ai.  1978.  A neurological electromyographic
and electroneurographic .study in subjects working at the
production of phthalate plasticizers:  Preliminary results.
Med. Law.  69^ 631.

Industrial Bio-Test.  1978.  Three generation reproduction
study with di-2-ethyl hexyl phthalate in-albino rats.  Plas-
tic Industry News, 24^,  201-203.

Jaeger, R.J., and R.J.  Rubin.  1970.   Plasticizers from
plastic devices:  Extraction, metabolism, and accumulation
by biological systems.   Science   170: 460.

Kevy, S.V., et al.  1978.  Toxicology of "plastic devices
having contact with blood.  Rep. N01 HB 5-2906, Natl. Heart,
Lung and Blood Inst. Bethesda, Md.

Lake, E.G., et al.  1975.  Studies on the hepatic effects  '
of orally administered di-(2-ethylhexyl)  phthaiate in the
rat.  Toxicol. Appl. Pharmacol.  32:  355.
                            JJl-lf  ,

-------
Lake, B.C., et al.  1977.  The in vitro hydrolysis of  some
phthalate diesters by hepatic and intestinal preparations
from various species.  Toxicol. Appl. Pharmacol.  39:   239.

Mallette, F.S., and E. Von Haam.  1952.  The toxicity  and
skin effects of compounds used in the rubber and plastics
industries.  II.  Plasticizers.  Arch. Ind. Hyg. Occup.
Med.  5: 231.

Mayer, F.L. Jr., and H.O. Sanders.  1973.  Toxicology  of
phthalic acid esters in aquatic organisms.  Environ. Health
Perspect.  3: 153.

Mehrle, P.M., and F.L. Mayer.  1976.  Di-2-ethylhexylphtha-
late:  Residue dynamics and biological effects  in rainbow
trout and fathead minnows.  Pages 519-524.  In,  Trace sub-
stances in. environmental health.  University of Missouri
Press, Columbia.

Milkov, L.E.., et al.  1973.  Health status of workers  ex-
posed to phthalate plasticizers in the manufacture of  artifi-
cial leather and films based on PVC resins.  Environ.  Health
Perspect. Jan.. 175.

Nikonorow, M., et al.  1973.  Effect of orally  administered
plasticizers and polyvinyl chloride stabilizers in the  rat.
Toxicol. Appl. Pharmacol.  26: 253.

Peakall, D.3.  1975.  Phthalate esters:  Occurrence and
biological effects.  Residue Rev. 54: 1.

Peters, J.W., and R.M. Cook.  1973.  Effects of phthalate
esters on reproduction of rats.  Environ. Health Perspect.
Jan. 91.

Rock, G. et al.  1978.  The accumulation of mono-2-ethyl
hexyl phthalate (MEHP) during storage of whole  blood and
plasm.  Transfusion 18_ 553.

Rubin, R.J., and R.J. Jaeger.  1973.  Some pharmacologic
and toxicologic effects of di-2-ethylhexyl phthalate (DEHP)
and other plasticizers.  Environ. Health Perspect. Jan. 53.

Rubin, R.J., et al.  1979.  Ames mutagenic assay of a  series
of phthalic acid esters: positive response of the dimethyl
and diethyl esters in TA 100.  Abstract. Soc. Toxicol. Annu.
Meet.. New Orleans, March 11.

Schulz, C.O., and R.J. Rubin.  1973.  Distribution, metabo-
lism and excretion of di-2-ethylhexyl phthalate in the  rat.
Environ. Health Perspect. Jan. 123..

-------
Seth, P.K., et al.  1976.  Biochemical changes  induced  by
di-2-ethylhexyl phthalate in rat liver.  Page 423  in  Environ-
mental biology.  Interprint Publications, New OelhIT  India.

Shaffer, C.B., et al.  1945.  Acute and subacute toxicity
of di(2-ethyhexyl) phthalate with note upon  its metabolism.
Jour. Ind.- Hyg. Toxicol.  27: 130.

Singh, A. et al.  .1972.  Teratogenicity of phthaiate  esters
in rats.  J. Pharm". Sci. 61, 51  (1972).

Singh, A.R., et al'.  1974.  Mutagenic and antifertility
sensitivities of mice to di-2-ethylhexyl phthalate  (DEEP)
and dimethoxyethyl phthalate (DMEP).  Toxicol. Appl.  Pharm-
acol.  29: 35.                                    .
                                                        L4
Singh, A.R., et al.  1975.  Maternal-fetal transfer of   C-
di-2-ethylhexyl phthalate and   C-diethyl phthalate in  rats.
Jour. Pharm. Sci.  64: 1347.

Swinyard, E.A., et al.  1976.  Nonspecific effect of  bis(2-
ethylhexvl) ohthalate on hexobarbital sleep  time.  J. Pharm.
Sci.  65l 733.

Toraita, I., et al.  1977.  Phthalic acid esters in various
foodstuffs anmd biological materials.  Ecotoxicology  and
Environmental Safety.  1: 275.

U.S. EPA.  1978.  In-depth studies on health and environ-
mental impacts of selected water pollutants.  U.S. Environ.
Prot.  .Agency, Contract No. 63-01-4646.

U.S. EPA.  1979.  Phthalate Esters:  Ambient Water Quality
Criteria Document.  (Draft).

U.S. International Trade Commission.  1978.  Synthetic  or-
ganic chemicals, U.S. production and sales.  Washington,
D.C.

Waddell, W.M., et al.  1977.  The distribution  in mice  of
intravenously administered   C-di-2-ethylhexyl phthalate
determined by whole-body autoradiography.  Toxicol. Appl.
Pharmacol.  39: 3"39.

Wallin, R.F., et al.  1974.  Di(2-ethylhexyl) phthalate
(DEHP) metabolism in animals and post-transfusion tissue
levels in man.  Bull. Parenteral Drug Assoc.'  28: 278.

'/agi, Y. at al.  1978.  Smbryotoxicity of phthalate esters
in mouse.  In: Proceedings of the First International Con- •
gress on Toxicology.  Plaa, G.  and Duncan. W. (ads.)

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                                      No. 147
         Phthalic Anhydride

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to  the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources,  this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented  by the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

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                            •  PHTHALIC ANHYDRIDE

                                    Summary



     Phthalic  anhydride failed  to produce  carcinogenic effects in  rats or

mice  in  a long  term National  Cancer Institute  (NCI)   feeding  study (7,500

ppm; 15,000.ppm).

     Information  on  the  mutagenic  effects  of phthalic  anhydride  was  not

found in the available  literature.

     The hydrolysis  product of phthalic anhydride,  phthalic  acid,  has shown

teratogenic effects in  the  developing  chick  embryo,  but not in any mammalian

tests.   Phthalic  anhydride inhalation  at  high  levels may  produce repro-

ductive impairment in male rats.
                                                                           4
     Chronic occupational  exposure to  phthalic anhydride  has been reported

to  produce  progressive  respiratory   damage  in  workers,  including  marked

fibrosis of the lungs.

     Data  concerning  the  effects of phthalic  anhydride to aquatic organisms

was not found in the available literature.

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                              PHTHALIC ANHYDRIDE
I.  INTRODUCTION
     This  profile is  based  on the  Preliminary  Environmental Hazard  Assess-
ment   of  Chlorinated   Naphthalenes,  Silicones,   Fluorocarbons,   Benzene-
poly carboxylates, and-Chlorophenols  (U.S. EPA, 1973).
     Phthalic  anhydride (molecular  weight  - 148.1) is  a white,  crystalline
solid  that melts  (sublimes) at  131°C,  has  a boiling  point of  284.5°C,  a
density  of  1.527,   and  a   solubility  of  0.62  gms/100 gms  water  at  25°C
(Towle,  et al.  1963).   This compound  is  soluble  in alcohol  and  sparingly
                                  «
soluble in ether.
     The  major uses  of phthalic anhydride are  in  the  synthesis of  plasti-
cizers,  alkyd  resins, unsaturated polyester  resins, -and  in the  preparation
of various classes of chemical dyes  (U.S. EPA, 1973).                        4'
     Production  of  phthalic anhydride in  1971 was  4  x  10  tons  (Blackford,
1970).
     Phthalic anhydride  is in  equilibrium with phthalic acid in aqueous  sys-
tems.  Under dry conditions, phthalic anhydride  is relatively  stable at  am-
bient  temperature (U.S.  EPA, 1973).   Elevated temperatures will produce  oxi-
dative degradation of phthalic anhydride.
     Phthalic anhydride  is biodegraded by microorganisms (Ribbons and Evans,
1960; Saegar and Tucker, 1973).
II.  EXPOSURE
     Phthalic anhydride  is used  in large  quantities and therefore has poten-
tial for  industrial  release  and environmental contamination.   NO monitoring
data are  available  to indicate ambient air or water levels of the compound.
Fawcett  (1970)  has  determined  40-200 pprn  by volume  in  phthalic anhydride

-------
off-gas process.   Phthalic  acid wastes have been  noted in waste waters  from
paint and  varnish  industries (Mirland and  Sporykhina,  1968)  and alkyd resin
plants (Minkovich, 1960).
     Human exposure to  phthalic anhydride from foods cannot be assessed, due
to a lack of monitoring data.
     Release of  phthalic  acid from parenterally-used plastic medical devices
(blood bags, plastic  tubings,, catheters, etc.) may  occur since these mater-
ials have been treated  with phthalate plasticizers;  however,  no data on  this
type of release are available (U.S. EPA,  1973).
     Bioaccumulation data on  phthalic. anhydride were not found in the avail-
able literature.
III. PHARMACOKINETICS
     Specific  information  on  the  metabolism,  distribution,  absorption, or
elimination of phthalic anhydride was not found in the  available literature.
IV.  EFFECTS
     A.  Carcinogenicity
          A long-term  carcinogenesis  bioassay in  rats  and mice fed phthalic
anhydride (7,500 ppm; 15,000 ppm)  has  been  conducted by the NCI (1979).  The
results indicate that  oral  administration  of  these levels of  the compound
produced no carcinogenic effects in either of the species used.
     8.  Mutagenicity
          Information on the mutagenic effects  of  phthalic anhydride was not
found in the available literature.
     C.  Teratogenicity
          Phthalic  acid was  shown  to  produce  an  increase  in  teratogenic
effects in the developing  chick embryo following  injection (Verrett, 'et al.

-------
1969).  Mammalian testing  of phthalic acid for teratogenicity failed  to  show
effects in mice (Koehler, et al. 1971).
     0.  Reproductive Effects
          Inhalation  exposure of  rats  to phthalic  anhydride  at high  levels
(100-200 mg/1)  has been  reported to  cause testicular  changes  and impaired
reproductive capability (Protsenko, 1970).
     E.  Chronic Toxicity      .  ...
          Markman and  Savinkina  (1964) have reported progressive respiratory
            .                                      r
damage  in  workers  exposed  to  phthalic  anhydride  for  two  years  or more.
Workers exposed for six years evidenced marked fibrosis  of the lungs.
     F.  Other Relevant Information
          Phthalic anhydride  has been implicated as a skin sensitizing agent
in some individuals  exposed for prolonged  periods of time  (Amer.  Ind.  Hygi
Assoc., 1967).
V.   AQUATIC TOXICITY
     Data concerning  the effects of  phthalic  anhydride  to aquatic organisms
were not found in the available literature.
VI.  EXISTING GUIDELINES
     The 3-hour,  TWA  occupational exposure  limit established  for phthalic
anhydride is 1 ppm (ACGIH, 1977).

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                              PHTHALIC ANHYDRIDE

                                  References
American  Industrial  Hygiene  Association.  1967.  Phthalic  anhydride:   Human
toxicity.  Amer. Ind. Hyg. Assoc. J.  28: 395.

ACGIH.   1977.   Threshold  limit values  for  chemical substances in  workroom
air.

Blackford,  J.L.   1970.    Isophthalic  acid.   Chemical  economics  handbook,
Stanford Research Institute.

Fawcett, R.L.   1970.  Air  pollution potential of  phthalic  anhydride.   J.  Air
Pollut. Contr. Assoc.  20: 461.

Koehler,  F.,  et  al.    1971.    Teratogenicity  of  thalidomide  metabolites.
Experientia.  27:. 1149.

Markman,  G.I.   and  R.A.   Savinkina.   1964.    The  condition of  the lungs  of
workers in phthalic anhydride production (an x-ray study).   Kemerovo.   35.

Mirland,  L.A.  and  V.A.  Sporykhina.   1968.   Polarographic determination  of
phthalic  acid  in  waste  waters   from   the  paint   and   varnish   industry.
Lakokrasch.. Mater. Ikh.  Primen.  1: 49.

Minkovich,  O.A.   1960.    The  recovery of phthalic  anhydride  wastes  in  the
manufacture of alkyd resins.   Lakokra. Mater,  i  ikh Primen.   1:  83.

NCI.   1979.   Bioassay of phthalic anhydride for possible  carcinogenicity.
NCI-CG-TR-159.

Protsenko,  E.I.   1970.   Gonadotropic action of  phthalic  anhydride.   Gig.
Sanit.  35: 105.

Ribbons, D.W. and  W.C. Evans.   1960.   Oxidative metabolism of  phthalic  acid
by soil pseudomonads.  Biochem. J.   76: 310.

Saegar,  V.W.  and  E.S.  Tucker.  1973.   Biodegradation  of phthalate  esters.
In:  Flexible vinyls  and human  safety:   An  objective analysis.   Conference
or the Society of Plastics Engineers,  Inc.,  March  20-22.   Kiamesha  Lake,  N.Y.

Towle, P.H.,  et al.  1968.  Phthalic  acids  and other benzenepolycarboxylic
acids.   Vol.  15,   p.  444.  In:  A Stauden (ed.)',  Encyclopedia of chemical
technology.  2nd ed.  J.  Wiley  & Sons, New York.

U.S. EPA.   1973.   Preliminary  hazard  assessment of chlorinated  naphthalenes,
silicones, fluorocarbons, benzenecarboxylates, and  chlorophenols.
                                                                      »
Verrett, M.J., et al.  1969.    Teratogenic effects of  captan and  related com-
pounds in the developing  chick  embryo.  Ann. N.Y.  Acad. Sci.  160:  334.
                                IV7-7

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                                       No.  148
             2-Picoline

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal..  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document  has undergone,  scrutiny to
ensure its technical accuracy.

-------
                             2-PICOLINE




                              Summary
     Pertinent data could not be found that defined 2-picoline as




a carcinogen or a omtagen.   Studies on cats ' indicated that the




structure and composition of the liver and the structure and growth




pattern of the skin were disrupted in the offspring of tested rats




who were given 157 mg per kg body weight daily during their pregnancy.




     2-picoline has been shown to produce biochemical and physical




changes in the liver, spleen, bone marrow, and lymph nodes.

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I.   INTRODUCTION
     2-picoline (alpha-picoline,  2-methylpyridine;  CAS No.
109-06-8) is a colorless liquid possessing a strong unpleasant
odor.  It has the following physical properties:
          Formula:                 CgHyN
          Molecular Weight:        93.12
          Melting Point:           -70°C
          Boiling Point:           129°C
         ..Vapor Pressure:          8 mm Hg at 20°C
          Vapor Density:           3.21
     2-picoline is freely soluble in water and miscible with alcohol
and ether (Windholz, 1976).  2-picoline is used as an organic
solvent and intermediate in the dye and resins industries.
II.  EXPOSURE"
     A.   Water
          Pertinent data could not be located in the available
li terature.
     B.   Food
          Pertinent data could not be located in the available
1i terature.
     C.   Inhalation
          2-picoline occurs in the working environment of coke oven
workers (Naizer and Mashek, 1974) and is present in cigarette smoke
(Brennemann et. al., 1979).
     D.   Dermal
                                                             »
          Pertinent data could not be located in the available
li terature.

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III. PHARMACOKINETICS


     A..   Absorption


          la racs, -2-picoline is rapidly absorbed and taken up by


the liver, heart, spleen, lungs, brain, and muscles during the


first .10 to 20 tninute-s after oral administration (Kupor, 1972).
   • •»' •*

     B.   Distribution


          Pertinent data could not be located in the available


literature.


     C.   Metabolism and -Excretion
          o

          Most of an administered dose in an acute toxicity study


was excreted in the urine within 48 hours (Kuper, 1972).


IV   EFFECTS


     A.   Carcinogenic!ty


          Pertinent data could not be found in the available


literature.


     B.   Mutagenicity


          Pertinent' data could not be found in the available


li terature.


     C.   Teratogenicity


          The structure and composition of the liver and the


structure and growth pattern of the skin were disrupted in the


offspring of treated rats who were administered 157 mg per kg body


weight of 2-picoline throughout their pregna-ncy (Hikiforova and


Taskaev, 1974).


     D.   Other Reproductive Effects


          Glycolytic processes and protein formation in the liver


was disturbed during the pregnancy of rats inhaling 2-picoline at
                                   -J'

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the maximum permissable concentration for 4 months.  The pregnancy




complicated toxicosis which without pregnancy was successfully




compensated by the liver (Taskaev, 1979).




     E.    Chronic Toxicity                           ,,




          The following biochemical and physical changes ha'v« been .




observed in rats after the administration of 2-picoline; changes




occurred in the liver carbohydrate metabolism (Taskaev, 1979; Kuper




and Gruzdeva, 1974) and changes occurred in protein synthesis of the




liver noted after chronic o.ral (Kuper and Gruzdeva, 1974) and




inhalation (Taskaev, 1979) exposure.   Administration of low doses




results  in changes in LDH isoenzyme distribution and activity




(Gruzdeva, 1976).  The major chronic  effects of 2-picoline are




injury to the liver (Ovchinnikova, 1978; Taskaev, 1979; Ovchinnikova,




1977) and spleen, bone marrow, and lymph nodes (Semchenko, 1973 and




1972).




     F.    Other Relevant Information




          Pertinent data could not be found in the available




li terature.



V.   AQUATIC TOXICITY




     A.    Acute Toxicity




          Pertinent information could not be found in the available




li terature.




     B.    Chronic Toxicity, Plant Effects and Residues




          Pertinent information could not be found in the available




1i terature.                                                  ,



     C.    Other Relevant Information




          Pertinent information could not be found in the available




li terature.
                             !<+
-------
VI.  EXISTING GUIDELINES AND STANDARDS




     A.   Human




          The 8-hour, time-weighted average occupational exposure




limit for alpha-picoline has been set in .Russia at 5 cag/m^




(Verschueren, 1977).  Maximum allowable concentration in Class I




waters for the production of drinking waters has been set in the




Netherlands at 0.05 mg/1 (Verschueren, 1977).




     3.   Aquatic




          Pertinent information could not be found in the available




literature.                  "

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                             REFERENCES
Brunnemann, K.D., and at. al., 1978.  Chemical Studies on Tobacco
Smoke: LXI. -Volatile Pyridines: Quantitative Analysis in Mainstream
and Sidestream Smoke of Cigarettes and Cigars.  Anal. Lett.  11(7):
545-560.

Gruzdeva. K.N., et. al.  1976.  Use of Electrophoretic Methods for
Determining Lactate Dehydrogenase Isoenzymes in Studying Chronic
Poisoning with Pyridine Derivatives.  Khromatogr. Elektroforeticheakia
Metody Issled. Biol. Aktiv.  Soedin.  44-7.

Kuper, V.G.  Distribution of alpha-Picoline in Rat Tissue During
Acute alpha-Picoline Intoxication.  Vop. Patokhimii  Biokhim.
Belkov.   Drugikh Biol. Aktiv.  Soedin.  51-2.

Kuper, V.G., and K.N. Gruzdeva.  1974.  Concerning the Question of
Carbohydrate Metabolism After Chronic Poisoning with alpha-Picoline.
Narusheniya Metab., Tr. Naucha.  Konf. Med. Inst. Zapadn.,  Sib.,
1st, 261-5.
Naizer ,  Y. ,  and
Its Homologs in
(5):76-78.
  V. Mashek..  1974.
  the Environment of
Determination of Pyridine and
Coke Plant Workers.  Gig. Sanit.
Nikiforova, A.A., and I.I. Taskaev.  1974.  Liver and Skin Morpho-
genesis in Some Laboratory Animal Embryos Following Poisoning with
Pyridine Bases.  Reakt. Plast. Epiteliya Soedin.  Tkani Norm, Eksp.
Patol.  Usloviyakh, Dokl. Mezhvuz, Gistol. Konf.  196-9.
Ovchinnikova,
in White Rats
2, 5-Lutidine.
L.S.  1977.  Morpholgical and Histochemical Changes
Liver After Acute Poisoning with alpha-Picoline and
 Gig. Aspekty Okhr. Zdorov'ya Naseleniya.  124-5.
Ovchinnikova, L.S., and Lambina.   1978.  Morphohistochemical
Changes in Liver of White Rats with Subacute Poisoning with Products
of Synthetic Rubber Production.  Deposited Doc., ISS. Viniti 2667-
78, 101-2.

Semchenko, V.V.  1972.  Regenerative Processes  in Blood-Forming
Organs of Experimental Animals During and After Chronic Intoxica-
tion by Methylpyridlne.  Mater.  Nauch. Sess.,  Posvyashch.  50-
Letiyu Obrazov.  SSSR, Omsk,  Cos. Med. Inst'.   896-8.

Semchenko, V.V.  1973.  Histological and Histo'chemical Characteristics
of Spleen and Lymph Nodes of Rats During Chronic Intoxication with
alpha-Picoline and 2,5-Lutidine.  Hezenkhima Tkanevya Proizvod. Evol.
Ontog., 56-8.

-------
                             REFERENCES
Taskaev, I.T.  1979.  Histological and Cytological  Changes  in  Rat
Liver During Experimental Poisoning and Subsequent  Pregnancy.
Arkh. Anat. .-'Gistol. Smbriol.  Vol. 76, ISS.  2,  49-54.

Verschueren, K.  1977.  Handbook of Environmental Data on Organic
Chemicals.  Van Nostrand Reinhold Company.  New  York.

Windholz, Martha et. al. (editors).  1976.  The  Merck  Index.   Merck
& Co., Inc.  Rahway, N.J.

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                                        No. 149
Polynuclear Aromatic Hydrocarbons (PAH)

    Health and Environmental Effects
  U.S. ENVIRONMENTAL PROTECTION AGENCY
         WASHINGTON, D.C.  20460

             APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                       SPECIAL NOTATION










U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated



polynuclear aromatic hydrocarbons and has found sufficient



evidence to indicate that this compound is carcinogenic.

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                    POLYNUCLEAR AROMATIC HYDROCARBONS'(PAH)
                                    SUMMARY
     The  first chemicals  ever shown  to be  involved  in  the development  of
cancer belong  to the polycyclic  aromatic hydrocarbon  (PAH)  class.   Several
PAH  are  well-known as  animal carcinogens  by all  routes  of administration.
Others are  not carcinogenic alone, but  in  certain cases  can enhance or  in-
hibit  the  tumorigenic  response  of carcinogenic  PAH.  Numerous  studies  of
workers  exposed to coal gas,  coal tars,  and coke oven emissions,  all  of
•which  have  large amounts  of PAH,  have demonstrated  a  positive association
between  their  exposures and lung cancer  development.   The carcinogenic  risk
of ingested PAH in humans, however, has not been extensively  studied.
     NO  standard toxicity  data  for  aquatic  organisms  are  available   for
freshwater  or  marine  life.   Limited information  concerning  toxic responses
                                                                            i
of freshwater  fish  reveals  that concentrations of  1,000 juq/l for six months
produced an 37% mortality in one warm water species.
                                  i

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 I.    INTRODUCTION



      This profile is based primarily upon  the  Ambient Water Quality Criteria



 Document  for  Polynuclear Aromatic  Hydrocarbons  (U.S.  EPA,  1979a) and  the



 Multi-media Health  Assessment Document  for  Polycyclic Organic  Matter (U.S.



 EPA,  1979b).



      Polycyclic aromatic hydrocarbons (PAH) are  a diverse class of compounds



 consisting of substituted and unsubstituted polycyclic  and heterocyclic aro-

•                                                                       *

 rnatic rings.   PAH are formed  as  a  result of  incomplete combustion of organic



 material  (e.g., fossil  fuels,  wood, etc.).  This leads to  formation  of C-H



 free  radicals which can polymerize  to form various PAH.  Among these PAH are.



 compounds  such  as  benzo(a)pyrene  (BaP) and  benz(a)anthracene  (8aA),  which



 are  ubiquitous  in  the  environment an.d well-known  for   their  carcinogenic



 activity.  The  presence  in ambient air of over  one hundred  individual  PAH



 has been  reported,  but quantitative data  on  only .26 PAH  are  available thus



 far.



      Most of the  PAH  are  high melting-point,   high  boiling-point  solids that



 are very insoluble in water.  As the ring  size increases,  the  volatility de-



 creases  significantly.   The PAH are strong  absorbers  of  ultraviolet  light,



 and  PAH  fluoresce  strongly;  both  of   these,  properties  lead  to  analytical



 methods  for  detection of trace  quantities.    Because  of  their high  melting



 points and low  water  solubilities  and   vapor pressures, most PAH  are  gener-



 ally  associated with particulate matter.   In air,  they  are adsorbed on small



 diameter particles that can be easily inhaled.   In  water,  PAH  appear  to also



 be  primarily  associated with  particulate  matter.   Based  upon water  treat-



 ability of PAH,  the compounds appear to exist  in equal proportions in three
                                                                        »


 forms;  bound to large suspended particles;  bound  to  finely dispersed  par-



 ticles,  and as  the dissolved form (U.S.  EPA, 1979a).

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     PAH adsorbed  to  airborne particulate matter  appear to be  fairly  stable
in  the environment.   Nevertheless, some  photooxidation  occurs with  atmos-
pheric  PAH  since quinone  derivatives  have been detected in the  atmosphere,
and their concentrations  increase during the summer when  the light  intensity
is greatest.
     Considerable  study on  the  microbial and  chemical  stability and  degra-
dation  of  PAH in  the  aquatic environment  has been  conducted.   In  general,
the  low .molecular  weight molecules appear  to  biodegrade relatively  rapidly
while  PAH  containing more  than  three rings  appear to  be extremely  stable.
The first step  in  the  microbial degradation process appears to  be the  forma-
tion of ortho-dihydrodiols which  rapidly  react to open  the ring.  PAH  also
appear  to be  light sensitive in aquatic systems, but the  rate of  degradation
is difficult  to  determine experimentally since the vast majority of  the  cars-
pounds  are  adsorbed to particulate matter.   Recent  studies have shown  that
adsorption of many PAH compounds  to  sediments is  a  major transport process
in aqueous  systems.  Studies in water treatment  of municipal and  industrial
sewage  indicate  that about two-thirds  of the PAH  can  be eliminated by sedi-
mentation  and biodegradation.   If this  secondary effluent  is  subjected  to
chemical  treatment (chlorination  or  ozonation)   the  remaining  PAH  can  be
degraded.
II.  EXPOSURE
     A.  Water
         3ased  upon work   by  Basu and Saxena  (1978)  the  average concentra-
tions of BaP, carcinogenic PAH  (BaP,  benzo(j)fluoranthene, indeno(l,2,3-cd)-
pyrene), and  total PAH (above 3  compounds  plus benzo(g,h.i)perylene, benzo-
                                                                       »
(b)fluoranthene, and fluoranthene)  in U.S. drinking water are 0.55 ng/1,  2.1
ng/1,  and  13.5  ng/1,  respectively.  NO drinking  water monitoring  data  on

-------
other PAH compounds  are available.   The low con-  centrations  are somewhat a-
reflection  of  the  extremely  low  water  solu-  bilities  of PAH  compounds.
Slightly higher drinking water values  have  been  reported in Europe (e.g. 3-5
ng/1 carcinogenic PAH  and  40-60  ng/1 total PAH),  but  these differences will
have  relatively negligible  effects on  the calculated  daily  intake  values
through  drinking  water  compared   to  other  sources  (U.S.   EPA,   1979a).
Assuming that  a human  consumes approximately 2  liters of water per day,, the
daily intake of PAH via drinking water would be:
         0.55 ng/1 x 2 liters/day = 0.0011 jug/day (BaP)
         2.1 ng/1 x 2 liters/day = 0.0042 jjg/day (carcinogenic PAH)
         13.5 ng/1 x 2 liters/day = 0.0270 jug/day (total PAH)
     8.  Food
         It is  difficult to evaluate the  human dietary intake  of PAH through
foods since the amount not only depends on  the food  habits of the individual
and the  style  of cooking,  but it also depends upon  the  origin  of the foods.
In order  to provide a  reasonably  accurate estimate of  the PAH  dietary 'in-
take,  average  concentrations  of  PAH in representative food  items  would have
to be available.  Unfortunately,  as of this date,  these data  have  not been
generated.  However, examination  of the available food  monitoring  data does
suggest that a  typical  range  of  concentrations  for PAH  and  BaP are 1.0-10.0
ppb and 0.1-1.0 ppb, respectively (U.S. EPA, 1979a).   Combining these ranges
with average total daily food consumption by man from all types of  foods of
1600 g/day, the following  estimates of dietary PAH  and  BaP intake  are  poss-
ible:
         0.1 - 1.0 ppb  x 1600 g/day  = 0.16 - 1.6jjg/day (BaP)
         1.0 - 10 ppb x 1600 g/day  = 1.6  - 16 jug/day"(PAH)
                                * / *} '/i -
                            111-TT

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         The  U.S. SPA  (1979a) has  estimated the  weighted average  bioconcen-
tration  factors for the  edible portion  of all aquatic  organisms consumed by
Americans.   These  range  from  120  to  24,000, and  are based  on the  octanol-
/water partition coefficients for each compound.
     C.  Ambient Air
         It  is not possible to determine  the average  intake of PAH  from in-
halation  of ambient  air  in the  United States  because the  monitoring  data
have focused  mostly on BaP concentrations.   However, by  making some  assump-
tions,  it  is  possible  to  provide estimates that  are  reasonably  close  to
probable  actual  values.    Using  the   1974-1975  Los  Angeles  monitoring  data
from Gordon (1976), the  relative amounts  to  carcinogenic PAH  and total PAH
compared to the average 3aP concentration  are presented below.
^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^^  »
                                        Carcinogenic         Total
                            BaP           PAH               . PAH
Ambient cone,  ng/m5    0.5-2.9             2.0              10.9
Inhalation  intake,
  micrograms/daya      O.C095-0.0435       0.038              0.207
aAssumed average air breathed per day was  19 m^
III. PHARMACQKINETICS
     There  are no  data available  concerning  the  pharmacokinetics  of PAH  in
humans.  Nevertheless, it  is possible to make  limited assumptions based  on
the results  of animal  studies conducted with several PAH, particularly  BaP.
The metabolism of PAH in human and animal tissues has  been  especially  well-
studied,  and   has  contributed  significantly  to  an .understanding   of  the
mechanisms of  PAH-induced cancer.
                                     / 9 7 6 -"

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     A.  Absorption
         Regardless  of  the  route  of exposure,  it  can be  demonstrated  in
laboratory animals  that PAH are readily  absorbed across all epithelia  which
are  in contact with  the external  environment  (Rees,  et al.  1971; Kotin,  et
al.  1969;  Vainio,  et  al.  1976).    The   fact  that  PAH are  generallly  high
lipid-soluble  neutral  molecules  greatly  facilitates  their  passage  through
the predominantly lipid-like cell membranes of  animals,  including  man.
     •B.  Distribution
         Upon  reaching the bloodstream,   PAH  are  rapidly distributed to  most
internal body  organs  (Kotin,  et al. 1969; Bock  and Dao,  1961;  Oao,  et  al.
1959;   Flesher,   1967).   Under  experimental   conditions  with   laboratory
animals, the  route  of exposure has  little apparent  influence  on the  tissue
localization  of  PAH.   Extensive  localization  in  the fat  and  fatty  tissues
(e.g.,  breast)  is observed  (Bock  and Dao, 1961;  Schlede,  et  al. 1970  a,b)
and suggests that these tissues may act  as a chemical  trap, creating a  situ-
ation  for  sustained release of the unchanged  substance.   In pregnant  rats,
it is  apparent that BaP and 7,12-dimethylbenz(a)anthracene, but probably  not
3-methylcholanthrene,  are capable of transplacental passage and localization
in the  fetus (Shendrikova and Aleksandrov, 1976).
     C.  Metabolism
         PAH  are   metabolized by   the   microsomal  mixed-function  oxidase
system, also  known  as  aryl  hydrocarbon  hydroxylase.   This enzyme system  is
readily  inducible  and  is  found  in  most mammalian  tissues,  although pre-
dominantly  in the  liver.   In conjunction  with  various  P-450  type  cyto-
chromes, this  enzyme  complex  is  involved in  detoxification of  many  xeno-


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biotics,  but may  also  catalyze  the  formation  of  reactive  epoxide metab-
olites, themselves  leading to  carcinogenesis.   A second  microsomal  enzyme,
epoxide hydrase,  converts epoxide metabolites  of PAH  to  vicinal glycols,  a
process  which  may  also  be  of  critical  importance  in  the  process   of
carcinogenesis.
         Because  of  the  importance of  metabolic  activation  for  the  ex-
pression  of  carcinogenic effects by  PAH,  the  chemical fate  of  many reore-
sentative  compounds in  mammalian  cells has been  extensively  explored  (U.S.
EPA, 'i979a).  By far the most widely  studied of the  PAH has been BaP, one of
the   principal   carcinogenic   products  from   the   combustion  of   organic
material.   The metabolites  of BaP   (and  all  PAH)   can  be  divided  into  a
water-soluble  and  an  organic solvent-soluble  fraction.   Components  of  the
latter  fraction  are  primarily ring-hydroxylated  products,   quinones,   and
Labile  spoxide intermediates,  for  BaP  there  are  at least  three dihydro-
diols,  three quinones,  and four phenols which  can be detected as positional
isomers.   The  K-region  (4,5-)  and  non-K-region   (7,3-; 9,10-)  epoxides  are
precursors  of   the  corresponding  vicinal  diols,  which  are  formed  by  the
action of  the epoxide  hydrase enzyme.   A subsequent  oxidative attack  by aryl
hydrocarbon  hydroxylase  may  convert the  non-X-region diols to  vicinal diol
epoxides,  one  of which  (7,8-diol-9,LQ-epoxide) is  an ultimate carcinogenic
form of BaP.
         In  the water-soluble fraction  containing BaP metabolites are mainly
conjugates of  hydroxylated products  with  glutathione, glucuronic  acid,  and
sulfate.  This group of  metabolies is  tentatively regarded to be composed  of
non-toxic excretion oroducts.

-------
         The  general  scheme  of  metabolism  for  unsubstituted  PAH  closely
parallels that  for  BaP,  although several other  major environmental PAH have
not been  studied.   It is  also evident that K-region derivatives of PAH may
be  preferred  targets  for  conjugation and  excretion,  whereas  non-K-region
epoxides   undergo   further   reductions   and   oxidative   attack   to   form
toxicologically  important  molecules.   For  PAH  bearing  alkyl  substituents
(e.g.,  DMBA,   MCA),   the  .primary  metabolites  formed  are  hydroxymethyl
derivatives.    Nevertheless,   epoxidation   reactions    at    K-region   and
non-K-region  aromatic  double  bonds  occur  which   are   catalyzed  by  aryl
hydrocarbon  hydroxylase.   Removal  of  activated  intermediates occurs  by
conjugation with glutathione  or glucuronic  acid, or by further metabolism to
tetrahydrotetrols.
     0.  Excretion                                                          4
         Over forty  years ago, researchers  recognized that various PAH were
excreted primarily  through the hepatobiliary system  and  the feces (Peacock,
1936;  Chalmers  and  Kirby,  1940).   However,  the  rate  of  disappearance  of
various PAH from  the  body,  and the principal routes  of excretion are influ-
enced  both  by  structure  of  the  parent  compound  and the  route  of adminis-
tration (Heidelberger  and Weiss,  1959; Aitio, 1974a,b).   Moreover,  the rate
of  disappearance  of a  PAH (i.e.,  benzo(a)pyrene)  from body  tissues  can be
stimulated markedly  by prior  treatment  with inducers of  microsomal enzymes
(e.g., benzo(a)pyrene,  7,12-dimethylbenz(a)anthracene, 3-methylcholanthrene,
chrysene) (Schlede,  et  al.  1970a,b).  Likewise,  it has been  shown  that in-
hibitors of microsomal  enzyme activity,  such as  parathion and paraoxon, can
decrease the rate of BaP  metabolism  in certain animal tissues (Weber,  et al.
                                                                       »
1976).  From  the  available data  concerning  excretion of  PAH  in  animals,  it
is apparent extensive bioaccumulation is not likely to occur.


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IV.  EFFECTS
     A.  Carcinogenicity
         PAH were  the  first compounds ever  shown  to be associated with  car-
cinogenesis.  As  of this date,  carcinogenic PAH  are  still distinguished  by
several unique  features:  (1)  several  of the PAH are among the most  potent
carcinogens  known  to   exist,   producing  tumors  by  single  exposures   to
microgram  quantities;  (2) they  act both  at the  site  of  application  and  at
organs distant  to the site  of absorption;  and  (3) their  effects  have  been
demonstrated in nearly  every  tissue and  species  tested,  regardless  of  the
route of  administration  (U.S. EPA,  1979a).   Among the more  common  PAH,  at
least one,  BaP, is ubiquitous  in the environment.  In animals, PAH produce
tumors  which resemble   human  carcinomas.   The  demonstration  that  organic
extracts  of particulate air pollutants  are  carcinogenic  to  animals  has
raised  concern  over  the  involvement  of  PAH   in human   cancer  formation
(Hoffmann and Wynder, 1976).
         Oral administration of PAH to  rodents  can result in tumors  of  the
fore-stomach, mammary  gland, ovary,  lung,  liver,  and iymphoid  and hemato-
poietic tissues (U.S.  EPA,   1979a).  Exoosure to  very  small doses  of  PAH  by
inhalation  or intratracheal instillation  can also be  an  effective means  of
producing  tumors  of the respiratory tract.   However,   for  both oral and  in-
tratracheal  routes  of  administration,  BaP is less  effective  than  other  PAH
(e.g., OM8A, MCA)  in  producing carcinomas.   However,  BaP  has a remarkable
potency for  the induction of skin tumors  in mice-'that cannot  be matched  by
any other  environmental  PAH.  Therefore,  caution  must be  exercised  in con-
sidering the carcinogenicity of  PAH  as  a class,  or in  using  BaP  as a  reore-
                                                                        »
sentative example in evaluating the carcinogenic risk of PAH.

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          The  presence  of PAH in the air, or as components  of soot,  tars,  and
 oils,  have  long been associated with an  excess  incidence of cancer  in  human
 populations  (U.S. EPA,  1979a,b).   However,  it  has  never  been possible  to
 study  a  population having exposure to PAH  in  the absence of other  potential
 carcinogens,  cocarcinqgens,  tumor initiators,  or  tumor promoters.
          Convincing  evidence from air pollution  studies indicates  an  excess
 of  lung cancer  mortality  among  workers exposed  to large  amounts of  PAH-
 containing  materials  such  as  coal gas,  tars,  soot,  and coke-oven  emissions
 (Kennaway,  1925;  Kennaway  and  Kennaway,  1936,  1947;  Henry,   et  al.  1931;
 Kuroda,  1937;  Reid  and  Buck,  1956; Doll,  1952;  Doll, et  al. 1965,  1972;
 Redmond,  et  al.  1972,  1976;  Mazumdar,   et  al.  1975;  Hammond,  et  al.  1976;
 Kawai, et al.  1967).  However,  no definite proof exists that  the  PAH present
 in  these materials  are  responsible  for the cancers  observed.   Nevertheless,
 our understanding of  the characteristics of  PAH-induced tumors in  animals,
 and their  close  resemblance to human carcinomas  of the same  target organs,
 suggests PAH pose a carcinogenic threat to man, regardless  of the  route  of
• exposure.
      8.  Mutagenicity
          The  demonstration  of mutagenicity  in bacterial and mammalian  cells
 by  exposure to PAH is generally  equated  with  the capability to induce  tumor
 formation.   This assumption is based on the participation of  a common  elec-
 trophilic metabolite in producing the  carcinogenic/mutagenic event, and  the
 common target site in the cell (i.e., ONA or  other components  of  the genome)
 for the effect  to be produced.
          In  recent years, considerable  research  effort has been directed  at
 determining the mutagenicity of  various PAH derivatives as a means  of  ident-
 ifying  structural features  associated  with the  biological effect  produced.

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Working  with bacterial  mutants  which  can  be  reverted  to  histidine  inde-
pendence by  a chemically-induced mutation, epoxides of  carcinogenic  PAH were
shown  to possess  significant mutagenicity (U.S.  EPA,   1979a).   Further work
with cultured mammalian  cells established that  carcinogenic PAH can produce
forward  mutations  wheri  a.  drug  metabolizing  enzyme  system  is  available
(Huberman and Sachs, 1974,  1976).
      -  .Numerous  attempts  have been made to  correlate exposure to  PAH with
the induction of  chromosomal aberrations.  Although variations  in  chromosome
number  and  structure" accompany  PAH-induced  tumors  in rodents,  it is  not
clear   whether   these  changes   are  consistently   observable  (U.S.   EPA,
1979a,b).   NO evidence  in  the  published literature  has  been  found to  in-
dicate  that  PAH  may produce  somatic mutations in  the  absence  of  neoplastic
transformation.                                                       '       ^
     C.  Teratogenicity
         PAH  are  not generally  regarded to  have  significant  teratogenic
activity.   3aP  showed no effect  on the  developing  embryo  in  several  mam-
malian and non-mammalian  species (Rigdon and Rennels, 196
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     E.  Chronic Toxicity
         Little  attention  has been  paid to the  non-carcinogenic affects  of
exposure to PAH.  Nevertheless,  it  is  known that tissues of the  rapidly  pro-
liferating  type  (e.g.,  intestinal epithelium,  bone marrow, lymphoid  organs,
testis) seem  to  be the preferred targets  for  PAH-induced cytotoxicity  (U.S.
EPA,  1979).  This  action  is  probably  due to a specific, attack  on- DMA  of
cells in the S phase of the mitotic  cycle  (Philips,, et  al.  1972).
         Acute and  chronic exposure to various carcinogenic PAH  has resulted
in  selective destruction of hematopoietic  and lymphoid elements, ovotoxicity
and anti-spermatogenic  effects,, adrenal necrosis,  and  changes in  the intes-
tinal and  respiratory  epithelia (U.S. EPA,  1979a)..  For the most  part,  how-
ever, tissue damage occurs at dose levels  that would also  be  expected to in-
duce  carcinomas,, and thus the  threat of   malignancy predominates  in  evalu-
ating  PAH  toxicity*  For  the- non-carcinogenic  PAH,, there  is a shortage  of
available data concerning-  their  involvement in toxic responses.•
V.   AQUATIC TOXICITY'
     A.  Acute Toxicity
     Standard  toxicity  determinations  for freshwater  or  marine   organisms
have   not   been  conducted   for  any   PAH.   The   marine  worm,.  Neanther
arenaceodenta,  was  exposed,  to  crude  oil extracts,  and  LC_Q' values   for
various PAH ranged  from 300 to l.OOO^g/l  (Neff,  et al. 1976a,b).  A  90  per-
cent lethality, -determined from photodynamic response,  was obtained  for  the
protozoa,  Paramecium caudatum at an  for anthracene, concentration of 0-1 /ug/1
in   one-hour  exposures   (Epstein,   1963).    Bluegill   sunfish   (Leoomis
macrochirus)  displayed  an 87%  mortality  at   concentration  of  1,000  yug/1
benzo-a-anthracene.                                                     *
                                    > ^n.^-
                                 ^  } / J J~
                                      A

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     a.  Chronic
         Standard  toxicity  studies using  either  freshwater or marine organ-
isms  have, not  been  conducted on  any  PAH.   A  six-month  study  of benzo-
(a)pyrene on  the bluegill sunfish  (Leoomis macrochiras)  produced 87 percent
mortality at a concentration of 1,000 ug/1 (Brown, et al. 1975).
     C.  Plants
         Studies of  the effects of PAH  on freshwater or marine plants could
                                                        »
not be located in the available literature.
     D.  Residues
         In  short-term modeling  of freshwater ecosystem  studies,  three-day
bioconcentration factors  for benzo(a)pyrene  of  930, 5,258,  11,536, 82,231,
and  134,248  were  obtained  for  the  mosquito-fish  (Gamousia  affinis),  the
algae Oedoqonimi eardiacum,  the mosquito Culex pipiens  quinquefasciatus, the
snail  Physa  sp.,  and  cladoceran  Daonia puiex,  respectively  (Lu, et  all
1977).  For anthracene, a 1-hour  bioconcentration factor of 200 was obtained
for  Daphnia  maqna  (Herbes,  1976).  For  marine molluscs,  bioconcentration
factor values  ranged from 8.2  for the  clam  (Rangia cuneata)  (Neff,  et al..
1976a) to 242 for  the eastern  oyster  (Crassostrea virginica) (Couch, et al.,
in press).
VI.  EXISTING GUIDELINES AND STANDARDS
     Neither  the  human health  and  aquatic  criteria  derived  by  U.S.  EPA
(1979a),  which  are summarized below,  have not gone through the  process  of
public review;  therefore, there  is a possibility-'that  these criteria may be
changed.
                                     )*•

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     A.  Human
        'TO  date,  one  recommended  standard  for  PAH  as a  class  has  been
developed.  The World Health Organization (1970)  recommends  a concentration
of PAH  in water not. to exceed  0.2/jg/l.   This recommended standard is based  .
on the  composite analysis of  six PAH  in drinking  water:  (1) fluoranthene,
(2)  benzo(a)pyrene,  (3)  benzo(g,h,i)perylene, (4) benzo(b)-fluoranthene,  (5)
benzo(k)fluoranthene, and  (6) indeno(l,2,3-cd)pyrene.
          In  the occupational environment,  a Federal  standard has been pro-
mulgated  for  coke  oven  emissions, based primarily on the presumed effects of
the  carcinogenic  PAH  contained  in  the  mixture  as measured  by  the  benzene
soluble fraction  of total particulate  matter.   Similarly,  the American Con-
ference of Governmental Industrial  Hygiensists  recommends a  workplace expo-
sure limit for coal tar pitch  volatiles,  based .on the benzene-soluble frac-
                                                                            *
tion containing carcinogenic PAH.  The  National  Institute for Occupational
Safety  and Health has  also  recommended  a  workplace criterion for coal  tar
products  (coal tar,  creosote,  and coal tar  pitch),  based on  measurements of
the  cyclohexane extractable  fraction.   These  criteria are summarized below:
Substance              Exposure  Limit             Agency
Coke Oven Emissions     0.150 mg/m3, 8-hr.         U.S. Occupational Safety
                        time-weighted average      and Health Administration
Coal Tar  Products       0.1  mg/rn^, 10-hr.          U.S. National Institute  for
                        time-weighted average      Occupational Safety  and
                                                   Health
Coal Tar  Pitch          0.2  mg/rn^, (benzene        American Conference  of
Volatiles               soluble.fraction) 8-hr.    Governmental Industrial
                        time-weighted average      Hygienists
          Based on  animal  bioassay data,  and  using the  "one-hit"  model,   the
U.S.  EPA   (1979a)  has  set  draft  ambient  water quality criteria for  BaP   and
                                                                       »
dibenz(a,h)anthracene (DBA)  which will result in  specified   risk  levels  of
human cancer as shown in the-table below.
                                      -17

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Exposure Assumptions
     (per day)
2 liters of drinking water
and consumption of 18.7
grams fish and shellfish.

Consumption of fish and
shellfish only.
Exposure-Assumptions
	(per day)
2 liters of drinking water
and consumption of 18.7
grams fish and shellfish.

Consumption of fish and
shellfish only.
              SaP

Risk Levels and Corresponding Draft Criteria
                  ng/1
0

0
10-7


0.275




1.25
10-6

 2.75



12.5
10-5

 27.5



125
                                             DBA
Risk Levels and Corresponding-Draft Criteria
0

0
10-7

0.43



1.96
10-6


 4.3




19.6
10-5

 43




196
     B.  Aquatic
                                                      •

         Criteria  have  not been  proposed  for  the  protection of  aquatic

organisms (U.S. EPA, 1979a).
                                   Ugt ft  f
                                  '/  /  <3 y

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                       POLYNUCLEAR AROMATIC HYDROCARBONS
                                  REFERENCES

Aitio, A.  '1974a  Different  elimination and effect on mixed function oxidase
of  20-methylcholanthrene  after  intragastric and  intraperitoneal  adminis-
tration.  Res. Commun. Chem. Path.. Phamacol.  9: 701.

Aitio,  A.   1974b.   Effect  of chrysene  and  carbon  tetrachloride  adminis-
tration  on  rat hepatic microsomal  monoxygenase  and udglucuronsyltransferase
activity.  FEBS Lett.  42: 46.


3asu O.K. and  J.  Saxena.   1978.   Polynuclear aromatic hydrocarbons in selec-
ted  U.S.  drinking  waters  and  their  raw  water sources.   Environ.  Sci.
Technol.  12:  75.

Bird,  C.C.,  et  al.   1970.   Protection  from  the  embryopathic  effects  of
7-hydroxymethyl-12-methylbenz(a) anthracene       by      2-methyl-l,2-bis-(3
pyridyl)-l-propanone(metapirone      ciba)     and     B-diethyl-aminoethyl-
diphenyl-n-propyl acetate  (SKR 525-A).  Br. Jour. Cancer 24: 548.

Bock, F.G.,  and T.L. Oao.   1961.   Factors affecting  the  polynuclear hydro-
carbon level in rat mammary glands.  Cancer Res. 21: 1024.                 ^

Brown, E.R., et al.   1975.-  Tumors  in fish caught in polluted waters:  poss-
ible  explanations.   Comparative Leukemia  Res.  1973,  Laukemogenesis.   Univ.
Tokyo Press/Karger, Basel, pp. 47-57.

Chalmers,. J.G.,  and A.H.M. Kirby.   1940.   The  elimination of 3,4-benzpyrene
from  the  animal  body  after  subcutaneous  injection.    I.  Unchanged  benz-
pyrene..  Biochem. Jour. 34: 1191.

Couch,  J.A.  et al.   The American oyster  as an indicator  of  carcinogens in
the  aquatic environment.   Inj   Pathobiology  of  environmental  pollutants -
animal models  and wildlife "or monitors.   Storrs,  Conn.   National Academy of
Sciences.  (In press).

Currie,  A.R.,   at  al.   1970.   Embryopathic   effects   of  7,12-dimethyl-
benz(a)anthracene  and its  hydroxymethyl  derivatives  in  the  Sprague-Oawley
rat.  Nature 226: 911.

Oao.  T.L.,   et al.   1959.   Level  of 3-methylcholanthrene  in  mammary glands
of  rats after intraaastric  instillation  of carcinogen.  Proc.   Soc.  Exptl.
Biol. Med.   102: 635.

Doll. R.   1952.   The causes  of  death among gas workers  with  special refer-
ence to cancer of the lung.  Br. Jour. Ind. Med.   9: 180.
                                                                      »
Doll. R., et  al.   1965.  Mortality of  gas workers with  special  reference to
cancers  of  the  lung  and  bladder,  chronic  bronchitis,  and  pneumoconiosis.
Br. Jour. Ind. Med. 22: 1.

Doll R.  at al.   1972.  Mortality of  gas  workers  - final  report  of  a  pros-
pective study.  Br. Jour. Ind. Med.   29: 394.

Epstein, S.S.,  et  al.   1967.  The  photodynamic  effect  of the carcinogen,
3,4-benzoryene, on Paramecium caudatum.  Cancer Res.  23:  35.

-------
Flesher,  J.S.   1967.   Distribution of  radioactivity in  the  tissues of  rats
after  oral administration  of  7,12-dimethyl-benz(a)anthracene~5H.  Biochem.
Pharmacol.  16: 1821.

Gordon,  R.J.    1976.   Distribution  of  airborne  polycyclic  aromatic   hydro-
carbons throughout Los Angeles.   Environ. Sci. Technol.  10:  370.

Hammond,  E.G.,  et al.   1976.  Inhalation  of benzpyrene  and cancer in  man.
Ann. N.Y. Acad. Sci.  271: 116..

Heidelberger, C., 'and S..M. Weiss.  1959.   The distribution  of  radioactivity
in  mice  following  administration  of  3,4-benzpyrene-5Ci4  and  1,2,5,6-di-
benzanthracene-9, 10-C14.  Cancer Res. 11: 885.

Henry, S.A.  et al.   1931.  The incidence  of cancer of the bladder and pros-
tate in certain occupations.  Jour. Hyg.  31:  125.

Herbes,  S.E.   1976.   Transport and  bioaccumulation of  polycyclic aromatic
hydrocarbons  (PAH)   in   aquatic   systems.    In:    Coal   technology  program
quarterly  progress report  for the  period  ending  December  31,  1975.  Oak
Ridge National Lab.,  Oak Ridge, TN.  ORNL-5120.  pp. 65-71.

Hoffmann  0. and E.L.  Wynder.   1976.   Re.spiratory  carcinogenesis.  In:   chem-
ical  carcinogens  C.E.  Searle  (ed.)   ACS  Monograph 173,  Amer.  Chem.   SocV
Washington, O.C-                                                            4

Huberman,  E.,  and L.  Sachs.   1974.   Cell-mediated  mutagenesis  of mammalian
cells with chemical carcinogens.  Int. Jour. Cancer  13: 32.

Huberman, £., and  L.  Sachs.  1976,  Mutability of different  genetic loci in
mammalian  cells by  metabolically activated  carcinogenic  polycyclic  hydro-
carbons.  Proc.  Natl. Acad. Sci.  73: 188.

Kawai, M.,  at al.   1967.   Epidemiologic study of  occupational  lung cancer.
Arch. Environ. Health 14: 859.

Kennaway,  E.L.    1925.    The   anatomical distribution  of the  occupational
cancers.  Jour.  Ind. Hyg.  7:  69.

Kennaway, E.L., and  N.M.  Kennaway.   1947.   A  further  study of the incidence
of cancer of the lung and larynx.   3r. Jour. Cancer.  1: 260.

Kennaway, N.M., and E.L.  Kennaway.   1936.   A  study  of  the incidence of can-
cer of the lung and larynx.  Jour. Hyg. 36:  236.

Kotin,  P.,  et   al.    1969.   Distribution,  retention,   and  elimination  of
C^4-3,4  benzpyrene after  administration  to  mice  and  rats.   Jour.  Natl.
Cancer Inst.  23: 541.

Kuroda, S.   1937.   Qccuoational pulmonary  cancer of generator  gas  workers.
Ind. Med. Surg.   6: 304.

Lu, P.  at ai.   1977.   The environmental  fate of three  carcinogens;  benzo-
(a)-oyrene, benzidine, and vinyl chloride evaluated  in  laboratory  model eco-
systems.  Arch.  Environ.  Contam.  Toxicoi.  6: 125.

Mazumdar, S., et  al.   1975.  An  eoidemiolcgicai  study of exposure to  coal
tar pitch volatiles among coke oven workers.   APCA Jour. 25: 382.
                                    f-JL*

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Neff,  J.M.,  et  al.   1976a.  Effects  of petroleum  on survival,  respiration
and  growth  of marine  animals.   In:   Sources,  Effects  and  Sinks of  Hydro-
carbons  in  the  Aquatic Environment.   Proceedings of  a symposium,  American
University, Washington, O.C., American Institute of Biological Sciences,   p.
520.

Neff,  J.M.,  et  al.   1976b.  Accumulation and  release of  petroleum-derived
aromatic  hydrocarbons  by  four  species  of  marine   animals.    Mar.   Siol.
38: 279.

Peacock,  P.R.f  1936.   Evidence regarding  the-"mechanism  of elimination  of
1,2-benzpyrene,  1,2,5,6-dibenzanthracene,  and  anthracene  from  the  blood-
stream of injected animals.  Br. Jour. Exptl. Path.  17: 164.

Philips,  E.F.,  et  al.   1972.   In  vivo  cytotoxicity of  polycyclic  hydro-
carbons.,  Vol.  2.  p.  75 In:   Pharmacology  and the-Future of Man. Proc.  5th
Intl. Congr. Pharmacology,~T972, San Francisco..

Redmond,  C.K.,. et al.   1972.   Long  term mortaility  study of  steelworkers.
Jour. Qccup. Med.  14: 621.

Redmond,  C.K.,  et  al.   1976.   Cancer  experience  among  coke- by-product
workers..  Ann. N.Y. Acad. Sci.  p. 102.

Rees, E.O., et al.   1971.   A  study of the mechanism of  intestinal absorption
of benzo(a)pyrene..  aiochem. Biophys.  Act.  225: 96.

Reid, O.O., and  C.  Buck,   1956..  Cancer in  coking plant workers.  Br. .Jour.
Ind. Med.  13: 265.

Rigdon, R.H.,  and J. Neal.   1965.   Effects of feeding benzo(a)pyrene on fer-
tility, embryos, and young  mice.  Jour.  Natl. Cancer Inst.   34: 297.

Rigdon, R.H.,  and  E.G.  Rennels.  1964.  Effect  of feeding benzpyrene on  re-
production in  the rat.  Experientia 20:  1291.

Schlede, E., et  al.  1970a. Stimulatory effect of benzo(a)pyrene and pheno-
barbital pretreatment on  the biliary  excretion  of benzo(a)pyrene  metabolites
in.the rat.  Cancer Res. 30: 2898.

Schlede, E. et al.   1970b.  Effect  of enzyme induction on  the metabolism  and
tissue distribution of benzo(a)pyrene..   Cancer Res. 30:2893.

Shendrikova,  I.A., and  V.A. Aleksandrov.  1974.  Comparative characteristics
of  penetration  of  polycyclic   hydrocarbons  through  the  placenta into   the
fetus in rats.  Byull.. Eksperiment. Biol. i Medit.  -77: 169.

U.S. EPA.  1979a.  Polynuclear  Aromatic  Hydrocarbons:   Ambient Water Quality
Criteria (Draft).

U.S.  EPA.   1979b.   Multi-media  Health  Assessment  of  Polycyclic  Organic
Matter.  (Draft) prepared under contract to U.S.  EPA  by J.  Santodonato,  et
al., Syracuse Research Corp.
                                , *-r 0.0,
                              *) '  U )

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Vainio,  H.-,  et  al.   1976.   The  fate of  intratracheally  installed  benzo-
(a)pyrene in the  isolated perfused rat  lung  of both  control  and 20-methyl-
cholanthrene pretreated rats.  Res. Commun. Chem. Path. Pharmacol.  13: 259.

Weber, R.P., et  al.  1976.   Effect of the  organophosphate insecticide para-
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World  Health  Organization.   1970.   European  standards  for drinking  water.
2nd ed.  Geneva.
                                      -   a-

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                                      No.  150
              Pyridine

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts  presented by  the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

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                                   PYRIDINE
                                    Summary

     Pyridine'has  not shown carcinogenic  effects following repeated subcuta-
neous administration  to ,rats;-, the compound  did not show  mutagenic activity
in the Ames Salmonella assay.
     A single  study has  indicated that pyridine produced developmental  ab-
normalities when administered to chicken embryos.
     Chronic exposure  to.  pyridine produces CNS disturbances and may  produce
adverse hepatic and renal effects.
     Pyridine has  been shown to  be  toxic to  freshwater fish at  concentra-
tions ranging from 100,000 to 1,580,000 /jg/1.   For  freshwater  invertebrates,
toxic concentrations of pyridine range  from 575,000  to  2,470,000 >ig/l.
                                    r-^ffl ^
                                 '   I > ) J*
                                   I So-3

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 I.    INTRODUCTION
      Pyridine  (CAS  number  110-86-1)  is  a  colorless  liquid  possessing  a
 sharp, penetrating odor.  It has the  following physical  properties:
           Formula:                         C-H-N
           Molecular Weight:                79.1
           Melting Point:                   -42°C
           Boiling Point:                   115.3°c
           Density:                         0.982
           Vapor Pressure:                  10 mm Hg at 13.2°C
                                             (Sax, 1975)
           Solubility:                      misqible with water, alcohol,
                                             ether, and other organic
                                             solvents (Windholz, 1976)
     Pyridine is a weak  base and  forms salts with strong  acids.   It is used
as a  solvent  for anhydrous mineral salts,  in  various organic synthetic pre-
parations, and in  analytical chemistry (Windholz, 1976).   The  estimated an-
nual  production  of  pyridine is in excess  of  60  million pounds (Federal Reg-
 ister 43:16638, April 19, 1978).
 II.  EXPOSURE
     A.   Water
          Pertinent data could not be located in the available literature.
     B.   Food
          Reported levels of  pyridine in  foods  include:  from 0.02 to 0.12
ppm,  ice cream;  0.4 ppm,  baked goods; 1.0  ppm, non-alcoholic  beverages; 0.4
ppm,  candy.  Pyridine  has also been  found to occur naturally  in  coffee and
tobacco (Furia, 1975).
     C.   Inhalation
          Pyridine may  be  produced   and  released during  the  combustion  of
                                                                         »
coke and as a combustion product in cigarette smoke (Graedel, 1978).


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           The major release of pyridine is  from  emissions  from manufacturing
 and chemical  processes.  Based  on  total  annual  production,  the  U.S.  EPA
 (1976)  has  estimated a significant  potential emission of pyridine  during
 manufacture.  •
     0.    Dermal
           Pertinent data could not be located  in  the  available literature.
 III.. PHARMACOKINETICS
     A.    Absorption
           Absorption  of  pyridine occurs  through  the respiratory and gastro-
intestinal tracts,  but probably not through  the skin  (Gosselin,  et al.  1976).
     B.    Distribution
           Pertinent data could not be located  in  the  available  literature.
     C.    Metabolism  and Excretion
           Pyridine  may be  partly excreted unchanged  or  may be methylated at
 the N-position  (Patty,.  1963) and excreted as  N-methyl pyridinium hydroxide,
 its chief metabolite (Browning,. 1965).-   Methylation  occurs in  mice but not
 in  rats,  and it may  occur to  some  extent in man   The  fate of the majority
 of absorbed pyridine  is  not known (Browning, 1965).
 IV.  EFFECTS
     A.    Carcinogenicity
           Subcutaneous injection of  pyridine   at  levels of  3  to  100  mg/kg
 twice weekly  for a  year  did not produce tumors  in rats (Mason, et al. 1971).
     B.    Mutagenicity
           Pyridine  did   not  show mutagenic  effects'with  activation in  the
 Ames Salmonella assay (Commoner, 1976).
     C.    Teratogenicity
           Pyridine  caused  chick embryo  abnormalities in  one  limited  study
 (Federal Register 43:16688, April 197 1978).

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     0.   Other Reproductive Effects
          Pertinent data could not be  located  in  the  available literature.
     E.   Chronic Toxicity
        .  Prolonged  daily  exposure to pyridine  at levels  from  6 to 12 ppm
causes mild  central  nervous system (CNS)  disturbances in workers, while ex-
posure from  15 to 330 ppm causes  insomnia,  nervousness, and  low-back or ab-
dominal pain accompanied by frequent urination (Gosselin, et al.  1976).
          In animals,  the  major effects of  repeated  feeding of pyridine are
hepatic and  renal injury  (Patty,  1963).   Chronic  exposure  to 10  or 50 ppm
pyridine  vapors  causes  increased  liver/body weight  ratios  in  rats  (ILQ,
1971).
     F.   Other Relevant Information
          Symptoms  in humans  associated  with  inhalation  or ingestion  of
pyridine are CNS  depression,  arid  liver and kidney damage  (Federal  Register
4:16688,   April 19,  1978;   Gosselin,   et  al.  1976;  Sax,  1975;  ILO, 1971).
Vapors are also irritating  to  eyes,  skin, and nasal  membranes (ACGIH,  1977;
Sax, 1975),  Skin eruptions  induced  by pyridine  may  be provoked by  exposure
to  light  (Arena,  1974).   Ingestion of pyridine causes CNS depression,  heart
and  gastrointestinal  distress,  fever,  and, at  high  doses, death;  and may
stimulate bone marrow production  of  platelets  in low doses  (ACGIH,   1977;
Gosselin,  et al. 1976).  Death may be  due to either hepatic or renal damage,
or  from pulmonary injury  (Gosselin, et al. 1976;  ACGIH, 1977).
          Exposure to vapors of  pyridine  from  1,250 to 10,000 ppm for 1 to 7
hours did not  cause  mortality in  rats, but a 0.1 percent  diet  of pyridine
induced rapid weight loss and death in two weeks  (ILO, 1971).

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 V.   AQUATIC TOXICITY
      A.   Acute Toxicity
           McKee  and Wolf  (1963)  have  reviewed the  effects of  pyridine on
 several aquatic organisms.  The freshwater  minnow,  bleak (Alburnus lucidus),
 was the most sensitive sp.ecies tested  with  threshold toxicities ranging from
 100,000 to  160,000 pg/1.  Tests with  the  freshwater  mosquitofish (Gambusia
 affinis)  revealed  a  96-hour  LC5Q value  of  1,300,000  ug  of pyridine  per
 liter of turbid water.  Orange-spotted  sunfish (Lepomis  humilis)  were killed
 in  one hour  from exposure  to  pyridine  at  concentrations  ranging  from
 1,480,000  to 1,580,000 ug/1,  while goldfish (Carassius  auratus)  were killed
 after 10 to 30 hours' exposure to  pyridine.  Verschueren (1979) has reported
 a 24-hour  LC5Q value  of 1,350,000 ^ig/1 for  mosquitofish exposed to  pyri-
 dine.
           Oowden  and  Bennett  (1965)  demonstrated  a  48-hour  LC5Q value  of
 2,114,000  ;ug/l for Daphnia magna exposed to pyridine.  McKee and  Wolf (1963)
 reported a threshold effect of 40,000 ug/1  for Daphnia sp.  Canton and Adema
 (1978)  determined  48-hour LC5Q .values ranging  from 1,130,000 to 1,755,000
'ug/1. for  Daphnia magna,  and  48-hour  LC5Q  values  of 575,000  and 2,470,000
 ug/1 for Daphnia pulex and Daphnia  cucullata,  respectively.
      B.   Chronic Toxicity,  Plant Effects and Residues
           Pertinent data  could not  be  located in the available literature.
      C.   Other Relevant  Information
           Thomas (1973) reports that  pyridine exposure  levels  of  5,000 jjg/1
 impart  an  off-flavor to fish flesh.
                                     >~rGrr
                                  * I /  ' /

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VI.  EXISTING GUIDELINES AND STANDARDS
     A.   Human
          The  8-hour,  time-weighted-average occupational  exposure limit for
pyridine recommended  by the American  Conference of  Governmental Industrial
Hygienists is 5 ppm (ACGIH, 1977).
     8.   Aquatic
          Based on  96-hour LC5Q data,  Hahn  and Jensen  (1974)  have assigned
pyridine an aquatic toxicity rating of from 100,000 to 1,000,000 pg/1.


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                                   REFERENCES
 American Conference of Governmental Industrial Hygienists.   1977.   Threshold
 limit values  for chemical  substances and  physical agents  in the  workroom
 environment with intended changes for  1977,   Cincinnati,  Ohio.

 Arena,   J.M.   1974.   Poisoning:  Toxicology-Symptoms-Treatments.   3rd  ed.,
 Charles  C.  Thomas:  Springfield,  Illinois.

 Browning,  E.  1965..  Toxicity and Metabolism- of  Industrial. Solvents.   Ameri-
 can Elsevier f • New York.

 Canton,  H.J.,  and D.D.M.  Adema.  1978.   Reproducibility  of short-term  and
 reproduction toxicity experiments with Daphnia  maqna and  comparison of  the
 sensitivities  of Daphnia magna  with  Daphnia pulex  and  Daphnia cucullata  in
 short-term  experiments,  Hydrobioligia.  59:  13T!

 Commoner,.  B.   1976.   Reliability of bacterial  mutagenesis  techniques  to
 distinguish  carcinogenic  and non-carcinogenic  chemicals.   U.S.  EPA,   NTIS
 PB-259 934.

 Dowden,  B.  and H. Bennett.   1965.  Toxicity of selected chemicals  to certain
 animals.  Jour.  Wat.  Poll. Cont.  Fed.  37: 1308.

 Furia, T.   1975.  Fenaroli's  Handbook of Flavor  Ingredients.   2nd ed.   CRC
 Press,. Boca Raton, Fla.

 Gosselin,  R.E.,  et  ai.   1976.  Clinical Toxicology  of  Commercial Products.
 4th ed.  Williams and Wilkins, Baltimore.

 Graedel,  T..  1978.    Chemical  Compounds  in the Atmosphere.   Academic Press,
 New York.

 Hahn, R.  and P.   Jensen.   1974.   Texas A and M University, College Station,
 Texas.   Water  Quality Characteristics  of Hazardous  Materials.   Prepared  for
 the  National Oceanic and  Atmospheric Administration.   NOAA-78013001.   NTIS
PB-285 946/OST.

 International Labour  Office.   1971.  Encyclopedia of Occupational Safety and
Health, Vol. 2.  McGraw-Hill Book Co.,  New York.

Mason, M., et al.   1971..  Toxicology and  carcinogenesis  of various chemicals
used in the preparation of vaccines.   Clin. Toxicol.  4:  185.

McKee, J.E.  and  H.W.  Wolf.   1963.  Water  Quality Criteria.   The  Resources
Agency of California.  State Water Quality Control Boar.d Publication 3-A.

Patty, F..   1963.   Industrial  Hygiene  and Toxicology: Volume  2,  Toxicology.
2nd ed.  John Wiley and Sons, New York.

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Sax,  N.I.    1975.   Dangerous  Properties of  Industrial Materials.   4th ed.
Van Nostrand Reinhold Co., New York.

Thomas, N.A.   1973.   Assessment  of fish tainting substances.  In: Biological
Methods for  the Assessment of Water Quality.  American  Society  for Testing
and Materials, ASTM-STP-528, p. 178.

U.S.  EPA.   1976.   Preliminary  scoring  of  selected organic  air  pollutants.
U.S. Environ.. Prot.. Agency, EPA 450/3-77-008a.

Verschueren, K.  1979.  Handbook  of Environmental Data on Organic Chemicals.
Van Nostrand Reinhold Co., New York.

Windholz,  M. (ed.)  1976.  Merck  Index.   9th  ed-  Merck and Co.,  Rahway, New
Jersey.

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                                      No. 151
              Qulnones

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460


           APRIL 30, 1980
               LJ^ ~  f
             To v f

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                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document has undergone  scrutiny to
ensure its technical accuracy.

-------
                           QUINONE




                           Summary
     Quinone has been reported to produce neoplasms,  but insufficient



data are available to assess its carcenogenic potential.  Quinone




was not mutagenic to Orosophila melanogaster, human leukocytes,



nor Neurospora.



     Quinone is very toxic to fish and plants.  Exposure to humans



causes conjunctiyal irritation and, in some cases,  corneal edema,



ulceration, and scarring; transient eye irritation  was  noted



above 0.1 ppm.   Quinoae is highly toxic to mammals  via  the oral




and inhalation route.
                                -7***

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I.   INTRODUCTION

     Quinone (p-Benzoquinone;  Gas No. 106-51-4) is a yellow,

crystalline•solid with chlorine-like irritating odor.  It has the

following physical properties:

          Formula:
          Physical State:          large, yellow, monoclinic
                                     prisms

          Molecular Weight:        108.09

          Specific Gravity:        1.318 (20*C)

          Melting Point:           112.9"C

          Boiling Point:           sublimes

          Vapor Pressure:          considerable; sublimes readily
                                   upon gentle heating (Patty,1967)

     Quinone is soluble in. alcohol, ether, and alkali; and slightly

soluble in hot water.  Quinone can be prepared by oxidation starting

with aniline or by the reduction of hydroquinone with broaic acid.    .

The compound has found wide application in the dye,  textile, chemical,

tanning, photography, and cosmetic industries primarily because of

its ability to transform certin nitrogen-containing  compounds into

a variety of colored substances (Patty, 1967).

II.  EXPOSURE

     A.   Water

          Pertinent data could not be located in the availabe

literature.

     B.   Food

          Pertinent data could not be located in the available

literature.

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     C.   Inhalation




          Because of its ability to sublime, quinone becomes an air




contaminant problem at Che production sice.




     D.   Dermal




          Pertinent data could not be located in the available literature..




ILL PHARMACOKINETICS




     A.   Absorption



          Quinone is readily abs.orbed from the gastroenteric trace



and subcutaneous tissues (Patty, 1967).   Sax, 1979, reports quinine



as capable of causing death or permanent injury due Co Che exposures



of normal use via absorption through oral and inhalation routes.




Quinone affects the eyes (Procter, 1978).



     B.   Distribution




          Pertinent data could not be located in the available literature.




     C.   Mebalolism. and Excretion



          Quinone is partially excreted  unchanged; but the bulk is




eliminated in conjugation with hexuronic, sulfuric, and other acids




(Patty, 1967).



IV.  EFFECTS



     A.   Carcinogenicity




          Quinone has been reported to produce neoplasms but upon



review by the International Agency for Research on Cancer, it was



determined that there was insufficient data to conclude that it-was



a carcinogen (IARC, 1977)                   '' ,



     B.   Mutagenlcity



          Quinone did not produce mutagenic effects in studies with

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Orosophila melanogaster and human leukocytes (Lueers and Obe, 1972).




Another study reported quinone as nonmutagenic to Neurospora



(Reissig, 19'63).




     C.   Teratogenicity




          Pertinent data could not be located in the available




literature.




     D.   Other Reproductive Effects




          Pertinent data could not be located in the available literature.




     E»   Chronic Toxicity




          Quinone has been reported to oxidize with the lens protein




SH groups in rabbits (Ikemota and Augusteyn, 1976).  Chronic exposure




causes Che gradual development of changes characterized as follows:




brownish discoloration of the conjunctiva and cornea confined to




the intrapalpebral fissure; small opacities of the cornea; and




structural cornea! changes which result in loss of visual acuity



(Sterner, et al., 1947; Anderson and Oglesby, 1938).




     F.   Other Relevant Information




          Acute exposure causes conjunctival irritation and, in



some cases, corneal edema, ulceration, and scarring; transient eye



irritation may be noted above 0.1 ppm and becomes marked at 1 to 2




ppm (AIHA, 1963).  Ulceration of the cornea has resulted from one



brief exposure to a high concentration of the vapor of quinone, as



well as from repeated exposures to moderately .high concentrations



(Patty, 1967).  Absorption of large doses of quinone from the gas



troenteric tract or from subcutaneous tissues of animals induces



chronic convulsions, respiratory difficulties, drop in blood pres-



sure, and death by paralysis of the medullary centers (Patty, 1967).

-------
     Oral rat LDSOs have been reported for quinone ranging from



130 to 296 mg per kg body weight (Verschueren,  1977).   Inhalation     '




of quinone at concentrations ranging from 230 to  270 ng per cu.m.




for 2 hrs was lethal to 100 percent  of the test  population of




rats.



IV..  AQUATIC TOXICITY




     A.   Acute Toxicity




          Quinone has been reported to be toxic to invertebrate



Daphnia at 0.4 ppm (Verschueren, 1977).  Also,  quinone has an LD50




for perch ranging from 5 to 10 mg/1 (Verschueren,  1977).



     B.   Chronic Toxicity, Plant Effects, and  Residues



          Quinone inhibits photosynthesis in the  fresh water algae




S. capricornutum (Gidding, 1979), decreases chlorophyll flourescence



and cyclosis (protoplasmic streaming)  of  Nitella  cells (Apartsin,



et al, 1979; Stom,. 1977; Stom and Kuzevania, 1976; Stom and Rogozina,



1976), and inhibits carbon metabolism in  Ghloralla pyrenoidosa



(Printavu, 1975).




VI.  EXISTING GUIDELINES AND STANDARDS



     A.   Human



          The 8-hour, time-weighted average occupational  exposure




limit for quinone has been set in the  United States at a  concentration



of 0.1 ppm and in the U.S.S.R. at a concentration  of 0.01  ppm



(Verschueren, 1977).
                             /sf-7

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                             REFERENCES
Procter, M.H., and James Hughes.  Chemical Hazards of  the Workplace.
J.3. Lippincott Company.  Philadelphia.  1978.

Reissig, J.L.  1963.  Induction of Foward Mutants in the Pyr-3
Region of Neorospora.  J. Gen. Microbial.  30:317-325.

Sterner,. J.H.t at al.  1947.  Quinone Vapors and Their Harmful
Effects to Corneal and Conjunctival Injury.  J. Ind. Hyg.
Toxicol.  29:60.

Stom, D.I.  1977. 'Influence of Polyphenols and Quinones on Aquatic
Plants and Their Blocking of Sulfhydryl Groups.  Acta  Hydrochim.
Hydrobiol.  -Vol. 5, ISS. 3, 291-8.

Stom, D.I, and E.N. Kuzevanova.  1976.  The Distribution of
Sulfhydryl Groups in Nitella Cells and the Effects on  Them
of Polyphenols and p-Benzoquinone.  Tsitologiya.  Vol. 18,
ISS. 2, 230-2.

Stom, D.I., and N.A. Rogozina.  1976.  Possible Mechanism of
Action of Quinone Pesticides on the Photoplasmic Streaming in
Marine Plants.  Eksp. Vodn. Toksikol.  Vol. 6, 111-118.

Verschueran, K.  1977.  Handbook of Environmental Data on Organic
Chemicals.  Van Nostrand Reinhold Co.  New York.

AIHA.  1963  Hygenic Guide Series: Quinone.  Am. Ind.  Hyg. Assoc. J.
24:194.  1963.

Anderson, 3., and F. Oglesby.  1938.  Corneal Changes  from Quinone-
Hydroquinone Exposure.  A.M.A.  Arch. Opthalmol.  59:495.

Apartsin, M.S., et al.  1979.  Mechamism of the Effect of
Pyrocatechol and p-Benzoquinone on Nitella cells.  Dokl. Akad. Nauk.
S.S.S.R.  Vol. 244, ISS. 2, 510-12.

Giddings, J. M.  1979.  Acute Toxicity to Selenastrum  capricornutum
of Aromatic Compounds from Coal Conversion.  Bull. Environ. Coneam.
Toxicol.  Vol. 23, ISS.  3, 360-4.

IARC.  1977.  Monographs on the Evaluation of Carcinogenic Risk of
Chemicals to Man, Vol. 15.  World Health Organization.

Ikemoto, F., and R.C. Augusteyn.  1976.  The Reaction  of Lens
Proteins and Amino Acids with 1,4-Benzoquinone.  Jpn.  J. Ophtnalmol.
(Japan).  Vol. 20, ISS.  4, 457-65.

Lueers, H., and G. Obe.   1972.  On the Possible Mutagenic Activity
of p-Benzoquinone.  Mutut. Res.  15:77-80.

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Patty,  F.A.   1967.   Industrial Hygiene and  Toxicology.  Inter-
science Publishers.  New York.

Pristavu,  N.   1975.  Action of p-Benoquinone  on  the Radioactive
Carbon  Metabolism  in Chlorella pyrenoidosa.   Proc. Int. Congr
Photpsynth. ,'  3rd.  Vol. 2.  1541-6.
                             IF/'?

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                                      No. 152
             Resorclnol

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a. survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal*  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and   environmental impacts,  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

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                                   RESORCINOL
                                    Summary

      Resorcinol,  1,3-dihydroxybenzene,  is a  phenolic compound.   Resorcinol
 is weakly  antiseptic and resorcinol  compounds are  used  in  Pharmaceuticals
 and hair dyes for human use.  Major industrial uses  are  as adhesives in rub-
 ber products and tires,  wood adhesive  resins, and as ultraviolet  absorbers
 in polyolefin plastics.  Resorcinol  is also  a byproduct of coal, conversion
 and is a component of cigarette smoke.  Thus,,  substantial opportunity exists
 for human exposure.
      Many phenolic  compounds, including resorcinol, are  strong  mitotic spin-
 dle poisons in plants.  This, evidence of mutagenic  activity  and the  strong
 oncogenic activity  in plants  have not  been adequately  tested  in animals  to
 provide  an understanding of  the processes.    In animals  the only cocarcino-
 genic activity (in cigarette smoke condensate)  demonstrated has  been- as  a
 protective  agent against benzo(a)pyrene  carcinogenicity.'
      Resorcinol has been  demonstrated  to result in chronic toxicity:   reduc-
 ing  growth  rate  in  an insect species  and causing  chronic health complaints
 from  workers in a tire manufacturing plant.
      Acute  toxicity through oral,  eye,  skin  penetration,  and skin irritation
has  been  demonstrated by all  tests.   Values vary  in  the literature and are
 inadequate  to draw  a quantitative conclusion.   Resorcinol  has also been
shown to  be acutely toxic to  both freshwater and marine aquatic organisms in
96-hour LC5Q tests.

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     No standards  or guidelines exist for resorcinol.   ACGIH's Committee on
Threshold  Limits has  proposed  a  TLV of 5  ppm but has not  finalized that
recommendation.  Industry has suggested this  value  is lower than is required
for safety, citing  existing workplace levels of 9.6  ppm without worker com-
plaint or evidence of acute or chronic toxicity..

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 I.    INTRODUCTION
      Resorcinol  is  a  phenolic compound  (molecular  weight,  110.1;  boiling
 point,  276^0;  melting  point, 110.0°C).   Synonyms  are  m-dihydroxybenzene,
 1,3-benzenediol,  3-hydroxyphenol,  and resorcin.   Resorcinol occurs as  white
 or  nearly white  needle-shaped crystals  or powder.   It has a faint,  charac-
 teristic  odor and a  sweetish  taste with  a  bitter aftertaste.   One gram  is
 soluble in 1  ml of water and in 0.1 ml of  alcohol.                   -  ...
     Resorcinol is a  weak  antiseptic and is used in  antiseptics,  keratolytic
 disease treatments  and fungicides (Wilson, et  al.  1577).  Major  uses of re-
 sorcinol  are:  in tires and other rubber  products;  wood adhesive resins;  as
 an  ultraviolet absorber in  polyolefin plastics;  as an  intermediate  in  dye
 manufacture (especially hair dyes);  and in the production of synthetic  tan-
 ning  agents,  explosives, and  specialty  adhesives.   The  tire  and rubber  in-
 dustries  accounted for  43  percent  of the use of resorcinol in 1974, primar-
 ily  as  adhesives in  fabricating  belting,  rubberized hose,  and rubberized
 textile sheets  (Stanford Research Institute, 1975).
     Resorcinol is expected  to be a. component  of various waste streams  from
 coal  conversion  facilities.   The  potential  for removal  through  existing
 waste treatment  processes is  currently under  assessment  (Herbes and Beau-
 champ', 1977).
 II.  EXPOSURE
     Resorcinol is used in substantial quantities in industry and frequently
 in small  quantities in  the home.   Although the  potential for human exposure
 exists,  very  little exposure information is  available.  The Koppers Company,
 Inc., Monroeville, Pennsylvania, is  the  major supplier of resorcinol  in the
                                                                       »
United States.  They report substantial testing  of the plant environment in-
dicating  resorcinol concentration up  to  9.6 ppm  in  ambient air  (Flickinger,
 197$).                           .

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      Resorcinol is currently  sold and transported  as a solid,  although the
 Koppers Company  reports  increasing  inquiries  regarding bulk  shipments  of
 molten resorcinol.  They  indicate that  this  would increase  the  opportunity
 for industrial and public  exposure to the compound (Flickinger,  1976).
      In an epidemiological study  of  rubber workers at a hexamethylenetetra-
•mine-resorcinol (HR)  resin system tire manufacturing plant, all environment-
 al  samples in  the  study were less  than  1  mg/m3 (Gamble, et al. 1976)..
      Resorcinol has been  shown  to be  present in  cigarette  smoke  and  is  a
 component  of the weakly acidic  fraction  of cigarette smoke condensate  which
 has  been  shown to  have  tumor-promoting capability  (Schlotzhauer,  et al..
 1978).
 III.  PHARMACOKINETICS
   _  Despite the presence  of resorcinol and resorcinol compounds in  numerous
 pharmaceutical, preparations, no  specific  information on the metabolism, dis—
 tribution, absorption,, or  excretion of resorcinol was found in the available
 literature.
 IV..   EFFECTS
      A.   Carcinogenicity
          The  available  data dealing  with the potential carcinogenicity of
 resorcinol are  at  this time inadequate  to formulate, a clear understanding of
 resorcinol's  oncogenic potential.   In a study  of commonly  used  cutaneous
agents, Stenback (1977)  showed no tumor  induction in rabbits and  mice from
topically  applied  resorcinol.   Resorcinol was selected because of  its pre-
sence in hair dyes.
          Van Duuren and Goldschmidt  (1976),  in a study  of  21 tobacco smoke
                                                                       »
components, found  that resorcinol  reduced the  carcinogenic potential of ben-
zo(a)pyrene  (BaP)  in dermal application  to mice.   Thus, fewer tumors  were

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 induced by BaP in1 the presence of resorcinol,  indicating  possible inhibition
 of carcinogenic activity.
           Substantial  evidence appears  to exist  for the oncogenic  activity
 of resorcinoi in plants.  Anderson  (1573)  reports that the "strong  carcino-
 genicity"  of resorcinol, tested in Nicotiana  hybrids suggests that "an onco-
 genic  reactivity of phenols is common to  plant and  animal  tissues but with
 differences  in  strength of reaction  to a derivative in a given system".
     B..-  Mutagenicity
          Dean  .(1578)  reports that  most phenolic  compounds including• resor-
 cinol are mitotic spindle poisons in plant tissues.   He further reports that
 considering  the severity of  effects  on plant  chromosomes that it is surpris-
 ing  that ir^ vivo  plant and  animal tests  have not  been  done  to determine
 their clastogenic properties.
          By  micronucleus test, Hossack and Richardson (1977)  were unable to
 find evidence of mutagenicity in resorcinol  or a; number,  of other hair  dye
 constituents  tested.
          The Ames  assay for resorcinol was^negative in  a  test of commonly
 used cutaneous agents (Stenback,  1977).
     C-   Teratogenicity and Other Reproductive Effects
          Pertinent data could not be located in the available literature.
     0.   Chronic Toxicity
          In  a study of chronic toxicity effects on  the  black cutworm,  Aqro-
 tis epsilon,  Reese  and  Beck  (1976)   found no  significant correlation  between
 resorcinol concentration and pupation or survival  but found  correlation with
body weight  at  various  stages of development.   They report  that  resorcinol
 is the  only  compound among  those tested which had "no adverse effect on  any
 of the nutritional indices and yet reduced growth.  It is  also the only com-
                               IS

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 pound which inhibited  growth  but did  not inhibit pupation."   They  hypothe-
 sized that resorcinol  may act  through a  temporary  inhibition of  ingestion
 but that the  insects continued to eat.  regularly,  allowing  pupation  on a nor-
 mal schedule  (Reese and Beck,  1976).
           In  the epiderrcLological study of the HR resin system  tire  manufac-
 turing plant,  Gamble,,  et al.  (1976) reported that HR exposed workers consis-
 tently showed  an excess of respiratory symptoms and that there  was a consis-
 tent association of alcohol  consumption  with increased incidence-  of  symp-
 toms.   The reported symptoms  included rash,  itch,, difficult  breathing  at
 work, cough,- chest  tightness,,  burning eyes, running nose, and burning  sensa-
 tion  in the heart region.
      E~   Acute  Toxicity
          with one  exception,  all acute toxicity  data  in the readily  avail-
 able  literature  are: supplied  by  Flickinger (1976) for the Koppers  Company,
 the primary  manufacturer  and  supplier of  resorcinol  in the  United States.
 Lloyd, et al.  (1977)  independently reported  the LD5Q  for  acute  oral  toxi-
 city  to be 370 mg/kg for resorcinol..
          In  a  review  of  the  industrial toxicology  of  the  benzenediols,
 Flickinger  (1976)  reports  various  acute  toxicity data for resorcinol.   A
 summary of relevant results follows:
          An  acute  oral  LD5Q for  resorcinol  was reported  by  Flickinger
 (1976) as 0.98 gm/kg in the rat..  Rats  dying during  the period  showed hyper-
emia  and  distension of  the stomach and  intestines..   Surviving rats showed
normal weight and no gross lesions at necropsy.
          The  ^50  f°r derma-L application in  the rat  was 3.36  gm/kg.  At
higher levels,   resorcinol  produced skin  necrosis.   At 1.0  gm/kg  levels,

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 moderate to severe irritation  was followed in 24  hours by slight hyperkera-
 tosis.  Surviving  rats  snowed reduced, weight  but no  internal gross lesions
 upon necropsy..
           Flickinger (1576)  reported that resorcinol  is a severe  eye irri-
 tant (0.1 gm  in  eye  of male, albino rabbits).   No recovery was seen  in the
 14-day follow-up period  with all exposed individuals  exhibiting keratoconus
 and pannus formation.
           Resorcinol  is a primary skin irritant.   Contact  with 0.5  gm of re-
 sorcinol on intact and abraided  skin produced moderate  irritation  on intact
 skin and varying  reactions  including  necrosis on  abraided skiru
           Inhalation  of  up  to  2,800  mg/nP  of  resorcinol  aerosol  for  8
 hours resulted in no observable toxic effects to  the  rats (Flickinger,  1576).
 V.    AQUATIC TOXICITY
      The possibility that resorcinol  may  be present in  some quantity  in  coal
 conversion process  effluents  requires further investigation as to the  feasi-
 bility  of control  technology.   Hertoes and  Beauchamp  (1577)  compared toxic
 interactions of two coal  conversion effluents,  resorcinol and  6-methylquina-
 line.  With Daphnia magna as a test  species,  they found mixtures of the  two
compounds to be less toxic  than either  pure compound tested alone.  They  re-
port a 48-hour LC5Q for resorcinol alone to be 1.28 mg/1.
     Curtis, et  al.  (1579)  reported the  acute  toxicity  of  resorcinol to
 freshwater  and saltwater organisms.    In  freshwater,   the  LC5Q values   for
 fathead  minnow  are  as follows:   24 hours,  88.6  mg/1;  48 hours,  72.6 mg/1;
and  96. hours,   53.4 mg/1.  In  saltwater, the  LC5Q values  for Palaemonetes
gugio or Penaeus  setiferus  are:   24 hours,  169.5 mg/1;  48  hours,  78.0 mg/1;
and 96 hours,  42.4  mg/1.   Thus, resorcinol was found to be toxic to aquatic
 life in both freshwater and saltwater.

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 VI..  EXISTING GUIDELINES AND STANDARDS
      There  are no  OSHA regulations,  NIOSH  recommendations,  or other  guide-
 lines concerning  resorcinol.  In 1974, ACGIH's Committee on Threshold  Limits
 proposed  a  TLV for resorcinol of 5 ppm.  Flickinger (1976)  reports of cur-
 rent  industrial 8-hour  workday exposures at 9.6 ppm "without.signs of  intox-
 ication or skin or  respiratory irritation" and recommends TLV industrial ex-
posures of "at least 10 ppm, perhaps even 20 ppm or  higher".   ACGIH has not
issued a formal TLV  for resorcinol.
     Information  regarding  existing  guidelines and standards  to  protect
aquatic life  from the effects of resorcinol was  not found in  the  available
literature.

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                                    REFERENCES
 Anderson,  R.A.   1973.  Carcinogenicity  of phenols,  alkylating agents,  ure-
 than,  and  a  cigarette-smoke  fraction  in  Nicotiana  seedlings.   Career  Re-
 search  33: 2,450.

 Curtis, M.W.,  et al.   1979.  Acute  toxicity of  12 industrial chemicals  to
 freshwater and saltwater, organisms,  water Research   13:  D7.

 Dean, 8.0.  1978.  Genetic  toxicology of benzene,  toluene, hylenes, and  phe-
 nols.  Mutation Research  47: 75.

 Flickinger, C.W.   1976.  The  benzenediols:  catechol,  resorcinol and hydro-
 quinone — a review  of  the  industrial toxicology  and current industrial  ex-
 posure limits.  Am. Ind. Hyg. Assoc. Jour.  37: 596.

 Gamble, J.F., et  al.   1976.  Respiratory function and symptoms:  an environ-
 mental - epidemiological study of  rubber workers exposed to a phenol-formal-
 dehyde type resin.  Am.  Ind. Hyg. Assoc. Jour.  (September 1976): 499.

 Heroes, S.E.  and  J.O. Beauchamp.   1977.   Toxic interactions  of  mixtures of
 two coal  conversion  effluent  components, (resorcinol  and 6-fliethylquinoline)
 to Daohnia magna.   Bull. Env. Contain. Toxicol..  17: 25.

 Hossack, O.O.N. and  J.C. Richardson.  . 1977.   Examination  of  the  potential
 mutagenicity of hair dye constituents using  the  micronucleus test.   Exper-
 mentia  33: 377.

 Lloyd, G.K.,.  et  al.   1977.   Assessment of the acute  toxicity of  potential
 irritancy  of hair  dye constituents.  Food Cosmet. Toxicol.  15: 607.

 Reese, J.C.  and  S.D. Beck.  1976.   Effects  of allelochemics on  the  black
 cutworm,  Aqrostis  ipsilon:  effects  of resorcinpl,  phloroglucinoal, and  Gal-
 lic acid on larval growth,  development,  and utilization of food.   Ann.  Ento-
 mol. Soc.  Am.   69: 999.

 Schlotzhauer,  W.S.,  et  al.   1978.   Characterization  of  catechols,  resorci-
 nols,  and  hydroquinones  in an acidic  fraction of cigarette smoke  condensate.
 Jour.  Agric.. Food  Chem.   26: 1277.

 Stanford Research Institute.   1975.  Chemical Economics  Handbook-  Stanford
 Research Institute,"Menlo Park, California.

 Stenback,  F.  1977.   Local  and systemic  effects of commonly  used  cutaneous
 agents:  lifetime studies of  16 compounds in mice and rabbits.   Acta.  Pharma-
 col.. et Toxicol.   41:  417.

.van Ouuren, B.L.  and 8.M. Goldschmidt.   1976.  Cocarcinogenic  and tumor-pro-
 moting agents  in   tobacco  carcinogenesis.   Jour.   Natl.  Cancer Institute
 56: 1237.

 Wilson,  C.Q.,  et  al.  (eds.)  1977.   Textbook of Organic and  Pharmaceutical
 Chemistry.   J.3.   Lippincott Co.,  Philadelphia,  Pennsylvania,  pp.  72,   181,
 194.
                                    — f f?*] A -
                                     ^ If IA. •'

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                                 No. 153
            Selenium

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
      WASHINGTON, D.C.  20460

         APRIL 30, 1980
         1*3-1

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                         DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards  from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations'of such sources, this short profile
may not reflect  all available  information  including  all  the
adverse health  and  environmental impacts  presented  by  the
subject chemical..  This document  has undergone scrutiny  to
ensure its technical accuracy.
                            7/3-3k-

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                             SELENIUM



                             SUMMARY






      Human  daily  intake  of selenium  has  been  estimated  at  50



to 150 jag/day.  While selenium is an. essential nutrient foe humans



and. other  species,  it  is  toxic  in  excessive amounts.   Selenium



poisoning  produces  symptoms in  man  similar to  those  produced



by  arsenic.    Although it  has  been  shown  to  produce  tumors  in
                                           v


animals/  the  Food  and  Drug  Administration,  the  International



Agency  for Research on Cancer  and the  National  Academy of Science



have  concluded that  the  available animal  data are  insufficient



to allow an evaluation of the  carcinogenicity of selenium compounds.



      The data base  for selenium  for aquatic life is quite limited.



No  chronic data are  available  for  marine  fish.   'Selenium does



not bioconcentrate to a great extent in  freshwater  species,  indi-



cating  that  tissue  residues  should not be a  hazard  to freshwater



organisms.  This information  is not available for marine organisms.
                                xl


                             •77*9-

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                             SELENIUM  •
I.    INTRODUCTION
      This profile  is  based on the Ambient Wa.ter Quality  Criteria
Document for Selenium  (U.S. EPA, 1979).
      Selenium  (Se;  atomic weight 78.96) is a naturally occurring
element  which reacts  with metals  to  form ionic  selenides  with
a valence of  minus  2,  and with most  other  chemials to form  cova-
lent  compounds.    It  may assume  any of  several  valence  states
ranging from minus 2 to plus 6.  Selenium is used in photocopying,
the  manufacture  of glass,  electronic  devices,   pigments,   dyes
ana  insecticides  (Dept.   Interior,  1974).    It  is also  used  in
veterinary medicine  (U.S.  EPA,  1979)  and in antidandruff  shampoos
(Cummings and Kimura,  1971).    The major  source  of  selenium  in
the  environment  is  the weathering  of rocks  and soils  (Rosenfeld
and  Beath,   1964)   but human  activities  contribute  about  3,500
metric tons per  year  (U.S. EPA,  1975a).   Selenium  is an essential
nutrient for humans and other species  (Schroeder, 1970).
II.   EXPOSURE
      Selenium  is  not  present in  measurable quantities  in  most
U.S.  drinking water  supplies.    Of  3,676  residences  located  in
33  geographically dispersed areas,  only 9.96  percexit of  the  sam-
ples  had selenium  levels above  the  detection  limits of 1  jug/1
(Craun,  et  al.  1977).    However,  in  seleniferous  areas  of  South
Dakota, levels of 50 to 330  ug/1  were measured  in  drinking waters
(Smith and Westfall, 1937).  Sewage plant effluents may contribute
to  tne  selenium content  of water;  as much as 28U  pg/1 have  been
reported in caw  sewage, 45 jag/1 in primary effluent, and 50 wg/1

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in secondary  effluent. (Baird, et  al.  1972).   Selenium concentra-
tions in  plants  depend largely  on the concentration  in  the  soil
where the plants are grown.   High selenium concentration in  vegeta-
tion is  transmitted  to other food sources, e.g.,  meats and  eggs.
The EPA  (1979) has estimated  the weighted  average bioconcentration
factor  for  selenium  to  be  18   for  consumed  fish  and shellfish.
Zoller  and  Reamer  (1976)  reported  that  most urban  regions  have
concentrations  of particulate'  selenium  ranging from  0.1  to  10
ng/m .
III.  PHARMACOKINETICS
      A.  Absorption
          Selenium  appears   to  be  effectively   absorbed   by  the
gastrointestinal  tract.     Thomson  and  Stewart  (1974)  reported
absorptions . of.  70,  64,   and 44  percent  for  sodium  selenite  in
three young  women.   Data from  rats are  similar with  absorptions
ranging  from 81  to  97 percent  for  a  number of organic selenium
compounds and sodium selenite (Thomson  and Stewart, 1973; Thomson,
et al.  1975).   The literature contains no information on  absorp-
tion by inhalation or dermal  exposures  (National Research Council,
1976).
      B.  Distribution
          The primary  disposition sites  for selenium  in  the  body
are  the  liver,   kidney,  spleen,  and   middle,  and  lower  sections
of  the  small intestine  (U.S. EPA,  1979) .   Based on  the  work  of
Kincaid,  et  al.   (1977)  it is apparent that  tissue concentration
levels of selenium can be affected both by dose  and normal  dietary
intake, although  the primary  deposition sites remain the same.

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      C.  Metabolism
          Selenium  is  an  essential  element  and  at  nutritional
levels  it  is  incorporated  into   specific  functional proteins;
at higher concentrations,  it is substituted for sulfur  in  sulfur-
containing  compounds.   Selenium  analogs  are  often  less   stable
than  sulfur  compounds,  ana  this   lability  may  be  the  basis of.
toxicity  (Stadtman,  1974).   Selenite and  selenate are  methylated
by  mammalian  tissues  in  an  apparent  detoxification  process.
Mouse liver, lung and kidney (Ganther, 1966)^are  active in methyla-
tion, but muscle, spleen, and. heart have little  activity.
      0.  Excretion
          Thomson  and Stewart  (197<±)  studied  selenium excretion
by feeding  three women selenite.   It was apparent that the  pri-
mary routes of  excretion  were  in the  feces ana  urine,  with  little
loss through the skin or lungs.
IV.   EFFECTS
      A.  Carcinogenicity
          Only  six  studies  have  been  performed  to  specifically
investigate whether  selenium  is  carcinogenic.   From  these  studies
there is  no conclusive evidence that  selenium has induced  tumors
in the  test animals.   The Food .and  Drug Administration  has de-
clarea  that selenium  poses no  carcinogenic risk  (Food and  Drug
Administration, 1973) .
      B.  Mutagenicity
          Selenium  has  been  shown  to affect  the  genetic  process
in barley  (Walker  and Ting, 1967)   and  in  Drosophila  melanog-aster
(Ting and Walker, 1969; Walker ana Braaley,  1969).  However,  these
                             / ^3 -I

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ana  other genotoxic effects  are  not true  mutagenic  effects.
There is  no study  in which a true mutagenic activity for  selenium
has been demonstrated.
      C..  Teratogenicity
          The  consumption ' of  seleniferous diets  interfered with
the  normal,  development of  the. embryo in  many mammalian  species,
including rats, pigs, sheep and cattle (U.S. EPA,  1979) .. Robertson
(1970)  suggested that  selenium may  be a  teratogen in  man from
the examination of the older literature whiqh  correlated malformed
babies and the consumption of  toxic grains  by  people  in Columbia.
      D.  Other Reproductive Effects
          Vesce  (1947)  noted  changes in  endocrine,  glands,  espe-
cially  the  ovaries,   following oral  administration  of  5  to 12.5
mg sodium selenide to guinea pigs over two  periods of 20 days.
      E.  Chronic Toxicity
          Chronic effects  from prolonged feeding of diets  contain-
ing added selenium in amounts of 5 to  15 ug/g  include liver  damage
in  the  form   of  atrophy,  necrosis,   cirrhosis,   and hemorrhage,
and  marked and  progressive  anemia   in  some  species   (Fishbein,
1977).   In man  hepatic  necrosis has  not  been observed following
chronic exposure; however,  lassitude,  loss of hair,   discoloration
ana loss of fingernails were symptoms  (Beath,  1962).
      P.  Other Relevant  Information
          The  essentiality of  selenium  for   several  animals,  has
been  known  since the  1950's  (Ganther, 1970;  Schwarz,  1961) with
                                                             •
selenium deficiency resulting in white muscle disease  in ruminants,
hepatic degeneration and  peridontal disease in other  mammals.

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Synergism/antagonism  exists between  the actions  of selenium  and


other metals  such  as  arsenic,  mercury,  cadmium,  silver and  thal-


lium (Diplock, 1976).


V.    AQUATIC TOXICITY


      A.  Acute Toxicity


          Cardwell, et  al.  (1976)  exposed 6 species of  freshwater


fish  to selenium  dioxide  and  observed  the 96-hour  LCcQ values


to  range from  2,060  to  28,500  ug/1.    The 96-hour  LC5Q values


for  fathead  minnow fry  and juveniles  are  2,060  and  5,200  ug/1,
                                           \

respectively,  indicating an  apparent  decrease in  toxicity with


age.   With  the  invertebrates  Daphnia   magna  and scud,  the LC5Q


values  are  430  and 313 ug/1 respectively (U.S. EPA, 1978; Adams,


1976) .


          The  96-hour LCeg  values for  marine  species  are  6,710


pg/1 for  the sheephead minnow  (U.S.  EPA, 1978) and  600 ug/1  for


mysid shrimp  (U.S. EPA, 1978).


      B.  Chronic Toxicity


          No pertinent  data are available on the chronic  toxicity


of  selenium  to freshwater  organisms  (U.S.  EPA, 1979).   The only


data  available, in  marine  species is  that  of  the mysid shrimp


(Mysidopsis  oahia).    It  has  been  exposed to selenium  for  its


life cycle and the chronic  value is 135  ug/1.


      C.  Plant Effects


          Selenium  is  toxic,  to  two  freshwater  algal  species,


Chlorella  vulgaris  and  Haematoccus  cupensis,  with growth  being

                                                             »

retarded at  50 ug/1  (Hutchinson  and Stokes,  1975).  For the  salt-


water alga, Skeltonema costaturn, the  96-hour ECcQ values for

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chlorophyll a^  and  cell numbers are  7,930  and 8,240 pg/1, respec-
tively (U.S.  EPA, 1978).
      0.   Residues
          Bioconcentration  factors  have  been determined  for  the
rainbow trout,  fathead  minnow  and bluegill.   These factors range
from  Z to  20  (Adams,  1976,- U.S.  EPA,  1978).   The  tissue half-
life  for  the  bluegill is between  1  and 7  days  (U.S.  EPA, 1978).
These  results  show  that  tissue  accumulation of  selenium should
not present a hazard to freshwater aquatic, organisms.
          No residue  data  are  available,  for  marine species (U.S.
EPA, 1979). •
VI.   EXISTING GUIDELINES
      A..  Human
          The  U.S.   Environmental  Protection  Agency  (1975b)  has
established  the maximum  permissible,  level of  selenium  at  0.01
mg/1 for U.S. drinking waters.   A time-weighted average concentra-
tion threshold limit value  (TLV) of 0.2 mg/m  has been established
by  the American  Conference of  Government  Industrial. Hygienists
(ACGIH,  1977).    The minimum  toxic,  dose  for  selenium  has  been
calculated  to  be 16.1  mg/day.  The  U.S. EPA  (1979)  draft water
criterion for selenium is 10 ug/1.  As a result of public comments
received,  additional review and consideration  of  the recommended
criterion is required.
      B.  Aquatic
          For  selenium  in  freshwater,  the  draft criterion* to
protect  aquatic life  is  9.7  ug/1 as a  24-hour  average  and  the
concentration .should not exceed  22  pg/1  at any  time  (U.S.  EPA,
                             "•»
1979) .. .In saltwater the criterion is 4.4 ug/1 as a  24-hour average
and  the concentration should not exceed  10 ug/1 at any time.
                             -T&r-

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                             SELENIUM

                            REFERENCES
Adams, W.J.  '1976.   The toxicity and residue dynamics of  selenium
in fish and aquatic  invertebrates»  Diss. Abstr.  Int.  p.  121.

American Conference  of  Industrial  Hygienists.   1977.   Threshold
limit  values  for chemical  substances  in  workroom  air  adapted
by ACGIH for 1977..

Baird,  R.B.,  et  al.   1972..   Determination of  trace  amounts  of
selenium in  wastewaters by  carbon  rod atomization.   Anal.  Chem.
44: 1887.

Beath, O.A.  1962.   The story  of selenium in^ Wyoming.  University
of Wyoming» Lararaie.

Cardwell, R.D., et al.   1976.   Acute toxicity of selenium dioxide
to freshwater fishes.  Arch. Environ. Contam. Toxicol. 4:  129.

Craun,  G.F.,  et  al.   1977.   Preliminary report of an epidemio-
logic  investigation  of  the  relationship(s)  between  tap  water
constituents and  cardiovascular  disease.    Proc.  Am.  Water Works
Assoc. Meet.

Cummins,  L.M.   and   E.T.  Kimura.    1971.    Safety  evaluation  of
selenium sulfide antidandruff shampoos.  Toxicol. Appl. Pharmacol.
20: 89.

Department of  Interior.   1974.   Minerals  yearbook,  1972.   Bureau
of Mines, Washington, D.C.

Diplock, A.T.   1976.   Metabolic  aspects  of  selenium action and
toxicity.  CRC Crit. Rev. Toxicol. 271.

Fishbein, L.   1977.   Toxicology  of  selenium and tellurium.    Adv.
Mod. Toxicol.,  Vol.  2, .1 ISS Trace Elem., 191.

Food  and  Drug   Administration.    1973.   Selenium  in animal feed.
Federal Register, Vol. 38, No.  81.

Ganther,  H.E.    1966.    Enzymic synthesis  of  dimethyl   selenide
from  sodium selenite  in  mouse  liver  extracts.    Biochemistry
5: 1089.

Ganther,  H.E.    1970.    In Trace  element  metabolism  in   animals,
ed. C.F. Mills, Edenburgh:Livingstone, 212.

Hutchison,  T.C. and  P.M.  Stokes.   1975.   Heavy  metal   toxicity
and algal bioassays.  Water quality parameters.   ASTM: 320.

Kincaid, R.L.,  et al.   1977.   Effect of added dietary  selenium
on  metabolism  and tissue  distribution  of radioactive  and stable
selenium in calves.  Jour. Anim. Sci.  44:1: 147.

                                J_

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National  Research Council.    1976.    Selenium.  Comm.  Meci. Biol.
Effects Environ.  Pollut.,  Subcomm.  Selenium.    Natl.  Acad. Sci.,
Washington, D.C.

Robertson, D.S.F.   1970.  Selenium,  a possible  teratogen?  Lancet
L: 518.

Rosenfeld, I.   1964.   Excretion and  retention of   Se in relation
to  modes   of  administration,  toxicity  and  pregnancy  in  rats.
Metabolic  effects and  metabolism of  selenium  in  animals.   Part
IV, Bull..  414..  Agric.. Exp. Sta., University of  Wyoming.

Rosenfeld, I.  and O.A.  Beath.    1964.   Selenium;  geobotany, bio-
chemistry, toxicology and nutrition.   Academic Press, New York.

Schroeder, H.A-   1974.   Selenium. Page 101 in The poisons around
us..  Indiana University  Press, Bloomington.,.

Schroeder, H.A.,  et  al.  1970.   Essential trace  metals  in man:
selenium.  Joar. Chron.  Dis.  23: 227.

Schwarz, K.   1961.   Development and  status  of  experimental work
of factor  3 selenium.  Fed. Proc.  20.: 666.

Smith,  M.I.  and  B.3.  Westfall.    1937.   Further  field  studies
on the  selenium problem in relation  to  public, health.   U.S. Pub.
Health Rep..  52: 1375.

Stadtman,  T.C.   1974.  Selenium biochemistry.  Proteins containing
selenium are essential  components of certain bacterial and mamma-
lian enzyme systems.  Science  133:  915-

Thomson, C.D.  and R.D.H.  Stewart.    1973.   Metabolic  studies  of
( °Se)  selenomethionine and   (   Se)   selenite in   the  rat.   Br.
Jour. Nutr.  30: 139.

Thomson, C..D. and R.D.H. Stewart.  1974.  The metabolism of  (  Se)
selenite in young women.  Br. Jour.  Nutr.  32: 47.

Thomson, C.D-75et  al.   1975.   Metabolic studies  of  (  Se)  seleno-
cystine and  (   Se)  selenomethionine  in the rat.    Br.  Jour. Nutr.
34: 501.

Ting,  K.P. and. G.W.R..  Walker.   1969..   The  distributive  effect
of  selenoamino  acid  treatment  on  crossing-over   in  Drosophila
melanogaster.  Genetics  61: 14iw

U.S.  EPA.    1974.    Safe drinking water act,  puolic  law  93-523,
93rd Congress,  S. 433.
                                                             »•
U.S.  EPA.   1975a.    Preliminary  investigation of  effects  on  the
environment  of boron,   indium,  nickel,   selenium,   tin,  vanadium
and  their compounds.    Selenium.   U.S.  Environ.   Proc.  Agency,
Washington, D.C.

U.S.  EPA.   l975b.   National  interim primary drinking water regu-
lations.   Fed.  Reg.  40:248: 59566.

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U.S.  EPA.    1978.   In-depth  studies  on  health  and environment
impacts of selected water pollutants.  -U.S. Environ.  Prot. Agency.
Contract No. 68-01-4646.

U.S.  EPA.    1979.   Selenium:  Ambient Water  Quality  Criteria.
Environmental Protection Agency, Washington, D.C.

Vesce, C.A.  1947.  Intossicazione spermentale da.selenio.  Intos-
sicazione Sperimentale da Selenio, Folia Med.  (NAPOLI)  33: 209.

Walker,  G.W.R-   and  A.M.  Bradley.    1969.    Interacting effects
of  sodium  monohydrogen  arsenate and  selenocystine  on   crossing-
over in Drosophila Melanogaster.  Can. Jour. Genet. Cytol.  11:  677.

Walker,  G.W.R.  and  K.P.  Ting.   1967.   Effects  of  selenium in
recombination in barley.  Can. Jour. Genet. Cytol.  9: 314.

Windho
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                                     No.  154
               Silver

  Health and Environmental Effects
U.S.. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980
              . /- * 4 ^
            -/a J-'-

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

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                           SILVER
                           SUMMARY
     While metallic silver in  the zero  valence  state is not
considered to be toxic, most of  its  salts  are  toxic to a
large number of organisms.  Silver salts  can  combine with
certain biological-molecules and subsequently  alter their
properties.  Upon ingestion, many silver  salts  are  absorbed
in the human circulatory system  and  deposited  in  various body
tissues, resulting in generalized or sometimes  localized gray
pigmentation of the skin and mucous  membranes known as argy-
ria.  Silver has not been shown  to be a carcinogen  (except by
the mechanism of solid state tumorigenisis); however,  there
is some evidence that silver salts can  effect  the growth of
tumors..  The acceptable daily  intake for  silver has been de-
termined to be 1.6 mg per day  for a  70  kg  man..
     Silver is acutely lethal  to aquatic  species  in the ug/1
range.  In terms of acute lethality,  Daphnia magna  appears to
be the most sensitive species, with  a 48-hour EC5Q  of  1.5
ug/1..  At levels as low as 0,17  ug/1/ silver caused premature
egg hatching and reduced, fry growth  in  fathead  minnows.

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                           SILVER
I.   INTRODUCTION
     This .profile is based on the Ambient Water Quality Cri-
teria Document for Silver (U.S. EPA, 1979).
     Silver (Ag; atomic weight 107.87) is a white ductile
metal occuring naturally in the pure form and  in ores.
Silver can exist in two valence states, Ag"1" and Ag"1"*",
The solubility of common silver salts varies greatly, with
silver nitrate having a solubility of 2,,5-x 109 p.g/1 and
silver iodide having a solubility of 30 ug/1 (Windholz,
1976).  Many silver salts are light-sensitive..  Water or
atmospheric oxygen have no effect on metallic  silver; how-
ever, ozone, hydrogen sulfide, and sulfur react with it.  The
principle uses of. silver are in photographic materials, elec-
troplating, dental alloys, solder and brazing  alloys, paints,
jewelry,  silverware, coinage, mirror production.
II.  EXPOSURE
     Exposure to silver is mainly through food and water
intake with only minimal contribution from ambient aerosols.
Concentrations of silver in.surface waters have been shown to
vary from 0-38 ug/1 with a mean of 2.6 ug/1  in samples
containing silver.  High silver concentrations are obtained
in high silver mineralized areas or in waters*  receiving
effluent from industries that use silver.
     The average intake of silver from food has been calcu-»
lated to be 40 ug/day (Tipton, et al. 1966) to 88 ug/day
(Kehoe, et al. 1940) in the U.S.  Although silver is detected
in neats and vegetables, the concentrations in fish, shell-
                                      y-

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fish, .and Crustacea are greater.  Marine  animals  accumulate
silver in concentrations which are higher than  their environ-
ment.  This is particularly significatnt  in  areas such  as
sewage-sludge dumping sites, which contain high concentra-
tions of silver  in "the sediment.  The  dead bodies of animals
in reducing environments wall contribute-  their  silver to sed-
iments,, a major  factor in  the geochemical cycle of silver
(Boyle, 1968).
     Exposure to high levels of  silver has also occurred by
inhalation in specific industries (e.g.,  silver smelting and
photography) and from mechanical uses  of  silver compounds.
Steel mills do not seem to contribute  to  ambient  air concen-
trations of silver (Harrison, et al. 1971).
III.. P HARM AGO KINETICS
     A.   Absorption
          Silver may enter the body via the  respiratory
tract, the gastrointestinal tract, mucous membranes,  or bro-
ken skin.  The efficiency  of absorption by any  of these
routes is poor.  Colloidal silver given orally  to rats  showed
two to five percent -absorption by the  gastrointestinal  tract
(U.S. EPA, 1979).  Dogs receiving orally  a tracer quantity of
silver nitrate   absorbed ten percent.   It was shown in  hu-
mans who accidently inhaled silver that the  biological  half-
life of silver was about one day, pcobably due  to rapid muco-
ciliary clearance, swallowing, and fecal  excretion (Newton .
and Holmes, 1966).  Some absocpotion did  take place since
there was localization of  silver in the liver,  but quantifi-
cation was impossible.  In human burn  patients  treated  with
                           _ ;o i *r..
                            / (J J>

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silver nitrate dressing, only 0.008 percent of  the  silver  was
absorbed (U.S. EPA, 1979).
     B.   Distribution
          The amount of silver, its chemical  form,  and  the
route by which it is administered affects the tissue content
and distribution of silver within the body (Furchner, et al.
1968).  Table 1 summarizes data on the distribution of  silver
in rats.
     Table 1:  Distribution of Silver in the Rat and Day 6
               Following Intramuscular Injections of Differ-
               ent Doses of Silver (percent of  dose per or-
               gan) (Scott and Hamilton, 1950).

Percent of Dose
Absorbed
Absorbed
Heart and Lungs
Spleen
Blood
Liver
Kidney •
G.I. tract
Muscle
Bone
Skin
Urine
Feces
Unabsorbed

Carrier- Free

92.1

0.06
0.01
0.50
0.36
0.07
1.12
0.27
0.18
0.24
(1.64
96.56
7.9
Dose
0 .1 ng

63.7

0.13
0.13
0.95
2.24
0.92
4.22
0.56
0.35
0.67
0.88
88.95
36.3

1.0 ma

53.5

0.59
2.69
3.03
33.73
0.63
8.21
2.39
2.20
7.39
1.82
37.33
46.5
Silver administered to other species appears to generally
follow this distribution pattern.
  C.  Metabolism
     Inhaled silver particles that are not removed from the
lungs by the mucociliary reflex and coughing are probably
                              ^r
                         ^^^Q <7 ft
                            iA JU

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phagocytized and transported  via  the  protein fractions of the


blood plasma to the liver, from which they are  eventually ex-


creted in the bile.  Formation of silver  selinide deposits in


the liver, as well as the formation of metallic silver,


silver sulfide, or silver complexes with  sulfur amino acids


may be a method of detoxifying silver.  -In the  kidney, com-


plexation with metallothionein may be another detoxification


pathway (U.S. EPA, 1979).

     D.   Excretion
                                          ».

          Regardless of  route and chemical form of silver


administered, fecal excretion always  predominates over uri-


nary excretion.  Most absorbed silver is  excreted into the


intestines by the liver  via the bile.. Phalen and Morrow


(1973) exposed beagle dogs to an  atmosphere  containing silver


aerosols and showed the  biological half-life to be 8.4 to


12.9 days..


IV.  EFFECTS


     A.   Carcinogenicity


          Implanted foils and disks and injected colloidal


suspensions of metallic  silver have been  found  to produce


tumors or hyperplasia in several  studies.  These tumors  may


be due to the particular physical form of  the metal  or  to its


being an exogenous irritant.  There is no  evidence that


silver or its salts produce tumors by any'other mechanisms.


In one study, intratumoral injections of  colloidal silver ap-


peared to stimulate cancer growth (Guyer  and Mohs, 1933),  and


in another study silver  nitrate 'appeared  to  act as a  promoter


with DMBA (7,12-dimethylbenz(a)anthracene)  initiated  mice

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(Saffrotti and Shubik, 1963).  On the other  hand,  Taylor  and
Carraichael (1953) showed a tumor growth  inhibitor  effect  of  •
silver chloride.  The evidence for any carcinogenic  effect of
silver is'very tenuous (U.S. EPA, 1979).
     B.   Mutagenicity
          Silver nitrate (Demerec, et al. 1951), silver chlo-
ride (Nishioka, 1975), and silver sulfadiazine  (Fox,  et al.
1969) have been examined for rautagenicity in microorganisms
and shown to be nonmutagenic in these test systems.
     C.  • Teratogenicity
          Few associations between silver and birth  defects
have appeared in the literature and one  is apparently erro-
neous.  Kukizaki (1975) found only weak  cytotoxic  effects
when silver-tin alloy powder was incubated in seawater  with
fertilized eggs or early embryos of the  sea  urchin Hemicen-
trotus pulcherrinus.  Silver salts were  tested  for toxicity
to 4- and 8-day-old chick embryos but did not produce abnor-
malities  in development (Ridgway and Karnofsky, 1952).
     D.   Other Reproductive Effects
          Pertinent information could not be located  in the
available literature concerning any other reproductive  ef-
fects due to exposure to silver.
     E.   Chronic Toxicity
          In rats, chronic exposure to 0.4 mg/1 of silver in
drinking water causes hemorrhages in the kidney.   Larger
doses cause changes in conditioned-reflex activity,  lowering
of immunological resistance  (0.5 mg/1),  and growth depression
(20 mg/1).  In humans, the most common noticeable  effect  of

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chronic exposure to silver or silver compounds  is  generalized

argyria (generalized gray pigmentation).

     F.   Other Relevant Information

         'Silver exhibits antagonism to selenium,  vitamin E,

and copper, inducing deficiency symptoms  in  animals  fed  ade-

quate diets: or aggravating deficiency symptoms  when  the  ani-

mal's diet lacks one- or more of the nutrients.  The  effects

have been described in dogs, sheep, pigs,  rats, chicks,  tur-

key, poults, and ducklings (U.S. EPA, 1979).

V.   AOUATIC TOXICITY

     A.   Acute Toxicity

          Davies, et al. (1978) conducted  96-hour  tests  with

rainbow trout in both hard (350 mg/1 as CaC03)  and soft

water (26 mg/1 as CaCC>3) water.  The LC50  values were

6.5 and 13 uq/1 for soft and hard water,  respectively.   There

are too few data" to assess the relative importance of  hard-

ness and experimental variability on these nonreplicated  re-

sults.

          The 48-hour static EC5Q for Daphnia magna  in

soft water (40 mg/1 as CaC03) is 1.3 ug/1  (U.S. EPA, 1978),

indicating that this species is the most  sensitive freshwater

invertebrate species tested.

          Acute toxicity data are available  only for four

saltwater invertebrate species and range  from 5.8  to 262  ug/1,

(Calabrese, et al. 1973; Calabrese and Nelson,  1974; nelson,
                                                           »
et al. 1976; Sosnowski and Gentile in; U.S.  EPA, 1979).   The

American oyster is the most sensitive saltwater species  test-

ed, and the mysid shrimp is the most resistant.

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     B..   Acute Toxicity
          Davies, et al.  (1978) conducted  an  18-month  mortal-
ity test with rainbow trout and found  the  no-effect  concen-
tration of silver to be 0.09 - 0..17 ug/1  (17.2%  mortality at
0.17 ug/1 and no mortality at 0.09 ug/1).  There was also
premature hatching of eggs and reduced growth of fry at  0.17
ug/1.
          The chronic toxicity of silver  to mysid  shrimp has
been determined, based on a flow-through,  life-cycle  exposure
                                          «.
(Sosnowski and Gentile _in: U.S. EPA, 1979).   No  spawning
occurred at 103 ug/1-  The time of spawning was  delayed  to
seven days at 33.3 ug/1.  Brood size was  statistically
smaller at 33.3 ug/1 when compared to  the  controls,  although
larval survival was not affected.  The highest concentration
of silver tested that had no effect on growth, reproduction,
or survival was 10.2 ug/1/ which is approximately  0.04 times
the 96-hour LCgg determined for adult  shrimp.
     C.   Plant Effects
          Hutchinson and Stokes (1975) observed  growth retar-
dation in the freshwater alga, Chlorella vulgaris, at  silver
concentrations, between 10 and 60 ug/l»  A  concentration  of
2,000 ug/1 was determined to be toxic  to  six  additional  algal
species (Gratteau, 1970).
          The only marine algal species tested,  Skeltonema
costatum, showed growth inhibition after a 96-hour exposure
                                                           f
to 130 ug/1 (U.S. EPA, 1978).
     D.   Residues
          Bioconcentration factors of  17  to 368  were deter-
mined for three species of insects exposed to  silver

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(Nehring, 1973).  Bluegills showed no  bioconcentration  of

silver at a water concentration of 0.03 ug/1  after  a. 28-day

test (U.Sk EPA, 1978).  Pertinent  information on  residues  in

saltwater species could not be located  in  the available

Literature.

VI.  EXISTING GUIDELINES AND STANDARDS

     A.   Humans

          Both the U.S. standard for silver  in drinking water

and in workplace air have been based on a'presumed  1  g  mini-

mum dose of silver that has caused agryia.

The existing standards for silver are:

             Existing Standards Regarding  Silver

Medium                  Silver Concentration      Authority

Drinking water                50 ug/1          U.S..  EPA  (1976) Ra-
                                               tional  Academy  of
                                 ,             Sciences  (1977)

Drinking water                 0.5 ug/1        State of  Illinois
                                               (cited  in National
                                               Academy of Sci-
                                               ences ,1977)

Drinking water                10 ug/1          State of  California
                                               (cited  in National
                                               Academy of Sciences,
                                               1977)

Workplace air, thresh—         0.01 mg/m^      Occupational  Safety
  old limit value                              and Health Adminis-
  time-weiahted                                tration (1974)
                                               (39 FR  23541)

Short-tern exposure            0.03 mg/i?.3     .American  Conference
  limit  (_>. 15 minutes)                         of  Governmental In-
  4 tines per day                              dustrial  Hygiensts
                                               (1977)

The acceptable daily intake (ADI) for  silver  is 1.6 mg/day.   The

U.S.. EPA draft water criterion for silver  is  10 ug/1  for the

protection of human health.  This criterion  is presently

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undergoing further evaluation and review  before  final  recom-


mendation.


     B.   Aquatic


          For silver the draft criterion  to protect fresh-


water aquatic life-is 0.009 ug/1 as a  24-hour average;  the


concentration should not exceed 1.9 ug/1  at any  time (U.S.


EPA, 1979)..


          To protect saltwater aquatic life, the draft  cri-


terion is 0.26 ug/1 as a 24-hour average; the concentration
                                          «.

should not exceed 0.58 ug/1 at any time (U.S. EPA,  1979).

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                              SILVER

                            REFERENCES
American  Conference  of  Governmental Industrial  Hygienists,  1977
TLV  Airborne  Contaminants  Committee.    1977.    TLVs  threshold
limit values. for  chemical  substances  and physical agents  in the
workroom  environment  with intended  changes  for 1977.   Am.  Conf.
Govt. Ind.. Hyg., Cincinnati, Ohio.

Boyle, R.W.   1968.   Geochemistry of silver  and its  deposits with
notes on  geochemical prospecting  for  the  element.    Geol.  Surv..
Can., Bull. No. 160.   p. 1.

Calabrese, A.,  and D.A.  Nelson.    1974.   Inhibition  of embryonic
development  of the  hard  clam,   Mercenaria  mercenaria by  heavy
metals*  Bull. Environ. Con tarn. Toxicol.  Lit- 92.

Calabrese, A.,  et al.   1973.   The toxicity of heavy  metals  to
the  embryos,  of   the  American  oyster   (Crassostrea   virginica).
Mar. Biol.  18t 162.

Davies. P.H., et  al.   1978.  Toxicity of silver to  rainbow trout
(Salmo gairdneri).  Water Res.  12: 113.

Demerec,  M..,  et al.   1951.   A survey of chemicals  for mutagenic
action on S.  coli.  Am. Nat.  85: 119»

Fox,  C.L.,  et  al.    1969.   Control of Pseudomonas  infection  in
burns by silver sulfadiazine.  Surg.. Gynecol.7 Obstet.   128: 1021.

Furchner, J.E.,  et  al.   1968.   Comparative  metabolism of radio-
nuclides  in mammals.   IV.  Retention of silver-llOm in the mouse,
rat, monkey and dog.   Health Phys.  15: 505.

Gratteau,  J.C.    1970.    Potential  algicides  for  the  control  of
algae.  Ref.  No..  1970..  Water Sewage Works p. R-24.

Guyer, M.F.  and  F.E.  Mohs.    1933.    Rat  carcinoma  and injected
colloidal platinum.   Arch. Pathol.  15: 796.

Harrison,  P.R.,  et  al.   1971.   Areawide trace  metal concentra-
tions  measured  by   multielement  neutron  activation analysis:
A one day study  in  northwest Indiana.   Jour.  Air  Pollut.  Contro.
Assoc.  21: 563~

Hutchinson, T.C., and P.M.  Stokes.   1975.    Heavy  metal toxicity
and algal bioassays.  Water Quality Parameters. ASTM SPT  573:  320.

Kehoe, R.A., et al.   1940.   A spectrochemical study  of the normal
ranges  of concentration  of certain  trace  metals  in biological
materials.  Jour.  Nutr.  19: 579.
                              AT/-/3

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Kukizaki,  S.    1975.    Studies on  the  effects  of  dental amalgam
upon  the  fertilization and  early  development of  sea urchin  eggs
(Hemicentrotus pulcherrimus).  Jour. Jap. Soc. Dent.  Appar. Mater.
16: 123.

National Academy  of  Sciences.  1977.   Drinking  water and health.
U.S. Environ. Prot.. Agency,  Washington, D.C., PB-269 519.   Natl.
Tech. Inf. Serv., Springfield, Va.

Nehring,  R.V.    1973.    Heavy  metal toxicity  in two  species of
aquatic  insects.    Master's  Thesis,  Colorado-  State  Univ.,   Fort
Collins, Colorado.  82 p.

Nelson,  D.A.,  et al.   1976.  Biological effects  of heavy metals
on  juvenile bay  scallops,  Argopecten  irradians,   in short-term
exposures.  Bull. Environ. Contain.. Toxicol.   16: 275.

Newton, D. and A. Holmes.   1966..   A case of accidental inhalation
of zinc-65 and silver-HOm.  Radiat. Res.  29: 403.

Nishioka,  H.    1975.   Mutagenic  activities  of metal  compounds
in bacteria.  Mutat. REs.  31: 135.

Phalen,  R.F.  and P.E.  Morrow.    1973.   Experimental inhalation
of metallic silver.   Health  Phys.  24: 509.

Ridgway, L.D.,  and  D.A.  Karnofsky.  1952.   The  effects  of metals
on  the  chick  embryo:   Toxicity  and production of   abnormalities
in development.  Ann. N.Y. Acad. Sci. 55: 203.

Saffrotti, U.  and P.  Shubik.   1963.   Studies  on promoting action
in skin carcinogenesis.  Natl. Cancer Inst.  Monogr.   10: 489.

Scott,  K.G.  and J.G. Hamilton.   1950.   The  metabolism  of silver
in  the  rat with radiosilver  used  as an indicator.   Univ.  Calif.
(Berkeley) Publ. Pharmacol.  2: 241.

Sosnowski,  S.L.  and J.H.  Gentile.   Chronic  toxicity  of  copper
and  silver to  the  raysid  shrimp  Mysidopsis bahia.    EPA-Environ-
mental Research Lab., Narragansett, R.I.Manuscript.

Taylor,  A. and  N.  Carmichael.    1953.   The  effect of metallic
chlorides  on  the growth  of  tumor and  non-tumor tissue.   Univ.
Texas  Publ.  No..  5314,  Biochem.  Inst.  Stud.  5, Cancer  Stud. 2-
p. 36.

Tipton,  I.H., et  al.   1966.   Trace elements in diets and excreta.
Health Phys.  12: 1683.
                                                              »
U.S.  EPA.   1976.  National interim primary  drinking  water regula-
tions.   EPA-570/9-76-003.   U.S.   Environ.  Prot. Agency,  Off. of
Water Supply.

U.S.  EPA.   1978.   In-depth  studies on health  and   environmental
impacts  of selected  water  pollutants.   U.S.  EPA  Contract  No.
68-01-4646.  U.S. Environ. Prot. Agency, Washington,  D.C.

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U.S. EPA.   L979.   Silver:  Ambient  Water  Quality  Criteria.   Envi-
ronmental Protection  Agency, Washington, D.C.

Windholz, M.  (ed.)    1976.   The Merck  Index.   9th  ed.'    Merck
and Co., Inc.,  Rahway,  N.J.
                         IS'H- If

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                                       No.  155
                TCDD

  Health and Environmental Effects
a.S- ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This 'report, represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal fc  The information contained in  the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental  impacts  presented by  the
subject chemical..   This  document-  has undergone scrutiny  to
ensure, its. technical accuracy.:
                        Iff-*-

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          2,3,7,8-TETRACHLORODIBENZO-P-DIOXIN (TCDD)

                           SUMMARY

     2,3,7,8-Tetrachlorodibenzo-p-dioxin  (TCDD) has been

found to induce heptocellular carcinomas  and tumors in  two

rat feeding studies.  TCDD has also produced fetotoxic  and

teratogenic effects in laboratory animals.  The positive

rautagenicity of TCDD has been, demonstrated in three bacte-

rial bioassay systems.  TCDD is also a potent inducer of

hepatic and renal microsomal drug metabolizing enzymes.
                                v
     No standard tests for acute or chronic toxicity in

aquatic life have been conducted_with TCDD.  Other studies,

however, have shown adverse effects over  a period of 96

hours to concentrations as low as 0.000056 jug/1.  The weighted

average bioconcentration factor for TCDD  for edible portion

of all aquatic organisms consumed by Americans has been

calculated to be 5,800.

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          2,3,7,8-TETRACHLORODIBENZO-P-DIOXIN (TCDD)

I.    INTRODUCTION
     2,3,7,8-Tetrachlorodibenzo-p-dioxin  (TCDD)  is a contami-
nant unintentionally formed during the production of 2,4,5-
trichlorophenol (TCP) from 1,2,4,5-tetrachlorobenzene.
TCDD is also found as a contaminant of 2,-4,5-trichlorophenoxy-
acetic acid. (2,4,5-T) (U.S. EPA, 1979).
     Characteristically, TCDD  (C^^C^f^) is a  white crystal-
line solid with the. following physical properties: melting
point, 302-305°C; solubility in water, 0.2 to 0.6 ug/1;
lipiphilic, and non-volatile (U.S.. EPA, 1979) ..
     TCDD is considered a relatively stable compound which
can: be degraded at temperatures in excess of. 500°C, or by
irradiation 'with UV light or sunlight under certain condi-
tions (U.S. EPA, 1979).   It has. been shown to disappear
slowly from soil with residues persisting for ten years
after application.  TCDD bio-accumulates  in aquatic organisms.
II.  EXPOSURE
     A.   Water
          The amount of human exposure that can  be directly
attributed to drinking water alone is difficult  to determine
(U.S. EPA,. 1979).  It has been stated that no TCDD has ever
been detected in drinking water, with limits of  detection
in the parts per trillion range (National Research Council,
1977).  Underground water supplies would probably not be
contaminated with TCDD under most conditions since vertical
movement of TCDD has not been demonstrated in soil (Kearney,
et al., 1972).

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     B.   Food
          The occurence of TCDD in food could  result from
(1) accidental spraying of plant crops;  (2) contaminated
forage or (3) food chain magnification  (U.S. EPA,  1979).
          TCDO is neither absorbed by oat and  soybean seeds
after spraying, nor taken up from the soil  into  the  mature
plants (Isensee and Jones, 1971; Matsumura  and Benezet,
1973).  Aqueous solutions of pure TCDO exposed to  either
artificial light or sun light, do not decompose, whereas
TCDD photodecomposes rapidly when applied to leaf  surfaces
as a contaminant of the herbicides Agent Orange  and  Esteron
(Crosby,  at al., 1971; Crosby and Wong, 1977).
          TCDD has been detected in the adipose  tissue of
cattle feeding on contaminated forage (Kocher, et  al., 1978)
Studies conducted for the U.S. EPA also found TCDD in fat
of cattle previously exposed to 2,4,5-T  (U.S. EPA, 1979).
No TCDD,  however, was detected in liver samples.
          The U.S. EPA (1979) has estimated the  weighted
average bioconcentration factor of TCDD at  5,800.  This
estimate  is based on measured steady state  bioconcentration
studies in channel catfish containing 3.2 percent  lipids
(Isensee  and Jones, 1975).
     C.   Inhalation
          Pertinent information could not be located  in
the available literature.

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III. PHARMACOKINETICS
     A.   Absorption
         • Approximately 83-86 percent of  the TCDD  administered
in a single oral'dose, following activation with multiple
oral doses, is absorbed from the intestinal tract  (Rose,
et al., 1976).
     B.   Distribution
          The excretion of a single oral  dose of TCDD  in
rats occurred via the feces (53 percent), urine  (13 percent),
and expired air  (two percent)  (Piper, et  al., 1973).   An-
alysis after three days showed the highest percent of  the
administered dose per gram in the liver (3-18 percent) and
adipose (2.60 percent).
          Rose, et al. (1976)  found that  22 days, after a single
             L4
oral dose of   C labeled TCDD, 1.26 and 1.25 percent, of
    14
the   C was retained per gram of liver and adipose tissue,
respectively..  After repeated oral doses, however, the liver
was found to have five times as much radioactivity as  adi-
pose tissue.  Single oral doses of TCDD were excreted  through
the feces, whereas significant amounts of radioactivity
were found both  in the urine and the feces after repeated
oral doses -
     C..   Metabolism
          There  is no complete agreement  as to whether or
not TCDD is actually metabolized (U.S. EPA, 1979).  Rose, •
et al. (1976) found unchanged 14C-labeled TCDD in  the  liver
after oral administration, but noted that most of  the  radio-

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activity in the feces came  from compounds  other  than TCDD.
The slow elimination of TCDD from  rats  and monkeys suggests
that it is not readily metabolized (Van Miller,  et al.,
1976).
     0.   Excretion
          See also section  B., Distribution.
          Differences in TCDD elimination  have been observed
between the sexes and between species.  Rose,  et al.  (1976)
found male rats excreted 3.1 percent of the cumulative dose
in the urine while females  excreted 12.5 percent in the
urine.
          The half-life of.  radioactive  TCDD following a
single oral dose to rats was 31— 6 days, while that follow-
ing repeated oral doses was 23.7 days  (Rose, et  al.,  1976).
IV.  EFFECTS
     A.   Carcinogenicity
          Three studies have reported data  concerning the
carcinogenicity of TCDD. Van Miller, et al. (1977)  fed rats
dietary levels of TCDD ranging from 0.001  to 1000 jug/kg
of diet for up to 78 weeks.  In 50 animals  receiving  diets
ranging from 0.005 jug/kg to 5 jug/kg 13  benign  and  15  malig-
nant tumors were observed~  No tumors were  found in controls
or those fed a. dietary level of 0.0001 pg/kg.  Animals fed
diets of 50 jug/kg or more died between  the  second  and fourth
week of treatment.
          Toth, et al. (1977) administered  TCDD  to  mice
at levels of 0.007, 0.7, and 7 ,ug/kg per week  for  12  months.
No tumors were noted at any dose.

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          Kociba, et al.  (in. press)  administered 0.1,  0.01,
and 0.001 jug/kg ofTCDD per kg of body weight  to  male and
female rats.  Males at the- 0.1 ;ug/kg dose  exhibited  a statis-
tically significant increased incidence of. squamous  cell
carcinomas of the hard palate (4 out of 50) and  of the tongue
(3 out of 50).  Mo carcinomas were observed in the male
controls (0 .out of 35)..  Females at  the 0.1 jug/kg dose had
a statistically significant  increase in incidence of car-
cinomas at three sites:  -squamous cell carcinoma of  the
hard palate  (4 out of 49), squamous cell carcinoma of the
lung (7 out of 40), and hepatocellular carcinoma of  the
liver (11 out of 49).  Only  one carcinoma  of  these three
sites occurred in the female controls (1 out  of  86),  and
that was hepatocellular carcinoma of the liver»   Five  sites,
pancreas., adrenal gland, pituitary gland,  uterus,, and mam-
mary gland, had. a statistically significant, decrease in
their tumor incidence at certain dose levels  (Kociba,  et
al.., in press) .
     B.    Mutagenicity
          Multiple, oral doses of TCDD over  5  weeks resulted
in vacuolization of liver cell nuclei, increased mitotic
rate, and a polyploid chromosome number (Vos, et al.,  1974).
          TCDD administered  by intubation  intraperitonealy,
or orally did not cause chromosomal aberrations  in bone
marrow cells  (Green and Moreland 1975).   However, repeated
                                                          •
dosing of TCDD over 13 weeks produced an increase in chro-
mosomal breaks in rat bone marrow (Green,  et  al.  1977).
                        I

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          Some studies have been conducted  showing  that
TCDD might be a dominant lethal inducing agent, while others
have found no evidence of this effect  (U.S. EPA,  1979).
         , Bacterial assays with E. coli, and S. typhimurium
have found TCDD to be mutagenic via intercalation with DNA
(Hussain, et al-, 1972).  Some strains of Salmonella, how-
ever, have yielded negative mutagenic results when  tested
(Seiler, 1973),
          Tenchini, et al. (1977)  found no  significant differ-
ences in chromosome .number or chromosomal abnormalities
in maternal or abortive fetal samples from  pregnant women
exposed to TCDD during the explosion of a 2,4,5-T factory
in Italy.
     C. -  Teratogenicity
          Teratogenic effects from TCDD have been reported
in several studies.  Both teratogenic and fetotoxic effects
were observed in mice and cats administered 2,4,5-T contain-
ing 30 ppra TCDD (Courtney, et al., 1970).   Smith, et al.
(1976) found the incidence of cleft palate  to be  signifi-
cantly higher in mice receiving 1 /ag/kg and 3 /ag/kg per
day of TCDD for 10 days during gestation.  At 3 jug/kg, the
incidence of bilateral dilated renal pelvis among fetuses
was also significantly greater.  TCDD levels of 0.12S to
2.0 pg/kg/day given orally to cats, on days 6 to 15 of gesta-
tion produced dose-related increases in fetal mortality,
fetal intestinal hemorrhages, and early and late  cesorptions
(Sparschu,  et al., 1971).
                           AS1*--?

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      D»   Other Reproductive Effects
           Pertinent information could not be  located  in
 the available literature.
      E.   Chronic Toxicity
           Chronic studies involving administration of TCDD
 to rats> guinea, pigs and mice, have reported  toxic effects.
 to the liver and thymus (U.S. EPA, 1979).  Female rhesus.
 monkeys fed a diet containing 500 ppt TCDD for up to nine
 months, exhibited symptoms of facial hair and eyelash loss,
 edema, accentuated hair follicles, and dry scaly skin (Allen,
 et al., 1977).
           A large number of studies have reported the inci-
 dence of chloracne among workers exposed to TCDD during
 the production of 2,4,5-trichlorophenol  (TCP,. 2,4-D or 2,4,5-
 T) (U.S.  EPA, 1979).  Other chemical manifestations among
 exposed workers include muscular weakness, loss of. appetite
'and weight, sleep disturbances, orthostatic hypotension,
 abdominal pain, liver impairment, hyperpigmentation of the
 skin,, hirsutism, and. psychopathological changes (U.S. EPA,
 1979) ..
      F.   Other Relevant Information
           No synergistic effect was detected when 2,4,5-
 T and TCDD were administered to mice alone, or in combina-
 tion with each other (U.S. EPA, 1979).  Both compounds are
 capable of producing cleft palates and kidney anomalies
 in fetuses.
           The International Agency for Research on Cancer
 (1977) has reviewed the literature and concludes that TCDD
 is a potent inducer of hepatic and renal microsomal drug

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metabolizing enzymes.  TCDD  intoxication  results  in  a  marked
increase in the cellular smooth endoplasmic  reticulum  con-
tent of hepatic and renal cells.  This compound is also
capable of simultaneously activating and  suppressing certain
microsome associated foreign compound and steroid-hormone-
metabolizing enzyme systems.  It has' been found to increase
the activity of renal and hepatic glutathione-S-transferase,
*
and hepatic 
-------
     D.   Residues
          TCDD has a. high affinity  for  the tissues of aquatic
species.  Isensee and Jones  (1975)  conducted  a  model fresh—
water ecosystem study'on TCDD and observed bioconcentration
factors between 3,600 and 26,000 over a 3  to  31 day period.
The highest bioconcentration factors were  reported for. Dyphnia
maqna  (26,000), the mosquito fish,  Gambusia affinis (25,000),
and the snail,- Physa sp.  (20,000).
VI.  EXISTING GUIDELINES AND STANDARDS
     A..   Human .
          The calculated acceptable daily  intake (ADI)  for
          -4
TCDD is 10   jug/kg/day.  This ADI does  not consider TCDD
to be a known or suspected carcinogen (NRC, 1977) .
          The draft, ambient, water quality criterion has
been set by the U,S. EPA  (1979) at  levels  intended to reduce
the human carcinogenic risk to rhe-  range of 10"3,  10~ ,
and 10~ .  The. corresponding draft  criteria are 4.55 x 10~
7 jug/1, 4.55 x 10~8 ;ug/l, and 4.55  x. 10~9 pg/1,  respectively.
     B.   Aquatic
          No drafted criterion is available to  protect fresh
and saltwater species from TCDD toxicity.

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                            TCDD

                         REFERENCES                         i

Allan, J.R., et al.  1977..  Morphological  changes  in  monkeys
consuming a d-iet containing low levels of  TGDD.   Pood Cosrae-t.
Toxicol. " 15: 401.

Courtney, K. Diane,- -et al.  1970,  Teratogenic  evaluation of
2,4,5-T..  Science' 168: 864-.

Crosby, D.G., .and A.S.. Wong.  1977.  Environmental degrada-
tion of 2,3,7,8-tetrachlorodibenzo-p-dioxin.  Science 195:
1337.

Crosby, D.G., et al.  1971.  Photodecomposition  of chlori-
nated dibenzo-p-dioxin.  Science  173: 748-.

Green, S., and F.S. Morelandi  1975.  Cytogenetic  evaluation
of several dioxins in the rat- .Toxicol. Appl.  Pharmacol.
33: 161.  •

Green, S., et al.  1977.- Cytogenetic effect of
2,3,7,8tetrachlorodibenzo-p-diox_in on rat  bone marrow cells.
FDA By-lines  7r 292.-  Food. Drug Admin., Washington,  D.C.

Hussain, S., et al.  1972.  Mutagenic effects of TCDD on
bacterial systems.  Ambico.  1:32.

International Agency for Research on Cancer, Chlorinated.
1977.  Dibenzodioxins.  IARC Monographs on The Evaluation  of
Carcinogenic Chemicals to Man.  Vol. 15, Lyon, France.

Isensee, A.R., and G.E. Jones.  1971.  Absorption  and trans-
location of  coot and foliage applied 2,4-dichlorophenol,  2,7—
dichlorodibenzo-p-dioxin and TCDD.  Jour.  Agri.  Food  Chem.
19: 1210.

Isensee, A.R., and G.E. Jones.  1975.  Distribution of
2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in aquatic model
ecosystems.   Environ. Sci. Technol.  9: 668.

Kearney, P.C., et al.  1972.  Persistence  and metabolism of
chlorodioxins in soils.  Environ. Sci. Tech.  5: 1017.

Kocher, et al.  1978.  A search for the presence of
2,3,7,8tetrachlorodibenzo-p-dioxin in beef fat.  Bull.
Environ. Cont. Toxicol.  19: 229.

Kociba, et al.  Toxic, and App. Pharm. In press.

Matsumura, F., and H.J. Senezet.   1973.  Studies on the bio-
accumulation and microbial degradation of TCDD.  Environ.
Health Perspect.  5: 253.

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Miller,. R.A., et al.   1973..   Toxicity of  2,3,7,8-tetrachloro-
dibenzo-p-diox in (TCDD)  in aquatic  organisms.   Environ..
Health Perspect.   5:  177..

National'Research  Council, Safe  Drinking  Water Committee.
1977.  Drinking water and health: Part II.  Natl.  Acad. Sci.,
Washington, D.C.

Piper, W.N., et al.   1973.   Excretion and tissue distribution
of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the  rat.   Environ.
Health Perspect..   5:  241.

Rose., J.Q., et al.. 1976..  The. fate of 2,3,7,8-TCDD following
single and  repeated oral doses to the rat.  Toxicol. Appl.
Pharmacol.  36: 209.

Seiler, J.P.  1973.   A survey on -the  mutagenicity of various
pesticides.  Experientia  29: 622.

Smith, F.A., et al.   1976.   Teratogenicity  of  TCDD  in DF-1
mice.  Toxicol. Appl.  Pharmacol^ 38:  519

Sparschu, G.L.., et al- 1971.  Study  of the teratogenicity  of
TCDD in the rat..   Food Cosmet. Tech*   9: .405.

Tenchini, M..L., et al.  19.77.. Approaches  to examination of
genetic damage after  a major hazard in the  chemical, industry:
Preliminary cytohenic findings on TCDD exposed subjects after
the Seveso  accident.   Presented  at  the Expert  Conference on
Genetic Damage Caused by Environmental Factors,  Oslor Norway,
May 11-13,  1977.

Toth, K., et al.   1977.  Carcinogenic bioassay of the herbi-
cide 2,4,5-TCPE with  different TCDD content in Swiss mice.
In: International  Conference on  Ecological  Perspectives on
Carcinogens and Cancer Control,  Cremona,  1976,  Basel, Karger,
A.G..

U.S. EPA.   1979.   2,3,7,8-Tetrachlorodibenzo-p-dioxinr  Am-
bient Water Quality Criteria. (Draft).

Van Miller,. J.P. ,  and  J.R.. Allen.   1977.  Chronic toxicity  of
2,3,7,8-tetrachlorodibenzo-p-dioxin in rats.   Fed.  Proc.  35:
396.

Van Miller, J.P.,  et  al..  1976.  Tissue distribution and ex-
cretion of  tritiated  tetrachlorodibenzo-p-dioxin in  nonhuman
primates and rats.  Food Cosmet. Toxicol.   14:  31.

vos, J.G.,  et al.  1974.  Toxicity  of  2,3,7,8-tetrachlorodi-
benzo-p-dioxin in  C-57BL-6 mice.  Toxicol.  Appl.  Pharmacol.
29: 229.

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                                       No,  156
     1,1,1,2-Tetrachloroethane

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20A60

           APRIL 30, 1980

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                          DISCLAIMER
     This report, represents  a survey of the- potential health
and environmental hazards from exposure to the subject chemi-
cal*  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short- profile-
may not reflect  all available  information  including all the
adverse health  and  environmental impacts  presented by the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

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                                    SUMMARY

     1,1,1,2-Tetrachloroethane  is  potentially formed  during chlqrination  of
drinking  water  and  has. been  identified  at a  concentration of  0.11 /jg/1.
Although  inhalation  is the major  route of exposure  to chlorinated ethanes,
specific  information on  1,1,1,2-tetrachloroethane inhalation is  not avail-
able.
     Literature  reporting  adverse occupational  exposures  to  this  chloro-
ethane cannot be  found.   Animal experiments  measuring the acute and subacute
effects indicate,  however, that chronic  exposure may  produce liver damage.
1,1,1,2-Tetrachloroethane  is  currently  being tested  by the  National Cancer
Institute   for   possible   carcinogenicity.   The  compound   not  mutagenic
according  to one report.   Data  could  not be   located  in  the  available
literature showing it to be teratogenic.       •       .                  . .
     Pertinent information  could not be  found  in the available  literature
regarding the adverse effects of this compound on aquatic animals-or plants.

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                           1,1,1,2-TETRACHLOROETHANE
 I.    INTRODUCTION
      This  profile  is based  primarily on the  Ambient Water Quality  Criteria
 Document for Chlorinated Ethanes  (U.S. EPA, 1979a).
      The chloroethanes- are  hydrocarbons  in which one or more of  the  hydrogen
 atoms  of ethane are  replaced by  chlorine atoms.   In general,  water  solu-
 bility  and vapor pressure  decrease with increasing chlorination, while  dens-
 ity and melting point increase.   1,1,1,2-Tetrachloroethane (molecular weight
 167.9)  is  a  liquid at  room temperature  with a  boiling point  of 129°C,. a
melting  point of  -68°c,  a  specific  gravity  of  1.553,  and a  solubility in
water of 2.85 mg/1 (U.S. EPA, 1979a).
     1,1,1,2-Tetrachloroethane  is used as  a  solvent and  in  the manufacture
of a  number of widely  used products, as  are  the other  chloroethanes  (U.S.
EPA,  1975).  In general,  these compounds  form  azeotropes with  water  (Kirk
and Othmer,  1963)  and  are  very soluble  in organic  solvents  (Lange, 1956).
Pearson and McConnell  (1975) were unable to  demonstrate microbial degrada-
tion  of these compounds,, but did report  chemical  degradation.  For  a more
general, treatment  of  the  chlorinated ethanes  as  a class,  the  reader  is
referred to the EPA/ECAO  Hazard. Profile on  Chlorinated Ethanes  (U.S.. EPA,
1979b)..
II.  EXPOSURE
     1,1,1,2-Tetrachloroethane is  potentially formed during chlorination  of
drinking water and has- been  identified at a concentration of 0.11 jug/1 (U.S.
EPA,   1974).   Information on the levels of  1,1,1,2-tetrachloroethane  in food
are not available although other chloroethanes have  been  detected (U.S.  EPA,
1979a).  Inhalation is  the major  route of exposure  to chlorinated ethanes.
However, specific  information, on  1,1,1,2-tetrachloroethane  exposure is  not

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available  (U.S.  EPA,  1979a).   As with  most solvents,  chloroethanes can  be
absorbed through  the skin.  This is  not,  however,  a major route of  exposure
(U.S. EPA, 1979a).
     The U.S.  EPA (1979a) has estimated  a weighted average bioconcentration
factor of  18 for .1,1,1,2-tetrachloroethane  for the edible  portions  of fish
and  shellfish  consumed by Americans.   This  value was  based on an estimated
steady-state bioconcentration  factor  of 62,  which was  determined  from   an
octanol/water partition coefficient of 457.
III. PHARMACOKINETICS
     A.   Absorption
          Specific  information  on  the  absorption  of  1,1,1,2-tetrachloro-
ethane is not available.   In general,  the chloroethanes are absorbed rapidly
following ingestion or inhalation (U.S. EPA,  1979a).
     8.   Distribution
          Inhalation  or  ingestion  of  1,1,1,2-tetrachloroethane results  in
the presence of high levels of solvent  in  the  fetuses  of the exposed animals
(Truhaut, et al. 1974).   Other studies  indicate a widespread distribution of
chloroethanes throughout the body after administration (U.S. EPA, 1979a).
     C.   Metabolism
          After  oral  administration  to  rats,   guinea  pigs,  and  rabbits,
1,1,1,2-tetrachloroethane  underwent  hydrolytic  dehalogenation resulting  in
formation of trichloroethanol,  which was  eliminated primarily in  the  urine
in the form  of  a conjugated glucuronic derivative,  urochloralic  acid.   Oxi-
dation to trichloroacetic acid was considerable  only in  rats (Nguyen, et al.
1971; Truhaut and Nguyen,  1973).   In the latter  study  monochloroacetic acid
and mercaptan derivatives were not found  in  the urine.   The only halogenated

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compound  found  in the  expired  air  was  untransfarmed  1,1,1,2-tetrachlara-
ethane.   Trichloroethanol. and trichloroacetic acid have  also  been  identified
in  the urine 'of  rats  following  interperitoneal  (i.p.)  injection or  vapor
inhalation  of 1,1,1,2-tetrachloroethane (Ikeda  and. Ohtsuji, 1972), and  have
been  identified  in the urine of mice following  i.p.. injection of  the  parent
compound  (Yllner,  1571).
          In  general, the  metabolism of  chloroethanes  involves  both enzy-
matic  dechlorination  and hydroxylation  and .non-enzymatic oxidation   (U.S.
EPA,  1979a)..   Oxidation reactions may  produce unsaturated metabolites.which
are then  transformed  to the alcohol and ester (Yllner,  1971).
     0.   Excretion
          Murine  studies  show  that,  after i.p. injection of 1,1,1,2-tetra-
chloroethane, approximately 78  percent  of the dose  is  excreted  in 72  hours;
from 21 to  62 percent of this  dose is excreted  in the  breath  and from 18 to
56 percent  as  metabolites in the  urine  (Yllner,  1971).   Other  studies  also
indicate  that   1,1,1,2-tetrachloroethane   is   excreted   in  the  urine  as
metabolites and in the expired breath as the parent compound (see above).
IV.  EFFECTS
     A.   Carcinogenicity
          1,1,1,2-Tetrachloroethane  is  currently  being  tested  by NCI  for
possible  carcinogenicity;  results are  not available   (NTCTP,  I960).-  Other
information    relative     to     the     potential     carcinogenicity     of
1,1,1,2-tetrachloroethane was not located in the available literature.
     B.    Mutagenicity
          Simmon,  et  al.  (1977) tested 71  chemicals identified in  the W.S.
drinking  water  for  mutagenesis  with an  Ames  Salmonella/microsome  assay.
1,1,1,2-Tetrachloroethane was found not  to be  mutagenic  in this study.


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     C.   Teratogenicity and Other Reproductive  Effects
          The  isomer of 1,1,1,2-tetrachloroethane, syn-tetrachloroethane,  is
a  weak teratogeh in  two  strains of mice  (Schmidt and Reiner,  1976).   Both
tetrachloroethanes  are embryotoxic  (Schmidt and  Reiner,  1976;  Truhaut,  et
al., 1974).  Other pertinent data have not been  found.
     0.   Chronic Toxicity
          Adverse occupational exposure  to 1,1,1,2-tetrachloroethane has not
been  reported by  NIOSH.  (U.S.  EPA,  1979a).   Animal  experiments measuring
acute,  and subacute  effects indicate  that  chronic "inhalation  exposure may
produce liver damage  (see below).
     E.   Acute and Subacute Toxicities
          At 24 hours after, the oral administration  of Q',5 g 1,1,1,2-tetra-
chloroethane/kg to  rabbits, the blood  cholesterol  and  total  lipid  levels
were  increased and   the  glutamic-pyruvic transaminase,  glutamic-oxalacetic
transaminase, creatine  phosphokinase,  lactate dehydrogenase,  and a-hydroxy-
butyrate dehydrogenase  activities  were enhanced.   Except for creatine phos-
phokinase, these  enzyme levels  remain elevated  at 72  hours  after poisoning
(Truhaut, et al.   1973).   Subsequent studies  by  this  research  group  found
that in  rabbits,  1,1,1,2-tetrachloroethane was  only slightly  irritating to
the  skin  and ocular  mucous membrane, and its  cutaneous LD    was 30  g/kg.
Its acute toxicity by inhalation, for an exposure  of  4  hours, was similar in
rats  and  rabbits,   with   the  LC5Q  being  2500 mg/m3.    The  oral  LD5Q
values in rats and mice were 800 and  1500  mg/kg, respectively.  Histological
examination  revealed hepatotoxic  activity,  including   formation of  micro-
vacuolizations  and  centrolobular  necrosis.   1,1,1,2-Tetrachloroethane .was
from two  to  three times less  toxic than 1,1,2,2-tetrachloroethane (Truhaut,
et al. 1974).

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          Recent  studies exploring subacute effects  indicate that in  female
Wistar  rats,  1,1,1,2-tetrachloroethane (0.30 g/kg, 5 days/week, for 2  weeks,.
orally)  induced hepatic steatosis  by accumulation  of txiglycerides,  accom-
panied  by a  decrease  in liver  lactate  dehydrogenase, malate dehydrogenase,
and  glutamic  pyruvic transaminase  activities.   The tetrachloroethane  caused
no  changes in  the  liver of  male  rats (Truhaut,  et al.  1975).  However,
another  team  of investigators  found that 1,1,1,2-tetrachloroethane (from  100
to  800  pmoles/kg/day  for  7  days,  i.p.)   to  male  rats   increased  liver
succinate   dehydrogenase,.  acid  phosphatase   and   glucose  6-phosphatase
activities  and  decreased  liver DMA  content.   In  addition,  the  white cell
count was increased and  the  red  cell count  and. blood  cholesterol  content
were decreased  (Chieruttini, et al. 1976).
V.   AQUATIC TOXICITY
     Pertinent  data could  not be  located  in  the available  literature  re-
garding  either  the  acute  and  chronic  toxicity to  aquatic  animals, or  the
aquatic residues of 1,1,1,2-tetrachloroethane.
VI.  EXISTING GUIDELINES AND STANDARDS
     Guidelines  for occupational  exposure  to  1,1,1,2-tetrachloroethane  do
not exist (International  Labor Office, No.  77,  1977;  NIOSH,  1978);  however,
1,1,2,2-tetrachloroethane exposure  is  limited  in the workplace to 5  ppm  (35
mg/cu m) as an 8-hour time-weighted average (TWA) concentration.

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                           1,1,1,2-TETRACHLOROETHANE

                                  REFERENCES
Chieruttini, M.E.,  et al.  1976.  The  toxicology of the tetrachloroethanes.
Br.. Jour. Pharmacol.  57: 421.

Ikeda,  M.   and  H.  Ohtsuji.   1972.   Comparative study  of the  excretion of
Fujiwara reaction-positive  substances in urine  of humans  and  rodents given
trichloro- or  tetrachloro-derivatives  of • ethane and  ethylene.   Br.  Jour.
Ind. Med.  29: 99.

International Labor Office.   1977.   Occupational exposure limits for airborn
toxic substances  — A tabular compulation  of  values for selected  countries.
Occupational Safety and Health Series, NO.  37.  Geneva .

Kirk, R.  and 0.  Othmer.   1963.  Encyclopedia of chemical technology.  2nd
ed. John Wiley and Sons, Inc., New York.

Lange,  N.   (ed.)    1956.   Handbook  of  chemistry.   9th  ed.   Handbook Pub-
lishers, Inc., Sandusky, Ohio.

National  Toxicology  Carcinogenesis   Testing  Program.   1980.   Chemicals  on.
Standard Protocol.

National Institute  for Occupational. Safety and  Health.   197&.   Current in-
telligence bulletin, No. 27, OHEW Pub. No.  78-181, p. 4..

Nguyen,  P.,  et  al.   1971.   1,1,1,2-Tetrachloroethane metabolism.   C.R. Acad.
Sci., Ser.  0. 272: 1173.

Pearson, C.R.,   and  G.. McConnell.   1975.   Chlorinated  hydrocarbons  in the
marine environment.  Proc. R. Soc. London.  Ser.  8.  189: 305.

Schmidt and Reiner.   1976.   The  embryotoxic and teratogenic effect of  tetra-
chloroethane - experimental studies.   Biol. Rundsch.  14: 220.

Simmon,   V., et  al.  1977.   Mutagenic  activity  of  chemicals  identified  in
drinking water.  Dev. Toxicol. Environ.  Sci.  2:  249.

Truhaut,   R.   and    P.   Nguyen.    1973.    Metabolic   transformations   of
1,1,1,2-tetrachloroethane  in  the rat,  guinea pig,  and  rabbit.   Jour. Eur.
Toxicol.  6: 211.

Truhaut, R.,  et  al.   1973.   Serum  enzyme  activities and  biochemical blood
components  in  subacute  1,1,1,2-tetrachloroethane poisoning  in  the  rabbit.
Jour. Eur.  Toxicol.  6: 81.

Truhaut,  R.,  et al.   1974.   Toxicological  study  of  1,1,1,2-tetrachloro-
ethane.   Arch. Mai. Prof. Med. Trav.  Secur. Soc.  35: 593.

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Truhaut,  R.,  et  al.   1975.   Preliminary biochemical  study of  the hepato-
toxicity  of 1,1,1,2-tetrachloroethane  in the  Wistar  rat.   Effect  of sex.
Eur. Jour. Toxicol. Environ. Hyg.  3: 175.

U.S. EPA.   1974.   "Draft analytical report  - New Orleans  area  water supply
study," EPA 906/10-74-002.   Lower Mississippi  River  Facility,  Slidell, La.,
Surveill.  Anal. Oiv. Region VI, Dallas, Tex.

U.S. EPA;   1975.   Identification of organic compounds  in effluents from in-
dustrial sources.  EPA. 560/3-75-002.

U.S.  EPA.   1979a.   Chlorinated Ethanes:  Ambient. Water  Quality  Criteria.
(Draft)

U.S. EPA.  1979b..  Hazard Profile:  Chlroinated Ethanes (Draft).

Yllner,  S.   1971.   Metabolism  of  1,1,1,2-tetrachloroethane  in the  mouse.
Acta Pharmacol. Toxicol.  29(5-6):  471-480, 1971.

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                                    No.  157
     1,1,2,2-Tetrachloroethane

  Health and Environmental Effects
U.S.  ENVIRONMENTAL PROTECTION AGENCY
      WASHINGTON, D.C.  20460

          APRIL 30,  1980
          1*7-1

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such  sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental  impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                       SPECIAL. NOTATION


U.S.. EPA's Carcinogen Assessment Group (GAG) has evaluated
1,1,2,2,-tetrachloroethane and has found sufficient evi-
dence to indicate that this compound is carcinogenic.

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                  1,1,2,2-TETRACHLOROETHAN E
                           SUMMARY
     An increased incidence of hepatocellular  carcinomas  has
been shown in mice following oral administration of  1,1,2,2-
tetrachloroethane.' Mutagenic effects have been reported  in
the Ames Salmonella assay and in E~ coli..  There is  no  avail-
able evidence to indicate that. 1,1,2,2-tetrachloroethane"pro-
duces teratogenic effects.  Occupational exposure  to 1,1,2,2-
tetrachloroethane has produced several toxic effects includ-
ing neurological symptomsr liver and kidney damage,  pulmonary
edema, and fatty degeneration of heart muscle.
     The toxicity of 1,1,2,2-tetrachloroethane has been exam-
ined in one species each of freshwater and marine  fish,  in-
vertebrates, and plants..  Freshwater invertebrates appear to
be the most sensitive species examined, with acute toxic- con-
centrations of 9,320 ug/1 being reported.

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                  1,1,2,2-TETRACHLOROETHANE
I.   INTRODUCTION
     This profile is based on the Ambient Water Quality Cri-
teria Document for Chlorinated Ethanes  (U.S. EPA,  1979a).
     The chloroethanes are hydrocarbons  in which one  or more
of the hydrogen atoms of ethane are replaced by chlorine
atoms-  In general, water solubility and vapor pressure
decrease with increasing chlorination, while density  and
melting point increase.  1,1,2,2-Tetrachloroethane  (molecular
weight 167-9) is a liquid at room temperature with  a  boiling
point of 146.3°C, a melting point of -36°C, a specific
gravity of 1.596, and a solubility in water of 2.9  gm/1 (U.S.
EPA, 1979a) .
     The chloroethanes are used as solvents, cleaning and de-
greasing agents, and in the chemical synthesis of a number of
compounds-
     The chlorinated ethanes form azeotropes with water (Kirk
and Othmer, 1963).  All are very soluble in organic solvents
(Lange, 1956)..  Microbial degradation of the chlorinated
ethanes has not been demonstrated (U.S. EPA, 1979a).  For
additional information regarding the chlorinated ethanes  in
general, the reader is referred to the Hazard Profile on
Chlorinated Ethanes (U.S- EPA, 1979b).
II.  EXPOSURE
     The chloroethanes present in raw and finished  waters are
due primarily to industrial discharges.  Small amounts  of
chloroethanes may be formed by chlorination of drinking water
or treatment of sewage.  Atmospheric chloroethanes  result


                       ..  -/fo

-------
from evaporation of volatile  chloroethanes  during use as
degreasing agents or  in dry cleaning operations  (U.S.  EPA,
1979a). '  '
     Routes of human  exposure  to  chloroethanes  include water,
air, contaminated foods and. fish, and dermal  absorption-
Fish and shellfish have shown  level's of  chloroethanes in the
nanogram range (Dickson and Riley, 1976)..   Information on the
levels of 1,1,2,2-tetrachloroethane  in foods  is  not avail-
able.
     The EPA (1979a)  has estimated a weighted average biocon-
centration factor for 1,1,2,2-tetrachloroethane  to  be 18 for
the edible portions of fish and shellfish consumed  by Ameri-
                                                 i
cans..  This estimate  was based on steady-state bioconcentra—
tion studies in the bluegill.
III.. PHARMACOKINETICS
     A-   Absorption
          The chloroethanes are absorbed rapidly following
ingestion or inhalation (U.S.  EPA, 1979a).  Morgan,  et al.
(1972) have determined that 1,1,2,2-tetrachloroethane  has a
high octanol/water partition coefficient, high rate  of pul-
monary absorption, and low rate of elimination by exhalation.
     B..   Distribution
          Pertinent data could not be located in the avail-
able literature on 1,1,2,2-tetrachloroethane.  The  reader is
referred to a more general treatment of  chlorinated  ethanes*
(U.S. EPA, 1979b), which indicates widespread distribution  of
these compounds throughout the body..

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      C.    Metabolism
           The metabolism of chloroethanes involves both enzy-
 matic dechlorination and hydroxylation and non-enzymatic oxi-
 dation (U.S.  EPA,  1979a).  Oxidation reactions may produce
 unsaturated metabolites which are then transformed to the
•alcohol  and ester  (Yllner,  1971).  Trichloroethanol and tri-
 chloro acetic acid have been identified in the urine of rats
       k
 following inhalation of 1,1,2,2-tetrachloroethane vapor
 (Ikeda and Ohtsuji, 1972).   Metabolism of ..this compound ap-
 pears to involve the activity of the mixed-function oxidase
 system (Van Dyke and Wineman, 1971).
      D.    Excretion
       \
           The chloroethanes are excreted primarily in the
 urine and expired  air.   Murine studies indicate that, after
 intraperitoneal (i.p.)  injection of 1,1,2,2-tetrachloro-
 ethane,  approximately 80 percent of the dose is excreted in
 72  hours. Half of this dose is excreted as carbon dioxide in
 the breath and one-fourth as metabolites in the urine (Yllner,
 1971).  Human studies (Morgan, et al.  1972)  indicate that
 after inhalation exposure of 1,1,2,2-tetrachloroethane the
 amount expired in  the breath is less than that observed in
 animal studies,  although a  different radioactive tracer was
 used.
 IV.  EFFECTS
      A.    Carcinogenicity
           Results  of a  National Cancer Institute (NCI) car-
 cinogenesis bioassay for 1,1,2,2-tetrachlocoethane show that
 oral administration produced an increased incidence of hepato-

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cellular carcinomas in exposed mice  (NCI, 1978).  No  sta-
tistically significant tumor  increase was seen  in cats.
     B.  .Mutagenicity
          The mutagenic activity of  1,1,2,2-tetrachloroethane
has been shown in the Ames Salmonella assay and  in  a  ENA
polymerase-deficient strain of JJ. coli  (Brem-, et al»,  1974).
     C.   Teratogenictty and Other Reproductive Effects
          Embryo toxicity and weak teratogenicity have been
reported in two strains of mice exposed with 1,1,2,2-tetra-
chloroethane (Schmidt and Reiner, 1976) .  Other pertinent  in-
formation could not be located in the available literature.
D..   Chronic Toxicity
          Occupational exposure to 1,1,2,2-tetrachloroethane
has produced toxic effects including neurological symptoms,
liver and kidney damage, pulmonary edema, and fatty degenera-
tion of heart muscle (U.S. EPA, 1979a) .
          Animal experiments have indicated that chronic  in-
halation exposure may produce liver  and kidney degeneration
(U.S.. EPA, 1979a) .
V.   AQUATIC TOXICITY
     A.   Acute Toxicity
          Toxicity studies on one species from each category
of freshwater and marine fish and invertebrates have  been  re-
ported  (U.S. EPA, 1978)..  In freshwater fish, the study
yielded a 96-hour static LCsg. value  of  21,300 ug/1  for
the blueg.ill (Lepomis macrochirus) .  For freshwater inverte-
brates, the study yielded a 48-hour  static LCgQ value  of

-------
9,320 ug/1 for the caldoceran Daphnia- magna.  In marine fish


and invertebrates, the studies yielded a 96-hour static LC50


value of 12,300 ug/1 for the sheepshead minnow (Cyprinodon


variegatus), and of 9,020 ug/1 for the mysid shrimp (Mysi-


dopsis bahia).


->••	 B.   Chronic Toxicity


          Pertinent information could not be located in the
                        •

available literature.


     C.   Plant Effects


          When the freshwater algae Selenastrum capricornutum


was tested for adverse effects of 1,1,2,2-tetrachloroethane


on chlorophyll and cell numbers EC50 values of 136,000


and 146,000 ug/1 were obtained.  When the marine algae SSkele-


tonema costatum was tested for these adverse effects, 96-hour


ECgg values were 6,440 and 6,230 ug/1, respectively.


     D.   Residues


          A bioconcentration value of 8 was reported for the


bluegill (U.S.. EPA, 1979a).


VI.  EXISTING GUIDELINES AND STANDARDS


     Neither the human health nor aquatic criteria derived by


U.S. EPA (1979a), which are summarized below, have gone


through the process of public review; therefore, there is a


possibility that these criteria will be changed.


     A.   .Human


          Based on the NCI carcinogenic data, and using a


linear, nonthreshold model, the U.S. EPA (1979a) has esti-


mated the level of 1,1,2,2-tetrachloroethane in ambient water
                                 /S7-?

-------
that will result in an additional cancer  risk  of  10~5  to
be 1.8 ug/1.
          The exposure standard determined  by  OSHA for 1/1,-
2,2-tetrachloroethane. is 5 ppm as an eight-hour time-weighted
average concentration..
     B^   Aquatic                                	
          The draft, criterion for protection of freshwater
aquatic life is 170 ug/1 as a 24-hour average, not to  exceed
380 ug/1-  The draft criterion to protect, marine  life  from
1,1,2,2-tetrachloroethane is 70 ug/1 as-a 24-hour average,
not to exceed 160 ug/1 (U»S. EPA, 1979a) .

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                           1,1,2,2-TETRACHLOROETHANE

                                  REFERENCES
Brem, H.,  et  al.   1974.  The mutagenicity  and ONA-modifying effect of halo-
alkanes.  Cancer Res.  34: 2576.

Oickson, A.G.  and  J.P. Riley.  1976.  The  distribution of short-chain halo-
genated aliphatic hydrocarbons  in  some marine organisms.  Mar. Pollut. Bull.
79: 167."

Ikeda,  M.  and  H.  Ohtsuji.   1972.   Comparative study  of the  excretion of
Fujiwara reaction-positive  substances in urine  of humans and rodents given
trichloro- or  tetrachloro-derivatives of  ethane  and  ethylene.   Br.  Jour.
Ind. Med.  29: 99.
                                                 «.
Kirk, R.  and D.  Othmer.   1963..   Encyclopedia of chemical  technology.   2nd
ed. John Wiley and Sons, Inc., New York.

Lange,  N.  (ed.)   1956.  Handbook  of chemistry.  9th  ed.   Handbook Publish-
ers, Inc.,  Sandusky,  Ohio.

Morgan,  A., et al.   1972.  Absorption of halogenated  hydrocarbons and their
excretion  in  breath  using chlorine-38 tracer techniques.  Ann.  Occup.  Hyg.
15: 273.

National Cancer Institute.   1978.   Bioassay of 1,1,2,2-tetrachloroethane for
possible carcinogenicity.  Natl. Inst. Health,  Natl. Cancer  Inst. OHEW Pub I.
NO. (NIH) 78-827.   Pub. Health Serv., U.S. Oept. Health Edu. Welfare.

Schmidt  and  Reiner.   1976.    The  embrvotoxic  and   teratogenic  effect "of
tetrachloroethane - experimental studies.  Biol. Rudsch 14: 220.

U.S.  EPA.   1978.   In-depth  studies on  health and environmental  impacts of
selected water pollutants.   Contract NO.  68-01-4646,  U.S.  Environ.  Prot.
Agency.

U.S.  EPA.    1979a.   Chlorinated  Ethanes:  Ambient Water Quality  Criteria.
(Draft)

U.S.  EPA.   1979b.   Environmental  Criteria and Assessment  Office.   Chlori-
nated Ethanes: Hazard Profile.  (Draft)

Van Dyke,  R.A. and C.G. Wineman.   1971.  Enzymatic  dechlorination: Oechlor-
ination  of  chloroethanes  and  propanes   in  vitro.    Biochem.  Pharmacol.
20: 463.                                    ~~

Yllner,   S.    1971.    Metabolism   of  l,l,2,2-tetrachloroethane-14C  in  the
mouse.  Acta. Pharmacol. Toxicol.   29: 499.

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                                    No.  158
      Tetrachloroethylene

Health and Environmental Effects
 - ENVIRONMENTAL PROTECTION AGENCY
     WASHINGTON, D.C..  20460

         APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal*  The information contained in the report is drawn chiefly
from secondary  sources  and available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect'all available  information  including all the
adverse health  and  environmental impacts  presented by  the.
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy.
                            .1 fr 7 V
                           TV fl'7

-------
                       SPECIAL NOTATION


U.S. EPA's Carcinogen Assessment Group (GAG) has evaluated
tetrachloroethylene and has found sufficient evidence to
indicate that this compound is carcinogenic.

-------
                      TETRACHLOROETHYLEN E
                            SUMMARY
      Tetrachloroethylene is widespread in the environment,
 and is  found in trace amounts in water, aquatic organisms,
 air,  foodstuffs, and human tissue.  Tetrachloroethylene
 causes  mild intoxication and liver dysfunction following
 chronic exposure to high levels associated with certain in-
 dustries »   Tetrachloroethylene has not been shown to be tera-
 togenic, but it has been shown to be mutagenic in bacterial
 assays  and carcinogenic in mice.
      The bluegill (Lepomis macrochirus) is the most sensitive
 freshwater species to acute tetrachloroethylene toxicity
 with  a  reported 96-hour LC50 of 12,900 ug/1.  In the only
 acute toxicity study for saltwater species the mysid shrimp
.(Mysidopsis bahla) has an observed 96-hour LCgg value of
 10,200  ug/1-  The chronic value for this shrimp is 448 ug/1.
 A  freshwater algae has a reported no-effect concentration of
 tetrachloroethylene at 816,000 ug/1.  A marine alga, however,
 was adversely affected at the considerably lower level of
 10,000  ug/1.  Tetrachloroethylene is only slightly bioconcen-
 trated  by  the bluegill (49 times) after 21 days of exposure,
 and has an elimination half-life of less than  one day.

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I.   INTRODUCTION
     This profile is based on the Ambient Water Quality Cri-
teria Document for Tetrachloroethylene (U.S. EPA, 1579)..
     Tetr'achloroethylene (02^4, 1,1,2,2-tetrachloroethy—
lene, perchloroethylene, PCE? molecular weight 165.85) is a
colorless, nonflammable liquid..  It has the following physi-
cal/chemical properties (Patty, 1963)t
          Melting-Point:          -23.25°C
          Density:                1.623 g/ml
          Vapor Pressure:         19 mm Hg
          Water Solubility:       150 ug/ml
          Octanol/Water
          Partition Coefficient:  339
     Tetrachloroethylene is primarily used as a solvent in
the dry cleaning industry and,, to a lesser extent, as a de-
greasing solvent in metal industries (Windholz, 1976).
II...  EXPOSURE
     The National Organics Monitoring Survey (U.S. EPA, 1978)
detected tetrachloroethylene in 9 out of 105 drinking water
samples between November 1976 and January 1977 (range, <0.2
to 3.1 ug/1;  median <0,2 ug/1).  No data exist for ingestion
of tetrachloroethylene from food for the United States.  How-
ever, in England, tetrachloroethylene concentrations in foods
ranged from nondetectable amounts in orange juice to 13 ugA<3
in butter (McConnel, et al,, 1975).  The U.S. EPA (1979)  has
estimated the weighted bioconcentration factor of tetrachlo-
roethylene to be 110 for the edible portion of consumed fish
and shellfish.  This estimate is based on measured steady—
                                                          »
state bioconcentration studies in bluegills.  Generally,
                       IS?-5

-------
environmental tetrachloroethylene  concentrations in air tend
to be low.  A survey of eight  locations  in the U.S. indicated
concentrations up to 6.7 ug/m^  in  urban  areas and less than
0.013 ug/m-3 in rural areas  (Lillian,  et  al.,  1975).  By far
the most significant exposure  to tetrachloroethylene is in
the industrial environment  (Fishbein,  1976)»  Significant der-
mal exposure would be confined  to  occupational settings.
III. PHARMACOKINETTCS
     A.   Absorption
          Using  inhalation  exposure,  Stewart, et al.  (1961)
found that tetrachloroethylene  reached near steady-state
levels in the blood of human volunteers  with  two hours of
continuous exposure.  However,  steady-state conditions in
this study were  probably obtained  by  a redistribution phenom-
enon, since the  biological  half-life  of  tetrachloroethylene
metabolites in humans has been  measured  to be 144 hours
(Ikeda and Imamura, 1973).
     B.   Distribution
          In humans (McConnell, et  al.,  1975)  and rats (Savo—
lainen, et al.,  1977), tetrachloroethylene tends to accumu-
late in the body fat, and to a  lesser extent  in  the brain and
liver..  Measurements in the rat suggests  that the level of
PCS in the liver and blood  remains  constant after three hours
of exposure.
     C.   Metabolism
          In a qualitative  sense, metabolic products  appear
to be similar in humans (ikeda, et  al.,  1972;   Ikeda,  1977)
and experimental animals (Yllner, 1961;  Daniel,  1963;  Ikeda

-------
and Ohtsuji, 1972).  The metabolism of  tetrachloroethylene
leads to the production of trichloroacetic  acid,  and  is  ap-
parently saturable  (Ikeda, 1977)..  The-  enzyme  systems  respon-
sible for this metabolism are  indueible with phenobarbital
(Ikeda and Imamura, 1973) and  polychlorinated  biphenyls
(Moslen, et al., 1977).
     D~   Excretion.
          In humans tetrachloroethylene is  primarily elimi-
nated from the body via the lungs with a half-life of  elimi-
nation estimated to be 65 hours  (Stewart, et al., 1961,  1970r
Ikeda and Imamura,  1973)-  Its metabolite,  trichloroacetic
acidr is eliminated in the urine of humans  with a half-life
estimated to be- 144 hours (Ikeda and Imamura,  1973).
IV..  EFFECTS
     A.   Carcinogenicity
          Tetrachloroethylene  caused hepatocellular carcino-
mas- in B6C3-F1 mice of both sexes (NCI, 1977)..  An experiment
in Osborne-Mendel rats produced  negative results, although
early mortality precluded the  use of this data in evaluating
the carcinogenicity of PCE: (NCI, 1977) ..
          Greim, et. al- (1975) could not demonstrate an  in-
crease in the mutation rate 'of _§. coli K12 with tetra-
chloroethylene..  Howeverr Cerna and Kypenova (1977) tested
PCE and found elevated rautagenic activity in Salmonella
strains sensitive to both base pair substitution and frame-
shift mutations.
     C.   Teratogenicity
          Only one  report has appeared concerning possible

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  tetrachloroethylene—induced teratogenesis (Schwetz, et al.
  1975).   Female rats and mice were exposed to 2000 mg/rn^ 7
  hours daily on days 6 to 15 of gestation.  Significant de-
  creases  in fetal body weight and resorption, subcutaneous
  edema and delayed ossification of skull bones and sternabone
  in the pups were noted*  These effects were mild, however,
  and led  the authors to conclude that PCE was not teratogenic.
  Additional work is necessary to determine whether PCE is ter-
  atogenic (U.S. EPA, 1979).
      D.    Other Reproductive Effects
            No information available.
      E.    Chronic Toxicity
            Repeated exposure to tetrachloroethylene has re— .
  suited in damge to liver and kidney in dogs (Klaassen and
  Plaa, 1967).  Toxic nephropathy has also been observed in
  mice and rats (NCI, 1977)»  In humans, chronic exposure to
.  1,890 to 2,600 mg PCE/m-^ caused three of seven men to have
  impaired liver function (Coler and Rossmiller, 1953).   Occa-
  sional reports have even associated tetrachloroethylene expo-
  sure with the symptomatology of more serious chronic diseases
  such as  Raynaud's disease  (Lob, 1957; Sparrow, 1977)..  Spar-
  row (1977)  .reported a case which involved depressed immune
  function, mildly depressed liver function, polymyopathy and
  severe acrocyanosis*  In a group of workers  occupationally
  exposed  to lower concentrations of tetrachloroethylene at ap-
  proximately 400 mg/m3 (one for 15 years), subjective com- ,
  plaints,  such as headache,  fatigue,  somnolence,  dizziness,

-------
and a sensation of  intoxication  were  noted  (MedeJc and
Kovarik, 1973).
     F..   Other Relevant'Information
         , Intolerance, of alcohol has  been  reported with tet-
rachloroethylene exposure  (Gold,  1969).,
V..   AQUATIC TOXICITY
     A..   Acute Tox:icity   	"""'"'	
          Ninety—six hour  LCgg values foe  flow—through
and static tests are 18,400  and  21,400 ug/lr  respectively,
with the fathead minnow, Pimephales promelas  (Alexander,  et
al.. 1978)... With the bluegill, Lepomis macrochirus,  the 96-
hour LC5Q value is  12,900  ug/1 (U.S..  EPA, 1978).   For
Daphnia maqna, an observed 48-hour LCSO value of  17,700
ug/1 has been recorded  (U.S. EPA, 1978).
          No acute  data are  available for saltwater fish.
The mysid shrimp (Mysidopsis bahia) has an observed 96-hour
LC50 of 10,200 ug/1  (U..S.  EPA, 1978) »
     B.   Chronic Toxicity
          Chronic test data  are  not available for freshwater
species..  A chronic value  for the. saltwater mysid shrimp  in a
life cycle test is  448 ug/1  (U.S, EPA, 1978).
     C.   Plant Effects
          No adverse effects on  chlorophyll ^ concentration
or cell numbers with the alga, Selenastrum capricornutum,
were observed at exposure  concentrations as high  as  816,000
ug/1 (U.S. EPA, 1978)-  Two 96-hour EC50 values were  re-   .
ported for the marine micro alga, Skeletonema costaturo:
504,000 ug/1 based  on cell numbers and 509,000  ug/1  based  on

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chlorophyll £ concentration  (U.S. EPA,  1978).   The  raacroalga,

Phaeodectylum tricornuturn, was considerably  more  sensitive to

tetrachloroethylene toxicity with a  reported EC50 of

10,500 ug/1 (Pearson and McConnell,  1975).

     0.   Residues

 _        The bioconcentration factor for bluegills, Lepomis

macrochirusr has been reported to be 49  (U.S.  EPA,  1978).

Equilibrium was reached within 21 days  and the depuration

rate was rapid with a half-life of less  than one  day.

VI.G EXISTING GDIDLINES AND STANDARDS

     A.   Human

          Based on the NCI mice data, and using the "one-hit"

model, the U.S. EPA (1979) has estimated levels of  tetrachlo-

roethylene in ambient water which will  result  in  specified

risk levels of human cancer:

                               Risk  Levels and
Exposure Assumptions      Corresponding  Draft  Criteria
     (per day)     .      0  10 710~610"5

2 liters of drinking
water and consumption
of 18.7 grams fish and
shellfish.                0 0.020 ug/1  0.20  ug/1  2.0 ug/1

Consumption of fish and.
shellfish only.           0 0.040 ug/1  0.40  ug/1  4.0 ug/1

The present American Governmental Conference on Industrial

Hygiene (AGCIH, 1977) threshold limit value  (TLV) is 670
                            i^fTj^
                           *} 0 I IA£

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     B.   Aquatic
          For tetrachloroethylene,  the  draft  criterion to
protect saltwater aquatic life  is 79 ug/1  as  a 24-hour aver-
age; the' concentration should never exceed 180 ug/1 at any
time (U.S. EPA, 1979).
          For freshwater aquatic life,  the draft criterion is.
310 ug/1 as a 24-hour average;  the  concentration should never
exceed 700 ug/L at any time  (U.S. EPA,  1979).
          This: draft criteria to protect aquatic life  is
presently being" reviewed before final recommendation.

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                         TETRACHLOROETHYLENE

                              REFERENCES

    Alexander, H., et  al.   1978.  Toxicity  of  perchloroethylene,
    trichloroethylene,  1,1,1-trichloroethane,  and  methylene  chlo-
    ride to fathead minnows.  Bull..  Environ. Contam.  Toxicol.
    20: 344.

    American Conference of Governmental Industrial Hygienists.
    1977.  Documentation of the threshold limit  value.   3rd  ed.

    Cerna, M., and H.  Kypenova.   1977.  Mutagenic  activity of
    chloroethylenes analyzed  by screening system tests.   Mutat.
    Res.  46: 214.

    Coler, H.R., and H.R. Rossmiller.  1953.   Tetrachloroethylene
    exposure in a small industry.  Arch. Ind.  Hyg.  Occup.  Med,
    8: 227.

    Daniel, J.W.  1963.  The  metabolism of  36cl-labelled  tri-
    chloroethylene and  tetrachloroethylene  in  the  rat.  Biochem.
    Pharmacol.  12: 795.

    Fishbein, L..  1976.  Industrial  mutagens and potential rauta—
    gens I.  Halogenated aliphatic hydrocarbons.  Mutat.  Res.
    32: 267.

    Gold, J.H.  1969..  Chronic perchloroethylene poisoning.  Can.
    Psychiat. Assoc.. Jour.  14: 627,
I
    Greim, H., et al..  1975..  Mutagenicity  in  vitro and potential
    carcinogenicity of chlorinated ethylenes as  a  function of
    metabolic oxirane  formation.  Biochem.  Pharmacol.  24: 2013.

    Ikeda, M.  1977.  Metabolism of  trichloroethylene and  tetra-
    chloroethylene in human subjects.  Environ.  Health Perspect.
    21: 239.

    Ikeda, M., and T. Imamura.  1973.  Biological  half-life  of
    trichloroethylene and tetrachloroethylene  in human subjects.
    Int. Arch. Arbeitsmed.  31: 209.

    Ikeda, M., and H. Ohtsuji.  1972.  A comparative study of the
    excretion of Fujiwara - reaction-positive  substances  in  urine
    of humans and rodents given trichloro-  or  tetrachloro- deriv-
    atives of ethane and ethylene.   Br. Jour.  Ind. Med.   29:  99.

    Ikeda, M., et al.  1972.  Urinary excretion  of total  tri-
    chloro-compounds, trichloroethanol and  trichloracetic  acid as
    a measure of exposure to  trichloroethylene and tetrachloro—
    ethylene.  Br. Jour.. Ind. Med.   29: 328.

-------
Klaassen, C.D..,  and  G.L..  Plaa.   1967.  Relative effects of
chlorinated hydrocarbons  on  liver and kidney function in
dogs.  Toxicol.  Appl..  Pharmacol.   10: 119.

Lillian, D.,  et  al..  19.75.   Atmospheric fates of halogenated
compounds.  Environ. Sci. Technol.   9:  1042.

Lob, M.  1957.   The  dangers  of  perchloroethylene.  Int.. Arch.
Gewerbe—patholog... und  Gewerfahyg.   16: 45..

McConnell, G-., et  al.   1975. Chlorinated hydrocarbons and
the environment.  Endeavour   34:  13.

Medek, V.r and J"..  Kovarik.   1973.  The  effects of perchloro- .
ethylene on the  health of workers.   Pracovni Lekarstvi  25:
339.

Moslen,. M.T., .et al.  1977.   Enhancement of the metabolism
and hepatoxicity of  trichloroethylerie and perchloroethylene.
Biochem.. Pharmacol.  26:  369.

National Cancer  Institute.   1977.  Bioassay of tetrachloro-
ethylene for possible  carcinogenicity.   CAS No. 127-18-4- NCI —
CG-TR-13 DREW Publication No. (NIH)  77-813.

Pa.tty, F.  1963.   Aliphatic  halogenated hydrocarbons..  Ind.
Hyg.. Toxicol.  2r  1314..
            4

Pearson, C.R., and G.  McConnell.   1975.  Chlorinated Ci  and
C2 hydrocarbopns in  the marine-  environment.  Proc. R. Soc.
London B.  189:  305-

Savolainen, H..,  et al. • 1977.   Biochemical  and behavioral
effects of inhalation  exposure  to. tetrachloroethylene and
dichloromethane.   Jour. Neuropathol. Exp. Neurol.   36:. 941.

Schwetzv B.A., et  al.   1975. The effect of maternally in-
haled trichloroethylene, perchloroethylene, methyl chloro-
form, and methylene  chloride on embryonal and  fetal develop-
ment in mice and rats. Tox.icol.  Appl.  Pharmacol.   32: 84.

Sparrow, G.P.  1977.   A connective  tissue disorder similar to
vinyl chloride disease in a  patient  exposed to perchloroethy-
lene.  Clin. Exp.  Dermatol.   2r 17.

Stewart,. R.D, et al..  1970.  Experimental human exposure to
tetrachloroethylene.   Arch.  Environ- Health  20:  225.
                           i t TT~l7 f-
                           I U I ->

-------
Stewart, R.D., et al..  1961..  Human exposure  to  tetrachloro-
ethylene vapor.  Arch..,Environ. Health  2: 516.
U.S, EPA.  1978.  In-depth studies on health  and environmen-
tal impacts of selected water pollutants.  Contract No.   68-
01-4646.  U.S- Environ. Prot. Agency.
U.S. EPA.  1979.  Tetrachloroethylenet Ambient Water Quality
Criteria (Draft).
Windholz, M., ed.  1976.  The Merck Index.  9th  ed.  Merck
and Co., Rahway, tf.J^
Yllner, S..  1961^  Urinary metabolites of 14c-tetrachloro-
ethylene. in rnice^  Nature (Lond.)  191: 820.
                              KT

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                                      No.  159
              Thallium

  Health and Environmental Effects
U.S.. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse- health  and  environmental impacts  presented by  the
subject chemical.   This  document has undergone  scrutiny to
ensure its technical accuracy.

-------
                           THALLIUM



                            Summary



     Thallium  is a highly  toxic  element to many organisms,



including humans.  Symptoms of acute  exposure to thallium



include alopeciar_ataxiar  and tremors,  occasionally leading



to irreversible- coma- and death..   There-  is. no information



available on the mutagenic and carcinogenic properties of



thallium*  Although thallium has  been reported to be terato—



genier the evidence is not convincing..   The acceptable daily



intake (ADI) of thallium has beea determined to be*IS.4 mg



per day.  Thallium can be  chronically toxic to fish  at con-



centrations as low as 20 ug/1.  -Algae are also sensitive,



with effects produced at concentrations as low as 100 ug/1..'

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                           THALLIUM
 I.   INTRODUCTION
      This profile  is based on  the Ambient  Water  Quality Docu-
 ment for Thallium  (U.S. EPA, 1979).
      Thallium  (Tl;- atomic weight 204.37)  is  a soft,  malle-
 able, heavy metal  with a silver-white  luster (Lee,  1971).
 Thallium exists  in either the  monovalent  (thallous)  or tri-
 valent (thallic) form, the former being the  more common and
 stable and therefore forming more numerous and stable  salts
 (Harapel, 1968).  Thallium reacts chemically  with moisture  in
 air to form oxides.  Thallous  oxide  is easily oxidized to
 thallic oxide, a very hygroscopic compound,  or reduced to
 thallium.  While thallium itself is  relatively insoluble in
 water (Windholz, 1976), thallium compounds exhibit a wide
 range of solubilities.
      Current production and use of thallium  and  iti  compounds
 approximated 680 kg in 1976 (U.S. Dept. Interior, 1977).   In-
 dustrial uses of thallium include the manufacture of alloys,
 electronic devices, and special glass.  Many thallium-con-
 taining catalysts  have been patented for industrial organic
 reactions (Zitko,  1975).
 II *  EXPOSURE
      There is little information on  the extent of thallium
 contamination of water.  In a  single study ..by Greathouse
 (1978)  evaluating  drinking water from 3,834  households  ran-
 domly selected from 35 geographic areas, thallium was  dete'ct-
' able in only 0.68  percent of the samples (detection limit  was
 0.3 ppb), with the average concentration at  detection  of 0.89

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ppb.  Assuming a water  consumption of 2 Liters per day for
the average adult,  over 99  percent of adults would consume <
1 ug per day.  The  only study pertaining to natural water
measured the thallium content of  run-offs from mining and
smelting operations involving copper,, gold, zinc, and cadmium
with which, thallium is  associated in trace quantities (U.S.
EPA, 1978) ..  The highest, concentrations reported were- 30 ppb
in slag- run-off near Kellog-,  Idaho and. 21 ppb in the Colorado
River below drainage from a copper mine.
     Ingestion of thallium  from food is mainly due to the
consumption of vegetables..   Little data is available, al-
though Geilmann, et al.  (1960) found an average of 68.2 ppb
dry weight thallium in  four vegetables, analyzed*  This may be
high due to the small sample size.,  Breads contain 0.75 ppb
dry weight thallium, and  the thallium content of meats has
not been adequately determined.   The EPA (1979)  estimated
the weighted average bioconcentration factor for thallium to
be 61 for the edible portions of  fish and shellfish consumed
by Americans..  This estimate is. based on measured steady-state
bioconcentration studies  in  bluegill.  A daily intake from
food has been calculated  at  3.8 ug/day.   However,  due to the
sparse data,, this is probably not an accurate estimate-
     The contribution of  thallium in air to exposure is, in
most instances, small..  However,  thallium is  a contaminant in
flyash, and in a worst  case  situation in the  vicinity of a
                                                           »-
coal-fired plant, daily absorption could be. as high as 4.9
ug (Carson and Smith, 1977).   Due to possible high concentra-
tions in vegetable  matter, cigarette smoke may be  a signifi-

-------
cant source of thallium,  with  urinary excretion of thallium

in smokers being twice  that  in non-smokers (Weinig and Zink,

1967).

III. PHARMACOKINETICS

     A.   Absorption

          Gastrointestinal absorption of  trace quantities of

thallium appears to be  almost  complete in  both man (Barclay,

et al. 1953) and rats  (Lie,  et al.  1960).  No information was

found in the available  literature  concerning  the deposition

and clearance of inhaled  thallium  aerosols.   The skin would

not be expected to be a significant route  of  absorption of

thallium; however, systemic  pois_oning has  resulted from oint-

ments containing- 3-8 percent thallium acetate applied to the

skin (Munch, 1934) .

     B.   Distribution

          Thallium is widely distributed  in the body  in the

intracellular space.  Active transport of  thallium, mediated

by Na/K ATPase into erythrocytes has  been  demonstrated

(Gehring and Hammond, 1964;  Cavieres  and Ellroy,  1974).

Other factors besides active transport into cells must be

operating, since in both  conditions of normal thallium expo-

sure and fatal exposure in man,, there is a tendency for thal-

lium to concentrate in  the kidneys, colon  and hair (Weinig

and Zink, 1967; Cavanagh, et al. 1974).,

          Thallium crosses the placenta freely  from the  ma—
                                                           »
ternal circulation to the fetus.   In  studies  using rats  and

mice, steady state maternal/fetal  ratios of 0.84  and  0.46,

-------
respectively, were obtained  (Gibson,  et al..  1967);  and under
non-steady state conditions, wide  variations in dosage (0.2-
6.4 mgAg/roin) did not  alter, the. distribution from mother to
fetus (Gibson and Becker, 1970)..  Richeson (1958)  cites one
report in which thallium was found in the  tissue of a baby
whose mother had taken  1.2 g thallium at term..
     CV   Metabolism
          Pertinent information  could, not  be located in the
available literature.
     D~   Excretion
          Human excretion of thallium has  been estimated from
two studies, one involving a tracer dose of  204 Tl given to a
middle—age woman with osteogenic carcinoma raetastatic to the
lungs (Barclay, et al.  1953) and the  other involving a woman
suffering from thallium poisoning-  (Innis and. Moses,  1978).
From these two less than ideal studies,  total excretion of
thallium per day in. adults not exposed  to  unusual  sources of
thallium is probably as follows:
          Excretory route               ug  Tl/day
          Qrine                          1.20
          Feces                          0.06
          Hair                           0..32
          Skin and Sweat                 0.06
                           Total        1/64

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 IV.   EFFECTS'

      A..    Carcinogenicity and Mutagenicity

           Information regarding the carcinogenic and muta-

 genie potential  of thallium could not be located in the

 available  literature.

      B..    Teratogenicity

           There  are- two  reports of the teratogenicity of

 thallium,  one  involving-  chicken embryos (Karnofsky, et al.
„  r
 1950)  and  the  other rats (Gibson and Beckerr 1970).  In both

 cases, overt fetal toxicity due to thallium was noted, making

 it impossible  to. distinguish teratogenicity from a more

 general  toxic  effect.

      C.    Other  Reproductive Effects

           The  only known reproductive effect is fetal toxic-

 ity  in cases of  acute poisoning of the mother.

      D.    Chronic Toxicity

           There  are few  reports of chronic thallium poisoning

 in man-  In  one  brief  report concerning 13 men  exposed 3 to 4,

 months,  the  signs and  symptoms  were pains  in the legs, weari-

 ness,  loss of  hair,  disturbance of sensation, psychic trouble

 albuminuria  and  nephritis  (Meyer,  1928) ..

      Rats  fed  thallous acetate  in their diet for 105 days ex-

 perienced  no reduction in  weight gain at concentrations of 5

 and  15 ppm;  30 ppm,  however,  proved fatal  to approximately

 half the animals (Downs,  et al.  1960).

      E.    Other  Relevant Information

           Potassium  has  been shown to markedly  enhance the
      %

 rate of  thallium excretion (primarily urinary)  in both rats
                              A

-------
and dogs (Gehring and Hammond, 1967) ..  Potassium  also in-
creased somewhat the acute LO^Q of  thallium-   In  humans.,
potassium also increases urinary excretion with accompanying
temporary- accentuation of the neurological signs  and  symptoms
(Innis and Moses, 1978; Pappr et al. 1969)»
V.   AQUATIC TOXICITY
     A.   Acute Toxicity
          The bluegill appears to be- extremely resistant  to   y
thallium under renewal and static test conditions with  96-
hour LC5Q values of 132rOOO and 121,000 ug/1,  respectivel
(U.S.. EPA,'1979).  The fathead minnow was tested under  flow- ).
through conditions with measured^ concentrations, and  the  96- 0
hour LC50 value was found to be 860 ug/1  (U.S. EPA,  1978     .'
Atlantic salmon, when exposed to thallium for  as long as  2,60
hours, experienced 40 and 70 percent mortality at. approxi—    zO
mately 20 and 45 ug/1, respectively, with mortality occuring
throughout the test (Zitko, et al»  1975)-  The 48-hour  LC5
for Daphnia magna is 2,180 ug/1 (U.S.. EPA, 1978).
     B_   Chronic Tox icity
          An embryo—larval test with the fathead minnow indi-1—
cated adverse effects at the lowest thallium concentration
tested of 40 ug/1 (U.S. EPA, 1978).  No chronic data  are  avai
able for freshwater invertebrate speciesr and  no chronic  ef-
fects of thallium on saltwater organisms have  been reported
(U.S.. EPA, 1979) .

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     C.   Plant Effects;



          There is a 40 percent  inhibition  of oxygen evolu-



tion by the alga, Chlamydomonas  reinhardi,  exposed to a con-



centration of 40,800 ug/1  (Overnell,  1975).   The  96-hour



EC5Q values for chlorophyll £  inhibition  and  cell number



are 110 and 100 ug/lf respectively.



     0.-...  Residues



          The bluegill bioconcentrated  thallium 34 times



(whole body),, and the'Atlantic salmon bioconcentrated this



heavy metal '130 times above that of the ambient water (ZltJco,



et al. 1975; U.S. EPA, 1978).,



VI.  EXISTING GUIDELINES



     A.   Human



          The American Conference of Governmental Industrial



Hygienists (ACGIH, 1971) and the Occupational Safety and



Health Administration (OSHA) adopted a  threshold  limit value



of 0.1 mg/m3 for thallium.  The acceptable daily  intake-



(ADI) of thallium has been calculated to  be 15.4  rag  per day.



The U.S. EPA (1979) draft water criterion document for



thallium recommends a criterion of 4 ug/1 for  the protection



of human health.



     B.   Aquatic



          A criterion for the protection  of aquatic  species



from excess thallium exposure has not been .derived.


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                          THALLIUM

                         REFERENCES

American Conference of Governmental Industrial Hygienists.
1971.  Documentation of threshold limit values for substances
in workroom air..  3rd ed.

Barclay, R.K., et al.  1953.  Distribution and excretion of
radioactive thalli'um in the chick, embryo, rat and man.  Jour.
Pharmacol. Ex. Therap..  107: 178.

Carson, B.L.., and r..C. Smith.  1977..  Thallium.  An- appraisal
of environmental, exposure.  Tech. Rep. No. 5, Contract No.
N01-ES-2-2090.  Natl. Inst^ Environ. Health Sci.

Cavanagh, J.B., et al.»  1974.  The effects of thallium salts
with particular reference to the nervous system changes.
Jour. Med.  43: 293.

Cavieres, J.D.., and J.C. Ellroy.  1974.  Thallium and the
sodium pump in human red cells.  Jour. Physiol.  (London)
243: 243.-

Downs, W.L., et al~  1960.  Acute and subacute toxicity
studies of thallium compounds.  Am. Ind. Hyg- Assoc. Jour.
21: 399.

Geilmann,. W., et al.  1960.  Thallium ein regelmassig vor-
handenes spurenelement ,im tierschen und pflanzlichen or—
ganismus.  Biochem.. Zeit.  333: 62..

Gehring, P.J., and P.B.. Hammond.  1964.  The uptake of thal-
lium bv rabbit erythrocvtes..  Jour. Pharmacol.  Exp. Therap.
145: 215.

Gehring, P.J., and P.B. Hammond.  1967.  The interrelation-
ship between- thallium and potassium in animals.  Jour.
Pharmacol. Exp. Therap.  155: 137.

Gibson, J.E., et al.  1967.  Placental transport and distri-
bution of. thallium-204 sulfate in newborn rats and mice.
Toxicol. Appl. Pharmacol.  10: 408 (Abstract).

Gibsonr J.E., and B.A. Becker.  1970.  Placental. transfer,
embryo toxicity and teratogenicity of thallium sulfate- in
normal and postassium-deficient rats.  Toxicol.  Appl..
Pharmacol..  16: 120.

Greathouse, D.G.  1978.  Personal communication.
                                                          9
Hampel, C.A., ed.  1968.  The encyclopedia of chemical ele-
ments.  Reinhold Pub.,  New York.
                         Itf-ll

-------
Innis, R., and H. Moses.   1978.   Thallium poisoning.   Johns
Hopkins Med- Jour.   142:  27.

Karnofsky, D.A.., et  al.   1950.   Production of  achondroplasia
in the chick embryo  with  thallium.   Proc.  Soc.  Exp. Biol.
Med..  73:. 255.

Lee, A.G.  1971.  The chemistry  of  thallium.   Elsevier Pub-
lishing Co., Amsterdam.

Lie, R.,  et al.  I960*  The distribution  and excretion of
thallium-204 in the  rat,  with suggested MPC's  and  a bio-assay
procedure..  Health Physv  ..2: 334.

Meyer, S.  1928.  Changes  in the  blood as  reflecting
industrial damage.   Jour.  Ind. Hyg.  ^10;  29.

Munch, J.C.  1934.   Human  thallotoxicosis.  Jour.  Am.  Med.
Assoc»  102t 1929.

Overnell, J.  1975.  Effect of some  heavy  metal ions on
photosynthesis in a  freshwater alga.  Pest. Biochem. Physiol.
51 19.

Papp, J.P., et al.   1969.  Potassium chloride  treatment in
thallotoxicosis.  Ann. Intern. Med.  71:  119.

Richeson, E.M.  1958.  Industrial thallium intoxication.
Ind. Med. Surg.  2:  607..

U.S. Department of the Interior..  1977.  Commodity data sum-
maries.  Bur. Mines.

U.S. EPA.  1978.  In-depth studies on health and environmen-
tal impacts of selected water pollutants.   Contract No.  68—
01-4646.  U.S. Environ. Prot. Agency.

U.S. EPA.  1979.  Thallium:  Ambient Water Quality Criteria.
(Draft)  EPA PB-292444.  National Technical Information Ser-
vice, Springfield, Va.

Weinig, E., and P. Zink.   1967.  Uber die  quantitative mas-
senspektroraetrische  bestimmung des normalen thallium-gehalts
im menschlichen organismus.  Arch. f. Topxikol.  22: 255.

Windholr, M., ed.  1976.  The Merck  Index.'  9th ed.  Merck
and Co., Inc., Rahway, N.J..

Zitko, V. 1975.  Toxicity  and pollution potential  of thallium
Sci. Total Environ.  4: 185.

Zitko, V., et al.  1975.  Thallium:  Occurrence in the
environment and toxicity to fish.  Bull. Environ.  Contam.
Toxicol.  13: 23.

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                                     LB43-01
                                     No. 160
              Toluene

  Health and Environmental  Effects
U.S. ENVIRONMENTAL PROTECTION AGENCT
      WASHINGTON, D.C.   20460

         OCTOBER 30, 1980
               160-1

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                             DISCLAIMER
     This report represents a survey of the potential health and
environmental hazards from exposure to the subject chemical,  The
information contained in the report is drawn chiefly from secondary
sources and available reference documents.  Because of the limita-
tions of such sources, this short profile may not reflect all
available information including all the adverse health and environ-
mental impacts presented by the subject chemical*  This document
has undergone scrutiny to ensure its technical accuracy.
                               160-2

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                              TOLUENE




                              Summary




     The available studies that describe the carcinogenic or




mutagenic potential of toluene are inadequate for drawing any con-




clusions about its carcinogenicity.  There is  suggestive evidence,




based on skin painting experiments in mice, that toluene has a




weak promoting effect on DMBA-initiated skin carcinogenesis.  Three




studies in rats indicate that toluene damages chromosomes in bone




marrow cells.  Toluene-exposed workers showed an increase (not




statistically significant) of chromosome breaks in peripheral




lymphocytes.  Some neuromuscular deficiencies have been reported in




women exposed chronically to toluene in the workplace.  Subacute




and chronic studies on experimental animals have failed to show




evidence of severe cumulative toxicity.  Acute exposure to high




levels of toluene causes CNS depression.  The U.S.  EPA (1979) has




calculated an ADI of 29.5 mg for toluene.




     Toluene is acutely toxic to freshwater fish at concentrations




of 6,940 to 32,400 ug/1 and to marine fish at concentrations from




4,470 to 12,000 ug/1.   A single chronic value of 2,166 ug/1 has




been reported for marine fish.   Aquatic plants appear to be resist-




ant to the action ot toluene with effective concentrations ranging




from 8,000 to 433,000 ug/1.
                               160-3

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I.   INTRODUCTION

     This profile is based primarily on.  the Ambient Water Quality

Criteria Document for Toluene  (U.S. EPA, 1979) and to a lesser

extent on Criteria for a Recommended Standard:  Occupational Exposure

to Toluene (NIOSH, 1973) and its update  (NIOSH, 1977).

     Toluene. (CgHsCHs;' molecular weight  92.13) is a clear, colorless,

noncorrosive liquid with a sweet pungent odor.  It has the following

physical and chemical properties (Kirk and Othmer, 1963; Sutton'and

Calder, 1975; Shell and Ettre, 1971; Weast, et al. 1971):



          Boiling Point              110.6°C
          Freezing Point             -94.9°C
          Flash Point                6-108C
          Vapor Pressure             28  mm Hg at 25°C
          Solubility                 Water:  534.8 + 4.9 mg/1 in
                                     fresh water and" 379.3 + 2.8 mg/1
                                     in  seawater.  Miscible with
                                     alcohol, chloroform, ether,
                                     acetone, glacial acetic acid,
                                     carbon disulfate and other
                                     organic solvents.
          Production                 7.3 x 103 tons/year (USITC, 197.7)


     Approximately 85 percent of the toluene produced is converted

into benzene and other chemicals.  The remainder is used as a

solvent and as a gasoline additive (NIOSH, 1973).

     Little is known about the transport and persistence of toluene

in the environment.   Toluene is volatile and can evaporate into the

atmosphere from bodies of water (MacKay and-Wolkoff,  1973),  In the

atmosphere, toluene is photochemically degraded to benzaldehyde and

traces of perbxybenzoyl nitrate.   Toluene can re-enter the hydrosphere

in rain (Walker, 1976).


                               160-4

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II.  EXPOSUR.E




     A.   Water




        ' No estimates of average daily uptake of toluene from




water, food, and air are available.  In nationwide surveys of




organic chemicals in the drinking water of representative U.S.




communities, toluene was found to contaminate 1 raw and 11 finished




water supplies out of the 133 water supplies surveyed (U.S. EPA,




I975a; 1975b; 1977).  Quantitative analyses of five of the above




finished waters revealed levels of toluene ranging' from 0.1 ug/1 to




19 ug/1.  Benzaldehyde and benzole acid, metabolites of toluene,




were also detected.  Benzaldehyde was found in the water of five




cities, and in two of the cities was measured at levels of 0.1 and




0.5 ug/1.  Benzoic acid at 15 ug/1 was found in the water of another




city.




     B.   Food




          Little data on levels of toluene in food are available.




Toluene was detected in sea water and fish obtained near petroleum




and petrochemical plants in Japan (Ogata and Miyake,  1973).  The




muscle of one representative fish contained five ug toluene/g of




tissue.  Benzaldehyde, a metabolite of toluene, occurs naturally in




some foods and is intentionally added to others as a flavoring agent.




Benzoic acid, another metabolite of toluene, is added to some foods




as a preservative.



     The U.S. EPA (1979) has estimated the weighted average




bioconcentration factor for toluene to be 20 for the edible portions



of fish and shellfish consumed by Americans.  This estimate is based






                               160-5

-------
on the octanol/water partition coefficient of toluene and on




estimates of fish and shellfish consumption.



     C.'  Inhalation




          Toluene has been detected in urban air at concentrations



many times lower than vapor levels considered to be potentially



harmful in occupational settings.  An average level of 37 ppb and a



maximum level of 129 ppb were measured in the air of Los Angeles




(Lonneman, et al. 1968).  Comparable levels were found in the air



of Toronto, Canada (Filar and Graydon, 1973), and the air of Zurich,



Switzerland (Grob and Grob, 1971).  In these latter studies,



atmospheric toluene in urban areas appeared to arise primarily from



motor vehicle emissions.



III.  PHARMACOKINETICS



     A.   Absorption



          No reports are available on oral administration of toluene'



to humans (U.S. EPA, 1979).   Toluene concentrations in arterial



blood of persons continuously inhaling toluene vapors appeared to



approach equilibrium after 20 to 30 minutes, at which time blood



lev'els were about 1 ug/ml in persons inhaling 100 ppm, and 2 ug/ml



in persons inhaling 200 ppm toluene (Astrand, et al. 1972).   Systemic



uptake of toluene was doubled by exercise, due primarily to increased



ventilation rate (Astrand, et al. 1972).   This increased uptake of



toluene upon exercise was also noted by Carlsson and Lindqvist (1977),



who in addition noted that obese persons retain more toluene than



thin ones.  In their study,  the average uptake of toluene vapor



during exercise was approximately 49 percent for obese subjects





                               160-6

-------
versus 37 percent for thin subjects.  The rate of percutaneous




toluene absorption in humans was reported to be 14 to 23 mg/cm^/hour




(Dutkiewicz and Tyras, 1968).




          Rats absorbed toluene much more rapidly and developed




substantially higher peak blood and tissue toluene concentrations




when toluene wa's administered to the lungs, rather than to the gas-




trointestinal tract (Pyykko, et al. 1977).  Toluene absorption




through the skin of experimental animals occurred to a considerably




lesser degree than through the lungs or gut (Wahlberg, 1976).




     B.   Distribution




          Toluene is rapidly taken up from the blood into body




tissues according to their lipid content and blood perfusion




(U.S. EPA, 1979).  Partition coefficients (tissue:blood) for toluene




in homogenates of rabbit tissues have been determined.-  The parti-




tion coefficient for adipose tissue was 50 times greater, the




coefficient for bone marrow was approximately 15 times greater,




and those for brain and liver were roughly 2 times greater than




the partition coefficients for lung, kidney, heart, and muscle




(Sato, et al.  1974).  Saturation of liver and brain tissue of mice




was not reached even after 3 hours of inhalation of concentrations




as high as 4000 ppm toluene (Bruckner and Peterson, 1976).




     C.   Metabolism




          In humans and experimental animals,  toluene is thought to




be enzymatically converted by the mixed function oxidase (MFO)




system to benzyl alcohol,  which is subsequently oxidized to benzal-




dehyde and benzoic acid.   Benzoic acid is then conjugated with






                               160-7

-------
glycine to form hippuric acid (U.S. EPA, 1979).  There has also




been a report, however, of glucuronide conjugation of benzoic acid




in rabbits given large doses (Bray, et al. 1951).  Toluene toxicity




is diminished in rats by MFO inducers (Ikeda and Ohtsuji, 1971) and




enhanced by MFO inhibitors (Koga and Ohmiya, 1978), suggesting that




metabolism of toluene 'results in detoxication.

         . •


     D.   Excretion




          Toluene is rapidly excreted from the body following




inhalation exposure.  Most of the estimated absorbed dose of toluene




can be accounted for within the first.12 hours as the parent compound




in expired air and as hippuric acid in the urine (U.S. EPA, 1979).




Elimination rates are slower for women than for men, perhaps because




of the larger proportion of fatty tissue in women (U.S. EPA, 1979).




          Excretion of toluene in experimental animals is similar




to that found in man.  In the rat, for example, elimination of




toluene occurs more slowly from adipose tissue than from any other




(Pyykko, et al. 1977; Carlsson and Lindqvist, 1977), including bone




marrow, from which elimination is also relatively slow (U.S. EPA,




1979).   Toluene is rapidly lost from the brain, as reflected in




rapid recovery from toluene-induced CNS depression (Peterson and




Bruckner, 1976; Savolainen, 1978).
      •7



IV.  EFFECTS




     A.   Carcinogenicity




          The data base on the carcinogenic!ty of toluene is




extremely limited.   No inhalation studies have been done.  No




accounts have been found in the literature in which cancer in






                               160-8

-------
humans has been attributed specifically  to  toluene. . It  is difficult




to link cancer -induction with any single solvent, as persons having




occupational exposure to solvents are characterized by considerable




job mobility and exposure to a variety of chemicals (D.S. EPA, 1979).




Toluene was .not demonstrated to be carcinogenic when applied to the




skin of mice for one year (Doak, el al.  1976) or throughout a life-




time (Poel, 1963) (since toluene evaporates rapidly, this method




is not appropriate).  Toluene has not shown carcinogenicity when




administered to rats by inhalation at concentrations of  up to




300 ppm, 6 hours/day, 5 days/week for as long as 18 months (Gibson,




1979).  Frei and Kingsley (1968) reported that toluene has a weak




promoting effect on DMBA-initiated skin  carcinogenesis in Swiss




mice.  The major metabolite of toluene,  benzole acid, is not




carcinogenic, however', .about 1 percent of toluene can be metabolized




to o- and p-cresol, which are cancer promoters (Boutwell and Bosch,




1959).



     B.   Mutagenicity




          There is no conclusive evidence that toluene is mutagenic,




although it has been reported to cause chromosome damage.  For



example, the incidence of chromosomal abnormalities in peripheral




blood lymphocytes of humans who had been exposed to an average of




200 ppm toluene for as long as 15 years  was no greater than in




controls (Forni, et al. 1971).   However, there have been two reports




that toluene induced chromosomal aberrations in the bone marrow



cells of rats (Lyapkalo, 1973;  Dobrokhotov and Enikeev,  1977), and



typographers exposed to toluene have a slightly increased frequency






                               160-9

-------
of chromosome breaks as compared to controls (Funes-Cravloto et al.




1977).  Toluene has not been tested in bacterial screening systems



(Dean, 1978).




     C.   Teratogenicity



          Although toluene should readily pass the placenta, no




reports of teratogenic effects in humans are linked to toluene



exposure (U.S. EPA, 1979).  Toluene is teratogenic and embryotoxic.



in mice (Nawrot and Staples, 1979).  It was shown to be teratogenic



at 1..0 mg/kg, embryolethal at 0.3 ml/kg, and decreased fetal weight



occurred at 0.5 ml/kg.  There was no maternal toxicity at any dose



level on days 6-15, but maternal weight gain was noted for doses



given on days 12-15.



     D.   Other Reproductive Effects



          Women occupationally exposed to multiple solvents including



toluene through the use of varnishes had a relatively high incidence^



of menstrual disorders.  Their offspring were said to experience



more frequent fetal asphyxia, to be more underweight, and not to



nurse as well as "normal" infants (Syrovadko, 1977).  Dysmenorrhea



was a frequent subjective complaint of female shoemakers Chronically



exposed to 60-100 ppm toluene (Matsushita, et al, 1975).  In a



single study, some retardation of body weight and skeletal growth



were seen in fetuses of rats exposed continuously to 399 ppm toluene



on days 1 to 8 of gestation; inhalation of lower levels of toluene



had no effect (Hudak and Ungvary, 1978).
                               160-10

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     E.   Chronic Toxicity




          The toxicity of toluene was recently reviewed (Lohr and




Stokholm, 1979).  Its major toxic effects are on the central nervous




system, causing depression, headaches, confusion, dizziness, insomnia




and (at high exposure levels) death.




          A  st'udy of 38 female shoemakers exposed chronically to




solvents including toluene at 60 to 100 ppm for about three years




revealed abnormal tendon reflexes, reduced grasping power, and




decreased finger agility when compared to controls (Matsushita,




et al., 1975).  Reports reviewed by the National Institute for




Occupational Safety and Health (1973) have failed to demonstrate




adverse effects on the hematopoietic, hepatic, renal, or other




physiologic systems of workers routinely inhaling approximately 100




ppm toluene.  Chronic exposure may also lead to disturbances in the




immune system, dermatitis, and permanent damage to the central




nervous system (Cohr and Stockholm., 1979; U.S. EPA, 1979).




     F.   Other Relevant Information




          The primary hazard associated with acute exposure to high




levels of toluene is excessive CNS depression (U.S. EPA, 1979).




Toluene is capable of altering the metabolism and bioactlvity of




other chemicals which are metabolized by the mixed function oxidase




system.  For example, simultaneous administration of toluene and




trichloroethylene or toluene and benzene to experimental animals




resulted in suppression of metabolism of both compounds (Ikeda,




1974; Ikeda, et al., 1972).   Another showed marked reduction in the




concentration of benzene metabolites in various tissues, including






                               160-11

-------
bone marrow, after simultaneous administration of toluene, and



data that suggested that toluene might protect against benzene



myelotoxicity (Andrews, et al. , 1977).



V.   AQUATIC TOXICITY



     A.   Acute Toxicity



          For freshwater fish, 96-hour static LCso values ranged



from 12,700 ug/1 for the bluegill (Lepomis 'macrochirus) to 59,300



ug/1 for the guppy (Poecilia reticulatus)  (U.S. EPA, 1978; Pickering



and Henderson, 1966).  Only a  single 48-hour LC5Q value for Daphnia



magna of 313,000 ug/1 has been obtained for toluene.  In marine



fish, two 96-hour static LCso values of 6,300 and 10,000-50,000



ug/1 were obtained for striped bass (Morone saxatilis) and coho



salmon Oncorhynchus kisutch (Benville, et  al., 1977).  Among four



species of marine invertebrates, the bay shrimp (Crago franciseorum)



was most sensitive, with a 96-hour static  LC5Q value of 3,700 ug/1



(Benville, et al., 1977), while the mysid  shrimp Mysidopsis bahia



was most resistant, with a 96-hour static LC5Q value of 56,300 ug/1



(U.S. EPA, 1978).



     B.   Chronic Toxicity



          No. freshwater chronic data could be found in the available



literature.   The only marine chronic value reported was 2,166 ug/1
                                       •i


for the sheepshead minnow (Cyprinodan variegatus) (U.S. EPA, 1978).



     C.   Plant Effects



          The freshwater algae Chlorella vulgaris and Selenastrum



capricornutum were fairly insensitive to the action of toluene: EC5Q



values for cell numbers range from 245,000 ug/1 for Chlorella





                               160-12

-------
(Kauss and Hutchinson, 1975) to 433,000 ug/1 for Selenastrum

(U.S. EPA, 1978).  Among five marine algal species tested, Skeletonema
                                                                    /
costatum was the most sensitive with an adverse effect on growth at

8,000 ug/1 (Donstan, et al., 1975).

     D.   Residues

          No bioconcentration factors are available for toluene in

freshwater or marine organisms.

VI.  EXISTING GUIDELINES AND STANDARDS

     Both the human health and aquatic criteria derived by U.S. EPA

(1979), which are summarized below, have not yet gone through the

process of public review; therefore, there is a possibility that

these criteria may be changed.

     A.   Human

          The NIOSH (1973) recommended standard for exposure to

toluene is 100 ppm, determined as a time-weighted average for an

8-hour workday, with a ceiling of 200 ppm.

          The U.S. EPA (1979) draft criterion for toluene in ambient

water is 12.4 mg/1, corresponding to a calculated acceptable daily

intake of 29.5 mg.  This criterion is based on chronic toxicological

test data for rats (maximum no-effect level of 590 mg/kg, 5 days/wk)

and the application of an uncertainty factor of 1000.

     B.   Aquatic

          The draft criterion for the protection of freshwater

organisms is 2,300 ug/1, as a 24-hour average, not to exceed 5,200

ug/1; and for marine life the draft criterion Is 100 ug/1, as a

24-hour average, not to exceed 230 ug/1.


                               160-13

-------
                               TOLUENE

                             REFERENCES
       *
Andrews, L.S,, et al.   1977.   Effects  of  toluene  on  the  metabolism,
disposition and hemopoletic  toxicity of  (^H)  benzene.  Biochera.
Pharmacol.  26: 293.

Astrand, I., et al.  1972.   Toluene exposure.   I.  Concentration
in alveolar air and  bl'ood at rest and  during  exercise.   Work Environ,
Health  9:- 119.

Benville, P.E., Jr., et al.  1977.  The acute  toxicity of six
monocyclic aromatic  crude oil  components  to striped  bass  (Morone
saxatilis) and bay shrimp (Crago franeiscorum).   Calif.   Fish
Game.  63: 204. •

Boutwell, R.K., and D.K. Bosch.  1959.  The tumor-promoting action
of phenols and related  compounds for mouse skin.  Cancer  Res.
19: 413-424.

Bray, H.G., et al.   1951.  Kinetic studies of  the metabolism of
foreign organic compounds.   I.  The formation  of  benzole  acid from
benzamide, toluene, benzyl alcohol and benzaldehyde  and  its conju-
gation with glycine  and glucuronic acid in the  rabbit.   Biochem.
Jour.  48: 88.

Bruckner, J.V.t and R.G. Peterson.  1976.  Evaluation of  toluene
toxicity utilizing mouse as  an animal model of  solvent abuse.
Pharmacologist  18:  244.

Carlsson, A., and T. Lindqvist.  1977.  Exposure  of  animals and
man to toluene.  Scand. Jour. Work Environ. Health   3: 135.

Cohr, K.H., and J. Stockholm, 1979.  Toluene,  a toxicologic review.
Scand. J. Environ, and Health.  5: 71-90.

Dean, B.J.  1978.   Genetic toxicology of benzene, toluene, xylene,
and phenols.  Mutat. Res.  47: 75.

Doak, S.M.A., et al.  1976.   The carcinogenic  response in-mice to
the topical application of propane sultone to  the skin.
Toxicology  6: 139.

Dobrokhotdv, V.B., and M.I.  Enikeev.  1977.  Mutagenic effect of
benzene,  toluene,  and a mixture of these hydrocarbons in a chronic
experiment.  Gig.  Sanit.  1: 32.
                               160-14

-------
                              TOLUENE

                             REFERENCES  (Continued)

Dunstan-, W.M. , et al.  1975.  Stimulation and  inhibition of phyto-
plankton growth by low molecular weight  hydrocarbons.  Mar. Biol.
31: 305.

Dutkiewicz, T., and H. Tyras.  1963.  The quantitative estimation
of toluene skin absorption in man.  Int. Arch. Gewerbepath.
Gewerbehyg.   24': 253

Forni, A., et al.  1971.  Chromosome studies in workers exposed to
benzene or toluene or both.  Arch. Environ. Health   22: 373.

Frei, J. , and W.F. Kingsley.  1968. 'Observations on the chemically
induced regressing tumors of mouse epidermis.  J.  Natl. Cancer
Inst.  41: 1307-1313.

Funes-Cavioto, F., et al.  1977.  Chromosomal  aberrations and
sister-chromatid exchange in workers in  chemical laboratories
and a rotoprinting factory and in children of  women laboratory
workers.  Lancet  2: 322-325.

Gibson, J.E.  1979.  Chemical Industry Institute of Toxicology -
Two year vapor inhalation toxicity study with  toluene in
Fischer-344 albino rats:  18-month status summary.  (Personal
communication).

Grob, K. , and G. Grob.  1971.  Gas-liquid chromatographic/mass
spectrometric investigation of Cg-C20 organic  compounds in an
urban atmosphere.  Jour. Chromatogr.   62: 1.

Hudak, A., and G. Ungvary.  1978.  Embryotoxic effects of benzene
and its methyl derivatives:  toluene and xylene.  Toxicology
11: 55.

Ikeda, M.  1974.  Reciprocal and metabolic inhibition of toluene
and trichloroethylene in vivo and in vitro.  Int.  Arch. Arbeitsmed.
33: 125.

Ikeda, M., and H. Ohtsuji.  1971.  Phenobarbitol-induced protection
against toxicity of toluene and benzene  in the rat.  Toxicol. Appl.
Pharmacol.  20: 30.

Ikeda, M. , et al.  1972.  In vivo suppression  of benzene and styrene
oxidation by co-administered toluene in  rats and effects of pheno-
barbitol.  Xenobiotica  2: 101.
                               160-15

-------
                              TOLUENE
     «
                             REFERENCES  (Continued)

Kauss, P.B., and T.C. Hutchinson.  1975.  The effects of water-
soluble petroleum components on  the grown of Chlorella vulgaris
Beijernck.  Environ. Pollut.  9: 157.

Kirk, R.E.', and D. Othmer.  1963.  Kirk-Othmer Encyclopedia of
Chemical Technology.  2nd ed.  John Wiley and Sons, Inc.,
New York.

Koga, K., and Y. Ohmiya.  1978.  Potentiation of toluene toxicity
by hepatic enzyme inhibition in  mice.  Jour. Toxicol.  Sci.  3: 25.

Lonneman, W.A., et al.  1968.  Aromatic  hydrocarbons in the atmos-
phere of the Los Angeles Basin.  Environ. Sci. Technol.  2: 1017.

Lyapkalo, A.A.  1973.  Genetic activity  of benzene and toluene.
Gig. Tr. Prof. Zabol.  17: 24.

Mackay, D., and A.W. Wolkoff.  1973.  Rate of evaporation of low-
solubility contaminants from water bodies to atmosphere.  Environ.
Sci. Technol.  7: 611.

Matsushita, T., et al•  1975.  Hematological and neuro-muscular
response of workers exposed to low concentration of toluene vapor.
Ind. Health  13: 115.

Nawrot, P.S., and R.E. Staples.  1979.   Embryofetal toxicity and
teratogenicity of benzene and toluene in the mouse Teratology.
19: 419 (abstract).

National Institute for Occupational Safety and Health.  1973.
Criteria for a recommended standard...occupational exposure to
toluene.  HEW Publ. No. HSM 73-11023.  U.S.  Government Printing
Office.  Washington, D.C.

National Institute for Occupational Safety and Health.  1977.
Review, summarization and evaluation of  literature to support the
update and revision of criteria  documents.  V.  Toluene.

Ogata, M. , and Y. Miyake.  1973.  Identification of substances in.
petroleum causing objectionable  odor in  fish.  Water Res.
7: 1493.

Peterson, R.G., and J.V. Bruckner.  1976.  Measurement of toluene
levels in animal tissues.  Proc. Int. Symp.   Deliberate Inhalation
of Industrial Solvents, Mexico City.
                               160-16

-------
                              TOLUENE

                             REFERENCES  (Continued)

Pickering, Q.H., and C. Henderson.  1966.  Acute  toxicity of  some
important petrochemicals to  fish.  Jour. Water Follut.  Control
Fed.  38: 1419.

Filar, S., and W.F. "Graydon.  1973.  Benzene and  toluene distribu-
tion in Toronto atmosphere.  Environ. Sci. Technol.  7: 628.

Poel, W.E.  1963.  Skin as a test site for the bioassay of carcinogens
and carcinogen precursors.  Natl. Cancer Inst.  Monogr.  10:  611.
                                                   *
Pyykko, K. , e t. al.  1977.  Toluene concentrations in various  tissues
of rats after inhalation and oral administration.  Arch. Toxlcol.
38: 169.

Sato, A., et al.  1974.  Pharmacokinetics of benzene and toluene.
Int. Arch. Arbeitsmed.  33: 169.

Savolainen, H.  1978.  Distribution and  nervous system binding of
intraperitoneally injected toluene.  Acta Pharmacol. Toxicol.
43: 78.

Shell, F.D., and L.S. Ettre. eds.  1971.  Encyclopedia of Industrial
Chemical Analysis;  Interscience Publishers, John Wiley and Sons,
Inc., New York.

Sutton, C., and J.A. Calder .  1975.  Solubility of alkylbenzenes
in distilled water and seawater at 25 C.  Jour. Chem. Eng. Data.
20: 320

Syrovadko, O.N.  1977.  Working conditions and health status  of
women handling organosiliceous varnishes containing toluene.
Gig. Tr . Prof. Zabol.  12:15.

U.S. EPA.  1975a.  New Orleans area water supply  study.  Analysis
of carbon and resin extracts.  Prepared  and submitted to the  lower
Mississippi River Branch, Surveillance and Analysis Division,
Region VI, by the Analytical Branch,  Southeast Environ. Res.  Lab.,
Athens, Ga.

U.S. EPA.  1975b.  Preliminary assessment of suspected carcinogens
in drinking water.  Report to Congress,  Washington, D.C.

U.S. EPA.  1977.  National Organic Monitoring Survey, general review
of results and methodology:  Phases I-III.
                               160-17

-------
                              TOLUENE

                             REFERENCES (Continued)

U.S. EPA.  1978.  In-depth studies on health and environmental
impacts of 'selected water pollutants.  Contract No. 68-01-4646.

U.S. EPA.  1979.  Toluene:  ambient water quality criteria.
(Draft)   •  '

U.S. ITC, Annual.  1977.  Synthetic Organic Chemicals, U.S.
Production and Sales.  U.S. International Trade Commission,
Washington, D.C.

Wahlberg, J.E.  1976.  Percutaneous toxicity of solvents.  A
comparative Investigation in the guinea pig with benzene, toluene
and 1,1,2-trichloroethane.  Ann. Occup. Hyg.  19: 115.

Walker, P.  1976.  Air pollution assessment of toluene.  MTR-7215,
Mitre Corp., McLean, Va.

Weast, R.C., et al.  1971.  Handbook of chemistry and physics.
52nd ed. CRC Press, Cleveland, Ohio.
                               160-18

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                                    No. 161
         2* 4-Toluenedlamine

  Health and Environmental Effects
U.S.  ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.   20460

          APRIL 30, 1980
          -fttt-
         lt>H

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal*  The information contained in the report is drawn chiefly
from secondary  sources  and available reference  documents..
Because of the limitations of such sources, this short profile
may not reflect all available  information  including all the
adverse health  and  environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny to
ensure its technical accuracy..      .                   !

-------
                              2,4-TOLUENEDIAMlNE
                                   Summary

     2,4-Toluenediamine produced carcinogenic effects in  rats  and  mice in a
long-term National Cancer  Institute  (NCI) feeding study  (50 pom;  100 ppm).
2,4-Toluenediamine was  found to be  mutagenic,  using mutants,  of Salmonella
typhimurium. hamster embryo cell systems,  and Orosophila melanoqaster.
     2,4-Toluenediamine was also found to be hepatotoxic to rats and mice in
the NCI  study on carcinogenicity.   The  compound valso hastened  the develo-
pment of chronic  renal  disease  and accelerated  animal morbidity.  Data con-
cerning the teratogenicity of 2,4-toluenediamine was not found  in the avail-
able  literature.    However,   a  closely  related  compound,  the  2,5-diamino
analog, is teratogenic in mice.
                                   111-3

-------
I..   INTRODUCTION
          2,4-Toluenediamine  (molecular weight  122.17)  is white  solid that
melts  at 99,°C,  has a  boiling  point of  292°C,  a.  density of 1.047  g/cm at
100°C,  heat of  vaporization of 27.975  kj/mol,  heat  of fusion  of 19.874,
and  a specific  heat of  2*572  J/g  at 150°C  (Milligan and  Gilbert,  1978).
This  compound  is very  soluble  in  hot  benzene,  in hot  water,'and  in "both
alcohol and ether (Weast,  1971).   The major use for 2,4-toluenediamine is in
the manufacture of  2,4-toluenediisocyanate  (TOI),  the  major raw material for
the  producton  of flexible polyurethane  foams and elastomers  (Milligan and
Gilbert,  1978).   The  production  of 2,4-toluenediamine  has  increased  more
than  100 percent since  1966 and  was  reported in  1976 at 2.05 X 10  tons,
with a predicted growth rate of 8-12 percent per year (Milligan and Gilbert,
1973).   2,4-Toluenediamine can  also be used in the manufacture, of dyes and
was an  important  ingredient in human hair  dyes of the  permanent,, oxidative
type  until  1971,  when its use  was restricted after being implicated in the
induction of liver carcinomas in rats (Ito,  et al.   1969).  Using  mutants of
Salmonella  typhimurium,  Ames,   et  al.  (1975)  found 2,4-toluenediamine  to be
mutagenic.
II.  EXPOSURE
          Two potential sources of exposure to 2,4-toluenediamine  are in its
manufacture and its use as an intermediate  in the production  of 2,4-toluene-
diisocyanate.  2,4-Toluenediamine  is manufactured by  seven U.S.  companies at
nine  U.S.  locations (Muller, 1979;  Gunn  and Cooke,\1976), and most of the
corresponding  diisocyanate is  produced  by  the  same  companies  at  the  same
locations.   Capacity  for  the  latter compound  is 3.75 X  10   tons  yearly
(Muller,  1979).   Some  additional amounts are consumed in the  production of
dyes  or  are  exported to manufacturers of  2,4-toluenediisocyanate outside the
United States.  The amount consumed as  a  dye  intermediate is believed  to be

-------
quite small,  and the magnitude  of the exports  of 2,4-toluenediamine is un-
known (Gunn  and  Cooke,  1976).  Monitoring data  are not available concerning
exposure  to, 2,4-toluenediamine  dermally  or  by  water,   food,  inhalation.
Dermal carcinogenicity  in  mice is discussed  below under "Effects" ("Chronic
Toxicity").
III. PHARMACOKINETICS
          Information on the  absorption,  distribution, metabolism,  and ex-
cretion of 2,4-toluenediamine was not found in the available literature..
IV.  EFFECTS
     A.    Carcinogenicity
          Carcinoma  of  the liver  with invasion and  metastases  was observed
in rats fed diets containing  0.1 or 0.06 percent 2,4-toluenediamine (Ito, et
al. 1969)..   When the compound was  fed at  levels of 50 and 100 ppm to inbred
barrier-raised F344  rats for  2  years,  a  statistically  significant increase
was observed  in  the  incidence of hepatic  neoplasia in males,  and it induced
a  significant dose-related positive trend  in  the  incidence  of  liver  neo-
plasms  in both  sexes.   Hepatocellular  changes  considered to  be associated
with neoplasia were  increased  at  a  high  level of statistical significance in
both sexes.   The compound  also caused statistically significant increases in
the  incidence of  mammary  tumors  in  females,  and  an  increase of  mammary
tumors  in  males,  although  not  significant   statistically,  was  believed
related to the chemical (Cardy,  1979;  Ulland, 1979).   2,4-Toluenediamine was
also   carcinogenic   for   female   B6C3F1   mice,. ' inducing   hepatocellular
carcinomas.   The incidence of lymphomas in  the female mice  suggested  that
these tumors  may have been related  to administration of the test chemical as
well (Ulland, 1979).

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     8.   Mutagenicity
          Fahmy and Fahmy  (1977)  conducted a comparative assay in Drosophila
melanqaster for the  assessment of the  mutagenic efficiency of  the hair dye
components  2,4-toluenediamine  and  4-nitro-o-phenylenediamine  relative   to
benzidine,  a  human  carcinogen which,  like  2,4-toluenediamine,  is  also  an
aromatic amine..   All compounds showed  mutagenicity activity..  Although-act-
ivities of  the chemicals on the different  genetic  sites varied between com-
pounds and  as a  function  of cell stage, mutagenic  activity  did not vary  in
response to changes in  dose.  The  mutagenicitiesf and  selectivities  of the
test compounds for ribosomal ONA gradually  decreased  in the order benzidine
greater  than  2,4-toluenediamine  greater   than  4-nitro-o-phenylenediamine.
For 2,4-toluenediamine  a good correlation  was  found between mutagenicity  in
the Salmonella/microsome test and morphological transformation  in a hamster
embryo cell system (Shah,  et al.  1977).   For  mutagenesis,  the compound re-
quired  metabolic  activation by  a  rat  liver  microsomal  enzyme  (S9)  pre-
paration.   In  contrast,  transformation of  hamster  cells was induced without
activation  by  external  enzymes..  In the Ames  assay there was  no mutagenic
activity in the strain TA1QOV indicating that the product is not a base pair
mutagen.  The  dose response  curves  obtained with tester  strain TA1538  and
TA98 show  that 2,4-toluenediamine is metabolized  by the  S9  to a frameshift
mutagen  (Shah, et al.  1977).  In a study  of  the  mutagenic effect  of 2,4-
toluenediamine in mice,  Scares  and Lock (1978)  found no significant increase
in dominant lethal mutations  (seven  weeks post-treatment) on males.
     C.   Teratogenicity
          Data concerning  the teratogenic effects  of 2,4-toluenediamine. were
not  found  in the  available  literature.    However,  2,5-toluenediamine,  a
closely  related compound which is a hair  dye constituent,  was found terato-
genic in mice (Inouye and Murakami,  1977).

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     D»   Other Reproductive Effects
          Information  on other  reproductive  effects- was  not found  in the
available literature.
     E..   Chronic. Toxicity
          Two  reports  primarily  dealing with  carcinogenicity  provide infor-
mation on chronic  toxicity.   Cardy (1979) found  that 2, 4-toluenediamine was
hepatotoxic when  fed at levels  of 50 and 100 ppm  to inbred,  barrier-raised
F344  rats  for 2  years.   The  compound also  accelerated  the  development of
chronic renal  disease  in the strain,  an effect  that contributed to a marked
decrease  in  the   survival   rate.   Giles  and Chung  (1976),  in  a  chronic
toxicity study of 2,.4-toluenediamine  alone  or in  combination with selected.
hair dye. complexes,  found the  compound, to  be nontoxic and noncarcinogenic to
the skin of mice.
     F.   Acute Toxicity '
          Lewis and Tatken (1979) summarize the available information:
         Oral-human LOQ: 50 mg/kg     Subcutaneous-rat LDLQ:  50 mg/kg
         Oral-rat LD :  500 mg/kg     Subcutaneous-dog TD^: 200 mg/kg
         Oral-rat TD:  11  g/kg     Subcutaneous-dog LD: AGO mg/kg
where  LDQ— lethal  dose  to  all  animals;  TDLQ—lowest toxic  dose  (other
than  inhalation);   LD^-- the  lowest   published  lethal  dose  (other  than
LD50) introduced by any other route than inhalation.
     G..   Other Relevant Information
          Except  as  reported above, no additional information was  found  on
the effects of 2, 4-toluenediamine.

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V.   AQUATIC. TOXICITY
     A.   Acute Toxicity, Chronic Toxicity, Plant Effects, and Other
          Relevant Information.
          NO  information was  found  in  the  available  literature  on acute
toxicity, chronic toxicity,  plant effects, and other relevant information.
     B.   Residues
          Veith,.  et al..  (1979),  in  a method  of estimating  the bioconcen-
tration   factor   of  organic  chemicals   in   fathead   minnows  (Pimephales
promelas), report a log bibcentration factor of 1.96 and log n-octanol/water
partition coefficient of  3.16* for the fathead minnow in  32 days'  exposure.
A  structure-activity  correlation between the bioconcentration  factor (BCF)
and  the  n-octanol/water partition  coefficient   (P)  is  expressed   by  the
equation—log BCF = 0.85 log  P-70..  According to  the  authors,, this  permits
the estimation of the bioconcentration  factor of  chemicals to within  60 per-
cent before laboratory testing..
VI..  EXISTING GUIDELINES AND STANDARDS
          NO  existing  guidelines or  standards were  found  in  the  available
literature..
*Under the same conditions  the  log  n-octanol/water partition coefficient for
heptachlor was  5.&A;  for hexachlorobenzene,  5.23;  for mirex,  6.89;  and for
dipheylamine, 3.42.

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                                  REFERENCES
Ames,  B.N., et  al.   1975.  Hair  dyes are  mutagenic:  Identification  of a
variety of  mutagenic  ingredients.  Proc. Nat. Acad. Sci. U.S.A.   72: 2423.

Cardy,  R.H.   1979.   Carcinogenicity  and chronic  toxicity  of  2,4-toluene-
diamine in  F344  rats.  Jour. Natl. Cancer Inst.  62:  1107.

Fahmy  ,  M.J.  and O.G. Fahmy.   1977.  Mutagenicity  of hair dye components
relative  to  the  carcinogen  benzidine- in  Drosphila  melanoqaster.   Mutat.
Res.   56: 31.

Giles,  A.L. and  C.W.  Chung.   1976.   Dermal carcinogenicity  study by mouse-
skin  painting with 2,4-toluenediamine alone or  in  representative  hair dye
formulations.  Jour.  Toxicol.  Environ. Health.  1:  433.
                                                  \

Gunn,  T.C.  and  s. Cooke.   1976..  Toluene In:   Chemical Economics Handbook.
Stan-  ford  Research Institute,  p. 696.5033.

Inouye, M.  and U. Murakami.   1977.   Teratogenicity of 2,5-diaminotoluene, a
hair-dye constituent  in mice.  Fd. Cosmet. Toxicol.   15: 447.

Ito,  N.,  et al.   1969.   The development  of carcinoma  in  liver of  rats
treated with m-toluylenediamine  and  the synergistic effects with other chem-
icals.  Cancer Res.   29: 1137.

Lewis,  R.J. and  R.L.  Tatken.   1979.   Registry of  Toxic Effects of Chemical
Substances.    National  Institute  for   Occupational   Safety   and  Health,
Cincinnati,  Ohio.

.Milligan,  8. and  K.E.  Gilbert.   1978.   Amines,  aromatic (diaminotoluenes).
Vol.  2.,  p.  321.  In:   M.  Grayson  (ed.),  Encyclopedia  of  Chemical  Tech-
nology.  3rd ed.  John Wiley 4  Sons, New York.

Muller, R.G.  1979.   Directory of  Chemical  Producers.  Stanford Research In-
stitute.

Shah,  M.J.,  et  al.   1977.   Comparative studies of bacterial  mutation and
hamster  cell  transformation  induced  by  2,4-toluenediamine.   Am.  Assoc.
Cancer Res.  Proc.  18: 23.

Scares,  E.R. and L.F. Lock.   1978.   The mutagenic effect  of  2,4-dinitro-
toluene and 2,4-diaminotoluene in mice.  Pharmacologist.  20: 155.

Ulland,  B.   1979.   Bioassay  of  2,4-diaminotoluene, for possible  carcino-
genicity.   NCI-CG-TR-162.   U. S.  Department of  Health, Education  and  Wel-
fare.  National  Institute  of Health.  U.S. OHEW Pub. No. (NIH) 79-1718.  -

Veith,  G.D.,  et  al.   1979.  Measuring and estimating bioconcentration factor
of chemicals in  fish.  Jour. Fish. Res. Board Canada.  36: 1040.

Weast,  R.C.  1971.  Handbook  of  Chemistry  and Physics.   51st  ed.  Chemical
Rubber Co.,  Cleveland, Ohio.
                                fl77 (\ —
                               "fr~)j -

                                 161-1

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                                        PS:33-01
                                        No. 162
        Toluene Diisocyanate

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
      WASHINGTON, D.C.  20460

         OCTOBER 30, 1980
              162-1

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                          DISCLAIMER
     This report, represents a survey of the potential health
and environmental hazards from exposure to the subject
chemical.  The information contained in the report is drawn
chiefly from secondary sources and available reference
documents.  Because of the limitations of such sources/ this
short profile may not reflect all available information
including all the adverse health and environmental impacts
presented by the subject chemical.  This document has undergone
scrutiny to ensure its technical accuracy.
                            162-^

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                                                   PS:33-01
                     TOLUENE DIISOCYANATE



                           Summary





     Toluene diisocyanate (TDI) is used in the manufacture of



polyurethane foam.  TDI is formed through the reaction of 2,4-



toluenediamine with phosgene.  The TDI is then reacted with



di- and poly-functional hydroxy compounds to form polyurethane



foam.



     TDI is readily reactive in water, forming carbon dioxide



and polyurea derivatives.  Environmental occurence of TDI is



unlikely due to its high reactivity with hydroxy compounds



and peroxy radicals.



     Information on the teratogenicity of toluene diisocyanate



was not found in the available literature.   TDI after being



tested by the National Cancer Institute for carcinogenicity



using a standard bioassay protocol,  was found not be



carcinogenic.  Additionally, toluene diisocyanate did not



show mutagenic activity on testing of Salmonell typhimurium
*                                      ~


strains with and without a mammalian liver microsome activating



system.



     Extensive toxicologic data exists for TDI,  primarily



from occupational exposure studies.   TDI produces respiratory



effects,  including mucous membrane irritation,  bronchoconstriction,



coughing,  and wheezing.   Exposure to high concentrations can



result in pulmonary edema or death.
                           162-3

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     The effects of chronic, low-level exposure to TDI vary.
Decreased lung function has been reported from inhalation
of 0.003 ppm TDI, but. other investigators have not seen these
respiratory effects from inhalation of 0.02 ppm TDI.
Hypersensitivity'to TDI has also been observed from occupational
respiratory exposure*  Immunologic and pharmacologic reactions
have been proposed as the mechanism of action of TDI.
     Other reported effects include memory loss,  psychological
disturbances, and skin irritation.  Uncertainty exists regard-
ing the frequency of these effects in those occupationally
exposed.  Maintaining exposure below 0.005 ppm has proven
effective in protecting health of unsensitized workers.
Where an individual has previously been sensitized, a no-
threshold effect is indicated upon subsequent exposure to
TDI.
                            162-4

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                     TOLUENE DIISOCYANTE
         •             Environmental Fat®

     Toluene diisocyanate (TDI) readily reacts with hydroxy
compounds.  Its atmospheric half-life is approximately three
days (Brown-, et al. 1975).  TDI readily hydrolyzes in neutral
aqueous media, or more rapidly under acidic or basic conditions,
to give unstable carbamic acids (Tennant, 1979).  These acids
tend to lose carbon dioxide, giving the corresponding amine
which,  in turn, reacts with the starting isocyanate to produce
a urea derivative.  This reaction produces a concurrent
decrease in pH (Curtis, et al. 1979).  TDI readily hydrolyzes
in water, and has a half-life of 0.5 seconds to 3 days, depending
on pH (Brown, et al., 1975).  As temperature increases the
reaction becomes more vigorous (Tennant, 1979).
     Brown, et al. (1975) concluded that because of the
short lifetime of toluene diisocyanate in water, its occurrence
in this medium is unlikely.
'     Toluene diisocyanate is persistent in the atmosphere.
Under atmospheric conditions reaction with ozone leads to an
atmospheric half-life of 3,981 days.  The reaction of TDI
with peroxyradical groups has an environmental half-life of
approximately 7.94 x 105 days in the water phase.
                            162-5

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I.   INTRODUCTION
     This profile is based upon relevant literature identified
through bibliographic searches in TOXLINE and Chemical
Abstracts, and through manual searches.  The National Insti-
tute for Occupational Sefety and Health (NIOSH) has published
a criteria document for diisocyanates (NIOSH-, 1978).  This
report represents a comprehensive review of the available
toxicologic literature on toluene diisocyanate (TDI) and was
the source for much of the data described below.
     Toluene diisocyanate is also reported as 2,4-diisocyanate-
1-methylbenzene, tolylene diisocyanate, raethylphenylene
isocyanate, diisocyanotoluene, and stilbene diisocyanate.
The compound is a colorless-to-pale yellow liquid.  The
chemical formula is CgHgN202.  Physical properties of TDI are
as follows:  molecular weight, 174.16; melting point,  20 to
22°C; boiling point, 251eC; vapor pressure,  0.05 ram Hg at
25'C; and specific gravity, 1.22 at 25°C (NIOSH,  1978).  TDI
is soluble aromatic hydrocarbons,  nitrobenzene, acetone,
ethers, and esters.
     The most common method of synthesizing toluene diisocyanate
is through the primary reaction of diaminotoluene with
phosgene.  Toluene diisocyanate is then reacted with di- and
poly-functional hydroxy compounds  to form poly-urethane foams»
coatings, elastomers,  and spandex  fibers (NIOSH,  1978).
                            162-6

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     Toluene diisocyanate production in the U.S. was 605
million pounds  (Predicasts, Inc., 1980) in 1978, with an
estimated 6.4 percent annual growth in production.  Production
capacity amounted to 775 million pounds per year in 1978.

II.  EXPOSURE '
     Respiratory and dermal exposure to toluene diisocyanate
has been well documented in occupation environments (NIOSH,
1978).  Sources of occupational exposures include production
processes of basic TDI manufacture, production of polyurethane
foam, and accidental releases or spills in product synthesis,
transportation, use, or disposal.
     Non-occupational exposure to TDI through ingestion of
contaminated food or water is unlikely since TDI released to
the environment would readily react with other compounds,
forming stable polyurea end products.   For example,  Curtis,
et al. (1979) conducted acute aquatic toxicity studies of TDI
and reported the immediate reaction of TDI with water resulting
in the production of carbon dioxide and a polyurethane foam-
like solid.   Human exposures would most likely occur to these
polyurea compounds and not TDI.  Accidental releases and
spills may result in respiratory TDI exposure vof persons in
the immediate vicinity.   Dermal exposure may also occur in
persons coming in direct contact with the compound.
                            162-7

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III. PHARMACOKINETICS
     Information on the abosrption, distribution/  metabolism,
and excretion of TDI was not identified in the available
literature.  NIOSH (1978), in describing the sensitization
phenomenon of TDI exposure, hypothesized that this response
may be the result of TDI reaction with in vivo hydroxyl,
amino, sulfhydryl, or similar compounds which form a hapten
complex with TDI.  This complex is believed to be responsible
for the sensitization of individuals to TDI.
                            162-8

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IV.  EFFECTS


     A.  •Carcinogenicity


         TDI did not show carcinogenic activity after being


tested by NCI using a standard bioassay protocol.


     B.  Mutagenicity


         Toluene diisocyanate did not show mutagenic activity


on testing Salmonella typhimurium strains with or without a


mammalian liver microsome activating system -(NIOSH, 1978).


     C.  Teratogenicity and Other Reproductive Effects


         Information on teratogenic or other reproductive


effects of toluene diisocyanate was not found in the available


literature.


     D.  Chronic Effects


         Inhalation of toluene diisocyanate represents the


primary route of exposure and produces chronic effects; the


mechanism of the chronic respiratory changes is uncertain.


     Toluene diisocyanate induces a hypersensitive reaction


in specific individuals.  Predisposing factors may include
*

both environmental and endogenous host factors (AdTcinson,


1977).  Intensity and duration of exposure are important in


eliciting a hypersensitive reaction.  Genetic factors


controlling immune responsiveness,  metabolic aberration were


suggested as factors influencing the allergic reaction


(Adkinson,  1977).  However,  Butcher, et al.  (1976) found no


pattern of prior hay fever or asthama,  or of skin sensitization


in clinically sensitized individuals.



                            162-9

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     Exposure to high concentrations has caused respiratory
sensitization in workers  (Walworth and Virchow, 1959; Bruckner,
et al. 1968).  These sensitization reactions were described
earlier.  The sensitization can progress to a condition
resembling chronic bronchitis and pulmonary edema.  Individuals
sensitized to TDI present an asthmatic reaction upon reexposure
to very low concentrations of TDI.  Butcher, et al. (1979)
described four specific types of responses in hypersensitive
workers:  (1) immediate;  (2) late; (3) dual; and (4) dose-
related.  The responses were measured as percent change in
one-second Forced Expiratory Volume (PEV]_) over time.
Immediate response occurred within one hour of exposure,
whereas late response exhibited a gradual decline in FEVi
over five hours.  The dual response elicited an early response
within one hour and a late response after eight hours.  The
dose-related response was exhibited at 0.01 ppm,  whereas
exposure to 0.005 ppm did not. show a significant decrease in
FEVi-.  T*16 author suggested a pharmacologic basis for the
hypersensitivity, but noted that an allergic mechanism could
not be ruled out.
     Porter, et al. (1975) reported sensitization correlated
with the frequency and severity of significant exposures
greater than 0.05 ppm.   Once sensitized, an individual exposed
to very low concentrations of TDI will produce asthmatic
reactions upon subsequent TDI exposure.

                           162-10

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     Wegman (1977)  reported decrements  in FEVi in both



 sensitized  and unsensitized workers.  However, Adams  (1975)



 and Butcher, et,al.  (1977) did not  show decreased FEV^ after



 occupational exposures of 11 and 2.5 years, respectively.



 TDI concentrations were 0.02 ppm and below, with occasional



 excursions  above this level»  Consequently, the National



 Institute for Occupational Safety and Health  (NIOSH)  recommended



 an eight hour time-weighted average limit of  5 ppb, noting



 that the above studies and others had not reported significant



 effects on  lung  function at concentrations of 14-50 ug/m3



 (2.0-7.0 ppb).



     Some authors have reported skin sensitization in persons



 occupationally exposed to TDI (Nava, et al. 1975; Karol, et al.



 1978), but  other investigators have not observed such skin



 sensitization reactions (Munn, I960.; Bruckner, et al. 1968).



     Other  chronic effects from TDI exposure  include neurologic



 effects, eye irritation, and psychological symptoms.   Le



 Quesne, et  al. (1976) reported memory loss lasting 4 years in
«


 workers exposed to massive concentrations of TDI while fighting



 a fire at a polyurethane foam factory.





     F.  Acute Effects



         Inhalation of TDI is the primary route of exposure



which has demonstrated acute effects.   Several authors have



 reported daily and cumulative decreases in lung function



 following respiratory exposure to TDI.   Investigations of









                            162-11

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acute  effects  from TDI exposure have produced contradictory
results. * Peters, et al.  (1968) reported significant decreases
in  lung  function upon exposure to 0.1-3.0 ppb, whereas Adams
(1975) noted no significant decrease in lung function at 20
ppb.
     Occupational exposure to high concentrations of TDI
causes direct  irritation of the respiratory tract (Walworth
and Virchow, 1959; Maxon, 1964; Axford, et al. 1976? Gandevia,
1963).
     Eye, nose, and throat irritation was observed upon
atmospheric exposures to 500 ppb (Henschler, lf62).   Nausea,
vomiting, and  abdominal pain may also occur (Key, et al.
1977).  Dermal contact with liquid TDI may produce redness,
swelling, and blistering.  Contact with eyes may produce
severe irritation and permanent damage.  Ingestion of TDI may
cause burns of the mouth and stomach (Key, et al. 1977).
     Lewis and Tatken (1979) reported an inhalation LCso for
rats' of 600 ppm following a 6-hour exposure; and an inhalation
     f°r mice. of- 10 ppm following a 4-hour exposure.
V.   AQUATIC TOXICITY
     A.  Acute Toxic ity
         Curtis, et al. (1979) reported a 96-hour LCso of
164.5 mg/1 in the fathead minnow (Pimephales promelas).  No
significant mortality was noted in grass shrimp (Palaemonetes
pugio) exposed to 508.3 mg/1.  The authors noted that TDI


                           162-12

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reacted with water of dilution, and concluded that TDI was
toxic to 'the fathead minnow in the unreacted form only, as
evidenced by all mortalities occurring during the first 12
hours of the test.  However, the authors did note that a
concurrent decrease in pH was observed as a result of carbon
dioxide formation from TDI reactivity.  Lewis and Tatken
(1979) reported an aquatic toxicity rating, TLmgg (equivalent
to a 96-hour LCso)' of 1.0-10.0 ppm.  Thus TDI has moderate
acute toxicity to aquatic organisms.

     B.  Chronic Toxicity, Plant Effects, and Residues
         Pertinent data could not be located in the available
literature.
VI.  EXISTING GUIDELINES AND STANDARDS
     The Occupational Safety and Health Administration (OSHA)
regulates TDI by specifying a PEL for airborne TDI of 0.14
mg/ m3  (40 CPR 1910.1000) as a 15-minute exposure.
     The American Conference of Governmental Industrial
Hygienists (1979) has recommended a threshold limit value-time
weighted average for toluene diisocyanate of 5 ppb (0.04
mg/m3).  NIOSH (1978)  recommended a time-weighted-average
limit-for airborne toluene diisocyanate of 5 ppb,  with a
ceiling value of 20 ppb.   NIOSH (1978) also reported occupa-
tional exposure limits for TDI in numerous countries.   These
limits ranged from 0.07 to 0.5 mg/m3.
                            162-13

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                     TOLUENE DIISOCYANATE
                          References
Adams, W.G.F.   1975.  Long-term effects on the health of
men engaged in  the manufacture of tolylene di-isocyanate.
Br.  Jour. Ind. Med.  32: 72.
        . *
Adkinson, N.F.  1977.  Environmental influences on the
immune system and allergic responses.  Environ. Health
Perspect.  20:  97.

American Conference of Governmental Industrial Hygienists.
1979.  Threshold limit values for chemical substances and
physical agents in the workroom environment with intended
changes for 1979.  American Conference of Governmental
Industrial Hygienists.  Cincinnati, Ohio, p. 94.

Axford, A.T., et al.  1976.  Accidental exposure to
isocyanate fumes in a group of firemen.  Br. Jour. Ind.
Med.  3: 65.

Brown, S.L., et al. 1975.  Research program on hazard
priority ranking of manufactured chemicals.   Phase II-
Final Report.   NTIS PB-263162.

Bruckner, H.C. , et al.  1968.  Clinical and inununologic
appraisal of workers exposed to diisocyanates.  Arch.
Environ.  Health  16: 619.

Butcher, B.T.,  et al. 1976.  Toluene diisocyante (TDI)
pulmonary disease—immunologic and inhalation challenge
studies.  Jour.  Allergy.  Clin.  Immunol.  58: 89.

Butcher, B.T.,  et al. 1977.  Longitudinal study of workers
employed in the manufacture of toluene-diisocyanate.  Am.
Rev. Resp. Dis.  116: 411.

Butcher, B.T., et al. 1979.  Inhalation challenge and
pharmacologic studies of toluene diisocyanate (TDI)—
sensitive workers.  Jour.  Allergy.  Clin. "Immunol.
64: 146.•

Curtis, M.W.,  et al. 1979.  Acute toxicity of 12 industrial
chemicals to freshwater and saltwater organisms.  Water
Res.  13:  137.
                            162-14

-------
Gandevia,  B.  1963.   Studies of ventilatory capacity and
histamine  response during exposure to isocyante vapour
in polyurethane 'foam manufacture.  Br. Jour. Ind. Med.
20: 204.

Henschler, D'., et al.   1962.  The toxicology of the
toluene diisocyanates.  Arch. Toxicol.  19: 364.

Karol, M.H.,  et al.  "1978.  Tolyl-specific IgE antibodies
in workers with hypersensitivity to toluene diisocyanate.
Am. Ind. Hyg. Assoc. Jour. 39: 454.

Key, M.M., et al. 1977.  Occupational diseases—a guide
to their recognition.   National Institute for Occupational
Safety and Health.   Cincinnati, Ohio. p. 233.

LeQuesne, P.M., et al.  1976.  Neurological complications
after a single exposure to toluene diisocyanate.  Br.
Jour. Ind. Med.  33: 72.

Lewis, R.J. and R.L. Tatken (ed.) 1979.  Registry of
toxic effects of chemical substances.  National Institute
for Occupational Safety and Health.  Cincinnati, Ohio-.
U.S. Government Printing Office, Washington, D.C., p. 180.

Maxon, F.L. 1964. Respiratory irritation from toluene
diisocyanate.  Arch.  Environ. Health. 8: 755.

Munn, A. 1960.  Experience with diisocyanates.  Trans.
Assoc. Ind. Med. Off. 9: 134.

National Institute for  Occupational Safety adn Health.
1978.  Criteria for  a recommended standard:  Occupational
exposure to diisocyanates.  National Institute for
Occupational Safety  and Health.   Cincinnati, Ohio, p. 138.

Nava, C., et al. 1975.  Pathology produced by isocyanates-
raehtods of immunological investigation.   Ric. Clin. Lab.
5: 135.

Peters, J.M., et al. 1968.  Acute respiratory effects in
workers exposed to low levels of toluene diisocyanted (TDI.
Arch. Environ. Health.  16:642.

Porter, C.V., et al. 1975.  A retrospective study of
clinical, physiologic and immunologic changes in workers
exposed to toluene diisocyanate.   Am. Ind.  Hyg. Assoc.
Jour. 36: 159.

Predicasts, Inc. 1980.  Predicast No. 78 (2nd Quarter)
Jan. 18, 1980.  Predicasts Inc.,  Cleveland Ohio.
                            162-15

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Tennant, G. 1979. Imines, nitrones, nitriles and
isocyanates.  In; D. Barton, W.d. Ollis (eds.)  Compre-
hensive Organic Chemistry, Vol. 2: Nitrogen compounds,
carboxylic acids, phosphorus compounds.  Pergamon Press*
New York, p. 521.

Walworth, H.T. and W.E. Virchow.  1959.  Industrial
hygiene experience with toluene diisocyanate.  Am. Ind.
Hyg. Assoc. Jour.  20: 205.

Wegman, D.H,, et al. 1977.  Chronic pulmohary function
loss from exposure to toluene diisocyanate.  Br. Jour.
Ind. Med. 34: 196.
                            162-16

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                                      No. 163
             Toxaphene

  Health and Environmental Effects
U.S.. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

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                          DISCLAIMER
     This report represents  a survey of the potential health
and environmental hazards front exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and  available reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and environmental impacts  presented by  the
subject chemical.   This  document  has undergone  scrutiny  to
ensure its technical accuracy.

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                       SPECIAL NOTATION










O.S. EPA1s Carcinogen Assessment Group (GAG) has evaluated



toxaphene and has found sufficient evidence to indicate



that this compound is carcinogenic..

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                          TOXAPHENE
                           SUMMARY
     Toxaphene is a mixture of polychlorinated camphenes.
It is obtained from camphene by photochemical chlorination,
which produces a heterogeneous mixture  of  chemicals (177)
containing 67 to 69" percent chlorine.   Toxaphene has not
produced tec.at.og.en.ic effects in laboratory animals, but
has been found to be mutagenic in two strains of Salmonella
typhimurium with metabolic activation.   A  National Cancer
—i»a^—^—«-"^—•                               \
Institute (NCI) 1979 -study found that toxaphene signifi-
cantly increased the incidences of hepatocellular carcinomas
in mice.
     The. insecticide, toxaphene has been demonstrated to
be a potent toxin to a variety of aquatic  life.   For both
freshwater and marine fish species,  acute  toxicity values
of 0.8 to 28 ug/1 were reported.  Marine invertebrate species
displayed considerable interspecies  variation,  with LC5Q
values ranging from 0.08  to 2,700 jig/1..

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                          TOXAPHENE
I.   INTRODUCTION
     This profile is based on the Ambient Water Quality
Criteria document for Toxaphene  (U.S. EPA, 1979).
     Toxaphene is a commercially produced, broad spectrum,
chlorinated hydrocarbon consisting primarily of chlorinated
camphene and related compounds and. isomers.  It is currently
the most heavily used insecticide in the U.S., with an annual
production rate exceeding 50 x 10   tons  (-U.S. EPA, 1979).
     On May 25, 1977, because of its carcinogenic effects,
aquatic toxicity, and high bioconcentration factor, the
U.S.. EPA issued a notice of rebuttable presumption against
registration and continued registration of pesticide pro-
ducts containing toxaphene.
     Toxaphene is an amber, waxy solid with a mild terpene
odor and an average molecular weight of 414.  Its physical
properties include: melting point of 65-90°C; vapor pres-
sure, 0.17-0.40 ram Hg at 25°C; solubility in. water, 0.4-
3.0 mg/1; and is soluble in relatively non-polar solvents,
with an octanol/water partition coefficient of 825 (U.S
EPA, 1979).
     The commercial product is relatively stable but may
dehydrochlorinate upon prolonged exposure to sunlight, alkali,
or temperatures above 120°C (Metcalf, 1966';-. Brooks, 1974).
In natural water systems, toxaphene tends to be absorbed
by the particulates present or to be taken up by living
organisms and bioconcentrated.  Thus, it is seldom found
as a soluble component in receiving waters but can persist
                              y

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in sediments oc remain absorbed on  suspended  solids  for
prolonged periods  (U.S. EPA, 1979).
II.  EXPOSURE
     A.   Water
          Toxaphene has been monitored  in  the U.S. since
1959.  Although it has been detected at  several  locations,
it is not found in all'waters  (U.S. EPA, 1979).   Seven  rou-
tine monitoring studies of U.S., surface  water prior  to  1975
did not detect toxaphene  (U.S. EPA, 1.B19) ."
          Nicholson, et al.  (1964,  1966) detected toxaphene
in the drink-ing water obtained from Alabama at levels rang-
ing from 0.01-0.1 ug/1.  A survey of commercial  drinking
water samples by the U.S. EPA  (1976a) during  1975 and 1976
found no detectable levels of toxaphene  (limit of detection
0.05 ug/1).
          Toxaphene has been detected in water around areas
where it is applied to crops as an  insecticide.   For example,
it has been detected in surface waters  in  California at
levels ranging from 0.02 to 7.9 jig/1, and  in  drainage ef-
fluents at levels of 0.130 to 0.950 ;ag/l (Johnston,  et  al.
1967; Bailey and Hammon, 1967).  Several studies of  an  agri-
cultural watershed in Alabama found that treatment of drink-
ing water, did not reduce toxaphene  concentrations (U.S.
EPA, 1979) .                                 ';.
          Toxaphene has been detected in the  sediment samples
of various waters even when  it is not found in samples  of  *
the surface waters (Mattraw, 1975).  Concentrations  as  high
as 2.46 ug/1 have been found in sediments  (U.S.  EPA, 1979).

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Sediment samples at three locations downstream of a plant
producing toxaphene had a maximum- residue level of 15 ;ig/l
toxaphene before dredging (Reimold and Durant, 1972) .
     B.   Food
          The best available estimate of dietary intake
of toxaphene is 0.021 pg/kg/day, based on the U.S. Food
and Drug Administration basket survey between 1964 and 1970
(Duggan and Corneliussen, 1972) .  Based on recent market
basket, surveys indicating a decrease in the incidence of
toxaphene contamination, a stable incidence of toxaphene
in raw meat since 1969, and a two-fold increase in the inci-
dence of toxaphene in unprocessed food samples between 1972
and 1976, the U.S. EPA  (1979) estimates the current dietary
intake to be 0.042 jag/kg/day.
          The U.S. EPA  (1979) has estimated the weighted
average bioconcentration factor for toxaphene to be 18,000
for the edible portions of fish and shellfish consumed by
Americans.  This estimate was based on the measured steady-
state bioconcentration studies in five species of fish and
shellfish.
     C.   Inhalation
          The highest toxaphene residues in air have been
found in areas where toxaphene is applied for agricultural
purposes  (especially cotton production in the Southern U.S.)
(U.S. EPA, 1979).  Studies indicate that airborne residues
are highest during cotton growing season and decrease to
low levels after harvesting, but spring tilling releases
                        /61-7

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soil residues to the air.  Concentrations  ranging  from 0
to 2520 ng/m  have been measured  in southern  agricultural
areas  (Arthur, et al. 1976; Stanley, et al.. 1971.)  Mean
monthly concentrations have been  measured  as  high  as  167
ng/m3  (Arthur, et al.  1976).
          Toxaphene* has also been monitored in  the atmos-
phere over the east coast near Bermuda and the  open ocean
(Bidleman and Olney, 1975).  The  mean concentrations  were
0.79 and 0.53 ng/m , respectively.  Using .the maximum mean
monthly concentration of 167 ng/m  (Arthur, et  al.  1976),
the average daily dose of toxaphene from air  is. approximately
0.057 ug/kg (U.S. EPA, 1979)*  This amount would reflect
intake at a high toxaphene use area, whereas  a  more conserva-
tive value using a concentration  of 0.53 ng/m  monitored
over open ocean  (Bidleman and Olney, 1975)' would be an aver-
age da-ily intake of 0.18 ng/kg of toxaphene from air  (U.S.
EPA, 1979).
     0.   Dermal
          Toxicity studies with laboratory animals  indicate
that toxaphene can be absorbed across the  skin  in  toxic
amounts by humans (U.S. EPA, 1979).  Incidence  of  dermal.
absorption of toxaphene by humans is restricted to  occupa-
tional or accidental exposure.
III. PHARMACOKINETICS
     A.   Absorption
          The recently completed  U.S. EPA  (1978) study sug-
                                                           »
gests  that inhalation exposures to toxaphene  do not result
in sufficient absorption by humans to cause quantifiable
levels in the blood.

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          Animal studies show absorption of  toxaphene  across
the alimentary tract, skin, and respiratory  tract,  as  indi-
cated by adverse effects elicited by oral, dermal,  and in-
halation exposures (U.S. EPA, 1979).  The. vehicle and  mode
of administration, as well as individual differences,  affect
the rate of absorption of toxaphene.  The ratio of  oral
LDeg- to. dermal LDgQ  (in comparable lipophilic solvents)  is
about 0.1 (Lackey, 1949a,b; Conley, 1952; U.S. EPA, 1979).
     B.   Distribution                    N
          Toxaphene  is readily distributed throughout  the
body, with highest residues found in fat tissue.  Three
hours after single intubations of Cl-36 labelled toxaphene,
rats had .detectable  levels of Cl-36 activity in all tissues
examined (kidney, muscle, fat, testes, brain, blood, liver,
intestines, esophagus, spleen, and stomach), with the  highest
levels being found in the stomach and blood  (Crowder and
Dindal, 1974.)  After 9 to 14 days, most of  the activity
is found in the fat, blood, kidney, liver, and intestines
(Crowder and Dindal, 1974; Ohsawa, et al. 1975).  The  pre-
dominance of fat storage had been demonstrated in 12-week
feeding studies with rats, and 2-year feeding studies  with
rats and dogs (Clapp, et al. 1971; Lehman, 1952; Hercules,
Inc., undated).  In  the above studies, toxaphene residues
were highest in fat  tissues but always remained below  the
levels administered  in the diet, thus suggesting that  toxa-
phene is not biomagnified in terrestrial organisms  (U.S.
EPA, 1979).

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     C.   Metabolism
          Toxaphene undergoes reductive dechlorination,
dehydrochlorination, and hydroxylation in mammalian  systems
(U.S. EPA,' 1979).  Studies by Crowder and Dindal  (1974),
Ohsawa, et al.  (197.5) and Khalifa, et al.  (1976)  have ob-
served 50 percent dechlorination of toxaphene after  adminis-
tration by intubation to rats, or iri vitro with rat  liver
microsomes and NADPH under anaerobic conditions.  Toxaphene
has been suggested as a substrate for the hepatic microsomal
mixed-function oxidases because of type I binding spectra
with cytochrome P-450, and NADPH dependence  (Kulkarni, et
al- 1975; Chandurkar, 1977).
          Several investigators have noted that fat  residues
of toxaphene resemble whole toxaphene, while residues in
both the liver and feces are consistently more polar (Pollock,
1978; Saleh, et al. 1977).
     D.   Excretion
          The half-life of C-14 or Cl-36 labelled toxaphene
in rats after single oral doses appears to be from one to
three days, with most of the excretion occurring  via the
urine and feces  (Crowder and Dindal, 1974; Ohsawa, et al.
1975),  Only a small portion of the: urine and fecal metabo-
lites is eliminated as glucuronide or sulfate conjugates
(Chandurkar, 1977).
          A study of the blood levels of toxaphene in an
individual consuming contaminated fish (52 pg toxaphene/g
fish) revealed levels of 142 ppb, 47 ppb,<30 ppb on day
1, day 11, and day 14 of measurement (U.S. SPA, 1978).

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IV.  EFFECTS
     A.   Carcinogenicity
          The National Cancer Institute  (1979) has recently
completed a carcinogenicity bioassay of  toxaphene.  The
80-week feeding study did not follow current NCI standards;
only ten animals were used in each matched control groupr
and: matched-fed control groups were not  utilized (NCI, 1977).
The feeding schedule was as follows: for rats - males, time
weighted average (TWA) doses at 556 mg/kg and 1,112 mg/kg,
                                          %
and females, TWA doses at 540 mg/kg and  1,080 mg/kg; and for
mice, males and females, TWA doses at 99 mg/kg and 198 mg/kg.
          In male rats in the high dose  group, a significant
increase was noted in the incidence of follicular-cell car-
cinomas and adenomas of the thyroid.  Of the nine thyroid
tumors which were found in this group, two were carcinomas.
A^ significant increase of follicular-cell. adenomas of the
thyroid was also noted in the high-dose  group of female
rats.  No carcinomas of the thyroid were found in this group.
In both of these groups, the development of thyroid tumors
was dose-related-
          In both male and female mice,  significant increases
were noted in the incidence of hepatocellular carcinomas
and in the incidence of hepatocellular carcinomas combined
with neoplastic nodules of the liver.
          Based on the results of this study, the National
Cancer Institute has concluded that "Toxaphene was carcino-..
genie in male and female B6C3F1 mice, causing increased

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incidences of hepatocellular carcinomas.  The  test  results
also suggest carcinogenicity of tbxaphene for  the thyroid
of male and female Osborne-Mendel rats"  (NCI,  1979).
          Litton Bionetics, Inc.  (1978)  also reported  a
significant excess of hepatocellular tumors  (hepatocellular
adenoma plus hepatocellular carcinoma) in male mice  fed
dietary levels of 50 ppm toxaphene.
     B.   Mutagenicity
          The mutagenicity of toxaphene  has been tested
in bacterial systems using Salmonella typhimurium strains
TA153S, TA1S37, TA1538, TA98, and TA100  (Hill, 1977).  Posi-
tive test results were obtained for strains TA98 (frameshift
mutation) and TA100  (base pair substitution) only in tests
without metabolic activation.  All other tests were nega-
tive.  A "high temperature" toxaphene has elicited positive
dose response increases in strains TA98  and TA100 only with
metabolic activation.  In other studies, toxaphene and toxa-
phene subfractions have been found to be mutagenic  to  strain
TA100 with or without metabolic activation  (Hill, 1977).
          A study conducted by the U.S.  EPA  (1973)  found
no significant differences in the rates  of chromosomal aber-
rations in leukocytes between groups of  workers occupation-
ally exposed to toxaphene and those not  exposed.
     C.   Teratogenicity
          Toxaphene did not produce teratogenic effects
when administered in the diet of rats, mice, and guinea
pigs (U.S. EPA, 1979).  Kennedy, et al.  (1973) found no
indication of teratogenic effects in F3  weanlings of rats

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fed toxaphene at levels of 25 mg/kg diet and 100 mg/kg diet.
Pregnant rats and mice fed 15 to 35 mg/kg/day of toxaphene
produced young with no teratogenic effects as did pregnant
guinea pigs fed 15 mg/kg body weight  (Chernoff and Carver,
1976;. DiPasquale, 1977)..
     D.   Other Reproductive Effects
          Adverse effects on fertility, gestation, viability,
lactation, or survival indices were not observed in male
and female rats fed dietary levels of 25 mg/kg and 100 mg/kg
toxaphene (Kennedy, et al. 1973), or in mice fed dietary
levels of 25 mg/kg toxaphene (Keplinger, et al. 1970).
     E.   Chronic Toxicity
          Long term exposures to low dietary levels of toxa-
phene have been, investigated in several studies involving
rats, dogs, and monkeys (U.S. EPA, 1979).  All studies noted
some form of liver pathology in rats at dietary levels of
100 mg/kg or above-  At 100 mg/kg, cytoplasmic vacuolization
was noted by Kennedy, et al. (1973).  Increased liver weight
with minimal liver cell enlargement was noted in rats at
dietary levels of 25 rag/kg (Fitzhugh and Nelson, 1951).
The lowest dietary level of toxaphene producing unequivocal
liver damage over a two-year feeding period was 20 mg/kg
(U-S. EPA, 1979)-  Only at high concentrations, i.e., 1,000
mg/kg diet, does toxaphene elicit central nervous system
effects (Hercules, Inc., undated).
                                                           »•
     F.   Other Relevant Information
          Induction of hepatic microsomal mixed-function
oxidase (MFO) appears to account for most of the interactions

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of toxaphene with other compounds  (U.S. EPA,  1979).   Pre-
treatment with known MFO inducers, such as DDT,  aldrin,
and dieldrin, increases oral LCgg's  two to three-fold (Deich-
man and Ke'plinger, 1970) .  Piperonyl butoxide, which  inhibits
the metabolism of many toxicants by  MFO, has  been  shown
to potentiate the toxicity of toxaphene in houseflies (Saleh,
et al. 1977).
          Keplinger and Deichmann  (1967) found that equitoxic
combinations of toxaphene with parathion, diazinon, or tri-
thion were less toxic than expected  based on  the assumption
of simple similar action.
          Acute human intoxication by toxaphene-lindane
mixtures produces signs and symptoms that are not  character-
istic of toxaphene or lindane poisoning (Pollock,  1958;
Masumura, 1975).                            .
V-.   AQUATIC TOXICITY
     A.   Acute
          Acute toxicity data of toxaphene to freshwater
fish are derived from 52 96-hour LC^Q values  for 18 species
resulting from 48 static and 4 flow-through assays.   Observed
LCgg values for these species of fish range from 0.8  pg/1
for the channel catfish (Ictalurus punctatus) to 28 ug/1
for the goldfish, (Carassius auratus) (U.S. EPA, 1979).
No single family or species appeared to be dramatically
more resistant or sensitive to toxaphene.  For freshwater
invertebrates, 17 static bioassays on 13 species resulted  '
in reported LC^Q values of 1.3 ug/1  for the stonefly  (Cla-
asenia sabulosa) to 178 pg/1 for the crayfish (Procambarus
simulans) (U.S.  EPA, 1978).

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          For the marine fish,  toxicity  data  were  determined
from five flow-through and two  static assay procedures  repre-
senting six species.  Observed  LC50 values ranged  from  0.5
ug/1 for the pinfish  (Lagodon rhomboides) to  4.7 ^ig/1 for
the threespine stickleback (Gasterosteus aculeatus)  (U.S.
EPA, 1979).  The toxicity of toxaphene to marine inverte-
brates shows considerable interspecific variation  in 31
assays (10 flow-through and 21  static) with reported LC5Q
values ranging from 0.054 pg/1  for larval stages of the
driftline crab (Sesarma cineseum) to 2,700 pg/l-for the
blue crab (Callinecten sapilus).
     B.   Chronic
          Chronic life cycle toxicity tests have produced
chronic values of 0.037 and 0.059 ^ag/1 for the fathead  min-
now (Pimephales promelas) and channel catfish  (Ictalurus
punctatus) , respectively (Mayer, et al.. 1977) .  Growth  ef-
fects were noted in brooktrout chronically exposed to concen-
trations of 0.038 ug/1.  Life cycle tests on  freshwater
invertebrates have been performed on three species with
chronic values of 0.09, 0.18, and 1.8 jig/1 reported for
Daphnia magna; the scud (Gammarus pseudolimnaeus); and midge
larvae (Chironomus plumosus), respectively (Sanders, in
press).-  An embryo-larval test on the marine  fish sheeps-
head minnow (Cyprinodon variegatus)  produced a chronic value
of 0.83 pg/1  (Goodman, et al. 1978).  A chronic value of
0.097 pg/1 was obtained for the marine mysid  shrimp (Mysi- .
dopsis bahia)  (Nimmo, 1977).
                               -is

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     C.   Plant Effects
          No data for the effects of toxaphene were  found
for freshwater species.  Effective concentrations  for  five
species of marine plants ranged from 0.15 ug/1 for reduced
growth in the dinoflagellate  (Monochrysis lutheri) to  ISO
ug/1 for lethality in the dinoflagellate  (Danaliella euchlora)
arid no growth of the algae  (Protococcus) sp.  (U.S.. EPA,
1978).
     0.   Residues                        >.
          Bioconcentration  factors for three  species of
fish were reported (Mayer,  et al. 1975; Mayer, et al.  1977).
Brooktrout fry (Salvelinus  fontinalis) had the highest fac-
tor of 76,000 in 15 days, while yearling brooktrout had
the lowest factor of 16,000 in 161 days.  In  the marine
longnose killif ish (Fundulas similis) , bioconcentrations
for a number of different life stages were reported as 29,450
for juveniles, 27,900 for fry, 5,400 for adults, and 1,270
to 3,700 for ova of exposed adults (Schimmel, et al. 1977).
VI.  EXISTING GUIDELINES AND STANDARDS
     Neither the human health nor aquatic criteria derived
by U.S. EPA  (1979), which are summarized below, have gone
through the process of public review; therefore, there is
a possibility that, these criteria will be changed.
     A.   Human
          The standards for toxaphene in air , water , and
food which have been established or recommended by various *
groups and agencies were set before the results of the NCI
bioassay for carcinogenicity were available  (U.S. EPA, 1979).
                           ~in/Li-
                         m i t * >

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The ACGIH (1977) recommends a time weighted average value
of 500 mg/m  for the working environment and a tentative
short-term exposure limit of 1 mg/m .  The national interim .
primary drinking water standard for toxaphene is 5 ug/1
(40 PR 11990; U.S. EPA, 1976b, 1976c).  The National Academy
of Sciences  (1977) .estimated the acceptable daily intake
of toxaphene for man at 1.25 jig/kg and suggested no-adverse-
effect levels from water at 8.75 pg/1 (assigning 20 percent
of the total ADI to water) or 0.44 jag/1 (assigning 1 percent
of the total ADI to water).  Effluent standards for toxa-
phene manufacturers have been set at 1-5 pg/1 for existing
facilities and 0.1 p;/l for new facilities (U.S.  EPA, 1976a).
Tolerances established by the U.S. Pood and Drug Administra-
tion for toxaphene in various agricultural products range
from 0.1 mg/kg- in sunflower seed's to 7 rag/kg in meat fat
(U.S. EPA, 1979).
          The U.S. EPA (1979) draft water quality criterion
                                       —4
for toxaphene is 0.467 ng/1 or 4.7 x 10   pg/1.  This cri-
terion is based on the NCI (1979)  study that reported hepato-
cellular carcinoma and neoplastic nodules in mice fed toxa-
phene; the criterion was calculated to keep the lifetime
cancer risk below 10   for humans.
     B»   Aquatic
          A drafted criterion for the protection of fresh-
water aquatic organisms is 0.007 ug/1 for a 24-hour average
concentration, not to exceed 0.47 ug/1 at any time.  For
marine aquatic life, the drafted criterion is 0..019 pg/1
for a 24-hour average concentration not to exceed 0.12 yag/1
at any time  (U.S. EPA, 1979).
                        162-17

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                          TOXAPHENE

                          REFERENCES

American Conference of Governmental Industrial Hygienists.
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and physical agents in the workroom environment with  intended
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Arthur, R.D.., et al,,  1976.  Atmospheric levels of  pesti-
cides in the Mississippi delta.  Bull. Environ. Contam.
Toxicol.  15: 129.

Bailey, T.E., and J.R. Hannum.  1967.  Distribution of pesti-
cides in California.  Jour. San. Eng. Div. Proc. Am.  Soc.
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                                          v
Bidleman, T.F., and C.E.. Olney.  1975.  Long range  transport
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Brooks, G.T.  1974.  Chlorinated insecticides.  CRC Press,
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Chandurkar, P.S.  1977.  Metabolism of toxaphene components
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                          s

Chernoff, N., and B.D. Carver.  1976.  Fetal toxicity of
toxaphene in rats and mice.  Bull. Environ. Contam. Toxicoi.
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Clapp, K.L., et al.  1971..  Effect of toxaphene on  the hepatic
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Conley, B.E.  1952.  Pharmacological properties of  toxaphene,
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Crowder, L.A., and E.F. Dindal. ' 1974.  Fate of chlorine-
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Deichmann, W.B., and M.L.. Keplinger .  1970.  Protection
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DiPasquale, L.C.  1977.  Interaction of toxaphene with ascor-
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Weight State University, 1976.  EPA in-house rep. 1977.
Summarized by K. Diane Courtney, Environ. Toxicoi.  Div.,
Health Eff.  Res. Lab., U.S. Environ. Prot. Agency, in a
Toxaphene ceview dated Nov. 16, 1977.

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Duggan, R.E., and P.E. Corneliussen.  1972.  Dietary  intake
of pesticide chemicals in the United States  (III) .  June
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Fitzhugh, O.G., and A.A. Nelson.  1951.  Comparison of chronic
effects produced ia rats by several chlorinated hydrocarbon
insecticides.  Fed. Proc. 10: 295.

Goodman, L.R., et al.  1978-  Effects of heptachlor and
toxaphene on laboratory-reared embryos and  fry of  the sheeps-
head minnow..  30th Ann. Conf. Southeast Assoc. Game Fish
Coitun.

Hercules Inc.  Undated.  Hercules toxaphene  insecticide.
Bull. T-lOSc.
                                          ».
Hill, R.N.  1977V  Mutagenicity testing of  toxaphene.  Memo
dated Dec. 15, 1977, to Fred Hageman.  Off.  Spec.  Pestic.
Rev. U.S. Environ. Prot. Agency, Washington, D.C.

Johnston, W.R., et al.  1967.  Insecticides  in tile drainage
effluent.  Water Resour. Res. 3: 525.

Kennedy, G.L., Jr., et al.  1973.  Multigeneration reproduc-
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Pharmacbl.  25: 589.

Keplinger, M.L., et al.  1970.  Effects of. combinations
of pesticides on reproduction in mice.  In;  Pestic. Symp.
Collect. Pap.. Int. Am. Conf.. Toxicol. Occup. Med..  6th, .7th.

Khalifa, S., et al.  1976.  Toxaphene degradation  by  iron
(II) protoporphyrin systems.  Jour. Agric. Food Chem. 24:
277.

Kulkarni, A.P., et al.  1975..  Cytochrome P-450 optical
difference spectra of insecticides.  Comparative study.
Jour. Agric. Food. Chem. 23: 177.

Lackey, R.W.  1949a.  Observations on the acute and chronic
toxicity of toxaphene in the dog.  Jour. Ind. Hyg. Toxicol.
31: 155.          '

Lackey, R.W.  1949b.  Observations on the percutaneous absorp-
tion of toxaphene in the rabbit and dog.  Jour. Ind. Hyg.
Toxicol.  31: 155.

Lehman, A.J.  1952.  Oral toxicity of toxaphene.  U.S. Q.
Bull. Assoc. Food Drug Off. 16: 47.

Litton Bionetics, Inc.  Carcinogenic evaluation in mice.
Toxaphene..  Final Report.  LSI Project No.  20602.  Kensington,
MD.  Submitted to Hercules, Inc., Wilmington, Del.  Nov.
1978.

-------
Matsumura, F.  1975.  Toxicology of insecticides.  Plenum
Press.

Mattraw, H.C.  1975.  Occurrence of chlorinated hydrocarbon
insecticides - southern Florida - 1963-1972.  Pestic. Monitor.
Jour. 9: 106.

Mayer, F.L., Jr., et al.  1975.  Toxaphene:  Effects on
reproduction, growth, and mortality of brook trout.  EPA-
600/3-75-013.  U.S."Environ. Prot. Agency.

Mayer, F.L., et al.  1977.  Toxaphene:  Chronic toxicity
to fathead minnows and channel catfish.  EPA-600/3-77-069.
U.S. Environ.. Prot. Agency.

Metcalf, R.L.  1966.  Kirk-Othmer encyclopedia of chemical
technology.  John. Wiley and Sons, Inc., New York.

National Academy of Sciences.  1977.  Drinking water and
health.  A report of the Safe Drinking Water Committee Ad-
visory Center on Toxicology Assembly of Life Sciences, National
Research Council.  Washington, D.C.

National Cancer Institute.  1977.  Guidelines for carcino-
genesis bioassays in small rodents.  Tec. Rep. No. 1. Publ.
No.  017-042-00118-8.  U.S. Government Printing Office,
Washington, D.C.

National Cancer Institute.  1979.  Bioassay of toxaphene
for possible carcinogenicity.  DHEW Publ. No. (NIH) 73-837.

Nicholson, H.P., et al.  1964.  Water pollution by insecti-
cides in an agricultural river basin.  I.  Occurrence of
insecticides in river and treated water.  Limnol. Oceanog.
9: 310.

Nicholson, H.P., et al.  1966.  Water pollution by insecti-
cides:  A six and one-half year study of a watershed.  Proc.
Symp.  Agric. Waste Waters Rep. No. 10 of Water Resour.
Center.  University of California.

Nirarao, D.W.  1977.  Toxaphene:  Its effects on mysids.
Memo to Fred Hagman, U.S. Environ.. Prot. Agency, Washington,
D.C.             .                        ^

Ohsawa, T., et al.  1975.  Metabolic dechlorination of toxa-
phene in rats.  Jour. Agric. Food Chem. 23: 38.

Pollock, G.A.  1978.  The toxicity and metabolism of toxa-
phene.  University of California, Davis.

Reimold, R.J., and C.J. Durant.  1972.  Monitoring toxaphene
contamination in a Georgia estuary.  Natl. Tech. Inf. Serv.
COM 73-1072.  Springfield, Va.

-------
Saleh, M.A., et al.  1977.  Polychlorobornane components
of toxaphene:  Structure-toxicity relations and metabolic  .
reductive dechlorination..  Science 198: 1256.

Schimmel, S.C., et al.  1977.  Uptake and toxicity of toxa-
phene in several estuarine organisms.  Arch. Environ. Contam.
Toxicol. 5: 353.

U.S. EPA.  1976a.  Laboratory examination of drinking water
pesticide analysis.- Unpublished.  Summarized in U.S. EPA
1977.
U.S. EPA.  1976b.  National interim primary drinking water
regulations.  EPA-570/9-76-003.  Off. of Water Supply.

U.S. EPA.  1976c.  Quality criteria for water.  Report No.
EPA-440/9-76-023.                       .  *

U.S. EPA.  1978.  Occupational exposure to toxaphene.  Final
Rep. by the Epidemiol. Stud. Progr. Off. Tox. Subst. Wash-
ington, D.C.'(Draft).

U.S. EPA.  1979.  Toxaphene:  Ambient Water Quality Criteria
(Draft).

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                                        SJ-46-01


                                        No. 164
       1,1,1-Trichloroethane

      (Methyl Chloroform (MC))

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
      WASHINGTON, D.C.   20460
          October 30, 1980
               164-1

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                          DISCLAIMER
     This report represents a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal. - The information contained in this report is drawn chiefly
from secondary sources and available reference documents.
Because of the limitations of such sources, this short profile
may not reflect all available information including all the
adverse health and environmental impacts presented by the
subject chemical.  This document has undergone scrutiny to
ensure its technical accuracy.
                            164-2

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                  1,1,1-TRICHLOROETHANE  (MC)




SUMMARY






Results of  an NCI carcinogenesis bioassay of MC was inconclu-




sive due to experimental problems.  NCI  and a manufacturer




are currently re-evaluating its carcinogenic potential.  In




vitro studies ha-ve indicated that MC is  slightly mutagenic




with or without activation, and can cause mammalian cell




transformation.  Studies of the teratogenic potential of MC




are suggestive; however, more studies are needed to make a




conclusive statement.  Inhalation exposure of healthy adults




to the current PEL for MC (350 ppm) has generally resulted




only in untoward psychophysiologic effects.  Animal studies,




as well as accidental human exposure have shown that MC, at




high inhalation concentrations, produces microscopic pathology




of liver and kidneys which is much less severe than that




produced by carbon tetrachloride or tetrachloroethylene.  MC




is moderately toxic to aquatic life.

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  I.   INTRODUCTION
        %
       The chlorinated ethanes are hydrocarbons In which one or

  more of the hydrogen atoms of ethane are replaced by chlorine

  atoms.  Water solubility and vapor pressure decrease with

  increasing chlorination, while density and melting point

  increase.  At room temperature, 1,1,1-trichloroethane (M.W.

  133.4) is a liquid with a boiling point 74.1'C,  a vapor

  pressure (20°C)  of 100 torr, a melting point of  -33°C,  a

  specific gravity of 1.3492, and a low solubility in water

  (U.S.  EPA,  1980a).

       The chloroethanes are used as solvents, cleaning and

  degreasing agents, and in the chemical synthesis of a number

  of compounds.

       The 1976 production of 1,1,1-trichloroethane was:

  3  X 103  kkg/year (U.S. EPA,1980a).

       The chlorinated ethanes form azeotropes with water (Kirk

  and Othmer,  1963).  All are very  soluble in organic solvents

'  (Lange,  1956).   Microbial. degradation of the chlorinated

  ethanes  has  not  been demonstrated (U.S.  EPA, 1980a).

      The reader  is referred to  the Chlorinated Ethanes  Hazard

  Profile  for  a more general discussion of chl.orinated  ethanes

  (U.S.  EPA,  1980b).

  II.  EXPOSURE

      The chloroethanes present  in raw and finished waters are

  due  primarily to industrial discharges.   Small amounts  of the

  chloroethanes may be formed by  chlorination of drinking water

-------
 or treatment of  sewage.   Air levels  of  chloroethanes  are



 produced  by evaporation  of  these  compounds,  widely  used as



 degreasing  agents  and  in dry cleaning operations  (U.S. EPA,



 1980a).   Occupational  air monitoring studies have indicated



 1,1,1-trichloroethane  levels ranging from  1.5 to  396  ppm



 (U.S.  EPA,  1980a).



     Sources of  human  exposure  to  chloroethanes include water,



 air, ingestion .of  contaminated  foods and fish, and  dermal



 absorption.



     Human  exposure  to MC was estimated from ambient  air



 monitoring  data.   At 8 cities values of 0.02 to 1.86  ug/kg/day



 were calculated.   At one  city,  however, where an  Me manufacturing



 facility  is  located, 12-86 ug/kg/day was calculated (USEPA 1980),



 Drinking  water showed only traces  (0.05-1.0  ppb)  of  MC,



 except near  a MC producing facility  (USEPA,  1980).



     An ana'lysis of  several  foods  indicated  1,1,1-trichloroethane



 was present  at levels of  1-10 ug/kg  (Walter,  et al.,  1976).



 Fish and  shellfish have shown levels of 1,1,1-trichloroethane



 in the nanogram range (Dickson  and Riley, 1976).



     The  U.S. EPA  (1980a) has estimated the  weighted  average



 bioconcentration factor for  1,1,1-trichloroethane to  be 21



 for the edible portions of fish and  shellfish consumed by



Americans.  This estimate is based on the measured steadystate



bioconcentration studies  in bluegills.

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III. ATMOSPHERIC  FATE  AND TRANSPORT

     Because  of Its volatility,  its  transformation  into other
     *
potentially harmful atmospheric  components,  its  tropospheric

chemical reactivity, and its diffusion into  the  stratosphere,

MC is  thought to  pose  a hazard to human health.

     The. volatilization of MC from water can be  reversible

because it is stable in the atmosphere and is transported

back to surface water  via rainfall.  Tropospheric half-lives

of twenty weeks (Pearson and McConnell, 1975) to 8  years

(McConnell and Schiff, 1978) indicate that MC is highly stable

in the troposphere.  Billing et. al., (1976) estimated the de-

composition rate  of MC under simulated atmospheric  conditions

to be less than 5% in  23.5 hours.  It is generally  accepted

that the larger the tropospheric residence time of  a chemical

species, the  greater is the likelihood of its diffusion into

the stratosphere  (U.S. EPA, 1980b).  In a recent study of

the impact of chloro and chlorofluoro compounds on  stratospheric

ozone, based  on atmospheric measurement data, the NAS concluded

(1979) that MC contributes one quarter to one half  as many

chlorine atoms to the stratosphere as do CFC's 11 and 12;  at

the 1976 global emission rate MC is estimated (NAS, 1979)  to

destroy 8 to  15 percent as much ozone as do both CFG 11 and

12.  Thus,  release from improperly disposed solid wastes

containing MC may pose a possible threat to the environment.

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 IV.   PERSISTENCE



     .MC  is  inert  to  reaction with oxygen under normal



 conditions,  except at  high  temperatures.   There are  two



 laboratory  studies of  the hydrolysis  of  MC (Billing,  1975;



 Pearson  and  McConnel',  1975).   These studies  used two  different




 methods'for  calculating  the hydrolytic half  life (U.S. EPA,



 1980c).   The hydrolytic  half-life is  about 5-9  months in



 freshwater,  and about  39 months  in sea water (Pearson and



 McConnel, 1975; U.S. EPA, 1980c).



      At  ambient temperatures  MC  hydrolyzes to  acetic  and



 hydrochloric acids.  Vinylidene  chloride  (a  CAG listed



 carcinogen)  is a minor product,  except at  10°C  and at slightly



 alkaline  pH  when it  is the  major  product  of  hydrolysis (Pearson



 and McConnell 1975,  U.S. EPA,  1980c).



     MC undergoes photochemical  oxidation  (Dilling et. al.,



 1975; Appleby, 1976; U.S. EPA, 1980c), yielding  estimates for



 global average residence time  of  1.4  to 12 years  (U.S. EPA,



JL980c) .   From these  estimated  lifetimes it was  inferred that



 between 10 and 20 percent of  the MC molecules produced will



 reach the stratosphere.



V.   PHARMACOKINETICS



     A.   Absorption



          The chloroethanes are rapidly absorbed  following



oral or inhalation routes of exposure (U.S. EPA,  1980a).



Slow dermal  absorption of 1,1,1-trichloroethane has been



demonstrated in humans (Stewart and Oodd, 1964).

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      B.    Distribution


           Stahl,  et  al.  (1969)  have  noted  the  presence  of
     %

 1,1,1-trichloroethane  in the  liver,  brain,  kidney,  muscle,


 lung  and  blood  in post-mortem tissue samples following  high


 levels of exposures.   MC accumulates in  the liver,  kidney,


 and brain of  the  mouse following  inhalation exposure  (Holmberg,


 et. al.,  1977).


      C.    Metabolism


           The metabolism of chloroethanes  involves  both


 enzymatic dechlorination and  hydroxylation  to  corresponding


 alcohols  (U.S. EPA, 1979a).   Oxidation reactions may  produce


 unsaturated metabolites  which are then transformed  to the


 alcohol and ester (Yllner, 1971a,b,c,d).   Trichloroethanol


 and trichloroacetic acid have been identified  in the  urine


 of rats following inhalation  exposure to 1,1,1-trichloro- .


 ethane (Ikeda and Ohtsuji, 1972).  Metabolism  appears to


 involve the mixed-function oxidase system  (Van Dyke and


 Wineman,  1971).


     D.    Excretion


           The chloroethanes are excreted primarily  in the


 urine and  expired  air  (U.S. EPA, 1980).  Monster and co-


 workers (1979) reported  that  60-80 percent  of 1,1,1-trichloro-


 ethane inhaled by  volunteers  was expired unchanged; two


 urinary metabolites represented 3 percent of the uptake.


Excretion  of the  chloroethanes is generally rapid,  the major-


 ity of compound being  eliminated within 24  hours (U.S. EPA,


 1980a).

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VI.   EFFECTS



      A.    Carcinogenicity



           The NCI  (1977) conducted  a bioassay of MC  using




mice  and rats.  Although a variety  of neoplasms were observed,



no relationship was established bewteen dosage groups, species,



sex type of neoplasm or site of occurence.  The shortened




life  spans of the  test animals due  to the  toxicity of the



chemical made an assessment of carcinogenicity impossible



(NCI, 1977).  The NCI and a manufacturer of MC are currently



retesting  the compound for carcinogenicity.



     Price et. al., (1978), have demonstrated in vitro  trans-



formation  of rat embryo cells by MC.  Injection of these



cells in vivo produced undifferentiated fibrosarcomas at the



site of inoculation in all tested animals.



     B.   Mutagenicity



          Several groups have investigated the mutagenicity



of MC in the Ames assay.   Henschler et. al. (1977) found MC



inactive both with and without addition of microsomes, using



TA-100 strain of S. typhimurium.   Simmon et. al.  (1977)  used



slightly different assay conditions and reported that MC is



slightly mutagenic to this strain.   A dose response was



evident, and metabolic activation did not 'alter mutagenicity.



     C.    Teratogenicity



          Schwetz  et.  al.  reported  (1974,  1979)  on inhalation



studies (at 87.5 ppm)  on  pregnant mice  and rats.   A number of

-------
skeletal  abnormalities were noted, but  these were  of marginal
     *

statistical  significance.  Additional studies  are  needed.


     MC,  when  injected into the air  space of fertilized chicken


eggs at 2, 3 and 6 days of incubation is embryotoxic (LD5Q


of 50-100 mM/egg), and induces a variety of birth  malfor-


mations (Elovaora et. al.: 1979).  Both these  studies suggest


that MC has  teratogenlc potential, and  that further experiments


should be performed to confirm this  potential  toxicity.


     D.   Other Reproductive Effects


          Pertinent information could not be located in the


available literature on other reproductive effects of 1,1,1-


trichloroethane.


    ' E.   Chronic Toxicity (U.S. EPA, 1980b)


          Inhalation exposure of healthy adults to the current


TLV for MC (350 ppm) generally does not result in  significant


untoward physiologic effects.   Studies of human


exposure to 100-500 ppm have shown only subjective symptoms


of light-headedness, syncope,  mild headache and nausea, and


objective symptoms of eye, nose and throat irritation.   No


significant clinical chemistry organ function tests (e.g.


liver function) have been noted.  However,  adverse effects


on the performance of manual tasks have been documented.


     At higher exposures (>10,000 ppm) MC produces anesthesia


and cardiovascular effects which can be lethal.  Animal


studies,  as well as accidental human exposure,  have shown  that

-------
 MC,  at  these  high  concentrations,  produces  a. "chlorinated



 hydrocarbon"  type  of  microscopic pathology  of  liver  and




 kidneys  (fatty  infiltration  and cellular  necrosis) which is



 much less  severe than that produced by  carbon  tetrachloride.




 VI.   AQUATIC.TOXICITY



      A. '   Acute Toxicity



           For freshwater  fish, 96-hour  static  LCso values of



 69,700 ug/1 for the bluegill Lepomis maerochirus and 150,000



 ug/1 for the  fathead  minnow, Pimephales promelas, while a



 single 96-hour  flow-through LCso value  of 52,800 ug/1 was



 obtained for  the fathead  minnow, Pimephales  promelas,



 (Alexander, at. al. 1978).  For marine  organisms, 96-hour



 static LC5Q values ranged from 31,200 ug/1  for the mysid



 shrimp, Mysidopsis bahia, to 70,900 ug/1 for the sheepshead



 minnow, Cyprinodon variegatus, (U.S. EPA, 1978).



     B.    Chronic Toxicity and Plant Effects



          Pertinent information could not be located in the



 available  literature.



     C.   Residues



          A bioconcentration factor of 9 was obtained for the



bluegill (U.S. EPA, 1980a).



VIII. EXISTING GUIDELINES AND STANDARDS



      Neither the human health nor aquatic criteria derived



by U.S. EPA.(1980a),  which are summarized below, have gone

-------
through the process of public review; therefore, there is a



possibility that these criteria will be changed.



     A.   Human



          Based on mammalian toxicology data, the EPA (1979a)



has prepared a draft ambient water quality criterion to



protect human health at the level of 15.7 mg/1 for 1,1,1-



trichloroethane.



     The 8-hour, TWA exposure standard established by OSHA



for 1,1,1-trichloroethane is 350 ppm.



     B.   Aquatic



          The freshwater criterion has been drafted as 5,300



ug/1 as a 24-hour average, not to exceed 12,000 ug/1; while



the criterion to protect marine life has been drafted as a 24-



hour average concentration of 240 ug/1,  not to exceed 540



ug/1.

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                     1,1,1-TRICHLOROETHANE

                           REFERENCES
 1.    Alexander, H.C.,  et  al.   1978.   Toxicity  of  perchloro-
      ethylene, trichloroethylene,  1,1,1-trichloroethane  and
      methylene chloride to  fathead minnow.   Bull. Environ.
      Contain. -Toxicolr  20:344.

 2.    Dickson, A. 6.  and J.  P.  Riley.   197.6.  The  distribution
      of  short-chain  halogenated aliphatic hydrocarbons in some
      marine organisms.  Mar. Pollut.  Bull.   79:167.

 3.    Dilling, W. L., N. B.  Tefertiller and 6.  J.  Kallos.  1975.
      Evaporation rates and  reactivities of methylene chloride,
      chloroform, 1,1,1-trichloroethane, trichloroethylene,
      tetrachloroethylene  and other chlorinated compounds in
      dilute\\aqueous  solutions.  Environ. Sci.  Technol.   9:833-
      838.  -\\\V
           \
 4.    Elovaara, E. K., K.  Hemminki  and H. Vainio.  1978.  Effects
      of methylene chloride, trichloroethane, trichloroethylene,
      tetrachloroethylene  and toluene  on the  development  of
      chick embryos.  Toxicol.  12:111-119.

 5.    Henschler, D., T. Ednee, T. Mendecker and M. Metzler.  1977,
      Carcinogenlclty of trichloroethylene: fact or artfact?
      Arch. Toxicol.  37:233.

 6.    Holmberg, B., et al.   1977.  A study of the distribution
      of methylchloroform  and n-octane in the mouse during and
      after Inhalation.  Scand. Jour. Work Environ. Health.
      3:43.
 t
 7.    Ikeda, M. and H. Ohtsuji.  1972.  Comparative study of
      the excretion of Fujlwara reaction-positive substances
      in urine of humans and rodents given trichloro- or
      tetrachloro-derivatives of ethane and ethylene.  Br.
     Jour. Ind. Med.  29:99.

8.   Kirk, R. and D. Othmer.  1963.  Encyclopedia of chemical
      technology.   2nd ed., John Wiley and .Sons, Inc., New York.

9.   Lange, N. (ed.)  1956.  Handbook of chemistry.   9th ed.,
     Handbook Publishers,  Inc., Sandusky,  Ohio.

10.   McConnell,  J.  C. and H. I. Schiff.  1978.   Methyl chloro-
     form: impact on stratospheric ozone.   Science  199:174-177.

-------
 11.   Monster,  A.  C.,  et  al.   1979.  Kinetics of 1,1,1-tri-.
      chloroethane in  volunteers; influence of exposure con-
     .centration'and work load.  Int. Arch. Occup. Environ.
      Health  42:293.

 12.   National  Academy of Sciences.  1979.  Stratospheric
      ozone depletion  by  halocarbons: chemistry and transport.
      Panel on  Chemistry  and Transport.  Washington, D.C.

 13.   National  Cancer  Institute.  1977.  Bioassay of 1,1,1-
      trichloroethane  for possible carcinogenicity.  Carcinog.
      Tech. Rep.  Ser.  NCI-CG-TR-3.

 14.   Pearson,  C.  R. and  G. McConnell.  1975.  Chlorinated C^
      and C2 hydrocarbons in the marine environment.  Proc.
      Roy. Soc. London Ser. B. 189:305-332.

 15.   Price, P. J., et. al.  1978.  Transforming activities of
      trichloroethylene and proposed industrial alternatives.
      In Vitro  14:290.

 16.   Schwetz,  B.  A.,  et. al.  1974.  Embryo- and fetotoxicity
      of inhaled  carbon tetrachloride, 1,1-dichloroethane, and
     methyl ethyl ketone in rats.  Toxicol. Appl. Pharmacol.
      28:452.

 17.   Schwetz,  B.  A. , B. K. J. Leong, and P. J. Gehrig.  1979.
     The effect  of maternally inhaled trichloroethylene, per-
     chloroethylene,  methyl chloroform and methylene chloride
     on embryonal and fetal developmenting mice and rats.
     Toxicol.  Appl. Pharmacol. 32:84-96.

18.  Simmon, V. F., K. Karhaven, and R. G. Tardiff.  1977.
     Mutagenic activity  of chemicals identified in drinking
     water in: Progress  in Genetic Toxicology, I. D.  Scott,
     B. A. Bridges and F. H. Sobels ed.  Elsevier, New York.
    , pp. 249-258.

19.  Stahl, C. J., et. al.  1969.  Tirchloroethane poisoning:
     observations on  the pathology and toxicology in six fatal
     cases.  Jour. Forensic. Sci.  14:393.

20.  Stewart, R. D. and H. C. Dodd.  1964.  Absorption of
     carbon tetrachloride, trichloroethylene, tetrachloro-
     ethylene, methylene chloride,  and 1,1,1-trichloroethane
     through the human skin.  Am Ind. Hyg. Assoc. Jour.  25:439.

21.  U.S. EPA.  1978.   In-depth studies on health and environ-
     mental impacts of selected water pollutants.  Contract No.
     68-01-4646, D.S.  Environ. Prot. Agency.

-------
22.  U.S. EPA 1980a.  Chlorinated Ethanes: Ambient Water
     Quality Criteria.  (Draft)

23.  *U.S. EPA 1980.  Environmental Criteria and Assessment
     Office.  Chlorinated Ethanes: Hazard Profile.  (Draft)

24.  U.S. EPA 1980c.  Final Report on Risk Assessment of
     1,1,1-tirchloroethane: Contract No. 68-01-0543.  Tech-
     nical Directive- 2.  Batelle Columbus Laboratories,
     Columbus, Ohio 43201; August

25.  Van Dyke, R. A. and C. 6. Wineman.  1971.  Enzymatic
     dehlorination: dechlorination of chloroethanes and
     propanes in vitro.  Biochem. Pharmacol.  20:463.

26.  Walter, P., et. al.  1976.  Chlorinated hydrocarbon
     toxicity (1,1 , 1-trlchloroethane, trichloroethylene, and
     tetrachloroethylene) ; a monograph.  PB-257 185.  Natl.
     Tech. Inf. Serv, Springfield, Va.

                         ••^'
27.  Yllner, S.  1971a.  Metabolism of 1,2 dichloroethane-14C
     in the mouse.  .Acta. Pharmacol.  Toxicol.   30:257.

28.  Yllner, S.  1971b.  Metabolism of 1, 1 , 2-trichloroethane-
             in the mouse.  Acta. Pharmacol. Toxicol.  30:248.
29.  Yllner, S.  1971c.  Metabolism of 1 , 1 , 1, 2-tetrachloro-
     ethane in the mouse.  Acta. Pharmacol.  Toxicol.  29:471.

30.  Yllner, S.  1971d.  Metabolism of 1 , 1 , 2 , 2-tetrachloro-
     ethane in the mouse.  Acta. Pharmacol.  Toxicol.  29:299.

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                                       No.  165
       1,1,2,,-Trlchloroethane

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained 171  the report is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of sucik sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental  impacts  presented by  the
subject chemical.   This  document  has undergone scrutiny to
ensure its technical accuracy.  '

-------
                       SPECIAL NOTATION






U.S.. EPA's Carcinogen Assessment Group (CA6) has evaluated


1,1,2-trichloroethane and has found sufficient evidence  to


indicate that this compound is carcinogenic.
                               .r»f>f -
                             -7) u-1

-------
                             1.1.2-TRICHLOROETHANE
                                    Summary
     Results of  a National Cancer Institute carcinogenesis bioassay  indicate
that  oral administration  of 1,1,2-trichloroethane  produces  an  increase  of
several tumor types in rats and mice.
     Information  is  not available  to indicate  if 1,1,2-trichloroethane  has
any mutagehic effects, teratogenic effects,"or adverse reproductive effects.
     Animal  studies  have  indicated  that  exposure  to 1,1,2-trichloroethane
may produce liver and kidney toxicity.
     Aquatic toxicity  data  for  1,1,2-trichloroethane is  limited,  with only
two  acute studies in  freshwater  fish and invertebrates  available.  Toxic
doses ranged from 18,000 to 40,200 jjg/1.
                                    ' Anl/
                                   7 /IS"

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                             1,1, 2-TRICHLOROETHANE
I.   INTRODUCTION
     This  profile  is  based  on  the  Ambient Water  Quality Criteria.  Document
for Chlorinated Ethanes  (U.S. EPA, 1979a).
     The chloroethanes' are hydrocarbons in which one or more of  the  hydrogen
atoms of ethane  are  replaced by .chlorine  atoms.   Water solubility and  vapor
pressure decrease  with increasing chlorination, while both density and  melt-
ing  points increase-   1,1,2-Trichloroethane  (molecular  weight  133.4)  is  a
Liquid  at  room  temperature  with a  boiling point of  113°C,  a melting  point
of  -37.4°C,  a  specific  gravity of 1.4405,  and  slightly soluble  in  water
(U.S. EPA, 1979a).
     The chloroethanes are used as solvents, cleaning and degreasing  agents,
and in the chemical synthesis of a number of compounds.                      4
     The chlorinated  ethanes  form  azeotropes with water  (Kirk  and  Othmer,
1963) and all are  very soluble  in organic solvents (Lange, 1956).  Microbial
degradation of the chlorinated  ethanes has  not been  demonstrated (U.S. EPA,
1979a) .
     The reader  is referred  to  the  Chlorinated Ethanes  Hazard Profile  for  a
more general discussion of chlorinated ethanes (U.S. EPA,  1979b).
II.  EXPOSURE.
     The chloroethanes are present in raw and finished, waters primarily from
industrial discharges-   Small  amounts  of  chloroethanes  may  be  formed, by
chlorination of drinking  water  or treatment of sewage.  A metropolitan water
monitoring study has shown finished water levels from 0.1  to 8.5 ^ug/1 for
1,1,2-trichloroethane  (U.S.  EPA,  1979a).   Air  levels of  chloroethanes are
produced by evaporation  of  volatile chloroethanes widely  used  as degreasing
agents and in dry-cleaning operations (U.S. EPA, 1979a).
                                   •_ • 7 r*  f_
                                   *1 I * <**

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     Sources  of human exposure to chloroethanes include water,  air,  contami-
nated  foods and fish, and dermal  absorption.   Fish and shellfish  have shown
levels of chloroethanes in the nanogram  range  (Dickson and Riley,  1976).
     Pertinent  information  was  not  found  in  the  available  literature  on
1,1,2-trichloroethane .levels in  food.
     The U.S.  EPA (1979b) has estimated the weighted  bioconcentration factor
for  1,1,2-trichloroethane to be 6.3.  This estimate  was  based on the octa-
nol/water partition coefficient  for 1,1,2-trichloroethane.'
III. PHARMACOKINETICS
     A.  Absorption
         The  chloroethanes  are absorbed  rapidly following oral  or  inhalation
routes of exposure (U.S.  EPA,  1979a). .Dermal absorption of  1,1,2-trichloro-
ethane may be extensive  as indicated by  lethal toxicity in animals following
dermal exposure  (Smyth, et al. 1969).
     B.  Distribution
         Specific  information on  the distribution  of 1,1,2-trichloroethane
has not been  found in the available  literature.   The reader is  referred to  a
more general  treatment of  the  chloroethanes  (U.S. EPA,  1979b) which  indi-
cates widespread distribution of these compounds throughout the  body.
     C.  Metabolism
         The  metabolism of  chloroethanes involves both enzymatic dechlorina-
tion and hydroxylation to corresponding alcohols  (U.S.  EPA,  1979aK   Oxida-
tion reactions  may produce  unsaturated metabolites which are then  transform-
ed to the alcohol  and ester (Yllner,  1971).  Trichloroethanol and  trichloro-
acetic acid have been identified  in  the urine  of rats following  inhalation
exposure to  1,1,2-trichloroetnanol  (Ikeda and  Ohtsuji,  1972).   Metabolism
appears to  involve the  activity of  the mixed  function oxidase system  (Van
Dyke and Wineman, 1971).

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     0.  Excretion
         The  chloroethanes are  excreted primarily  in  the urine  and in  ex-
pired  air  (U.S. EPA,  1979a)  with excretion  being generally rapid.   Experi-
ments  conducted by  Yllner (1971) indicate that following  intraperatoneal  in-
jection of  1,1,2-trichloroethane into mice, more  than  90 percent of  the  ad-
ministered  dose is excreted  in 24 hours,  with more than  half found in  the
urine.  Ten to  twenty  percent of injected compound is found in  expired air.
IV.  EFFECTS
     A.  Carcinogenicity
         Results of an NCI carcinogenesis bioassay  for 1,1,2-trichloroethane
show  that  oral  administration  of compound  produced an  increase  of  several
tumor  types  (NCI,  1978).   Rats, showed adrenal carcinomas,  kidney carcinomas,
and varied hemangiosarcomas,  while mice showed an increase in  hepatocellular
carcinomas.
     B.  Mutagenicity, Teratogenicity and Other Reproductive Effects
         Available  information   on this compound  is very  limited  in these
areas.  A search of the literature did not reveal  any pertinent data.
     C.  Chronic Toxicity
         Animal  studies  have indicated  that  exposure  to 1,1,2-trichloroeth-
ane may produce liver  and kidney toxicity (U.S. EPA, 1979a).
V.   AQUATIC TOXICITY
     A.  Acute Toxicity
         The only aquatic toxicity data for 1,1,2-trichloroethane are single
static bioassays  on the  bluegill. (Lepomis  macrochirus)  and Daphnia magna.
The acute  96-hour  LC5Q  value  for the bluegill was 40,200 ug/1,  while  the
48-hour  LC-Q value for  Daphnia magna  was  18,000 ug/1  (U.S. EPA,  *1979).
Marine studies are presently not available.

                                   ^/<7o*y_
                                  ") f U >
                                      If

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     B.  Chronic Toxicity, Plant Effects and Residues
         Available  information on this compound  is very  limited in  these
areas.  A search of the  literature did not reveal any pertinent data.
VI.  EXISTING GUIDELINES AND STANDARDS
     Neither  the  human  health  nor the  aquatic  criteria derived by U.S.  EPA
(1979)., which are  summarized below, have  gone through the process of public
review;  therefore,  there  is   a   possibility that  these  criteria  will  be
changed.
     A.  Human
         Based  on the  NCI  carcinogenesis  data,  and  using a linear,  non-
threshold model, the  U.S.  EPA (1979a)  has estimated the level  of  1,1,2-tri-
chloroethane  in ambient  water that will result in an additional cancer risk
of ICT5 to be 2.7 jjg/1.                                                     i
         The 8-hr, TWA exposure standard for 1,1,2-trichloroethane  is  10 ppm.
     B.  Aquatic
         The  draft criterion  for  protection of  freshwater  aquatic  life  is
310 pg/1 as a 24-hour average;  the concentration  should not exceed 710 ug/1
at any time  (U.S.  EPAt  1979a).   NO  criterion for  protection of saltwater
aquatic life has been found.

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                             1,1,2-TRICHLOROETHANE

                                  REFERENCES
Oickson, A..G., and  J.P.  Riley.   1976.   The distribution of short-chain  halo-
genated  aliphatic  hydrocarbons  in some  marine  organisms.    Mar.   Pollute
Bull.  79: 167.

Ikeda,  M.,  and  H.  Ohtsuji.   1972.,  Comparative  study of the excretion  of
Fujiwara reaction-positive  substances  in  urine  of humans and rodents  given
trichloro- or  tetrachloro-derivatives  of  ethane and  ethylene.   Br.  Jour.
Ind.. Med.  29: 99.

Kirk,  R.  and Qthmer.  0.  1963.   Encyclopedia of  Chemical  Technology.  2nd
ed- John Wiley and Sons, Inc. New York.

Lange,  N.   (ed.)   1956.    handbook   of  Chemistry.    9th   ed.   Handbook
Publishers, Inc.  Sandusky, Ohio.

National  Cancer  Institute.   1978.   Bioassay  of  1,1,2-trichloroethane for
possible carcinogenicity.  Natl.  Inst.  Health,  Natl.  Cancer Inst.. OHEW  Publ.
NO. (NIH) 78-1324.^  Pub. Health Serv. U.S. Oep. Health Edu.. Welfare.

Smyth,  H.F.,  Jr.,  et  al.   1969.   Range-finding  toxicity data:   list VII.
Am.. Ind. Hyg. Assoc. Jour.  30: 470.

U.S.  EPA.    1979a.   Chlorinated  Ethanes:  Ambient  Water Quality Criteria.
(Draft).

U.S.   EPA.    1979b.    Environmental    Criteria   and   Assessment    Office.
Chlorinated Ethanes:  Hazard Profile (Draft).

Van  Dyke,   R.A.,  and  C.G.   Wineman.    1971.    Enzymatic  dechlorination:
Oechlorination of chloroethanes  and  propanes  in vitro  Biochem.  Pharmacol.
20: 463.

Yllner,  S.   1971.    Metabolism  of   l,l,2-trichloroethane-l,2-14c   in  the
mouse.  Acta. Pharmacol. Toxicol.  30:  248.

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                                      No. 166
         Trichloroethylene

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, O.C.  20460

           APRIL 30, 1980

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                        SPECIAL NOTATION




 U.S. EPA1s Carcinogen Assessment Group (GAG)  has evaluated

 trichloroethylene and has found sufficient evidence to

 indicate that this compound is carcinogenic.
                          DISCLAIMER
     This report  represents  a survey of the  potential  health
and environmental hazards from exposure  to  the subject  chemi-
cal.  The information contained in the report is  drawn chiefly
from secondary  sources  and  available  reference documents.
Because of the limitations of such sources,  this  short profile
may not reflect all available  information  including all  the
adverse health  and  environmental impacts  presented  by  the
subject chemical.   This document  has undergone  scrutiny  to
ensure its technical accuracy.

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                               TRICHLOROETHYLBC
                                    SUMMARY

     Trichloroethylene  is  a  colorless  liquid used  mainly  as  a  degreasing
solvent.   Both  acute and  chronic  exposure  to  high  levels  of  trichloro-
ethylene  produce central  nervous system  depression and  other  neurological
effects.  Trichloroethylene also causes some  kidney  and liver damage.  Tri-
chloroethylene has not been shown to be a teratogen, and the data  suggesting
mutagenicity and carcinogenicity are weak.  The  studies of mutagenicity and
carcinogenicity  have been  complicated  by the presence  of contaminants with
known carcinogenic  and. mutagenic activity.   However, the  cancer  assessment
group has determined that Trichloroethylene is carcinogen<|tic.
     Only a  few  studies  have been reported  on trichloroethylene toxicity'to
                                                                           i
aquatic species.  Fathead minnows,  when  exposed  in  flow  through and static
tests,  had  96  hour  LC5Q values  of 40,700  and  66,800  ug/1,. respectively.
The 96  hour LC5Q for the bluegill was  44,700 ug/1 in static tests.  The  48
hour  LC5Q  for  the  freshwater  invertebrate,  Daohnia  maqna,  was  35,200
ug/1.  In the  only  reported chronic test,  no adverse effects  were observed
in Daohnia  maqna exposed to  10,000  yg/1.  Photosynthesis  was  reduced by  50
percent in  the alga, Phaedactylan tricomutum, at a  concentration  of  8,000
ug/1.  Trichloroethylene  was  bioconcentrated  17-fold by the  bluegill after
14. days exposure.  The half life of  this compound in  tissues  was less than 1
day.

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                               TRICHLOROETHYLENE

I.   INTRODUCTION
     This  profile is  based on the  Ambient Water  Quality Criteria  Document
for  Trichloroethylene  (U.S. EPA, 1979).
     Trichloroethylene   ((^HCl-,   1,1,2-trichloroethylene,   TCE,  molecular
weight  131.4)  is a  clear,  colorless liquid.   Trichloroethylene has a  water
solubility  of  1,000 ug/ml; a vapor  pressure of 77 mm Hg and a  melting  point
of  83°C (Patty,  1963).  Trichloroethylene  is mainly  used as  a  degreasing
solvent, and is used  to  lesser extents  as  a household  and industrial  dry-
cleaning  solvent,  an  extractive  solvent   in  foods,  and  as  an inhalable
anesthetic during certain short-term sorgical procedures  (Huff,  1971).
     Current  Production:   Annual  production  of  trichloroethylene  in  the
United  States  approximates 234,000 metric tons  (U.S.  EPA, 1979).  The  vola-
tilization  of  trichloroethylene   during production and  use  is  the  major
source  of  environmental levels of this compound.   Trichloroethylene is  not
expected to  persist in the  environment because  of its rapid photooxidation
in air, its  low water  solubility,  and  its volatility (Pearson and  McConnell,
1975; Oillings, et al. 1976; Patty, 1963).
II.  EXPOSURE
     A.  water
         The National  Organics Monitoring Survey observed trichloroethylene
in 28  of  113  drinking waters  at a mean concentration  of  21 ug/1  in  May
through July, 1976 (U.S. EPA, 1979).   Trichloroethylene may be formed during
the chlorination of water  (National Academy  of Science, 1977;  Bellar,.et  al.
1974).

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     8.  food
         There  is  Little information  concerning the occurence of  trichloro-
ethylene  in U.S.  foodstuffs.   In  England,  trichloroethylene has  been  ob-
served at  concentrations up to 10  ug/kg in meats,  up  to 5 ug/kg in  fruits,
vegatables,  and beverages  (McConnell,  et  al., 1975);  packets of  tea  were
found to  contain 60 ugAg  (Fishbein,  1976)-.   Little trichloroethylene  would
be expected  in  other foodstuffs, except in- the case  where it is  used as a
solvent  for  food  extractions.   The  U.S.  EPA  (1979)   has  estimated  the
weighted bioconcentration  factor  of trichloroethylene to be 39.   This  esti-
mate is based on measured steady-state bioconcentration  studies in. bluegills
and estimates of fish and shellfish consumption.
     C.  Inhalation
         The only  significant  exposure to trichloroethylene in air occurs to
a relatively small, industrially exposed population  (Fishbein, 1976).
III. PHARMACOKINETICS
                            »
     A.  Absorption
         Trichloroethylene  is  readily  absorbed by  all  routes of  exposure.
In humans exposed  to the compound by  inhalation, steady state conditions are
approached within  two  hours.   Absorption of  trichloroethylene following in-
gestion has not  been studied in humans.  In rats,  at least  80 percent of an
orally administered dose is systemically absorbed (U.S. EPA, 1979).

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     8.  Distribution
         In  humans,  trichloroethylene  is  distributed  mainly  to  body  fat
 (McConnell,  et  al.  1975).   Laham  (1970)  demonstrated  transplacental  dif-
 fusion of trichloroethylene  in  humans.
     C.  Metabolism
         Qualitatively  the  metabolism  . of  trichloroethylene  appears to  be
 similar  across  species  (Kimmerle and Eben, 1973).  The principal products of
 trichloroethylene  metabolism measured  in urine  are,  trichloroethanol,  tri-
 chloroacetic  acid,  and  conjugated derivatives  (glucuronides)  of  trichloro-
 ethanol.  A  reactive epoxide, trichloroethylene  oxide,  has been shown to be
 formed during.the metabolism of trichloroethylene; it  can alkylate  nucleic
 acids and proteins  (Van Duureen and  Banerjee,  1976;  Bolt  and Filser, 1977).
 Patterns of  metabolism  of trichloroethylene in  humans  differ  between  male
 and  female  (Nomiyama  and Nomiyama,  1971),  and  with age  (U.S.  EPA,  1979).
 Increased microsomal enzyme activity enhances  the conversion of  trichloro-
 ethylene to  trichloroacetaldehyde (U.S. EPA, 1979).  Ethanol interferes  with
 the metabolism  of  trichloroethylene, causing ethanol  intolerance in  exposed
 workers (U.S. EPA, 1979).
     D.  Excretion
         Trichloroethylene and  its metabolites are excreted in exhaled  air,
urine, sweat,  feces, and  saliva  (Kimmerla and  Eben  1973;  U.S.  EPA, 1979).
Trichloroethylene is lost  from  the body with a half-life of about 1.5 hours
                               »
 (Stewart, et  al.  1962);  however,  its  metabolites  have  longer  half-lives
ranging from 12 to 73 hours (Ikeda and Imamura,  1973f Ertle, et al. 1972).

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IV.  EFFECTS
     A.  Carcinogenicity
         The  National  Cancer Institute  (NCI,  1976)  observed. an  increased
incidence of  hepatocellular carcinoma in mice  (strain  B6C3-F1)  treated with
trichloroethylene.   Similar  experiments in  Osbome-Mendel  rats  failed  to
increase the  incidence of tumors in  this  species.  It has been pointed out
that trichloroethylene  used in the  NCI bioassay  (1976)  contained traces of
iranofunctional  alkylating agents, epichlorohydrin and epoxibutane,  as sta-
bilizers, and they might  account  for the observed carcinogenicity (U.S. EPA,
1979).   No  systematic  study  of  humans  exposed  to trichloroethylene have
revealed a correlation with cancer (Axelson, at al. 1978).
     B.  Mutagenicity
         Trichloroethylene  has been  reported  to  be mutagenic,  in  the pee-
sence- of mammalian  liver  enzymes, to  a number of bacterial  strains.  These
include  E..  coli  K12,  and  S..  typhimurium strain TA  100  (U.S.  EPA,  1979:
Simmon,  et  al.  1977),  in  addition  to ,the  yeast  Saccharomyces  cerevisiae
(Shahin and VonBarstel, 1977).  However, there is some doubt as to the muta-
genicity of trichloroetnylene  due to  epichlorohydrin and  epoxibutane contam-
ination.  Henscher,   et al.  (1977)   observed  that  these contaminants were
potent mutagens in §_. typhimurium strain TA100.   Pure  trichloroethylene was
weakly mutagenic.
     C.  Teratogenicity
         Exposure  of mice  and  rats  to  1600 .mg/m  trichloroethylene  for
seven hours a  day on days  6  through 15 of  gestation did not produce  tera-
togenic effects (Schwetz,  et al. 1975).

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       D.  Other Reproductive Effects
           Pertinent data could not be located in the available literature.
       E.  Chronic Toxicity
           Disturbances of  the nervous system,  which continue for  at least a
  year after  final exposure,  were observed  following industrial  exposure to
  trichloroethylene  (Nomiyama  and   Nomiyama,   1977;  Bardodej  and.  Vyskoch,
  1956).   Symptoms included headaches, insomnia,  tremors, severe neuroasthemic
  syndromes   coupled   with   anxiety  states,   and   bradycardia.     Prolonged
  occupational exposures  to trichloroethylene have been also  associated with
  impairment  of  the peripheral  nervous system.   This can  include persistent
  neuritis (Bardodej and  Vyskoch,  1956),  temporary loss  of  tactile sense, and
  paralysis of  the fingers (McBirney, 1954).   Rare  cases  of hepatic  damage
  have been   observed  following  repeated   abuse of  trichloroethylene  (Huff,
  1971).
       F.  Other Relevant Information
          ' Long-term toxicity  of  trichloroethylene appears  to depend largely
'  on  its metabolic  products  (U.S.   EPA,  1979).   Chemicals  that   enhance  or
  depress the mixed function oxidase  system will have a synergistic or antago-
  nistic  effect,  respectively,  on the toxicity of trichloroethylene.
           Trichloroethylene has   been  shown  to  induce transformation   in  a
  highly  sensitive in vitro Fischer  rat  embryo cell  system  (F1706)  (U.S. EPA,
  1979).   Following  exposure  of  cells  to 1  M  trichloroethylenet the  cells
  formed  progressively growing foci  made up of  cells lacking  contact inhibi-
  tion, and the cells  gained the  ability to grow  in semi-solid agar.

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V.  AQUATIC TOXICITY
     A.  Acute Toxicity
         Alexander,   et   al.  (1978)  exposed   fathead   minnows  (Pimephales
promelas)  'to  trichloroethylene   in  flow-through  and   static  tests.   The
observed  96-hour  LCep  values were  40,700  and  66,800  ug/1,  respectively.
The observed  96-hour LC5Q  for the bluegill  depends macrochirus)  is  44,700
;jg/l in static tests: (U.S.  EPA,  1978).  The 48 hour LC5Q for Daphnia rcaqna
and  is 85,200  ug/1  (U.S.  £PA,   1978).   No  saltwater fish  or  invertebrate
acute toxicity data were found in the  available  literature.
     8~  Chronic Toxicity
         In the only  reported chronic test, no adverse effects were observed
with Daphnia  magria at the  highest test  concentration of  10,000 ;jg/l  (U.S.
EPA, 1978).
     C.  Plant Effects
                                                      14
         There was a  50 percent decrease  noted in   C  uptake by  the  salt-
water  alga,  Phaedectylum tricornutum,   at a  concentration  of  8,000 ,ug/l
(Pearson and McConnell,  1975).
     0.  Residues
         Bioconcentration by  bluegills  was studied  (U.S.  EPA,  1978)  using
radiolabeled  trichloroethylene.   After 14  days  the  bioconcentration  factor
was 17.  The half-life of this compound in  tissues was less than one day.
VI.  EXISTING GUIDELINES AND STANDARDS
     A.  Human
         The  Food  and  Drug  Administration (1974)v has  limited  the concen-
tration of  trichloroethylene  in  final food products to 10 mg/kg  in instant

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coffee,  25  mg/kg  in ground  coffee and  30 mg/kg  in  spice  extracts.  The
American Conference of Governmental Industrial Hygienists (ACGIH) TLV is 535
mg/m .
         The  Cancer Assessment  Group  (CAG)  has  determined  that,  at the
present time, under existing policy, TCE  is a carcinogen.  The NCI bioassay
(the results from which CAG has made their  determination) is being  repeated.
When the data is available,  it should be reviewed.
     B.  Aquatic
         For  trichloroethylene,  the draft  criterion to  protect freshwater
aquatic life  is  1,500  ug/1 as  a  24-hour average; the  concentration should
not exceed 3,400 jug/1  at  any  time.   Criterion for saltwater species has not
been developed because  sufficient data could not be located in the  available
literature.
                               /it-IP

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                       TRICHLOROETHYLEN E

                          REFERENCES

Alexander,  H.C..,  et al.   1978.  Toxicity of perchloroethy-
lene,  trichloroethylene, 1,1,1-trichloroethane, and methylene
chloride ,to fathead minnows.  Bull. Environ. Contam. Toxicol.
In press.

Axelson,. 0.,  et al.  1978..  A cohort study on trichloroethy-
lene exposure and cancer mortality..  Jour. Occup. Med.  20:.
.19.4;. ........

Bardodej,  Z., and J.  Vyskocil.  1956.  The problem of
trichloroethylene in occupational medicine.  AMA Arch. Ind.
Health  13:  581.

Bellar,  T.A., et al.  1974.   The occurrence of organohalides
in chlorinated drinking  waters.  Jour. Am. Water Works Assoc.
66r  703.

Bolt,  H.M-,  and J.G.  Filser.  1977.  Irreversible binding of -
chlorinated ethylenes to macromolecules.   Environ.  Health
Perspect.   21:  107.

Dill ings,  et al.   1976.   Simulated atmospheric photodecomposi -,
tion rates  of methylene  chloride, 1,1,1-trichloroethane, tri-
chloroethylene,  and other compounds.   Environ. Sci. Technol.
10:  351.

Ertle, T.,  et al.  1972.  Metabolism of trichloroethylene in
man.   I. The  significance of trichloroethanol in long—term
exposure conditions.   Arch.  Toxicol.   29:  171.

F.ishbein,  L,   1976.  Industrial mutagens  and potential muta-
gens.  I.  Halogenated aliphatic derivatives.  Mut.  Res.  32:
267.

Food and Drug Administration.   1974.   Code of Federal Regula-
tions, Title  21,  121.1041.  Trichloroethylene.

Henschler,  D.,  et al. 1977.  Short communication:   Carcino-
genicity of  trichloroethylene: fact or artifact? Arch.  Toxi-
col. 233.

Huff,  J.E.   1971.  New evidence on the old problems of
trichloroethylene.  Ind.  Med.   40:  25.

Kimmerle,  G., and A.  Eben.  1973.  Metabolism, excretion and
toxicology  of trichloroethylene after inhalation.  2.  Exper-
imental  human exposure.   Arch. Toxicol.  30: 127.

Laham, S.   1970.   Studies on placental transfer trichloro-
ethylene.   Ind.. Med.   39: 46.

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McBirney, B.S.  1954.  Trichloroethylene  and  dichloroethylene
poisoning.  AMA Arch. Ind. Hyg.  10:  130.

McConnellf G., et al.  1975.  Chlorinated  hydrocarbons  and
the environment.  Endeavour.  34: 13.
National'Academy of Science.  1977.   Drinking water  and
health.  Safe Drinking Water Comm., Adv.  Center  on Toxicol.,
Assembly of Life Sci., Natl. Res. Council, Washington,  D.C.

National Cancer Institute.  1976.  Carcinogenesis bioassay of
trichloroethylene.  CAS No. 79-01-6, NCI-CG-TR-2.

Nomiyama, K., and H. Nomiyama.  1971.  Metabolism of  tri-
chloroethylene in human sex differences in urinary excretion
of trichloroacetic acid and trichloroethanol.  Int.   Arch.
Arbeitsmed.   28: 37.

Patty, P.A.   1963.  Aliphatic halogenated  hydrocarbons. Ind.
Hyg. Tox.  2: 1307..

Pearson, C.,  and G. McConnell.  1975.  Chlorinated C^ and
C2 hydrocarbons in the marine environment.  Proc. R.  Soc.
London B.  189: 302.

Schwetz, B.A.., et al.  1975.  The effect  of maternally  in-
haled trichloroethylene, perchloroethylene, methyl chloroform
and methylene chloride on embryonal and fetal development  in
mice and rats.  Toxicol. Appl. Pharmacol.  32:   84.

Shahin, M., and R. von Barstal.  1977.  Mutagenic and lethal
effects of benzene hexachloride, dibutyl,  phatalage and
trichloroethylene in Saccharomvces cervisae.   Mut. Res. 48:
173.

Simmon, V.F., et al.  1977.  Mutagenic activity  of chemicals
identified in drinking water.  Paper presented at 2nd Int.
Conf. Environ. Mutagens, Edinburgh, Scotland,  July 1977.

Stewart, R.D., et al.  1962.  Observations on  the concentra-
tions of trichloroethylene in blood and expired  air following
exposure to humans.  Am. Ind. Hyg. Assoc. Jour.  23:  167.

O.S. EPA.  1978.  In-depth studies on health  and environmen-r
tal impacts of selected water pollutants..  Contract No.  68-
01-4646.  U.S. Environ. Prot. Agency.

U.S. EPA.  1979.  Trichloroethylene: Ambient Water Quality
Criteria.  U.S. Environ. Prot. Agency.

Van Duuren, B.L.,  and S. Banerjee.  1976.  Covalent interac-
tion of metabolites of the carcinogen trichloroethylene in
rat hepatic microsomes.  Cancer Res.  36: 2419.

-------
                                                SJ-46-06
                                          No. 167
Trichlorofluoromethane, Dichlorodifluorome thane



           and Trichlorotrifluoroethane




         Health and Environmental Effects
        U.S. ENVIRONMENTAL PROTECTION AGENCY




              WASHINGTON, D.C.  20460




                  October 30, 1980

-------
                           DISCLAIMER
     This report represents a survey of the potential health
and environmental hazards from exposure to the subject chemical,
The information contained in the report is drawn chiefly from
secondary sources and available reference documetns.  Because
of the limitations of such sources, this short profile may not
reflect all available information including all the adverse
health and environmental impacts presented by the subject
chemical.  This document has undergone scrutiny to ensure its
technical accuracy.
                             167-1

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        TRICHLOROFLUOROMETHANE, DICHLORODIFLUOROMETHANE   .




    .          '   AND TRICHLOROTRIFLUOROETHANE






SUMMARY



     Trichlorofluoromethane (F-ll) , dichlorofluoromethane (F-



12) and l;l,2-trichloro-l,2,2-trifluoroethane (F-113) are not



easily degraded in the environment.  After release at the




surface of the earth, F-ll and F-12 and F-113 mix with the



atmosphere and rise slowly into the stratosphere where they are



decomposed by ultraviolet radiation to release chlorine atoms.



The chlorine atoms react with ozone, thereby reducing the total



amount of ozone in the stratosphere and permitting an increased



amount of biologically active ultraviolet radation to reach the



earth's surface*  The accumulation of F-ll, F-12 and F-113 in



the atmosphere also increases the  absorption and emission of



infrared radiation (the "greenhouse" efect).



     F-ll, F-12 and F-113 are absorbed via the lungs, gastrointestinal



tract, and skin, however, most of  that which is absorbed is



eliminated unchanged in expired air.



     F-ll was not found carcinogenic in a long-term mouse study.



The carcinogenic potential of F-113 has not been tested by NCI,



and few specific studies have been documented.   F-ll, F-12



and F-113 were negative in the Ames Salmonella test; F-12 was



positive in a Neurospora crassa test system.
                             167-2

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     At high concentrations in air, these compounds have
     •
been shown to induce cardiovascular and pulmonary effects in

animals.

     In March 1979, fully halogenated chlorofluoroalkanes

(including-F-ll, F-12 and F-113) were banned as propellents in

the United States except for essential uses.  The action was

taken because the chlorofluoroalkanes may deplete the strato-

spheric ozone, leading to various adverse effects.


1.   INTRODUCTION

     This paper is based on an EPA report entitled "Environmental

Hazard Assessment Report:  Major One- and Two-Carbon Saturated

Fluorocarbons" (U.S. EPA, 1976a).

     Trichlorofluoromethane and dichlorofluoromethane are commonly

referred to by their fluorocarbon numbers, which are F-ll and

F-12, respectively.  This convention will be followed in this

paper.   1,1,2-Trichlo-1,2,2-trifluoroethane is dubbed F-113.

     F-ll, a colorless volatile liquid, F-12, a colorless gas,

and F-113, a non-flammable colorless liquid, have the following

physical/chemical properties (U.S. EPA, 1976a, Downning, 1966).

                                    F-ll      F-12     F-113

     Molecular Formula              CClsF'..    CC12F2   CC12F-CC1F2

     Molecular Weight               137       120      187

     Boiling Point (°C)             23.8      -29.8     47.6

     Freezing Point (°C)            -111      -158

     Solubility
     (gm/lOOgm H20, 0°C, 1 atm.)     soluble in water   0.017
                                    and many organic
                                    solvents

                             167-3

-------
     A review of the production range (includes importation)

statistics for trichlorofluoromethane (CAS No. 75-69-4) which

is listed in the initial TSCA Inventory (1979) has shown that

between 100 million and 200 million pounds of this chemical

were produced/imported in 1977.*

     A review of the production range (includes importation)

statistics for dichlorodifluoromethane (CAS No. 75-71-8) which

is listed inthe initial TSCA Inventory (1979) has shown that

between 200 million and 300 million pounds of this chemical

were produced/imported in 1977*

     The major uses of F-ll and F-12 are as aerosol propellants,

refrigerants, and foaming agents (U.S. EPA, 1976a).

II.  EXPOSURE

     A*   Environmental Fate

          Although F-ll and F-12 volatilize quickly from water

and soils, they are considered persistent in the environment

due to their resistance to biodegradation, photodecomposition,

and chemical degradation (U.S. EPA, 1975a).  AFter release at

the surface of the earth, F-ll, F-12 and F-113 (as well as

other chlorofluoromethanea) mix with the atmosphere and rise
*This production range Information does not include any
 production/importation data claimed as confidential by the
 person(s) reporting for the TSCA Inventory, nor does it In-
 clude any information which would compromise Confidential
 Business Information.  The data submitted for the TSCA
 Inventory, including production range information, are sub-
 ject to the limitations contained in the Inventory Reporting
 Regulations (40 CFR 710).
                             167-4

-------
slowly Into the stratosphere where they are decomposed by
     •
ultraviolet radiation to release chlorine atoms.  Chlorine

atoms and a subsequent reaction product, chlorine oxide, react

with ozone and oxygen atoms, thereby reducing the total amount

of ozone in the stratosphere and somewhat shifting the distribution

of ozone toward lower altitudes.  As a consequence, there is an

increase in the amount of biologically active ultraviolet

radiation (below 295 nm) reaching the earth's surface.  In

addition, the temperature distribution in the stratosphere is

somewhat altered*

     The accumulation of chlorofluoroalkanes in the atmosphere,

at all levels, also increases the absorption and emission of

infrared radiation (the "greenhouse" effect).  This retards

heat loss from the earth and thus affects the earth's temperature

and climate.  The amount of change in infrared absorption and

emission is well known, however, the amount and details of the

further effects on the earth's climate are uncertain.  This

effect is inevitably combined with the effects due . to increased

carbon dioxide in the atmosphere and works in the same direction

(HAS, 1976, 1979).

   ' B.   Bioconcentration

          While chlorofluoroethanes are quite lipophilic and

have the potential to bioaccumulate in organisms, their high

volatility appears to preclude significant bioaccumulation

(U.S. EPA, 1975a).
                             167-5

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     C«   Environmental Occurrence



        •  Trichlorofluoromethane has been detected  in  finished




drinking water, effluents from raw sewage and sewage treatment



plants, and in rivers and lakes (U.S. EPA, 1976b).  F-ll  is



formed in small quantities during chlorination and  fluoridation



of drinking water (U.S. EPA, 1975b).



     The major routes by which the fluorocarbons reach the



environment involve their commercial applications.  Because of



their characteristic high vapor pressures and low boiling points,



it is expected that all losses of fluorocarbons would ultimately



the atmosphere (U.S. EPA, 1976a).



III. PHARMACOKINET1CS



     The available data on fluorocarbon absorption  and elimination



indicate that they are absorbed across the alveolar membrane,



gastrointestinal tract, and. skin.  Inhaled fluorocarbons are



taken up readily by the blood.  Fluorocarbons absorbed by any



route are eliminated through expired air (U.S. EPA, 1976a).



     Data from Allen and Hansbury, Ltd. (1971) show that sub-



sequent to a five-minute exposure in ambient air, F-ll and



F-12 are concentrated to the greatest extent in the adrenals,



fat, and the heart of rats.



     Eddy and Griffith (1971) observed metabolism in rats



following oral administrations of 14C-labelled F-12.  About 2Z



of the total dose was exhaled as 002 and about 0.52 was excreted



in urine;  the balance was exhaled unchanged.    Within



thirty hours after administration, the fluorocarbon and its






                             167-6

-------
 metabolites  were  no  longer  present  in  the  body.   Blake  and




 Mergner  (1974) -have  indicated  that  the  apparent resistance of




 F-ll  and F-12  to  biotransformation  may  be  more a  function of




 their rapid  elimination rather  than their  general  stability.




 IV.   HEALTH  EFFECTS'




      A.'    Carcinogenicity




           A  bioassay of F-ll for  possible  carcinogenicity was




 conducted using rats and mice.  Animals were  subjected  to F-ll




 by gavage for  78  weeks.  The results of the bioassay  in rats




 were  not  conclusive because an  inadequate  number  of animals




 survived  to  the end of the  study.   Under the  conditions of the




 bioassay,  F-ll was not carcinogenic in  mice (NCI,  1978).  The




 carcinogenic potential of F-113 has not been  dected by  NCI, and




 few specific studies have been  documented.  Epstein et. al.




 (1967) observed a synergistic  effect when  piperonyl butoxide




 and F-113  were sumultaneously  injected  in  mice, producing an




 increase  in  hepatoms.




     B.    Mutagenicity




           Mutagenicity data on  the  fluorocarbons  are  scant.




 Neither  of the compounds was mutagenic  in  Salmonella  tester




 strains TA1535 or TA1538 with  activation (Uehleke  et  al., 1977).




 Sherman  (1974) found no increase  in mutation  rates over controls




 in a rat  feeding  study of F-12.   Stephens  et  al.,  (1970) reported




a significant mutagenic activity  of F-12 in a Neurospora crassa




 test system.  F-113 has not been  shown  to  be  positive in the




Ames test, and was reported not to  be mutagenic in the  dominant






                             167-7

-------
lethal test in* the mouse.



     C.   Other Toxicity




          Taylor (1974) noted that exposure to 7% oxygen-15Z



tr ichlorof luor ometh-ane (F-ll) caused cardiac arrhythmias in all



rabbits exposed.  F-ll was subsequently shown to exert its




toxicity at air concentrations of 0.5-5% in the monkey and dog,



and from 1-10% in the rat and mouse.  In all these animals it



induced cardiac arrhythmias, sensitized the heart to epinephrine-



induced arrhythmias, and caused tachycardia (increased heart



rate), myocardial depression, and hypertension.  The concentrations



of F-12 that sensitized the dog to epinephrine and that influenced



circulation in the monkey and dog were similar to those reported



for F-ll, however, F-12 differed in its effects on the respiratory



parameters.  It caused early respiratory depression and broncho-



constriction which predominated over its cardiovascular effects



(Aviado, 1975a,b).



     A possible increased sensitivity to the fluorocarbons in



humans with cardiac or respiratory illness may exist, but this



is difficult to determine definitively on the basis of animal



studies.  Azar et al. (1972) noted that human inhalation of 1,000



ppm (4,949 mg/m3) F-12 did not reveal any adverse effect, while



exposure to 10,000 ppm resulted in a 7% reduction in a standardized



psychomotor test score.



V.   AQUATIC EFFECTS



     No data were found.
                             167-8

-------
VI.  EXISTING GUIDELINES
    «
     As of March 17, 1979, fully halogenated chlorofluoroalkanes

were banned as propellents in the United States except  for

essential uses.  Th.fi action was taken because the chlorofluoro-

alkanes. (including F-ll, F-12 and F-113) may deplete the stratos-

pheric ozone, leading to an increase in skin cancer , climatic

changes, and other adverse effects (43 CFR 11301).
                             167-9

-------
                      REFERENCES
Allen and Hansburys, LTD.  1971.  An Investigation of
possible Cardio-Toxic Effects of the Aerosol Propellanta,
Arctons 11 and 12.  Vol. 1, Unpublished Report.  (As
cited in U.S. EPA, 1976a.)

Avlado, D. M.  1975a.  Toxicity of aerosol propellants
on the respiratory and circulatory systems.  IX.  Summary
of the most toxic: trichlorofluoromethane (FC-11).  Toxic-
cology ^, 311-314.  (As cited in the U.S. EPA, 1976a.)

Aviado, D. M.  197Sb.  Toxicity of 'aerosol propellants
on the respiratory system and circulatory systems.  X.
proposed classification.  Toxicology 3, 321-332.  (As
cited in U.S. EPA, 1976a.)           ~~

Azar, A., C. F. Reinhardt, M. E. Maxfield, P. E. Smith,
and L.S. Mullin.  1972.  Experimental human exposure to
fluorocarbon 12 (dichlorofluoromethand).  Amer.. Indust.
Hyg. Assoc. J. 33/4), 207-216.  (As cited in U.S. EPA,
1976a).

Blake, D. A. and G. W. Mergner.  1974.  Inhalation studies
on the biotransformation and elimination of [^*C]-trichloro-
fluoromethane and [14C]-dichlorodifluoromethane in beagles.
Tox. Appl Pharm. 30. 396-407.  (As cited in U.S. EPA,
1976a.)

Downing, R. C. Aliphatic Chlorofluorohydrocarbons.  In;
Kirk-Othmer's Encyclopedia of Chemical Technology, Volume
9,  2nd Edition, 1966.

Eddy, C. W. and F. D. Griffith.  1961.  Metabolism of di-
chlorodif luoromethane-C1* by rats.   Presented at Amer. Indust.
Hyg. Assoc. Conf., Toronto, Canada, May 1971.  (As cited in
U.S. EPA, 1976a.)

National Academy of Sciences, National Research Council.
1976.  Halocarbons: Environmental Effects of Chloromethane
Release.

National Academy of Sciences, National Research Council.
1979.  Stratospheric Ozone Depletion by Halocarbons: Chemistry
and Transport.

National Cancer Institute.  1978.  Bioassay of trichloro-
fluoromethane for possible carcinogenicity.   PB-286-187.
                        167-10

-------
 Sherman,  H.   1974.   Long-term  feeding  studies  in rats  and
 dogs  with dichlorodif luor omethane  (Freon  12 Food Freezant)
unpublished  Report,  Haskell  Laboratory.   (As cited  in  U.S.
 EPA,  1976a.)

 Stephens, S.  et  al.  1970.   Phenotypic and genetic  effects
 in  Neurospora crassa produced  by selected gases and  gases
 mixed with oxygen.   Dev.  Ind.  Microbiol.  12 ,  346.

 Taylor, O.C.   1974.. Univ. of  California, Riverside, Un-
 published Data.   (As cited in  U.S.  EPA, 1976a.)

 Uehleke,  H.  et al .   1977.  Metabolic activiation of  halo-
 alkanes and  tests  in vitro for mutagenicity .   Xenobiotica
U.S.  EPA.   1975a.   Environmental  Hazard Assessment  of One-
and Two-Carbon  Fluorocarbons.   EPA-560/2-75-003.

U.S.  EPA.   1975b.   Identification of Organic Compounds
in Effluents  from  Industrial  Sources.  EPA-560/3-75-002 .

U.S.  EPA.   1976a.   Environmental  Hazard Assessment  Report:
Major  One-  and  Two-Carbon  Saturated f luorocar bons ,  Review
of Data.  EPA-560/8-76-003.

U.S.  EPA.   1976b.   Frequency  of Organic Compounds Identified
in Water.   PB-265-470.

U.S.  EPA.   1979.   Toxic  Substances Control Act Chemical  Sub-
stances Inventory,  Production  Statistics  for Chemical on the
Non-Confidential Initial TSCA  Inventory.
                         167-11

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                                        LB:42-2


                                        No. .168
       2,4,6-Trichlorophenol

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

          OCTOBER 30, 1980
               168-1

-------
                            DISCLAIMER
     This report represents a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary sources and available reference documents.  Be-
cause of the limitations of such sources, this short profile
may not reflect all available information including all the
adverse health and environmental impacts presented by th«
subject chemical.  This document has undergone scrutiny to
ensure its technical accuracy.
                              168-2

-------
                      2,4,6-TRICHLOROPHENOL

                             Summary
         «
     In  a 1979 study N.C.I.-concluded that 2,4,6-trichlorophenol

is carcinogenic in rats and mice.  EPA's carcinogen Assessment

Group has determined that there is substantial evidence that

2,4,6-trichlorophenol is carcinogenic in man.

     2,4,6-Trichlorophenol is a convulsant and an uncoupler of

oxidative phosphorylation.

     2,4,6-Trichlorophenol is acutely toxic to freshwater fish

with LCso values ranging from 320 to 9,040 ug/1.  No chronic or

marine studies were available.  Tainting of rainbow trout flesh

has been noted at concentrations in water greater than 52 ug/1.
                              168-3

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                       2,4,6-TRICHLOROPHENOL

 I.   INTRODUCTION

     TJiis profile  is based on  the Ambient Water Quality Criteria

 Document for Chlorinated Phenols  (U.S. EPA, 1980).

     2,4,6-Trichlorophenol (2,4,6-TCP) is a colorless, crystalline

 solid with the empirical formula CgHsClsO and a molecular weight

 of 197.5 (Weast, 1978).  It has the following physical and chemical

 properties (Weast, 1978):


          Melting  Point:      69.5°C
          Boiling  Point:       246°C
          Vapor Pressure:      1 mm Hg at 76 °C
          Solubility:          slightly soluble in water; soluble in
                               alcohol and ether

     Trichlorophenols  are used as antiseptics and disinfectants,

 as well as for intermediates in the synthesis of other chemical

 products (U.S. EPA 1980).

     It is generally accepted  that chlorinated phenols will

 undergo photolysis in  aqueous  solutions as a result of ultraviolet

 irradiation and that photodegradation leads to the substitution

 of hydroxyl groups in  place of the chlorine atoms and subsequent
 t
 polymerization (U.S. EPA, 1980).  For additional information

 regarding the chlorinated phenols, the reader is referred to the

Hazrd Profile on Chlorinated Phenols (U.S. EPA, 1980).  '

 II.  EXPOSURE

     Unspecified isomers of trichlorophenols have been detected

 in surface waters in Holland at concentrations of 0.003 to 0.1

 ug/1 (Piet and DeGrunt, 1975).  2,4,6-Trichlorophenol can be

 formed from the chlorination of phenol in water (Burttschell, et

al. 1959).   Exposure to other chemicals such as 1,3,5-trichlorobenzene,

                              168-4

-------
 lindane, the alpha- and delta-isomers of 1,2,3,4,5,6-hexachloro-



 cyclohexane, and hexachlorobenzene could result in exposure  to



 2,4,6-trichlorophenol via metabolic degradation of the parent



 compound.



      The U.S. EPA (1980) has estimated the bioconcentration  factor



 2,4,6-trichlorophenol to be 110 for the edible portion of aquatic



 organisms.   This estimate is based on the octanol/water partition



 coefficient for this chemical.



      Trichlorophenols are found in flue gas condensates from



 municipal  incinerators (Olie, et al.  1977).



      A.   Absorption, Distribution and Metabolism



           Information regarding the absorption,  distribution and



 metabolism of 2,4,6-trichlorophenol could not be located in  the



 available  literature.



      B.   Excretion



           In rats,  82 percent of an administered dose  (1 ppm in



.the diet for 3  days)  of 2,4,6-trichlorophenol was eliminated in



 the urine  and 22 percent in the feces.   Radiolabelled  trichlorophenol



 was not detected in liver,  lung,  or fat obtained five  days after



 the last dose (Korte, et al.  1978).



 IV.   EFFECTS



      A.   Carcinogenicity



          Early studies on  the tumor-promoting or-initiating



 capacities  of 2,4,6-trichlorophenol were negative or inconclusive



 (U.S.  EPA,  1980).   Based on the results of its recent  study,



 however, the NCI concluded  that this compound is  carcinogenic  in



 male F344 rats  (inducing lymphomas and  leukemias),  and in both





                               168-5

-------
  sexes of BgCsFi mice,  inducing hepatocellular  carcinomas and
  adenomas.  (National  Cancer  Institute,  1979).
       B.  •  Mutagenicity
            Ames tests using  Salmonella, with and without mammalian
  microsomal activation, were negative for 2,4,6-trichlorophenol
  (Rasanen,  et al.'1977).   2,4,6-Trichlorophenol increased the rate
  of mutations,  but not  the rate of intragenic"recombination in a
  strain of  Saccharomyces cerevisiae  (Pahrig, et al. 1978).  In
  addition,  two  of the 340  offspring  from female mice injected with
  50 mg/Xg of 2,4,6-trichlorophenol during gestation were reported
\\..
 •to have  changes in hair coat color  (spots) of  genetic significance.
  At 100 mg/Xg,  1 out  of 175  offspring exhibited this response
  (U.S.  EPA,  1980).
       C.    Teratogenicity, Other Reproductive Effects and Chronic Toxicity
            Information  regarding teratogenicity, other reproductive
  effects  and chronic  toxicity of 2,4,6-trichlorophenol could not
  be located  in  the available literature.
       D.    Other Relevant  Information
            2,4,6-Trichlorophenol is  a convulsant (Farquharson, et
  al. 1958)  and  an uncoupler  of oxidative phosphorylation (Weiribach
  and Garbus,  1965; Mitsuda,  et al. 1963).
            2,4,6-Trichlorophenol affects glucose metabolizing
  enzymes  at low concentrations (U.S. EPA, 1980).  At relatively
 high  concentrations  it'affects the microsoraal  oxidizing system
  in vitro, which may have  implication with respect to the liver's
 detoxification or cancer  inducing abilities (U.S. EPA, 1980).
                               168-6

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V.   AQUATIC TOXICITY  (U.S. EPA, 1980)
     A.   Acute Toxicity
          Three assays have been conducted with 2,4-trichlorophenol
to determine its acute toxicity to freshwater fish.  A 96-hour
static LCso value of 600 ug/1 has been obtained for the fathead
minnow (Pimephal.es promelas).  In a flow-through assay, a 96-hour
LCso value 9,040 ug/1 was obtained for juvenile fathead minnows.
The bluegill (Lepomis macrochirus) has been shown to be the most
sensitive species studied, with a 96-hour static LCso of 320
ug/1.  Only one acute study has been performed on a freshwater
invertebrate species.  The result of a 48-hour static assay
produced an LCso value of 6,040 ug/1 for Daphnia magna.  There
were no acute studies for any species of marine life.
     B.   Chronic Toxicity
          2,4,6-Trichlorophenol is moderabely toxic to the fathead
minnow (720 ug/1) (U.S. EPA,  1980).
     C.   Plant Effects
          Complete destruction of chlorophyll in the algae,
Chlorella pyrenoidosa,  has been reported at concentrations of 10
ug/1.  A chlorosis LCso value of 5,923 ug/1 was obtained for the
duckweed, Lemna minor.   Studies of the effects of 2,4,6-
trichlorophenol on marine plants have not been reported.
     D.   Residues
          No actual bioconcentration factors have been determined,
but, based upon the octanol/water partition coefficient of 4,898,
a bioconcentration factor of 380 has been estimated for those
aquatic organisms having an eight percent lipid content.

                              168-7

-------
The weighted  average  bioconcentration  factor for  the  edible

portions  of all organisms  consumed by  Americans is  estimated  to
      *
be 110.

     E.   Miscellaneous

          The tainting of  fish  flesh by  2,4,6-trichlorophenol has

been observed in the  rainbow trout (Salmo gairdneri).  The highest

estimated concentration of  2,4,6-trichlorophenol  that will not

impair the flavor of  trout  exposed for 48 hours to  the chemical

is 52 ug/1.

vi.  EXISTING;GUIDELINES AND STANDARDS
            ••\W
     A.   Human

          Based on carcinogenicity the U.S.  EPA (1980) has

recommended 12 ug/1 as the  ambient for the 2,4,6-trichlorophenol

water quality, criterion, for the ingestion  of both fish and  water

(10~6 excess  risk).

          No  other existing guidelines or standards were found

for human exposure to 2,4,6-trichlorophenol.

     B.   Aquatic
 •
          The  •     criterion to protect freshwater .organisms 970

ug/1,  is the  chronic exposure value.    Data were insufficient  to

derive a criterion for marine organisms  (U.S. EPA,  1980).
                              168-8

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                      2,4,6-TRICHLOROPHENOL

                            REFERENCES

Fahrig, R., et al.  1978.  Genetic activity of chlorophenols and
chlorophenol impurities.  Pages 325-328.  In;  Pentachlorophenol:
Chemistry, pharmacology and environmental toxicology.  K. Rango
Rao, Plenum Press, New York.

Farquharson, M.E.., et al. 1958.  The biological action of
chlorophenols.  Br. Jour. Pharmacol.  13:20.

Kohli, J., et al. 1976.  The metabolism of higher chlorinated
benzenes. Can. Jour. Biochem.  54:203.  .

Korte, P., et al.  1978.  Ecotoxicologic profile analysis, a
concept for establishing ecotoxicologic priority list for chemicals
Chemosphere  7:79.

National Cancer Institute 1979.  Broassay of 2,4,6-trichlorophenol
for possible carcinogenicity. NCI-CG-TR-155.; PB223-159.

Piet, G.J. and F. DeGrunt. 1975.  Organic chloro compounds in
surface and drinking water of the Netherlands.  Pages 81.-92 In:
Problems raised by the contamination of man and his environment.
Comm. Eur. Communities, Luxembourg.

Rasanen, L., et al.  1977.  The mutagenicity of MCPA and its soil
metabolites, chlorinated phenols, catechols and some widely used
slimicides in Finland.  Bull. Environ. Contain. Toxicol.  18:565.

U.S. EPA.  1978.   Ambient Water Quality Criteria for Chlorinated
Phenols:  EPA 440/5-80-032.

Weast R.C. (ed.)   1978.  Handbook of Chemistry and Physics. 59th
ed. CRC Press, Cleveland, Ohio.
                              168-9

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                                    No. 169
       1,2,3-Trlchloropropane

  Health and Environmental  Effects
U.S.  ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.   20460

          APRIL 30,  1980
         J61-I

-------
                          DISCLAIMER
     This report represents  a  survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the report \is drawn chiefly
from secondary  sources  and  available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect all available  information  including all the
adverse health  and environmental  impacts  presented by the
subject chemical.   This document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                            1,2,3-TRICHLOROPROPANE
                                   Summary

     Pertinent data are not available on the  possible  carcinogenicity, muta-
genicity,  teratogenicity,  or  chronic  toxicity  of  1,2,3-trichloropropane.
Acute toxicity  studies with animals  suggest  harmful  effects  to the liver.
1,2,3-Trichloropropane is reported to  be irritating to  the eyes and mucous
membranes of humans.
     Pertinent data on the toxicity of trichloropropane to  aquatic organisms
are not available.
                                     -<
                                 H,  7-3

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                            1,2.3-TRICm.OROPROPANE
I.   INTRODUCTION   .
          1,2,3-Trichloropropane  (CAS  registry  96-18-4)  is  a  colorless,
clear liquid made  from the chlorination of propylene..  It has the following
chemical and physical properties  (Windholz,  1976; Hawley, 1971; Verschueren,
1977):
                    Formula:                       C^H5C13
                    Molecular Weight:              147.43
                    Melting Point:                 -14.7°c
                    Boiling Point:                 1$6.85°C
                    Density:                       1.388920
                    Vapor Pressure:                 2.0 torr @ 20°C          \V.
                    Solubility:                    Sparingly soluble in      \v
                                                   water, soluble in alcohol
                                                   and ether.
     1,2,3-Trichloropropane is used as a paint and  varnish remover,  solvent,
and  decreasing  agent  (Hawley,  1971), in  addition to  its use  as a  cross-
linking agent in the elastomer Thiokol ST (Johnson,  1971).
II.  EXPOSURE
     A.   Water
          1,2,3-Trichloropropane  has  been detected  in drinking water  (U.S.
EPA, 1975) and also in 6 of 204 surface water samples  taken in various loca-
tions throughout  the  United  States  (U.S.  EPA,  1977).   No information  con-
cerning concentration was available.
     8.   Food
          Pertinent data were not found in the available literature.

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      C.    Inhalation
           Pertinent data were  not  found  in the  available literature;  how-
 ever,  fugitive emissions from  manufacturing and  production facilities  pro-
 bably  would account for the major portion of 1,2,3-trichloropropane if found
 in air.
     D.    Dermal
           Pertinent data were not found in the available literature.
 III. PHARMACOKIhETICS
     Pertinent data were not found in the available literature.
 IV.  EFFECTS
     A.    Carcinogenic!ty, Mutagenicity, Teratogenicity,
           Reproductive Effects, Chronic Toxicity.
           Pertinent data were not found in the available literature.
     B.    Acute Toxicity
           Exposure  to  trichloropropane  at high  concentrations  is irritating
to the eyes and mucous membranes and causes narcosis.
           McOmle  and Barnes (1949)  exposed 15  mice  to 5000 ppm trichloro-
propane for 20 minutes.  Seven  of the mice  survived  exposure;  however, four
of these mice  died from liver damage 7 to 10 days later.   Seven of ten mice
exposed to 2500  ppm trichloropropane  for  10 minutes  per  day  for 10 days
died.  McOmie  and  Bames (1949)  found  that liquid  trichloropropane applied
to the skin of rabbits produced irritation and erythema,  followed by slough-
ing and cracking.   Repeated application of 2 ml  of trichloropropane caused a
painful reaction,  including  subdermal  bleeding,  and the  death  of one  of
seven rabbits treated.
          Silverman, et al.  (1946) reported  eye  and  throat  irritation  and an
objectional odor  to human volunteers exposed to  100  ppm  trichloropropane for

-------
15 minutes.  McOmie  and  Barnes  (1549)  found that ingestion of  3g of  tri-
chloropropane by humans caused drowsiness, headache,  unsteady gait, and  lum-
bar pain.
V.   AQUATIC TOXICITY
     Pertinent  data were not found in the  available literature~
VI.  EXISTING GUIDELINES AND STANDARDS
     A.   Human
          The American Conference  of Governmental  Industrial Hygienists re-
commends a  threshold limit  value of 50  ppm for  occupational  exposure to
1,2,3-trichloropropane (ACGIH,  1977).
     8.   Aquatic
          No guidelines  on  standards  to  protect aquatic organisms  from
1,2,3-trichloropropane toxicity have been  established because of the lack of
pertinent  data.
                                161-1

-------
                             1,2,3-TRICHLOROPROPANE
                                  References
 American  Conference of Governmental  Industrial Hygienists.  1977.   Documen-
 tation  of the  Threshold  Limit Values  for Substances  in  Workroom  Air,  3rd
 ed.   American  Conference  of Governmental  Industrial Hygienists Cincinnati,
 OH.
 Hawley, G.G.  (ed.)   1971.   The Condensed Chemical  Dictionary.   8th ed.   Van
 Nostrand Reinhold Co., New York.
 Johnson,  R.N.  1971.  Polymers containing  sulfur.   In;   Kirk-Othmer Encyclo-
 pedia of Chemical Technology.  John Wiley and Sons, New York, p. 253.
 McOmie,  W.A.  and  T.R.  Barnes.    1949.    Acute and subacute  toxicity  of
 1,2,3-trichloropropane in mice and rabbits.  Fed. Proc.  8: 319.
 Silv^rman, L.,  et al.   1946.   Further studies on sensory response to certain
 industrial solvent vapors.  Jour. Ind. Hyg. Toxicol.  28: 262.
 U.S.  EPA.   1975.   New Orleans  area  water supply  study, draft  analytical
 report.  U.S. Environ. Prot. Agency.  April update.
 U.S. EPA.  1977.  Monitoring  to detect previously unrecognized pollutants in
 surface waters.  U.S. Environ. Prot. Agency.  NTIS PB273-349.
 Verschueren,   K.  1977.   Handbook  of  Environmental  Data  on Organic  Chem-
 icals.  Van Nostrand Reinhold Co., New York.
Windholz,  M.  (ed.)   1976.   The  Merck Index, 9th  ed.  Merck and  Co.,  Inc.,
Rahway, NJ.
                                   )

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                                      No. 170
  o,6\o-Trlethyl Phosphorothioate

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents a survey of  the  potential  health
and environmental hazards from exposure  to the  subject  chemi-
cal.  The information contained in the report is drawn chiefly
from secondary  sources  and   available  reference  documents.
Because of the limitations of such sources, this short profile
may not reflect  all  available information including all  the
adverse health  and  environmental  impacts presented  by  the
subject chemical.  This  document has undergone scrutiny  to
ensure its technical accuracy.
                        /76-a.

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                        0,0,0-TRIETHYL PHOSPHOROTHIOATE
                                    Summary

     There  is no information  available  on the possible carcinogenic,  muta-
genic,. teratogenic, or-adverse reproductive effects of  0,0,0-triethyl  phos-
phorothioate..   Triethyl phosphate,.-a- possible metabolite of  the  compound,
has shown weak mutagenic activity in Salmonella,  Pseudomonas, and Drosophila.
     Like other organophosphates, 0,0,0-triethyl  phosphorothioate may be ex-
pected to produce cholinesterase inhibition in humans.
     No pertinent data are available on the aquatic effects of the compound.

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                        0,0,0-TRIETHYL PHOSPHOROTHIOATE
 I.    INTRODUCTION
      0,0,0-Triethyl  phosphorothioate  (CAS  registry number  126-68-1),  also
 known as triethyl  thiophosphate,  is a colorless liquid with a  characteristic
 odor.  It has the  following physical and chemical properties (Hawley, 1971):
                    Formula:                 C^oys
                    Molecular Weight:        198
                    Boiling Point:           93.5°C-94°C (10 torr)
                    Density:                  1.074
     0,0,0-Triethyl phosphorothioate  is used as a plasticizer, lubricant ad-
ditive,  antifoam  agent,  hydraulic  fluid,   and  as  a chemical intermediate
 (Hawley, 1971).
 n.  EXPOSURE
     A.   Water and Food
          Pertinent data were not found in the available literature.
     8..   Inhalation
          Pertinent data  were-  not  found in the available  literature; how-
ever, fugitive  emissions  from production and  use would probably  constitute
the major source of contamination (U.S. EPA, 1977).
     D.   Dermal
          Pertinent data were not found in the available literature.
in. PHARMACQKINETICS
     A.   Absorption
          Pertinent data were  not found in  the  available  literature.   Acute
toxicity studies with a number of organophosphate  insecticides  indicate that
these  compounds   are  absorbed   following   oral  or  dermal  administration

-------
 (Gaines, 1960).  March, et  al.  (1955) have reported  rapid absorption of  the
 structurally similar  insecticide demeton from  the  gastrointestinal tract of
 mice following oral administration.
      8.   Distribution
           Pertinent data were not found in the available- literature.
      C.   Metabolism                      .
           Pertinent data,  were not  found in  the available  literature.   The
 thiono isomer of the insecticide  demeton may  be metabolized  via oxidative
 desulfuration by the  liver  at the P=S bond  in  mammals (March, et al.. 1955)
 to  form the thiolo derivative.  Thus, 0,0,0-triethyl  phosphorothioate may  be
 converted to triethylphosphate in vivo (Matsumura, 1975).        >.V\^'
      0-   Excretion
           Pertinent data were not found in the  available  literature.   March,
 et  al.. (1955) have reported that fallowing  oral administration of  demeton,
 the large majority  of compound  was eliminated  as urinary metabolites,  with
.small quantities detected in the  feces.  Elimination was  rapid  following
 oral administration.
 IV..  EFFECTS
      A.   Carcinogenicity
           Pertinent data were  not found in the available literature.
      B.    Mutagenicity
           Pertinent data  were not found  in   the  available literature.  The
 insecticide   oxydemeton  methyl  has   been shown  to   produce   mutations,  in
 Drosoohila, £. coli and Saccharomyces (Fahrig, 1974).'  Triethyl phosphate, a
 possible  metabolite of  0,0,0-triethyl phosphorothioate,  has  produced weak
 mutagenic effects in Salmonella and  Pseudomonas  (Dyer and Hanna,  1973) and
 recessive lethals in Drosophila (Hanna and Dyer, 1975).

-------
      C.    Teratogenicity
           Pertinent data  were  not  found  in  the  available literature.   A
 single intraperitoneal  injection of  demeton (7  to 10  mg/kg)  between  days
 seven and twelve of gestation  has been reported to  produce  mild  teratogenic
 effects in mice  (Budreau and Singh,  1973).
      0.    Other  Reproductive Effects	-	
           Pertinent data were  not  found  in the  available literature.   Em-
 bryotoxic effects  (decreased fetal  weights,  slightly  increased  fetal  mor-
 tality) have been reported following intraperitoneal administration  of deme-
 ton  (7 to 10 mg/kg) to pregnant mice (Budreau and Singh,  1973).
      E.    Chronic Toxicity
           Pertinent data   were  not  found  in  the available   literature.
 0,0,0-triethyl  phosphorothioate,  like other organophosphates,  may  be  ex-
 pected  to produce  symptoms  of  cholinesterase inhibition  in  humans  (MAS,
 1977).
 v.   AQUATIC TOXICITY
     Pertinent data were not found in the available literature.
VI.  EXISTING GUIDELINES AND STANDARDS
     Pertinent   data  were  not   found   in   the  available   literature.


-------
                        0,0,0-TRIETtiYL PHOSPHOROTHIOATE

                                  References


 Budreau,  C.  and R.  Singh..  1975.   Teratogenicity and embryotoxicity of deme-
 ton and fenthion in CT 1 mouse embryos.  Toxicol.. Appl. Pharmacol.  24: 324.

 Dyer,  K. and P.  Hanna-  1973..  Comparative mutagenic activity  and toxicity
 of triethyl phosphate and dichlorvos in bacteria and. Orosophila.   Mut. Res*
 21: 175.            '                         -

 Fahrig,  0.   1974.-  Comparative mutagenicity studies  with  pesticides.   Chem-
 ical Carcinogenesis Assays, IARC Scientific Publication.   10:• 161-

 Harma,  P. and  K. Dyer.   1975.  Mutagenicity of organophosphorous  compounds
 in bacteria and Drosophila..  Mut» Res..  28: 405.

 Gaines, T.   1960.  The acute  toxicity  of pesticides  to rats.   Toxicol. Appl.
 Pharmacol..  2:  88.

 Hawley, G.G.  (ed.)   1971.   The Condensed Chemical. Dictionary.  8th ed.  Van
 Nostrand Reinhold Co..f New- York.

 March,  R.,.  et al.  1955.  Metabolism  of  syston in white mouse and American
 cockroach.  Jour- Econ. Entom.  48r 355.

 Matsumura,  F.   1975.  Toxicology of  Insecticides.   Plenum Press, New  York,.
 223.

.National  Academy of Sciences.  1977.   Drinking Water and Health.  National
 Research Council, Washington,  DC,  p. 615.

 U.S. EPA.   1977.  Industrial  process profiles for environmental user   Plas-
 ticizer industry.  U.S.. Environ. Prot.  Agency, NTIS PB-291-642.
                               170-J

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                                      No. 171
          Trinltrobenzene

  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
       WASHINGTON, D.C.  20460

           APRIL 30, 1980

-------
                          DISCLAIMER
     This report represents a survey of the potential health
and environmental hazards from exposure to the subject chemi-
cal.  The information contained in  the report is drawn chiefly
from secondary  sources  and  available reference  documents..
Because of the limitations of  such sources, this short profile
may not reflect  all available  information  including all the
adverse health  and  environmental  impacts  presented by  the
subject chemical-   This  document  has undergone  scrutiny to
ensure its technical accuracy.

-------
                               TRINITROBENZENE
                                   Summary

     Information on  the  carcinogenicity,  mutagenicity,  teratogenicity,  or
adverse reproductive effects of trinitrobenzene was not found  in  the  avail-
able literature.
     Trinitrobenzene  has been reported to produce liver damage, central ner-
vous system damage,  and  methemoglobin  formation in animals.
     Slight irritant effects have been reported  for marine fish  exposed  to
trinitrobenzene at concentrations of 100 ug/1.
                                  I7/-3

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                                TRINITROBENZENE
t.   INTRODUCTION
     This  profile  is  based  on  the  Investigation  of.  Selected  Potential
Environmental Contaminants:  Nitroaromatics (U.S. EPA, 1976).
     Trinitrobenzene  (1-,3,5-trinitrobenzene,  molecular  weight, 213.1)  is a
crystalline  solid with  the following  physical properties:   melting point,
122.5°C;  specific  gravity,  1.76.   The  compound   is  explosive  upon  rapid
heating.   Trinitrobenzene  is insoluble in  water,  but soluble in alcohol  or
                                                   «u
ether (Windholz,  1976).
     Trinitrobenzene  is  used as an explosive, and as a vulcanizing agent  for
natural rubber  (U.S.  EPA, 1976).
     Hydrolysis  of   trinitrobenzene  under   neutral  pH   conditions   is  not
expected  to  be rapid; as pH increases, hydrolysis  would  be  favored  (Murto,
1966).   Photolytic degradation of trinitrobenzene  has not been demonstrated
in aqueous solutions  (Burlinson, et al.  1973).
     A  bioconcentration factor is not  available  for trinitrobenzene;  how-
ever,  the work of  Neely,   et  al.  (1974)  on  several  nitroaromatics  would
suggest a  low theoretical bioconcentration of the compound.
     Biodegradation of  trinitrobenzene  by  acclimated microorganisms has been
reported by Chambers, et al.  (1963).
II.  EXPOSURE
     Pertinent  information  on levels  of exposure to  trinitrobenzene  from
occupational  contact or from  non-occupational 'sources  of  exposure  (air,
water, food) was  not  found  in the  available literature.
III. PHARMACOKINETICS
                                                                         »-
     Pertinent  information   on  the absorption, distribution,  metabolism,  or
excretion of trinitrobenzene was not  found  in the available literature.   The

-------
reader  is referred  to  a  discussion of  the  pharmacokinetics  of  dinitro-
benzenes, which may show pharmacokinetic similarities (U.S. EPA,  1979).
     Acute oral toxicity  studies  conducted with dogs indicate that trinitro-
benzene is effectively absorbed by this route (Fogleman, et al. 1955).
IV.  EFFECTS
     Pertinent  information on  the carcinogenic, mutagenic,  teratogenic, -or
adverse reproductive  effects  of trinitrobenzene was not  found in the avail-
able literature.
                                                   *.
     A series  of  toxicity studies in rats,  mice,  and guinea pigs have indi-
cated  that  orally  administered   trinitrobenzene   causes   liver  damage  and
central  nervous system damage  (Korolev,   et  al. 1977).   The acute toxicity
study of  Fogleman, et al. (1955) has shown  that trinitrobenzene, like dini-
trobenzenes, induces methemoglobin formation in  vivo.
V.   AQUATIC TOXICITY
     The  only  study  reporting  the effects of trinitrobenzene to  aquatic  life
has been  presented by Hiatt, et  al.  (1957).  Slight  irritant effects i.e.,
excitability,  violent  swimming,  opercular movement increases suggesting  res-
piratory   distress  upon   short   term  exposure   to  marine   fish  Kuhlia
sandvicensis were  observed at  exposure levels  of  100 ug/1,  while moderate
and violent  reactions to  the  chemical were  produced at  exposures  of 1,000
and 10,000 pg/1.   No  effects  were noted on exposures to concentrations of 50
or 10 ug/1.
VI.  EXISTING GUIDELINES
     There is  no  available  8-hour,  TWA exposure limit  for trinitrobenzene.
The compound  has  been  declared a hazardous chemical  by  the  Department of
                                                                        •
Transportation      (Federal      Register.      January      24,      1974).

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                                TRINITROBENZENE

                                  References
Burlinson, N.E.  et al.   1973.   Photochemistry of TNT:  Investigation of  the
•pink water' problem.  U.S. Nat. Tech. Inform. Serv. Ace. No. AD  769-670.

Chambers, C.W.,  et al.  1963.  Degradation  of aromatic compounds by phenol-
adapted bacteria.  Jour.'Water Pollut. Contr.  Fedr.  35: 1517.

Fogleman, R.W.,  et al.   1955.   Toxicity of trinitrobenzene-aniline  complex,
a rodent repellent.  Agric. Food Chem.   3(11): 936.

Hiatt, R.W.,  et  al.  1957.  Relationship  of chemical  structure  to  irratient
response in marine fish.  Nature.  179:  904.
                                                   >.
Korolev,  A.,  et al.   1977.   Experimental data  for  hygienic standardization
of  dinitrotoluene  and  trinitrobenzene  in  reservoir  waters.    Gig.  Sanit.
10: 17.

Murto, J.   1966.  Nucleophilic reactivity.   VII.  Kinetics of the  reactions
of  hydroxide  ion  and water  with picrylic  compounds.   Acta Chem.  Scand.
20: 310.

Neely,  W.B.,  et  al.   1974..   Partition  coefficient  to  measure bioconcen-
tration  potential of  organic  chemicals  in  fish.   Environ.  Sci.  Technol.
8: 1113.  .

U.S.  EPA.   1976.  Investigation of selected potential environmental contam-
inants :  Nitroaromatics.

U.S.  EPA.   1979.   Environmental Criteria  and Assessment  Office.  Dinitro-
benzene:  Hazard Profile. (Draft).

Windholz, M..  (ed.).   1976.   The Merck  Index,  9th ed.  Merck  and'Co.,  Inc.
Rahway, N.J. p. 9392.
           SPO 998-oae

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                                         SJ40-6






                                         No.  172
              Aniline



  Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY




       WASHINGTON, D.C. 20460




          October 30, 1980











               172-1

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                          DISCLAIMER
     This report represents a survey of the potential health
and environment hazards from exposure to the subject chemi-
cal.  The information contained in the report is drawn chiefly
from secondary sources and available reference documents.
Because of the limitations of such sources, this short profile
may not reflect all available information including all the
adverse health and environmental impacts presented by the
subject chemical.  This document has undergone scrutiny to
ensure its technical accuracy*
                            172-2

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                       ANILINE SUMMARY








     Aniline is an aromatic ami tie.  Like many members of this




group it is characterized by an outstanding property: the



ability to form methemoglobin in mammalian organisms.  Aniline



and some of its analogues have been suspected as carcinogens



since the turn of the century.  An increased incidence of



urinary bladder tumors has been noted in workers in the



aniline and aniline dyestuff industries.  Oral feeding studies



conducted by the National Cancer Institute have shown aniline



to be carcinogenic in rats.  Aniline is reported not to be



mutagenic to six strains of S. typhimurium, however, several



aniline analogues and derivatives are mutagenic.
                            172-3

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                           ANILINE
I.   INTRODUCTION

     Aniline (aolnobenzene) is a liquid at room temperature

Its physical properties can be summarized as follows:

(Hawley, 1977; Weast, 1977-78; Strecher, 1968):
     Molecular formula

     Molecular weight

     Melting point

     Boiling point

     Flash point

     Solubility in cold water
     Temperature at which
     vapor pressure equals 1 mtn/Hg

     pKa
 C6H5NH2

 93.13

 -6.2°C

184.4°C

158.0°F (Closed Cup)

 35.0 grams/liter
 (0.38 M).



 34.8eC

  4.63
Of particular interest is aniline's water solubility - i.e.,
                  *                           •
aniline is soluble even in cold water.

     The Stanford Research Institute's Chemical Economics

Handbook cites the nitrobenzene reduction process as the

current method of aniline synthesis (McCaleb, 1976).  It is

estimated that 270,000 kkg of aniline was produced in 1978

(Slimak, et. al., 1980).  Aniline was reported to be used

for the following uses:


                            172-4

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s.ynthesis of isocyanate (50Z of total consumption) 9 of rubber
                                          «
chemicals (27%), dyes and Intermediates (6%), hydroquirxone

(5Z), drugs (37%), and miscellaneous chemicals including herbicides

(9%).

II.  EXPOSURE

     A.   Water

          Aniline levels in surface or drinking water was not

reported in the available literature.  However, overall

emission of aniline to receiving waters as a result of the

production of aniline, consumptive use, carry-over as impurities

in manufactured products, and degradation of manufactured

products was estimated in 1978, to be 9970 kkg (Slimak, et.

al., 1980).

     B.   Food

          Pertinent data on aniline concentrations could not

be located in the available literature.

     C.   Inhalation

          Pertinent data on aniline concentrations could not

be located in the.available literature.  However,  overall

emissions of aniline to air as a result of aniline, isocyanates,

rubber chemicals, dyes and intermediates, hydroqulnone, and

miscellaneous products production is estimated to  be 69.4 kkg

in 1978 (Slimak, et. al., 1980).  Aniline is also  reported

to occur in cigarette smoke (Gosselin, et. al., 1976).
                            172-5

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PHARMACOKINETICS



     A.   Absorption




          Both inhalation and dermal absorption are important



exposure routes in humans (Pietrowski, 1957).  At air concen-



trations of up to 20 mg/m3, absorption is about equal by




both routes, that is, 6 mg/hr; at higher concentrations the



respiratory pathway becomes progressively a more important



factor.  Dermal contact with liquid aniline also results in



rapid systemic absorption: 0.2-0.7 mg/cm^ of skins/hr has



been shown to occur (Pietrowski, 1957).



     B.   Distribution



          Aniline is rapidly absorbed into the blood stream;



its subsequent systemic distribution has not been reported.



Its metabolic transformations (see below) are mainly dependent



on liver enzymes.



     C.   Metabolism



          Aniline is metabolized in the liver by oxidation



and conjugation.  Hepatic microsomal oxidizing enzymes cause



metabolic transformation to N-acetylaminophenol and to o- and



p-aminophenol.  Conjugating enzymes then cause metabolic



transformation to the glucuronide and sulfonate (Casarett and




Doull, 1975).



     D.   Excretion



     The administration of aniline leads to urinary excretion



of glucuronic acid and sulfonic acid conjugates and of its



metabolites, o- and p-aminophenol (Williams, 1959;  Parke,
                            172-6

-------
1960).  In addition, small  amounts of  free aniline, phenyl




sulfamic acid, aniline-  glucuronide,  aminophenyl and




acetylaminophenyl, mercaptonic  acids,  phenylhydroxylamine,




and acetanilide are excreted in varying amounts by different




species tested (Park'e, 1960).   The conjugates of p-amino-




phenol are the most important urinary  metabolites of aniline




(Williams, 1959).  The urinary  excretion of these




metabolites gives an accurate measure  of the absorption of




aniline vapor (Pietrowski,  1972).  It  is probable that the




other aniline metabolites mentioned above also appear in the




urine of people exposed to  aniline (IARC, 1974).




IV.  EFFECTS




     A.   Carcinogenicity




          The carcinogenic  potential of aniline has been of




great interest because, since 1895, an increased incidence of




urinary bladder tumors has  been noted  in workers in the




aniline dye industry (IARC, 1974).  It has subsequently been




shown that other amines which occur in the environment, such




as 2-naphthylamine, 4-aminobiphenyl and benzidine, are probably




more important in the causation of these occupational cancers




(IARC, 1974).  Although most animal studies appear to have



exonerated aniline as a human carcinogen (IARC, 1974), an NCI




study showed a dose-related increase in fibrosarcomas or




sarcomas in the spleen and  in several  organs of the body



cavity.  Although the results were not statistically signifi-




cant,  the rarity of these tumors and their dose-dependency






                           172-7

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led to the conclusion that aniline is carcinogenic in female




Fisher 334 rats*  The male rats showed a statistically



significant increase in the incidence of hemangiosarcomas of



the spleen and a significant increase in the combined incidence



of fibrosarcomas and  sarcomas of the spleen and in multiple




body organs (NCI, 1978).



     B.   Mutagenicity



          In the Ames assay aniline is not mutagenic toward



any of the six standard S. typhimurium strains, when tested



in the presence or absence of microsomes (Geomet, 1980).



However, in the presence of nor-harman (a 2-carboline derivative),



significant mutagenicity has been observed (Nagao, et. al.,



1977;  Sugimura, 1979).



     C.   Teratogenicity



          Pertinent data could not be located in the available



literature.  However, cyanotic effects such as those produced



by aniline can adversely affect the fetus, leading the ITC



to recommend that reproductive effect tests be conducted (44



FR 31871).



     D.   Other Reproductive Effects



          Aniline can cross the placenta to form methemoglobin



in the fetus, affecting its development (Gosselin, 1976).



Courtney, (1979), reported that, after treatment of CD-I



mouse  dams with aniline (150 to 200 mg/kg) during gestation



and lactation,  CPK and LDH isozyme patterns in serum and
                            172-8

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cardiac  tissue  in  offspring were  altered  on  days  1  and  20




postpartua.  The serum CPK pattern  was  markedly altered on




day 1 postpar.tum and  by day 20, an  additional  enzyme  appeared.




     E.   Acute and Chronic Effects




          Aniline  is  moderately lethal  to  the  rat:  the  oral




LD50 is  440 mg/kg.  The LD50 value  via  the dermal route is




1400 mg/kg in the  rat.  For human beings  the oral LDLO  is




reported as 50 mg/kg  (NIOSH, 1979).  The no-effect  level for




humans has been estimated as 0.25 mg/kg (NIOSH, ref.  169).




     Aniline absorption causes anoxia due  to the  formation of




methemoglobin.  Most  of the signs and symptoms of overexposure




to aniline can be  attributed to methemoglobin  formation.




Such symptoms include fatigue, headache, irritability and




dizziness (Proctor and Hughes, 1979).   In  addition  there may




be direct effects  of  aniline on the central nervous system




(e.g., insomnia, paresthesias) and cardiotoxic effects




(Patty,  1979;  Sax, 1979).




     Chronic exposure induces amenia (Patty, 1979;  Sax,  1979;




Proctor  and Hughes, 1979).  Other chronic effects of aniline




exposure are hepatic  injury, (perhaps caused by an  aniline




metabolite) (Jenkins, 1972;  Geomet, 1980) and splenic hemosi-




derosis  (NCI,  1978).




V.   AQUATIC TOXICITY




     The lethal concentration and threshold concentration for




aniline with respect  to Chironomus doraalis Meig. larvae are




6  and 3 mg/1,  respectively (Puzikova and Markin, 1975).






                            172-9

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Aniline also produces toxic effects in Daphnia, at 0.4 mg/1


(Verschueren, 1977).


     B.   Chronic Toxicity


          Lakhnova 1975, reported that aniline at a.


concentration of 0.2 mg/1 is lethal to Daphnia magna Straus


within nine days.  Therefore, Lakhnova recommended a maximum


permissible limit of 0.02 mg/1.


VI.  EXISTING GUIDELINES AND STANDARDS


     A.   Human


          The maximum allowable concentration in class I


waters for the production of drinking water in the U.S. is 5
                    o

mg/1.  Several European countries have set lower limits: 0.1
                                                t

mg/1 (USSR), 5 ppm (Federal German Republic), 2.6 ppm


(Deutsche Demokratische Republik), and 1.3 ppm (CSSR)


(Verschueren, 1977).


     B.   Aquatic


          Data were insufficient to draft a criterion for


protection of freshwater or marine life.  Lakhnova (1975)


recommended a maximum permissible limit of 0.02 mg/1.
                            172-10

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Casarett, L. J. and J. Doull, 1975.  Toxicology;
The. Basic Science of Poisons, MacMillan Press, New
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Courtney, K. D., 1979.  Postpartun CPK (Creatine
Phosphokinase) and'LDH (Lactic Dehydrogenas) Cardiac
and Serum Isozymes After 2,4,5-T, Carbaryl, or Aniline
Treatment, Toxicol. Appl. Pharmacol.  48(1): A139, 1979.

Czajkowska, T., B. Krysiak, and J. Stetkiewicz, 1977.
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Deichmann, W. B. and H. W. Gerarde, 1969.  Toxicology
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Geomet Technologies Inc., 1980.  Aromatic Amines (Draft),
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Gosselln, R. E., H. C. Harold, R. P. Smith, and M. N.
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Hawley, G. G., 1977.  The Condensed Chemical Dictionary,
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IARC, 1974.  Monographs on the Evaluation of Carcinogenic
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Jenkins, F.P., J.A. Robinson, J.B. Gellatly, and GWA
Salmond, 1972, the no-effect doses aniline in human
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the rat.  Food Cosmet. Toxicol.  10:671-679.

Lakhnova, V.A., 1975.   Effect of aniline  on Daphnia
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McCaleb, K. E. 1979.  Chemical Economics  Handbook.  "Aniline
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Park, CA.
                       172-11

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Nagao, M. T. Yahagi, M. Honda, U. Seino, T. Matsutshima,
and T. Sugimura.  1977.  Demonstration of Mutagenicity
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NCI, 1978.  Carcinogenesis Technical Report No. 130,
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Parke, D. V., 1960.  The metabolism of (14C)- aniline
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Patty, F., 1979.  Industrial Hygiene and Toxicology,
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Pietrowski, J. K., 1972.  Certain problems of expsoure
tests for aromatic compounds.  Pracov. Lek, 24, 94.

Pietrowski, J. 1957, Quautitative estimation of aniline
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Proctor, N.H. and J.P. Hughes, 1979, Chemical Hazards
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Sax, NI, 1979, Dangerous Properties of Industrial
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Puzikova, N. B. and V. N. Markin, 1975.  Effect of
aniline and aniline hydrochloride on Chironomus dorsalis
Meig. larvae. Tr. Safat. Ofd. Cos NIORKh; Vol. 13, 104-9.

Stecher, P. G. (ed.) 1968.  The Merck Index, 8th ed.,
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Slimak, K, Robert Hall, and Ronald Burger, 1980.  Level
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mental Protection Agency (EPA-560/13-80-013).
                       172-12

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Suglmura, T. and M. Hagao, 1979.  Mutagenic factors in
cooked food, CRC Grit. Rev. Toxciol. Vol. 6, ISS 3,
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Verschueren, K., 1977.  Handbook of Environmental Data
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Weast, R. C. (e'd.) 1977.  Handbook of Chemistry and
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Wisniewska - Knypl, J. M., 1978.  Activity of drug -
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Env. Xenobiotics, 141-4.
                       172-13

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