BACKGROUND DOCUMENT
RESOURCE CONSERVATION AND RECOVERY ACT
SUBTITLE C - IDENTIFICATION AND LISTING OF HAZARDOUS WASTE
APPENDIX A - HEALTH AND ENVIRONMENTAL EFFECT PROFILES
APRIL 30, 1980
U.S. ENVIRONMENTAL -PROTECTION AGENCY
OFFICE OF SOLID WASTE
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Preface
These health and environmental effect profiles have
been compiled to support the listing of approxmately 170
of the hazardous constituents identified on Appendix VIII
in the regulations (40 CFR, Part 261). These profiles are
also being used to support the listing of hazardous wastes
in Subpart D of Part 261, due to the presence in the
wastes, of these hazardous constituents. Many of these
profiles have been summarized from the water quality criteria
documents prepared in support of various programs under
the Clean Water Act. In each case, however, the document
is based on information and references available to the
Agency and which are referenced in each Individual document.
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Table of Contents
Chemical Substance(Document Number) Pjige
Acetaldehyde(1) 1
Acetonitrile(2) 10
Acetophenone(3) 22
Acetyl Chloride(4) 29
Acrolein(5) 35
Acrylamlde(Re served)
Acrylonltrlle(7) 51
Aldrin(8) 65
Allyl Alcohol(9) 79
Antimony(10) 87
Arsenic(ll) 104
Asbestos(12) 125
Barium(13) 145
Benzal Chloride(14) 156
Benzene(15) 163
Benzidlne(16) 179
Benz(a)anthracene( 17) 193
Benzo(b)fluoranthene( 18) 205
Benzo(a)pyrene(19) 216
Benzotrichloride(20) 228
Benzyl Chloride(21) 235
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Chemical Substance(Document Number) Page
Beryllium(22) - " 247
Bis(2-chloroethoxy) Methaae(23) 263
Bis(2-chloroethyl) Ether(24) 269
Bis(2-chloroisopropyl) Ether(25) 280
Bis(chloromethyl) Ether(26) 288
Bis(2-ethylhexyl) Phthalate(27) 298
Bromoform(28) 312
Bromotnethane(29) 322
4-Bromophenyl Phenyl Ether(30) 332
Cadmium(31) 339
Carbon Disulfide(32) 366
Carbon Tetrachloride (Tetrachloromethane)(33) 374
Chloral(34) 387
Chlordane(35) 400
Chlorinated Benzenes(36) 418
Chlorinated Ethanes(37) 435
Chlorinated Naphthalenes(38) 453
Chlorinated Phenols(39) 464
Chloroacetaldehyde(40) 486
Chloroalkyl Ethers(4l) 497
Chlorobenzene(42) 510
p-Chloro-m-cresol(43) 520
Chloroethane(44) 526
Chloroethene(Vinyl Chloride)(45) 533
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Chemical SubstanceCDocument Number) Page
2-Chloroethyl Vinyl Ether(46) 550
Chloroform (Carbon Trichloromethane)(47) 558
Chloromethane(48) 574
2-Chloronaphthalene(49) 584
2-Chlorophenol(50) 595
Chromium(51) 607
Chrysene(52) 626
Cresote(53) 637
Cresols and Cresylic Acid(54) 653
Crotonaldehyde(55) 684
Cyanides(56) 694
Cyanogen Chlorlde(57) 707
DDD(58) 713
DDE(59) 724
DDT(60) 734
Dibromochloromethane(61) 751
Di-n-butyl Phthalate(62) 758
Diben2o(a,h)anthracene(63 ) 767
l,2-Dichlorobenzene(64) 779
l,3-Dichlorobenzene(65) 790
l,4-Dichlorobenzene(66) 798
Dichlorobenzenes(67) 809
3,3'-Dichlorobenzidine(68) 823
l,l-Dichloroethane(69) ' 836
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Chemical Subsjijince(Document Number) Page
l,2-D'ichloroethane(70) 843
l,l-Dichloroethylene(71) 855
trans-l,2-Dichloroethylene(72) 866
Dichloroethylenes(73 ) 874
Dichloromethane(74) 887
2,4-Dichlorophenol(75) 898
2,6-Dichlorophenol(76) 911
2,4-Dichlorophenoxyacetic Acid (2,4-D)(77) 918
l,2-Dichloropropane(78) 935
Dichloropropanes/Dichloropropenes(79) 944
Dichloropropanol(SO) 955
l,3-Dichloropropene(81) 962
Dieldrin(82) 970
o,o-Diethyl Dithiophosphoric Acid(83) 991
o,o-Diethyl-S-methyl Phosphorodithioate(84} 999
Diethyl Phthalate(85) 1006
Dimethylnitrosamine(86) 1014
2,4-Dimethylphenol(87) 1024
Dimethyl Phthalate(88) 1035
Dinitrobenzenes(89) 1043
4,6-DInltro-o-cresol(90) 1052
2,4-Dinltrophenol(91) 1060
DInitrotoluene(92) 1070
2,4-Dinitrotoluene(93) '1083
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Chemical Substance(Document Numb_er_) Page
2,6-Dinitrotoluene(94) 1095
Di-n-octyl Phthalate(95 ) 1104
l,2-Diphenylhydrazine(96) 1111
Disulfoton(97) 1121
Endosulfan(98) 1132
Endrin(99) 1149
Epichlorohydrin <1-Chloro-2,3-epoxypropane)(100) 1167
Ethyl Methaerylate(lOl) 1181
Ferric Cyanide{102) 1189
Fluoranthene<103) 1195
Formaldehyde(104) 1206
Formic Acid(105) 1221
Fumaronitrile(106) 1231
Halomethanes(107 ) 1237
Heptachlor(108) 1252
Heptachlor Epoxide(109) 1271
Hexachlorobenzene(110) 1283
Hexachlorobutadiene(111) 1297
Hexachlorocyclohexane(112) 1310
gamma-Hexachlorocyclohexane(113) 1330
Hexachlorocyclopentadiene(114) 1349
Hexachloroethane(115) 1361
Hexachlorophenc(116) 1369
Hydrofluoric Acid(117) '1378
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Cjiemical Subs tance(Document Number) Page
Hydrogen Sulfide(118) 1390
Indeno (1,2,3-cd) Pyrene(119) 1400
Isobutyl Alcohol(120) 1410
Lead (121) 1415
Malelc Anhydride(122) 1434
Malononltrile(123) 1441
Mercury(124) 1451
Methomyl(125) 1475
Methyl Alcohol(126) 1491
S,S'-methylene-o.o,o',o'-Tetraethyl Phosphorodithioate(127) 1513
Methyl Ethyl Ketone(128) 1520
Methyl Isobutyl Ketone(129) 1526
Methyl Methacrylate(130) 1532
Naphthalene(131) 1543
l,4-Naphthoquinone(132) 1556
Nickel(133) 1563
Nitrobenzene(134) 1579
4-Nitrophenol(135) 1591
NItrophenols(136) 1600
Nitrosamines(137) 1616
N-NItrosodiphenylamine(138) 1633
N-Nitrosodi-n-propylamine(139) 1643
Paraldehyde(140) 1657
Pentachlorobenzene(141) 1666
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Chemical Substance(Document Number) Page
Pentachloronitrobenzene(142) 1675
Pentachlorophenol(143) 1690
Phenol(144) 1706
Phorate(145) 1722
Phthalate Esters(146) 1737
Phthalic Anhydride(147) 1753
2-Picoline(148) 1760
Polynuclear Aromatic Hydrocarbons(PAHs)(149) 1769
Pyridine(150) 1791
Quinones(151) 1801
Resorcinol(152) 1810
Selenium(153) 1821
Silver(154) 1833
TCDD(155) 1848
l,l,l,2-Tecrachloroethane(156) 1862
1,1,2,2-Tetrachloroethane{157) 1872
Tetrachloroethylene(Perchloroethylene)(158) 1883
Thallium(159) 1897
Toluene(160) 1909
2,4-Toluenediamine(161) 1926
Toluene Dilsocyanate{162) 1935
Toxaphene(163) 1949
1,1,l-Trichloroethane(164) 1970
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Chemical Substance(Doeument Number) Page
l,l,2-Trichloroethane(165) 1981
Trichloroethylene(166) 1990
Trichlorofluoromethane and Dichlorodifluoromethane(167) 2003
2,4,6-Trichlorophenol(168) 2014
l,2,3-Trichloropropane(169) 2026
0,0,o-Trlethyl Phosphorothloate(170) 2033
Trinitrobenzene( 171) 2040
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No. 1
Acetaldehyde
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 acc-uracy.
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ACETALDEHYDE
Summary
An increased incidence of malignant neoplasms was reported in
workers in an aldehyde factory. Acetaldehyde was found in
concentration of 1 to 7 mg/m-* but there was no indication that
acetaldehyde was the causative factor for the cancers.
Equivacol results were obtained from a number of mutugenicity
assays.
I. INTRODUCTION
Acetaldehyde (CH3COH) is a clear, flammable liquid with a
pungent, fuity odor. It has the following physical/chemical
properties (Hawley, 1977; U.S. EPA, I976a):
^0
Chemical Structure; CH3 - C'^
^H
CAS No.: 75-07-0
Molecular Formula: C2H40
Boiling Point: 20.2°C
Melting Point: -123.5°C
Vapor Pressure: 740 mm (20"C)
Density: 0.7834 at 18°C/4°C
Octanol/Water
Partition Coefficient: 0.43
Vapor Density: 1.52
Solubility: soluble in water and most
organic solvents
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A review of the production range "(includes importation)
statistics for acetaldehyde (CAS No. 75-07-0) which was listed in
the initial TSCA Inventory (1977) has shown that between 1 billion
and 2 billion pounds of this chemical were produced/imported in
1977. *_/
Acetaldehyde is used mainly as a chemical intermediate in the
production of paraldehydes, acetic acid, acetic anhydride, and a
variety of other chemicals (Hawley, 1977),
II. EXPOSURE
The NIOSH National Occupational Hazard Survey estimates that
2,430 workers are exposed to acetaldehyde annually (1976).
A. Environmental Fate
The available data do not Indicate a potential for persis-
tence and accumulation in the environment. While there is little
information on the environmental fate of acetaldehyde, the BOD/COD
of 0.72 confirms that acetaldehyde will readily biodegrade
(Verschueren, 1978).
As to its fate in air, aldehydes are expected to photodisso-
ciate rapidly and competively with their oxidation for a half-life
of 2 to 3 hours. Aldehydes do not persist in the atmosphere but
the fact that acetaldehyde is a component of vehicle exhaust may be
significant in its contribution to smog (U.S. EPA, I977b).
*/ This production range information does not include any production/
~~ importation data claimed as confidential by the person(s) report-
ing 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 Regulation (40 CFR 710).
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B. Bioconcentration
Acetaldehyde has an octanol/water partition coefficient
of 0.43 indicating that it is highly hydrophilic and should not
accumulate (U.S. EPA, 1976).
C. Environmental Occurrence
Acetaldehyde is a normal intermediate product in the
respiration of higher plants; it occurs in traces in ripe fruits
and may form in alcoholic beverages after exposure to air. It has
been reported that acetaldehyde is found in leaf tobacco, ciga-
rette smoke, and automobile and diesel exhaust (U.S. EPA, I977a).
Acetaldehyde has been reported in both finished drinking water
supplies and effluents from sewage treatment plants in several
locations throughout the U.S. (EPA, I976b).
III. PHARMACOKINETICS
Acetaldehyde which is the first occurring metabolite of ethanol
in mammals is produced in the liver and is often found in various
tissues after the consumption of alcohol (Obe and Ristow, 1977).
It is an intermediate product in the metabolism of sugars in the
body and hence occurs in traces in blood (EPA, 1977b).
IV. HEALTH EFFECTS
A. Careinogenicity
Watanabe and Sugimoto (1956) administered 0.5-5% acetalde-
hyde subcutaneously to rats for a period of 489 to 554 days. Four
of the 14 animals developed spindle cell carcinomas at the site of
injection.
•
An increased incidence of malignant neoplasms has been observed
in workers at an aldehyde factory who were exposed to acetaldehyde,
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butyraldehyde, crotonaldehyde, aldol, several alcohols, and longer
chain aldehydes. Acetaldehyde was found in concentrations of
1-7 mg/m3. Of the 220 people employed in this factory, 150 has
been exposed for more than 20 years. During the period 1967 to
1972, tumors were observed in nine males (all of whom were smokers).
The tumor incidences observed in the workers exceeded incidences of
carcinomas of the oral cavity and bronchogenic lung cancer expected
in the general population and, for the age group 55-59 years, the
incidence of all cancers in chemical plant workers. There is no
indication that acetaldehyde was the causative factor in the excess
incidence of cancer (Bittersohl, 1974; Bittersohl, 1975).
Acetaldehyde has been found positive in a variety of mutagenicity
tests: siter chromatid exchange in cultured human lymphocytes and
a Chinese hamster (ovary) cell line (Ristow and Obe, 1978; Obe and
Ristow, 1977); S. typhimirium (Ames Test); (Pol A~) E. coli
(Rosenkranz, 1977); and WP2 uvrA trp~) E. coli (Veghelyi et a^.,
1978). It has, however, also been reported negative by other
investigators: S. typhimurium, with and without activation (Cotruvo
et al., 1977; Commoner, 1976; Laumbach et al., 1977); Saccharomyces
cerevis tae test for recombination (Cotruvo et al., 1977); and
B_a_c_i_llus sub t i113 repair essay (Laumbach et al., 1977). Thus, of
tan reports of in vj.tro tests for the mutagenicity of acetaldehyde,
5 were positive and 5 were negative. Acetaldehyde was also found
to cross-link isolated calf thymus DNA (Ristow and Obe, 1978).
C. Other Toxlcity
1. Acute
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A table summarizing the acute toxicity of acetaldehyde
In rats and mice Is found below:
Species
rat
rat
rat
rat
rat
mouse
mouse
Dose
Route
Result
Reference
I6,000ppm x 4 hrs.
4,000ppm x 4 hrs.
640 mg/kg
20,000ppm x 30 min.
1,930 mg/kg
560 mg/kg
1,232 mg/kg
ihl
ihl
s . c .
ihl
oral
s . c .
oral
lethal
lethal
LD50
LC50
LD50
LD50
LD50
Smyth, 1956
NIOSH, 1977
Skog, 1950
Skog, 1950
NIOSH, 1977
Skog, 1950
NIOSH, 1977
D. Other Relevant Data
Acetaldehyde Is a mucous membrane irritant in humans
(Verschueren, 1978).
V. AQUATIC EFFECTS
A. Acute
The 24-hour median threshold limit (TLm) for acetaldehyde
pinperch is 70 mg/1. The 96-hour TLm in sunfish is 53 mg/1
(Verschueren, 1978).
VI. EXISTING GUIDELINES
A. Humans
The American Conference of Governmental and Industrial
Hygienists (ACGIH) has adopted a Threshold Limit Value (TLV) of
100 ppm for acetaldehyde. The OSHA standard In air Is a Time
Weighted Average (TWA) of 200 ppm.
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REFERENCES
ACGIH (1977). American Conference of Governmental and Industrial
Hygienists, Threshold Limit Values for Chemical Substances and
Physical Agents in the Workroom Environment, Cincinnati, Ohio.
Bittersohl, G. (1974). Epdemiological investigations on cancer
in workers exposed to aldol and other aliphatic aldehydes. Arch.
Geschwalstforsch. 43:172-176.
Bittersohl, G. (1975). Env. Qual. Safety. 4:285-238 (as cited
in NCI, 1978).
Commoner, B. (1976). Reliability of bacterial mutagenesis
techniques to distinguish carcinogenic and non-carcinogenic
chemicals. EPA-600/1-76-002.
Cotruvo, J.A. e_t^ a_l_-> (1977). Investigation of mutagenic effects
of products of ozonation reactions in water. Ann. N.Y. Acad.
Scl. 298:124-140.
Hawley, G.G. (1977). Condensed Chemical Dictionary, 9th edition.
Van Nostrand Reinhold Co.
Laumbach, A.D., et al. (1977). Studies on the mutagenicity of
vinyl chloride metabolites and related chemicals. Prev. Select.
Cancer. (Proc. Int. Symp.) 1:155-170.
NIOSH (1976). National Occupational Hazard Survey.
NIOSH (1977). Registry of Toxic Effects of Chemical Substances.
Obe, G. , and H. Ristow. (1977). Acetaldehyde, But Not Ethanol,
Induces Sister Chromatid Exchanges in Chinese Hamster Cells in
Vitro. Mutation Research. 56:211-213.
National Cancer Institute, Chemical Selection Working Group,
September 28, 1978.
OSHA (1976). Occupational Safety and Health Standards (29 CFR
1910), OSHA 2206.
Ristow, H., and G. Obe. (1978). Acetaldehyde Induces Cross-Links
In DNA and Causes Sister-Chromated Exchanges in Human Cells.
Mutation Research 58:115-119,
-------
Rosenkranz, H.S. (1977). Mutageniclty of halogenated alkanes and
their derivatives. Env. Hlth. Perspect. 21:79-84.
Skog, E. (1950). A toxicological Investigation of lower aliphatic
aldehydes I. Toxicity of formaldehyde, acetaldehyde, propionaldehyde,
and butyraldehyde; as well as of acrolein and crotonaldehyde.
Acta Pharmacol. 6:29-318.
Smyth, H.F. (1956). Am. Ind. Hyg. Assn. Quarterly, 17:144.
U.S. EPA (1976a). Preliminary Scoring of Selected Organic Air
Pollutants. EPA-450/3-77-008. PB 264-443.
U.S. EPA (1977a). Potential Industrial Carcinogens and Mutagens.
EPA-560/5-77-005.
U.S. EPA (1977b). Review of the Environmental Fate of Selected
Chemicals. EPA-560/5-77-003.
U.S. EPA (19790. Toxic Substances Control Act Chemical Substances
Inventory, Production Statistics for Chemicals on the Non-Confidential
Initial TSCA Inventory.
Veghelyi, P.V. £££!_• (1978). The fetal alcohol syndrome: symptoms
and pathogenesis. Acta Pediatr. Acad. Sci. Hung. 19:171-189.
Verschueren, K. (1978). Handbook of Environmental Data on Organic
Chemicals. Van Nostrand Reinhold Co., New York.
Watanabe, F. and S. Sugimoto (1956). Study on the carcinogeniclty
of aldehyde. 3rd Report. Four cases of sarcomas of rats appearing
in areas of repeated subcutaneous Injections of acetaldehyde.
Gann. 47:599-601.
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No. 2
Aceton!trlie
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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ACETONITRILE
SUMMARY
Depending on the amount absorbed, acetonitrile may cause
disorders in the central nervous system, Liver, kidneys, car-
diovascular system and gastrointestinal system, regardless of
the route of administration. These effects are attributed to
the metabolic release of cyanide from the acetonitrile mole-
cule, although the parent molecule itself may cause these ef-'
f ects .
This Hazard Assessment Profile was based Largely on in-
formation obtained from NIOSH and its Criteria for a Recom-
mended Standard: Occupational Exposure to Nitriles, (NIOSH,
1978).
The NIOSH 1972-1974 National Occupational Hazards Survey
estimates that about 26,000 workers are occupationaLly ex-
posed to nitriles.
Major occupational exposures to nitrile occur by inhala-
tion of vapor or aerosols and by skin absorption. Adverse
effects of nitriles are also found from eye contact.
There is no available evidence to indicate that acetoni-
triLe has mutagenic or carcinogenic activity. Two studies
have reported teratogenic effects in rats.
Unlike the immediate onset of cyanide toxicity, nitrile
poisoning displays a delayed onset of symptoms.
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I. INTRODUCTION
Acetoni t ri le (CH^CN) is a mononitrile and falls into
the saturated aliphatic class of nitrites. It is a colorless
liquid and has a vapor pressure of 73 mm Hg at 20' C. It has
a molecular weight of 41.05 and a specific gravity of 0.786
(NIOSH, 1978).
When heated to decomposition, nitrites emit toxic fumes
containing cyanides (Sax/. 1968).
Acetonitrile was introduced to the commer i ca I market in
1952, and its industrial uses lie in the manufacture of plas-
tics, synthetic fibres, elastomers, and solvents. Acetoni-
trile is used as a solvent in the extractive distillation
that separates olefins from diolefins, butadiene from buty-
Lene, and isoprene from isopentane.
In 1964, 3.5 million pounds of acetonitrite were con-
sumed industrially.
" II. EXPOSURE
A. Water and Food
Pertinent data were not found in the available lit-
erature.
. B. Inhalation
Acetonitrile can be readily absorbed from oral mu-
cosa (McKee, et al. 1962; Dalhamn, et al. 1968).
1 In the workplace, acute poisoning and death have
f*-l*JVt-
been reported following the inhalation of acetonitrile (De-
quidt, et al.. 1 974) .
Studies have demonstrated that acetonitrile is ab-
sorbed by lung tissue (Dequidt, et al. 1974; Grabois, 1955;
Amdur, 1959).
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C. DermaL
Dermal exposures to acetonitriLe have caused ad-
verse reactions including death in some cases (NIOSH, 1978).
Acetonitri le has been reported to have been absorb-
ed through the intact skin of rabbits, yielding a dermal
LDgo of 980 mg/kg (Pozzani, et a 1. 1959).
III. PHARMACOKINETICS
A. Ab sorpti on
Acetonitrile is a component of cigarette smoke and
is absorbed by the oral tissues (McKee, et al. 1962; Dalhamn,
Jet a I. 1968) .
Humans have been shown to absorb acetonitrile di-
rectly through the skin and respiratory tract (Zeller,. et al.
1969; Amdur, 1959; Dequidt, et al. 1974).
B. Distribution
Studies by McKee, et al. (1962) and Dalhamn, et aL.
(1968) show that acetonitrile from cigarette smoking is re-
tained by the lungs.
Tissue distribution studtes indicated that mononi-
triles (and acetonitrile, in particular) are distributed uni-
formly -in the internal organs of humans and that cyanide me-
tabolites are found predominantly in the spleen, stomach and
skin, and to a lesser extent, in the liver, lungs, kidneys,
hearts, brain, muscle, intestines, and testes COequidt, et
aL. 1974).
*
Haguenoer, et al. (1975) exposed three rats to
2,300 or 25,000 ppm acetonitrile by inhalation. At 25,000
ppm, all three rats died after 30 minutes. Chemical analysis
-/*/-
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of the organs showed that the mean concentration of
acetonitrile in mus c Le "was 136 ^ig/100 g of tissue and 2,438
ug/100 g of kidney tissue. High acetonitrile excretion or
possible renal blockage were postulated as the causes for the
high renal concentration.
Nitriles and their metabolic products have been de-
tected in urine, blood and tissues (McKee, et al. 1962).
C. Metabolism
Since human and animal studies report symptoms
characteristics of cyanide poisoning, it is reasonable to
assume that a portion of the effects of exposure to acetoni-
trile is due to the release of the cyanide ion from the par-
ent compound (Zeller, et al. 1969; Amdur, 1959; Pozzani,
1959).
After absorption, nitriles may be metabolized to an
alpha cyanohydrin or to inorganic cyanide, which is oxidized
.. to thiocyanate and is excreted in the urine. The C=N group
may be converted into a carboxylic acid derivative and ammon-
ia, or may be incorporated into cyanocobalamine. Ionic cya-
nide also reacts with carboxyl groups and with disulfides
(McKee, et al. 1962).
Haguenoer, et al (1975) injected white male Wistar
rats with varying levels of acetonitrile ranging from 600
mg/kg to 2,340 mg/kg. At autopsy, the internal organs showed
that the combined hydrogen cyanide consisted essentially of
thiocyanates , cyanohydrins and cyanocoba lamines , •
D. Excretion
Acetonitrile is found in the morning urine of cigar-
ette smokers. Concentrations of acetonitrile range from 2.2
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of the organs showed that the mean concentration of
acetoni tri Le in muscle was 136 jjg/100 g of tissue and 2,438
jjg/100 g of kidney tissue. High acetonitrile excretion or
possibLe renal blockage uere postulated as the causes for the
high renal concentration.
Nitrites and their metabolic products have been de-
tected in urine, blood and tissues (WcKee, et al. 1962).
C. Metabo Lism
Since human and animal studies report symptoms
characteristics of cyanide poisoning, it is reasonable to
assume that a portion of the effects of exposure to acetoni-
trile is due to the release of the cyanide ion from the pai—
ent compound (Zeller, et al. 1969; Amdur, 1959; Pozzani,
1959).
After absorption, nitriles may be metabolized to an
alpha cyanohydrin or to inorganic cyanide, which is oxidized
- to thiocyanate and is excreted in the urine. The C=N group
may be converted into a carboxylic acid derivative and ammon-
ia, or may be incorporated into cyanocoba lamine. Ionic cya-
nide also reacts with carboxyl groups and with disulfides
(McKee, et al. 1962).
Haguenoer, et al (1975) injected white male Wistar
rats with varying levels of acetonitrile ranging from 600
mg/kg to 2,340 mg/kg. At autopsy, the internal organs showed
that the combined hydrogen cyanide consisted essentially of
thiocyanates, cyanohydrins and cyanocobaLamines.
D. Excretion
Acetonitrile is found in the morning urine of cigar-
ette smokers. Concentrations of acetonitrile range from 2.2
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jug/100 ml urine for those smoking three cigarettes per day up
to 20 jjg/100 ml urine for heavy smokers (up to 2.5 packs per
day). The results showed that acetonitrile, once absorbed
into the body, can be excreted unchanged in the urine (McKee,
et al. 1962).
Acetonitrile is also excreted unchanged in exhaled
air (Haguenoer, et al. 1975).
IV. EFFECTS
A. Carcinogenicity
Dorigan, et al. (1976) failed to show significant
carcinogenic effects in a two-year exposure study conducted
with rats.
8 . Mutag en i c i ty
Pertinent data were not found in the available lit-
erature.
C. Teratogenicity
IntraperitoneaI (i.p.) administration of acetoni-
trile to pregnant rats produced fetal malformations (Dorigan,
et al. 1976). Schmidt, et. al. (1976) have determined skele-
tal abnormaIities in rats following i.p, exposure to acetoni-
tri le.
D. Other Reproductive Effects
Pertinent data were not found in the available lit-
eratu re.
E . Chronic Toxicity
In an experiment to stimulate chronic occupational
exposure (seven hours per day, five days per week), 30 rats
were exposed to a concentration of 655 ppm acetonitrile for
K
- 17-
-------
90 days. The rats exhibited bronchial inflammation, desqua-
matization and hypersecretion of mucus, and hepatic and renal
Lesions. Monkeys exposed by the same regimen, but to 350 ppm
acetonitrile for 91 days, experienced bronchitis and moderate
hemorrhage of the superior and inferior sagittal sinuses of
the brain (Pozzani, et al. 1959).
Dogs exposed to acetonitrile at a concentration of
300 ppm for 91 days showed a reduction in body weight as well
as a reduction in hemoglobin and hematocrit values (Pozzani,,
et al. 1959).
Monkeys exposed to 660 ppm acetonitrile per day
showed poor coordination during the second week of exposure
and a monkey exposed to 330 ppm showed hyperexcitabiLity
toward the end of the 13th week (Pozzani, et al. 1959).
The same investigators reported chronic LOgg
values of 0.85 and 0.95 ml/kg for female rats which i.p, ad-
ministration of acetonitrile.
G. Other Relevant Information
Dogs exposed with lethal quantities of acetonitriLe
(16,000 ppm for four hours) showed blood cyanide Levels rang-
ing from 305-433 pg/100 ml of blood after three hours (Poz-
zani, et al. 1959).
V. AQUATIC TOXICITY
A. Acute
Observed 96-hour LCgg values for the fathead
minnow (Pimephales promelas) are 1020 mg/L in hardwater an'd
1000 mL/l in softwater (Bringmann, 1976). For bLuegills,
(Lepomi s maereehi rus) and guppies (LebistGS reti cu latus), the
-------
respective 96-hour values in softwater are 1850 mg/I and 1650
mg/L (Jones, 1971; Henderson, et al. 1960).
B. Chronic/ Plant Effects, and Residue
Pertinent data were not found in the available lit-
erature.
C. Other Relevant Information
Acetonitrile has been observed to damage the bron-
chial epithelium of fish (Belousov, 1969). This compound,
when added to the aqueous environment of roaches and fil-
berts, disrupted blood circulation and protein metabolism and
induced hyperemia, hemorrhages, and the appearance of small
granules in the heart, brain, liver, and gills of fish. The
hepatic glycogen Level decreased sharply. CH^CN induced-
death apparently resulted from circulatory disturbances and
necrohiotic changes in the cerebral neurons (Belousov, 1972).
Acetonitrile at a concentration of 100 mg/l inhib-
ited nitrification in saprophytic organisms (Chekhovskaya,
1966).
VI. EXISTING GUIDELINES
A. Human
A federal occupational standard exists for acetoni-
trile and is based on the TLV for workplace exposure pre-
viously adopted by American Conference of Governmental and
Industrial Hygienists. This TLV is 40 ppm (70 mg/m3) and
is an eight-hour TWA.
3. Aquat i c
Pertinent data were not found in the available lit-
erature.
-------
REFERENCES
Amdur, M.L. 1959. Accidental group exposure to acetoni-
triLes - A clinical study. J. Occup. Med. 1: 627.
American Conference of Governmental Industrial Hygienists.
Threshold Limit values for chemical substances and physical
agents in the workroom environment, with intended changes for
1979. Cincinnati, Ohio. 94 pp.
Belousov, Y.A. 1969. Effects of some chemical agents on the
histophysiological state of the bronchial epithelium. CUch.
Zap. Yoroslav. Gos. Pedagog. Inst. USSR 62:126-129). Chem.
Abst. 97853c.
Belousov, Y.A. 1972. Morphological changes in some fish
organs during poisoning. Vlujanie Pestits. Dikikh Zhivotn.
41-45. Chem. Abst. 141567d, Vol. 80.
Bringmann, 6. 1976. Vergleichende Vefunde der Schadwirkung
wassergefahrdender. Stoffee gezen Bakterien (Speudomomas
putida) und Blaualgen (Microcystis aeruginosa) nwfaL Iwasser.
117-119.
Checkhovskaya, E.V., et a 1. 196"6. Data for experimental
studies of toxicity of waste waters from aery lorn" triLe pro-
duction. (Vodosnabzh. Kanaliz. Gidrotekh. Sooruzh. Mezhved.
Resp. Nauch. USSR SB 1: 83-88). Chem. Abst. 88487k.
DaLhamn, T., et al. 1968. Mouth absorption of various com-
pounds in cigarette smoke. Arch. Environ. Health 16: 831.
Dequidt, J . , et al. 1974. Intoxication with acetonitrile
with a report on a fatal case. Eur. J. Toxicol. 7: 91.
Dirigan, et al. 1976. Preliminary scoring of selected
organic air pollutants. Environ. Prot, Agency, Contract No.
68-02-1495.
Grabois, B, 1955. Fatal exposure to methyl cyanide. NY
State Dep. Labor Div. Ind. Hyg. Mon. Rev. 34: 1,7,8.
Haguenoer, J.W., et al. 1975. Experimental acetonitrile
intoxications - I. Acute intoxicatios by the intraperitoneal
route. Eur. J. Toxicol. 8: 94.
Henderson, C., et al. 1960. The effect of some organic
cyanides (nitriles) on fish. Purdue Univ. Eng. Bull. Exp.
Ser. 106: 120.
Jones, H. 1971. Environmental control in the organic and.
petrochemical industries. Noyse Data Corp.
-20-
-------
McKee, H.C., et aL. 1962. Acetonitrile in body fluids re-
Lated to smoking. Public Health Rep. 77: 553.
NIOSH. 1978. NIOSH Criteria for a Recommended Standard:
Occupational Exposure to Nitrites. U.S. DHEW, Cincinnati.
Pozzani, V.C., et at. 1959. An investigation of the mammal-
ian toxicity of acetonitrile. J. Occup. Med. 1: 634.
Sax. N.I. 1968. Dangerous Properties of Industrial Materi-
als, 3rd ed. NY Van Nostrand Reinhold Co.
Schmidt, W., et at. 1976. Formation of skeletal abnormali-
ties after treatment with aminoacetonitri le and cycy lophosph-
amide during rat fetogenesis. (Verh. Anat. 71:635-638 Ger.)
Chem. Abst. 1515w.
Sunderman, F.W., and J.F. Kincaid. 1953. Toxicity studies
of acetone cyanohydrin and ethylene cyanohydrin. Arch. Ind.
Hyg. Occup. Med. 8: 371.
ZeLler, H.V., et aL. 1969. Toxicity of nitriles. ZentralbL
Arbirtsmed ArbeitsschutE. 19: 255.
-------
No. 3
Acetophenone
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.
-------
ACETOPHENONE
Summary
Acetophenone is present in various fossil fuel processes and products,
particularly coal and petroleum products. It is used as a flavoring agent
in products for human consumption and as an intermediate in organic
synthetic processes, particularly plastics manufacturing.
No data on the potential for carcinogenic, mutagenic, or teratogenic
effects or on the chronic toxicity of acetophenone were found in the
available literature.
There are no existing OSHA, NIOSH, or ACGIH standards or guidelines.-
Acetophenone is a skin irritant and has been shown to cause severe eye irri-
tation in rabbits at microgram quantities. Acetophenone is highly toxic to
aquatic life
-------
I. INTRODUCTION
Acetophenone (1-phenylethanone, phenyl methyl ketone, acetyl-
benzene, benzoyl methide, hypnone, C6H5COCH3; molecular weight 120.15)
is a liquid with a melting point of 20.5°C and is slightly soluble in
water. Acetophenone is used to impart a pleasant jasmine or orange-
blossom-like odor to perfumes, as a catalyst for the polymerization of ole-
fins, and in organic syntheses, especially as a photosynthesizer (Windholz,
1976). Additionally, it is used as a tobacco flavoring, as a solvent or
intermediate in the synthesis of Pharmaceuticals, and as a by-product of the
coal processing industry. Acetophenone is present in gasoline exhaust at
less than 0.1 to" 0.4 ppm (Verschueren, 1977).
II. EXPOSURE
No data on levels of acetophenone in food or water or on other
potential (inhalation or dermal) exposures were found in the readily avail-
able literature.
III. PHARMACOKINETICS
Information on the absorption, distribution, metabolism, or ex-
cretion of acetophenone was not found in the readily available literature,
despite the fact that it is used in pharmaceutical preparations and in
tobacco, perfume, and other products for human comsumption.
IV. EFFECTS
A. Carcinogenicity, Mutagenicity, Teratogenicity, and Chronic Toxicity
Readily available data are extremely limited. One paper suggests
the possible mutagenicity of acetophenone due to its ability to cause DNA
breakage in bacterial systems following DNA photosensitization (Rahn, et
al. 1974). Because of the particular sensitivity of the bacterial system
to DNA breakage, this information by itself is insufficient to establish
acetophenone as a mutagenic agent.
-------
There is no additional data readily available on the potential for
carcinogenic, mutagenic, or teratogenic activity by acetophenone. No data
are available on chronic toxicity. '
B. Acute Toxicity
Skin irritaion was observed in the rabbit at 10 mg/24 hrs. using
the draize procedure and at 515 mg when applied to the skin in the absence
of the absorbent gauze patch. Severe eye irritation was obtained in the
rabbit following application of 771 ug of acetophenene. The oral l_D,-n in
rats was 900. mg acetophenone/kg, while the lethal dose following intra-
peritoneal injection in mice was 200 mg/kg (NIOSH, 1978). Acetophenone is a
hypnotic in high concentrations and was used as an anesthetic in the last
century before less toxic substances were found (Kirk and Othtner, 1963).
C. Other Relevant Information
Based upon the retention time in a gas chromatographic/mass spec-
trographic column, Veith and Austin (1976) suggest a potential for bio-
accumulation of acetophenone. There is no additional information available
to verify this situation, however.
Microbial metabolism of acetophenone as the sole source of carbon
and energy has been demonstrated in pure culture (Cripps, 1975).
V. AQUATIC TOXICITY
Based upon reported values in the literature, acetophenone has
been shown to be highly toxic to aquatic life, (U.S. • EPA, 1979). LC5Q
values for fathead minnow are reported for the following time periods: 1
hour, greater than 200 mg/1; 24 hours, 200 mg/1; 48 hours, 163 mg/1; 72
hours, 158 mg/1; and 96 hours, 155 mg/1 (U.S. EPA, 1976).
Acetophenone has been reported to be a major constituent (36 per-
cent) of a weathered bunker fuel. This suggests that it may be present in
large quantity following spills of some bunker fuels (Guard, et al. 1975).
-------
Bunker fuels are highly variable form refinery to refinery; thus, a blanket
statement as to percentage composition of acetophenone or other constituents
cannot be made.
VI. EXISTING GUIDELINES AND STANDARDS
There are no existing guidelines and standards from OSHA, NIOSH,
or ACGIH. Similarily, no ambient water quality standards for acetophenone
exist.
-------
REFERENCES
Cripps, R.E. 1975. The microbial metabolism of acetophenone: metabolism
of acetophenone and some chloroacetophenones by an Arthrobacter species.
Biochem. Jour. 152: 233.
Guard, H.E., et al. 1975. Identification and potential biological effects
of the major components in the seawater extract of a bunker fuel. Bull.
Environ. Contam. Toxicol. 14: 395.
Kirk, R.E. and D.F. Othmer. 1963. Kirk-Othmer. Encyclopedia of Chemical
Technology. 2nd ed. J. Wiley and Sons, Inc., New York.
National Institute for Occupational Safety and Health. 1978. Registry of
Toxic Effects of Chemical Substances. E. Fairchild (ed.). U.S. Department
of Health, Education, and Welfare. Cincinnati, Ohio.
Rahn, R.O., et al. 1974. Formation and chain breaks and thymine dimers in
DMA upon photqsensitization at 313 nm with acetophenone, acetone, or benzo-
phenone. Photochem. Photobio. 19: 75.
U.S. EPA. 1976. Acute Toxicity of Selected Organic Compounds to Fathead
Minnows. EPA-600-3-76-097. U.S. EPA Environmental Research Lab., Duluth,
Minnesota.
U.S. EPA. 1979. Biological Screening of Complex Samples From In-
dustrial/Energy Processes. EPA-600-8-79-021. U.S. EPA, Research Triangle
Park, North Carolina.
Veith, G.O. and N.M. Austin. 1976. Detection and isolation of bioaccumu-
latable chemicals in complex effluents. In: L.H. Keith (ed.), Identifi-
cation and Analysis of Organic Pollutants in Water. Ann Arbor Science
Publishers, Inc., Ann Arbor, MI. p. 297.
Verschueren, K. 1977. Handbook of Environmental Data on Organic Chem-
icals. Van Nostrand Reinhold Company, New York.
Windholz, M. (ed.) 1976. Merck Index. 9th ed. Merck and Co., Rahway,
N.O,
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No. 4
Acetyl Chloride
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.
-30-
-------
ACETYL CHLORIDE
Summary
Acetyl chloride is an irritant and a corrosive. Cutaneous exposure
results in skin burns, while vapor exposure causes extreme irritation of the
eyes and mucous membranes. Inhalation of two ppm acetyl chloride has been
found irritating to humans. Death or permanent injury may result after
short exposures to small quantities of acetyl chloride. An aquatic toxicity
rating has been estimated to range from 10 to 100 ppm.
However, acetyl chloride reacts violently with water. Thus, its half-
life in ambient water should be short and exposure from water should be nil,
The degradation products should likewise pose no exposure problems if the pH
of the water remains stable.
-31-
-------
ACETYL CHLORIDE
I. INTRODUCTION
Acetyl chloride (ethanoyl chloride; CFLCOC1; molecular weight, 78.50)
is a colorless, fuming liquid with a pungent odor, a boiling point of
51-52°C, and a melting point of -112°C (Windholz, 1976). It is used as
an acetylating agent in testing for cholesterol and in the qualitative
determination of water in organic liquids. It is miscible with benzene,
chloroform, ether or glacial acetic acid (Windholz, 1976). In the presence
of water or alcohol, however, acetyl chloride hydrolyzes violently to form
hydrogen chloride and acetic acid. Phosgene fumes, which are highly toxic,
are emitted when acetyl chloride is heated to decomposition (Sax, 1975).
The 1975 U.S. annual production of acetyl chloride was approximately
4.54 x 10 grams (SRI, 1976). During transportation, this chemical should
be stored in a cool, well-ventilated place, out of direct sunlight, and away
from areas of high fire hazard; it should periodically be inspected (Sax,
1975). Acetyl chloride must be protected from water (Windholz, 1976).
II. EXPOSURE
Acetyl chloride reacts violently with water (see above). Thus, its
half-life in ambient water should be short and exposure from water should be
nil. The degradation products should likewise pose no exposure problems if
the pH of the water remains stable. Internal exposure to acetyl chloride
will most likely occur through inhalation of the vapor, or, on rare occa-
sions, through ingestion. Skin absorption is very unlikely although severe
burns would be expected.
III. PHARMACOKINETICS
Pertinent data could not be located in the available literature.
-------
IV. EFFECTS
Acetyl chloride is an irritant and a corrosive. Cutaneous exposure
results in skin burns. Vapor exposure causes extreme irritation of the eyes
and mucous membranes (Windholz, 1976). Inhalation of 2 ppm acetyl chloride
was found irritating to humans (Handbook of Organic Industrial Solvents,
1961). Death or permanent injury may result after very short exposures to
small quantities of acetyl chloride (Sax, 1975).
Because the toxicity of acetyl chloride might be expected to pattern
that of its breakdown product hydrogen chloride (HCL), LC. value (the
lowest concentration of a substance in air which has been reported to cause
death in humans or animals) for HC1 might be indicative of its toxicity.
This value in humans is 1000 ppm for one minute (Mason, 1974).
Pertinent information could not be located in the available literature
regarding the carcinogenicity, mutagenicity, teratogenicity and chronic
toxicity of acetyl chloride.
V. AQUATIC TOXICITY
Acetyl chloride has been shown to be toxic to aquatic organisms in the
ranges of 10 to 100 ppm (Hann and Jensen, 1974). No other information has
been found in the literature.
VI. EXISTING GUIDELINES AND STANDARDS
No standards for acetyl chloride have been reported. However, a
ceiling limit of 5 ppm has been reported for hydrogen chloride (the most
irratating hydrolysis product of acetyl chloride) in industrial exposures.
(Mason, 1974).
-33-
-------
ACETYL CHLORIDE
REFERENCES
Handbook of Organic Industrial Solvents, 2nd ed. 1961. Cited in: Registry
of toxic effects of chemical substances. NI05H (DHEW) Pub. No. 79-100, p. 4.
Hann, W. and P.A. Jensen. 1974. Water quality characteristics of hazardous
materials. Vol. 2. Texas A&M University.
Mason, R.V. 1974. Smoke and toxicity hazards in aircraft cabin furnish-
ings. Ann. Occup. Hyg. 17: 159.
Sax, N.I. 1975. Dangerous properties of industrial materials, 4th ed. Van
Nostfand Reinhold Co., New York, p. 355.
Stanford Research Institute. 1976. Chemical economics handbook.
Windholz, M. (ed.) 1976. The Merck Index, 9th ed. Merck and Co., Inc.,
Rahway, N.J., p. 11.
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No. 5
Acroleln
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.
-------
ACROLEIN .
SUMMARY
Acrolein has not been shown to be a carcinogen or cocarcinogen in in-
halation experiments. Acrolein is mutagenic in some assay systems. Infor-
mation on teratogenicity is not available. The only reported chronic effect
of acrolein in humans is irritation of the mucous membranes. Chronic expo-
sure of Syrian golden hamsters to acrolein in the air caused reduced body
weight- gains and inflammation and epithelialv metaplasia in the nasal
cavity. In addition, females had decreased liver weight, increased lung
weight, and slight hematologic changes.
Acrolein has been demonstrated to be acutely toxic in freshwater organ-
isms at concentrations of 57 to 160 pg/1. A single marine fish tested was
somewhat more resistant with a 48-hour LC-- of 240 jug/1. Toxicity to
marine invertebrates was comparable to that of freshwater organisms.
-37-
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ACROLEIN
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Acrolein (U.S. EPA, 1979).
Acrolein (2-propenal; CI-L=CHCHO; molecular weight 56.07) is a flamm-
able liquid with a pungent odor. It has the following physical and chemical
properties (Weast, 1975; Standen, 1967):
Melting Point -86.95°C
Boiling Point Range 52.5 - 53.5°C
Vapor Pressure 215mm Hg..at 20°C
Solubility Water: 210.8 percent by weight
at 20°C .
Density 0.8410 at 20°C
' Production (Worldwide) 59 kilotons (Hess, et al. 1978)
Capacity (Worldwide) 102 kilotons/year
Capacity (United States) 47.6 kilotons/year
Acrolein is used as a biocide, crosslinking agent, and tissue fix-
ative. It is used as an intermediate throughout the chemical industry.
The fate of acrolein in water was observed in natural channel waters
(Bowmer and Higgins, 1976). No equilibrium was reached between dissipating
acrolein and degradation products, with the dissipating reaction apparently
being continued to completion. Degradation and evaporation appear to be the
major pathways for loss, while a smaller amount is lost through absorption
and uptake in aquatic organisms and sediments (Bowmer and Sainty, 1977;
Hopkins and Hattrup, 1974).
II. EXPOSURE
There is no available evidence that acrolein is a contaminant of pot-
able water or water supplies (U.S. EPA, 1979).
Acrolein is a common component of food. It is commonly generated
during cooking or other processing, and is sometimes produced as an unwanted"
-------
by-product in the fermentation of alcoholic- beverages (Izard and Libermann,
1978; Kishi, et al. 1975; Hrdlicka and Kuca, 1965; Boyd, et al. 1965;
Rosenthaler and Vegezzi, 1955). However, the data are insufficient to
develop a conclusive measure of acrolein exposure from food processing or
cooking.
The U.S. EPA (1979) has estimated the weighted average bioconcentration
factor for acrolein to be 790 for the edible portions of fish and shellfish
consumed by Americans. This estimate is based on measured steady-state bio-
concentration studies in bluegills.
Atmospheric acrolein is generated as a combustion product of fuels and
of cellulosic materials (e.g., wood and cigarettes), as an intermediate in
atmospheric oxidation of propylene, and as a component of the volatiles pro-
duced by heating organic substrates (U.S. EPA, 1979). Acrolein is present
in urban smog; average concentrations of 0.012 - 0.018 mg acrolein/m and
peak concentrations of 0.030 - 0.032 mg acrolein/m were noted in the air
of Los Angeles (Renzetti and Bryan, 1961; Altshuller and McPherson, 1963).
Diesel exhaust emissions contained 12.4 mg acrolein/m ; trace amounts of
acrolein were present in samples taken from an area of traffic; and no acro-
lein was detected in ambient air from an open field (sensitivity of measure-
ment was below one part per million) (Bellar and Sigsby, 1970). Acrolein
content of smoke from tobacco and marijuana cigarettes ranged from 85 to 145
ug/cigarette (Hoffman, et al. 1975; Horton and Guerin, 1974). Acrolein was
detected at levels of 2.5 - 30 mg/m at 15 cm above the surface of pota-
toes or onions cooking in edible oil (Kishi, et al. 1975).
-------
III. PHARMACOKINETICS
A. Absorption
Total respiratory tract retention of acrolein in anesthetized dogs
was 77 to 86 percent (Egle, 1972).
B. Distribution
Pertinent data were not found in the available literature.
C. Metabolism
Relatively little direct information is available on the metabolism
of acrolein. In vitro, acrolein can serve as a substrate for alcohol dehy-
drogenases from human and horse liver (Pietruszko, et al. 1973). In vivo
studies in rats indicate that a portion of subcutaneously administered acro-
lein is converted to 3-hydroxylpropylmercapturic acid (Kaye and Young, 1972;
Kaye, 1973). Acrolein undergoes both spontaneous and enzymatically cata-
lyzed conjugation with glutathione (Boyland and Chasseaud, 1967; Esterbauer,
et al. 1975). The low pH's encountered in the upper portions of the gastro-
intestinal tract probably would rapidly convert acrolein to saturated alco-
hol compounds (primarily beta propionaldehyde) (U.S. EPA, 1979). As several
of the toxic effects of acrolein are related to the high reactivity of the
carbon-carbon double bond, saturation of that bond should result in detoxi-
fication (U.S. EPA, 1979).
D. Excretion
In rats given single subcutaneous injections of acrolein, 10.5 per-
cent of the administered dose was recovered in the urine as 3-hydroxy-
propylmercapturic acid after 24 hours (Kaye and Young, 1972; Kaye, 1973).
-------
IV. EFFECTS
A. Carcinogenicity
One-year and lifespan Inhalation studies with hamsters indicate
that acrolein is not a carcinogen or cocarcinogen (Feron and Kruysse, 1977;
National Cancer Institute, 1979).
B. Mutagenicity
Both positive and negative results have been obtained in muta-
genicity assays. Acrolein induced sex-linked mutations in Drosophila
melanogaster (Rapoport, 1948) and was mutagenic for DNA polymerase-deficient
Escherichia coli (Bilimoria, 1975) and Salmonella typhimurium (Bignami, et
al. 1977). Mutagenic activity was not detected in the dominant lethal assay
in ICR/Ha Swiss mice (Epstein, et al. 1972) or in a strain of E. coli used
for detecting forward and reverse mutations (with or without microsomal
activation) (Ellenberger and Mohn, 1976; 1977). Acrolein was weakly muta-
genic for Saccharomyces cerevisiae (Izard, 1973).
C. Teratogenicity
Pertinent data were not found in the available literature.
C. Other Reproductive Effects
Exposure of male and female rats to 1.3 mg/m acrolein vapor for
26 days did not have a significant effect on the number of pregnant animals
or the number and mean weight of fetuses (Bouley, et al. 1976).
E. Chronic Effects
*
Little information is available on the chronic effects of acrolein
on humans. An abstract of a Russian study indicates that occupational expo-
s.ure to acrolein (0.8 to 8.2 mg/m ), methylmercaptan (0.003 to 5.6
3 "? *
mg/nr), methylmercaptopropionaldehyde (0.1 to 6.0 mg/nr}, formaldehyde
(0.05 to 8.1 mg/m ), and acetaldehyde (0.48 to 22 mg/m3) is associated
-4J-
-------
with irritation of the mucous membranes. This effect is most frequent in
women working for less than one and greater than seven years (Kantemirova,
1975). Acrolein is known to produce irritation of the eyes and nose (Albin
1962; Rattle and Cullumbine, 1956; Sim and Pattle, 1957) and is thought to
be responsible, at least in part, for the irritant properties of
photochemical smog (Altshuller, 1978; Schuck and Renzetti, 1960) and
cigarette smoke (Weber-Tschopp, et al. 1976a; 1976b; 1977).
In the only published chronic toxicity study on acrolein in animals
(Feron and Kruysse, 1977), male and female Syrian golden hamsters were ex-
posed to acrolein at 9.2 mg/m in air, seven hours per day, five days per
week, for 52 weeks. During the first week only, animals evidenced signs of
eye irritation, salivated, had nasal discharge, and were very restless.
During the exposure period, both males and females had reduced body weight
gains compared to control groups. Survival rate was unaffected. Slight
hematological changes, increased hemoglobin content and packed cell volume,
decreases in liver weight (-16 percent), and increases in lung weights (+32
percent) occurred only in females. In both sexes, the only pathological
changes in the respiratory tract were inflammation and epithelial metaplasia
in the nasal cavity.
In a study of subacute oral exposure, acrolein was added to the
drinking water of male and female rats at 5 to 200 mg acrolein/1 for 90 days
(Newell, 1958). No hematologic, organ-weight, or pathologic changes could
be attributed to acrolein ingestion.
F. Other Relevant Information
Acrolein is highly reactive with thiol groups. Cysteine and other
compounds containing thiol groups antagonize the toxic effects of actolein
-------
(Tillian, et al. 1976; Low, et al. 1977; Sprince, et al. 1978; Munsch, et
al. 1973;1974; Whitehouse and Beck, 1975). Ascorbic acid also antagonizes
the toxic effects of acrolein (Sprince, et al. 1978).
The effects of acrolein, on the adrenocortical response of rats
unlike those of DDT and parathion, are not inhibited by pretreatment with
phenobarbital and are only partially inhibited by dexamethason (Szot and
Murphy, 1970). Pretreatment of rats with acrolein significantly prolongs
hexobarbital and pentobarbital sleeping time (Jaeger and Murphy, 1973).
V. AQUATIC TOXICITY
A. Acute Toxicity
A relatively narrow range of acute toxicity to six species of
freshwater fish has been reported for acrolein (U.S. EPA, 1979). LC^Q
values ranged from 61 to 160 ^ig/1 with fathead minnows, (Pimeghales
promelas), being most sensitive and largemouth bass, (Micropterus .
salmoides), the most resistant of the species tested. Results from 7 static
bioassays varying from 24 to 96 hours in duration were reported. The fresh-
water invertebrate D_aphnia magna was as sensitive to acrolein as freshwater
fish with 48-hour static LC_n values of 59 and 80 jjg/1 being reported in
two individual studies. The longnose killifish, (Fandulus similis), was the
only marine species tested for acute toxicity of acrolein; a 48-hour flow-
through LC^Q of 150 jug/1 was obtained. The eastern oyster, (Crassostrea
virginica), and adult brown shrimp, (Penacus aztecus), were the most sensi-
tive species tested an EC5Q value of 55 jug/1 based on 50% decrease in
shell growth of oysters and an EC,-- value of 100 based on loss of equi-
librium of brown shrimp (Butler, 1965). Adult barnacles were more 'resistant
in static assays with 48-hour LC5Q values of 1,600 and 2,100 jug/l' being
reported.
-------
B. Chronic Toxicity
In a chronic life cycle test with the freshwater fathead minnow,
Pimephales promelas, survival of newly hatched second generation fry was
reduced significantly at 42 but not 11 ug/1, leading to a chronic value of
21.8 ug/1 (Macek, et al. 1976). A comparable value of 24 jjg/1 was obtained
from reduced survival of three generations of Daphnia magna. Chronic data
for marine organisms was not available.
C. Plant Effects
Pertinent data relating the phytotoxitity of freshwater marine
plants could not be located in the available literature.
D. Residues
A bioconcentration factor of 344 was obtained for radio labeled
acrolein administered to bluegills, (Lepomis macrochivas). A biological
half-life greater than seven days was indicated (U.S. EPA. 1979).
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
Based on the use of subacute toxicological data for rats (no
observable effect level of 1.56 mg/kg body weight) and an uncertainty factor
of 1000, the U.S. EPA (1979) has derived a draft criterion of 6.50 jjg/1 for
acrolein in ambient water. This draft criterion level corresponds to the
calculated (U.S. EPA, 1979) acceptable daily intake of 109 ug.
»
The ACGIH (1977) time-weighted average TLV for acrolein is 0.1 ppm
(0.25 mg/m ). The same value is recommended by OSHA (39 FR 23540). This
-------
standard was designed to "minimize, but not. entirely prevent, irritation to
all exposed individuals" (ACGIH, 1974).
The FDA permits acrolein as a slime-control substance in the manu-
facture of paper and paperboard for usage in food packaging (27 FR 46) and
in the treatment of food starch (28 FR 2676) at not more than 0.6 percent
acrolein.
B. Aquatic
The draft criterion for protecting freshwater organisms is 1.2 pg/1
as a 24-hour average not to exceed 2.7 pg/1. For marine life, the draft
criterion has been proposed as 0.88 pg/1, not to exceed 2.0 ug/1.
-4JT-
sf
-------
ACROLEIN
REFERENCES
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and toxicology, in acrolein. John Wiley and Sons, Inc.,
New York.
Altshuller, A. P. 1978. Assessment of the contribution
of chemical species to the eye irritation potential of photo-
chemical smog. Jour. Air Pollut. Control Assoc. 28: 594.
Altshuller, A. R., and S. P. McPherson. 1963. Spectrophoto-
metric analysis of aldehydes in the Los Angeles atmosphere.
Jour. Air Pollut. Control Assoc. 13: 109.
American Conference of Governmental Industrial Hygienists.
1974. Documentation of the threshold limit value. 3rd ed.
American Conference of Governmental Industrial Hygienists.
1977. Threshold limit values for chemical substances in
workroom air.
Bellar, T. A., and J. E. Sigsby. 1970. Direct gas chromato-
graphic analysis of low molecular weight substituted organic
compounds in emissions. Environ. Sci. Technol. 4: 150.
Bignami, M., et al. 1977. Relationship between chemical
structure and mutagenic activity in some pesticides: The
use of Salmonella tyghj.murj.um and Aspergillus nidulans.
Mutat. Res'.' 46 : Z43T~
Bilimoria, M. H. 1975. Detection of mutagenic activity
of chemicals and tobacco smoke in bacterial system. Mutat.
Res. 31: 328.
Bouley, G., et al. 1976. Phenomena of adaptation in rats
continuously exposed to low concentrations of acrolein.
Ann. occup. Hyg. 19: 27.
Bowmer, K. H., and M. L. Higgins. 1976. Some aspects of
the persistence and fate o£ acrolein herbicide in water.
Arch. Environ. Contam. Toxicol. 5: 87.
Bowmer, K. H., and G. R. Sainty. 1977. Management of aqua-
tic plants with acrolein. Jour. Aquatic Plant Manage. 15:
40.
Boyd, E. N., et al. 1965. Measurement of monocarbonyl classes
in cocoa beans and chocolate liquor with special reference
to flavor. Jour. Food Sci. 30; 854.
-------
Boyland, E., and L. F. Chasseaud. 1967. Enzyme-catalyzed
conjugations of glutathione with unsaturated compounds.
Biochem. Jour. 104: 95.
Butler, P. A. 1965. Commercial fisheries investigations.
Effects of pesticides on fish and wildlife, 1964 research
findings Fish Wildl. Serv. U.S. Fish Wildl. Serv. Circ.
Egle, J. L., Jr. 1972." Retention of inhaled formaldehyde,
propionaldehyde, and acrolein in the dog. Arch. Environ.
Health 25: 119.
Ellenberger, J., and G. R. Mohn. 1976. Comparative mutageni-
city testing of cyclophosphamide and some of its metabolites.
Mutat. Res. 38: 120.
Ellenberger, J., and G. R. Mohn. 1977. Mutagenic activity
of major mammalian metabolites of cyclophosphamide toward
several genes of Escherichia coli. Jour. Toxicol. Enviorn.
Health 3: 63.7.
Epstein, S. S., et al. 1972. Detection of chemical mutagens
by the dominant lethal assay in the mouse. Toxicol. Appl.
Pharmacol. 23: 288.
Esterbauer, H., et al. 1975. Reaction of glutathione with
conjugated carbonyls. Z. Naturforsch. C: Biosci. 30c:
466.
Feron, V. J., and A. Kruysse. 1977. Effects of exposure
to acrolein vapor in hamsters simultaneously treated with
benzo (a)pyrene or diethylnitrosamine. Jour. Toxicol. Environ.
Health 3: 379.
Hess, L. B., et al. 1978. Acrolein and derivatives. In
Kirk-Othmer Encyclopedia of Chemical Technology. 3rd ed.
Interscience Publishers, New York.
Hoffman, D., et al. 1975. On the carcinogenicity of mari-
juana smoke. Recent Adv. Phytoc.hem. 9: 63.
Hopkins, D. M., and A. R. Hattrup. 1974. Field evaluation
of a method to detect acrolein in irrigation canals. U.S.
PB Rep. No. 234926/4GA. Natl. Tech. Inf. Serv.
Horton, A. D., and M. R. Guerin. 1974. Determination of
acetaldehydes and acrolein in the gas phase of cigarette
smoke using cryothermal gas chromatography. Tob. Sci. 18:
19.
•
Hrdlicka, J., and J. Kuca. 1965. The changes of carbonyl
compounds in the heat-processing of meat. Poultry Sci.
44:27.
-------
Izard, C. 1973. Recherches sur les effets mutagenes de
1' acroleine et des ses deux epoxydes: le glycidol et le
glycidal, sur Saecharomyces cerevisiae, C.R. Acad. Sci.
Ser. D. 276: 3(737"!
Izard, C., and C. Libermann. 1978. Acrolein. Mutat. Res.
47: 115.
Jaeger, R. J., and S. D. Murphy. 1973. Alterations of
barbiturate action following 1,1-dichloroethylene, corti-
costerone, or acrolein. Arch. Int. Pharmacodyn. Ther. 205:
281.
Kantemirova, A. E. 1975. Illness with temporary work dis-
ability in workers engaged in acrolein and methylmercaptopro-
pionaldehyde (MMP) production. Tr. Volgogr. Gos. Med. Inst.
26: 79. Chem. Abst. 88; 109868g.
Kaye, C. M. 1973. Biosynthesis of mercapturic acids from
allyl alcohol, allyl esters, and acrolein. Biochem. Jour.
134: 1093.
Kaye, C. M., and L. Young. 1972. Synthesis of mercapturic
acids from allyl compounds in the rat. Biochem. Jour. 127:
87.
Kishi, M., et al. 1975. Effects of inhalation of the vapor
from heated edible oil on the circulatory and respiratory
systems in rabbits. Shokuhin Eiseigaku Zasshi. 16: 318.
Low, E. S., et al. 1977. Correlated effects of cigarette
smoke components on alveolar macrophage adenosine triphos-
phatase activity and phagocytosis. Am. Rev. Respir. Dis.
115: 963.
Macek, K. J., et al. 1976. Toxicity of four pesticides
to water fleas and fathead minnows: Acute and chronic toxi-
city of acrolein, heptachlor, endosulfan, and tribluralin
to the water flea (Daphnia magna) and the fathead minnow
(Priinephales prgmelas'n EPA~1>UU/3-76-099 . U.S. Environ.
Prot. Agency.
Munsch, N., et al. 1973. Effects of acrolein on DNA syn-
thesis ir\ vitro. Fed. Eur. Biochem. Soc.' Lett. 30: 286.
Munsch, N., et al. 1974. jCri vitro binding of tritium labeled
acrolein to regenerating ra~lTve"r~DNA polymuase. Experi-
mentia 30: 1234.
National Cancer Institute. 1979. Personal communication
ftom Sharon Peeney.
Newell, G. w. 1958. Acute and subacute toxicity of acro-
lein. Stanford Res. Ins. SRI Project No. 5-868-2. Summar-
ized in Natl. Acad, Sci. 1977.
-------
Pattle, R. E., and H. Cullumbine. 1956. Toxicity of some
atmospheric pollutants. Brit. Med. Jour. 2: 913.
Pietruszko, R., et al. 1973. Comparison of substrate specifi-
city of alcohol dehydrogenases from human liver, horse liver,
and yeast towards saturated and 2-enoic alcohols and alde-
hydes. Arch. Biochem. Biophys. 159: 50.
Rapoport., I. A. 1948. Mutations under the influence of
unsaturated aldehydes. Dokl. Akad. Nauk. (U.S.S.R.), 61:
713. Summarized in Izard and Libermann, 1978.
Renzetti, N. A., and R. J. Bryan. 1961. Atmospheric samp-
ling for aldehydes and eye irritation in Los Angeles smog
- 1960. Jour. Air Pollut. Control Assoc. 11: 421.
Rosenthaler, L., and G. Vegezzi. 1955. Acrolein in alco-
holic liquors. Z. Lebensm.-Untersuch. u. - Forsch. 102:
117.
Schuck, E. A., and N. A. Renzetti. 1960. Eye irritants
formed .during photooxidation of hydro-carbons in the pre-
sence of oxides of nitrogen. Jour. Air Pollut. Control
Assoc. 10: 389.
Sim, V. M., and R. E. Pattle. 1957. Effect of possible
smog irritants on human subjects. Jour. Am. Med. Assoc.
165: 1908.
Sprince, H., et al. 1978. Ascorbic-acid and cysteine pro-
tection against aldehyde toxicants of cigarette smoke.
Fed. Proc. 37: 247.
Standen, A., ed. 1967. Kirk-Othmer Encyclopedia of Chemi-
cal Technology. Interscience Publishers, New York.
Szot, R. J., and S. D. Murphy. 1970. Phenobarbital and
dexamethasone inhibition of the adrenocortical response
of rats to toxic chemicals and other stresses. Toxicol.
Appl. Pharmacol. 17; 761.
Tillian, H. M., et al. 1976. Therapeutic effects of cys-
teine adducts of alpha, beta-unsaturated aldehydes on ehr-
lich ascites tumor of mice. Eur. Jour. Cancer 12: 989.
U.S. EPA. 1979. Ambient Water Quality Criteria: Acrolein.
(Draft)
Weast, R. C., ed. 1975, Handbook of chemistry and physics.
56th ed. CRC Press, Cleveland, Ohio. »
Weber-Tschopp, A., et al. 1976a. Air pollution and irri-
tation due to cigarette smoke. Soz.-Praeventivmed 21: 101.
-------
Weber-Tschopp, A., et al. 1976b. Objective and subjective
physiological effects of passive smoking; Int. Arch. Occup.
Environ. Health 37: 277.
Weber-Tschopp, A., et al. 1977. Experimental irritating
effects of acrolein on man. Int. Arch. Occup. Environ.
Health 40; 117.
Whitehouse, M. W., and F.W.J. Beck. 1975. Irritancy of
cyclophosphamide-derived aldehydes (acrolein, chloracetalde-
hyde) and their effect on lymphocyte distribution _in VJ.VQ:
Protective effect of thiols and bisulfite ions. Agents
Actions 5: 541.
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No. 7
Acrylonltrlle
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
acrylonitrile and has found sufficient evidence to indicate
that this compound is carcinogenic.
-------
ACRYLONITRILE
Summary
•
-------
ACRYLQNITRILE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Acrylonitrile (U.S. EPA, 1979).
Acrylonitrile (CH2=CHCN) is an explosive, flammable liquid having a
normal boiling point of 77CC and a vapor pressure of 80 mm Hg (20°C).
Currently, 1.6 billion pounds per year of acrylonitrile are manufactured in
the United States. The major use of acrylonitrile is in the manufacture of
copolymers for the production of acrylic and modacrylic fibers. Acryloni-
trile has been used as a fumigant; however, all U.S. registrations for this
use were voluntarily withdrawn as of August 8, 1978 (U.S. EPA, 1979).
II. EXPOSURE
A. Water
While no data on monitoring of water supplies for the presence of
acrylonitrile were found in the literature, potential problems may exist.
Possible sources of acrylonitrile in the aqueous environment are: (a) dump-
ing of chemical wastes, (b) leaching of wastes from industrial landfills,
(c) leaching of monomers from polymeric acrylonitrile, and (d) precipitation
from rain. Acrylonitrile is short-lived in the aqueous environment; a 10
ppm solution was completely degraded after 6 days in Mississippi River water
(Midwest Research Institute, 1977).
B. Food
There is no data on the levels of acrylonitrile in food. However,
acrylonitrile may contaminate food by leaching of the monomer from polyacry-
lonitrile containers (National Resources Defense Council, 1976). The U.S.
EPA (1979) has estimated the weighted average bioconcentration factor* for
-------
acrylonitrile to be 110 for the edible portions of fish and shellfish con-
sumed by Americans. This estimate is based on steady-state bioconcentration
studies in bluegills. "' '
C. Inhalation
NIOSH (1978) estimated that 125,000 workers are exposed to acrylo-
nitrile each year. Acrylonitrile may be liberated to the atmosphere via
industrial processes or by the burning of polyacrylonitrile fiber (Monsanto,
1973). Data could not be found in the available literature regarding the
concentrations of acrylonitrile in ambient air,
III. PHARMACQKINETICS
A. Absorption
When orally administered to rats, essentially all of the acryloni-
trile is absorbed (Young, et al. 1977).
B. Distribution
In rabbits, after administration of a 30 mg/kg dose, acrylonitrile
rapidly disappeared from the blood; only 1 mg/kg remained after 4 hours
(Hashimoto and Kanai, 1965). In rats the metabolites of acrylonitrile dis-
tributed to the stomach wall, erythrocytes, skin, and liver (Young, et al.
1977). .
C. Metabolism
Earlier reports (Giacosa, 1883; • Meurice, 1900) indicated that most
aliphatic nitriles are metabolized to cyanide which is then detoxified to
thiocyanate. A more recent report concluded that acrylonitrile exerts its
toxicity by the metabolic release of cyanide ion, and that the relative abi-
lity of various species to convert CN~ to SCN~ determined their suscep-
tibility to the toxic action of acrylonitrile (Brieger, et al. 1952). Other
facts, however, suggest that acrylonitrile toxicity is due in part to the
-------
acrylonitrile molecule itself or other unknown metabolite(s) rather than
just to the cyanide functional group (U.S., EPA, 1979). In a comprehensive
tracer study with rats Young, et al. (1977) found three uncharacterized
metabolites as well as C02 after acrylonitrile administration. Also, cya-
noethylated mercapturic acid conjugates have been detected after administra-
tion of acrylonitrile (U.S. EPA, 1979).
D. Excretion
Urinary excretion of thiocyanate after acrylonitrile administration
ranges from 4-33 percent of the administered dose --depending on the species
(U.S. EPA, 1979). Urinary excretion also depends on route of administration
(Gut, et al. 1975).
IV. EFFECTS
A. Carcinogenicity
In two studies rats received acrylonitrile in the drinking water at
concentrations of 0, 35, 100 and 300 mg/1, which is equivalent to daily dos-
ages of approximately 4, 10, 30 mg/kg body weight respectively, excess mam-
mary tumors and tumors of the ear canal and nervous system were noted (Mor-
ris, 1977; Quast, et al. 1977). Both the intermediate and the highest doses
produced increased tumor incidences. In rats administered acrylonitrile in
,olive oil by stomach tube at 5 mg/kg body weight 3 times per week for 52
weeks, a slight enhancement of the incidence of mammary tumors, forestomach
papillomas and acanthomas, skin carcinomas, and encephalic tumors has been
reported (Maltoni, et al. 1977). Also, exposure of rats by inhalation (40,
20, 10, and 5 ppm for 4 hours daily, 5 times/week) for 52 weeks caused in-
creases in tumor incidence (Maltoni, et al. 1977). It should be pointed out
that possible impurities found in the acrylonitrile used by various investi-
gators might determine the carcinogenic effect. The specific role of these
impurities has not yet been determined (U.S. EPA, 1979).
1
-57-
-------
Retrospective studies on workers in a textile fiber plant (Q'Berg,
1977) and on workers in the polymerization recovery and laboratory areas of
a B.F. Goodrich plant (Monson, 1977) have shown higher than expected inci-
dences of cancers of all sites in workers exposed to acrylonitrile. The
greatest increase was noted with lung cancer. It should be noted that these
workers were exposed to other chemicals in their working environment.
8. Mutagenicity
Acrylonitrile is a weak mutagen in Drosophila melanogaster (Benes
and Sram, 1969); although toxicity limited this testing. Milvy and Wolff
(1977) reported mutagenic activity for acrylonitrile in Salmonella typhimur-
jum with a mammalian liver-activating system. In Escherichia coli mutagenic
activity was observed without an activating system (Venitt, et al. 1977).
C. Teratogenicity
Studies in pregnant rats demonstrated that acrylonitrile adminis-
tered by gavage at 65 mg/kg/day caused fetal malformations (Murray, et al.
1976). These malformations included acaudea, short-tail, short trunk, miss-
ing vertebrae, and right-sided aortic arch. In a subsequent study, Murray,
et al. (1978) concluded that in pregnant rats exposed to 0, 40, or 80 ppm of
acrylonitrile by inhalation, teratogenic effects in the offspring were seen
at 80 ppm but not 40 ppm. Significant maternal toxicity was found at both
80 and 40 ppm, as well as in the previous study at 65 mg/kg/day.
D. Other Reproductive Effects
Pregnant rats receiving 500 ppm acrylonitrile in their drinking
water showed reduced pup survival, possibly due to a maternal toxicity
(Beliles and Mueller, 1977).
-------
E. Chronic Toxicity
Knoblock, et al. (1972) observed a perceptible change in peripheral
blood pattern, functional disorders in the respiratory and cardiovascular
systems, and the excretory system, as well as signs of neuronal lesions in
the central nervous system of rats and rabbits breathing acrylonitrile (50
mg/m air) for 6 months. Babanov, et al. (1972) reported that inhalation
of acrylonitrile vapor (0.495 mg/m , 5 hours/day, 6 days/week) for 6
months resulted in central nervous system disorders, increased erythrocyte
count, and decreased leukocyte count in rats. Workers exposed for long per-
iods of time to acrylonitrile have subjective complaints including headache,
fatigue, nausea- and weakness, as well as clinical symptoms of anemia, jaun-
dice, conjunctivitis and abnormal values of specific gravity of whole blood,
blood serum and cholinesterase values, urobilinogen, bilirubin, urinary pro-
tein and sugar (Sakarai and Kusimoto, 1972). In another study, functional
disorders of the central nervous system, cardiovascular and hemopoietic sys-
tems were noted (Shustov and Mavrina, 1975). Sakarai and Kasumoto (1972)
concluded that acrylonitrile exposures at levels of 5-20 ppm caused mild
liver injury and probably a cumulative general toxic effect.
F. Other Relevant Information
HCN and CO were found to enhance acrylonitrile toxicity in experi-
mental animals (Yamamoto, 1976) as well as in workers engaged in acryloni-
trile production (Ostrovskaya, et al. 1976).
V. Aquatic Toxicity
A. Acute Toxicity
The 96-hour LC^ values of fathead minnows (Pimephales promelas)
were 10,100 and 18,100 jjg/l for flow-through and static tests, respectively,
and 14,300 and 18,100 jjg/l for hard (380 mg/1) and soft (29 mg/1) waters,
-------
respectively (Henderson, et al. 1961). A reported 48-hour LCen for Daph-
nia magna is 7,550 ug/1 (U.S. EPA, 1978). The saltwater pinfish (Lagodon
rhomboides) has an observed 96-hour LCV- value of 24,500 ug/1 in a static
concentration unmeasured test (Daugherty and Garrett, 1951).
B. Chronic Toxicity
Daphnia maqna has been exposed for its life cycle and the results
indicate no adverse effects at concentrations as high as 3,600 jug/1 (U.S.
EPA, 1978). Henderson, et al. (1961) observed a 30-day LC5Q value of
2,600 pg/1 with Pimephales promelas (fathead minnows). No chronic test data
are available for saltwater species.
C. Plant Effects
Pertinent data could not be located in the available literature on
the sensitivity of plants to acrylonitrile.
D. Residues
In the only reported study, the bluegill (Lepomis macrochirus) was
exposed for 28 days and the determined whole body bioconcentration factor
was 48, with a half-life between 4-7 days (U.S. EPA, 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
The American Conference of Governmental Industrial Hygienists
threshold limit value (TLV) (ACGIH, 1974) for acrylonitrile is 20 ppm. In
»
January, 1978, the Occupational Safety and Health Administration (OSHA) an-
nounced an emergency temporary standard for acrylonitrile of 2 ppm averaged
-to-
-------
over an eight-hour period. Based on rat data (Norris, 1977; Quast, et al.
1977; Maltoni, et al. 1977), and using the "one-hit" model, the U.S. EPA
(1979) has estimated levels of acrylonitrile in ambient water which will re-
sult in specified risk levels of human cancer:
Exposure Assumptions Risk
(per day)
0
2 liters of drinking water
and consumption of 18.7
grams of fish and shellfish.
Consumption of fish and
shellfish only..
Levels and Corresponding Draft Criteria
10-7
0.008 x
0.016 x
10-4 ng/i
ID"6 ID"5
0.08 x 0.8 x
10-4 ng/i io-4 ng/1
0.16 x 1.6 x
B. Aquatic
For acrylonitrile, the draft criterion to protect freshwater aquat-
ic life is 130 ug/1 as a 24-hour average, and the concentration should not
exceed 300 jug/1 at any time. To protect saltwater species, the draft cri-
terion is 130 pg/1 as a 24-hour average, with the concentration not to exceed
290jug/l at any time (U.S. EPA, 1979).
-------
ACRYLONITRILE
REFERENCES
Babanov, G.P., et al. 1972. Adaptation of an organism
to acylonitrile at a low concentration factor in an indus-
trial environment. Toksikol. Gig. Prod. Neftekhim. 45:
58.
Beliles, R.P., and S. Mueller. 1977. Three-generation
reproduction study of rats receiving acrylonitrile in drink-
ing water. Acrylonitrile progress report second generation.
Submitted by Litton Bionetics, Inc. to the Manufacturing
Chemists Association. LBI Project No. 2660. November, 1977.
Benes, V., and R. Sram. 1969. Mutagenic activity of some
pesticides in Drosophila melanogaster. Ind. Med. Surg.
38: 442.
Brieger, et al. 1952. Acrylonitrile: Spectrophotometric
determination, acute toxicity and mechanism of action.
Arch. Indust. Hyg. Occup. Med. 6: 128.
Daugherty, F.M., Jr., and J.T. Garrett. 1951. Toxicity
levels of hydrocyanic acid and some industrial by-products.
Tex. Jour. Sci. 3: 391.
Giacosa, P. 1883. Toxicity of aliphatic nitriles. Hoppe-
Seyle 2: 95.
Hashimoto, K., and R. Kanai. 1965. Toxicology of acrylo-
nitrile: metabolism, mode of action, and therapy. Ind. Health
3: 30.
Henderson, C.,, et al. 1961. The effect of some organic
cyanides (nitriles) on fish. Eng. Bull. Ext. Ser. Purdue
Univ. No. 106: 130.
Knobloch, K., et al. 1972. Chronic toxicity of acryloni-
trile. Med. Pracy 23: 243.
Maltoni, C., et al. 1977. Carcinogenicity bioassays on
rats of acrylonitrile administered by inhalation and by
ingestion. La Medicina del Lavoro 68: 401.
Meurice, J. 1900. Intoxication and detoxification of dif-
ferent nitriles. Arch. Internat. de Pharmacodynamie et
de Therapie 7: 2.
Midwest Research Institute. 1977. Sampling and analysis ,
of selected toxic substances. Section V. Sampling and
analysis protocol for acrylonitrile. Progress Report No.
13, Oct. 1-31, 1977. EPA Contract No. 68-01-4115, MRI Pro-
ject NO. 4280-C (3) .
-------
Milvy, P., and M. Wolff. 1977. Mutagenic studies with
acrylonitrile. Mutation Res. 48: 271.
Monsanto Company. July 19, 1973. Environmental Impact
of Nitrile Barrier Containers, LOPAC: A case study. Monsanto
Co. St. Louis, Missouri.
Monson, R.R. November 21, 1977. Mortality and Cancer Mo-
bidity among B.F. Goodrich White Male Union Members who
ever worked in Departments 5570 through 5579. Report to
B.F. Goodrich Company and to the United Rubber Workers.
Federal Register No. 43FR45762 (see OSHA Dockit H-108, ex-
hibits 67 and 163}.
Murray, F.J., et al. 1976. Tertologic evaluation of acrylo-
nitrile monomer given to rats by gavage. Report from Toxi-
cology Research Lab., Dow Chem. v
Murray, F.J., et al. 1978. Teratologic evaluation of in-
haled acrylonitrile monomer in rats. Report of the Toxi-
cology Research Laboratory, Dow Chemical U.S.A. Midland,
Michigan. May 31, 1978.
National Resources Defense Council. 1976. Pop bottles:
The plastic generation—a study of the environmental and
health problems of plastic beverage bottles, p. 33.
NIOSH. 1978. A Recommended Standard for Occupational Expo-
sure to Acrylonitrile. DHEW (NIOSH) Publication No. 78-116,
U.S. Government Printing Office.
Norris, J.M. 1977. Status report on two-year study incor-
porating acrylonitrile in the drinking water of rats. Health
Environ. Res. The Dow Chemical Company.
O'Berg, M. 1977. Epidemiologic studies of workers exposed
to acrylonitrile: Preliminary results. E.I. Du Pont de
Nemours & Company.
Ostrovskaya, R.S., et al. 1976. Health status of workers
currently engaged in production of acrylonitrile. Gig.
T. Prof. Zabol. 6: 8.
Quast, J.F., et al. 1977. Toxicity of drinking water con-
taining acrylonitrile in rats: Results after 12 months.
Toxicology Res. Lab., Health and Environmental Res. Dow
Chemical U.S.A.
Sakarai, H., and M. Kusimoto. 1972. Epidemiologic Study
of Health Impairment Among AN Workers. Rodo Kagaku. 48:
273.
Shustov, V.Y., and E.A. Mavrina. 1975. Clinical picture
of chronic poisoning in the production of nitron. Gig. Tr.
Prof. Zabol 3: 27.
-A3-
-------
Threshold Limit Values. 1974. TLV's: Threshold Limit
Values for Chemical Substances and Physical Agents in the
Work Room Environment with Intended Changes for 1974. Am.
Conf. Govern. Ind. Hyg.
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. 1979. Acrylonifcrile: Ambient Water Quality
Criteria (Draft).
Venitt, S., et al. 1977. Mutagenicity of acrylonitrile
(cyanoethylane) in Escherichia coli. Mutation Res. 45:
283.
Yamamoto, K. 1976. Acute combined effects of HCN and CO,
with the use of combustion products from PAN (polyacrylo-
nitrile)—gauze mixtures. Z. Rechtsmed. 78: 303.
Young, J.D./.et al. 1977. The pharmacokinetic and metabolic
profile of C-acrylonitrile given to rats by three routes.
Report of the Toxicological Research Laboratory. Dow Chemi-
cal. Midland, Michigan.
-------
No. 8
Aldrin
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-uracy.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated
aldrin and has found sufficient evidence to indicate that
this compound is carcinogenic.
-C.7-
-------
ALDRIN .
Summary
Aldrin Is a man-made compound belonging to the group of cyclodiene in-
secticides. The chronic toxicity of low doses of aldrin include shortened
lifespan, liver changes, and teratogenic effects. The induction of hepato-
cellular carcinoma in both male and female mice from the administration of
aldrin leads to the conclusion that it is likely to be a human carcinogen.
Aldrin has not been found mutagenic in several te'st systems although it did
induce unscheduled DMA synthesis in human fibroblasts. The - World Health
Organization acceptable daily intake level for aldrin is 0.1 jjg/kg/day.
Aldrin is rapidly converted to dieldrin by a number of fresh and salt-
water species. The overall toxicity of aldrin is similar to dieldrin. The
96-hour LC5Q values for freshwater fish vary from 2.2 to 37 jug/1 with in-
vertebrates being one order of magnitude less sensitive. Both marine fish
and plants were susceptible to levels of aldrin corresponding to those of
freshwater fish.
If
-------
ALDRIN
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Aldrin and Dieldrin (U.S. EPA, 1979a).
Aldrin is a white crystalline substance with a melting point of
104°C. It is soluble in organic solvents. The chemical name for aldrin
is I,2,3,4,10,10-hexachloro-lf4,4a,5,8,8a-hexahydro-l,4,:5,8-exo-dimethano-
naphthalene. Aldrin is biologically altered in the environment to dieldrin,
a more stable and equally toxic form. For information concerning dieldrin
refer to the dieldrin hazard profile or the draft Ambient Water Quality Cri-
teria Document for Aldrin and Dieldrin (U.S. EPA, 1979a,b).
Aldrin was primarily used as a broad spectrum insecticide until 1974
when the U.S. EPA restricted its use to termite control by direct soil in-
jection, and non-food seed and plant treatment (U.S. EPA, 1979a). From 1966
to 1970 the use of aldrin in the United States dropped from 9.5 x 103 to
5.25 x 103 tons (U.S. EPA, 1979a). This decrease in use has been attri-
buted primarily to increased insect resistance to aldrin and to development
of substitute materials. Although the production of aldrin in the United
States is restricted, formulated products containing aldrin are imported
from Europe (U.S. EPA, 1979a).
II. EXPOSURE
A. Water
Aldrin has been applied to vast areas of agricultural 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 (U.S. EPA,
1979a). Levels of aldrin, ranging from 15 to 18 ng/1 or as high as 40"? ng/1
-------
have been found in waters of the United States (U.S. EPA, 1976; Leichten-
berg, et al. 1970). The half-life of aldrin in water one meter in depth has
been estimated to be 10.1 days (MacKay and Wolkoff, 1973).
B. Food
The estimated daily dietary intake of aldrin in 16 to 19 year old
males was estimated to be 0.001 mg in 1965 and only a trace amount in 1970
(Natl. Acad. Sci., 1975).
No direct measured bioconcentration factor for aldrin can be ob-
tained because it is rapidly converted to dieldrin by aquatic organisms
(U.S. EPA, 1979a). The U.S. EPA (1979a) has estimated the weighted average
bioconcentration factor of aldrin at 32. This estimate is based on the
octanol/water partition coefficient for aldrin.
C. Inhalation
Aldrin enters the air through various mechanisms such as spraying,
wind action, water evaporation, and adhesion to particles (U.S. EPA, 1979a).
Ambient air levels of 8 ng/m of aldrin have been reported .(Stanley, et
al. 1971).
D. Dermal
Dermal exposure to aldrin is limited to workers employed during
its manufacture and use as a pesticide. Wolfe, et al. (1972) reported that
exposure in workers is mainly through dermal absorption rather than inhala-
tion. The ban on the manufacture of aldrin in the United States has greatly
reduced the risk of exposure.
•III. PHARMACOKINETICS
A. Absorption
»
Pertinent data could not be located in the available literature
concerning the absorption of aldrin (U.S. EPA, 1979a).
-i
-70-
-------
B. Distribution
The distribution of aldrin in humans or animals has not been ex-
tensively studied because aldrin is readily converted to dieldrin in vivo
via epoxidation (U.S. EPA, 1979a). For example, the blood plasma levels of
aldrin were lower than the corresponding blood plasma levels of dieldrin in
six workers just after chronic exposure to aldrin for five weeks (Mick, et
al. 1971).
C. Metabolism
The epoxidation of aldrin to dieldrin.. has been reported in many
organisms including man (U.S. EPA, 1979a). The reaction is NADPH-dependent
and the enzymes are heat-labile (Wong and Terriere, 1965). The metabolic
products of aldrin include dieldrin, as well as aldrin diol, and polar meta-
bolites excreted in the urine and feces (U.S. EPA, 1979a).
D. Excretion
Aldrin is excreted mainly in the feces and to some extent in the
urine in the form of several polar metabolites (U.S. EPA, 1979a). Ludwig,
et al. (1964) reported nine times as much radioactivity in the feces as in
the urine of rats chronically administered 14C-aldrin. A saturation level
was reached in these animals and concentrations of radioactivity in the body
decreased rapidly when feeding was terminated.
Specific values for the half-life of aldrin in humans were not
found in the available literature. However, in humans exposed to aldrin
and/or dieldrin the half-life of dieldrin in the blood was estimated to be
266 days (Jager, 1970). In another study with 12 volunteers ingesting vari-
ous doses of dieldrin,. Hunter, et al. (1969) estimated the average dieldrin
half-life to be 369 days.
-71-
-------
IV. EFFECTS
A. Carcinogenicity
Aldrin has induced liver tumors in males and females in various
strains of mice according to reports of four separate feeding studies (Davis
and Fitzhugh, 1962; Davis, 1965; 43 FR 2450; Song and Harville, 1964). Ac-
cording to reports of five studies in two different strains of rats, aldrin
failed to induce a statistically significant carcinogenic response at all
but one site (Deicnmann, et al. 1967, 1970; Fitzhugh, et al. 1964; Cleve-
land, 1966; 43 FR 2450).
The only information concerning the carcinogenic potential of
aldrin in man is an occupational study by Versteeg and Jager (1973). The
workers had been employed in a plant producing aldrin and dieldrin with a
mean exposure time of 6.6 years. An average time of 7.4 years had elapsed
since the end of exposure. No permanent adverse effects including cancer
were observed.
B. Mutagenicity
Aldrin was found not to be mutagenic in two bacterial assays (S.
typhimurium and E^_ coli) with metabolic activation (Shirasu, et al. 1977).
Aldrin did, however, produce unscheduled DNA synthesis in human fibroblasts
with and without metabolic activation (Ahmed, et al. 1977).
C. Teratogenicity
Aldrin administered in single oral doses to pregnant hamsters
caused significant increases in hamster fetal death and increased fetal ano-
malies (i.e., open eye, webbed foot, cleft palate, and others). When a sim-
ilar study was done in mice at lower doses, teratogenic effects were also
f
observed, although these effects were less pronounced (Qttolengni, et al.
1974).
-------
D. Other Reproductive Effects
Deichmann (1972) reported that aldrin and dieldrin (25 mg/kg diet)
fed to mice for six generations affected fertility, gestation, viability,
lactation and survival of the young.
E. Chronic Toxicity
The other effects produced by chronic administration of aldrin to
mice, rats, and dogs include shortened lifespan, increased liver to body
weight ratios, various changes in liver histology, and the induction of
hepatic enzymes (U.S. EPA, 1979a).
F. Other Relevant Information
Since aldrin and dieldrin are metabolized by way of mixed function
oxidase (MFO), any inducer or inhibitor of the MFO enzymes should affect the
metabolism of aldrin and dieldrin (U.S. EPA, 1979a).
When aldrin is administered with DDT, or after a plateau has been
reached in dogs with chronic DDT feeding, the retention of DDT by the blood
and fat increases considerably (Deichmann, et al. 1969). Clark and Krieger
(1976) found that tissue accumulation of C-aldrin was significantly in-
creased when an inhibitor of the epoxidation of aldrin to dieldrin was admi-
14
nistered prior to C-aldrin.
V. AQUATIC TOXICITY •
A. Acute Toxicity
Aldrin is rapidly converted to dieldrin- in the environment. How-
ever, a number of acute studies haved been done with aldrin, although the
test concentrations have not been measured after the bioassays. Reported
96-hour static LC5Q values are as follows: bluegill (Lepomis macrochirus)
4.6 to 15 JJQ/I (Henderson, et al. 1959; Macek, et al. 1969); rainbow'trout
(Salmo gairdneri) 2.2 to 17.7 jug/1 (Macek, et al. 1969; Katz, 1961); and
-73-
-------
fathead minnows (Pimephales promelas) 32 and 37 jjg/1 (Henderson, et al.
1959). Acute toxicity varies greatly in freshwater invertebrates. In bio-
assays in which the aldrin concentrations were not measured, the observed
48-hour LC5- value for Daphnia pulex was 28/jg/1 (Sanders and Cope, 1966),
and the observed 96-hour LCv- values ranged from 4,300 to 38,500 jug/1 for
scud, Gammarus spp. (Sanders, 1969, 1972; Gaufin, et al. 1965).
In flow-through exposures to aldrin, the 48 and 96-hour LC-n
values for six saltwater fish species ranged from 2.0 to 7.2 pg/1. Inverte-
brate LC5Q values ranged from 0.37 to 33.0 jjg/1 (U;-S. EPA, 1979a).
8. Chronic Toxicity
No entire cycle or embryo-larval tests have been reported for any
fresh or saltwater species (U.S. EPA, 1979a).
C. Plant Effects
An aldrin concentration of 10,000 pg/1 reduced the population
growth in 12 days for water meal, Wolffia papulifera (Worthley and Schott,
1971). The productivity of a phytoplankton community was reduced 85 percent
after four hour exposure to 1,000 jug/1 aldrin (Butler, 1963).
D. Residues
No freshwater or saltwater residue studies have been reported for
aldrin (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.
-74-
-------
A. Human
The current exposure level for aldrin set by the Occupational
Safety and Health Administration is a time-weighted average of 250 jjg/m
for skin absorption (37 FR 22139). In 1969, the U.S. Public Health Service
Advisory Committee recommended that the drinking water standard for aldrin
be 17 ;jg/l (Mrak, 1969). The U.N. Food and Agricultural Organization/World
Health Organization acceptable daily intake for aldrin is 0.1 ^ig/kg/day
(Mrak, 1969).
The carcinogenicity data of the National Cancer Institute (1976)
(43 FR 2450) were used to calculate the draft water quality criterion for
aldrin which keeps the lifetime cancer risk for humans below 10. The
concentration for aldrin is 4.6 x 10"2 ng/1 (U.S. EPA, 1979a).
B. Aquatic
Draft criterion has not been proposed directly for aldrin because
of its rapid conversion to dieldrin (U.S. EPA, I979a).
-73T-
-------
ALDRIN
REFERENCES
Ahmed, F.E., et al. 1977. Pesticide-induced DNA damage and its repair in
cultured human cells. Mutat. Res. 42: 161.
Butler, P.A. 1963. Commercial fisheries investigations. In: Pesticide and
wildlife studies; A review of Fish and Wildlife Service investigations
during 1961 and 1962. U.S. Fish Wildl, Serv. Circ. 167: 11.
Clark, C.R and R.I. Krieger. 1976. Beta-diethylaminoethyldiphenyl-
propylacetate (SKF 525-A) enhancement of tissue accumulation of aldrin in
mice. Toxicol. Appl. Pharmacol. 38: 315.
Cleveland, F.P. 1966. A summary of work on aldrin and dieldrin toxicity at
the Kettering Laboratory. Arch. Environ. Health. ,13: 195.
Davis, K.J., 1965. Pathology report on mice for aldrin, dieldrin, hepta-
chlor, or heptachlor epoxide for two years. Internal Memorandum to Dr. A.J.
Lehman. U.S. Food Drug Admin.
Davis, K.J. and O.G. Fitzhugh. 1962. Tumorigenic potential of aldrin and
dieldrin for mice. Toxicol. Appl. Pharmacol. 4: 187.
Deichmann, W.B. 1972. Toxicology of DDT and related chlorinated hydro-
carbon pesticides. Jour. Occup. Med. 14: 285.
Deichmann, W.B., et al. 1967. Synergism among oral carcinogens in the
simultaneous feeding of four tumorigens to rats. Toxicol. Appl. Pharmacol.
11: 88.
Deichmann, W.B., et al. 1969. Retention of dieldrin and DDT in the tissues
of dogs fed aldrin and DDT individually and as a micture. Toxicol. Appl.
Pharmacol. 14: 205.
Deichmann, W.B., et al. 1970. Tumorigenicity of aldrin, dieldrin and en-
drin in the albino rat. Ind. Med. Surg. 39: 426.
Fitzhugh, O.G., et al. 1964. Chronic oral toxicity of aldrin and dieldrin
in rats and dogs. Food Cosmet. Toxicol. 2: 551.
Gaufin, A.R, et al. 1965. The toxicity of ten organic insecticides to var-
ious aquatic invertebrates. Water Sewage Works 12: 276.
Henderson, C.t et al. 1959. Relative toxicity of ten chlorinated hydro-
carbon 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 epidemio-
logical and toxicological study of long-term occupational exposure.
Elsevier Publishing Co. Amsterdam.
-------
Katz, M. 1961. Acute toxicity of some organic insecticides to three
species of salmonids and to the threespine stickleback, Trans. Am. Fish.
Soc. 90: 264.
Leichtenberg, 3.3., et al. 1970. Pesticides in surface waters in the
United States - A five-year summary, 1964-1968. Pestic. Monitor. Jour.
4: 71.
Ludwig, G., et al. 1964. Excretion and distribution of aldrin-l^C and
its metabolites after oral administration for a long period of time. Life
Sci. 3: 123.
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.
i_
MacKay, D. and A.W. Wolkoff. 1973. Rate of evaporation of low-solubility
contaminants from water bodies to atmosphere. Environ. Sci. Technol.
7: 611.
Mick, D.L., et al. 1971. Aldin and dieldrin in human blood components.
Arch. Environ. Health 23: 177.
Mrak, E.M. 1969. Report of the Secretary's commission on pesticides and
their relationship to environment health. U.S. Dept. Health, Edu. Welfare,
Washington, O.C.
National Academy of Sciences, National Research Council. 1975. Vol. 1 Pest
control: An assessment of present and alternative technologies. Contem-
porary pest control practices and prospects. Natl. Acad. Sci. Washington,
D.C.
Ottolenghi, A.O., et al. 1974. Teratogenic effects of aldrin, dieldrin and
endrin in hamsters and mice. Teratology 9: 11.
Sanders, H.O. 1969. Toxicity of pesticides to the crustacean, Gammarus
Lacustris. Bur. Sport Fish. Wildl. Tech. Pap. No. 25.
Sanders, H.O. 1972. Toxicity of some insecticides to four species of mala-
costracan crustaceans. Bur. Sport Fish. Wildl. Tech. Pap. No. 66.
Sanders, H.O. and O.B. Cope. 1966. Toxicities of several pesticides to two
species of cladocerans. Trans. Am. Fish. Soc. 95: 165.
Shirasu, Y., et al. 1977. Mutagenicity screening on pesticides and modifi-
cation products: A basis of carcinogenicity evaluation. Page 267 in H.H.
Hiatt, et al. (eds.). Origins of Human Cancer. Cold Spring Harbor Lab. New
York.
»
Song, J. and W.E. Harville. 1964. The carcinogenicity of aldrin and diel-
drin on mouse and rat liver. Fed. Proc. 23: 336.
-77-
-------
Stanley, C.W., et al. 1971. Measurement of atmospheric levels of pesti-
cides. Environ. Sci. Technol. 5; 430.
U.S. EPA. 1976. National interim primary drinking water regulations. U.S.
Environ. Prot. Agency. Publ. No. 570/9-76-003.
U.S. EPA. 1979a. Aldrin/Dieldrin Ambient Water Quality Criteria Document.
Washington, D.C. (Draft).
U.S. EPA. 1979b. Environmental Criteria and Assessment Office. Dieldrin:
Hazard Profile. (Draft).
Versteeg, J.P.J. and K.W. Jager. 1973. Long-term occupational exposure to
the insecticides aldrin, dieldrin, sndrin, and telodrin, Br. Jour. Ind.
Med. 30: 201.
Wolfe, H.R., et al. 1972. Exposure of spraymen to pesticides. Arch.
Environ. Health. 25: 29.
Wong, D.T. and L.C. Terriere. 1965. Epoxidation of aldrin, isodrin, and
heptachlor by rat liver microsomes. Biochem. Pharmacol. 14: 375.
Worthley, E.G. and C.D. Schott. 1971. The comparative effects of CS and
various pollutants on freshwater phytoplankton colonies of Wolffia
papulifera Thompson. Dep. Army. Edgewood Arsenal Biomed. Lab. Task
IW662710-AD6302.
-------
No. 9
Allyl 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 £rom 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.
-------
AULYL ALCOHOL
Summary
Allyl alcohol is a severe irritant to the mucous membranes at high con-
centrations. Hepatotoxicity has been seen after oral and inhalation
exposures, however, results indicate that this effect may not be
cumulative. Allyl alcohol is also absorbed percutaneously.
Information on the carcinogenic, mutagenic, teratogenic or other repro-
ductive effects of allyl alcohol was not found in the--available literature.
Data concerning the effects of allyl alcohol to aquatic organisms were
not found in the available literature.
-S/-
-------
I. INTRODUCTION
This profile is based on computerized searches of Toxline, Biosis
and Chemical Abstracts, and a review of other available appropriate
information sources as available.
Allyl alcohol (molecular weight-58.08) is a limpid liquid with
pungent odor. It is soluble in water, alcohol and ether, has a melting
point of -50°c and a boiling point of 96-97°C (Sax, 1979).
The major uses of allyl alcohol are in the manufacture of allyl
compounds, war gas, resins, and plasticizers (Windfnolz, 1976). Sixty kt.
are used in this country per year, of which 50 kt. are used to manufacture
glycerol (Kirk and Othmer, 1963).
After several years of storage, allyl alcohol polymerizes into a
substance that is soluble in chloroform but not water. When treated with
ether this substance becomes brittle (Windholz, 1976).
II. EXPOSURE
Pertinent data were not found in the available literature on air
or water exposure.
Esters of allyl alcohol are used as food flavorings. Natural de-
rivatives of allyl alcohol are widely distributed in vegetable material
(Lake, et al. 1978).
III. PHARMACOKINETICS
A. Absorption and Distribution
Pertinent data were not found in the available literature.
B. Metabolism
It has been suggested that allyl alcohol is completely metabolized
*
and that acrolein might be an intermediate metabolite (Browning,-1965). The
rate of metabolism in rats was found to be about 23 mg/kg/hr. during con-
stant intravenous infusion (Carpanini, et al. 1978).
-------
C. Excretion
Allyl alcohol was not found in the urine of animals that had been
dosed subcutaneously or intravenously with the compound (Browning, 1965).
Other pertinent data were not found in the available literature.
IV. EFFECTS
A. Carcinogenicity, Mutagenicity, Teratogenicity, and Reproductive
Effects
Information on the carcinogenic effects of allyl alcohol was not
found in the available literature.
B. Chronic Toxicity
Lake, et al. (1978) administered allyl alcohol to rats by gastric
intubation. The rats were dosed daily for 1, 10, or 28 days. Liver homo-
genates from treated animals were analyzed for enzyme activity. Adminis-
tration for one day produced marked periportal necrosis, but repeated ad-
ministration for 10 or 28 days did not seem to increase the damage.
Allyl alcohol administration in the drinking water at a dose of 72
mg/kg/day caused weight loss, transient pulmonary rales, crustiness of the
eyelids, and local areas of liver necrosis (Browning, 1965).
Rats exposed to 40, 60, or 100 ppm of allyl alcohol by inhalation
showed signs of acute mucous membrane irritation, such as gasping and nasal
discharge. At the 100 ppm dose, the animals died after 10 exposures
(Browning, 1965). No gross toxicity was seen at 5 or 10 ppm, 5 days a week
for 13 months in rats, rabbits, guinea pigs, and dogs. However, mild
reversible degenerative changes in the liver and kidney were seen at the
seven ppm dose. A dose of 50 ppm was lethal to rats after 30 days
(Torkelson, et al. 1959).
-------
Carpanini, et al. (1978) gave rats doses of allyl alcohol 50, 100,
200, or 800 ppm im the drinking water for 15 weeks. Weight loss was seen in
males given 100, 200, or 800 ppm and females given 800 ppm. Food
consumption values were lower than the controls in males at 200 ppm and 800
ppm and females at 800 ppm. A dose-related decrease in water consumption
was seen in all treated animals. Minor changes were seen in the liver,
kidneys, and lungs of both treated and control groups upon histological
examination.
C. Acute Toxicity ^
Oral l-D^g's of allyl alcohol have been found to be
64-100 mg/kg for rats, 96-139 mg/kg for mice, and 52-71 mg/kg for rabbits;
43 mg/kg was lethal to dogs. Intraperitoneal LD5g's were 42 mg/kg for
rats and 60 mg/kg for mice. In rabbits an LD5n Of 53-89 mg/kg was found
by percutan- eous absorption (Carpanini, et al. 1978). Inhalation of 1000
ppm was lethal to rabbits and monkeys after 3 to 4 hours. Erythema of the
conjunctiva and swelling of the cornea are seen in the eye after exposure to
allyl alcohol, however, no permanent damage was noted. Application to the
skin caused only mild erythema. Intravenous injection produced a drop in
blood pressure. Injection of 40 minims in a 20 percent saline solution
caused fluctuations in the blood pressure, of rabbits resulting in violent
convulsions. Vomiting, diarrhea, convulsions, apathy, ataxia, lacrimation
and coma are seen after oral administration. Few cases of serious injury
due to inhalation have been reported, however, because concentrations that
would cause severe damage in a short period of time are painful to the eyes
and nose. Five ppm are detectable by irritation and 2 ppm by odor
#
(Browning, 1965).
Moderate air contamination has been found to cause lacrimation,
pain around the eyes and blurred vision in man lasting up to 48 hours
(Carpanini, et al. 1978).
-------
0. Other Relevant Information
Allyl alcohol has an unusual effect on the central nervous system
of mice and rats. The effect is seen as apathy, unwillingness to move,
anxiety, and no interest in escaping. It is apparently different from nar-
cosis seen with other agents (Dunlap, et al. 1958).
V. AQUATIC TOXICITY
Pertinent data were not found in the available literature.
VI. EXISTING GUIDELINES
The recommended maximum atmospheric concentration (8 hours) is 2
ppm (Indust. Hyg. Assoc., 1963).
-------
REFERENCES
Qrownlng, E.C. 1965. Toxicity and Metabolism of Industrial Solvents.
Elsevier Publishing Co., Amsterdam, p. 739.
Carpanini, F.M.B., et al. 1978. Short-term toxicity of allyl alcohol in
rats. Toxicol. 9: 29.
Dunlap, M.K., et al. 1958. The toxicity of allyl alcohol. A.M.A. Archives
of indust. Health. 18: 303.
Industrial Hygiene Association. 1963. Hygienic Guide Series: Allyl
Alcohol. Indust. Hyg. Assoc. Jour. 24: 636.
Lake, B.C., et al. 1978. The effect of repeated administration on allyl
alcohol-induced hepatotoxicity in the rat. Biochem. Soc. Trans. 6: 145.
Sax, N.I. 1979. Dangerous Properties of Industrial Materials. 5th ed.
Von Nostrand Reinhold Co., New York.
Torkelson, T.R., et al. 1959. Vapor toxicity of allyl alcohol as deter-
mined on laboratory animals. Indust. Hyg. Assoc. Jour. 20: 224.
Verschueren, K. 1977. Handbook of Environmental Data on Organic Chem-
icals. Von Nostrand Reinhold Co., New York.
Windholz, M. (ed.) 1976. Merck Index. 9th ed. Merck and Co., Inc.,
Rahway, New Jersey.
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No. 10
Antimony
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.
-------
ANTIMONY
Summary
The adverse health effects most commonly associated with exposure to
antimony are pulmonary, cardiovascular, dermal, and certain effects on
reproduction, development, arid longevity. Cardiovascular changes have
been well-established with exposure to antimony and probably represent
^,
the most serious threat to human health. Antimony has not been assoc-
iated with carcinogenic effects. The lowest observed effect level for
antimony in the drinking water of rats was 5 ppm. A draft criterion of
145 jug/1 has been recommended for antimony in water based on an accep-
table daily intake of antimony from water, fish, and shellfish for man of
294 jug.
Antimony is highly toxic to aquatic organisms at a concentration
ranging from 19 mg/1 to 530 mg/1. Chronic values for antimony in fresh-
water organisms range from 0.8 mg/1 to 5.4 mg/1.
if
-------
ANTIMONY
I. INTRODUCTION
This profile is based primarily on the Ambient Water Quality Criteria
Document for Antimony (U.S. EPA, 1979). The health hazards of antimony
and its compounds have also been recently reviewed by the National Ins-
titute for Occupational Safety and Health (NIOSH, 1978).
Antimony (Sb; molecular weight 121.8) is a silvery, brittle, solid
belonging to group VB of the periodic table and lies between arsenic and
bismuth. It is classified as both a metal and a metalloid, and its prin-
cipal oxidation states are +3 and +5. Antimony has a boiling point of
1366°C and a melting point of 636°C. Most inorganic compounds of an-
timony are either only slightly water soluble or decompose in aqueous
media.
Antimony reacts with both sulfur and chlorine to form the tri-and
pentavalent sulfides and chlorides. Oxidation to antimony trioxide
(stibine), the major commercial oxide of antimony, is achieved under
controlled conditions.
Consumption of antimony in the United States is on the order of
40,000 metric tons per year (Callaway, 1969), of which half is obtained
from recycled scrap and the balance mainly imported. Use of antimony in
the United States is directed chiefly to the manufacture of ammunition,
storage batteries, matches and fireworks, and in the fire-proofing of
textiles.
-------
II. EXPOSURE
A. Water
Schroeder (1966) compiled data from surveys of municipal water
supplies in 94 cities and reported that levels averaged less than 0.2
ug/1 in finished water. In a related study, Schroeder and Kraemer (1974)
noted that tap water levels of antimony can be elevated in soft water
supplies due to leaching from plumbing.
B. Food
Because of the wide range of antimony levels in various types of
foods, it is not possible to accurately estimate an average dietary in-
take. Tanner -and Friedman (1977) concluded that dietary intake of
antimony is negligible, based upon trace metal food monitoring data from
the U.S. Food and Drug Administration. However, in earlier studies, cal-
culated average dietary intakes were reported at 100 ug per day for man
(Schroeder, 1970) and in the range of 0.25 to 1.28 mg per day for insti-
tutionalized children (Murthy, et al. 1971). In one study on antimony
levels in Italian diets a mean daily value of several micrograms was re-
ported (Clemente, 1976).
C. Inhalation
Antimony is not- generally found in ambient air at measurable
concentrations. National Air Sampling Network data for 1966 showed pos-
sibly significant levels at only four urban stations (0.042 to 0.085
jug/m3) (Schroeder, 1970; Woolrich, 1973).
D. Other Routes
The total body burden of antimony arising from all environmental
media is apparently very small relative to other trace metals (i.e.',
lead, mercury, cadmium) in the environment. Clemente (1976) published
-------
limited data on fecal and urinary levels of antimony in selected Italian '
populations and concluded that daily intakes were less than 2.0 ug/day.
In addition, data on the bioconcentration potential of antimony in fish
(U.S. EPA, 1978) indicate that no bioaccumulation is likely to occur.
The U.S. EPA (1979) has calculated the weighted average bioconcentration
factor (BCF) for antimony to be 1.4 for the edible portions of fish and
shellfish consumed by Americans. This estimate was based on 25-day bio-
concentration studies in bluegill.
III.PHARMACOKINETICS
Absorption of antimony in man and animals is mainly via the respir-
atory and gastro-intestinal tracts. The extent of absorption is dependent
on factors such as solubility, particle size, and chemical forms
(Felicetti, et al. 1974a; 1974b). Absorption via the GI tract is of the
order of several percent with antimony trioxide, a relatively insoluble
compound , and presumably would be much greater with soluble antimonials.
Blood is the main carrier for antimony, the extent of partition
between blood compartments depending on the valence state of the element
and the animal species studied (Felicetti, et al. 1974a). The rodent ex-
clusively tends to concentrate trivalent antimony for long periods in the
erythrocyte (Ojuric, et al. 1962). Whatever the species, it can gener-
ally be said that pentavalent antimony is borne by plasma and trivalent
antimony in the erythrocyte. Clearance of antimony from blood to tissues
is relatively rapid, and this is especially true in the case of paren-
teral administration and the use of pentavalent antimony (Casals, 1972;
Abdalla and Saif, 1962; El-Bassouri, et al. 1963).
The tissue distribution and subsequent excretion of antimony is a
function of the valence state.
-------
In animals, trivalent antimony aerosols lead to highest levels in the
lung, skeleton, liver, pelt, and thyroid while pentavalent aerosols show
a similar distribution, with the exception of slower uptake by the liver
(Felicetti, et al. 1974a; 1974b; Thomas, et al. 1973).
Parenteral administration to animals shows trivalent antimony accumu-
lating in the liver and kidney as well as in pelt and thyroid (Molkhia
and Smith, 1969; Waltz, et al. 1965).
In man, non-occupational or non-therapeutic exposure shows very low
antimony levels in various tissues with little--evidence of accumulation
(Abdalla and Saif, 1962). Chemotherapeutic use leads to highest accumu-
lation in liver, thyroid, and heart for trivalent antimony.
The biological half-life of antimony in man and animals is a function
of route of exposure, chemical form, and oxidation state. The rat
appears to be unique in demonstrating a long biological half-time owing
to antimony accumulation in the erythrocyte. In other species, including
man, moderate half-times of the order of days have been demonstrated.
While most soft tissues do not appear to accumulate antimony, the skin
does show accumulation, perhaps because of its high content of sulfhydryl
groups. With respect to excretion, injection of trivalent antimony leads
mainly to urinary excretion in guinea pigs and dogs, and mainly fecal
clearance in hamsters, mice and rats'.
Pentavalent antimony is mainly excreted via the kidney in most
species owing to its higher levels in plasma.
Unexposed humans excrete less than 1.0 jug antimony daily via urine,
while occupational or clinical exposure may result in markedly increased
»
amounts.
X
-IS-
-------
IV. EFFECTS
A. Carcinogenic!ty
Antimony has not been tested for carcinogenic activity using an
appropriately designed chronic bioassay protocol. However, Shroeder
(1970) indicated that the chronic administration of antimony at 5 ppm in
the drinking water of rats, had no apparent tumorigenic effect. However,
the shortened life span of treated animals (average 106 to 107 days less
than controls) limits the usefulness of these data. Similar results were
also observed in a study with mice chronically exposed to antimony at 5
ppm in the drinking water (Kanisawa and Schroeder, 1969).
A single-epidemiologic investigation has been conducted into the
role of antimony in the development of occupational lung cancer (Davies,
1973). This retrospective study, which was limited in scope, provided no
definitive information to support the possible role of antimony in lung
cancer development.
B. Mutagenicity
Antimony has not been tested for activity in standard muta-
genicity bioassays.
C. Teratogenicity
Little information is available concerning possible teratogenic
effects of antimony. In one study, Casals (1972) observed no effects,
i.e., no fetal abnormalities, following administration of a solution of
antimony dextran glycoside containing 125 or 250 mg Sb/kg to pregnant
rats on days 8 to 15 of gestation.
0. Other Reproductive Effects
Aiello (1955) observed a higher rate of premature deliveries*
among female workers engaged in antimony smelting and processing. In
-------
addition, dysmenorrhea was frequently reported among women workers.
Similarly, Belyaeva (1967) reported that a greater incidence of gyneco-
logical disorders was found among antimony smelter workers than in a con-
trol group (77.5 percent vs. 56 percent; significance unknown). Spon-
taneous late abortions occurred in 12 percent of the exposed females com-
pared to 4.1 percent among controls. Average urine levels of antimony
for exposed workers, however, were extremely high, ranging from 2.1 to
2.9 mg/100 ml. Antimony was also found in breast milk (3.3+ 2 mg/10),
placental tissue (3.2 to 12.6 mg/100 mg), amhiotic fluid (6.2 to 2.8
mg/100 mg), and umbilical cord blood (6.3 + 3 mg/100 ml).
In studies with rats exposed either to antimony dust (50 mg/kg,
i.p.) or to antimony trioxide dust (250 mg/m , 4 hours per day for 1.5
to 2 months), Belyaeva (1967) reported increased reproductive failure,
fewer offspring, and damage to the reproductive tissues (ovary and
uterus).
E. Chronic Toxicity
The toxic effects of exposure to antimony have been repeatedly
observed in both humans and experimental rodents. Pulmonary, cardio-
vascular, dermal, and certain effects on reproduction, development, and
longevity are among the health effects most commonly associated with an-
timony exposure.
Cardiovascular changes have been well established following ex-
posure to antimony and probably represent the most serious human health
effects demonstrated thus far (U.S. EPA, 1979). Air concentrations of
-------
antimony trisulfide exceeding 3 mg/cu m were associated with the induc-
tion of altered ECG patterns and some deaths attributed to myocardial
damage among certain antimony workers {Brieger, et al. 1954). Also, in
parallel studies on animals, Brieger and coworkers (1954) observed ECG
alterations in rats and rabbits exposed to antimony in air at levels of
3.1 to 5.6 mg/fn , 7 hours/day, 5 days/week for at least 6 weeks.
Gross and coworkers (1955) presented evidence for growth retardation
occurring when rats were chronically fed diets containing two percent
antimony trioxide. Other investigators (Schroeder, et al. 1970; Kanisawa
and Schroeder, 1969) reported that oral exposure to 5 ppm of antimony in
drinking water had no effect on the rate of growth of either rats or
mice. However, the 5 ppm exposure level was effective in producing
slight but significant lifespan shortening in both rats and mice, and
altered blood chemistries in exposed rats. Therefore, the 5ppm exposure
level has been considered the "lowest observed effect level" in animals
that likely approximates the "no effect" level for antimony-induced ef-
fects on' growth and longevity.
V. AQUATIC TOXICITY
A. Acute Toxicity
The data base for antimony and freshwater organisms is small and
indicates that plants may be more sensitive than fish or invertebrate
species.
A 96-hour LC5Q of 22,000 /jg/1 was reported for antimony tri-
chloride with the fathead minnow, whereas the value for bluegills and
antimony trioxide is above 530,000 ug/1 (U.5. EPA, 1979). For Daghnia
»
magna a 48-hour 1_C value of 19,000 jug/1 and a 64-hour EC5Q value of
19,800 pg/1 have been reported for antimony trichloride. Another 48-hour
-------
ECcn value for antimony trioxide and Daphnia maqna has been reported to
be above 530,000 jug/1 (U.S. EPA, 1979).
B. Chronic Toxicity
NO adverse effects on the fathead minnow were observed during an
embryo-larval test with antimony trioxide at the highest test concen-
tration of 7.5 jug/1 (U.S. EPA, 1978). However, a comparable test with
antimony trichloride produced limits of 1,100 and 2,300 jug/1 for a
chronic value of 800 jjg/1. A life cycle test with Daphnia magna and an-
timony trichloride produced limits of 4,200 and.. 7,000 jug/1 for a chronic
value of 5,400/jg/l (U.S. EPA, 1979). Pertinent information could not be
located in the available literature regarding chronic effects of antimony
on saltwater organisms.
C. Plants Effects
The 96-hour EC5Q values for chlorophyll a inhibition and re-
duction in cell number of the freshwater alga, Selenastrum capricornutum
are 610 and 630 pg/1, respectively. This indicates that aquatic plants
may be more sensitive than fish or invertebrate species (U.S. EPA,
1978). No inhibition of chlorophyll § reduction or in cell numbers of
the marine alga, Skeletonema costatum, were observed at concentrations as
high as 4,200 /jg/1 (U.-S. EPA, 1978).
D. Residues
There was no bioconcentration of antimony by the bluegill above
control concentrations during a 28 day exposure to antimony. No data
have been reported on bioconcentration of antimony in marine species.
c?
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
-97-
-------
public review; therefore, there is a possibility that these criteria may
be changed.
A. Human
Existing occupational standards for exposure to antimony are
reviewed in the recently released NI05H criteria document, Occupational
Exposure to Antimony (U.S. Department of Health, Education and Welfare,
1978). As stated in the NIOSH (1978) document, the American Conference
of Governmental Industrial Hygienists (ACGIH), in 1977, listed the TLV
for antimony as 0.5 mg/m along with a notice1- of intended change to a
proposed TLV of 2.0 mg/m for soluble antimony salts. The proposed TLV
was based mainly on the reports of Taylor (1966) and Cordasco (1974) on
accidental poisoning by antimony trichloride and pentachloride, respec-
tively. Proposed limits of 0.5 mg/m for handling and use of antimony
trioxide and 0.05 mg/m for antimony trioxide production were included
in the ACGIH (1977) notice of intended changes.
The Occupational Safety and Health Administration earlier adopted the
1968 ACGIH TLV for antimony of 0.5 mg/m3 as the Federal standard (29
CFR 1910.1000). This limit is consistent with limits adopted by many
other countries as described in Occupational Exposure Limits for Airborne
Toxic Substances - A tabular Compilation of Values from Selected Coun-
tries, a publication released by " the International Labour Office in
1977. The NIOSH (1978) document also presented table of exposure limits
from several countries, reproduced here as Table 1; the typical
standard adopted was 0.5 mg/m .
-------
TABLE 1
HYGIENIC STANDARDS OF SEVERAL COUNTRIES FOR
ANTIMONY AND COMPOUNDS IN THE WORKING ENVIRONMENT
CountryStandardQualifications
(mg/m3)
FinlandONot stated
Federal Republic of Germany 0.5 8-hour TWA
Democratic Republic of Germany 0.5 Not stated
Rumania 0.5 Not stated
USSR 0.5 For antimony dust
0.3 For fluorides and
chlorides (tri-and
pentavalent); obli-
gatory control of HF
and HC1
1.0 For trivalent oxides
and sulfides
1.0 For pentavalent
oxides and sulfides
Sweden 0.5 Not stated
USA 0.5
8-hour TWA
Yugoslavia 0.5 Not stated
Modified from Occupational Exposure Limits in Airborne ToxicSufa-
stances, International Labour Office.
The 0.5 mg/m level was also recommended as the United States occupa-
tional exposure standard by the NIOSH (1978) criteria document, based
mainly on estimated no-effect levels for cardiotoxic and pulmonary ef-
fects.
Based upon the data presented in the Ambient Water Quality Criteria
Document for Antimony (U.S. EPA, 1979), a recommended draft criterion of
145 jug/1 has been established. This value is based upon an acceptable
daily intake for man of 294 ug, derived from experimental animal studies
in which 5 ppm of antimony produced a slight shortening of lifespan with
no other deserved effects. An uncertainty factor of 100 was used in, ex-
trapolating from animal data to human health effects.
-------
B. Aquatic
The draft criterion for Antimony to protect freshwater aquatic
life as derived using the Guidelines is 120 /jg/1 as a 24 hour average and
the concentration should not exceed l,000jug/l at any time.
A saltwater criterion was not derived (U.S. EPA, 1979}
-too-
-------
ANTIMONY
REFERENCES
Abdalla, A., and M. Saif. 1962. Tracer studies with anti-
mony- 124 in man. In; G.E.W. Walstenhalne and M. O'Conner,
eds., Bilharziasis. Little Brown and Co., Boston, p. 287.
Aiello, G. 1955. Pathology of antimony. Folia Med., Naples
38: 100.
American Conference of Governmental Industrial Hygienists.
1977. Threshold limit values for chemical substances in
workroom air.
Belyaeva, A.P. 1967. The effect of ant.imony on reproduc-
tion. Gig. Truda Prof. Zabol 11: 32.
Brieger, H., et al. 1954. Industrial antimony poisoning.
Ind. Med. Surg. 23: 521.
Callaway, H.M. 1969. Antimony. In: The Encyclopedia Britan-
nica. Ency. Brit., Inc., 2: 20. "Chicago.
Casals, J.B. 1972. Pharmacokinetic and toxicological studies
of antimony dextran glycoside (RL-712). Brit. Jour. Pharmac.
46: 281.
Clemente, G.F. 1976. Trace element pathways from environ-
ment to man. Jour. Radioanal. Chem. 32: 25.
Cordasco, E.M. 1974. Newer concepts in the management
of environmental pulmonary edema. Angiology 25: 590.
Davies, T.A.L. 1973. The health of workers engaged in
antimony oxide manufacture—a statement. London, Department
of Employment, Employment Medical Adivsory Service, p. 2.
Djuric, D., et al. 1962. The distribution and excretion
of trivalent antimony in the -rat following inhalation.
Arch. Gewerbepath. Gewerbehyg. 19: 529.
El-Bassouri, M. , et al. 1963. Treatment of active urinary
schistosomiasis in children with sodium antimony dimercapto
succinate by the slow method. Trans. Roy. Soc. Trop. Med.
Hyg. 57: 136.
Felicetti, S.W., et al. 1974a. Metabolism of two valence
states of inhaled antimony in hamsters. Amer. Ind. Hyg.
Assoc. Jour. 355: 2S»2.
Felicetti, S.W., et al. 1974b. Retention of inhaled anti-
mony-124 in the beagle dog as a function of temperature
of aerosol formation. Health Phys. 26: 525.
-ttl-
-------
Gross, et al. 1955. Toxicological study of calcium halo-
phasphate phosphors and antimony trioxide. In: Acute and
chronic toxicity and some pharmacological aspects. Arch.
Indust. Health 11: 473.
International Labour Office. 1977. Occupational exposure
limits for airborne toxic substance - a tabular compilation
of values from selected countries. Occupational Health
Series No. 37. United International Labour Office, Geneva.
p. 44.
Kanisawa, M., and H.A. Schroeder. 1969. Life term studies
on the effect of trace elements of spontaneous tumors in
mice and rats. Cancer Res. 29: 892.
Molokhia, M.M., and H. Smith. 1969. Tissue distribution
of trivalent antimony in mice infected, with Schistosoma
Mansoni. Bull. WHO 40: 123.
Murthy, G.K., et al. 1971. Levels of antimony, cadmium,
chromium, cobalt, manganese and zinc in institutional total
diets. Environ. Sci. and Tech. 5: 436.
NIOSH. 1978. Criteria for a recommended standard: Occupa-
tional exposure to antimony. DHEW (NIOSH) G.P.O. No. 017-
033-00335-1.
Schroeder, H.A. 1966. Municipal drinking water and cardio-
vascular death rates. Jour. Amer. Med. Assoc. 195: 81.
Schroeder, H.A. 1970. A sensible look at air pollution
by metals. Arch. Environ. Health 21: 798.
Schroeder, H.A., and L.A. Kraemer. 1974. Cardiovascular
mortality, municipal water and corrosion. Arch. Enviorn.
Health 28: 303.
Schroeder, H.A., et al. 1970. Zirconium, niabium, antimony
and lead in rats: Life term studies. Jour. Nutr. 100: 59.
Tanner, J.T., and M.H. Friedman. 1977. Neutron activation
analysis for trace elements in foods. Jour. Radioanal.
Chem. 37: 529.
Taylor, P.J. 1966. Acute intoxication from antimony tri-
chloride. Br. Jour. Ind. Med. 23: 318.
Thomas, R.G., et al. 1973. Retention patterns of antimony
in mice following inhalation of particles formed at different
temperatures. Proc. Soc. Exp. Biol. Med. 144(2): 544.
-------
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. 1979. Antimony: Ambient Water Quality Criteria.
U.S. Environ. Prot. Agency, Washington, D.C.
Waitz, J.A., et al. 1965. Physiological disposition of
antimony after administration of Sb-labeled tartar emetic
to rats, mice and monkeys and the effects of tris (p- amino
phenyl} carbonium pamoate on this distribution. Bull. WHO
33: 537.
Woolrich, P.P. 1973. Occurrence of trace metals in the
environment: an overview. Amer. Ind. Hyg. Assoc. Jour. 34:
217.
-------
No. 11
Arsenic
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.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (GAG) has evaluated
arsenic and has found sufficient evidence to indicate that
this compound is carcinogenic.
-------
ARSENIC
SUMMARY
Epidemiological studies have shown increased death rates
from lung cancer in workers exposed to arsenic, probably
through inhalation. Other human studies have shown increased
skin cancers in non-occupationally exposed populations. In-
creased incidence of lymphonas and hemangioendotheliomas are
also occasionally reported.
Arsenicals have produced mutagenic effects in plants,
bacteria, _in vitro leukocyte cultures, and in the lymphocytes
of exposed humans. The teratogenic effects of arsenicals
have been demonstrated in many animal species. An increased
frequency of abortions in pregnant women exposed to arsenic
has been reported in a single study (U.S. EPA, 1979).
The chronic toxic effects of arsenic involve skin hyper-
keratosis, liver damage, neurological disturbances (including
hearing loss), and a gangrenous condition of the extremities
(Blackfoot disease). An increased mortality from cardiovas-
cular disease resulting from chronic arsenic exposure has
been suggested in two studies.
The data base for the toxicity of arsenic to aquatic or-
ganisms is more complete for freshwater organisms, where con-
centrations as low as 128 ug/1 have been acutely toxic to
freshwater fish. A single marine species produced an acute
value in excess of 8,000 ug/1. Based on one chronic life
cycle test using Daphnia magna, a chronic value for arsenic
was estimated at 853 ug/1.
t
-------
ARSENIC
I. INTRODUCTION
This profile is based on the Ambient Water Quality Cri-
teria Document for Arsenic (U.S. EPA, 1979).
Arsenic is a gray, crystalline metalloid with a molecu-
lar weight of 74.92, a density of 5,727, a melting point (at
28 atmospheres) of 817°C, and a boiling point (sublimates) of
613°C (Weast, 1975). Arsenic exists in a variety of valence
states; the most common forms include pentavalent .(arsenate) ,
trivalent (arsenite), and -3 valency (arsine). Properties of
some inorganic arsenic compounds are shown in Table 1.
Conditions of low pH, low oxidation-reduction potential,
and low dissolved oxygen in water favor formation of the
lower valency states {arsenite and arsine); more basic, oxy-
genated waters favor the presence of arsenate. Inorganic
arsenic can be converted to organic alkyl-arsenic acids and
to methylated arsines under both aerobic and anaerobic condi-
tions (U.S. EPA, 1979).
Arsenic and its compounds are used in the manufacture of
glass, cloth, and electrical semiconductors, as fungicides
and wood preservatives, as growth stimulants for plants and
animals, and in veterinary applications {U.S. EPA, 1976).
Production is currently 1.8 x 10^ metric tons per year
(U.S. EPA, 1979).
Arsenic will persist in some form in the environment.
Inorganic arsenate is thermodynaraically favored under normal
conditions over arsenite in water and is a more soluble form
(Ferguson and Gavis, 19-72). Both arsenate and arsenite may
be precipitated from water by adsorption onto iron and alum-
-------
Table 1. Properties of Some Inorganic Arsenic Compounds
(Standen, 1967; U.S. EPA, 1976)
Compound
Formula
Water Solubility
Specific Properties
•o
i
Arsenic trioxide
Arsenic pentoxide
Arsenic hydride
Arsenic(III) sulfide
Arsenic sulfide
Arsenic(V) sulfide
AS2°3
AsH
12 x io6 ug/i @ o°c
21 x IO6 ug/1 @ 25°C
Dissolves in water to form
arsenious acid (H3As03:
K = 8 x 10~10 @ 25°C)
2300 x 10^ ug/1 @ 20°C Dissolves in water to form
arsenic acid
K2 = 5.6 x IO"8;
K3 = 3 x 10~13}
= 2.5 10
-4
20 ml/100 g cold water
520 ug/1 @ 18°C
This compound and its methyl
derivatives are considered to
be the most toxic.
Burns in air forming arsenic
trioxide and sulfur dioxide;
occurs naturally as orpiment.
Occurs naturally as realgar.
1400 ug/1 8 0°C
-------
inum compounds (U.S. EPA, 1979). Methylated arsines appear
to be volatile and sparingly soluble. Waters containing high
organic matter may bind arsenic compounds to colloidal humic
matter (U.S. EPA, 1979).
II. EXPOSURE
Arsenic appears to be ubiquitous in the environment.
The earth's crust contains an average arsenic concentration
of 5 mg/kg (U.S. EPA, 1976), The major sources of arsenic in
the environment are industrial, such as those in the smelting
of non-ferrous ores and in coal-fired power plants that uti-
lize fuel containing arsenic. Substantial arsenic contamina-
tion of water can occur from the improper use of arsenical
pesticides (U.S. EPA, 1979).
Based on available monitoring data, the U.S. EPA (1979)
has estimated the uptake of arsenic by adult humans from air,
water, and food:
Source mg/day
Maximum Conditions Minimum Conditions
Atmosphere .125 - .001
Water 4.9 0.002
Pood Supply .9 .007
Total 5.925 .010
Contaminated well water, seafood, and air near smelting
plants all present sources of high potential arsenic intake.
The U.S. EPA (1979) has estimated the weighted average
bioconcentration factor (BCF) for arsenic to be 2.3 in the
edible portions of fish and shellfish consumed by Americans,
*
This estimate was based on bioconcentration studies in fresh-
water fish.
-------
III. PHARMACOKINETICS
A. Absorption
The main routes by which arsenic can enter the body
are inhalation and ingestion. Particle size and solubility
greatly influence the biological fate of inhaled arsenic.
Falk and Kotin (1961) have reported that the optimal range of
particle size for deposition in the lower tracheobronchial
tree is 0.1 to 2 u« Larger particles are trapped by the
mucous membranes of the nose and throat and swallowed;
V
following this, the particles may be absorbed from the
gastrointestinal tract (U.S. EPA, 1979).
Human inhalation studies in terminal lung cancer
patients (Holland, et al. 1959) have indicated that 4.8 to
8.8 percent of inhaled arsenic-74 in cigarette smoke may be
absorbed. Radioactive arsenite inhaled in an aerosol solu-
tion by two patients showed 32 and 62 percent absorption, re-
spectively. Pinto, et al. (1976) studied arsenic excretion
in 24 workers exposed to the compound during copper smelting;
urinary arsenic levels were found to correlate significantly
with average airborne arsenic concentrations.
Water soluble arsenicals are readily absorbed through
the gastrointestinal tract. Studies with radioactive arse-
nate administered orally to rats have shown 70 to 90 percent
absorption from the gastrointestinal tract (Urakubo, et al.
1975; Dutkiewicz, 1977). Arsenic trioxide is only slightly
soluble in water and is not well absorbed. Theoretically,
*
trivalent arsenicals should be less readily absorbed than
pentavalent forms due to reactivity with membrane components
-------
and lower solubility (U.S. EPA, 1979). However, investiga-
tors have reported high absorption of trivalent arsenic from
the gastrointestinal tract in humans (Bettley and O'Shea,
1975; Crecelius, 1977).
The absorption of arsenicals following dermal expo-
sure has been described in rats (Dutkiewicz, 1977) and humans
(Robinson, 1975; Garb and Hine, 1977).
Arsenic has been detected in the tissues (Kadowaki,
1960) and cord blood of newborns (Kagey, et al. 1977), and
thus transfers across the placenta in humans.
B. Distribution
Injection of radiolabelled arsenite in terminally
ill patients produced widespread distribution of the compound
(WHO, 1973). Hunter, et al. (1942) studied the distribution
of radioactive arsenicals in humans following oral and paren-
teral administration and found arsenic in the liver, kidney,
lungs, spleen, and skin during the first 24 hours after ad-
ministration. Levels of arsenic are maintained for long per-
iods in bone, hair and nails {Kadowaki, 1960; Liebscher and
Smith, 1968).
Tissue distribution of pentavalent arsenic has been
described in only a few animal studies; these studies indi-
cate only minor differences in distribution between trivalent
and pentavalent arsenicals (WHO, 1973).
-------
C. Metabolism
Studies with brain tumor patients given injections
of trivalent arsenic indicate that about 60 percent of the
total urinary arsenic was in the pentavalent state the first
day after dosing (Mealey, et al. 1959). Braman and Foreback
(1973) have analyzed human urine samples and detected high
amounts of methylated forms (dimethyl arsenic acid and methyl
arsenic acid). Analysis of the urine of one patient who in-
gested arsenic-contaminated wine indicated that 8 percent of
the initial dose was excreted as inorganic arsenic, 50 per-
cent was excreted as dimethyl arsenic acid, and 14 percent
was excreted as methyl arsenic acid (Crecelius, 1977).
The half-lives of inorganic and organic (methy-
lated) arsenicals in one patient have been reported as 10 and
30 hours, respectively (Crecelius, 1977).
D. Excretion
Arsenic is excreted primarily in the urine, with
small amounts removed in the feces and through normal hair
loss and skin shedding (U.S. EPA, 1979). Reports of minor
arsenic loss in sweat have also been made (Vellar, 1969).
Small amounts of radioactive arsenic (.003 to .35
percent) have been detected in expired air following adminis-
tration to rats (Dutkiewicz, 1977) and chickens (Overby and
Fredrickson, 1963).
IV. EFFECTS
A. Carcinogenicity
»
Epidemiological studies have shown an increased
mortality rate from respiratory cancer in workers exposed to
-------
arsenic during smelting operations (Lee and Fraumani, 1969;
Pinto and Bennett, 1963; Snegireff and Lombard, 1951; Kurat-
sune, et al. 1974). A retrospective study of Dow Chemical
employees indicated that workers exposed primarily to lead
arsenate and calcium arsenate showed increased death rates
from lung cancer and malignant neoplasms of the lymphatic and
hematopoietic systems (except leukemia) (Ott/ et al. 1974).
A similar trend was noted in a study of retired
Allied Chemical workers (Baetjer, et al. 1975),
High rates of development of skin cancers have been
reported in-several studies of populations exposed to high
concentrations of arsenic in drinking water (Geyer, 1898;
Bergogilio, 1964; Tseng, et al. 1968).
Hemangioendothelioma of the liver associated with
exposure to arsenicals through ingestion has been reported in
several case studies (Rothf 1957; Regelson, et al. 1968).
Extensive experiments in animal systems with arsen-
icals administered in the diet or drinking water, or applied
topically or by intratracheal instillation failed to show
positive tumorigenic effects (U.S. EPA, 1979). However, two
recent reports have shown effects in animals, Schrauzer and
Ishmael (1974) indicated that feeding of sodium arsenite in
drinking water accelerated the rate of spontaneous mammary
tumor formation. Osswald and Goerttler (1971) found an
increase in leukemias and lymphomas in mice injected
repeatedly with sodium arsenate.
Animal studies on the skin tumor-promoting or co-
carcinogenic effects of arsenicals have produced negative
results (Raposo, 1928; Baroni, et al. 1963; Boutwell, 1963).
-------
B. Mutagenicity
An increased incidence of chromosomal aberrations
has been found in persons exposed to arsenic occupationally
and medically (Petres, et al. 1970; Nordenson, et al. 1978;
Burgdorf, et al. 1977).
In vitro chromosomal changes following exposure to
arsenicals have been reported in root meristem cultures
(Levan, 1945) and in human leukocyte cultures (Petres and
Hundeiker, 1968; Petres, et al. 1970, 1972; Paton and
Allison, 1972).
Arsenate has been found to increase the frequency
of chromosome exchanges in Drosophila. Several organic ar-
senicals have a synergistic effect with ethylmethane sulfon-
ate in producing chromosome abnormalities in barley (Moutsh-
cen and Degraeve, 1965),
Sodium arsenate, sodium arsenite, and arsenic tri-
chloride produced positive mutagenic effects in a recombinant
strain of Bacillus subtillus (Nishioka, 1975). Loforth and
Ames (1978) were unable to show mutagenic effects of trival-
ent and pentavelent arsenicals in the Ames Salmonella assay.
Arsenite exposure decreased the survival of _E. coli after UV
damage of cellular PNA (Rossman, et al. 1975).
-------
C. Teratogenicity
Nordstroro, et al. (1979) have reported an increase
in the frequency of spontaneous abortions in pregnant women
living in the vicinity of a copper smelting plant; the expo-
sure environment was complex, involving several heavy metals
and sulfur dioxide.
Sodium arsenate has been shown to induce teratogen-
ic effects in the chick embryo (Ridgway and Karnofsky, 1952),
in golden hamsters (Perm and Carpenter, 1968; Ferm, et al.
*.
1971), in mice (Hood and Bishop, 1972), and rats (Beaudoin,
1974). Malformations noted included exencephaly, anenceph-
aly, renal agenesis, gonadal agenesis, eye defects, and rib
and genitourinary abnormalities. Sodium arsenite injected
intraperitoneally into mice produced a lower incidence of
malformations than an equivalent dose of sodium arsenate
(Hood and Bishop, 1972; Hood, et al. 1977}. Thacker, et al.
(1977) has noted that a higher oral dose of sodium arsenate
is needed to produce teratogenic effects in mice, when com-
pared to intraperitoneal doses.
Feeding of three generations of mice with low doses
of sodium arsenite in the chow failed to produce teratogenic
effects, but did decrease litter size (Schroeder and flitch-
ener, 1971).
D. Other Reproductive Effects
Pertinent information could not be located in the
available literature regarding other reproductive effects..
E. Chronic Toxicity
A variety of chronic effects of arsenic exposure
has been noted. This includes a characteristic palmar-
-------
plantar hyperkeratosis and a gangrenous condition of the
hands and feet called Blackfoot disease (U.S. EPA, 1979).
Several clinical reports of liver damage in patients treated
with arsenical medication have been published (WHO, 1979).
An increased mortality from cardiovascular disease has been
noted in two epidemiological studies of smelter workers ex-
posed to high airborne arsenic (Lee and Fraumeni, 1969; U.S.
EPA, 1979). Neurological disturbances, including hearing
loss, in workers exposed to arsenicals have been reported
«•.
(WHO, 1979).
Effects of arsenicals on the hematopoietic system
following chronic exposure have also been noted (WHO, 1979).
These include disturbed erythropoiesis and granulocytopenia,
which may lead to impaired resistance to viral infections.
V. AQUATIC TOXICITY
A. Acute Toxicity
Seven static and seven flow-through bioassays from
^48 to 96-hours in duration provide a range of LCgn values
for freshwater fish of 290 to 150,000 y.g/1- Hughes and Davis
(1967) demonstrated the most sensitive species as being blue-
gill fingerlings, Lepomis macrochirus, while Sorenson (1976)
reports that the most resistant species was the green sun-
fish, Lepomis cyanellus. Both species were tested in static
tests. Sanders and Cope (1966) provided the data for fresh-
water invertebrates in static bioassays. The cladoceran,
Simocephalus serrulatus, was the most sensitive with an 48-
"' ' " ' *
hour LC5Q value of 812 ug/1, while the stonefly, Ptaron-
arcys caljLfornj.ca, was the most resistant species with an
-------
LC5o value Of 22,040 ug/1. In marine, organisms, the chum
salmon, Onchorhynchus keta, had a 48-hour flow-through
value of 8,331 jig/1 (Alderdice and Brett, 1957). Two marine
invertebrates were tested in 96 or 48-hour static-renewal or
static assays and produced the following LC5Q values: bay
scallop, Argopecten irradlana, with 3,490 ug/1; and the em-
bryos of the American oyster, Crassostrea virgin lea, with a
value of 4,330 y.g/1.
B. Chronic Toxicity
One chronic life cycle freshwater test has provided
a chronic value of 853 ug/1 for arsenic to Daphhia magna.
Pertinent data could not be located in the available litera-
ture for the chronic toxicity of arsenic to marine organisms.
C. Plant Effects
The lowest effective concentration 'recorded was 100
percent kill levels of 2,320 ug/1 for four species of fresh-
water algae.
D. Residues
Bioconcentration factors for five freshwater inver-
tebrate species and two fish species ranged from less than 1
to 17 (U.S. EPA, 1979) .
VI. EXISTING GUIDELINES AND STANDARDS
A. Hunan
Criteria for organic and inorganic arsenicals have
been derived. However, due to public comment questioning the
*
relevancy and accuracy of the studies used in the development
of these criteria, further review is necessary before final
recommendation.
-------
The OSHA tine-weighted average exposure criterion
foe arsenic is 10 y.g/m^.
B, Aquatic
For arsenic, the draft criterion for freshwater or-
ganisms is 57 ug/1, not to exceed 130 ug/1. For marine or-
ganisms, the draft criterion is 29 vg/1, not to exceed 67
ug/1 (U.S. EPA,1^79).
-------
ARSENIC
REFERENCES
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arsenite to young chum salmon. Prog. Rep. Pacific Coast Stat.
Pish. Res. Board Can. 109: 27.
Baetjer, A.M., et al. 1975. Cancer and occupational exposure
to inorganic arsenic. Page 393 _in Abstracts. 18th Int. Cong.
Occup. Health Brighton, England, September 14-19.
Baroni, C., et al. 1963. Carcinogenesis tests of two inor-
ganic arsenicals. Arch. Environ. Health 7: 668.
Beaudoin, A.R. 1974. Teratogenicity of sodium arsenate in
rats. Teratology 10: 153.
Bergoglio, R.M. 1964. Mortalidad por cancer en zonas de
aguas arsenicales de la Provincia de Cordoba, Republica Argen-
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Bettley, P., and J. O'Shea. 1975. The absorption of arsenic
and its relation to carcinoma. Brit. Jour. Dermatol. 92:
563.
Boutwell, R. 1963. A carcinogenicity evaluation of potassium
arsenite and arsenilic acid. Jour. Agric. Food Chem. 11:
381.
Braman, R.S., and C.C. Eoreback. 1973. Methylated forms of
arsenic in the environment. Science 182: 1247.
Burgdorf, W. , et al. 1977. Elevated sister chromatic ex-
change rate in lymphocytes of subjects treated with arsenic.
Hum. Genet. 36: 69.
Crecelius, E.A. 1977. Changes in the chemical speciation of
arsenic following ingestion by man. Environ. Health Perspect.
19; 147.
Dutkiewicz, T. 1977. Experimental studies on arsenic absorp-
tion routes in rats. Environ. Health Perspect. 19: 173.
Falk, H.L., and P. Kotin. 1961. An assessment of factors
concerned with the carcinogenic properties of air pollutants.
Natl. Cancer Inst. Mon. 9: 81.
Ferguson, J.F., and J, Gavis. 1972. A review o£ the arsenic
cycle in natural waters. Water Res. 6: 1259.
-------
Perm, V.H., and S.J. Carpenter. 1968. Malformation induced
by sodium arsenate. Jour. Reprod. Fe.rtil. 17: 199.
. Ferm, V.H., et al. 1971. The teratogenic profile of sodium
arsenate in the golden hamster. Arch. Environ. Health 22:
557.
Garb, L.G., and C.H. Hine. 1977. Arsenical neuropathy: Res-
idual effects following acute industrial exposure. Jour.
Occup. Med. 19: 567.
Geyer, L. 1898. Uber die chronischen Hautveranderungen beim
Arsenicismus und Betrachtungen uber die Massenerkrankungen in
Reichenstein in Schlesien, Arch. Derm. Syphilol. 43: 221.
Holland, R.H., et al. 1959. A study of inhaled arsenic-74 in
man. Cancer Res. 19: 1154.
Hood, R.D., and S.L. Bishop. 1972. Teratogenic effects of
sodium arsenate in mice. Arch. Environ. Health 24: 62.
. Hood, R.D., et al. 1977. Effects in the mouse and rats of
prenatal exposure to arsenic. Environ. Health Perspect, 19:
219.
Hunter, F.T., et al. 1942. Radioactive tracer studies on
arsenic injected as potassium arsenite. Jour. Pharmacol. Exp.
Ther. 76: 207.
Hughes, J.S., and J.T. Davis. 1967. Effects of selected
herbicides on bluegill sunfish. Pages 480-482. In Proc. 18th
Ann. Conf., S.E. Assoc. Game Pish Comm., October T8, 19, 20
and 21, 1964. Clearwater, Fla. Columbia, S.C.: S.E. Assoc.
Game Fish Comm.
• Kadowaki, K. 1960. Studies on the arsenic contents in organ-
tissues of the normal Japanese. Osaka City Med. Jour. 9:
2083.
• Kagey, B., et al. 1977. Arsenic levels in maternal-fetal
tissue sets. Trace Subst. Environ. Health 11: 252.
Kuratsune, M., et al. 1974. Occupational lung cancer among
copper smelters. Int. jour. Cancer 13: 552.
Lee, A.M., and J.F. Fraumeni, jr. 1969. Arsenic and respira-
tory cancer in man: An occupational study. Jour. Natl. Can-
cer Inst. 42: 1045.
Levan, A. ^1945. Cytological reactions induced by inorganic
salt solutions. Nature 156: 751.
-------
Liebscher, K., and H. Smith. 1968. Essential and nonessen-
tial trace elements. -A method of determining whether an ele-
ment is essential or nonessential in human tissue. Arch.
Environ. Health 17: 881.
Lofroth, G., and B. Ames. 1978. Mutagenicity of inorganic
compounds in Salmonella typhimurium; arsenic, chronium, and
selenium. Mutat. Res. 53: 65.
Mealey, J., Jr., et al. 1959. Radioarsenic in plasma,
urine, normal tissues, and intracranial neoplasms. Arch.
Neurol. Paychiatry 81: 310.
Moutshcen, J,, and N. Degraeve. .1965. Influence of thiol—
nhibiting substances on the effects of ethyl methane sulphon-
ate (EMS) on chromosomes. Experientia 21: 200.
Nishioka, H. 1975. Mutagenic activities''of metal compounds
in bacteria. Mutat. Res. 31: 185.
Nordenson, I., et al. 1978. Occupational and environmental
risks in and around a smelter in northern Sweden. II. Chro-
mosomal aberrations in workers exposed to arsenic. Hereditas
88: 47.
Nordstrom, S., et al. 1978. Occupational and environmental
risks in and around a smelter in northern Sweden. III. Fre-
quencies of spontaneous abortion. Hereditas 88: 51.
Osswald, H., and Kl. Goerttler. 1971. Laukosen bei der Maus
nach diaplacentarer und postnataler Arsenik-Applikation.
Dtsch. Gesmte Path. 55: 289.
Ott, M.G., et al. 1974. Respiratory cancer and occupational
exposure to arsenicals. Arch. Environ. Health 29: 250.
Overby, L.R., and R.L. Predrickson. 1963. Metabolic stabil-
ity of radioactive arsanilic acid in chickens. Jour. Agric,
Food Chem. 11: 378..
Paton, G.R., and A.C. Allison. . 1972. Chromosome damage in
human cell cultures induced by metal salts. Mutat. Res. 16:
332.
Petres, J., and M. Hundeiker. 1968, "Chromosomenpulverisa-
tion" nach Arseneinwirkung auf Selljulturen in vitro. Arch.
Klin. Exp. Dermatol. 231: 366.
Petres, J., et al. 1970. Chromosomenaberrationen an mensch-
lichen Lymphozyten bei chronischen Arsenchaden. Dtsh. Med.
Wochenschr. 95; 79.
Petres, J., et al. 1972. Zum Einfluss anorganiechen Arsens
auf die DNS-Synthese menschlicher Lymphocyten _in vitro. Arch
Derm. Porsch. 242: 343.
-------
Pinto, S.S., and B.M. Bennett. 1963. Effect of arsenic tri-
oxide exposure on mortality. Arch. Environ. Health 7: 583.
•Pinto, S.S., et al. 1976. Mortality experience of arsenic
exposed workers, unpubl.
• Raposo, L. 1929. Le cancer a 1'arsenie. C.P. Soc. Biol.
(Paris) 98: 86.
Regelson, W., et al. 1968. Heraangioendothelial sarcoma of
liver from chronic arsenic intoxication by Fowler's solution.
Cancer 21: 514.
Ridgway, L.P., and D.A. Karnovsky. 1952. The effects of
metals on the chick embryo: Toxicity and production of abnor-
malities in development. Annu. N.Y. Acad. Sci. 55: 203.
. Robinson, T. 1975. Arsenical polyneuropathy due to caustic
arsenical paste. Brit. Med. Jour. 3: 139.
Rossman, T.-, et al. 1975. Effects of sodium arsenite on the
survival of UV-irradated Escherichia coli; Inhibition of a
rec A dependent function. Mutat. Res. 30: 157.
Roth, F. 1957. The sequelae of chronic arsenic poisoning in
Moselle vintners. German Med. Monthly 2: 172.
Sanders, H.O., and O.B. Cope. 1966. Toxicities of several
pesticides to two species of cladocerans. Trans. Am. Fish.
Soc. 95: 165.
Schrauzer, G., and D. Ishmael. 1974. Effects of selenium
and of arsenic on the genesis of spontaneous mammary tumors
in inbred CoH mice. Ann. Clin. Lab. Sci. 4: 441.
Schroeder, H.A., and M. Mitchener. 1971. Toxic effects of
trace elements on the reproduction of mice and rats. Arch.
Environ. Health 23: 102.
Snegireff, L.S., and O.M. Lombard. 1951. Arsenic and can-
cer. Observation in the metallurgical industry. AMA Arch.
ind. Hyg. 4: 199.
Sorenson, E.M.B. 1976. Toxicity and accumulation of arsenic
in green sunfish, Lepomis cyanellus, exposed to arsenate in
water. Bull. Environ. Contam. Toxicol. 15: 756.
Sram, R., and V. Bencko. 1974. A contribution to the evalu-
ation of the genetic risk of exposure to arsenic. Cesk Hyg.
19: 308.
Standen, A. (ed.) 1967. Kirk-Othmer encyclopedia of chemi-
cal technology. Interscience Publishers, New York.
-------
Thacker, G., et al. 1977. Effects of administration routes
on arsenate teratogenesis in mice. Teratology 15: 30.
Tseng, W.P., et al. 1968. Prevalence of skin cancer in an
endemic area of chronic arsenicism in Taiwan. Jour. Natl.
Cancer Inst. 40; 453.
U.S. EPA. 1976. Arsenic and its compounds. EPA 560/6-76-
016. U.S. Environ. Prot. Agency, Washington, B.C.
U.S. EPA. 1979. Arsenic: Ambient Water Quality Criteria.
U.S. Environ. Prot. Agency, Washington, D.C.
Urakubo, G., et al. 1975. Studies in the fate of poisonous
metals in experimental animals (V). Body retention and ex-
cretion of arsenic. Jour. Food Hyg. Soc. Jpn. 16: 34.
v.
Vellar, 0. 1969. Nutrient lopes through sweating. Thesis,
Universitetsforlaget, Oslo, Norway.
Weast, R.C. (ed.) 1975. Handbook of chemistry and physics.
56th ed. CRC Press, Cleveland, Ohio.
WHO. 1973. Environme nta1 Health Criteria; Arsenic. World
Health Organization. Geneva.
Vf
-------
No. 12
Asbestos
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.G. 20460
APRIL 30, 1980
-US'-
-------
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
asbestos and has found sufficient evidence to indicate that
this compound is carcinogenic.
-------
ASBESTOS
Summary
Numerous studies indicate that asbestos fibers introduced into the
pleura, peritoneum, and trachea of rodents have induced malignant tumors.
The strongest evidence for the carcinogenicity of ingested asbestos is pro-
vided by epidemiology of human populations occupationally exposed to high
concentrations of airborne asbestos dust. Inhalation exposure to asbestos
dust is accompanied by ingestion because a high percentage of the inhaled
fibers are removed from the lung by mucociliary action and subsequently
swallowed. Peritoneal mesothelioma, often in great excess, and modest ex-
cesses of stomach esophagus, colonrectal, and kidney cancer have been linked
to occupational exposure to asbestos.
Pertinent data on the acute or chronic effects of asbestos to aquatic
organisms were not found in the available literature.
-------
ASBESTOS
I. INTRODUCTION
This profile is based primarily upon the Ambient Water Quality Criteria
Document for Asbestos (U.S. EPA, 1979). In addition, valuable information
is available from recent reviews by the International Agency for Research on
Cancer (IARC, 1977) and the National Institute for Occupational Safety and
Health (NIOSH, 1977).
Asbestos is a broad term applied to numerous fibrous mineral silicates
composed of silicon, oxygen, hydrogen, and metal cations such as sodium,
magnesium, calcium, or iron. There are two major groups of asbestos, ser-
pentine (chrysotile or "white asbestos") and amphibole. Although chrysotile
is considered to be a distinct mineral, there are five fibrous amphiboles:
actinolite, amosite ("brown asbestos"), anthophyllite, crocidolite ("blue
asbestos"), and tremolite.. The chemical composition of different asbestos
fibers varies widely, and typical formulas are presented in Table 1. Some
typical physical properties of three different mineral forms of asbestos are
presented in Table 2.
TABLE 1
TYPICAL FORMULAS FOR ASBESTOS FIBERS
1. Serpentines
2. Amphiboles
chrysotile
amosite
crocidolite
anthophyllite
tremolite
actinolite
Na/2(Mg,
-------
TABLE 2
TYPICAL PHYSICAL PROPERTIES OF CHRYSOTILE (WHITE ASBESTOS),
CROCIDOLITE (BLUE ASBESTOS), AND AMOSITE
Units Chrysotile Crocidolite Amosite
(white asbestos) (blue asbestos)
Approximate
diameter of micron 0.01
smallest fibers
Specific - 2.55
gravity
Average
tensile , Ib./inch2 • 3.5 x 105
strength
Modulus of Ib./inch2 23.5 x 10.6
elasticity
0.08 0.1
3.37 3.45
\
5 x 10.5 1.75 x 105
27.0 x 106 23.5 x 10$
Asbestos minerals, despite a relatively high fusion temperature, are
completely decomposed at temperatures of 1,000°C. Both the dehydroxyla-
tion temperature and decomposition temperature increase with increased MgO
content among the various amphibole species (Speil and Leineweber, 1969).
The solubility product constants for various chrysotile fibers range
from 1.0 x HT11 to 3 x 10~lz. Most materials have a negative surface
charge in aqueous systems. However, since chrysotile has a positive U)
charge, it will attract, or be attracted to, most dispersed materials. The
highly reactive surface of asbestos causes many surface reactions which are
Intermediate between simple absorption and a true chemical reaction. The
absorption of various materials on the surface of chrysotile supports the
premise that the polar surface of chrysotile has a greater affinity for
polar molecules (e.g., H?0, NH ) than for non-polar molecules (Speil and
Leineweber, 1969).
-------
Of all the asbestos minerals, chrysotile is the most susceptible to
acid attack. It is almost completely destroyed within one hour in 1 N HCL
at 95°C. Amphibole fibers are much more resistant to mineral acids
(Lindell, 1972).
The resistance of the asbestos fibers to attack by reagents other than
acid is excellent up to temperatures of approximately 100°C with rapid
deterioration observed at higher temperatures. Chrysotile is completely de-
composed in concentrated KOH at 200°C. In general, organic acids have a
tendency to react slowly with chrysotile (Spell and Leineweber, 1969).
Chrysotile is the major type of asbestos used in the manufacture of as-
bestos products. These products include asbestos cement pipe, flooring pro-
ducts, paper products (e.g., padding), friction materials (e.g., brake lin-
ings and clutch facings), roofing products, and coating and patching com-
pounds. In 1975, the total consumption of asbestos in the U.S. was 550,900
thousand metric tons (U.S. EPA, 1979).
Of the 243,527 metric tons of asbestos discharged to the environment,
98.3 percent was discharged to land, 1.5 percent to air, and 0.2 percent to
water (U.S. EPA, 1979). Solid waste disposal by consumers was the single
largest contribution to total discharges. Although no process water is used
in dry mining of asbestos ore, there is the potential for runoff from asbes-
tos waste tailings, wet mining, and iron ore mining. Mining operations can
also contribute substantially to asbestos concentrations in water by air and
solid waste contamination. In addition to mining and industrial discharges
of asbestos, asbestos fibers, which are believed to be the result of rock
outcroppings, are found in rivers and streams.
-------
II. EXPOSURE
A. Water
Asbestos is commonly found in domestic water supplies. Of 775 re-
cent samples analyzed by electron microscopy under the auspices of the U.S.
EPA, 50 percent showed detectable levels of asbestos, usually of the chryso-
tile variety (Millette, 1979). Nicholson and Pundsack (1973) measured aver-
age asbestos levels of 0.3-1.5 ;jg/l in drinking water from two Eastern
United States river systems. Levels of 2.0 to 172.7 x 106 fibers/1 have
been reported in Canadian tap water, the highest levels being found in un-
faltered tap water near a mining area (Cunningham and Pontefract, 1971). In
other studies of Canadian drinking water levels of 0.1 to 4 x 106 fibers/1
have been reported (Kay, 1973). The U.S. EPA (1979) has concluded that
about 95 percent of water consumers in the United States are exposed to as-
bestos fiber concentrations of less than 10 fibers/1. The mass concen-
trations of chrysotile asbestos in the water of cities with less than 10
fibers/1 are likely to be less than 0.01 jug/1, corresponding to an adult
daily intake of less than 0.02 ug. Pertinent data 'on the ability of aquatic
organisms to bioconcentrate asbestos from water were not located in the
available literature.
B. Food
There are scant data on the contribution of food products to popu-
lation asbestos exposure. However, asbestos fibers and talc, which some-
times contains asbestos as an impurity, may be used in the manufacture of
certain processed foods such as sugar, coated rice, vegetable oil and lard
(IARC, 1977). Cunningham and Pontefract (1971) reported that certain beers
f
and wines could contain asbestos fibers at levels similar to those Found in
drinking water systems (106 to 107 fibers/I).
-------
C. Inhalation
Asbestos is present in virtually all metropolitan areas. Concen-
trations of asbestos in urban atmosphere are usually less than 10 ng/m ,
but may reach 100 ng/m (Nicholson, et al. 1971; Nicholson and Pundesack,
1973; Sebastian, et al. 1976; IARC, 1977). Construction sites and buildings
fireproofed with loose asbestos material showed the most significant contam-
ination with individual measurements as high as 800 ng/m (Nicholson, et
al. 1975).
III. PHARMACOKINETICS
There are contradictory data concerning whether ingested asbestos
fibers are capable of passage across the gastrointestinal mucosa (Gross, et
al. 1974; Cooper and Cooper, 1978; Cunningham and Pontefract, 1973;
Cunningham, et al. 1977). Most ingested asbestos particles are excreted in
the feces (Cunningham, et al. 1976). However, at least one recent study
(Cook and Olson, 1979) indicates that ingestion of drinking water containing
amphibole fibers may result in the appearance of these fibers in the urine,
thus providing evidence for passage of asbestos across the human gastro-
intestinal tract.
Ingestion of asbestos fibers is accompanied by swallowing of many
fibers cleared from the respiratory tract by mucociliary action. More than
half the asbestos inhaled will likely be swallowed (U.S. EPA, 1979). The
deposition of asbestos fibers in the lung is a function of their diameter
rather than length, as about 50 percent of particles with a mass median dia-
meter of less than 0.1 urn will be deposited on nonciliated pulmonary sur-
faces. Deposition on nasal and pharyngeal surfaces becomes important as
»
mass median diameter approaches 1 jUm and rises rapidly to become the domi-
nant deposition site for airborne particles 10 urn in diameter or greater
-------
(Brain and Volberg, 1974). Portions of inhaled asbestos fibers which are
not cleared by microciliary action may remain trapped in the lung for de-
cades (Pooley, 1973; Langer, 1973). However, the chrysotile content of the
lung does not build up as significantly as that of the amphiboles for simi-
lar exposure circumstances (Wagner, et al. 1974).
IV. EFFECTS
A. Carcinogenicity
All commercial forms of asbestos have demonstrated carcinogenic
activity in mice, rats, hamsters, and rabbits. 'Intraperitoneal injection of
various asbestos fibers has produced mesotheliomas in rats and mice (Maltoni
and Annoscia, 1974; Pott and Friedrichs, 1972; Pott, et al. 1976). In rats,
chronic inhalation of various types of asbestos have produced lung carcino-
mas and mesotheliomas (Reeves, et al. 1971, 1974; Gross, et al. 1967;
Wagner, et al. 1974;' Davis, et al. 1978). Intrapleural injection of asbes-
tos fibers has produced mesotheliomas in rats, hamsters, and rabbits (Donna,
1970; Reeves, et al. 1971; Stanton and Wrench, 1972; Stanton, 1973; Wagner,
et al. 1973, 1977; Smith and Hubert, 1974). The oral administration of as-
bestos filter material reportedly caused malignancies in rats (Gibel, et al.
1976) although other feeding studies have produced equivocal results.
Occupational 'exposure -to chrysotile, amosite, anthophyllite, and
mixed fibers containing crocidolite has resulted in high incidences of human
lung cancers (Selikoff, et al. 1979; Seidman, et al. 1979; Enterline and
Henderson, 1973; Henderson and Enterline, 1979; IARC, 1977), Occupational
exposure to crocidolite, amosite, and chrysotile have also been associated'
with a large incidence of pleural and peritoneal mesatheliomas. An excess
of gastrointestinal cancers has been associated in some studies with expo-
sure to amosite, chrysotile, or mixed fibers containing crocidolite (Seli
-------
koff, 1976; Selikoff, et al, 1979; Elmes -and Simpson, 1971; Henderson and
Enterline, 1979; Nicholson, et al. 1979; Seidman, et al. 1979; Newhouse and
Berry, 1979; McDonald and Liddell, 1979; Kogan, et al. 1972).
In the general environment, mesotheliomas have occurred in persons
living near asbestos factories and crocidolite mines and in the household
contacts of asbestos workers (Wagner, et al. 1960; Newhouse and Thomson,
1965). In addition, several studies have implicated asbestos in drinking
water with the development of cancer of the lung and digestive tract cancers
(Mason, et al. 1974; Levy, et al. 1976; Cooper,'" et al. 1978, 1979). There
is convincing evidence to support the contention that asbestos exposure and
cigarette smoking act synergistically to produce dramatic increases in lung
cancer over that from exposure to either agent alone (Selikoff, et al. 1968;
Berry, et al. 1972).
In a study by Hammond, et al. (1979) involving 17,800 insulation
workers, the death rate for non-smokers was 5.17 times that of a non-smoking
control population. The death- rate was 53.24 times that of the non-smoking
control population or 4.90 times the death rate for a comparable group of
non-exposed smokers. Cancers of the larynx, pharynx and buccal cavity in
insulators were also found to be associated with cigarette smoking, together
with some non-malignant asbestos effects such as fibrosis and deaths due to
asbestosis.
B. Mutagenicity
In cultured Chinese hamster cells, chrysotile and crocidolite have
produced genetic damage and morphologic transformation (Sincock and
Seabright, 1975; Sincock, 1977). On the other hand, chrysotile, amosite,
and anthophyllite showed no mutagenic activity toward tester strains of E^
coli or S^ typhimurium (Chamberlain and Tarmy, 1977).
1
-US'-
-------
C. Teratogenicity
Pertinent data on the possible teratogenic effects of asbestos were
not located in the available literature, although transplacental passage of
asbestos fibers has been.reported (Cunningham and Pontefract, 1971, 1973).
0. Other Reproductive Effects
It is not known whether asbestos exposure may impair fertility or
interfere with reproductive success (U.S. EPA, 1979).
E. Chronic Toxicity
The chronic ingestion of chrysotile by1-rats (0.5 mg or 50 mg daily
for 14 months) produced no effects on the esophagus, stomach, or cecum tis-
sue, but histological changes were seen in the ileum, particularly of the
villi (Jacobs, et al. 1978).
The long-term disease entity, asbestosis, results from the inhala-
tion of asbestos fibers and is a chronic, progressive pneumoconiosis. It is
characterized by fibrosis of the lung parenchyma and produces shortness of
breath as the primary symptom. Asbestos has accounted for numerous cases of
occupational disablement during life as well as a considerable number of
deaths among worker groups. In groups exposed at lower concentrations such
as the families of workers, there is less incapacitation and although asbes-
tosis can occur, deaths have not been reported (Anderson, et al. 1976).
Extrapulmonary chronic effects reported include "asbestos corns"
from the penetration of asbestos fibers into the skin. No chronic nonmalig-
nant gastrointestinal effects have been reported.
V. AQUATIC TOXICITY
Pertinent data concerning the effects of asbestos to either fresh-
»
water or marine organisms were not located in the available literature.
-------
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 current Occupational Safety and Health Administration (OSHA)
standard for an 8-hour time-weighted average (TWA) occupational exposure to
asbestos is 2 fibers longer than 5 microns in length per milliliter of air
(2f/ml or 2,000,000 f/m3). Peak exposures of up to 10 f/ml are permitted
for no more than 10 minutes (Fed. Reg., 1972). This standard has been in
effect since July 1, 1976, when it replaced an earlier one of 5 f/ml (TWA),
Great Britain also has a value of 2 f/ml as the accepted level, below which
no controls are required (BOHS, 1968). The British standard, in fact,
served as a guide for the OSHA standard (NIOSH, 1972).
The British standard was developed specifically to prevent asbes-
tosis among working populations; data were felt to be lacking that would
allow for determination of a standard for cancer (BOHS, 1968). Unfor-
tunately, among occupational groups, cancer is the primary cause of excess
death for workers (see "Carcinogenicity" section) with three-fourths or more
of asbestos-related deaths caused from malignancy. This fact has led OSHA
to propose a lower TWA standard of 0.5 f/ml (500,000 f/m3) (Fed. Reg.,
1975). The National Institute for Occupational Safety and Health (NIOSH),
in their criteria document for the hearings on a new standard, have proposed
a value of 0.1 f/ml (NIOSH, 1977). In the discussion of the NIOSH proposal,
»
it was stated that the value was selected on the basis of the sensitivity of
analytical techniques using optical microscopy and that 0.1 f/ml may not
neces
-------
sarily protect against cancer. Recognition that no information exists that
would define a threshold for asbestos carcinogenesis was also contained in
the preamble of the OSHA proposal. The existing standard in Great Britain
has been questioned by Peto (1978), who estimates that asbestos disease may
cause the death of 10 percent of workers exposed at 2 f/ml for a working
lifetime.
The existing federal standard for asbestos emissions into the en-
vironment prohibits "visible emissions" (U.S. EPA, 1975). No numerical
value was specified because of difficulty in monitoring ambient air asbestos
concentrations in the ambient air.or in stack emissions. Some local govern-
ment agencies, however, may have numerical standards (e.g., New York, 27
ng/m ).
No standards for asbestos in foods or beverages exist even though
the use of filtration of such products through asbestos filters has been a
common practice in past years. Asbestos filtration, however, is prohibited
or limited for human drugs (U.S. FDA, 1976).
The draft recommended water quality criterion for asbestos par-
ticles (U.S. EPA, 1979) is derived from the substantial data which exists
for the increased incidence of peritoneal mesothelioma and gastrointestinal
tract cancer in humans exposed occupationally to asbestos. This derivation
assumes that much or all of this increased disease incidence is caused by
fibers ingested following clearance from the respiratory tract. Several
studies allow the association of approximate airborne fiber concentrations
to which individuals were exposed with observed excess peritoneal and gas-
trointestinal cancer. All of the inhaled asbestos is assumed to be even-
*•
tually cleared from the respiratory tract and ingested.
Ml
-------
The draft criterion calculated to-keep the individual lifetime can-
cer risk below ICf , is 300,000 fibers of all sizes/liter. The corres-
ponding mass concentration for chrysotile asbestos is approximately 0.05
ug/1. This criterion has not yet gone through the process of public review;
therefore, there is a possibility that the criterion may-be changed.
B. Aquatic
Because no data are available on the aquatic toxicity of asbestos,
the U.S. EPA (1979) derived no aquatic criteria.
Jrf
-------
ASBESTOS
REFERENCES
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Lancet 2: 47b.
Brain, J.D., and P.A. Volberg. 1974. Models of lung retention
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Chamberlain, M. and E.M. Tarmy. 1977. Asbestos ana glass fibres
in bacterial mutation tests. Mutat. Res. 43: 159.
Cook, P.M. and G.F. Olson. 1979. Ingested mineral fibers: Elimi-
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Cooper, R.C. and W.C. Cooper. 1978. Public health aspects of
asbestos fibers in drinking water. Jour. Am. Water Works Assoc.
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Cooper, R.C., et al. 1978. Asbestos in domestic water supplies
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of Public Health, Univ. Calif. Berkeley, pp. 247.
Cunningham, H.M. and R.D. Pontefract. 1971. Asoestos fioers
in beverages and drinking water. Nature (Lond.) 232: 332.
Cunningham, H.M. and R.D. Pontefract. 1973. Asbestos fibers
in beverages, drinking water and tissues: their passage through
the intestinal wall and movement through the body. Jour. Assoc.
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Cunningham, H.M., et al. 1976. Quantitative relationship of
fecal asbestos to asbestos exposure. Jour. Toxicol. Environ.
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f-
Davis, J.M.G., et al. 1978. Mass and number of fioecs in the
pathogenesis o£ asbestos-related lung disease in rats. Br. Jour.
Can. 37: o7j.
-------
Donna, A. 1970. Tumori sperimentali da amiano di crisotilo,
crocidolite e amosite in ratto Sprague-Dawley. Med. Lavoro.
61: 1.
Elmes, P.C. ana M.J.C. Simpson. 1971. Insulation workers in
Belfast. III. Mortality 1940-66. Br. Jour. Ind. Med. 28: 226.
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respiratory cancer in the asbestos industry. Arch. Environ.
Health. 27: 312.
Federal Register. 1972. Standard for exposure to asbestos dust.
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Federal Register. 1975. Occupational exposure to asbestos;
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".
Gibel, W., et al. 1976. Tierexperimentelle untersuchungen uber
eine kanzerogene wirkung von asbestfiltermaterial nach oraler
aufnahrae. Arch. Geschwulstforsch. 46: 437.
Gross, P., et al. 1967. Experimental asbestosis: The develop-
ment of lung cancer in rats with pulmonary deposits of chrysotile
asbestos dust. Arch. Environ. Health 15; 343.
Gross, P., et al. 1974. Ingested mineral fibres. Do they pene-
trate tissue or cause cancer? Arch. Environ. Health 29: 341.
Hammond, E.G., et al. 1979. Cigarette smoking and mortality
among U.S. asbestos insulation workers. Ann. N.Y. Acaa. Sci.
(In press).
Henderson, V.I. and P.E. Enterline. 1979. Asbestos exposure
factors associated with excess cancer and respiratory disease
mortality. Ann. N.Y. Acad. Sci. (In press).
IARC Monographs on the Evaluation of Carcinogenic Risk of Chemi-
cals to Man. 1977. Asbestos. Vol. 14.
Jacobs, R., et al. 1978. Light and electron microscope studies
of the rat digestive tract following prolonged and short-term
ingestion of chrysotile asbestos. Br. Jour. Exp. Path. 59: 443.
Kay, G. 1973. Ontario intensifies search for asbestos in drinking
water. Water Pollut. Control 9: 33.
Kogan, F.M., et al. 1972. The cancer mortality rate among workers
of asbestos industry of the Urals. Gig. i Sanit. 37; 29.
Langer, A.M., et al. 1973. Identification of asbestos in .human
tissues. Jour. Occup. Med i5: 287.
-------
Levy, B.S., et al. 1976. Investigating possible effects of
asbestos in city water: Surveillance 'of gastrointestinal cancer
incidence in Duluth, Minn. Am. Jour. Epidemiol. 103: 362.
Lindell, K.v.' 1972. Biological effects of asbestos. Int.
Agency Res. Cancer, Lyon, France.
Maltoni, C. and C. Annoscia. 1974. Mesotheliomas in rats following
the intraperitoneal injection of crocidolite. In: W. Davis and
C. Maltoni, eds. Advances in tumour prevention, detection and
characterization. Vol. 1. Characterization of human tumours.
Excerpta Medica, Amsterdam.
Mason, T.J., et al. 1974. Asbestos-like fibers in Duluth water
supply. Relation to cancer mortality. Jour. Am. Med. Assoc.
228: 1019.
».
McDonald, J.C. and D.K. Liddell. 1979. Mortality in Canadian
miners and millers exposed to chrysotile. Ann. N.Y. Acad. Sci.
(In press).
Millette, J. 1979. Health Effects Res. Lab. (Personal communi-
cation) .
National Institute of Occupational Safety and Health. 1972.
Criteria for a recommended standard...Occupational exposure to
asbestos. DHEW (NIOSH) Pb. No. 72-10267.
Natioal Institute of Occupational Safety and Health. 1977.
Revised recommended asbestos standard. DHEW (NIOSH) Pub. No.
77-169.
Newhouse, M.L. and G. Berry. 1979. Patterns of disease among
long-term asbestos workers in the United Kingdom. Ann. N.Y.
Acaa. Sci. (in press).
Newhouse, M.L. and H. Thomson. 1965. Meaothelioma of pleura
and peritoneum following exposure to asbestos in the London area.
Dr. Jour. Ina. Mea. 22: 261.
Nicholson, W.J. 1971. Measurement of asbestos in ambient air.
Final report, Contract CPA 70-92. Natl. Air Pollut. Control
Admin.
Nicholson, W.J. and F.L. Pundsack. 1973. Asbestos in the envi-
ronment. Page 126 in P. Bogovski, et al. eds. Biological effects
of asbestos. lARC~~Sci. Publ. No. 8. Int. Agency Res. Cancer,
Lyon, France.
Nicholson, W.J., et al. 1971. Asbestos air pollution in New
York City. Page 136 ir^ H.M. England and W.T. Barry, eds. 'Proc.
Second Clean Air Cong. Academic Press, New York.
Nicholson, W.J., et al. 1975. Asbestos contamination of the
air in public buildings. Final report, Contract No. 63-0^-1346.
U.S. Environ. Prot. Agency.
-------
Nicholson, W.J., et al. 1379. Mortality experience of asbestos
factory workers: Effect of differing intensities of asbestos
exposure. Environ. Res. (In press}.
Peto, J. 1978. The hygiene standard for asbestos. Lancet 8062: 484.
Pooley, P.O. 1973. Mesothelioma in relation to exposure. Page
222 _in P. Bogovski, et al. eds. Biological effects of asbestos.
IARC Sci. Publ, No. 8. Int. Agency Res. Cancer, Lyon, France.
Pott, F. and K.H. Friedrichs, 1972. Tumoren der ratte nach
i.p.-injektion faserformiger staube. Naturwissenschaften. 59: 318.
Pott, F., et al. 1976. Ergebnisse aus tierversuchen 2ur kanzero-
genen wirkung faserformiger staube und ihre deutung im hinolick
auf die tumorentstehung beim menschen. Zbl. Bakt. Hyg., I Abt.
orig. B. 162: 467.
Reeves, A.L., et al. 1971. Experimental asbestos carcinogenesis.
Environ. Res.- 4: 496.
Reeves, A.L., et al. 1974. Inhalation carcinogenesis from various
forms of asbestos. Environ. Res. 8: 178.
Sebastien, P., et al. 1976. Les pollutions atraospheriques urbanies
par 1'asbeste. Rev. franc. Mai. resp. 4: 51.
Seidman, H., et al. 1979. Long-term observation following short-
term employment in an amosite asbestos factory. Ann. N.Y. Acad.
Sci. {In press).
Selikoff, I.J. 1976. Lung cancer and mesothelioma during prospec-
tive surveillance of 1249 asbestos insulation workers, 1963-1974.
Ann. N.Y. Acad. Sci. 271: 448.
Selikoff, I.J., et al. 1968. Asbestos exposure, smoking and
neoplasia. Jour. Am. Med. Assoc. 204: 106.
Selikoff, I.J., et al. 1979. Mortality experience of insulation
workers in the United States and Canada, 1943-1977. Ann. N.Y.
Acad. Sci. (In press).
Sincock, A.M. 1977. Iri vitro chromosomal effects of asbestos
and other materials. Ir± Origins of human cancer. Cold Spring
Harbour, 1976.
Sincock, A.M. and M. Seabright. 1975. Induction of chromosome
changes in Chinese hamster cells by exposure to asbestos fibers.
Nature (Lond.) 257: 56.
*
Smith, W.E. ana D.D. Hubert. 1974. The intrapleural route as
a means for estimating carcinogenicity, Pages 92-101 Ln E. Karbe
and J.F. Park, eds. Experimental lung cancer. Springer-Verlag,
Berlin. 92-101..
-------
Speil, S. and J.P. Leineweber. 1969. Asbestos minerals in modern
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Stanton, M.F. '1973. Some etiological considerations of fibre
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effects o£ asbestos. Int. Agency Res. Cancer. IARC Sci. Publ.
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induction with asbestos and fibrous glass. Jour. Natl. Cancer
Inst. 48: 797.
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Environmental.Protection Agency, Washington, D.C.
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Wagner, J.C., et al. 1960. Diffuse pleural mesothelioma and
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No. 13
Barium
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.
-------
BARIUM
SUMMARY
Water-soluble barium compounds are highly toxic to man. Fish and lower
species of marine organisms have been shown to bioaccumulate barium. The con-
centration of barium in sea water ranges around 20 ug/L, while that of drinking
water averages about 6 ug/L.
Soluble barium salts have a high acute toxicity. Small amounts of barium
can accumulate in the skeleton of humans and animals. Barium salts are strong
muscle stimulants: acute intoxication generally results in uncontrolled
contractions followed by partial or complete paralysis. Cardiac disturbances
including arrythmias can also occur. Barium dusts are irritant to nose,
throat and eyes. Baritosis (pneumoconiosis) occurs following chronic
inhalation of (fine) barium dusts. Barium sulfate used in barium enemas,
swallows and artificial orthopedic bones can result in tissue injury following
solubilizatioh of the barium sulfate and/or soluble impurities. Potassium
acts.as an antagonist for barium induced cellular disturbances. The TWA
for exposure to soluble barium compounds is 0.5 mg/m .
I. INTRODUCTION
Barium (Ba; atomic weight 137.34) is a yellowish-white metal of the alkaline
earth group. It is relatively soft and ductile and may be worked readily.
Barium has a melting point of 729°C and a boiling point of 1640°C; its density
is 3.51 g/cm3 (Kunesh 1978).
Barium characteristically forms divalent compounds. At room temperature,
it combines readily and exothermically with oxygen and the halogens. It reacts
vigorously with water to form barium hydroxide, Ba(OH)? (Kunesh 1978).
Barium occurs in nature chiefly as barite, crude BaSO,, and as witherite,
a form of BaCO_, both of which are highly insoluble salts. Only barite is
mined in this country (Kirkpatrick 1978).
A review of the production range (includes importation) statistics for
barium (CAS. No. 7440-39-3) whicb are listed in the initial TSCA Inventory,
(U.S. EPA 1979) has shown that between 100,000 and 900,000 pounds of this
chemical were produced/imported in 1977*.
*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).
-------
C. Environmental Occurrence
The flow of barium in the United States has been traced for the year 1969,
during which time consumption of barium totaled 1.87 billion pounds. It was
estimated that 30.8 million pounds of barium were emitted to the atmosphere.
Nearly 18 percent of the emissions resulted from the processing of barite, more
than 28 percent from chemical production, 26 percent from the combustion of
coal, and 23 percent from the manufacture of miscellaneous end products
(U.S. EPA 1972).
The concentration of barium in sea water is generally accepted as about
20 ug/L, with lower concentrations in the surface waters than at greater depths.
Barium ions are generally removed from solution quite rapidly by adsorption,
sedimentation and precipitation (U.S. EPA 1973). Concentrations of barium in
this country's drinking water supplies generally range from less than 0.6 ug/L
to about 10 ug/L, although a few midwestern and western states have had upper
limits of 100 to 300 ug/L (U.S. EPA 1976).
Due to the common use of barite as a weighting agent in drilling muds,
the resultant contamination of sediments near drilling sites was studied. The
average content of barium in benthic sediments from the Southern California Bight
was 637 parts per million (ppm), with a range from 43 to 1899 ppra. This area
includes active drilling sites where barium contamination is expected. The
concentration values were compared with the average 879 ppm barium found in
mainland intertidal sediments and the 388 ppm determined in the channel island
intertidal sediments. The lower barium content of the island sediments was
attributed to the volcanic soil of the islands; however, the higher barium
concentration of the mainland could not be traced to either natural or anthro-
pogenic origin. Due to variations in soil sources it is questionable whether
barium concentrations determined elsewhere could be used as reference values for
this study (Chow 1978).
In two studies correlating trace metal concentrations in the environment
with that in scalp hair of the inhabitants, barium was measured in the house
dust collected in four communities. Geometric mean values of barium determined
in house dust samples from the New York City area were as follows: 65.2 ug Ba/g
dust in Riverhead, 137.6 ug/g in Queens, and 312.4 ug/g in the Bronx (USEPA, 1978b)
The geometric mean value for barium measured in house dust in Ridgewooti, New
Jersey was 330.0 ug/g (U.S. EPA 1978c).
_ /"
-------
Barium and its compounds are used industrially as weighting agents in
oil and gas well drilling muds; as coloring agents in glass, ceramics, paint,
and pigments; as filler in rubber; and as antismoking agents in diesel
fuel (U.S. EPA 1972; NAPCA 1969). In medicine, barium sulfate is used as
an x-ray contrast medium because of its extreme insolubility and its ability to
absorb x-rays (Kirkpatrick 1978; U.S. EPA 1978a).
II. EXPOSURE
A. Environmental Fate
Due to the high reactivity of barium, it is not found in its elemental
state in the environment. In sea water, the naturally present sulfate and
carbonate tend to precipitate any water-soluble barium components. Thus, the
sediment usually has a higher concentration of barium than its corresponding
water source (Guthrie 1979).
B. Bioconcentration
Due to the toxicity of soluble barium salts to man, the bioaccumulation
of the element has been a concern. Barium can be concentrated in goldfish by
a factor of 150. Concentration factors for barium listed in one study are
17,000 in phytoplankton, 900 in zooplankton, and 8 in fish muscle (U.S. EPA 1973).
Thus, ingestion of fish by man can be a source of barium exposure.
Another study conducted on various species of marine organisms produced the
following results (Guthrie 1979): Barnacles bioaccumulated 'about
five times greater concentration of barium than was in the water, while oysters
and clams contained concentrations of the element similar to that present in
the water. Crabs and polychaetes were also analyzed for barium and were found
to contain a significantly smaller quantity than that present in the sediment
on which they dwell. However, no significant differences were noted between
the concentration of barium in the two organisms and the concentrations in
the water column.
In man, studies have been conducted to determine a correlation between barium
in the environment, measured as house dust, and the concentration of barium found
in scalp hair of the inhabitants. A significant positive correlation*has been
determined between the geometric mean concentrations of the element in house dust
and hair. Other covariants of significant value measured in the studies were sex,
hair length, and, in children less than 16 years old, age (U.S. EPA 1978b; U.S.
EPA 1978c).
-------
Ill. PHARMACOKINETICS
Soluble barium is retained by muscle tissue for about 30 hours, after
which the amount of retained barium decreases slowly (NAPCA 1969). Small
amounts of barium become irreversibly deposited in the skeleton. However,
the acceptance level is limited, as quantitative analysis of human bone
reveals no accumulation of barium from birth to death.... Barium levels
averaged 7 ug/g ashed bone. Very little barium is retained by the liver,
kidneys, or spleen, and practically none by the brain, heart, or hair.
Transient high concentrations are seen in the liver with lesser amounts in
lung and spleen following acute experimental dosing.
Barium administered orally or intraperitoneally as BaCl_ to weanling
male rats at doses of 1, 5, 25, or 125 mg/kg was taken up rapidly by the
soft tissues (30 mins), showed slow uptake by the skeleton (2 hrs) and was
excreted primarily in the feces (Clary and Tardiff, 1974). No retention
data were reported.
Pulmonary clearance rates of inhaled radioactive Ba salts ranged from
I i
several hours for the soluble Bad to hundreds of days for Ba in fused
clay . Large amounts of barium were excreted in the feces; a lesser amount
was excreted in the urine. Although BaSO, is "insoluble" in water, 50% of
133
BaSO, dissolved in a simulated biological fluid within 2-3 days, indicating
that solubilization is relatively rapid.
IV. HEALTH EFFECTS
A. Carcinogenicity
Bronchogenic carcinoma developed in rats injected with radioactive S
(unspecified dose) labelled barium sulfate (Patty 1963). BaSO, powder (particle
size undefined) injected intrapleurally in female and male mice produced a
mesothelioma in only 1 out of 30 animals. No other pathological lesions were
investigated or reported. Saline controls (32) resulted in no mesotheliomas.
Barium sulfate had an oncogenic potency similar to that of glass powder and
aluminum oxide. It therefore appears likely that the observed tumor vtas due
to foreign-body-oncogenesis (Wagner).
- ISO -
-------
B. Acute and Chronic Toxicity
The soluble salts of barium are highly toxic when ingested* Barium chloride
and barium carbonate, two of the soluble compounds, have been reported to
cause toxic symptoms of a .severe but usually nonfatal degree. Seven grams of
barium chloride (%4.5 g Ba) taken orally produced severe abdominal pain and
near-collapse, but not death (NAPCA 1969). However, Patty (1963) indicates
800 to 900 rag of barium chloride (550-600 mg Ba) to be a fatal.human dose.
Few cases of industrial poisoning from soluble barium salts have been reported:
Most of these have been cases of accidental ingestion (NAPCA 1969).
Ingested soluble barium compounds produce a strong stimulating effect on
all muscles of the body. The effect on the heart muscle is manifested by
irregular contractions followed by arrest of systolic action. Gastrointes-
tinal effects include vomiting and diarrhea. Central nervous system effects
observed include violent tonic and clonic spasms followed in some cases by
paralysis (NAPCA 1969).
Death resulting from barium exposure may occur in a few hours or a few
days, depending on the dose and solubility of the barium compound. A death
attributed to barium oxide poisoning has been reported. However, .the usual
effect of exposure to dusts and fumes of barium oxide, barium sulfide, and
barium carbonate is irritation of eyes, nose, throat and the skin (NAPCA 1969).
Some of the BaSO, used in orthopedic bone cements has been shown to escape
into surrounding tissues (Rae 1977). Mouse peritoneal macrophages exposed to
barium sulfate (10 particles.of unspecified size/macrophage) for periods up
to 144 hours showed a marked cytoplasmic vacuolization. Following cessation
of exposure only partial recovery occurred. No cell membrane damage was
observed (Rae 1977). The use of barium sulfate in barium swallows and
enemas \resulted in severe toxic ""affects on rupture of the intestinal tract
(Gardiner and Miller 1973, Bayer et al. 1974).
Inhalation of barium compounds is known to cause a benign respiratory
affliction (pneumoconiosis) called baritosis, which has been reported in
workers exposed to finely divided barium sulfate in Italy, in barite miners
in the United States, Germany, and Czechoslovakia, and among workers exposed
to barium oxide. Generally, baritosis produces no symptoms of emphysema or
bronchitis, and lung function tests show no respiratory incapacity, although
some afflicted workers complain of dyspnea upon exertion. In the majority
of cases nodulation disappears if exposure to the barium compound is stopped
(NAPCA 1969). Aspirated BaSO, can result in granulomas of the lung and other
sites in man (Patty 1963).
-------
Suicidal ingestion of a facial depilatory containing 15.8 g of BaS
resulted in paralysis of head, neck, arms, and trunk as well as respiratory
paralysis. Therapy with MgSO,, saline and potassium resulted in recovery
within 24 hours (Gould et al. 1973).
Acute oral toxicity values for barium carbonate were: mouse LD = 200 mg/
kg; rat LD = 50-200 mg/kg, LD5Q = 1480 + 340 mg/kg; rabbit LD = 170-300 mg/kg.
For barium chloride oral toxicity values were: mouse LD = 7-14 mg/kg; rat
LD = 355-533 mg/kg; rabbit LD = 170 mg/kg; dog LD = 90 mg/kg. For barium
flouride the acute oral LD for guinea pigs was 350 mg/kg (NAPCA 1969).
C. Other Relevant Information
Potassium acts as an in vitro antagonist of barium. Cardiac effects
such as arrythmias exerted by barium are also reversed rapidly by potassium.
Barium induces hypokalemia apparently by promoting a shift of potassium
from plasma into cells. The prolongation of action-potentials and depolariza-
tion of smooth and skeletal muscle by barium are thought to be due to
barium induced decreases in potassium conductance." In addition,, barium can
replace sodium to produce and/or prolong action potentials and can also
substitute for calcium in neurosecretory processes as described below (Peach 1975).
Barium chloride has been shown to cause arterial contractions in
_4
in vitro preparations of human digital arteries at concentrations of 10 to
10~ M (Jauernig and Moulds 1978). This activity was approximately 40 to 50
_2
fold more than that of potassium chloride/ At BaCl- concentrations above 10 M
contractions developed very slowly. The action of BaCl9 was inhibited by
-2
veraparmil, a calcium antagonist, at BaCl_ contractions below 10 M.
-------
V. AQUATIC TOXICITY
'•it.
;:-n
According Co an EPA report, experimental data indicate that in fresh
and marine waters, the soluble barium concentration would need to exceed
50 mg/L before toxicity to aquatic life would be expected (U.S. EPA 1976).
Furthermore, in most natural waters, sufficient sulfate or carbonate is present
to precipitate barium in the water to a virtually insoluble, non-toxic
compound.
Soluble barium salts, however, are quite toxic. It has been reported
that 10 to 15 mg/L of barium chloride (9.9 mg/L Ba) was lethal to an aquatic
plant and two species of snails (species and origin unspecified). Bioassay
with this same barium salt showed the LC for Coho Salmon to be 158 mg/L
(104 mg/L Ba) (U.S. EPA 1973).
VI. GUIDELINES
A. Hunan Health
The OSHA Time Weighted Average for exposure to barium (soluble compound)
is 0.5 mg/m3 (29 CFR 1910:1000).
B. Aquatic
There is no established criterion for barium in the aquatic environment.
The U.S. EPA (1973) suggests, however, that concentrations of barium equal
to or exceeding 1.0 mg/L constitute a hazard in the marine environment, and
levels less than 0.5 mg/L present minimal risk of deleterious effects.
-------
References
Bayer HP, Buhler F and Ostermeyer J, 1974. On the distribution of interstitial
and parenteral administered barium sulfate in the organism. Z. Rechtsmedizin
74: 207-215 (1974). (Ger.)
Chow T, Earl J, Reeds J, Hansen N, land Orphan V, 1978. Barium content of marine
sediments near drilling sites: A potent pollutant indicator. Marine Pollution
Bulletin. 9:97-99.
Clary JJ,.and Tardiff RG, 1974. The absorption, distribution and excretion of
orally administered 133-BaCl7 in weanling male rats. Toxicol. Appl. Pharmacol.
27:139.
Gardiner H and Miller RE, 1973. Barium peritonities. Am. J. Surgery 125:350-352.
Gould DB, Sorrell MB and Lupariello AD. 1973. Barium sulfide poisoning. Arch.
Jut. Med. 132:891-894.
Guthrie RK, Ernst M, Cherry D, Murray H, 1979. Biomagnification of heavy metals
by organisms in a marine microcosm. Bull. Environm. Contam. Toxicol. 21:53-61.
Jaueraig RA and Moulds RFW. 1978. A human arterial preparation for studying the
effects of vasoactive agents. Circ. Res. 42:363-3.68.
Kirkpatrick T. 1978. Barium Compounds In Kirk-Othmer1s Encyclopedia of
Chemical Technology, 3rd edition. John Wiley and Sons, Inc. New York. 3:463-479.
Kunesh CJ. 1978. Barium In Kirk-Othmer's Encyclopedia of Chemical Technology,
3rd edition. John Wiley and Sons, Inc. New York. 3:458-463.
NAPCA. 1969. Air Pollution Aspects of Barium and Its Compounds. National
Air Pollution Control Administration. PB 188 083.
Patty FA, Ed. 1963. Industrial Hygiene and Toxicology. Vol II. Toxicology.
2nd Edition. Interscience Publishers, New York; pp. 998-1002.
Peach MJ. 1975. Cations: Calcium, Magnesium, Barium, Lithium and Ammonium.
In: The Pharmaceutical Basis of Therapeutics. Goodman LS and Gilman A, Eds.
MacMillan Publishing Co., Inc. New York, pp. 791.
Rae.T, 1977. Tolerance of mouse macrophages in vitro to barium sulfate used in
orthopedic bone cement. Bioraed. Mater. Res. 11:839-846.
U.S. Dept. of Labor. General Industry Standards Table Z-l. 29 CFR 1910:1000.
U.S. EPA 1972. National Inventory of Sources and Emissions - Barium, Baron,
Copper, Selenium, and Zinc 1969-Barium Section I. PB 210 676. .
U.S. EPA 1973. Water Quality Criteria 1972. EPA-R-373-033.
- ; JTH -
-------
U.S. 1976. Quality Criteria for Water. EPA-440/9-76-023.
U.S. EPA 1978a. Source Assessment: Major Barium Chemicals. EPA-600/2-78-
0046. PB 280 756.
U.S. EPA. 1978b. Human Scalp Hair: An Environmental Exposure Index for Trace
Elements. I. Fifteen Trace Elements in New York, N.Y. (1971-1972). EPA-600/1-
78-037a. PB 284 434.
U.S. EPA. 1978c. Human Scalp Hair: An Environmental Exposure Index for Trace
Elements. II. Seventeen Trace Elements in Four New Jersey Communities (1972).
EPA-600/l-78-037b. PB 294 435.
U.S. EPA 1979. Toxic Substances Control Act Chemical Substance Inventory,
Production Statistics for Chemicals on the Non-Confidential Initial TSCA Inventory.
Wagner JC, Berry C, and Timbrell V. 1973. Mesotheliomas in rats after inocula-
tion with asbestos and other materials. Br. J. Cancer 28:173-185.
-------
No. 14
Benzal Chloride
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.
-------
8ENZAL CHLORIDE
Summary
Benzal chloride has been reported to induce papillomas, carcinomas, and
leukemia in mice. Details of this work were not available for assessment.
Mutagenic effects of benzal chloride exposure have been reported in
Salmonella, Bacillus, and E^ coli.
There is no available information an the teratogenic or adverse repro-
ductive effects of the compound.
-------
I. INTRODUCTION
Benzal chloride, CAS registry number 98-87-3, is a fuming, highly re-
fractive, colorless liquid. It is made by free radical chlorination of
toluene and has the following physical and chemical properties (Windholz,
1976; Verschueren, 1977):
Formula: C7H6C12
Molecular Weight: 161.03
Melting Point: -16°c
Boiling Point: 207°C
Density: 1.25614
Vapor Pressure: 0.3 torr i 20°C
Solubility: alcohol, ether
insoluble in water
Benzal chloride is used almost exclusively for the manufacture of ben-
zaldehyde. It can also be used to prepare cinnamic acid and benzoyl chlor-
ide (Sidi, 1971).
II. EXPOSURE
A. Water
Benzal chloride is converted to benzaldehyde and hydrochloric acid
on contact with water (Sidi, 1971).
B. Food
Pertinent data could not be located in the available literature.
C. Inhalation
It is likely that the only source of benzal chloride in the air is
production facilities. The compound will hydrolyze in moist air to give
benzaldehyde and hydrochloric acid. Inhaled benzal chloride will probably
produce effects similar to those of inhaled hydrogen chloride.
»
0. Dermal
Benzal chloride is irritating to the skin (Sidi, 1971).
-/£*»-
-------
III. PHARMACOKINETICS
Pertinent data on the pharmacokinetics of benzal chloride could not be
located in the available literature.
IV. EFFECTS
A. Carcinogenicity
In a study of Matsushito, et al. (1975) benzal chloride, along
with several other compounds, was found to induce carcinomas, leukemia, and
papillomas in mice. The details of the study were not available, but benzal
chloride was shown to possess a longer latency period than benzotrichloride
before the onset of harmful effects.
B. Mutagenicity
Yasuo, et al. (1978) tested the mutagenicity of several compounds
including benzal chloride in microbial assay systems which include the rec-
assay using Bacillus subtilIs, the reversion assay using E,_ coll, and the
Ames assay using Salmonella typhlmurium, with or without metabolic activa-
tion. Benzal chloride was positive in the rec-assay without activation and
in the reversion assays using ^ typhimurium and §^ cpli. with metabolic
activation.
C. Teratogenicity, Other Reproductive Effects and Chronic Toxicity
Pertinent data could not be located in the available literature.
D. Acute Toxicity
The oral LD50is for mice ancj rats exposed to benzal chloride are
2,4£2 mg/kg and 3,249 mg/kg, respectively (NIOSH, 1978).
V. AQUATIC TOXICITY
Pertinent aquatic toxicity data could not be located in the available
literature.
t
-JtaO-
-------
VI. EXISTING GUIDELINES AND STANDARDS
There are no existing guidelines or standards for exposure to benzal
chloride.
-------
REFERENCES
Matsushito, H., et al. 1975. Carcinogenicities of the related compounds in
benzoyl chloride production. 49th Annual Meeting Japan Ind. Hyg. Soc., Sap-
pro, Japan, p. 252.
National Institute for Occupational Safety and Health. 1978. Registry of
Toxic Effects of Chemical Substances. NIOSH, DHEW Publ. No. 79-100.
Sidi, H. 1971. Benzyl Chloride, Benzal.Chloride and Benzotrichloride. In:
Kirk-Otnmer Encyclopedia of Chemical Technology, 2nd ed. Vol. 5, John Wiley
and Sons, New York. p. 281.
Verschueren, K, 1977. Handbook of Environmental Data on Organic Chemicals.
Van Nostrand Reinhold Co., New York. p. 127.
Windholz, M. (ed.) 1976. The Merck Index. 9th ed., Merck and Co., Inc.,
Rahway, New Jersey.
Yasuo, K., et al. 1978. Mutagenicity of benzotrichloride and related com-
pounds. Mutation Research 58: 143.
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No. 15
Benzene
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, B.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.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated
benzene and has found sufficient evidence to indicate that
this compound is carcinogenic.
-------
BENZENE .
Summary
Benzene is a widely used chemical. Chronic exposure to it causes
hematological abnormalities. Benzene is not mutagenic to bacteria, but
recent evidence shows it to be carcinogenic in animals. Also, benzene has
been shown to be leukemogenic in humans. There is suggestive evidence that
benzene may be teratogenic and may cause reduced fertility.
Benzene has been shown to be acutely toxic "to aquatic organisms over a
concentration range of 5,800 to 495,000 /jg/1. The marine fish striped bass
was the most sensitive species tested.
-------
BENZENE .
I. INTRODUCTION
This profile is based on the draft Ambient Water Quality Criteria Docu-
ment for Benzene (U.S. EPA, 1979).
Benzene (Benzol CfiH6; molecular weight 78.1) is a volatile, color-
less, liquid hydrocarbon produced principally from coal tar distillation,
from petroleum by catalytic reforming of light naphthas, and in coal pro-
cessing and coal coking operations (Weast, 1972; Ayers and Muder, 1964; U.S.
EPA, 1976a). Benzene has a boiling point of -60.1°C, a melting point of
5.5°C, a water solubility of 1,780 mg/1 at 25°C, and a density of
0.87865 g/ml" at 20°C. The broad utility spectrum of benzene includes its
use as: an intermediate for synthesis in the chemical and pharmaceutical
industries, a thinner for lacquer, a degreasing and cleaning agent, a sol-
vent in the rubber industry, an antiknock fuel additive, a general solvent
in laboratories and in the preparation and use of inks in the graphic arts
industries.
Current production of benzene in the U.S. is over 4 million metric tons
annually, and its use is expected to increase when additional production
facilities become available (Fick, 1976).
II. EXPOSURE
A. Water
A report by the National Cancer Institute (1977) noted benzene
levels of 0.1 to 0.3 ppb in four U.S. city drinking water supplies. One
measurement from a groundwater well in Jacksonville, Florida showed levels
higher than 100 ppb. One possible source of benzene in the aquatic environ-
ment is from cyclings between the atmosphere and water (U.S. EPA, *1976b).
Concentrations of benzene upstream and downstream from five benzene
-------
production or consumption plants ranged from less than 1.0 to 13.0 ppb, with
an average of 4.0 ppb (U.S. EPA, 1977a).'
B. Food
Benzene has been detected in various food categories: fruits,
nuts, vegetables, dairy products, meat, fish, poultry, eggs, and several
beverages (Natl. Cancer Inst., 1977). NCI estimated that an individual
might ingest as much as 250 ug/day from these foods. The U.S. EPA (1979)
has estimated the weighted average bioconcentration factor of benzene for
the edible portion of fish and shellfish consumed by Americans to be 6.9.
This estimate is based on the octanol/water partition coefficient of benzene.
C. Inhalation
The respiratory route is the major source of human exposure to ben-
zene, and much of this exposure is by way of gasoline vapors and automotive
emissions. American gasolines contain an average of 0.8 percent benzene (by
weight) (Goldstein, 1977a), and automotive exhausts contain an average of 4
percent benzene (by weight) (Howard and Durkin, 1974). Concentrations of
benzene in the ambient air of gas stations have been found to be 0.3 to 2.4
ppm (Natl. Acad. Sci/Natl. Res. Council, 1977). Lonneman and coworkers.
(1968) measured an average concentration of 0.015 ppm in Los Angeles air
with a maximum of 0.057 ppm. The rural background level for benzene has
been reported as 0.017 ppb (Cleland and Kingsbury, 1977).
III. PHARMACOKINETIC5
A. Absorption
The respiratory absorption of benzene by humans has been measured
several times and found to be 40 to 50 percent retained on exposures to 110
r
ppm or less (Srbova, et al. 1950; Teisinger, at al. 1952; Hunter and Blair,
1972; Nomiyama and Nomiyama, 1974). Absorption was slightly less efficient,
-------
28 to 34 percent, on exposure to 6,000 ppm (Duvoir, et al. 1946).
Deichmann, et al. (1963) demonstrated that rats exposed to benzene (44 to 47
ppm) for long periods of time maintained blood benzene levels of approxi-
mately 4.25 mg/1.
B. Distribution
Free benzene accumulates in lipid tissue such as fat and bone
marrow, and benzene metabolites accumulate in liver tissue and bone marrow
(U.S. EPA, 1977b).
C. Metabolism
Benzene is metabolized by the mixed-function oxidase system to pro-
duce the highly reactive arene oxide (Rusch, et al. 1977). Arene oxide can
spontaneously rearrange to form phenol, undergo enzymatic hydration followed
by dehydrogenation to form catechol or a glutathione derivative, or bind
covalently with cellular macromolecules. Evidence has accumulated that a
metabolite of benzene is responsible for benzene toxicity, in light of the
fact that a protection from benzene toxicity is afforded by inhibitors of
benzene metabolism (Nomiyama, 1964; Andrews, et al. 1977). The specific
metabolite that produces benzene toxicity has not yet been identified, but
likely candidates are benzene oxide, catechol, and hydroquinone, or the cor-
responding semiquinones (U.S. EPA, 1977b).
D. Excretion
Phenol measurement (free plus combined) of the urine of human vol-
unteers indicated that 50 to 87 percent of the retained benzene was excreted
as phenol (Hunter and Blair, 1972). The highest concentration of phenol was
found in the urine within about 3 hours from termination of exposure.
*
Elimination via the lungs was no more than 12 percent of the retained dose.
-------
IV. EFFECTS
A. Carcinogenicity
On subcutaneous, dermal, oral, and inhalation exposure of rats and
mice to benzene, animal experiments have failed to support the view that
benzene is leukemogenic (U.S. EPA, 1979). Recent evidence suggests, how-
ever, that benzene is an animal carcinogen (Maltoni and Scarnato, 1979).
The evidence that benzene is a leukemogen for man is convincing and has re-
cently been reviewed by the Natl. Acad. Sci./Natl. Res. Coun. (1976), Natl.
Inst. Occup. Safety and Health (1977), and U.SJ- EPA (1977b). Vigliani and
Saita (1964) calculated a 20-fold higher risk of acute leukemia in workers
in northern Italy exposed to benzene. In some studies of acute leukemia
where benzene exposure levels have been reported, the concentrations have
generally been above 100 ppm (Aksoy, et al. 1972, 197Aa,b, 1976a,b; Vigliani
and Fourni, 1976; Vigliani and Saita, 1964; Kinoshita, et al. 1965; Sellyei
and Kelemen, 1971). However, other studies have shown an association of
leukemic evidence to benzene levels less than 100 ppm (Infante et al., 1977;
Ott et al., 1978).
6. Mutagenicity
Benzene has not shown mutagenic activity in the
Salmonella/microsome in vitro bioassay (Lyon, 1975; Shahin, 1977; Simmon, et
al. 1977).
C. Teratogenicity
With rats exposed to 100 to 2,200 ppm benzene during days 6 to 15
of gestation some skeletal deformities were observed in their offspring
(Amer. Pet. Inst., 1978). Pregnant mice given single subcutaneous injec-
#
tions of benzene (3 ml/kg) on days 11 to 15 of gestation produced fetuses
-/7G-
-------
with cleft palates, agnathia, and microagnathia, when delivered by caesarean
section on day 19 (Watanabe and Yashida, 1970).
0. Other Reproductive Effects
Gofmekler (1968) found complete absence of pregnancy in female rats
exposed continuously to 209.7 ppm benzene for 10 to 15 days prior to impreg-
nation. One of ten rats exposed to 19.8 ppm exhibited resorption of em-
bryos. The number of offspring per female exhibited an inverse relationship
to benzene exposure levels from 0.3 to 209.7 ppm.
E. Chronic Toxicity
In humans, pancytopenia (reduction of blood erythrocytes, leuko-
cytes, and platelets) has clearly been related to chronic benzene exposure
(Browning, 1965; Goldstein, 1977b; Intl. Labour Off., 1968; Snyder and
Kocsis, 1975). Also, impairment of the immunological system has been re-
ported with chronic benzene exposure (Lange, et al. 1973a; Smolik, et al.
1973). Wolf, et al. (1956) reported that the no-effect level for blood
changes in rats, guinea pigs, and rabbits was below 88 ppm in the air when
the animals were exposed for 7 hours per day for up to 269 days.
F. Other Relevant Information
In rabbits and rats injected subcutaneously with 0.2 mg/kg/day ben-
zene, the frequency of bone marrow mitosis with chromosomal aberrations in-
creased from 5.9 percent to 57.8 percent after an average of 18 weeks
(Kissling and Speck, 1971; Dobrokhotov, 1972). In patients with benzene
induced aplastic anemia, lymphocyte chromosome damage, i.e., abnormal
karyo-type and deletion of chromosomal material, has been found (Pollini and
Colombi, 1964).
-171-
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V. AQUATIC TOXICITY
A. Acute
Acute toxicity values for freshwater fish are represented by
96-hour static IC5Q values of 20,000 to 22,490 jjg/1 for the bluegill,
Legqmis macrochirus, to 386,000 jug/1 for the mosquitofish, Gambusia affinis,
.with goldfish, Carassius auratus, fathead minnows, Pimephales promelas, and
guppies, Ppecilia reticulatius, being somewhat more resistant than the blue-
gill (U.S. EPA, 1979). Only one study was available for the acute effects
of benzene to freshwater invertebrates. A 48-hour static LC5Q value of
203,000 tig/1 was obtained for the cladoceran Daphnia maqna. LCcn values
"'•"•"" — -"-' 2U
for marine fish were reported as 5,800 and 10,900 jug/1 for striped bass,
Morone saxatilis, and 20,000 to 25,000 /jg/1 for Pacific herring, Clupea
pallasi, and anchovy, Engraulis mordax, larvae. Marine invertebrates were
much more resistant with LC5Q values of 27,000, 108,000, and 450,000 jug/1
reported for grass shrimp, Palaemonetes pugio, dungeness crab, Cancer
magister, and the copepod, Tiqricopus californicus, respectively (U.S. EPA,
1979).
B. Chronic Toxicity
The only chronic toxicity test conducted on an aquatic species was
performed on the freshwater cladoceran, Daphnia magna. There were no ob-
served effects to these organisms at concentrations as high as 98,000 ug/1.
Pertinent information of the chronic effects of benzene on marine fish and
invertebrates could not be located in the available literature.
C. Plant Effects
A concentration of 525,000 ^g/1 was responsible for a 50 percent
>
reduction in cell numbers at 48-hours for the freshwater algae, Chlorella
vulqaris, while marine plants were reported as having growth inhibition at
-------
concentrations ranging from 20,000 to • 100,000 /jg/1 for the diatom,
Skeletonema costatum, with the dinoflagellate, Amphidinium carterae, and the
algae, Cricosphaera carterae, being intermediate in sensitivity with effec-
tive concentrations of 50,000/jg/l.
D. Residues
A bioconcentration factor of 24 was obtained for organisms with a
lipid content of 8 percent.
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
Existing air standards for occupational exposure to benzene include
10 -ppm, an emergency temporary level of 1 ppm by the U.S. Occupational
Safety and Health Administration (Natl. Inst. Occup. Safety Health, 1974,
1977), and 25 ppm by the American Conference of Governmental Industrial
Hygienists (ACGIH, 1971). Based on human epidemiology data, and using a
modified "one-hit" model, the EPA (1979) has estimated levels of benzene in
ambient water which will result -in specified risk levels of human cancer:
Exposure Assumptions Risk Levels and Corresponding Draft Criteria
(per day)
0 1Q-7 10-6 ip-5
2 liters of drinking water 0 0.15/jg/l • 1.5 jug/1 15 jjg/1
and consumption of 18.7
grams fish and shellfish.
Consumption of fish and 0 2.5/jg/l 25 pg/1 258 jug/1
shellfish only.
-03-
-------
B. Aquatic
Criterion for ' the protection of freshwater organisms have been
drafted at 3,100 jjg/1 as a 24-hour average concentration not to exceed 7,000
jug/1. For marine organisms criterion have been drafted as a 24-hour average
concentration of 920 fjg/1 not to exceed 2,100 pg/1.
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BENZENE
REFERENCES
ACGIH. 1971. Threshold limit values. American Conference of Governmental
Industrial Hygienists. Cincinnati, Ohio.
Aksoy, M., et al. 1972. Acute leukemia due to chronic exposure to benzene.
Am. Jour. Med. 52: 160.
Aksoy, M., et al. 1974a. Acute leukemia in two generations following
chronic exposure to benzene. Hum. Hered. 24: 70.
Aksoy, M., et al. 1974b. Leukemia in shoe workers exposed chronically to
benzene. Slood 44: 837.
v
Aksoy, M., et al. 1976a. Combination of genetic factors and chronic expo-
sure to benzene in the aetiology of leukemia. Hum. Hered. 26: 149.
Aksoy, M., et al. 1976b. Types of leukemia in chronic benzene poisoning.
A study in thirty-four patients. Acta Haematologica 55: 65.
American Petroleum Institute. 1978. Table . 6 in Submission to Environ.
Health Comm. of the Sci. Ad vis. Board, U.S. Environ. Prot. Agency. Jan. 13,
1978.
Andrews, L.S., et al. 1977. Biochem. Jour. 26: 293.
Ayers, G.W., and R.E. Muder. 1964. Kirk-Othmer encyclopedia of chemical
technology. 2nd ed. John Wiley and Sons, Inc., New York.
Browning, E. 1965. Benzene. In: Toxicity and metabolism of industrial
solvents. Elsevier Publishing Co., Amsterdam.
Cleland, J.G., and G.L. Kingsbury. 1977. Multimedia environmental goals
for environmental assessment. EPA 600/7-77-136. U.S. Environ. Prot.
Agency, Washington, D.C.
Oeichmann, W.B., et al. 1963. The hemopoietic tissue toxicity of benzene
vapors. Toxicol. Appl. Pharmacol. 5: 201.
Dobrokhotov, V.B. 1972. The mutagenic influence of benzene and toluene
under experimental conditions. Gig. Sanit. 37: 36.
Duvoir, M.R., et al. 1946. The significance of benzene in the bone marrow
in the course of benzene blood diseases. Arch. Mai. Prof. 7: 77.
J.E. 1976.. To 1985: U.S. benzene supply/demand. Hydrocarbon Pro-
cessing. 55: 127.
Gofmekler, v.A. 1968. Effect in embryonic development of benzene and for-
maldehyde. Hyg. Sanit. 33: 327.
-------
Goldstein, 8.0. 1977a. Introduction (Benzene toxicity: Critical review).
Jour. Toxicol. Environ. Health Suppl. 2: 1.
Goldstein, G.D. 19775. Hematotoxicity in humans. Jour. Toxicol. Environ.
Health Suppl. 2: 69.
Howard, P.M., and P.P. Durkin. 1974. Sources of contamination, ambient
levels, and fate of benzene in the environment. EPA 560/5-75-005. U.S.
Environ. Prot. Agency, Washington, D.C.
Hunter, C.G., and D. Blair. 1972. Benzene: Pharmakokinetic studies in
man. Ann. Occup. Hyg. 15: 193.
Infante, P.I., et al. 1977. Leukemia in benzene workers. Lancet. 2: 76.
International Labour Office. 1968. Benzene: Uses, toxic effects, substi-
tutes. Occup. Safety Health Ser., Geneva.
Kinoshita, Y., et al. 1965. A case of myelogenous leukemia. Jour. Japan
Haematol. Soc. 1965: 85.
Kissling, M., and B. Speck. 1971. Chromosomal aberrations in experimental
benzene intoxication. Helv. Med. Acta. 36: 59.
Lange, A., et al. 1973. Serum immunoglobulin levels in workers exposed to
benzene, toluene and xylene. Int. Arch. Arbeitsmed. 31: 37.
Lonneman, W.A., et al. 1968. Aromatic hydrocarbons in the atmosphere of
the Los Angeles basin. Environ. Sci. Technol. 2: 1017.
Lyon, J.P. 1975. Mutagenicity studies with benzene. Ph.D. thesis.
University of California.
Maltoni, C. and C. Scarnato. 1979. LaMedicina del Lavoro. 70(5): 352.
National Academy of Sciences/National Research Council. 1976. Health ef-
fects of benzene: A review. Natl. Acad. Sci., Washington, D.C.
National Academy of Sciences/National Research Council. 1977. Drinking
water and health. Natl. Acad. Sci., Washington, O.C.
National Cancer Institute. 1977. On occurrence, metabolism, and toxicity
including reported carcinogenicity of benzene. Summary rep. Washington,
D.C.
National Institute of Occupational Safety and Health. 1974. Criteria for a
recommended standard. Occupational exposure to benzene. U.S. Dep. Health
Edu. welfare, Washington, D.C.
National Institute of Occupational Safety and Health. 1977. Revised- recom-
mendation for an occupational exposure standard for benzene. U.S. Dept.
Health Edu. Welfare, Washington, D.C.
-------
Nomiyama, K. 1964. Experimental studies an benzene poisoning. Bull. Tokyo
Med. Dental Univ. 11: 297.
Nomiyama, K., and H. Nomiyama. 1974a. Respiratory retention, uptake and
excretion of organic solvents in man. Int. Arch. Arbertsmed. 32; 75.
Ott, M.G., et al. 1978. Mortality among individuals occupationally exposed
to benzene. Arch. Environ. Health. 33: 3.
Pollini, G., and R. Colombi. 1964. Lymphocyte chromosome damage in benzene
blood dyscrasia. Med. Lav. 55: 641.
Rusch, G.M., et al. 1977. Benzene metabolism. Jour. Toxicol. Environ.
Health Suppl. 2: 23.
Sellyei, M., and E. Kelemen. 1971. Chromosome study in a case of granu-
locytic leukemia with "Pelgerisation1 7 years 'after benzene pancytopenia.
Eur. Jour. Cancer 7: 83.
Shahin, M.M, 1977. Unpublished results. The University of Alberta,
Canada. Cited in Mutat. Res. 47: 75.
Simmon, V.F., et al. 1977. Mutagenic activity of chemicals identified in
drinking water. 2nd Int. Conf. Environ. Mutagens, Edinburgh, Scotland,
July, 1977.
Smolik, R., et al. 1973. Serum complement level in workers exposed to ben-
zene, toluene and xylene. Int. Arch. Arbeitsmed. 31: 243.
Snyder, R., and J.J. Kocsis. 1975. Current concepts of chronic benzene
toxicity. CRC Crit. Rev. Toxicol. 3: 265.
Srbova, J., et al. 1950. Absorption and elimination of inhaled benzene in
man. Arch. Ind. Hyg. 2: 1.
Teisinger, J., et al. 1952. The metabolism of benzene in man. Procovni
Lekarstvi 4: 175.
U.S. EPA. 1976a. Health effects of benzene: A review. U.S. Environ. Prot.
Agency, Washington, D.C.
U.S. EPA. 1976b. Air pollution assessment of benzene. Contract No, EPA
68-02-1495. Mitre Corp.
U.S. EPA. 1977a. Sampling in vicinity of benzene production and consump-
tion facilities. Preliminary report to Off. Tox. Subst. Battelle-Columbus
Laboratories.
U.S. EPA. 1977b. Benzene health effects assessment. U.S. Environ. Prot.
Agency, Washington, D.C.
U.S. EPA. 1978. Environmental sources of benzene exposure: source contri-
bution factors. Contract No. 68-01-4635, Mitre Corp.
-------
U.S. EPA. 1979. Benzene: Ambient Water Quality Criteria. (Draft).
Vigliani, E.G., and A. Fornl. 1976. Benzene and leukemia. Environ. Res.
11: 122.
Vigliani, E.G., and' G. Saita. 1964. Benzene and leukemia. New England
Jour. Med. 271: 872.
Watanabe, G.I., and S. Yashida. 1970. The teratogenic effects of benzene
in pregnant mice. Act. Med. Biol. 19: 285.
Weast, R.C. 1972. Handbook of chemistry and physics. The Chemical Rubber
Co., Cleveland, Ohio.
Wolf, M.A., et al. 1956. Toxicological studies of certain alkylated ben-
zenes and benzene. Arch. Ind. Health 14: 387.
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No. 16
Benzidlne
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.
-J80-
-------
SPECIAL NOTATION
U.S. EPA1s Carcinogen Assessment Group (CAG) has evaluated
benzidine and has found sufficient evidence to indicate that
this compound is carcinogenic.
-------
BENZIDINE
Summary
Benzidine is a known carcinogen and has been linked to an in-
creased incidence of bladder cancer in humans and to cancers and
tumors in experimental animals. Benzidine is mutagenic in the Ames
assay and gives positive results in a test measuring DNA synthesis
inhibition in HeLa cells.
Pertinent data could not be located in --the available litera-
ture concerning the toxic effects of benzidine to aquatic organ-
isms.
-------
BENZIDINE
I. INTRODUCTION
Benzidine (4,4'-diaminobiphenyl) is an aromatic amine with
a molecular weight of 184,24. It exists at environmental tempera-
ture as a grayish-yellow, white, or reddish-gray crystalline
powder. Its melting point is 128°C, and its boiling point is
400°C. Benzidine's amino groups have pKa values of 4.66 and
3.57 (Weast, 1972). Two and one-half liter's of cold water will
dissolve 1 g of benzidine, and its solubility increases as water
temperatures rise. Dissolution into organic solvents greatly
increases solubility. Benzidine is easily converted to and from
its salt. Diazotization reactions involving benzidine will result
in colored compounds which are used as dyes in industry (U.S.
EPA, 1979) .
II. EXPOSURE
A. Water
Residential water supplies could be contaminated with
benzidine and its derivatives if the industrial effluent contain-
ing these chemicals were discharged into water supplies, however,
to date U.S. EPA (1979) finds no reports of such contamination.
B. Pood
While food may become contaminated with benzidine due to
poor industrial hygiene, U.S. EPA (1979) reports that the ingestion
of contaminated food is not a real contribution to benzidine toxi-
f
city.
-------
The U.S. EPA (1979) has estimated a weighted average
bioconcentration factor (BCF) of 50 for benzidine, on octanol/water
partition coefficients and other factors.
C. Inhalation
Due to poor industrial hygiene and the use of open sys-
tems in the early days of the chemical and dye industries, inhala-
tion was formerly a principal route of entry for benzidine and its
derivatives into the body. At present workers wear respirators and
protective clothing to avoid exposure when cleaning equipment
(Haley, 1975).
D. Dermal
Skin absorption is the most important route for entry of
benzidine into the body. Intact skin is easily penetrated by the
powdery benzidine base and is penetrated less readily by 3,3'-di-
methoxybenzidine and 3,3 '-dichlorobenzidine. High environmental
air temperatures and humidity increase skin absorption of benzi-
dine, 3/3'-dimethoxybenzidine, 3,3'-dichlorobenzidine, and 3,3'-
dimethylbenzidine (U.S. EPA, 1979).
III. PHARMACOKINETICS
A. Absorption and Distribution
Benzidine is rapidly absorbed into the bodies of intra-
veneously injected rats, with maximum concentrations of free
and bound benzidine occurring at two and three hours, respectively.
The highest concentration of benzidine was found in the blood
followed by the liver, kidney, spleen, heart, and lung (Soloimskaya/
1968) . Four hours after rats received intraperitoneal injections
of 100 mg benzidine/kg, high concentrations of the compound were
£ound in the- stomach, stomach contents, and small intestine;
/
- I 8 4 -
-------
12 hours after administration, benzidine was found in the small
intestine and its contents. Benzidine levels in the liver, the
target organ for toxicity in the rat, remained relatively high
and constant throughout the 12-hour period. The conjugated material,
indicative of the presence of metabolites, was high in urine
and tissues at 12 hours (Baker and Deighton, 1953) . In rats
given 20 mg of 3,3 '-dimethylbenzidine subcutaneously once a week
. for eight weeks, amines were concentrated in the Zymbal's gland,
followed by the kidney, omentum, spleen, -and liver (Pliss and
Zabezhinsky, 1970).
B. Metabolism and Excretion
The urine of humans exposed to benzidine contained a num-
ber of metabolites: N-hydroxyacetylamino benzidine, 3-hydroxyben-
zidine, 4-amino-4-oxybiphenyl, and mono- and diacetylbenzidine
(Engelbertz and Babel, 1953; Troll, et al. 1963; Sciarini and
Meigs, 1961; Vigliani and Barsotti, 1962). Benzidine metabolites
in other species generally differ considerably from those in
humans, although 3-hydroxybenzidine and its conjugation products
are common to both animals and humans (Haley, 1975).
The half-life of benzidine in blood was 68 hours for
the rat and 88 hours for the dog. Rats, dogs, and monkeys ex-
creted 97, 96, and 83 percent, respectively, within one week
of an 0.2 rag/kg dose of benzidine. The respective excretion
rates for 3,3'-dichlorobenzidine were 98, 97, 88.5 percent.
Dogs and monkeys excreted free benzidine in the urine and dichloro-
benzidine in the bile while rats excreted both compounds via
the bile (Kellner, et al. 1973).
-------
Workers exposed to benzidine, who perspire freely and
have wet skin, contain a higher concentration of benzidine in the
urine (U.S. EPA, 1979).
IV. EFFECTS
A. Carcinogenicity
Benzidine is a proven human carcinogen. Its primary site
of tumor induction is the urinary bladder (U.S. EPA, 1979).
Workers exposed to benzidine have a carcinogenicity risk
14 times higher than that of the unexposed population (Case, et al.
1954). The incidence of bladder tumors in humans resulting from
occupational exposures to aromatic amines (benzidine) was first re-
searched in Germany in 1895. In the United States, the first cases
of this condition were diagnosed in 1931 and reported in 1934.
A number of studies document the high incidence of blad-
der tumors in workers exposed to benzidine and other aromatic
amines (Gehrman, 1936; Case, et al. 1954; Scott, 1952; Deichmann
and Gerarde, 1969; Hamblin, 1963; Rye, et al. 1970; Int. Agency
Res. Cancer, 1972; Riches, 1972; Sax, 1975; Zavon, et al. 1973;
Mancuso and El-Attar, 1966, 1967; Kuzelova, et al. 1969; Billiard-
Duchesne, 1960; Vigliani and Barsotti, 1962; Forni, et al. 1972;
Tsuchiya, et al. 1975; Goldwater, et al. 1965). Initial exposure
concentration, exposure duration, and years of survival following
exposure as well as work habits and personal hygiene are involved
in the development of carcinomas where benzidine appears to be im-
plicated (Rye, et al. 1970).
Benzidine has also produced carcinogenic effects' or
tumors in the mouse (hepatoma, lymphoma), the rat (hepatoma,
X
-/Sfc-
-------
carcinoma of the Zytnbal's gland, adenocarcinoma, sarcoma, mammary
gland carcinoma), the hamster (hepatoma, liver carcinoma, chol-
angioma), the rabbit (bladder tumor, gall bladder tumor) and -
the dog (bladder tumor) (Haley, 1975} .
At present, there is no evidence indicating that 3,3'-di-
methylbenzidine, 3 ,3 ' -dimethoxybenzidine, or 3 , 3 ' -dichlorobenzi-
dine are human bladder carcinogens (Rye, et al. 1970).
B. Mutagenicity
In the Ames test, benzidine is mutagenic to SalmonelLa
typhimurium strains TA1537, TA1538, and TA98. Benzidine produces
positive results in a DNA synthesis inhibition test using HeLa
cells (Ames, et al. 1973; McCann, et al. 1975; Garner, et al. 1975;
U.S. EPA, 1978; U.S. EPA, 1979).
C. Teratogenicity
No teratogenic effects of benzidine have been reported in
humans. Mammary gland tumors and lung adenomas occurred in progeny
of female mice that received 8 to 10 mg of 3 , 3 ' -diraethylbenzidine
in the last week of pregnancy. The tumors may have resulted from
transplacental transmission of the chemical or from its transfer to
neonates in milk from dosed mothers (Golub, et al. 1974}.
D. Other Reprodutive Effects
Pertinent data could not be located in the available
literature.
E. Chronic Toxicity
Glomerulonephr itis and nephrotic syndrome were produced
in Sprague-Dawley rats fed 0.043 percent N, N1 -diacetylbenzidin'e, a
metabolite of benzidine, for at least two months (Harman, et al.
1952; Harman, 1971) . Glomerulonephr itis also developed in rats fed
-/S7-
-------
benzidine (Christopher and Jairam, 1970), and in rats receiving in-
jections either 100 mg subcutaneously or 100 or 200 mg intraper-
tioneally of N,N'-diacetylbenzidine. The severity of the lesions
in the later study was dose-related {Bremner and Tange, 1966).
Mice fed 0.01 and 0.08 percent benzidine dihydrochloride
exhibited decreased carcass, liver, and kidney weights, increased
spleen and thymus weights, cloudy swelling of the liver, vacuolar
degeneration of the renal tubules, and hyperplasia of the rayeloid
elements in the bone marrow and of the lymphoid cells in the spleen
and thymic cortex. There was a dose dependent weight loss of 20
percent in males and 7 percent in females (Rao, et al. 1971).
F. Other Relevant Information
Dermatitis, involving both benzidine and its dimethyl
derivative, has been reported in workers in the benzidine dyestuff
industry. Individual sensitivity played a large role in the de-
velopment of this condition (Schwartz, et al. 1947).
V. AQUATIC TOXICITY
Pertinent data could not be located in the available litera-
ture concerning the toxic effects of benzidine to aquatic 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 ambient water concentration standard for benzirdine
is zero, due to potential carcinogenic effects of exposure to
-------
benzidine by ingestion of water and contaminated aquatic organisms.
U.S. EPA may set standards at an interim target risk level in
the range of 10 , 10~ , or 10 with respective corresponding
criteria of 1.67 x 10~3 jig/1, 1.67 x 10~4, and 1.67 x 10~5 ug/1.
B. Aquatic
Criteria for the protection of freshwater or marine
aquatic organisms were not drafted, due to a lack of toxicological
evidence (U.S. EPA, 1979).
-------
BENZIDINE
REFERENCES -'
Ames, B. et al. 1973- Carcinogens are mutagens: A simple test system
combining liver homogenates for activation and bacteria for detection.
Proc. Natl. Acad. Sci. 70: 2281
Baker, R.K., and J.G. Deighton. 1953- The metabolism of benzidine in
the rat. Cancer Res. 13: 529.
Billiard-Duchesne, J.L. I960. Cas Francais de tumeurs professionelles
de la vessie. Acta Unio Int. Contra Cancrum (Belgium) 16: 284.
Bremner, D.A., and J. D. Tange. 1966. Renal and neoplastic lesions
after injection of N,N'~diacetylbenzidine. Arch. Pathol. 81: 146.
Case, R.A.M., et al. 1954. Tumours of the urinary bladder in workmen
engaged in the manufacture and use of certain dyestuff intermediates in
the British chemical industry: Part I. The role of aniline, benzidine,
alpha-naphthylamine and beta-naphthylamine. Br. Jour. Ind. Med. 11: 75-
Christopher, K.J., and B.T. Jairam. 1970. Benzidine
poisoning in white rats. Sci. Cult. (India) 36: 511.
Deichmann, W.B., and H.tf. Gerarde. 1969. Toxicology of drugs and chemicals.
Academic Press, New York.
Englebertz, P., and E. Babel. 1953- Nachweis von benzidin und seinen
umwand lungs produkten im harn und in organteilen. Zentr. Arbeitsmed-
Arbeitsschutz 3: 161.
Forni, A., et al. 1972. Urinary cytology in workers exposed to carcinogenic
aromatic amines: A six-year study. Acta Cytol, 16: 142.
Garner, et al. 1975. Testing of some benzidine anologies for microsomal
activation to bacterial mutagens. Cancer Let. 1: 39.
Gehrman, G.H. 1936. Papilloraa and carcinoma of the bladder in dye workers.
Jour. Am. Med. Assoc. 107: 1436.
Goldwater, L.J., et al. 1965. Bladder tumors in a coal tar dye plant.
Arch. Environ. Health 11: 814.
Golub, N.I,, et al. 1974. Oncogenic action of some nitrogen compounds
on the progeny of experimental mice. Bull. Exp. Biol. Med. (USSR) 78:
1402.
Haley, T.J. 1975. Benzidine revisited: A review of the literature and
problems associated with the use of benzidine and its congeners. Clin.
Toxicol. 8: 13.
-------
Hamblin, D.O. 1963- Aromatic nitro and amino compounds. Page 2105 in
D.W. Fassett and D.D. Irish, eds. Industrial hygiene and toxicology.
Vol. II. Interscience Publishers, New York.
Harraan, J.W. 1971. Chronic glomerulonephritis and the nephrotic syndrome
induced in rats with N,N'-diacetylbenzidine. Jour. Pathol. (Scotland)
104: 119-
Harman, J.W., et al. 1952. Chronic glomerulonephritis and nephrotic
syndrome induced in rats by N,N'-diacetylbenzidine. Am. Jour. Pathol.
28: 529.
International Agency for Research on Cancer. 1972. IARC monographs on
the evaluation of carcinogenic risk of chemicals to man. Vol. I. Lyon,
France.
»_
Kellner, H .M., et al. 1973- Animal studies on the kinetics of
benzidine and 3,3'-dichlorobenzidine. Arch. Toxicol. (West Germany)
31: 61.
Kuzelova, M., et al. 1969- Sledovani pracovniku zamestnanych pri
vyrobe benzidinu. Prac. Lek. (Czechoslovika) 21: 310.
Mancuso, T.F., and A.A. El-Attar. 1966. Cohort studies of workers
exposed to betanaphthylamine and benzidine. Ind. Med. Surg. 35: 571.
Mancuso, T.F., and A.A. El-Attar. 1967- Cohort study of workers exposed
to betanaphthylamine and benzidine. Jour. Occup. Med. 9: 277.
McCann, J., et al. 1975. Detection of carcinogens as mutagens in the
Salmonella/microsome test: Assay of 300 chemicals.' Proc. Natl. Acad.
Sci. 72: 5135.
Pliss, G.B., and M.A. Zabezhinsky. 1970. Carcinogenic properties of
orthotolidine (3,3'-dimethylbenzidine). Jour. Natl. Cancer Inst. 45: 283-
Rao, K.V.N., et al. 1971. Subacute toxicity of benzidine in the young
adult mice. Fed. Proc^ Am. Soc. Exp. Biol. 30: 344.
Riches, E. 1972. Industrial cancers. Nurs. Mirror (Great Br.) 134: 21.
Rye, W.A., et al. 1970. Facts and myths concerning aromatic diamine
curing agents. Jour. Occup. Med. 12: 211.
Sax, N.I. 1975. Dangerous properties of industrial materials. 4th ed.
Van Nostrand Reinhold Co., New York.
Schwartz, L., et al. 1947. Dermatitis in synthetic dye manufacture.
Page 268 in Occupational diseases of the skin. Lea and Febiger, Philadelphia,
Pa.
-------
Sciarini, L.J., and J.W. Meigs. 1961. The biotransformation of benzidine.
II. Studies in mouse and man. Arch. Environ. Health 2: 423.
Scott, T.S. 1952. The incidence of bladder tumours in a dyestuffs factory.
Br. Jour. Ind. Med. 9: 127.
Soloimskaya, E.A. 1968. The distribution of benzidine in rat organs and
its effect on the peripheral blood. Vopr.-Onkol. (USSR) 14: 51.
Troll, W., et al. 1963- N-hydroxy acetyl amino compounds, urinary metabolites
of aromatic amines in man. Proc. Am. Assoc- Cancer Res. 4: 68.
Tsuchiya, K., et al. 1975. An epidemiological study of occupational
bladder tumours in the dye industry of Japan. Br. Jour. Ind. Med. 32:
203.
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, D.C.
U.S. EPA. 1979- Benzidine: Ambient Water Quality Criteria. (Draft).
Vigliani, E.G., and M. Barsotti. 1962. Environmental tumors of the
bladder in some Italian dyestuff factories. Acta Unio Int. Contra Cancrum
(Belgium) 18: 669.
Weast, R.C., ed. 1972. Handbood of chemistry and physics. 53rd ed. CRC
Press, Cleveland, Ohio.
Zavon, M.R., et al. 1973- Benzidine exposure as a cause of bladder
tumors. Arch. Environ. Health 27: 1-
-m-
-------
No. 17
Benz(a)anthracene
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.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated
benz(a)anthracene and has found sufficient evidence to indi-
cate that this compound is carcinogenic.
-------
BENZ(a? ANTHRACENE
SUMMARY
Benz(a)anthracene is a member of the polycyclic aro-
matic hydrocarbons (PAH) class. Although the PAH class
contains several well-known potent carcinogens, benz(a)an-
thracene displays only weak carcinogenic activity. Benz (a)-
anthracene apparently does not display remarkable acute
or chronic toxicity other than the capability to induce
tumors on the skin of mice. Although exposure to benz(a)-
anthracene in the environment occurs in conjunc'tion with
exposure to other PAH, it is not known how these compounds
may interact in human systems. Furthermore, the specific
effects of benz(a)anthracene in humans are not known.
The only toxicity data for any of the polycyclic aro-
matic hydrocarbons is an 87 percent mortality of freshwater
fish exposed to 1,000 ;ag/l benz (a) anthracene for six months.
-------
I. INTRODUCTION
This profile is based primarily on the Ambient Water
Quality Criteria Document for Polynuclear Aromatic Hydrocar-
bons (U.S. EPA, 1979a) and the Multimedia Health Assessment
Document for Polycyclic Organic Matter (U.S. SPA, 1979b).
Benz (a) anthracene (CIS^T?) ^s one °^ t'ne familir of
polycyclic aromatic hydrocarbons (PAH) formed as a result
of incomplete combustion of organic material. Benz(a)anthra-
cene has the following physical/chemical properties (U.S.
EPA, 1979b):
Melting point: 159.5-160.5°C
Boiling Point: 4CO°C ?
Vapor Pressue: 1.10 x 10 torr
PAH, including benz(a)anthracene, are ubiquitous in
the environment, being found in ambient air, food, water,
soils, and sediment (U.S. EPA, 1979b}. The PAH class con-
tains a number of potent carcinogens (e.g., benzo(a)pyrene),
weak carcinogens (e.g., benz(a)anthracene), and cocarcino-
gens (e.g., fluoranthene), as well as numerous non-carcino-
gens (U.S. EPA, 1979b).
PAH which contain more than three rings (such as benz(a)-
anthracene) are relatively stable in the environment, and
may be transported in air and water by adsorption to particu-
late 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, 1979c).
-------
II. EXPOSURE
A. Water
Benz(a)anthracene levels in surface waters or
drinking water have not been reported. However, the concen-
tration of six representative PAH (net including benz(a)-
anthracene) in U.S. drinking water averaged 13.5 ng/1 (Basu
and Saxena, 1977, 1978).
B. Food
Benz(a)anthracene has been detected in a wide
variety of foods including margarine (up to 29.5 ppb), smoked
fish {up to 1.7 ppb), yeast (up to 2C3 pcb), and cooked
or smoked meat (up to 33.0 ppb) (U.S. EPA, 1979a). The
total intake of all types of PAH thrcuch the diet has been
estimated at 1.6 to 16 ug/day (U.S. EPA, i979b). The U.S.
EPA (1979a) has estimated the bioconcer.tration factor for
benz(a)anthracene to be 3,100 for the edible portions of
fish and shellfish consumed by Americans. This estimate
is based on the octanol/water partition coefficient of ben2-
(a)anthracene.
C. Inhalation
Benz(a)anthracene has been repeatedly detected
in ambient air at concentrations ranging from 0.18 to 4.6.
ng/m3 (U.S. EPA, 1979a) . Thus, the hur-an daily intake of
benz(a)anthracene by inhalation of ambient air may be in
the range of 3.42 to 87.4 ng, assuming that a human breathes
19 m of air per day. •
-------
III. PHARMACOKINETICS
There are no data available concerning the pharntaco-
kinetics of benz(a)anthracene, or other PAH, in humans.
Nevertheless, it is possible to make limited assumptions
based on the results of animal research conducted with sev-
eral PAH, particularly benzo(a)pyrene.
A. Absorption
The absorption of benz(a)anthracene in humans
has not been studied. However, it is known (U.S. EPA, 1979a)
that, as a class, PAH are well-absorbed across the respira-
tory and gastrointestinal epithelia. In particular, benz(a)-
anthracene was reported to be readily transported across
the gastrointestinal mucosa (Rees, et al., 1971). The high
lipid solubility of compounds in the PAH class supports
this observation.
B. Distribution
The distribution of benz(a)anthracene in mammals
has not been studied. However, it is known {U.S. EPA, 1979a)
that other PAH are widely distributed throughout the body
following their absorption in experimental rodents. Rela-
tive to other tissues, PAH tend 'to localize in body fat
and fatty tissues (e.g., breast).
C. Metabolism
Benz(a)anthracene, like other PAH, is metabolized
by the microsomal mixed-function oxidase enzyme system in
mammals (U.S. EPA, 1979b). Metabolic attack on one or more
of the aromatic double bonds 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 inter-
mediates 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 benz(a)anthracene 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. However, the
rate of disappearance of various PAH from the body and
the principal routes of excretion are influenced both by
the structure of the parent compound and the route of admini-
stration (U.S. EPA, 1979a). It is unlikely that PAH will
accumulate in the body with chronic low-level exposures.
IV. EFFECTS
A. Carcinogenicity
Benz(a)anthracene is recognized as a weak carcino-
gen in mammals (U.S. EPA, 1979a,b). It is a tumor initiator
on the skin of mice, but failed to yield significant results
in the strain A mouse pulmonary tumor bioassay system. .
-200-
-------
B. Mutagenicity
Benz(a)anthracene has shown weak mutagenic activity
in several test system, including Ames Salmonella assay,
somatic cells in culture, and sister chromatid exchange
in Chinese hamster cells (U.S. EPA, 1979b).
C. Teratogenicity
Pertinent data could not be located in the avail-
able literature concerning the possible teratogenicity of
benz(a)anthracene. Other related PAH are apparently not
significantly teratogenic in mammals (U.S. EPA, 1979a).
D. Other Reproductive Effects
Pertinent data could not be located in the avail-
able literature.
E. Chronic Toxicity
The chronic toxicity of benz(a)anthracene has
not been extensively studied. The repeated injection of
benz(a)anthracene in mice for 40 weeks (total dose, 10 mg.}
had little apparent effect on longevity or organ weights
(U.S. EPA, 1979b).
V. AQUATIC TOXICITY
A. Acute
Pertinent data could not be located in the avail-
able information.
B. Chronic
No standard chronic toxicity data have been pre-
sented on freshwater or marine species. The only toxicity'
data available for benz(a)anthracene for fish is an SI per-
-2OJ-
-------
cent mortality on the freshwater bluegill sunfish, Lepomis
macrochirus, exposed to 1,000 ^g/1 for six months (Brown,
et al., 1975).
C. Plant Effects
Pertinent data could not be located in the avail-
able information.
VI. EXISTING GUIDELINES AND STANDARDS
Neither the human nor the aquatic criteria derived
by U.S. EPA (1979a), which are summarized below, have gone
through the process of review; therefore, there is a pos-
sibility that these criteria will be changed.
A. Human
There are no established exposure criteria for
benz (a) anthracene. However, PAH as 5. 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 fluorar.thene, benzo(g,h,i)-
perylene, benzo (b) f luoranthene, benzc•(;<) fluoranthene, indeno-
(l,2,3-cd)pyrene, and benzo (a) pyrene) r.ot to exceed 0.2
ug/1. Occupational exposure criteria have been established
for coke oven emissions, coal tar predicts, and coal tar
pitch volatiles, all of which contain large amounts of PAH
including benz(a)anthracene (U.S. EPA, 1979a) .
The U.S. EPA (1979a) draft recommended criteria
for PAH in water are based upon the extrapolation of animal
carcinogenicity data for benzo(a) pyrer.e and dibenz (a , h) anthra-
cene.
-------
B. Aquatic
Data were insufficient to propose criteria for
freshwater or marine environments.
X
-203-
-------
8ENZ(a}ANTHRACENE '
REFERENCES
Basu, O.K., and J. Saxena. 1977. Analysis of raw and drinking water sam-
ples for polynuclear aromatic hydrocarbons. EPA PO No. CA-7-2999-A, and
CA-8-2275-8. Expo. Evalu. 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.
Brown, E.R., et al. 1975. Tumors in fish caught in polluted waters: possi-
ble explanations. Comparative Leukemia Res. 1973. Leukemogenssis. Univ.
Tokyo Press/Karger, Basel, pp. 47-57.
Rees, E.O., et al. 1971. A study of the mechanism of intestinal absorption
of benzo(a)pyrene. Biochem. Biophys. Act. 225: 96.
U.S. EPA. 1979a. Polynuclear Aromatic Hydrocarbons: Ambient Water Quality
Criteria (Draft),
U.S. EPA. 1979b. Multimedia health assessment document for polycyclic or-
ganic matter. Prepared under contract by J. Santodonato, et al., Syracuse
Research Corp.
U.S. EPA. 1979c. Environmental" Criteria and 'Assessment Office. Polynu-
clear Aromatic Hydrocarbons: Hazard Profile (Draft).
World Health Organization. 1970. European standards for drinking water.
2nd. ed.j Geneva.
-------
No. 18
Benzo(b)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.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (GAG) has evaluated
benzo(b)fluoranthene and has found sufficient evidence to
indicate that this compound is carcinogenic.
-207-
-------
BEN 20( b ) FLUORANTHENE
SUMMARY
Benzo(b)fluoranthene is a member of the polycyclic aro-
matic hydrocarbon (PAH) class. Numerous compounds in the PAH
class are well known for their carcinogenic effects in ani-
mals. Benzo(b)fluoranthene is carcinogenic to the skin of
mice and produces sarcomas when injected in mice. Very
little is known concerning the non-carcinogenic effects pro-
duced by chronic exposure to benzo(b)fluoranthene. Although
exposure to benzo(b)fluoranthene in the environment occurs in
conjunction with exposure to other PAH, it is not known how
these compounds may interact in human systems. Furthermore,
the specific effects of benzo(b)fluoranthene in humans are
not known.
Standard acute or chronic toxicity testing for aquatic
organisms has not been found in the available literature.
-------
BEN ZO (b ) FLUORANTHEN E
I . INTRODUCTION
This profile is based primarily on the Ambient Water
Quality Criteria Document for Polynuclear Aromatic Hydrocar-
bons (U.S. EPA, 1979a) and the Multimedia Health Assessment
Document fo,r Polycyclic Organic Matter (U.S. EPA, 1979b).
Benzo(b)fluoranthene (C20H12^ ^S °ne °^ fc^e ^am-^v
of polycyclic aromatic hydrocarbons (PAH) formed as a result
of incomplete combustion of organic material. Its physical/
chemical properties have not been well-characterized, other
than a reported melting point of 167°C (U.S. EPA, 1979b).
PAH, including benzo(b)fluoranthene, are ubiquitous in
the environment, being found in ambient air, food, water,
soils, and sediment (U.S. EPA, 1979b). The PAH class con-
tains a number of potent carcinogens (e.g., benzo(a)pyrene),
moderately active carcinogens (e.g., benzo(b)fluoranthene),
weak carcinogens (e.g., benz(a)anthracene), and cocarcinogens
(e.g., fluoranthene), as well as numerous noncarcinogens
(U.S. EPA, 1979b).
PAH which contain more than three rings (such as benzo-
(b)fluoranthene) are relatively stable in the environment and
may be transported in air and water by adsorption to particu-
lar matter. However, biodegradation and chemical treatment
are effective in eliminating most PAH in the environment.
Refer to the PAH Hazard Profile (U.S. EPA, 1979c) for a more
general treatment of PAH.
-------
II. EXPOSURE
A. Water
In a monitoring survey of U.S. drinking water, Basu
and Saxena {1977, 1978} were unable to detect benzo(b)fluor-
anthene. However, the concentration of six representative
PAH (fluoranthene, benzo(a)pyrene, benzofg h ijperylene,
benzo(j)fluoranthene, benzo(k)fluoranthene, indeno(l,2,3-cd)
pyrene) averaged 13.5 ng/1.
B. Food
Levels of benzo{b}fluoranthene have not been re-
ported for food. 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
weighted average bioconcentration factor of benzo(b)fluor-
anthene to be 6,800 for the edible portion of fish and shell-
fish consumed by Americans. This estimate is based on the
octanoi/water partition coefficient of benzo(b)fluoranthene.
C, Inhalation
Benzo(b)fluoranthene has been detected in ambient
air at concentrations ranging from 0.1 to 1.6 ng/m^ (Gordon
and Bryan, 1973). Thus, the human daily intake of benzo(b)-
fluoranthene by inhalation of ambient air may be in the range
of 1.9 to 30.4 ng, assuming that a human breathes 19 m^ of
air per day.
III. PHARMACOKINETICS
Pertinent data could not be located in the available
literature concerning the pharmacokinetics of benzo(b)fluor-
anthene, or other PAH, in huraans. Nevertheless, it is pos-
-------
sible to make limited assumptions based on the results of
animal research conducted with several PAH, particularly
benzo(a)pyrene.
A. Absorption
The absorption of benzo(b)fluoranthene 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. The
high lipid solubility of compounds in the PAH class supports
\
this observation.
B. Distribution
The distribution of benzo{b)fluoranthene in mammals
has not been studied. 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 benzo(b}fluoranthene in mammals
has not been studied. Benzo(b)fluoranthene, like other PAH,
is most likely metabolized by the microsomal mixed-function
oxidase enzyme system in mammals (U.S. EPA, 1979b). Meta-
bolic attack on one or more of the aromatic double bonds
leads to the formation of phenols and isomeric dihydrodiols
by the intermediate formation of reactive epoxides. Dihydro-
diols are further metabolized by microsomal mixed-function •
oxidases to yield diol epoxides, compounds which are known to
be biologically reactive intermediates for certain PAH. Re-
moval of activated intermediates by conjugation with gluta-
2
-an-
-------
thione or glucuronic acid, or by further metabolism to tetra-
hydrotetrols, is a key step in protecting the organism from
toxic interaction with cell macromolecules.
D. Excretion
The excretion of benzo(b)fluoranthene by mammals
has not been studied. However, the excretion of closely re-
lated PAH is rapid and occurs mainly via the feces (U.S. EPA,
1979a). Elimination in the bile may account for a signifi-
cant percentage of administered PAH. It is unlikely that PAH
will accumulate in the body with chronic low-level exposures.
IV. EFFECTS
A. Carcinogenicity
Benzo(b)fluoranthene is regarded as a moderately
active carcinogen (U.S. EPA, 1979b). It is carcinogenic by
skin painting on mice, and by subcutaneous injection in mice
(U.S. EPA, 1979b; LaVoie, et al. 1979). The sarcomagenic
potency of benzofb)fluoranthene is similar to that of benzo-
(a)pyrene (Buu-Hoi, 1964).
B. Mutagenicity
Benzo(b)fluoranthene is mutagenic in the Ames Sal-
monella assay in the presence of a microsomal activating sys-
tem (LaVoie, et al. 1979). It is also positive in the induc-
tion of sister-chromatid exchanges by intraperitoneal injec-
tion in Chinese hamsters (U.S. EPA, 1979b).
C. Teratogenicity
Pertinent data could not be located in the litera-
ture available concerning the possible teratogenicity of
-------
benzo(b)fluoranthene. Other related PAH are apparently not
significantly teratogenic in mammals (U.S. EPA, 1979a).
D. Other Reproductive Effects
Pertinent information could not be located in the
available literature.
E. Chronic Toxicity
Published data are not available regarding the non-
carcinogenic chronic effects of benzo{b}fluoranthene. It is
known, however, that exposure to carcinogenic PAH may produce
widespread tissue damage as well as selective destruction of
proliferating tissues (e.g., hematopoietic and lymphoid sys-
tems) (U.S. EPA, 1979a).
V. AQUATIC TOXICITY
Pertinent information could not be located in the avail-
able literature.
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
There are no established exposure criteria for
benzo(b)fluoranthene. However, PAH as a class are regulated
by several authorities. The World Health Organization has
recommended that the concentration of PAH in drinking water
(measured as the total of fluoranthene, benzo(g,h,i)perylene,
benzo(b)fluoranthene, benzo(k)fluoranthene, indenof1,2,3-cd)
pyrene, and benzo{a)pyrene) not exceed 0.2 ug/1. Occupa-
tional exposure criteria have been established for coke oven
emissions, coal tar products, and coal tar pitch volatiles,
all of which contain large amounts of PAH including benzo(b)-
fluoranthene (U.S. EPA, 1979a).
-a/3-
-------
The U.S. EPA (1979a) draft recommended criteria for
PAH in water are based upon the extrapolation of animal car-
cinogenicity data for benzo(a)pyrene and dibenzo(a,h)anthra-
cene.
B. Aquatic
The criteria for freshwater and marine life have
not been drafted (U.S. EPA, 1979a).
-------
BENZO(b)FLUORANTHENE
REFERENCES
Basu, O.K. and J. Saxena. 1977. Analysis of raw and drinking water sam-
ples for polynuclear aromatic hydrocarbons. U.S. Environ. Prot. Agency,
P.O. No. CA-7-2999-A. 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.
Buu-Hoi, N.P. 1964. New developments in chemical carcinogenesis by
polycyclic hydrocarbons and related heterocycles: A review. Cancer Res.
24: 1511.
».
Gordon, R.J. and R.J. Bryan. 1973. Patterns of airborne polynuclear hy-
drocarbon concentrations at four Los Angeles sites. Environ. Sci. Technol.
7: 1050.
La Voie, E., et al. 1979. A comparison of the mutagenicity, tumor-initiat-
ing activity and complete carcinogenicity of polynuclear aromatic hydrocar-
bons. In; Polynuclear Aromatic Hydrocarbons, P.W. Jones and P. Leber (eds.)
Ann Arbor Science Publishers, Inc.
U.S. EPA. 1979a. Polynuclear Aromatic Hydrocarbons: Ambient Water Quality
Criteria. (Draft)
U.S. EPA. 1979b. Multimedia health assessment document for polycyclic or-
ganic matter. Prepared under contract by J. Santodonato, et al., Syracuse
Research Corp.
U.S. EPA. 1979c. Environmental Criteria and Assessment Office. Polynucle-
ar Aromatic Hydrocarbons: Hazard Profile. (Draft)
World Health Organization. 1970. European Standards for Drinking Water. 2nd
ed., Geneva.
-------
No. 19
Benzo(a)pyrene
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.
-an-
-------
SPECIAL NOTATION
U.S. EPA1s Carcinogen Assessment Group {GAG} has evaluated
benzo(a)pyrene and has found sufficient evidence to indicate
that this compound is carcinogenic.
-2.1?-
-------
BENZO(a)PYREN£
Summary
The first chemicals shown to be involved in the development of cancer
belong to the polynuclear aromatic hydrocarbons (PAH) class. Benzo(a)pyrene
is the most widely recognized and extensively studied of all carcinogenic
PAH. It is among the most potent animal carcinogens known and produces
tumors in virtually all species by all routes of administration.
Since humans are never exposed to only benzo(a)pyrene in the environ-
ment, it is not possible to attribute human cancers solely to exposure to
benzo(a)pyrene. However, numerous epidemiologic studies support the belief
that carcinogenic PAH, including benzo(a)pyrene, are also human carcinogens.
Measured steady-state bioconcentration factors are not available for
freshwater or saltwater aquatic species exposed to benzo(a)pyrene. Standard
toxicity data for freshwater and saltwater aquatic life have not been re-
ported.
-------
I. INTRODUCTION
This profile is based 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).
Benzo(a)pyrene (C20H12^ is one of the family of polynuclear aromat-
ic hydrocarbons (PAH) formed as a result of incomplete combustion of organic
material. Its physical and chemical properties have not been well-charac-
terized, other than a reported melting point of 178.8-179.3_°C and a vapor
pressure of 5.49 x 10"9 mm Hg (U.S. EPA, 1979b). ''
PAH, including benzo(a)pyrene, are ubiquitous in the environment, being
found in ambient air, food, water, soils and sediment (U.S. EPA, 1979a).
The PAH class contains a number of potent carcinogens (e.g., benzo(a)py-
rene), moderately active carcinogens (e.g., benzo(b)fluoranthene), weak car-
cinogens (e.g., benz(a)anthracene), and cocarcinogens (e.g., fluoranthene),
as well as numerous noncarcinogens (U.S. EPA, 1979a).
PAH which contain more than three rings (such as benzo(a)pyrene) 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.
II. EXPOSURE
A. Water
Basu and Saxena (1977, 1978) have monitored various United States
drinking water supplies for the presence of PAH, including benzo(a)pyrene.
They reported that the average level of benzo(a)pyrene in drinking water was
0.55 nanograms/liter. This would result in a human daily intake of benzo-
»
(a)pyrene from water of about 0.0011 jug.
-220-
-------
B. Food
Benzo(a)pyrene has been detected in a wide variety of foods by
numerous investigators (U.S. EPA, 1979a), Benzo(a)pyrene levels are espe-
cially high in cooked or smoked meats, where in certain cases (i.e., char-
coal-broiled steak) concentrations as high as 50 ppb have been reported
(Lijinsky and Ross, 1967). It has been estimated (U.S. EPA, 1979b) that the
daily dietary intake of benzo(a)pyrene is about 0.16 to 1.6 ug, and total
PAH intake is about 1.6 to 16 ug. The U.S. EPA Q979a) has estimated the
weighted average bioconcentration factor for benzo(a)pyrene to be 6,800 for
the edible portions of fish and shellfish consumed by Americans. This esti-
mate is based on the octanol/water partition coefficient for benzo(a)pyrene.
C. Inhalation
Benzo(a)pyrene levels have been routinely monitored in the ambient
atmosphere for many years. The average urban-rural ambient benzo(a)pyrene
concentration in the United States has been estimated at 0.5 nanograms/m
(U.S. EPA, 1979a). Thus, the human daily intake of benzo(a)pyrene by inhala-
tion of ambient air is about 9.5 nanograms, assuming that a human breathes
about 19 m of air per day.
III. PHARMACOKINETICS
Pertinent data could not be found in available literature concerning
the pharmacokinetics of benzo(a)pyrene, or other PAH, in humans. Neverthe-
less, it is possible to make limited assumptions based on the results of
animal research conducted with several PAH, particularly benzo(a)pyrene.
A. Absorption
Toxicity data indicate that, as a class, PAH are capable of passage
*
across epithelial membranes (Smyth, et al. 1962). In particular, benzo(a)-
pyrene was reported to be readily transported across the intestinal mucosa
t
-22J-
-------
(Rees, et al. 1971) and the respiratory membranes (Kotin, et al. 1969; Vai-
niok, et al. 1976).
B. Distribution
Benzo(a)pyrene becomes localized in a wide variety of body tissues
following its absorption (Kotin, et al. 1969). Due to its high lipid solu-
bility, benzo(a)pyrene localizes primarily in body fat and fatty tissues
(e.g., breast) (Schlede, et al. 1970a,b).
C. Metabolism
The metabolism of benzo(a)pyrene in mammals has been studied in
great detail (U.S. EPA, 1979a). Benzo(a)pyrene, like other PAH, is metabo-
lized 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 interme-
diate formation of reactive epoxides. Dihydrodiols are further metabolized
by microsomal mixed function oxidases to yield diol epoxides, compounds
which are known to be ultimate carcinogens 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 benzo(a)pyrene by mammals has been studied by sev-
eral groups of investigators. In general, the excretion of benzo(a)pyrene
and related PAH is rapid, and occurs mainly via the feces (U.S. EPA, 1979a;
Schlede, et al. 1970a,b). Elimination in the bile may account for a signi-
ficant percentage of administered PAH. It is unlikely that PAH will accumu-
»
late in the body as a result of chronic low-level exposures.
-------
IV. EFFECTS
A. Carcinogenicity
The carcinogenic activity of benzo(a)pyrene was first recognized
decades ago, and since that time it has become a laboratory standard for the
production of experimental tumors which resemble human carcinomas in ani-
mals. The carcinogenic activity of benzo(a)pyrene is distinguished by sev-
eral remarkable features: (1) it is among the most potent animal carcino-
gens known, producing tumors by single exposures to microgram quantities;
(2) it acts both at the site of application and at organs distant to the
site of absorption; and (3) its carcinogenicity has been demonstrated in
nearly every tissue and species tested, regardless of the route of admini-
stration (U.S. EPA, 1979a).
Oral administration of benzo(a)pyrene to rodents can result in
tumors of the forestomach, mammary gland, ovary, lung, liver, and lymphoid
and hematopoietic tissues (U.S. EPA, 1979a). Exposure to benzo(a)pyrene by
intratracheal instillation in rodents can also be an effective means of pro-
ducing respiratory tract tumors (Feron, et al. 1973). In addition, benzo-
(a)pyrene has remarkable potency for the induction of skin tumors in mice by
direct dermal application (U.S. EPA, 1979a).
Numerous epidemiologic studies support the belief that carcinogenic
PAH, including benzo(a)pyrene, are responsible for the production of human
cancers both in occupational situations and among tobacco smokers (U.S. EPA,
1979b).
B. Mutagenicity
8enzo(a)pyrene gives positive results in nearly all mutagenicity
test systems including the Ames Salmonella assay, cultured Chinese hamster
cells, the sister-chromatid exchange test, and the induction of DNA repair
synthesis (U.S. EPA, 1979a).
-223,-
-------
C. Teratogenicity and Other Reproductive Effects
Only limited data are available regarding the teratogenic effects
of benzo(a)pyrene or other PAH in animals. Benzo(a)pyrene had little effect
on fertility or the developing embryo in several mammalian and non-mammalian
species (Rigdon and Rennels, 1964; Rigdon and Neal, 1965).
D. Chronic Toxicity
As long ago as 1937, investigators knew that carcinogenic PAH such
as benzo(a)pyrene produced systemic toxicity as manifested by an inhibition
of body growth in rats and mice (Haddow, et al... 1937). The target organs
affected by chronic administration of carcinogenic PAH are diverse, due
partly to extensive distribution in the body and also to the selective de-
struction of proliferating cells (e.g., hematopoietic and lymphoid system,
intestinal epithelium, testis) (Philips, et al. 1973).
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
There are no established exposure standards specifically for benzo-
(a)pyrene. 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, benzo(g,h,-
Dperylene, benzo(b)fluoranthene, benzo(k)fluoranthene, indeno(l,2,3-cd)py-
rene, and benzo(a)pyrene) not exceed 0.2 ug/1. Occupational exposure cri-
-23H-
-------
teria have been established for coke oven emissions, coal tar products, and
coal tar pitch volatiles, all of which contain large amounts of PAH in water
based upon the extrapolation of animal carcinogenicity data for benzo(a)py-
rene and dibenz(a,h)anthracene. Levels for each compound are derived which
will result in specified risk levels of human cancer as shown in the table
below:
BaP
Exposure Assumptions Risk Levels and Corresponding Draft Criteria
(per day)
o io-7 ig-6 10-5
2 liters of drinking water 0 0.097 0.97 9.7
and consumption of 18.7
grams of fish and shellfish.
Consumption of fish 0.44 4.45 44.46
and shellfish only.
DBA
Exposure Assumptions Risk Levels and Corresponding Draft Criteria
(per day)
o io-7 • iQ-6 ig-5
2 liters of drinking water 0 0.43 4.3 43
and consumption of 18.7
grams of fish and shellfish.
Consumption of fish 1.96 19.6 196
and shellfish only.
B Aquatic
Guidelines are not available for benzo(a)pyrene in aquatic environ-
ments.
-------
BENZO(A)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, Expo. Evalu. Branch, HERL., Cincin-
nati.
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.
Feron, V.J., et al. 1973. Dose-response correlation for the induc-
tion of respiratory tract tumors in Syrian golden hamsters by in-
tratracheal instillations of benzo(a)pyrene. Europ. Jour. Cancer.
9: 387.
Haddow, A., et al. 1937. The influence of certain carcinogenic and
other hydrocarbons on body growth in the rat. Proc. Royal Soc. B.
122: 477.
Kotin, P., et al. 1969. Distribution retention and elimination of
C -3, 4-benzopyrene after administration to mice and rats. Jour.
Natl. Cancer Inst. 23: 541.
Lijinsky, W. , and A.E. Ross. 1967. Production of carcinogenic
polynuclear hydrocarbons in the cooking of food. Food Cosmet.
Toxicol. 5: 343.
Philips, F.S. et al., 1973. In_ yiyo cytotoxicity of polycyclic hy-
drocarbons. Iri: Pharmacology and the future of man. Proc. 5th
Intl. Congr. Pharmacology, 1972, San Francisco. 2: 75.
Rees, E.O. , et al. 1971. A study of the mechanism of intestinal
absorption of benzo(a)pyrene. Biochem. Biophys. Act. 255: 96.
Rigdon, R.H., and J. Neal. 1965. Effects of feeding benzo{a)py-
rene on fertility, embryos, and young mice. Jour. Natl. Cancer.
Inst. 32: 297.
Rigdon, R.H., and E.G. Rennels. 1964. Effect of feeding benzo-
pyrene on reproduction in the rat. Experientia. 20: 1291.
Schlede, E. , et al. 1970a. Stimulatory effect of benzo(a)pyrene
and phenobarbital 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.
-226,-
-------
Smyth, H.F., et al. 1962, Range - finding toxicity data: List II.
Am. Ind. Hyg. Jour. 23: 95.
U.S. EPA. 1979a. Polynuclear Aromatic Hydrocarbons: Ambient
Water Quality Criteria. (Draft).
U.S. EPA. 1979b. Multimedia Health Assessment Document for Poly-
cyclic Organic Matter. Prepared under contract by J. Santodonato
et al., Syracuse Research Corporation.
Vainioh, et .al. 1976. The fate of intracheally installed benzo-
(a)pyrene in the isolated perfused rat lung of both control and 20-
methylcholanthrene pretreated. Res. Commun. Chem. Path. Pharmacol.
13: 259.
World Health Organization. 1970. European standards for drinking
water. 2nd ed. Revised. Geneva. "•
-------
No. 20
Benzotrichloride
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.
-------
BENZOTRICHLORIDE
Summary
Benzotrichloride has been shown to be mutagenic in a number of micro-
bial tests with and without metabolic activation.. One study has described
the carcinogenicity of benzotrichloride in mice. The lowest concentration
producing a lethal effect (LCLO) has been reported at 125 ppm for rats in-
haling benzotrichloride for four hours. Pertinent data for the toxic
effects to aquatic organisms were not found in the available literature.
-------
I. INTRODUCTION
Senzotrichloride (CAS registry number 98-07-7), is a colorless, oily,
fuming liquid. It is made by the free radical chlorination of boiling
toluene (Sidi, 1964; Windholz, 1976). Benzotrichloride has the following
physical and chemical properties (Windholz, 1976; Sidi, 1964):
Formula: C6H5C13
Molecular Weight: 195.48
Melting Pointr -5°C
Boiling Point: 220.8°c
Density: 1.375620
4
Solubility: alcohol, ether, benzene,
insoluble in water
Benzotrichloride is used extensively in the dye industry for the
production of Malachite green, Rosamine, Quinoline red, and Alizarine yellow
A. It can also be used to produce ethyl orthobenzoate (Sidi, 1964).
II. EXPOSURE
A. Water
Benzotrichloride decompose in the presence of water to benzoic and
hydrochloric acids (Windholz, 1976).
B. Food
Pertinent data were not found in the available literature.
C. Inhalation
Pertinent data were not found in the available literature; how-
ever, significant exposure could occur in the workplace from accidental
spills. Benzotrichloride decomposes in moist air to benzoic and hydro-
chloric acids (Windholz, 1976).
D. Dermal
Benzotrichloride is irritating to the skin (Windholz, 1976).
-------
III. PHARMACOKINETICS
Pertinent pharmacokinetic data were not found in the available
literature.
IV. EFFECTS
A. Carcinogenicity
In a study by Matsushito and coworkers (1975), benzotrichloride
was found to induce carcinomas, leukemia, and papillomas in mice. The de-
tails of the study were not available for assessment.
B. Mutagenicity
Yasuo, et al. (1978) tested the mutagenicity of several compounds
including benzotrichloride in microbial systems such as the rec-assay using
Bacillus subtilis, reversion assays using E^_ coli, and the Ames assay using
Salmonella typhimurium, with or without metabolic activation. Benzo-
trichloride was positive for mutagenicity in the rec-assay and was highly
positive on certain strains of §_._ coli and S_._ typhimurium in the reversion
assay with metabolic activation. Without metabolic activation, however,
benzotrichloride was only weakly positive in the latter assay.
C. Teratogenicity, Reproductive Effects, and Chronic Toxicity
Pertinent data were not found in the available literature.
0. Acute Toxicity
The lowest lethal concentration (LCLQ) for rats inhaling benzo-
trichloride is 125 ppm for four hours (Smyth, et al. 1951).
Benzotrichloride was severely irritating to the skin of rabbits
that received dermal applications of 10 mg for 24 hours and to the eyes of
rabbits that received instillations of 50 ;jg to the eye (Smyth, et al. 1951).
2
-233-
-------
V. AQUATIC TOXICITY
Pertinent data were not found in the available literature.
VI. EXISTING GUIDELINES AND STANDARDS
Existing guidelines and standards were not found in the available
literature.
-------
REFERENCES
Matsushito, H., et al. 1975. Carcinogenicities of the related compounds in
benzoyl chloride production. 49th Annu. Meeting Japan Ind. Hyg. Soc.,
Sappro, Japan, p. 252.
Sidi, H. 1964. Benzyl chloride, benzal chloride, benzotrichloride. In:
Kirk-Othmer Encyclopedia of Chemical Technology. John Wiley and Sons, New
York, p. 281.
Smyth, H.F., et al. 1951. Range finding taxicity data: List IV. Amer.
Med. Assoc. Arch, of Ind. Health. 4: 119.
Windholz, M. (ed.) 1976. Merck Index, 9th ed. Merck and Co., Inc.,
Rahway, NJ.
Yasuo, K., et al. 1978. Mutagenicity of benzotrichloride and related com-
pounds. Mutat. Res. 58: 143.
-------
No. 21
Benzyl Chloride
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.
-------
BENZYL CHLORIDE
Summary
Benzyl chloride has been shown to produce carcinogenic effects in rats
following subcutaneous administration and in mice following intraperitoneal
administration.
Weak mutagenic activity of the compound has been demonstrated in the
Ames Salmonella assay and in £._ coli.
There is no available information on the possible teratogenic or ad-
verse reproductive effects of benzyl chloride.
Inhibition of cell multiplication in the alga, Microcystis aeruginosa,
started at 30 mg/1. Concentrations of K> mg/1 and 17 mg/1 caused paralysis
in two species of fish.
-------
I. INTRODUCTION
Benzyl chloride (alpha-chlorotoluene), CAS Registry number 100-44-7, is
a colorless-to-light yellow, clear, lachrymatory liquid and is made by free-
radical (photochemical) chlorination of tolene (Hawley, 1971; Austin, 1974).
It has the following physical and chemical properties (Windholz, et al.
1976; Hawley, 1971; Weast, 1972):
Formula:
. Molecular Weight: 126.59
Melting Point: -43°C
Boiling Point: 179°C
Density: 1'1002n
Solubility: Miscible in alcohol, chloroform,
ether; insoluble in water
Production: approximately 89 million Ibs. 1977
(NIOSH, 1977)
\ Benzyl chloride is a moderately volatile compound with a vapor pressure
of 1 mm Hg at 22°C (NIOSH, 1978). The compound decomposes relatively
slowly in water with a 15-hour half-life of pH 7 (25°C) (NIOSH, 1978).
Benzyl chloride is used to make benzaldehyde through additional chlori-
nation and hydrolysis, but modest amounts are also used as a benzylating
agent for benzyl benzoate, n-butyl benzyl phthalate, benzyl ethyl aniline,
benzyl cellulose, components of dyes and perfumes, and for production of
phenylacetic acid by benzyl cyanide (Austin, 1974).
II. EXPOSURE
A. Water
Gruber (1975) reports that no benzyl chloride enters the water
from production.
-------
B. Food
Pertinent data could not be located in the available literature.
C. Inhalation
Pertinent data were not found in the available literature; how-
ever, benzyl chloride is used exclusively as a chemical intermediate in
manufacturing and exposure and is most likely limited to the workplace. As
such, the level of exposure is reported to be less than 1 ppm (NIOSH, 1978).
D. Dermal
Pertinent data could not be located in the available literature.
III. PHARMACOKINETICS
A. Absorption and Distribution
Pertinent data could not be located in the available literature.
B. Metabolism and Excretion
The major excretion product following ingestion of benzyl chloride
is a cysteine conjugate, benzylmercapturic acid (Stekol, 1938, 1939; Witter,
1944; Barnes, et al. 1959; Knight and Young, 1958).
Bray, et al. (1958) administered benzyl chloride at 200 mg/kg body
weight orally to rabbits. Urine collected for 24 hours showed 86.4 percent.
of the administered dose in the soluble fraction, with 49 percent as benzyl-
mercapturic acid, 20 percent as a glycine conjugate, 0.4 percent as glucosi-
duronic acid, and 17 percent as unconjugated benzoic acid. Maitrya and Vyas
(1970) found 30 percent of the total oral dose of benzyl chloride to be ex-
creted by rats as hippuric acid.
Knight and Young (1958) found that benzyl chloride is converted
directly to benzyl mercapturic acid, unlike related compounds such as chlor-
inated benzenes, which form acid-labile precursors. '
-------
Barnes, et al. (1959) found that 27 percent of the total oral dose
of benzyl chloride administered to rats was excreted as benzyl mercapturic
acid. This value compares with 49 percent excreted in rabbits (Bray, et al.
1958) and 4 percent in guinea pigs (Bray, et al. 1959).
Several studies have indicated that glutathione is the source of
the thiol groups for mercapturic acid formation from benzyl chloride
(Barnes, et al. 1959; Simkin and White, 1957; Anderson and Mosher, 1951;
Waelsch and Rittenberg, 1942; Sray, et al. 1969; Beck, et al. 1964). The
turnover rate of glutathione in the liver was found to be 49 mg/100 g of
liver per hour.(Simkin and White, 1957). An in vitro study by Suga, et al.
(1966) revealed that conjugation with glutathione can occur both enzymatic-
ally and non-enzymatically in rat liver*preparations. The enzymic conjuga-
tion has also been observed in human liver preparations (Boyland and Chas-
seaud, 1969).
IV. EFFECTS
A. Carcinogenicity
Benzyl chloride was reviewed by IARC (1976) and found to be car-
cinogenic in rats. Druckrey, et al. (1970) injected 14 rats subcutaneously
with benzyl chloride at 2.1 g/kg body weight (total dose) and 8 rats with
3.9 g/kg body weight (total dose) during 51 weeks. Injection site sarcomas
were noted in three of the rats receiving the lower dose and six receiving
the higher dose; most of the tumors had metastasized to the lungs. The
vehicle of administration, arachis oil, did not produce local tumors.
Poirier, et al. (1975) administered intraperitoneal injections of
benzyl chloride in tricaprylin to three groups of 20 male and female A/Hes-
ton mice, three times per week for eight weeks, with total doses of 0.6,
1.5, and 2.0 g/kg body weight. After 24 weeks, all survivors were killed;
-------
lung tumors occurred In 4/15, 7/16, and 2/8 surviving mice in the three
"groups, respectively. The average number of tumors per mouse was 0.26,
0.50, and 0.25, respectively. The incidence of tumors in mice receiving the
benzyl chloride was not significantly different from the results recorded
for untreated mice on the tricaprylin-vehicle treated mice.
8. Mutagenicity
McCann, et al. (1975a,b) found benzyl chloride to be weakly muta-
genic (less than 0.10 revertants/nanomole) when tested using the Ames assay
(Salmonella/microsomal activation).
Rosenkranz and Poirier (1978), in a National Cancer Institute re-
port, found benzyl chloride to be marginally mutagenic in the Ames assay at
doses of 5 ul and 10 jjl/plate without activation. Microsomal activation had
an inactivating effect on benzyl chloride. The investigators also evaluated
the DNA-modifying activity in bacterial systems using Escherichia coli pol A
mutants. A dose of 10 ul benzyl chloride produced a positive mutagenic ef-
fect.
Benzyl chloride was found to be non-mutagenic in the Ames Salmo-
nella microsomal assay by Simmon (1979). The compound was mutagenic when
exposure was by vapor phase in a dessicator.
C. Teratogenicity, Other Reproductive Effects and Chronic Toxicity
Pertinent data could not be located in the available literature.
0. Acute Toxicity
A number of studies have been conducted on the acute toxicity of
benzyl chloride vapor to animals and were reviewed in a criteria document
prepared by NIOSH (1978). Respiratory tract inflammation and secondary in-
fections were observed in mice exposed to 390 mg/m3 (LCcn) for two hours
and rats exposed to 740 mg/m^ (LC5Q) for two hours (Mikhailova, 1965).
-------
Rabbits exposed to 480 mg/m3 of benzyl chloride for eight hours/day for
six days suffered mild eye and nasal irritation by the sixth day, while cats
exposed to the same regimen suffered a loss of appetite in addition to eye
and respiratory tract irritation (Wolf, 1912). Death of a dog occurred
within 24 hours of exposure to 1,900 mg/m3 of benzyl chloride for eight
hours. Corneal turbidity and irritation of the ocular, respiratory, and
oral mucosa were observed before death (Schutte, 1915). Mikhailova (1965)
observed hepatic changes and necrosis of the kidney in rats and mice exposed
to benzyl chloride at 100 mg/m3.
Landsteiner and Jacobs (1936) investigated the sensitizing proper-
ties of benzyl -chloride to guinea pigs. Benzyl chloride, in a saline solu-
tion (0.01 mg/animal) was injected intracutaneously twice per week for 12
weeks. Two weeks later, re-exposure revealed that benzyl chloride had a
sensitizing effect.
Occupational exposures to benzyl chloride have been reported by
several investigators (Wolf, 1912; Schutte, 1915; Mikhailova, 1971; Katz and
"Talbert, 1930; Watrous, 1947). Lacrimlnation, conjunctivitis, and irrita-
tion of the respiratory tract and eyes have been reported following exposure
to benzyl chloride vapor levels ranging from 6 to 8 mg/m3 for five minutes
to brief exposure at 23,600 mg/m3. Although no cases were reported in the
literature, liquid benzyl chloride has the potential for skin irritation
based on its release of hydrochloric acid upon hydrolysis. The odor thresh-
old and nasal irritation thresholds for benzyl chloride are 0.21 to 0.24
mg/m3 and 180 mg/m3, respectively (Katz and Talbert, 1930; Leonardos, et
al. 1969).
-------
V. AQUATIC TOXICITY
A. Acute and Chronic Toxicity
Pertinent data could not be located in the available literature.
B. Plant Effects
Inhibition of cell multiplication in Microcystis aeruginosa start-
ed at 30 mg/1 (Bringtnann and Kuhn, 1976).
C. Residues
Pertinent data could not be located in the available literature.
D. Other Relevant Information
Hiatt, et al. (1953) found that 1.0 mg/1 of benzyl chloride pro-
duced no irritant response in marine fish, but 10 mg/1 caused a slight irri-
tant activity. This compound caused paralysis in the fish Trutta iridea and
Cyprinus carpio at concentrations of 10 mg/1 and 17 mg/1, respectively
(Meinck, et al. 1970),.
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
The American Conference of Governmental and Industrial Hygienists
(ACGIH, 1977) recommends an occupational exposure limit of 1 ppm (5 mg/m^)
for benzyl chloride. The U.S. federal standard promulgated by OSHA is also
1 ppm (TWA) (29 CFR 1910.1000). NIOSH recommends an environmental exposure
.* '
limit of 5 mg/m3 as a ceiling value for a 15-minute exposure (NIOSH, 1978).
8. Aquatic
No guidelines to protect fish and saltwater organisms from benzyl
chloride toxicity have been established because of the lack of available
data.
-------
REFERENCES
American Conference of Governmental Industrial Hygienists. 1977. Threshold
Limit Values for Chemical Substances and Physical Agents in the Workroom
Environment. Cincinnati, Ohio.
Anderson, E.I. and W.A. Mosher. 1951. Incorporation of S35 fr0m dl-cys-
tine into glutathione and protein in the rat. Jour. Biol. Cham. 188: 717.
Austin, G.T. 1974. The industrially significant organic chemicals. Chem.
Eng. 81: 132.
Barnes, M.M., et al. 1959. The formation of mercapturic acids — I. Forma-
tion of mercapturic acid and the levels of glutathione in tissues. Biochem.
Jour. 71: 680.
Beck, L.V., et al. 1964. Effects of bromobenzene on mouse tissue sulfhy-
dryl and insulin -U31 metabolism. Proc. Soc. Exptl. Biol. Med. 116: 283.
Boyland, E. and !_. Chasseaud. 1969. Glutathione S-aralkyltransferase.
Biochem. Jour. 115:. 985.
Bray, H.G., et al. 1958. Metabolism .of some omega-halogenoalkylbenzenes
and related alcohols in the rabbit. Biochem. Jour. 70: 570.
Bray, H.G., et al. 1959. The formation of mercapturic acids — II. The
possible role of glutathionase. Biochem. Jour. 71: 690.
Bray, H.G., et al. 1969. Some observations on the source of cysteine for
mercapturic acid formation. Biochem. Pharmacol. 18: 1203.
Bringmann, G. and R. Kuhn. 1976. Vergleichende Befunde der Schadwirkung
wassergefahrdender Stoffe gegen Bakterien (Sgeudomonas putida) und Blaualgen
(Microcystis aeruginosa), nwf-wasser/abwasser,T (117)H.9.
Oruckrey, H., et al. 1970. Carcinogenic alkylating substances — III.
Alkyl-halogenides, -sulfates, -sulfonates and strained heterocyclic com-
pounds. (Trans, of German) Z Krebsforsch 74: 241.
Gruber, G.I. 1975. Assessment of industrial hazardous waste practices, or-
ganic chemicals, pesticides, and explosives industries. TRW Systems Group,
NTIS-P8-251-307.
Hawley, G.G. (ed.) 1971. The Condensed Chemical Dictionary, 8th ed. Van
Nostrand Reinhold Company, New York.
Hiatt, R.W., et al. 1953. Relation of chemical structure to irritant re-
sponses in marine fish. London Nature. 172: 904.
International Agency for Research on Cancer. 1976. Monographs on the Eval-
uation of the carcinogenic risk of chemicals to humans. Vol. 11: 217.
-------
Katz, S.H. and E.J. Talbert. 1930. Intensities of odors and irritating
effects of warning agents for inflammable and poisonous gases, Paper 480.
U.S. Department of Commerce, Bureau of Mines. 37 pp.
Knight, R.H. and L. Young. 1958. Biochemical studies of toxic agents —
II. The occurrence of premercapturic acids. Biochem. Jour. 70: 111.
Landsteiner, K. and J. Jacobs. 1936. Studies on the sensitization of ani-
mals with simple chemical compounds, II. Jour. Exp. Med. 64: 625.
Leonardos, G., et al. 1969. Odor threshold determinations of 53 odorant
chemicals. Jour. Air Pollut. Control Assoc. 29: 91.
Maitrya, B.B. and C.R. Vyas. 1970. Studies on conjugation of organic com-
pounds in the rat. Ind. Jour. Biochem. 7: 284.
McCann, .J., et al. 1975a. Detection of carcinogens as mutagens — Bacter-
ial tester strains with R factor plasmids. Proc. Natl. Acad. Sci., USA.
72: 979.
McCann, J., et al. 1975b. Detection of carcinogens as mutagens in the Sal-
monella/microsome test — Assay of 300 chemicals. Proc. Natl. Acad. Sci.,
USA. 72: 5135.
Meinck, F.., et al. 1970. Les eaux residuaires industrielles.
Mikhailova, T.v. 1965. Comparative toxicity of chloride derivatives of
toluene — Benzyl chloride, benzal chloride and benzotrichloride. Fed.
Proc. (Trans. Suppl.) 24: T877.
Mikhailova, T.V. 1971. Benzyl chloride In: ILO Encyclopedia of
Occupational Health and Safety, Vol. 1. Geneva, International Labour
Office: 169.
National Institute for Occupational Safety and Health. 1977. Information
profiles on potential occupational hazards, benzyl chloride. DHEW, 210-77-
0120.
National Institute for Occupational Safety and Health. 1978. Criteria for
a Recommended Standard...Occupational Exposure to Benzyl Chloride. DHEW
78-182.
Poirier, L.A., et al. 1975. Bioassay of alkyl halides and nucleotide base
analogs by pulmonary tumor response to strain A mice. Cancer Res. 35: 1411.
Rosenkranz, H.S. and L.A. Poirier. 1978. An evaluation of the mutagenicity
and DNA-modifying activity in microbial systems of carcinogens and noncarci-
nogens. Unpublished report from U.S. Oept. of Health, Education and Wel-
fare, Public Health Service, National Institute of Health, National Cancer
Institute. 56 pp.
»
Schutte, H. 1915. Tests with benzyl and benzal chloride. Dissertation
translated from German. Wurzburg. Royal Bavarian Julius-Maximilians Uni-
versity, Franz Staudenraus Book Printing. 27 pp.
-J'/S"-
-------
Simkin, J.L. and K. White. 1957. The formation of hippuric acid -- The
influence of benzoate administration on tissue glycine levels. Biochem.
Jour. 65: 574.
Simmon, V.F. 1979. Ln vitro mutagenicity assays of chemical carcinogens
and related compounds with Salmonella typhimuriurn. Jour. Natl. Cancer Inst.
62: 893.
Stekol, J.A. 1938. Studies on the mercapturic acid synthesis in animals —
IX. Jour. 8iol. Chem. 124: 129.
Stekol, J.A. 1939. Studies on the mercapturic acid synthesis in animals ~
XII. The detoxification of benzyl chloride, benzyl alcohol, benzaldehyde,
and S-benzyl homocysteine in the rabbit and rat. Jour. Biol. Chem.
128: 199.
Suga, T., et al. 1966. Studies on mercapturic acids, effect of some aro-
matic compounds on the level of glutathione and the activity of glutathion-
ase in the rat. Jour. Biochem. 59: 209.
Waelsch, H. and D. Rittenberg. 1942. Glutathione — II. The metabolism of
glutathione studied with isotopic ammonia and glutamic acid. Jour. Biol.
Chem. 144: 53.
Watrous, R.M. 1947. Health hazards of the pharmaceutical industry. Br.
Jour. Ind. Med. 4: 111.
Weast, R.C. 1972. Handbook of Chemistry and Physics, 53rd ed. Chemical
Rubber Company, Cleveland, Ohio.
Windholz, M., et al. 1976. Merck Index, 9th ed. Merck and Co., Inc., Rah-
way, New Jersey.
Witter, R.F. 1944. The metabolism of monobromobenzene, benzene, benzyl
chloride and related compounds in the rabbit. Ann Arbor, University of
Michigan, University Microfilms, Dissertation. 1-7, 32-35, 37-66, 93, 197,
113-118, 126-138.
Wolf, W. 1912. Concerning the Effect of Benzyl Chloride and Benzal Chlor-
ide on the Animal Organisms. Translation of dissertation from German, Wurz-
burg, Royal Bavarian Julius-Maximilians University. Franz Staudenraus Book
Printing, 25 pp.
-------
No. 22
Beryllium
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.
-------
SPECIAL NOTATION
U.S. EPA1s Carcinogen Assessment Group (CAG) has evaluated
beryllium and has found sufficient evidence to indicate
that this compound is carcinogenic.
-.249-
-------
BERYLLIUM
SUMMARY
Beryllium was shown to be carcinogenic in three animal
species, producing cancers of the lung and bone when admin-
istered by injection, inhalation, or intratracheal instilla-
tion. Ingestion of beryllium has failed to produce cancers
in animals, possibly due to its poor gastr'ointestinal absorp-
tion. Several epidemiology studies support the hypothesis
that beryllium is a human carcinogen.
..Beryllium is toxic to freshwater organisms at concentra-
tions as low as 5.3 ^ag/1. Pertintent data for marine or-
ganisms were not found in-the available literature (U.S. .
EPA, 1979).
-2SO-
-------
BERYLLIUM
.I. INTRODUCTION
This profile is primarily based upon the Ambient Water
Quality Criteria Document for Beryllium (U.S. EPA, 1979).
Recent comprehensive reviews on the hazards of beryllium
have also been prepared by the National Institute for Occupa-
tional Safety and Health (NIOSH, 1972) and the International
Agency for Research on Cancer (IARC, 1972).
Beryllium (Be; atomic weight 9.01) is^a dark gray metal
of the alkaline earth family. Beryllium has the following
physical-chemical properties (IARC, 1972):
Boiling point: 2970°C
Melting point: 1284 - 1300°C
Hardness: 60 - 125
Density: 1.84 - 1.85
Solubility: Soluble in acids and alkalis
World production of beryllium was reported as approximately
250 tons annually, but much more reaches the environment
as emissions from coal burning operations (Tepper, 1972).
Most common beryllium compounds are readily soluble in water.
,-
The hydroxide is soluble only to the extent of 2 mg/1 (Lange,
1956). Beryllium forms chemical compounds in which its
valence is +2. At acid pH, it behaves as a cation but forms
anionic complexes at pH greater than 8 (Krejci and Scheel,
1966). The major source of beryllium in the environment
is the combustion of fossil fuels (Tepper, 1972). Beryl-
lium enters the waterways through weathering of rocks and
•
soils, through atmospheric fallout and through discharges
from industrial and municipal operations.
-------
II. EXPOSURE
A. Water
Kopp and Kroner (1967) reported the results of
trace metal analyses of 1,577 drinking water samples obtained
throughout the United States. Beryllium was detected in
5.4 percent of the samples. Concentrations ranged from
0.01 to 1.22 ug/1, with a mean value of 0.19 ug/1.
B. Food
Beryllium has been detected in 3- variety of vege-
tables, and in eggs, milk, nuts, bread, and baker's yeast
(Meehan and Smythe, 1967; Petzow and Zorn, 1974). Measured
levels of beryllium were generally in the range of 0.01
to 0.5 ppm. Using the data for consumption and bioconcen-
tration for freshwater and saltwater fishes, mollusks, and
decapods, and the measured steady-state bioconcentration
factor (BCF) for beryllium in bluegills, the U.S. EPA (1979)
has estimated .a weighted average BCF for beryllium to be
19 for the edible portions of fish and shellfish consumed
by Americans.
C. Inhalation
The detection of beryllium in air is infrequent
and usually in trace amounts. In urban areas beryllium
levels may reach 0.008 jag/m3, while in rural areas beryllium
concentrations have been measured at 0.00013 pg/m (Tabor
and Warren, 1958; National Air Sampling Network, 1968).
•
At a beryllium extraction plant in Ohio, beryllium concen-
trations were generally around 2 pg/m over a seven year
period (Breslin and Harris, 1959).
-------
III. PHARMACOKINETICS
Ingested beryllium is poorly absorbed within the gastro-
intestinal tract, presumably due to solubility problems
in the alimentary canal (Hyslop, et al. 1943; Reeves, 1965).
When inhaled, soluble beryllium compounds are rapidly re-
moved from the lung, whereas insoluble beryllium compounds
can remain in the lung indefinitely (Van Cleave and Kaylor,
1955; Wagner, et al. 1969; Sprince, et al. 1976). When
parenterally administered, beryllium is distributed to all
tissues, although it shows preferential accumulation in
bone, followed by spleen, liver, kidney and muscle (Van
Cleave and Kaylor, 1955; Crowley, et al. 1949; Klemperer,
et al. 1952; Kaylor and Van Cleave, 1953; Spencer, et al.
1972). Absorbed beryllium tends to be either excreted in
the urine or deposited in kidneys and bone (Scott, et al.
1950). Once deposited in the skeleton, beryllium is removed
very slowly, with half-lives of elimination reported to
be 1,210, 890, 1,770 and 1,270 days in mice, rats, monkeys,
and dogs, respectively (Furchner, et al. 1973).
IV. EFFECTS
A. Carcinogenicity
Beryllium was shown to be carcinogenic in three
animal species. Intravenous injection of beryllium, zinc
beryllium silicate, and beryllium phosphate produced osteo-
sarcomas in the rabbit (Gardner and Heslington, 1946; Dutra
and Largent, 1950; Komitowski, 1969; Fodor, 1971; IARC,
1972). Inhalation and intratracheal'instillation of beryl-
-------
lium compounds have produced lung cancers in the rat and
monkey (Vorwald and Reeves, 1959; Vorwald, et al. 1966;
Reeves, et al. 1967). Ingestion of beryllium by rats and
mice has failed to induce tumors, possibly due to the poor
absorption of beryllium from the gastrointestinal tract.
Several epidemiological studies have failed to
establish a clear association between beryllium exposure
and cancer development {Stoeckle, et al. 1969; Mancuso,
1970; Niemoller, 1963). However, other recent studies sup-
port the hypothesis that beryllium is a human carcinogen
(Berg and Burbank, 1972; Wagoner, et al. 1978; Discher,
1978) .
B. Mutagenicity
Pertinent data were not found in the available
literature.
C. Teratogenicity
Beryllium has been implicated as a teratogen in
snails (Raven and Sprok, 1953) and has inhibited limb re-
generation in the salamander, Amblystoma punctatum (Thorton,
1950).
D. Other Reproductive Effects
Pertinent data were not found in the available
literature.
E. Chronic Toxicity
Chronic beryllium inhalation in humans produces
a progressive, systemic disease which may follow the ces-
sation of exposure by as long as five years (Tepper, et
-------
al. 1361; Hardy and Stoeckle, 1959}. -Symptoms include pneu-
monitis with cough, chest pain, and general weakness. Sy-
stemic effects include right heart enlargement with cardiac
failure, enlargement of liver and spleen, cyanosis, digital
clubbing, and kidney stones (Hall, et al. 1959). Chronic
beryllium disease can be produced in rats and monkeys by
inhalation of beryllium sulfate at 35 /ag/m (Schepers, et
al. 1957; Vorwald, et al. 1966).
V. AQUATIC TOXICITY
A. Acute Toxicity
Acute toxicity data for beryllium for freshwater
fishes are taken from 22 static and 5 flow-through bioassays,
all 96 hours in duration. U.S. EPA (1979) presents the
most sensitive species, the guppy Poecilia reticulata, with
LC^Q values ranging from 71 to 17,500 pg/1. The data re-
flect that the toxicity of beryllium to freshwater fish
is decreased in hard water. This has also been confirmed
by U.S. EPA (1979) in the fathead minnow, Pimephales prome^
lag, with LC5Q values ranging from 82 to 11,000 ug/1. Acute
toxicity for aquatic invertebrates provides two 48-hour
LCcQ values of 7,900 and 2,500 ug/1, with water hardness
values of 180 and 200 fig/1 as CaCo-,. The source of these
invertebrate studies is the same for chronic freshwater
studies. No data for acute toxicity to marine species was
found in the available literature.
Sf
-------
B. Chronic Toxicity
No chronic tests for freshwater fish were found
in the available literature. The cladoceran, Daphnia magna,
was the only freshwater species tested for chronic effects;
chronic values of less than 36 jag/1 and 5.3 ug/1 were ob-
tained by the U.S. EPA (1973). No chronic data for marine
species of fish or invertebrates was found in the available
literature.
C. Plant Effects
The only plant study available reveals that the
green algae, Chlorella vannieli, displayed growth inhibition
at a concentration of 100,000 ug/1 (U.S. EPA, 197y).
D. Residues
Exposure of the bluegill for 28 days produced
a bioconcentration factor of 19 (U.S. EPA, 1978). No other
data was found in the available literature.
E. Other Relevant Information
The only marine data presented showed reduced
alkaline phosphatase activity in the mummichog, Fundulus
heteroclitus, at concentrations as low as 9 )ug/l - A tera-
togenic response was observed by Evola-Maltese (1957) in
sea urchin embryos at concentrations of 9.010 ug/1.
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
The present standard for occupational exposure
to beryllium prescribes an 8-hour time-weignted average
-------
of 2.0 ug/ro with a ceiling concentration of 5.0 jig/m .
This is the same value recommended by the American Confer-
ence of Governmental Industrial Hygienists (1977}. The
National Institute for Occupational Safety and Health (NIOSH,
1972) recommends that occupational exposure to beryllium
and its compounds not exceed 1 ug/m {8-hour time-weighted
average) with a ceiling limit of 5 pg/m (measured over
a 15 minute sampling period).
National Emission Standards for Hazardous Air
Pollutants set their criterion as not more than 10 g in
24 hours or emissions which result in maximum outplant con-
centrations of 0.01 ^g/m3, 30-day average (U.S. EPA, 1977).
Based on animal bioassay data for beryllium to
which the linear model was applied, the U.S. EPA (1979)
has estimated levels of beryllium in ambient water which
will result in carcinogenic risk for humans. As a result
of the public comments received, additional review and re-
evaluation of the data base is required before a final cri-
terion level can be recommended.
B. Aquatic
The U.S. EPA proposed a water quality standard
of 11 ^ug/1 for the protection of aquatic life in soft fresh-
water; 1,100 pg/1 for the protection of aquatic life in
hard freshwater; and 100 ug/1 for continuous irrigation
on all soils, except 500 mg/1 for irrigation on neutral
to alkaline lime-textured soils (U.S. EPA, 1977).
-------
The National Academy of Science/National Academy
of Engineering (1973) Water Quality Criteria recommendation
for marine aquatic life is: hazard level - 1.5 ug/1; minimal
risk of deleterious effects - 0.1 mg/1; and application
factor - 0.01 (applied to 96-hour LC5Q). Their recommenda-
tion for irrigation water is: 0.10 mg/1 for continuous use
on all soils.
The U.S. EPA (1979) has derived a draft criterion
for beryllium to protect freshwater aquatic organisms.
The 24-hour average concentration in ug/1 is dependent on
water hardness and is derived by the following equation:
OR = e^1'24 ln (hardness> ~ 6.65)
The concentration not to be exceeded at any time is:
CR _ e (1.24 In (hardness) - 1.46)
No draft criterion was derived for marine organisms (U.S.
EPA, 1979).
-------
BERYLLIUM
REFERENCES
American Conference of Governmental Industrial Hygienists
1977. Threshold limit values for chemical substances in
workroom air adopted by ACGIH for 1977. ACGIH, P.O. Box
1937, Cincinnati, Ohio 45201.
Berg, J.W., and F. Burbank. 1972. Correlations between
carcinogenic trace metals in water supply and cancer mor-
tality. Ann. N.Y. Acad. Sci. 199: 249.
Breslin, A.J., and W.B. Harris. 1959. Health protection
in beryllium facilities. Summary of ten years of experience.
AMA Arch Ind. Health 19: 596.
Crowley, J.F., et al. 1949. Metabolism of carrier-free
radioberyllium in the rat. Jour. Biol. Chem. 177: 975.
Discher, D.P. 1978. Letter to W.H. Foege, Director, Center
for Disease Control HEW {published in SNA Occupational Safety
and Health Reporter) 8: 853.
Dutra, F.R., and F.J. Largent. 1950. Osteosarcoma induced
by beryllium oxide. Am. Jour. Pathol. 26: 197.
Evola-Maltese, C. 1957. Effects of beryllium on the develop-
ment and alkaline phosphatase activity of Paracentrotus
embryos. Acta Embryol. Morphol. Exp. 1: iTTI
Fodor, J. 1971. Histogenesis of bone tumors induced by
beryllium. Magyar Onkol. 15: 180.
Furchner, J.E., et al. 1973. Comparative metabolism of
radionucleotides in mammals. VIII: Retention of beryllium
in the mouse, rat, monkey, and dog. Health Physics 24:
293.
Gardner, L. U., and H.F. Heslington. 1946. Osteo-sarcoma
from intravenous beryllium compounds in rabbits. Fed. Proc.
5: 221.
Hall, T.C., et al. 1959. Case data from the beryllium
registry. AMA Arch. Ind. Health 19:100.
Hardy, H.L., and J.D. Stoeckle. 1959. Beryllium disease.
Jour. Chron. Dis. 9: 152.
Hyslop, F., et al. 1943. The toxicology of beryllium.
U.S. Pub. Health Serv. Natl. Inst. Health Bull. 181.
IARC. 1972. Monographs on the evaluation of carcinogenic
risk of chemicals to man. Beryllium: 1: 17.
-------
Kaylor, C.T. , and C.D. Van Cleave. 1953. Radiographic
visualization of the deposition of radioberyllium in the
rat. Anat. Record 117: 467.
Klemperer, P.W., et al. 1952. The fate of beryllium com-
pounds in the rat. Arch. Biochem. Biophys. 41: 148.
Komitowski, D. 1969. Morphogenesis of beryllium-induced
bone tumors. Patol. Pol (suppl.) 1: 479.
Kopp, J.F., and R.C. Kroner. 1967. A five year study of
trace metals in waters of the United States. Fed. Water
Pollut. Control Admin., U.S. Dep. Inter., Cincinnati, Ohio.
Krejci, L.E., and L.D. Scheel. 1966. In H.E. Stokinger,
ed. Beryllium: Its industrial hygiene aspects. Academic
Press, Inc., New York.
Lange, N.A. ed. 1956. Lange's handbook of chemistry.
9th ed. Handbook Publishers, Inc., Sandusky, Ohio.
Mancuso, T.F. 1970. Relation of duration of employment
and prior illness to respiratory cancer among beryllium
workers. Environ. Res. 3: 251.
Meehan, W.R., and L.E. Smythe. 1967. Occurrence of beryl-
lium as a trace element in environmental materials. Environ.
Sci. Technol. 1: 839.
National Academy of Sciences, National Academy of Engineer-
ing. 1973. Water quality criteria 1972. A report. Natl.
Acad. of Sci., Washington, D.C.
National Air Sampling Network, Air Quality Data. 1968.
National Air Sampling Network, Durham, N.C., U.S. Dep. Health
Education and Welfare, Pub. Health Serv.
National Institute of Occupational Safety and Health. 1972.
Criteria for a recommended standard...Occupational exposure
to beryllium.. DHEW (NIOSH) ?ubl. No. 72-10806.
Niemoller, H.K. 1963. Delayed carcinoma induced by beryl-
lium aerosol in man. Int. Arch. Gewerbepthol. Gewerbehyg.
20: 18.
Petzow, G., and H. Zorn. 1574. Toxicology of beryllium-
containing materials. Chemlker Vig. 98: 236.
Raven, C.P., and N.S. Spronk. 1953. Action of beryllium
on the development of Limnaea stagnalis. Chem. Abstr. •
47: 6561.
-2C.O-
-------
Reeves, A.L. 1965. Absorption of beryllium from the gastro-
intestinal tract. AMA Arch. Environ. Health 11: 209.
Reeves, A.L., et al. 1967. Beryllium carcinogenesis. I.
Inhalation exposure of rats to beryllium sulfate aerosol.
Cancer Res. 27: 439.
Schepers, G.W.H., et al. 1957. The biological action of
inhaled beryllium sulfate. A preliminary chronic toxicity
study in rats. AMA Arch. Ind. Health 15: 32.
Scott, J.K., et al. 1950. The effect of add^d carrier
on the distribution an<
Biol. Chem. 172: 291.
on the distribution and excretion of soluble Be. Jour.
Spencer, H.C., et al. 1972. Toxicological evaluation of
beryllium motor exhaust products. AMRL-TR-72-118. Aero-
medical Res. Lab. Wright-Patterson AFB, Ohio.
Sprince, N.L., et al. 1976. Current (1975) problems of
differentiating between beryllium disease and. sarcoidosis.
Stoeckle, J.D., et al. 1969. Chronic beryllium disease:
Long-term follow-up of 60 cases and selective review of
the literature. Am. Jour. Med. 46: 545.
Tabor, E.G., and W.V. Warren. 1958. Distribution of cer-
tain metals in the atmosphere of some American cities.
Arch. Ind. Health. 17: 145.
Tepper, L.B. 1972. Beryllium. CRC critical reviews in
toxicology. 1: 235.
Tepper, L.B., et al. 1961. Toxicity of beryllium compounds.
Elsevier Publishing Co., New York.
Thornton, C.S. 1950. Beryllium inhibition of regenerations.
Jour. Exp. Zool. 114: 305.
U.S. EPA. 1977. Multimedia environmental goals for environ-
mental assessment. Vol. II. MEG charts and background inform-
ation. EPA-60017-77-136b. U.S. Environ. Prot.Agency.
U.S. EPA. 1978. In-depth studies on health and environmental
impacts of selected water pollutants. U.S. Environ. Prot.
Agency, Washington, D.C.
U.S. EPA. 1979. Beryllium: Ambient Water Quality Criteria.
U.S. Environ. Prot. Agency, Washington, D.C.
Van Cleave, C.D., and C.T. Kavlor. 1955. Distribution,
retention and elimination of Be in the rat after intrat
cheal injection. AMA Arch. Ind. Health 11: 375.
-------
Vorwald, A.J., and A.L. Reeves. 1959. .Pathologic changes
induced by beryllium compounds. AMA. Arch. Ind, Health
19: 190.
Vorwald, A.J., et al. 1966. Experimental beryllium toxi-
cology. In H.E. Stokinger, ed. Beryllium, its industrial
hygiene aspects. Academic Press, New York.
Wagner, W.D., et al. 1969. Comparative inhalation toxicity
of beryllium ores bertrandite and beryl with production
of pulmonary tumors by beryl. Toxicol. Appl. Pharmacol.
15: 10.
Wagoner, J.K., et al. 1978. Beryllium: carcinogenicity
studies. Science 201: 298.
-------
No. 23
BIs(2-chloroethoxy)methane
Health and Environmental Effects
U.S. ENV.IRONMENTAL 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.
-------
BIS(2-CHLOROETHQXY)METHAME
Summary
Pertinent data could not be located in the available literature search-
es on the mutagenic, carcinogenic, teratogenic, or adverse reproductive ef-
fects of bis(2-chloroethoxy)methane (BCEXM) in mammals. A closely related
compound, bis(2-chloroethoxy)ethane (BCEXE) has been shown to produce skin
tumors and injection site sarcomas in animal studies.
Pertinent information could not be located in the available literature
on bis(2-chloroethoxyJmethane toxicity to aquatic organisms.
-------
BIS(2-CHLOROETHOXY)METHANE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Chloroalkyl Ethers (U.S. EPA, 1979a).
The Chloroalkyl ethers are compounds in which a hydrogen atom in one or
both of the aliphatic ether chains are substituted with chlorine. 8is(2-
chloroethoxy)methane (BCEXM, dichloroethyl formal, C1CH2CH2-0-CH2-
OCH2-CH2C1) is a colorless liquid at room temperature with a boiling
point of 218.1°C and a specific gravity of 1.2339. The compound is
slightly soluble in water but miscible with most organic solvents.
The Chloroalkyl ethers have a wide variety of industrial uses in organ-
ic synthesis, treatment of textiles, the manufacture of polymers and insec-
ticides, as degreasing agents and solvents, and in the preparation of ion -
exchange resins (U.S. EPA, 1979a).
The Chloroalkyl ethers, like BCEXM, have a higher stability in water
than the alpha Chloroalkyl ethers, which decompose. BCEXM is decomposed by
mineral acids.
II. EXPOSURE
No specific information on exposure to BCEXM is available. The reader
is referred to a more general treatment of Chloroalkyl ethers (U.S. EPA,
1979b). BCEXM has been monitored in rubber plant effluents at a maximum
level of 140 rag/1 (Webb, et al. 1973). 8is-l,2-(2-chloroethoxy)ethane
(BCEXE), a closely related compound, has been reported in drinking water at
a maximum level of 0.03 ug/1 (U.S. EPA, 1975). Data on levels of 8CEXM in
foods was not found in the available literature.
NO bioaccumulation factor for BCEXM has been derived.
-------
III. PHARMACOKINETICS
Pertinent information could not be located in the available literature
on BCEXM. The reader is referred to a more general treatment of chloroalkyl
ethers (U.S. EPA, 1979t>).
IV. EFFECTS
A. Carcinogenicity
Pertinent information could not be located in the available litera-
ture on carcinogenic effects of BCEXM. The reader is referred to a more
general treatment of chloroalkyl ethers (U.S. EPA, 1979b). A closely re-
lated compound, BCEXE, has been shown to produce skin tumors in mice and in-
jection site sarcomas (Van Duuren, et al. 1972).
B. Mutagenicity, Teratogenicity, Other Reproductive Effects and Chron-
ic Toxicity
Pertinent data could not be located in the available literature for
BCEXM.
V. AQUATIC TOXICITY
Pertinent information could not be located in the available literature
on the aquatic toxicity of BCEXM.
VI. EXISTING GUIDELINES AND STANDARDS
No standards or recommended criteria exist for the protection of human
health or aquatic organisms to bis(2-chloroethoxy)methane.
-------
BIS(2-CHLOROETHOXY} METHANE
REFERENCES
U.S. EPA. 1975. Preliminary assessment of suspected carcinogens in drink-
ing water: Interim report to Congress, Washington, O.C.
U.S. EPA. 1979a. Chloroalkyl Ethers: Ambient Water Quality Criteria.
(Draft)
U.S. EPA. 1979b. Environmental Criteria and Assessment Office. Chloro-
alkyl Ethers: Hazard Profile. (Draft)
Van Duuren, et al. 1972. Carcinogenicity of haloethers. II. Structure-
activity relationships of analogs of bis(chloromethyl)ether. Jour. Natl.
Cancer Inst. 48: 1431.
Webb, R.G., et al. 1973. Current practice in GC-MS analysis of organics in
water. Publ. EPA-R2-73-277. U.S. Environ. Prot. Agency, Corvallis, Oregon.
-------
No. 24
Bis(2-chloroethyl)ether
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-uracy.
-3.70-
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated
bis{2-chloroethyl)ether and has found sufficient evidence to
indicate that this compound is carcinogenic.
-------
BI5(2-CHLOROETHYL)ETHER
Summary
Oral administration of bis(2-chloroethyl)ether (8CEE) did not produce
an increase of tumors in rats. Male mice showed a significant increase in
hepatomas after ingestion of BCEE. BCEE has also shown activity as a tumor
initiator for mouse skin.
Testing of BCEE in the Ames1 Salmonella assay, in §_._ coli, and in
Saccharomyces cerevisiae has shown that this compound induces mutagenic
effects.
There is no available evidence to indicate that BCEE produces adverse
reproductive effects or teratogenic effects.
The data base for bis(Z-chloroethyl-)ether is limited to three studies.
The 96-hr LC5Q value for the bluegill is reported to be over 600,000 jug/1.
Adverse chronic effects were not observed with the fathead minnow at test
concentrations as high as 19,000 jjg/1. A bioconcentration factor of 11 was
observed during a 14-day exposure of bluegills. The half-life was 4-7 days.
-272-
-------
8I5(2~CHLORO£THYL)ETHER
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Chloralkyl Ethers (U.S. EPA, 1979a).
The chloroalkyl ethers are compounds in which a hydrogen atom in one or
both of the aliphatic ether chains are substituted with chlorine. Bis(2-
chloroethyDether (BCEE, molecular weight 143.01) is a colorless liquid at
room temperature with a boiling point of 176-178°C at 760 mm Hg, and a
density of 1.213. The compound is practically insoluble in water, but is
miscible with most organic solvents (U.S. EPA, 1979a).
The chloroalkyl ethers have a wide variety of industrial and laboratory
uses in organic synthesis, in textile treatment, the manufacture of polymers
and insecticides, as degreasing agents, and in the preparation of ion ex-
change resins (U.S. EPA, 1979a).
The B-substituted chloroalkyl ethers, such as BCEE, are generally more
stable and hence less reactive in aqueous systems than the a-substituted
compounds (U.S. EPA, 1979a).
For additional information regarding chloroalkyl ethers in general, the
reader is referred to the EPA/ECAO Hazard Profile on Chloroalkyl Ethers
(U.S. EPA 1979b).
II. EXPOSURE
The B-chloroalkyl ethers have been monitored in water. Industrial dis-
charges from chemical plants involved in the manufacture of glycol products,
rubber, and insecticides may contain high levels of BCEE (U.S. EPA, 1979a).
*
The highest concentration of BCEE in drinking water reported by the U.S. EPA
-------
(1975) is 0.5 ug/1. There is no evidence of the occurrence of the chloro-
alkyl ethers in the atmosphere; human exposure appears to be confined to
occupational settings.
Human exposure to chloroalkyl ethers via ingestion of food is unknown
(U.S. EPA, 1979a). The B-chloroalkyl ethers, due to their stability and low
water solubility, may have a high tendency to be bioaccumulated. The U.S.
EPA (1979a) has estimated the weighted average bioconcentration factor for
BCEE to be 25 for the edible portions of fish and shellfish consumed by
Americans. This estimate is based on a measured steady-state biocon-
centration factor using bluegills.
III. PHARMACOKINETICS
A. Absorption
Experiments with radiolabelled BCEE have indicated that the com- ;
pound is readily absorbed following oral administration (Lingg, et al.
1978). Information on inhalation or dermal absorption of chloroalkyl ethers
is not available (U.S. EPA, 1979a).
B. Distribution
Pertinent information on the distribution of 8CEE could not be
located in the literature.
C. Metabolism
The biotransformation of BCEE in rats following oral administration
appears to involve cleavage of the ether linkage and subsequent conjugation
with non-protein-free sulfhydryl groups, the major route, or with glucuronic
acid (Lingg, et al. 1978). Thiodiglycolic acid and 2-chloro-
ethanol-B-D-glucuronide were identified as urinary metabolites of BCEE in
rats.
-------
D. Excretion
8CEE administered to rats by intubation was eliminated rapidly in
the urine, with more than 60 percent of the compound excreted within 24
hours (Lingg, et al. 1978).
IV. EFFECTS
A. Carcinogenicity
8CEE has shown activity as a tumor initiator in mouse skin (U.S.
EPA, 1979a). Preliminary results of an NCI study indicate that oral admin-
istration of BCEE does not produce an increase in tumor incidence in rats
(U.S. EPA, 1979a); however, mice administered BCEE by ingestion showed a
significant increase in hepatomas (Innes, et al. 1969).
B. Mutagenicity
Testing of the chloroalkyl ethers in the Ames1 Salmonella assay and.
in §_._ coli have indicated that BCEE induces mutagenic effects (U.S. EPA,
1979a). BCEE has also shown mutagenic effects in Saccharomyces cerevisiae
(Simmon, et al. 1977), but none were found in the heritable translocation
test for mice (Jorgenson, et al. 1977).
C. Teratogenicity, Chronic Toxicity and other Reproductive Effects
Pertinent information could not be located in the available liter-
ature.
0. Other Relevant Information
Acute physiological responses of the guinea pig to inhalation of
high concentrations of BCEE were congestion, emphysema, edema and hemorrhage
of the lungs (Shrenk, et al. 1933). Brief exposure of man to BCEE vapor, at
levels 260 ppm, irritated the nasal passages and eyes with profuse lacri-
mation. Deep inhalation produced nausea. The highest concentration with no
noticeable effect was 35 ppm (Shrenk, et al. 1933).
-------
V. AQUATIC TOXICITY
A. Acute Toxicity
96-hr LC5Q value for the bluegill, Lepomis macrochirus, could not
be determined for bis(2-chloroethyl)ether with exposure concentrations as
high as 600,000 |jg/l (U.S. EPA, 1978).
B. Chronic Toxicity
An embryo-larval test has been reported with bis(2-chloroethyl)
ether and the fathead minnow, Pimephales proroelas. Adverse effects were not
observed at test concentrations as high as 19,000jjg/l (U.S. EPA, 1978).
C. Plant Effects
Pertinent data could not be located in the available literature.
0. Residues
A bioconcentration factor of 11 was determined during a 14-day ex- .
posure of bluegills to bis(2-chloroethyl)ether. The half-life was 4-7 days.
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
Based on the results of an animal carcinogenesis bioassay, and
using a linear, non-threshold model, the U.S. EPA (1979a) has estimated that
an ambient water level of 0.42 ug/1 will present an increased risk of 10"
or less for 8CEE, assuming water and the injection of contaminated aquatic
organisms to be the only sources of exposure.
-------
The 3-hour, time-weighted average threshold limit value (TLV-TWA)
for BCEE determined by the American Conference of Governmental Industrial
Hygienists (ACGIH, 1978) is 5 ppm for 8CEE.
8. Aquatic
Freshwater or saltwater criteria cannot be derived for bis(2-chlo-
roethyDether because of insufficient data (U.S. EPA, 1979a).
-277-
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BIS(2-CHLORDETHYL)ETHER
REFERENCES
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.
Fishbein, L. 1977. Potential industrial carcinogens and mutagens. Publ.
EPA-560/5-77-005, Off. Toxic Subst. Environ. Prot. Agency, Washington, O.C.
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.
Jorgenson, T.A., et al. 1977. Study of the mutagenic potential of
bis(2-chloroethyl) and bis(2-chloroisopropyl) ethers in mice by the heri-
table translocation test. Toxicol. Appl. Pharmacol. 41: 196.
Lingg, R.O., et al. 1978. Fate of bis (2-chloroethyl)ether in rats after
acute oral administration. Toxicol. Appl. Pharmacol. 45: 248.
Schrenk, H.H., et al. 1933. Acute response of guinea pigs to vapors of
some new commercial organic compounds. VII. Dichloroethyl ether. Pub.
Health Rep. 48: 1389.
Simmon, V.F., et al. 1977. Mutagenic activity of chemicals identified in
drinking water. In: D. Scott, et al. (ed.) Progress in genetic toxicology.
Elsevier/North Holland Biomedical Press, New York.
U.S. EPA. 1975. Preliminary assessment of suspected carcinogens in drink-
ing water. Rep. Cong. U.S. Environ. Prot. Agency, Washington, D.C.
U.S. EPA. 1977a. National organic monitoring survey. General review of
results and methodology: Phases I-III. U.S. Environ. Prot. Agency, Off.
Water Supply, Tech. Support Oiv. Presented before Water Supply Res. Div.
Phys. Chem. Removal Branch, Oct. 21.
U.S. EPA. 1977b. Potential industrial carcinogens and mutagens. Office of
Toxic Substances. EPA-560/5-77-005. Washington, D.C.
U.S. EPA. 1978. In-depth studies on health and environmental impacts of
selected water pollutants. U.S. Environ. Prot. Agency, Contract No.
63-1-4646.
U.S. EPA. 1979a. Chloroalkyl Ethers: Ambient Water Quality Criteria.
(Draft)
U.S. EPA. 1979b. Environmental Criteria and Assessment Office. Hazard
Profile: Chloroalkyl Ethers. (Draft).
Van Duuren, 3.L. 1969. Carcinogenic epoxides, lactones, and haloethers and
their mode of action. Ann. N.Y. Acad. Sci. 163: 633.
-------
Van Duuren, B.L. et al. 1969. Carcinogenicity of haloethers. Jour. Natl.
Cancer Inst. 43: 481.
Van Duuren, B.L., et al. 1972. Carcinogenicity of haloethers. II. Struc-
ture-activity relationships of analogs of bis(chloromethyl)ether. Jour.
Natl. Cancer Inst. 48: 1431.
-------
No. 25
Bis(2-Chloroisopropyl)ether
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
-3.9TO-
-------
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.
-------
BIS(2-CHLOROISOPROPYL)ETHER
Summary
Preliminary results from an NCI carcinogenesis bioassay do not show an
increase in tumors following oral administration of bis(2-chloroisopropyl)-
ether (BCIE).
BCIE has produced mutagenic effects in two bacterial test systems (Sal-
monella and §_._ coli) but has failed to show mutagenicity in one mammalian
study.
No information is available on the teratogenic or adverse reproductive
effects of BCIE.
Chronic exposure to BCIE has produced liver damage in animals.
Data on the toxicity of bis(2-chloroisopropyl)ether to aquatic organ-
isms are not available.
-------
BIS(2-CHLDROISOPROPYL)ETHER
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Chloroalkyl Ethers (U.S. EPA, 1979a).
The Chloroalkyl ethers are compounds in which a hydrogen atom in one or
both of the aliphatic ether chains are substituted with chlorine. Bis(2-
chloroisopropyDether (BCIE, molecular weight 171.07) is a colorless liquid
at room temperature with a boiling point of 187-188°C at 760 mm Hg. The
compound is practically insoluble in water but is miscible with organic sol-
vents.
The Chloroalkyl ethers have a wide variety of industrial and laboratory
uses in organic synthesis, treatment of textiles, the manufacture of poly-
mers and insecticides, as degreasing agents, and in the preparation of ion
exchange resins (U.S. EPA, 1979a).
The beta-chloroalkyl ethers, like BCIE, are more stable in aqueous sys-
tem than the alpha-chloroalkyl ethers, which decompose rapidly. For addi-
tional information regarding the Chloroalkyl ethers as a class, the reader
is referred to the Hazard Profile on Chloroalkyl Ethers (U.S. EPA, 1979b).
II. EXPOSURE
The beta-chloroalkyl ethers have been monitored in water. Industrial
discharges from chemical plants involved in the manufacture of glycol pro-
ducts, rubber, and insecticides may present high effluent levels (U.S. EPA,
1979a). The highest concentration of BCIE monitored in drinking water by
the U.S. EPA (1975) was reported as 1.58jug/l.
The concentrations of Chloroalkyl ethers in foods have not been moni-
f
tored. The beta-chloroalkyl ethers, however, due to their relative stabili-
ty and low water solubility, may have a high tendency to be bioaccumulated.
-J233-
-------
The U.S. EPA (1979a) has estimated the weighted average bioconcentration
factor for bis(2-chloroisopropyl)ether to be 106 for the edible portions of
fish and shellfish consumed by Americans. This estimate is based on the
octanol/water partition coefficient.
III. PHARMACOKINETICS
A. Absorption
Experiments with radio-labeled BCIE have indicated that the com-
pound is readily absorbed following oral administration (Smith, et al.
1977). No information on inhalation or dermal absorption of the chloroalkyl
ethers is available (U.S. EPA, 1979a).
B. Distribution
Species differences in the distribution of radio-labeled BCIE have
been reported by Smith, et al. (1977). Monkeys retained higher amounts of
radioactivity in the liver, muscle, and brain than did rats. Urine and ex-
pired air from monkeys also contained higher levels of radioactivity than
those determined in the rat. Blood levels of BCIE in monkeys reached a peak
within 2 hours following oral administration and then declined in a biphasic
manner (t 1/2 = 5 hours and 2 days, respectively).
C. Metabolism
Urinary metabolites of labeled BCIE identified in studies with rats
included l-chloro-2-propanol, propylene oxide, 2-(l-methyl-2-chloro-ethoxy)
propionic acid, and carbon dioxide (Smith, et al. 1977).
D. Excretion
Smith, et al. (1977) found that in the rat, 63.36 percent, 5.87
percent, and 15.96 percent of a 30 mg orally-administered dose of BCIE were
t
recovered after 7 days in the urine, feces, and expired air, respectively.
In the monkey, the corresponding figures were 28.61 percent, 1.19 percent,
and 0 percent, respectively.
-------
IV. EFFECTS
A. Carcinogenicity
Preliminary results of an NCI carcinogenicity bioassay indicate
that oral administration of BCIE does not produce an increase in tumor inci-
dence (U.S. EPA, 1979a).
B. Mutagenicity
Testing of BCIE in the Ames Salmonella assay and in E^ coli have
indicated that the compound shows mutagenic activity (U.S. EPA, 1979a).
BCIE did not show mutagenic effects in the murine heritable translocation
test (Jorgenson, et al. 1977).
C. Teratogenicity and Other Reproductive Effects
Pertinent data could not be located in the available literature.
0. Chronic Toxicity
Chronic oral exposures of mice to BCIE produced centrilobular liver
necrosis in mice. The major effects in rats were pulmonary congestion and
pneumonia (U.S. EPA, 1979a).
E. Other Relevant Information
Several chloroalkyl ethers show initiating activity and therefore
may interact with other agents to produce skin papillomas (Van Duuren, et
al. 1969, 1972); however, data specific to BCIE is not available.
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
8CIE is an isomer of a group of chloroalkyl ethers which have been
shown to have carcinogenic potential. BCIE has been shown to be mutagenic;
however, definitive proof of carcinogenicity has not been demonstrated. The
available data is presently under review and a definitive determination as
to the carcinogenicity of this isomer cannot be made at this time.
B. Aquatic
No draft .criteria to protect fish and saltwater aquatic organisms
from bis(2-chloroisopropyl)ether toxicity have been derived (U.S. EPA, 1979).
-------
BIS(2-CHLOROISOPROPYL)ETHER (BCIE)
REFERENCES
Jorgenson, T., et al. 1977. Study of the mutagenic potential of bis(2-
chloroethyl) and bis(2-chloroisopropyl) ethers in mice by the heritable
translocation test. Toxicol. Appl. Pharmacol. 41: 196.
Smith, C., et al. 1977. Comparative metabolism of haloethers. Ann. N.Y.
Acad. Sci. 298: ill.
U.S. EPA. 1975. Preliminary assessment of suspected carcinogens in drink-
ing water: Interim report to Congress, Washington, O.C.
U.S. EPA. 1979a. Chloroalkyl Ethers: Ambient Water Quality Criteria.
(Draft)
U.S. EPA. 1979b. Environmental Criteria and Assessment Office. Chloro-
alkyl Ethers: Hazard Profile. (Draft)
Van Duuren, B., et al. 1969. Carcinogenicity of haloethers. Jour. Natl.
Cancer Inst. 43: 481.
Van Duuren, B., et al. 1972. Carcinogenicity of haloethers. II. Struc-
ture-activity relationships of analogs of bis(chloroethyl)ether. Jour.
Natl. Cancer Inst. 48: 1431.
-------
No. 26
Bis(Chlororaethyl)ether
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-uracy.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated
bis(chloromethyl}ether and has found sufficient evidence to
indicate that this compound is carcinogenic.
-390-
-------
BIS(CHLOROMETHYL)ETHER
Summary
Sis(chloromethyl)ether (BCME) has been shown to produce tumors in' ani-
mals following administration by subcutaneous injection, inhalation, or der-
mal application. Epidemiological studies of workers in the United States,
Germany, and Japan who were exposed to BCME and chloromethyl methyl ether
(CMME) indicate that these compounds are human respiratory carcinogens.
BCME has produced mutagenic effects in the Ames1 Salmonella assay and
in E._ coli. Increased cytogenetic abnormalities have been observed in the
lymphocytes of workers exposed to BCME and CMME; this effect appeared to be
reversible.
There is no available evidence to indicate that the chloroalkyl ethers
produce adverse reproductive effects or teratogenic effects.
Information has not been found on the toxicity of bis(chloromethyl)
ether to aquatic organisms. The hazard profiles on the haloethers and the
chloroalkyl ethers should be consulted for the toxicity of related compounds.
-------
BIS (CHLOROMETHYL.) ETHER
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Chloroalkyl Ethers (U.S. EPA, 1979a).
The Chloroalkyl ethers are compounds in which hydrogen atoms in one or
both of the aliphatic ether chains are substituted with chlorine. Bis-
(chloromethyl)ether, (BCME; molecular weight 115.0), is a colorless liquid
at room temperature with a boiling point of 104°C at 760 mm Hg, and a den-
sity of 1.328. The compound immediately hydroly2es in water, but is misci-
ble with ethanol, ether, and many organic solvents (U.S. EPA, 1979a).
The Chloroalkyl ethers have a wide variety of industrial and laboratory
uses in organic synthesis, textile treatment, the manufacture of polymers
and insecticides, the preparation of ion exchange resins, and as degreasing
agents (U.S. EPA, 1979a).
While BCME is very unstable in water, it appears to be relatively sta-
ble in the atmosphere (Tou and Kallos, 1974). Spontaneous formation of BCME
occurs in the presence of both hydrogen chloride and formaldehyde (Frankel,
et al. 1974). For additional information regarding the Chloroalkyl ethers
in general, the reader is referred to the EPA/ECAO Hazard Profile on Chloro-
alkyl Ethers (U.S. EPA,' 1979b).
II. EXPOSURE
As might be expected from the reactivity of BCME in water, monitoring
studies have not detected its presence in water. Human exposure by inhala-
tion appears to be confined to occupational settings (U.S. EPA, 1979a).
Data for human exposure to Chloroalkyl ethers by ingestion of food is
*
not available; nor is data relevant to human dermal exposure to chloralkyl
ethers (U.S. EPA, 1979a).
-------
The U.S. EPA (1979a) has estimated the- weighted average bioconcentra-
tion factor for 8CME to be 31 for the edible portions of fish and shellfish
consumed by Americans. This estimate is based on the octanol/water parti-
tion coefficient.
III. PHARMACOKINETICS
There is no specific information relating to the absorption, distribu-
tion, metabolism, or excretion of BCME (U.S. EPA, 1979a). Because of the
high reactivity and instability of BCME in aqueous systems, it is difficult
to generate pharmacokinetic parameters.
IV. EFFECTS
A. Carcinogenicity
8CME has been shown to produce tumors in several animal systems.
Inhalation exposure of male rats to BCME produced malignant respiratory
tract tumors (Kuschner, et al. 1975), while dermal application to mouse skin
led to the appearance of skin tumors (Van Duuren, et al. 1968). Administra-
tion of BCME to newborn mice by ingestion has been shown to increase the
incidence of hepatocellular carcinomas in males (Innes, et al. 1969).
Epidemiological studies of workers in the United States, Germany,
and Japan who were occupationally exposed to BCME and CMME have indicated
that these compounds are human respiratory carcinogens (U.S. EPA, 1979a).
BCME has been shown to accelerate the rate of lung tumor formation
in strain A mice following inhalation exposure (Leong, et al. 1971). BCME
has also shown activity as a tumor initiating agent for mouse skin (Slaga,
et al. 1973).
B. Mutagenicity
Testing of the chloroalkyl ethers in the Ames Salmonella assay and
in §_._ coli have indicated that BCME produced direct mutagenic effects (U.S.
EPA, 1979a).
t
-293-
-------
The results of a study on the incidence of cytogenetic aberrations
in the lymphocytes of workers exposed to BCME and CCME indicate higher fre-
quencies in this cohort. Follow-up indicates that removal of workers from
exposure led to a decrease in the frequency of aberrations (Zudova and
Landa, 1977).
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
Chronic occupational exposure to CMME contaminated with BCME has
produced bronchitis in workers (U.S. EPA, 1979a). Cigarette smoking has
been found to act synergistically with this type of exposure to produce
bronchitis (Weiss, 1976, 1977).
E. Other Relevant Information
The initiating activity of several chloroalkyl ethers indicates
that these compounds will interact with other agents to produce skin papil-
lomas (Van Duuren, et al. 1969, 1972).
V. AQUATIC TOXICITY
Pertinent information could not be found in the available literature
regarding aquatic toxicity for freshwater or marine species.
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
Based on animal carcinogenesis data, and using a linear, non-
threshold model, the U.S. EPA (1979a) has recommended a maximum permissible
-------
concentration of BCME for Ingested water at .02 ng/1. Assuming water is the
only source of exposure, compliance to this level should limit the risk car-
cinogenesis to not more than 10~ .
Based on animal studies, the 8-hour, time-weighted threshold limit
value (TLV-TWA) has been recommended for BCME as one ppb by the American
Conference of Governmental and Industrial Hygienists (1978).
B. Aquatic
Criterion for the protection of freshwater or marine aquatic organ-
isms were not drafted due to lack of toxicological'-evidence.
if
-.296--
-------
BIS(CHLOROMETHYL)ETHER
REFERENCES
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.
Frankel, L.S., et al. 1974. Formation of bis(chloromethyl)
ether from formaldehyde and hydrogen chloride. Environ.
Sci. Technol. 8: 356.
Innes, J.R.M., et al. 1969. Bioassay of pesticides and
industrial chemicals for tumorigenicity in'-mice: A prelimi-
nary note. Jour. Natl. Cancer Inst. 42; 1101.
Kuschner, M., et al. 1975. Inhalation carcinogenicity
of alpha halo ethers. III. Lifetime and limited period
inhalation studies with bis(chloromethyl)ether at 0.1 ppm.
Arch. Environ. Health 30: 73.
Leong, B.K.J., et al. 1971. Induction of lung adenomas
by chronic inhalation of bis(chloromethyl)ether. Arch.
Environ. Health 22: 663.
Slaga, T.J., et al. 1973. Macromolecular synthesis fol-
lowing a single application of alkylating agents used as
initiators of mouse skin tumorigenesis. Cancer Res. 33:
769.
Tour J.C., and G.J. Kallos. 1974. Kinetic study of the
stabilities of chloromethyl methyl ether and bis(chloromethyl)
ether in humid air. Anal. Chem. 46: 1866.
U.S. EPA. 1979a. Chloroalkyl Ethers: Ambient Water Quality
Criteria (Draft).
U.S. EPA. 1979b. Environmental Criteria and Assessment
Office. Hazard Profile: Chloroalkyl Ethers (Draft).
Van Duuren, B.L., et al. 1968. Alpha-haloethers: A new
type of alkylating carcinogen. Arch. Environ. Health 16:
472.
Van Duuren, B.L., et al. 1969. Carcinogenicity of halo-
ethers. Jour. Natl. Cancer Inst. 43: 481.
»
Van Duuren, B.L., et al. 1972. Carcinogenicity of halo-
ethers. II. Structure-activity relationships of analogs
of bis(chloromethyl)ether. Jour. Natl. Cancer Inst. 48:
1431.
-------
Weiss, W. 1976. Chloromethyl ethers, cigarettes, cough
and cancer. Jour. Occup. Med. 18: 194.
Weiss, W. 1977. The forced end-expiratory flow rate in
Chloromethyl ether workers. Jour. Occup. Med. 19: 611.
Zudova, Z., and K. Landa. 1977. Genetic risk of occupa-
tional exposures to haloethers. Mutat. Res. 46: 242.
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No. 27
Bis(2-ethylexyl)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 accuracy.
-------
BIS-(2-ETHYLHEXYL)PHTHALATE
SUMMARY
Bis-(2-ethylhexyl)phthalate has been shown to produce
mutagenic effects in the Ames Salmonella assay and in the
dominant lethal assay.
Teratogenic effects in rats were reported following
interperitoneal (i.p.) administration and oral administra-
tion of bis-{2-ethylhexyl)phthalate. Additional reproductive
effects produced by bis-(2-ethylhexyl)phthalate include
impaired implantation and parturition, and decreased fertility
in rats. Testicular damage and decreased spermatogenesis
have been reported in rats, following i.p. or oral adminis-
tration, and in mice, given bis-(2-ethylhexyl)phthalate
by oral intubation.
Evidence has not been found indicating that bis-(2-
ethylhexyl)phthalate has carcinogenic effects. Chronic
animal feeding studies of bis-(2-ethylhexyl)phthalate have
shown effects on the liver and kidneys.
Bis-(2-ethylhexyl)phthalate is acutely toxic to fresh-
water invertebrates at a concentration of 11,000 ug/1.
The same species has been shown to display severe reproduc-
tive impairment when exposed to concentrations less than
3 ug/1.
-3OO-
-------
BIS-(2-ETHYLHEXYL)PHTHALATE
I. INTRODUCTION
This profile is based on the Ambient Water Quality
Criteria Document for Phthalate Esters (U.S. EPA, 1979).
Bis-{2-ethylhexyl)phthalate, most commonly referred
to as di- (2-ethylhexyl)phthalate, (DEHP) is a diester of
the ortho form of benzene dicarboxylic acid. The compound
has a molecular weight of 391.0, specific gravity of 0.985,
boiling point of 386.9°C at 5 mm Hg, and is insoluble in
water {U.S. EPA, 1979) .
DEHP is widely used as a plasticizer, primarily in
the production of polyvinyl chloride (PVC) resins. As much
as 60 percent by weight of PVC materials may be plasticizer
(U.S. EPA, 1979). Through this usage, DEHP is incorporated
into such products as wire and cable covering, floor tiles,
swimming pool liners, upholstery, and seat covers, footwear,
and food and medical packaging materials (U.S international
Trade Commission, 1978).
In 1977, current production was 1.94 x 10 tons/year
(U.S. EPA, 1979).
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
bacterial flora may be capable of metabolizing phthalates
to the monoester form (Englehardt, et al. 1975).
-30J-
-------
II. EXPOSURE
Phthalate esters appear in all areas of the environ-
ment. Environmental release of the phthalates may occur
through leaching of plasticizers from PVC materials, vola-
tilization of phthalates from PVC materials, and the inciner-
ation of PVC items. Sources of human exposure to phthalates
include contaminated foods and fish, and parenteral adminis-.
tration by use of PVC blood bags, tubings, and infusion
devices {U.S. EPA, 1979).
Monitoring studies have indicated that phthalate concen-
trations in water are mostly in the ppm range, or 1-2 jag/liter
(U.S. EPA, 1979). Air levels of phthalates in closed rooms
that have PVC tiles have been reported to be 0.15 to 0.26
mg/m (Peakall, 1975). Industrial monitoring has measured
air levels of phthalates from 1.7 to 66 mg/m (Milkov, et
al. 1973). Levels of DEHP have ranged from not detect-
able to 68 ppm in foodstuffs (Tomita, et al. 1977). Cheese,
milk, fish and shellfish present potential sources of high
phthalate intake (U.S. EPA, 1979). Estimates of parenteral
exposure of patients to DEHP during use of PVC medical appli-
ances have indicated approximately 150 mg DEHP exposure
from a single hemodialysis course. An average of 33 mg
DEHP exposure is possible during open heart surgery (U.S.
EPA, 1979).
The U.S. EPA (1979) has estimated the weighted average
*
bioconcentration factor for DEHP to be 95 for the edible
portions of fish and shellfish consumed by Americans. This
-------
estimate is based on the measured steady-state bioconcentra-
tion studies in fathead minnow.
III. PHARMACOKINETICS
A. Absorption
The phthalates are readily absorbed from the intes-
tinal tract, the peritoneal cavity, and the lungs (U.S.
EPA, 1979). Daniel and Bratt (1974) found that seven days
following oral administration of radiolabelled DEHP, 42
percent of the dose was recovered in the urine and 57 per-
cent recovered in the feces of rats. Hilary excretion of
orally administered DEHP has been noted by Wallin, et al.
(1974). Limited human studies indicate that 2 to 4.5 per-
cent of orally administered DEHP was recovered in the urine
of volunteers within 24 hours (Shaffer, et al. 1945). Lake,
et al. (1975) have suggested that orally administered phtha-
lates are absorbed after metabolic conversion to the mono-
ester form in the gut.
Dermal absorption of DEHP 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 DEHP
have shown that 60 to 70 percent of the administered dose
was detected in the liver and lungs within 2 hours after
administration (Daniel and Bratt, 1974). Wadell, et al.
r
(1977) have reported rapid accumulation of labelled DEHP
in the kidney and liver of rats after i.v. injection, fol-
lowed by rapid excretion into the urine, bile, and intes-
ar
-303-
-------
tine. Seven days after i.v. administration of labelled
DEHP to mice, levels of 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 who had received large volumes of transfused blood,
detected DEHP in the spleen, liver, lungs, and abdominal
fat (Jaeger and Rubin, 1970).
Injection of pregnant rats with labelled DEHP
has shown that the compound may cross the placental barrier
(Singh, et al. 1975).
C. Metabolism
Various metabolites of DEHP have been identified
following oral feeding to rats (Albro, et al. 1973). These
results indicate that DEHP is initially converted from the
diester to the monoester, followed by the oxidation of the
monoester side chain forming two different alcohols. The
alcohols are oxidized to the corresponding carboxylic acid
or ketone. Enzymatic cleavage of DEHP to the monoester
may take place in the liver or the gut (Lake, et al. 1977).
This enzymatic conversion has been observed in stored whole
blood indicating widespread distribution of metabolic activ-
ity (Rock, et al. 1978).
D. Excretion
Excretion of orally administered DEHP is virtually
complete in the rat within 4 days (Lake, et al. 1975).
iMajor excretion is through the urine and feces, with biliary
-------
excretion increasing the content of DEHP (or metabolites)
in the intestine (U.S. EPA, 1979). Schulz .and Rubin (1973)
have noted an increase in total water soluble metabolites
of labelled DEHP in the first 24 hours following injection
into rats. Within one hour, eight percent of the DEHP was
found in the liver, intestines and urine. After 24 hours,
54.6 percent was recovered in the intestinal tract, excreted
feces and urine, and only 20.5 percent was recovered in or-
ganic extractable form. Blood loss of DEHP showed a biphasic
pattern, with half-lives of 9 minutes and 22 minutes, respec-
tively (Schulz and Rubin, 1973).
IV. EFFECTS
A. Carcinogenicity
Pertinent data could not be located in the avail-
able literature.
B. Mutagenicity
Testing of DEHP in the Ames Salmonella assay has
shown no mutagenic effects (Rubin, et al. 1979). Yagi,
et al. (1978) have indicated that DEHP is not mutagenic
in a recombinant strain of Bacillus, but the monoester meta-
bolite of DEHP did show some mutagenic effects. Results
of a dominant lethal assay in mice indicate that DEHP has
a dose and time dependent mutagenic effect (Singh, et al.
1974).
C. Teratogenicity
DEHP has been shown to produce teratogenic effects
in rats following i.p. administration (Singh, et al. 1972).
-------
Following oral administration there was a significant reduc-
tion in fetus weight at 0.34 and 1.70 g/kg/day.
D. Other Reproductive Effects
Effects on implantation and parturition have been
observed in pregnant rats injected intraperitoneally with
DEHP (Peters and Cook, 1973). A three-generation repro-
duction study in rats has indicated decreased fertility
in rats following maternal treatment with DEHP (Industrial
Bio-Test, 1978).
Testicular damage has been reported in rats ad-
ministered DEHP i.p. or orally. Seth, et al. (1976) found
degeneration of the seminiferous tubules and changes in
spermatagonia; testicular atrophy and morphological damage
were noted in rats fed DEHP (Gray, et al. 1977; Yamada,
et al, 1975). Otake, et al. (1977) noted decreased sperma-
togenesis in mice administered DEHP by intubation.
E. Chronic Toxicity
Oral feeding of DEHP produced increases in liver
and kidney weight in several animal studies (U.S. EPA, 1979).
Chronic exposure to transfused blood containing DEHP has
produced liver damage in monkeys (Kevy, et al. 1978). Lake,
et al. (1975) have produced liver damage in rats by adminis-
tration of mono-2-ethylhexyl phthalate.
' F. Other Relevant Information
Several animal studies have demonstrated that
pre-treatment of rats with DEHP produced an increase in
hexobarbital sleeping times (Daniel and Bratt, 1974; Rubin
and Jaeger, 1973; Swinyard, et al. 1976).
-------
V. AQUATIC TOXICITY
A. Acute Toxicity
Only one acute study on the freshwater cladoceran
(Daphnija magna) has produced a 96-hour static LC5Q value
of 11,000 jag/1 (U.S. EPA, 1978). Freshwater fish or marine
data have not been found in the literature.
B. Chronic Toxicity
Chronic studies involving the rainbow trout (Salmo
gairdneri) provided a chronic value of 4.2 pg/1 in an embryo-
larval assay (Mehrle and Mayer, 1976). Severe reproductive
impairment was observed at less than 3 pg/1 in a chronic
Daphnia magna assay (Mayer and Sanders, 1973).
C. Plant Effects
Pertinent information could not be located in
the available literature.
D. Residues
Bioconcentration factors have been obtained for
several species of freshwater organisms: 54 to 2,680 for
the scud (Gamarus pseudolimnaeus); 14 to 50 for the sowbug
(Asceilus brevicaudus); 42 to 113 for the rainbow trout
(Salmo gairdneri); and 91 to 886 for the fathead minnow
(Pimephales promelas) (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.
-3,07-
-------
A. Human
Based on "no effect" levels observed in chronic
feeding studies in rats or dogs, the U.S. EPA has calculated
an acceptable daily intake (ADI) level for DEHP of 42 ing/day,
The recommended water quality criteria level for
protection of human health is 10 mg/1 for DEHP (U.S. EPA,
1979).
B. Aquatic
Criterion was not drafted for either freshwater
or marine environments due to insufficient data.
-------
BIS-{2-ETHYLHEXYL) PHTHALATE
REFERENCES
Albro, P.W., et al. 1973. Metabolism of diethylhexyl phthal-
ate by rats. Isolation and characterization of the urinary
metabolites. Jour. Chromatogr. 76: 321.
Autian, J. 1973. Toxicity and health threats of phthalate
esters: Review of the literature. Environ. Health Perspect.
June 3.
Daniel, J.W., and H. Bratt. 1974. The absorption, metabo-
lism and tissue distribution of di(2-ethylhexyl) phthalate
in rats. Toxicology 2: 51.
Engelhardt, G., et al. 1975. The microbial metabolism
of di-n-butyl phthalate and related dialkyl phthalates.
Bull. Environ. Contain. Toxicol. 13: 342.
Gray, J., et al. 1977. Short-term toxicity study of di-
2-ethylhexyl phthalate in rats. Food Cosmet. Toxicol. 65:
389.
Industrial Bio-Test. 1978. Three generation reproduction
study with di-2-ethylhexyl phthalate in albino rats. Plastic
Industry News 24: 201.
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. NO1 HB 5-2906, Natl. Heart,
Lung and Blood Inst. Bethesda, Md.
Lake, B.C., et al. 1975. Studies on the hepatic effects
of orally administered di-(2-ethylhexyl) phthalate in the
rat. Toxicol. Appl. Pharmacol. 32: 355.
Lake, E.G., et al. 1977. The in vitro hydrolysis of some
phthalate diesters by hepatic and* intestinal preparations
from various species. Toxicol. Appl. Pharmacol. 39: 239.
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-ethylhexyl phtha-1-
ate: 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.
Otake, T., et al. 1977. The effect of di-2-ethylhexyl
phthalate (DEHP) on male mice. I. Osaka-Fuitsu Koshu Eisei
Kenkyusho Kenkyu Hokoku, Koshu Eisei Hen 15: 129.
Peakall, D.B. 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
plasma. 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 Enviorn-
mental biology. Interprint Publications, New DehlTT India.
Shaffer, C.B., et al. 1945. Acute and subacute toxicity
of di(2-ethylhexyl) phthalate with note upon its metabolism.
Jour. Ind. Hyg. Toxicol. 27: 130.
Singh, A.R., et al. 1972. Teratogenicity of phthalate esters
in rats. Jour. Pharmacol. Sci. 61: 51.
Singh, A.R., et al. 1974. Mutagenic and antifertility
sensitivities of mice to di-2-ethylhexyl phthalate (DEHP)
and dimethoxyethyl phthalate (DMEP). Toxicol. Appl. Pharmacol,
29: 35.
Singh A.R., et al. 1975. Maternal-fetal transfer of 14C-
di-2-ethylhexyl phthalate and x C-diethyl phthalate in rats.
Jour. Pharm. Sci. 64: 1347.
-------
Swinyard, E.A., et al. 1976. Nonspecific effect of bis (2-
ethylhexyl) phthalate on hexobarbital sleep time. Jour.
Pharmacol. Sci. 65: 733.
Tomita, I., et al. 1977. Phthalic acid esters in various
foodstuffs and 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. 68-01-4646.
U.S. EPA. 1979. Phthalate Esters: Ambient Water Quality
Criteria (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: 339.
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.
Yagi, Y., et al. 1978. Embryotoxicity of phthalate esters
in mouse. Proceedings of the First International Congress
on Toxicology, Plaa, G. and Duncan, W. , eds. Academic Press,
N.Y. p. 59.
Yamada, A., et al. 1975. Subacute toxicity of di-2-ethyl-
hexyl phthalate. Trans. Food Hyg. Soc. Japan, 29th Meeting
p. 36.
-2JJ-
-------
No. 28
Broraofora
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-uracy.
-------
BROMOFORM
SUMMARY
Bromoform has been detected in finished drinking water in
the United States and Canada. It is believed to be formed by the
haloform reaction that may occur during water chlorination.
Broraoform can be removed from drinking water via treatment with
activated carbon. Natural sources (especially red algae) produce
significant quantities of bromoform. There is a potential for
broraoform to accumulate in the aquatic environment because of its
resistance to degradation. Volatilization is likely to be an
important means of environmental transport.
Bromoform gave positive results in mutagenicity tests with
Salmonella typhimurium TA100. In a short-term in vivo oncogen-
icity assay it caused a significant increase in tumor incidence
at one dose level.
Inhalation of bromoform by humans can cause irritation of
the respiratory tract and liver damage. Respiratory failure is
the primary cause of death in bromoform-related fatalities.
I. INTRODUCTION
This profile is based primarily on the Ambient Water Quality
Criteria document for halomethanes (U.S. EPA 1979b).
Bromoform (tribromomethane; CHBr3) is a colorless, heavy
p
liquid similar in odor and taste to chloroform. Bromoform has
the following physical/chemical properties (Weast, 1974):
-------
Molecular Weight; 252.75
Melting Point: 8.3'C
Boiling Point: 149.5'C (at 760 mm Hg)
Vapor Pressure: 10 mm Hg al 34'C
Solubility: slightly soluble in water;
soluble in a variety of
organic solvents.
A review of the production range (includes importation)
statistics for bromoform (CAS No. 75-25-2} which is listed in the
initial TSCA Inventory (1979a) has shown that between 100,000 and
900,000 pounds of this chemical were produced/imported in 1977.—/
Bromoform is used as a chemical intermediate; solvent for
waxes, greases, and oils; ingredient in fire-resistant chemicals
and gauge fluids (U.S. EPA 1978a; Hawley, 1977).
II. EXPOSURE
A. Environmental Fate
Bromoform gradually decomposes on standing; air and light
accelerate decomposition (Windholz, 1976). The vapor pressure of
bromoform, while lower than that for chloroform and other chloro-
alkanes, is, nonetheless, sufficient to ensure that volatiliza-
tion will be an important means of environmental transport. The
JV 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 CPR 710).
- 3/S--
-------
half-life for hydrolysis of bromoform is estimated at 686 years.
Bromoform should be much more reactive in the atmosphere. Oxi-
dation by HO radical will result in a half-life of a few months
in the troposphere (U.S. EPA, 1977).
B. Bioconcentration
The bioconcentration factor for bromoform in aquatic organ-
isms that contain about 8% lipid is estimated to be 48. The
weighted average bioconcentration factor for bromoform in the
edible portion of all aquatic organisms consumed by Americans is
estimated to be 14 (U.S. EPA, 1979b).
C. Environmental Occurence
The National Organics Reconnaissance Survey detected bromo-
forra in the finished drinking water of 26 of 80 cities, with a
maximum concentration of 92 ug/1. Over 90% of the samples con-
tained 5 ug/1 or less. Ho bromoform was found in raw water
samples (Symons et_ al_. , 1975). Similarly, the EPA Region V
Organics Survey found bromoform in 14% of the finished drinking
water samples and none in raw water (U.S. EPA, 1975). Using a
variety of sampling and analysis methods, the National Organic
Monitoring Survey found bromoform in 3 of 111, 6 of 118, 38 of
113, 19 of 106, and 30 of 105 samples with mean concentrations
ranging from 12-28 ug/1 (U.S. EPA, 1978b). A Canadian survey of
drinking water found 0-0.2 ug/1 with a median concentration of
0.01 ug/1 (Health and Welfare Can., 1977).
m
The National Academy of Sciences (1978) concluded that water
chlorination, via the haloform reaction, results in the produc-
tion of trihalomethanes (including bromoform) from the organic
precursors present in raw water.
-------
Significant quantities of bromoform are also produced from
natural sources, especially red algae. For example, the essen-
tial oil of Asparagopsis taxiformis (a red marine algae eaten by
Hawaiians) contains approximately 80% bromoform (Burreson et al. ,
1975).
III. PHARMACOKINETICS
Broraoform is absorbed through the lungs, gastrointestinal
tract, and skin. Some of the absorbed bromoform is metabolized
in the liver to inorganic bromide ion. Bromide is found in
tissues and urine following inhalation or rectal administration
of bromoform (Lucas, 1929). Metabolism of bromoform to carbon
monoxide has also been reported (Ahmed, 1977). Recent studies
show that phenobarbital-induced rats metabolize bromoform to. A
(cocU
carbonyl bromide (COB^)/ the brominated analog of phosgene fPctrT
_et_ a±., 1979).
IV. HEALTH EFFECTS
A. Carcinogenicity
Bromoform caused a significant increase in tumor incidence
at one dose level in a short-term in vivo oncogenicity assay
known as the strain A mouse lung adenoma test. The increase was
observed at a dose of 48 mg/kg/injection with a total dose of
1100 mg/kg. The tumor incidence was not increased significantly
at doses of 4 mg/kg (total dose of 72 mg/kg) or 100 mg/kg (total
dose of 2400 mg/kg) (Theiss ^et_ _al_. , 1977.
-------
B. Mutagenicity
Bromoform was mutagenic in S. typhimurium strain TA 100
(without metabolic activation) (Simmon, 1977).
C. Other Toxicity
Rats inhaling 250 mg/rn^ bromoform for 4 hr/day for 2 months
developed impaired liver and kidney function (Dykan, 1962).
In humans, inhalation of bromoform causes irritation to the
respiratory tract. Mild cases of bromoform poisoning may cause
only headache, listlessness, and vertigo. Unconsciousness, loss
of reflexes, and convulsions occur in severe cases. The primary
cause of death from a lethal dose of bromoform is respiratory
failure. Pathology indicates that the chemical causes fatty
degenerative and centrolobular necrotic changes in the liver
(U.S. PHS, 1955).
Acute animal studies indicate impaired function and
pathological changes in the liver and kidneys of animals exposed
to bromoform (Kutob and Plaa, 1962; Dykan, 1962).
V. AQUATIC EFFECTS
A. Fresh Water Organisms
The 96-hr LC5Q (static) in bluegill sunfish is 29.3 mg/1.
The 48-hr LC5Q (static) for Daphnia magna is 46.5 mg/1. The 96-
hr ECcnS for chlorophyll A production and cell number in S^_
capricornutum are 112 mg/1 and 116 mg/1, respectively (U.S. EPA,
*
1978a). (See also Section II.B.)
-------
B. Marine Organisms
The 96-hr LCf-Q (static) in sheepshead minnow is 17.9 rag/1.
The 96-hr LC5Q (static) in raysid shrimp is 20.7 mg/1. The EC5Qs
for chlorophyll A production and cell number in S. costatum are,
respectively, 12.3 mg/1 and 11.5 mg/1 (U.S. EPA, 1978a).
VI. EXISTING GUIDELINES
A. Human
The OSHA standard for bromoform in air is a time weighted
average (TWA) of 0,5 ppm (39CFR23540).
The Maximum Contaminant Level (MCL) for total trihalometh-
anes (including bromoform) in drinking water has been set by the
U.S. EPA at 100 ug/1 (44FR68624). The concentration of bromoform
produced by chlorination can be reduced by treatment of drinking
water with powdered activated carbon (Rook, 1974). This is the
technology that has been proposed by the EPA to meet this
standard.
B. Aquatic
The proposed ambient water criterion for the protection of
fresh water aquatic life from excessive bromoform exposure is 840
ug/1 as a 24-hour average. Bromoform levels are not to exceed
1900 ug/1 at any time. The criterion for the protection of
marine life is 180 ug/1 (24 hr avg), not to exceed 1900 ug/1
(U.S. EPA, 1979b).
-------
REFERENCES
Ahmed, A.E., _et_ _al_. 1977. Metabolism of haloforms to carbon
monoxide/ I. In vitro studies. Drug Metab. Dispos., _S:198. {as
cited in U.S. EPA, 1979b).
Burreson, B.J., R.E. Moore, P.P. Roller 1975. Haloforms in the
essential oil of the alga Asparagopsis taxiformis (Rhodophyta).
Tetrahedron Letters, _7_:473-476. {as cited in NAS, 1978).
Dykan, V.A. 1962. Changes in liver and kidney functions due to
methylene bromide and bromoform. Nauchn. Trucy Ukr Nauchn. -
Issled. Inst. Gigieny Truda i Profyabolevanii J29_:82. (as cited
in U.S. EPA, 1979b).
Hawley, G.G. ed. 1971. Condensed Chemical Dictionary. 8th ed.
Van Nostrand Reinhold Co.
Health and Welfare Canada 1977. Environmental Health Direc-
torate national survey of halomethane in drinking water. (as
cited in U.S. EPA, 1979b).
Kutob, S.D., G.J. Plaa 1962. A procedure for estimating the
hepatotoxic potential of certain industrial solvents, Tox. Appl.
Pharm., _4_:354. (as cited in U.S. EPA, 1979b) .
Lucas, G.H.W. 1929. A study of the fate and toxicity of bromine
and chlorine containing anesthetics, J. Pharm. Exp. Therap.,
_3£:223-237. (as cited in NAS, 1978).
National Academy of Sciences 1977. Drinking Water and Health,
Part II, Chapters 6 and 7, Washington, D.C.
National Academy of Sciences 1978. Nonfluorinated Halomethanes
in the Environment, Washington, D.C.
Pohl, L.R. ^t_ _al_. 1979. Oxidative bioactivation of haloforms
into hepatotoxins, prepublication.
Rook, J.J. 1974. Formation of haloforms during chlorination of
natural waters. J. Soc. Water Treat. Examin. 23 (Part 2}:234-
243.
Simmon, V.F. 1977. Mutagenic activity of chemicals identified
in drinking water. In Progress in genetic toxicology, S. Scott
et_ a±. eds. (as cited in U.S. EPA, 1979b).
Symons, J.M _et_ a±. 1975. National organics reconnaissance *
survey for halogenated organics (NORS). J. Amer. Water Works
Assoc. _6_7_:634-647. (as cited in MAS, 1978).
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, Can. Res., _3J7_:2717. {as cited in U.S.
EPA, 1979b).
7f
-320-
-------
U.S. EPA 1975. Formation of Halogenated Organics by Chlorina-
tion of Water Supplies. EPA-600/1-75-002, PB 241-511. (as cited
in NAS, 1978).
U.S. EPA 1977. Review of the environmental fate of selected
chemicals, EPA-560/5-77-0033.
U.S. EPA 1978a. Indepth studies on health and environmental
impacts of selected water pollutants, contract no. 68-01-4646,
Washington, D.C. (as cited in U.S. EPA, 1979b).
U.S. EPA 1978b. The National Organic Monitoring Survey, Office
of Water Supply, Washington, D.C.
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. Halomethanes, Ambient Water Quality Criteria.
PB 296 797.
U.S. Public Health Service 1955. The halogenated hydrocarbons:
Toxity and potential dangers. Ho. 414. (as cited in U.S. EPA,
1979b).
Weast, R.C. ed. 1972. CRC Handbook of Chemistry and Physics.
CRC Press, Inc., Cleveland, Ohio.
Windholz, M. ed. 1976. The Merck Index, 9th ed., Merck and Co.,
Inc., Rahway, N.J.
-------
No. 29
Bromoraethane
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.
-------
BROMOMETHANE
Summary
On acute exposure to bromomethane, neurologic and psychiatric
-abnormalities may develop and persist for months or years. There is
no information on the chronic toxicity, carcinogenicity, or terato-
genicity of bromomethane. Bromomethane has been shown to be mutagenic
in the Ames S_._ typhimurium test system.
Acute LC5Q values have been reported in two tests as 12,000 and
11,000 ps/1 for.a marine and freshwater fish, respectively.
-------
BROMOMETHANE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria
Document for Halomethanes (U.S. EPA, 1979&).
Bromomethane (CH^Br, methyl bromide, monobromomethane, and
embafume; molecular weight 9^-94) is a colorless gas. Bromomethane
has a melting point of -93.6°C, a boiling point of 3.56°C, a specific
gravity of 1.676 g/ml at -20°C, and a water solubility of 17-5 g/1
at 20°C (Natl. Acad. Sci., 1978). Bromomethane has been widely used
as a fumigant, fire extinguisher, refrigerant, and insecticide (Kantarjian
and Shaheen, 1963). Today the major use of bromomethane is as a
fumigating agent. Broraomethane is believed to be formed in nature,
with the oceans as a primary source (Lovelock, 1975). The other
major environmental source of bromomethane is from its agricultural
use as a soil, seed, feed and space fumigant. For additional information
regarding 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 bromomethane qualitatively
in finished drinking waters in the U.S. There are, however, no data
on its concentration in drinking water, raw water, or waste water
(U.S. EPA, 1979a).
B. Food
There is no information on the concentration of bromomethane
in food. Bromomethane residues from fumigation decrease rapidly
through loss to the atmosphere and reaction with protein to form
-------
inorganic bromide residues. With proper aeration and product processing,
most residual bromomethane will rapidly disappear due to raethylation
reactions and volatilization (Natl. Acad. Sci., 1978; Davis, et al.
1977). There are no bioconcentration data for bromomethane (U.S.
EPA, 1979a).
C. Inhalation
Saltwater atmospheric background concentrations of broraomethane
averaging about 0.00036 mg/m3 have been reported (Grimsrud and Rasmussen,
1975; Singh, et al. 1977). This is higher than reported average
continental background and urban levels and suggests that the "oceans
are a major source of global bromomethane (Natl. Acad. Sci., 1978).
Bromomethane concentrations of up to 0.00085 mg/m3 may occur outdoors
locally with light traffic, as a result of exhaust containing bromomethane
as a breakdown product of ethylene dibromide, which is used in leaded
gasoline (Natl. Acad. Sci., 1978).
III. PHARMACOKINETICS
A. Absorption
Absorption of bromomethane most commonly occurs via the
lungs, although it can also occur through the gastrointestinal tract
and the skin (Davis, et al. 1977; von Oettingen, 1964).
B. Distribution
Upon absorption, blood levels of residual non-volatile
bromide increase, indicating rapid uptake of bromomethane or its
metabolites (Miller and Haggard, 19^3). Bromomethane is rapidly
distributed to various tissues and is broken down to inorganic bromide-.
Storage, only as bromides, occurs mainly in lipid-rich tissues.
-------
C. Metabolism
Evidently the toxicity of bromomethane is mediated by the
bromomethane molecule itself. Its reaction with tissue (methylation
of sulfhydryl groups in critical cellular proteins and enzymes)
results in disturbance of intracellular metabolic functions, with
irritative, irreversible, or paralytic consequences (Natl. Acad.
Sci., 1978; Davis, et al. 1977; Miller and Haggard, 1913).
D. Excretion
Elimination of bromomethane is rapid initially, largely
through the lungs. The kidneys eliminate much of the remainder as
bromide in the urine (Natl. Acad. Sci., 1978).
IV. EFFECTS
Pertinent information relative to the carcinogenicity, teratogenicity
or other reproductive effects, or chronic toxcity of bromomethane
were not found in the available literature.
A. Mutagenicity • •
Simmon and coworkers (1977) reported that bromomethane was
mutagenic to Salmonella typhimurium strain TA100 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.
B. Other Relevant Information
In several species, acute fatal poisoning has involved
marked central nervous system disturbances with a variety of manifestations:
ataxia, twitching, convulsions, coma, as well as changes in lung, liver,
-------
heart, and kidney tissues (Sayer, et al. 1930; Irish, et al. 1940;
Gorbachev, et al. 1962; von Oettingen, 1964). Also, residual bromide
in fumigated food has produced some adverse effects in dogs (Hosenblum,
et al. 1960).
V. AQUATIC TOXICITY
Two acute toxicity studies on one freshwater and one marine
fish species were reported with LC5Q values of 11,000 ug/1 for freshwater
bluegill (Lepomis macrochirus) and an LCgo value of 12,000 jig/1 for
the marine tidewater silversides (Menidia beryllina) (U.S. EPA,
1979a). Pertinent information relative to aquatic chronic toxicity
or plant effects for bromomethane were not found 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
The current OSHA standard for occupational exposure to
bromomethane (1976) is 80 mg/m3; the American Conference of Governmental
Industrial Hygienist's (ACGIH, 1971) threshold limit value is 78
mg/m3. The U.S. EPA (1979a) draft water quality criteria for bromomethane
is 2 pg/1. Refer to the Halomethane Hazard Profile for discussion
of criteria derivation (U.S. EPA, 1979b).
-3.2%-
-------
B. Aquatic Toxicity
The draft criterion for protecting freshwater life is a
24-hour average concentration of 1^0 jig/1, not to exceed 320 >ig/l.
The marine criterion is 170 ;ig/l as a 24-hour average, not to exceed
380 pg/l.
-------
BROMOMETHANE
References
American Conference of Governmental and Industrial Hygienists. 1971.
Documentation of the threshold limit values for substances in workroom
air. Cincinnati-, Ohio.
Davis, L.N., et al. 1977. Investigation of selected potential 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.
Gorbachev, E.M. , et al. 1962. Disturbances in neuroendocrine regulation
and oxidation-reduction by certain commercial poisons. Plenuma Patofiziol
Sibiri i Dal'n. Vost. Sb. 88.
Grirasrud, E.P., and R.A. Rasmussen. 1975- Survey and analysis of halocarbons
in the atmosphere by gas chroraatography-mass spectrometry. Atmos. Environ. 9:'
1014.
Irish, D.D., et al. 1940. The response attending exposure of laboratory
animals to vapors of methyl bromide. Jour. Ind. Hyg. Toxicol. 22: 218.
Kantarjian, A.D. , and A.S. Shaheen. 1963. Methyl bromide poisoning with nervous
system manifestations resembling polyneuropathy. Neurology 13: 1054.
Lovelock, J.E, 1975. Natural halocarbons in the air and in the sea. Nature
256: 193-
Miller, D.P., and H.W. Haggard. 1943- Intracellular penetration of bromide as
feature in toxicity of alkyl bromides. Jour. Ind. Hyg. Toxicol. 25: 423.
National Academy of Sciences. 1978. Nonfluorinated halomethanes in the
enviornment. Washington, D.C.
Occupational Safety and Health Administration. 1976. General industry standards.
OSHA 2206, revised January 1976. ' U.S. Dep. Labor, Washington., D.C.
Rosenblum, I., et al. 1960. Chronic ingestion by dogs of methyl bromide-
fumigated foods. Arch. Enviorn. Health 1: 3 '6.
Sayer, R.R., et al. 1930. Toxicity of dichlorodiflourome thane. U.S Bur. Mines
Rep. R.I. 3013.
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, *
and other atmospheric constituents. Atmos. Environ. 11: 819.
U.S. EPA. 1975. Preliminary assessment of suspected carcinogens in drinking
water, and appendicies. A report to Congress, Washington, D. C.
U.S. EPA. 1979a. Halomethanes: Ambient Water Quality Criteria. (Draft).
U.S. EPA, 1979b. Environmental Criteria and Assessment Office. Halomethanes:
Hazard Profile.
- vso-
-------
von Oettingen, W.F. 1964. The halogenated hydrocarbons of industrial and
toxicological importance. Elsevier Publ. Co., Amsterdam.
-------
No. 30
4-Bromophenyl Phenyl Ether
Health and Environmental Effects
G.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, B.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-uracy.
-3-3,3-
-------
4-Bromophenyl phenyl ether
SUMMARY
Very little information on 4-bromophenyl phenyl ether exists. 4-Bromophenyl
phenyl ether has been identified in raw water, in drinking water and in river
water. 4-Bromophenyl phenyl ether has been tested in the pulmonary adenoma
assay, a short-term carcinogenicity assay. Although the results were negative,
several known carcinogens also gave negative results. No other health effects
were available. 4-Bromophenyl phenyl ether appears to be relatively toxic
to freshwater aquatic life: a 24-hour average criterion of 6.2 ug/L has been
proposed.
I. INTRODUCTION
4-Bromophenyl phenyl ether (BrC,H,OC,HC; molecular weight 249.11) is a
O 4 03
liquid at room temperature; it has the following physical/chemical properties
(Weast 1972):
Melting point: 18.72°C
Boiling point: 310.14°C (760 mm Hg)
163°C (10 mm Hg)
Density: 1.420820
Solubility: Insoluble in water; soluble in, ether
No information could be found on the uses of this substance.
A review of the production range (includes Importation) statistics
for 4-bromophenyl phenyl ether (CAS Nol 101-55-3) which is listed in the initial
TSCA Inventory (1979) has shown that between 0 and 900 pounds of this chemical
were produced/imported in 1977.*
*
'This production range information does not include any product ion/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).
-------
II. EXPOSURE
No specific Information relevant to the environmental fate of 4-bromophenyl
phenyl ether was found in the literature. A U.S. EPA report (1975a) included this
substance in a category with several other drinking water contaminants consid-
ered to be refractory to biodegradation (i.e., lifetime greater than two years
in unadapted soil; point sources unable to be treated biologically). However,
the authors did not present or reference experimental data to support the inclu-
sion of 4-bromopheny phenyl ether in this category. U.S. EPA (1975a) estimated
that three tons of 4-bromophenyl phenyl ether are discharged annually.
4-Bromophenyl phenyl ether has been identified as a contaminant in finished
drinking water on three occasions, in raw water on one occasion and in river
water on one occasion. No quantitative data were supplied (U.S. EPA, 1976). Fri-
loux (1971) and U.S. EPA (1972) have also reported the presence of 4-bromophenyl
phenyl ether in raw and finished water of the lower Mississippi River (New
Orleans area). Again, no quantitative data were supplied. U.S. EPA (1975) sugrr
gest that 4-bromophenyl phenyl ether may be formed during the chlorination of
treated sewage and drinking water.
III. PHAKMACOKINETICS
No information was located.
IV. HEALTH EFFECTS
A. Cajrc in o genie ity
Three groups of 20 male mice were administered intxaperitoneal doses
(23, 17 or 18 doses, respectively) of 4-bromophenyl phenyl ether in tricaprylin
vehicle three times a week for 8 weeks (Theiss et al. 1977). The total doses
were 920, 1700, or 3600 rag/kg, respectively. Animals were sacrificed at 24
weeks from the start of the experiment. Incidences of lung adenomas were not
significantly increased, as compared with vehicle controls. However, this short-
term assay should not be considered indicative of the nononcogenieity of 4-
bromophenyl phenyl ether as several known oncogens tested negative in this assay.
-------
V. AQUATIC TOXICITY
A. Acute
An unadjusted 96 hour LC of 4,940 ug/L was determined by exposing
"bluegills to 4-bromophenyl phenyl ether (Table 1). Adjusting this value for test
conditions and species sensitivity, a" Final Fish Acute Value of 690 ug/L is obtained
(U.S. EPA, undated).
Exposure of Daphnia magna, yielded an unadjusted 48 hour LC-., of 360 ug/L
(Table 2). The Final Invertebrate Acute Value (and the Final Acute Value) for
4-bromophenyl phenyl ether is 14 ug/L (U.S. EPA, undated).
B. Chronic
In an embryo-larval test using the fathead minnow (in which survival and
growth were observed), a chronic value of 61 ug/L was obtained for 4-bromophenyl
phenyl ether exposure (Table 3). Dividing by the species sensitivity factor
(6.7), a Final Fish Chronic Value of 9.1 ug/L is derived. Since no other
information is available, this value is also the Final Chronic Value (U.S. EPA,
undated).
VI. EXISTING GUIDELINES
A. Aquatic
A 24 hour average concentration of 6.2 ug/L (6.2 ug/L = 0.44 x 14 ug/L
(Final Acute Value)) is the recommended criterion to protect freshwater aquatic
life. The maximum allowable concentration should not exceed 14 ug/L at any
time (U.S. EPA, undated).
-------
Table 1. Freshwater fish acute values
Organ ism
Bluegill,
Lepomis macrochirus
Bioassay Test
Method* Cone
S U
Chemical
.** Description
4 -Bromophenyl-
phenyl ether
Time
(hrs)
96
LC50
(ug/L.)
4,940
Adjusted
LC50
(ug/L)
2,700
* S = static
** U = unmeasured
Geometric mean of adjusted values: 4-Bromophenylphenyl ether = 2,700 ug/L
- 690 ug/L
Table 2. Freshwater invertebrate acute values
Organism
Cladoceran,
Daphnia magna
Bioassay Test
Method* Cone.**
S U
Chemical
Description
4-Bromophenyl-
phenyl ether
Time
(hrs)
48
LC50
(ug/L)
360
Adjusted
LC50
(ug/L)
300
* S = static
** U = unmeasured
Geometric mean of adjusted values:
4-Bromophenyl phenyl ether = 300 ug/L
300
21
14 ug/L
Table 3. Freshwater fish chronic values, 4-Bromophenyl phenyl ether
Organism
Fathead minnow,
Pirnephales promelas
Limits
Test* (ug/L)
E-L
89-167
Chronic
Value
(ug/L)
61
* E-L = embryo-larva
Geometric mean of chronic values = 61 ug/L T—-,
Lowest chronic value = 61 ug/L
-•3,37-
-------
BIBLIOGRAPHY
Friloux J. 1971. Petrochemical wastes as a pollution problem in the lower
Mississippi River. Paper submitted to the Senate Subcommittee on Air and Water
Pollution, April 5 (as cited in U.S. EPA, 1975b).
Theiss JC, Stoner GD, Shimkin MB, Weisburger EK. 1977. Test for carcinogenicity
of organic contaminants of United States drinking waters by pulmonary tumor
response in strain A mice. Cancer Research 37:2717-2720.
U.S. EPA. 1972 Industrial pollution of the Lower Mississippi River in
Louisiana Region VI. Surveillance and Analysis Division (as cited in U.S.
EPA, 1975b) .
U.S. EPA. 1975a. Identification of organic compounds in effluents from industrial
sources. EPA-560/3-75-002, PB 241 641.
U.S. EPA. 1975b. Investigation of selected potential environmental contaminants:
Haloethers. EPA-560/2-75-006.
U.S. EPA. 1976. Frequency of organic compounds identified in water. EPA-
600/4-76-062.
U.S. EPA. 1979. Toxic Substances Control Act Chemical Substance Inventory.
Production Statistics for Chemicals on the Non-Confidential Initial TSCA Inventory.
U.S. EPA. (undated). Ambient Water Quality Criteria Document on Haloethers,
Criteria and Standards Division, Office of Water Planning and Management. PB
296-796.
Weast, RC (ed.). 1972, Handbook of Chemistry and Physics, 53rd. ed. The
Chemical Rubber Co., Cleveland, OH.
-------
No. 31
Cadmium
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.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated
cadmium and has found sufficient evidence to indicate that
this compound is carcinogenic.
- V-M -
-------
CADMIUM
Summary
The major non-occupational routes of human cadmium exposure are through
food and tobacco smoke. Drinking water also contributes relatively little
to the average daily intake.
Epidemiological studies indicate that cadmium exposure may increase the
mortality level for cancer of the prostate. Long-term feeding and inhal-
ation studies in animals have not produced tumors, while intravenous admin-
istration of cadmium has produced only injection site tumors. Mutagenic
effects of cadmium exposure have been seen in animal studies, bacterial sys-
tems, in vitro tests, and .in the chromosomes of occupationally exposed
workers.
Cadmium has produced teratogenic effects in several species of animals,
possibly through interference with zinc metabolism. Testicular necrosis and
neurobehavioral alterations in animals following exposure during pregnancy
have been produced by cadmium in animals.
Chronic exposure to cadmium has produced emphysema and a characteristic
syndrome (Itai-Itai disease) following renal damage and osteomalacia. A
causal relationship between chronic cadmium exposure and hypertension in
humans has been suggested but not confirmed.
Cadmium is acutely toxic to freshwater fish at levels as low as 0.55
jug/1. Freshwater fish embryo/larval stages tended to be the most sensitive
to cadmium. Marine fish were generally more resistant than freshwater
fish. The long half-life of cadmium in aquatic organisms has been postu-
r
lated, and severe restrictions to gill-tissue respiration have been observed
at concentrations as low as 0.5jug/l.
-------
CADMIUM
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Cadmium (U.S. EPA, 1979).
Cadmium is a soft, bluish-silver-white metal, harder than tin but
softer than zinc. The metal melts at 321°C and shows a boiling point of
765°C (U.S. EPA, 1978b). Cadmium dissolves readily in mineral acids.
Some of the physical/chemical properties of cadmium and its compounds are
summarized in Table 1 (U.S. EPA, 1978b).
Cadmium is currently used in electroplating, paint and pigment
manufacture, and as a stabilizer for plastics (Fulkerson and Goeller, 1973).
Current production: 6000 metric tons (1968) (U.S. EPA, 1978b)
Projected production: 12,000 metric tons (2000) (U.S. EPA, 1978b)
Since cadmium is an element, it will persist in some form in the
environment. Cadmium is precipitated from solution by carbonate, hydrox-
ide, and sulfide ions (Baes, 1973); this is dependent on pH and on cadmium
concentration. Complexing of cadmium with other anions will produce soluble
forms (Samuelson, 1963). Cadmium -is strongly adsorbed to clays, muds, humic
and organic materials and some hydrous oxides (Watson, 1973), all of which
lead to precipitation from aqueous media. Cadmium corrodes slightly in air,
but forms a protective surface film which prevents further corrosion (U.S.
EPA, 1978b).
II. EXPOSURE
Cadmium is universally associated with zinc and appears with it in
natural deposits (Hem, 1972). Major sources of cadmium release into the
»
environment include emissions from metal refining and smelting plants, in-
cineration of polyvinyl chloride plastics, emissions from use of fossil
-------
Table 1. Some Properties of Cadmium end its Important Compounds
Primary
uue or
Compound occurrence Formula
C;i,lt:il nm Cadmium nickel Cd
injijl lniiLurles
C.i, tin luin Smelling jil.int CdO
itjfliji: ii r conl coiiihuii-
l lun eral Htjion
Co din In in I'lgment for CdS
uulllili! pliiatlca and
i.'ii.imula; jiboM-
v^ pliors
1
r CiiJinliim Fruit tree CdSO.
f .MI If. Hi: fiirulcldu
1
L'adiu linn Turf treat~ CdCOj
r.arl»tm;it u uiuut
•
.
Solubility
Molecular Physical ' Melting Dolling in watar
uulght Density state, point point ZO°C
(g/mole) (g/ml) 20°C (*C) (*C) (g/llter)
112.4 6.6 Silver metal 321 76S Insoluble
128.4 7.0 Brown powder Decomposes 0.00015
'at 900
t
144.5 4.8 Yellow crystal 1750 be composes 0.0013
208.5 4.7 White 1000 755
crystalline
172.4 4.3 White powder Decomposes 0.001
or crystalline below 500
Solubility
In oilier
solvents
Soluble lit
acid and
Nil, HO
'
Soluble In
acid and
Nllj stilts
Soluble in
acid, very
allglitly
soluble In
NII.UII
4
Insoluble
In acid
and (ilocttol
Soluble In
acid and
KCN, Kll,
suits
Acute Lethal
duse"
j
9 mg/iu la ttic
approximate
letbul concen-
tration In nuin,
inti.-i 1 cil iiu fume
50 niu/iu is tbe
approximate
luthul concen-
1 1 ill Inn In tnun,
Inlialcd; Ti
mt/kB, rat.
I.H50 (oral)
27 my/kg dog,
cutaneous)
-------
Table 1. Some Properties of Cadmium anil Its Important Compounds (Cont'd)
Ci>ni|>< 1
t'adnil urn
til! 01 Jill!
Cixlinlum
\ml list; 1 inn
Cdiliuliuii
c y an 1 il u
Vftatafy Molecular
usu or weight Density
occurrence , Formula (g/ttiole) (^/nil)
Turf urasa CJC1,, 183.3 4.0
flllllC till!
tldCtioj.liil lug K2Cd(CN)4 294.7 1.85
EltM'.Llupt at lug t'J(CH)., 164.4
1
Physical
state,
20"C
Colorless
cryutul
Colorless,
glass
crystal
Colorless
crystal
Melting Dolling
point point
(*C) Cc)
568 960
Decomposes
at ZOO
Solubility
Jn water
ZO'C
(g/lltcr)
1400
333
17, soluble
In tiot water
Solubility
in oLhur
uolvoiits
15.2 g/llter
In alcohol
Insoluble In
alcohol
Soluble In
acid and KCK
Acute Lctlinl
flB mg/kg rat.
''KuHu-iM.!! unJ Coeller, l'J73
ILiliui; .nul KI,>IIMI)U. 1973
Ulhi-r Ji
-------
fuels, use of certain phosphate fertilizers, and leaching of galvanized iron
pipes (U.S. EPA, 1978b). The major non-occupational routes of human expo-
sure to cadmium are through foods and tobacco smoke (U.S. EPA, 1979).
Based on available monitoring data, the U.S. EPA (1979) has estimated
the uptake of cadmium by adult humans from air, water, and food:
Adult
Source jug/day
Maximum conditions
Air-ambient .008 rug/day
Air-smoking 9.0
Foods 75.0
Drinking water 20.0
Total 304.008
Minimum conditions
Air-ambient 0.00002
Air-smoking 0
Food 12.0
Drinking water 1.0
Total 13.00002
The variation of cadmium levels in air, food, and water is quite exten-
sive as indicated above. Leafy vegetables, contaminated water, and air near
smelting plants all present sources of high potential exposure. The U.S.
EPA (1979) has estimated the weighted average bioconcentration factor of
cadmium to be 17 in the edible portions of fish and shellfish consumed by
Americans.
III. PHARMACQKINETICS
A. Absorption
The main routes by which cadmium can enter the body are inhalation
and ingestion. Particle size and solubility greatly influence the biolog-
ical fate of inhaled cadmium. When a large proportion of particles are in
-------
the respirable range, up to 25% of the inhaled amount may be absorbed (EPA,
1979). Cadmium fumes may have an absorption of up to 50%, and it is esti-
mated that up to 50% of cadmium in cigarette smoke may be absorbed (WHO,
1977; Elinder, et al. 1976). Large particles are trapped by the mucous mem-
branes and may eventually be swallowed, resulting in gastrointestinal
absorption (EPA, 1979).
Only a small proportion of ingested cadmium is absorbed. Two human
studies using radiolabelled cadmium have indicated mean cadmium absorption
from the gastrointestinal tract of 6% and 4.6% (Rahola, et al. 1973;
McLellan, et al. 1978), Various dietary factors interact with cadmium ab-
sorption; these include calcium levels (Washko and Cousins, 1976), vitamin D
levels (Worker and Migicovsky, 1961), zinc, iron, and copper levels (Banis,
et al. 1969). and ascorbic acid levels (Fox and Fry, 1970). Low protean
diets enhance the uptake of cadmium from the gastrointestinal tract (Suzuki,
et al. 1969).
Dermal absorption of cadmium appears to occur to a small extent;
Wahlberg (1965) has determined that up to 1.8 percent of high levels of cad-
mium chloride were absorbed by guinea pig skin.
Cadmium levels have been determined in human embryos (Chaube, et
al. 1973) and in the blood of newborns (Lauwerys, 1978), indicating passage
of cadmium occurs across the placental membranes.
B. Distribution
Cadmium is principally stored in the liver, kidneys, and pancreas
with higher levels initially found in the liver (WHO Task Group, 1977).
continued exposure leads to accumulation in all of these organs; levels as
-------
high as 200-300 mg/kg wet weight may be found in the renal cortex. This
storage appears to be dependent on the association of cadmium with the
cadmium binding protein, metallothionen (Nordberg et al., 1975).
Animal studies indicate that following intraperitoneal or intra-
venous administration of cadmium most of the compound is found in the blood
plasma. After 12-24 hours the plasma is cleared and most of the compound is
associated with red blood cells (U.S. EPA, 1978b).
The cadmium body burden of humans increases with age (Friberg, et
al. 1974) from very minimal levels at birth to an average of up to 30-40 mg
by the age of 50 in non-occupationally exposed individuals. Liver accumu-
lation continues through the last decades of life, while kidney concen-
trations increase until the fourth decade and then decline (Gross, et al.
1976). The pancreas and salivary glands also contain considerable concen-
trations of cadmium (Nordberg, 1975). Smoking effects the body burden of
cadmium; levels in the renal cortex of smokers may be double those found in
non-smokers (Elinder, et al. 1976; Hammer, et al. 1971).
C. Metabolism
Pertinent data were not found in the available literature.
D. Excretion
Since only about 6 percent of ingested cadmium is absorbed, a large
proportion of the compound is eliminated by the feces (U.S. EPA, a or b).
Some biliary excretion of cadmium has been demonstrated in rats (Stowe,
1976); this represented less than 0.1 percent of a subcutaneously adminis-
tered dose.
Urinary excretion of cadmium is approximately 1-2 mg/day in the
*
general population (Imbus, et al. 1963; Szadkowski, et al. 1969). Occupa-
tionally exposed individuals may show markedly higher urinary excretion
if -
-------
levels (Friberg, et al. 1974). A modest increase in human urinary excretion
of cadmium has been noted with increasing age (Katagiri, et al. 1971).
Additional sources of cadmium loss are through salivary excretion
and shedding of hair (U.S. EPA, 1979).
Biological half-life calculations for exposed workers have given
values of up to 200 days (urine). Direct comparisons of urinary excretion
levels and estimated body burden using Japanese, American, and German data,
suggest a half-time of 13-47 years. Using more complex metabolic models,
Frieberg, et al. 1974 concluded that the biologic half-time is probably
10-30 years. The most recent estimate of biologic half-time is 15.7 years
by Ellis (1979).
IV. EFFECTS
A. Carcinogenicity
The results of several epidemiology studies of the relationship of
cancer to occupational exposure to cadmium are summarized in Table 3 (U.S.
EPA, 1978a). The only consistent trend seen in these studies is an
increased incidence of prostate cancer in cadmium-exposed workers. A recent
study by Kjellstrom, et al. (1979) of 269 cadmium-nickel battery factory
workers found increased cancer mortality from nasopharyngeal cancer (signif-
icant) and increased mortality trends for prostate, lung, and colon-rectum
cancers (not significant). After reviewing these studies, EPA (1979) has
concluded that cadmium cannot be definitely implicated as a human carcino-
gen with the available data.
Animal experiments with the administration of cadmium by subcu-
taneous or intravenous injection have demonstrated that cadmium produces
-------
injection site sarcomas and testicular tumors (Leydigiomas) (see Table 2;
U.S. EPA, 1978a). A large number of metals and irritants produced compar-
able injection site sarcomas. Long term feeding and inhalation studies with
cadmium have not produced tumors (Schroeder, et al. 1964, Levy, et al. 1973;
Decker, et al. 1958; Anwar, et al. 1961; Paterson, 1947; Malcolm, 1972)
At the present time, the draft ambient water quality criterion for
protection of human health is based on the toxicity of cadmium rather than
on any carcinogenic effects. Though the studies summarized above qualita-
tively indicate a carcinogenic potential for cadmium, quantitatively, the
issue has not been resolved.
B. Mutagenicity
An increased incidence of chromosomal aberrations has been noted in
workers occupationally exposed to cadmium and in Japanese patients suffering
cadmium toxicity (Itai-Itai disease) (Bauchinger, et al. 1976; Bui, et al.
1975; Oeknudt and Leonard, 1976; Shiraishi and Yoshida, 1972).
Cadmium has been shown to produce mutagenic effects in vitro and in
vivo in several systems (see Table 4; U.S. EPA, 1978 a or b). These effects
include induction of point mutations in bacterial systems, chromosome aberr-
ations in cultured cells and cytogenetic damage in vivo, and promotion of
error prone base incorporation in ONA in vitro. Several investigators have
been unable to show dominant lethal effects of cadmium in mice (Epstein, et
al. 1972; Gillivod and Leonard, 1975; Suter, 1975). Point mutation studies
with cadmium in Drosophila have also produced negative findings (Shabalina,
1968; Friberg et al., 1974; Sorsa and Pfeiffer, 1973).
C. Teratogenicity
*
Damage to the reproductive tract resulting from a single dose of
parenterally administered cadmium chloride (2 mg/kg) have been observed in
-ISO-
-------
TABLE 2
STUDIES ON CADMIUM CARCINOGENESIS IN EXPERIMENTAL ANIMALS*
Authors
Animals
Compounds and routes
Tumors
Heath £t a±. , 1962; Heath and Daniel, 1964 Rats
Kazantzls, 1963; Kazantzis and Hanbury, 1966 Rats
lladdow e£ ajL. , 1964; Roe ^t ^1., 1964 Rats
Cuthrie, 1964 Chickens
Guaa _et. £l. , 1963; 1964; 1965; 1967 Rats, Mice
V Schroedtjr e^ a^L. , 1965; Kanisawa and Rats, Mice
l Schroiidur, 1969
i*»
"»
7 Nazarl £t al_., 1967; Favion et al., 1968 Rats
Knorrc, 1970; 1971 ' ^ Rats
I.ucls e£ £l. , 1972; 1973 Rats
Reddy ^t aj_. , 1973 Rats
Levy tii_ jil_, , 1973' Rats
Levy and CJark, 1975; Levy e_t al,., 1975 Rats, Mice
Cd powder in fowl serum (im)
CdS, Cdo (sc)b
CdSO^, CdCl2 (sc)
CdCl2 (intratesticular)
(ira)
Cd-acetate (drinking water)
CdCl2 (sc)
CdCl2 (sc)
CdCl2 (sc, intrahepatic)
CdCl2 (sc)
CdSO, (sc)
CdSO, (gastric Intubation)
Sarcomas
i
Sarcomas
Sarcoms and Leydigiomas
Teratoma :
Sarcomas and Leydigiomas
No Tumorlgcnic Effect
Sarcomas and Leydigiomas
Sarcomas and Leydigiomas
Sarcomas and Leydigiomas
Leydigiomas
Sarcomas
No Tumorlgenic Effect
Adapted from Sunderman, 1977.
lar: im; subcutaneous: sc.
-------
TABLE 3
SUMMARY OF RESULTS OF HUMAN EPIDEMIOLOGY STUDIES OF CANCER EFFECTS
ASSOCIATED WITH OCCUPATIONAL EXPOSURES TO CADMIUM
Population
Croup Studied
liattery factory
uorkurs
Battery factory
workers
Cadmium smelter
worker a
Rubber industry
Cadmium Compound
Exposed To
Cadmium oxide
Cadmium oxide
Cadmium oxide,
others
Cadmium oxide
Incidences of
All Cancers
High
Normal
High
High
Incidences .of
Lung Cancer
Normal
Normal
High
Normal
Incidences of
Prostrate Cancer
High
High
High
High
Reference
Potts (1965)
Kipling and
Waterhouse
(1967)
Lemon et al.
(1976)
McMlchael et al.
workers
(1976)
-------
rats, rabbits, guinea pigs, hamsters, and mice (Parizek and Zahor, 1956;
Parizek, 1957; Meek, 1959). This susceptibility appears to be genetically
regulated since different strains of mice show differential susceptibility
(Wolkowski, 1975).
Teratogenic effects of cadmium compounds administered parenterally
have been reported in mice (Eto, et al. 1975), hamsters (Perm and Carpenter,
1968; Mulvihill, et al. 1970; Ferm, 1971; Gale and Ferm, 1973) and rats
(Chernoff, 1973; Barr, 1973). Oral administration of cadmium (10 ppm) has
demonstrated teratogenic effects in rats (Schroeder and Mitchener, 1971),
but no teratogenicity has been reported in rats and monkeys (Cuetkova, 1970;
Pond and Walker, 1975; Willis, et al. 1976; Campbell and Mills, 1974).
D. Other Reproductive Effects
Rats in late pregnancy are apparently more sensitive to cadmium
than non-gravid animals or those immediately post-partum. A single dose of
2-3 mg/kg of body weight given during the last 4 days of pregnancy resulted
in high mortality (76 percent).
In addition to the embryotoxic effects of cadmium indicated in
Section C, persisting effects of cadmium exposure during pregnancy on postu-
lated development and growth of offspring have been observed. This Includes
neurobehavioral alteration in newborn rats (Chowdbury and Lauria, 1976) and
growth deficiencies in lambs (U.S. EPA, 1978a).
E. Chronic Toxicity
Friberg (1948, 1950) observed emphysema in workmen exposed to cad-
mium dust in an alkaline battery factory. This finding has subsequently
been well documented (U.S. EPA, 1979).
-------
TABLE ft
SUMMARY OF MUTAGENICITY TEST RESULTS
Test System
Genetic Effect
Reported
Hutagenicity
References
Hitman cells
Chinese Hamster Cells
S. cerevibiae
~-SMllliIIii recombinant
assay
Polymicleotides
Systems in vitro
Chromosomal damage
Point mutation
Point mutation
Gene mutation
Base mispairing
•f
+
+
Shiraishi et_^l.', 1972
Costa et^ £l., 1976
Takahoshi, 1972
Nishioka, 1975
Sirover and Loeb, 1976
\
v>
(A
Human leukocytes
Human leukocytes
Human leukocytes
Human leukocytes
Itat spermatogonla
Mout/i; oocytes
Mouse breeding
Mou.se breeding
Mouse breeding
Mammals
1).^ melanogaster
Systems in vivo
Chromosomal damage
Chromosomal damage
Chromosomal damage
Chromosomal damage
Altered spermatogenesis
Cytogenetic damage
Dominant lethal mutations
Dominant lethal mutations
Dominant lethal mutations
Chromosomal abnormalities
Sex-linked recessive lethal
•Shlrashl and Yoshida, 1972
Bui e£ al., 1975
Deknudt and Leonard, 1975
Bauchinger et_ al., 1976
Lee and Dixon, 1973
Shimada et^ al., 1976
Epstein £t _al., 1972
Gilliavod and Leonard, 1975
Suter, 1975
Shimada et, al., 1976
Sorsa and Pfeifer, 1973
-------
Chronic cadmium exposure produces renal tabular damage that is
characterized by the appearance of a characteristic protein 9B2-micro-
globulin) in the urine. Renal damage has been estimated to occur when
cadmium levels in the renal cortex reach 200 mg/kg (Kjellstron, 1977).
Itai-Itai disease is the result of cadmium induced renal damage plus osteo-
malacia (U.S. EPA, 1978a).
Exposure to high ambient cadmium levels may contribute to the etio-
logy of hypertension (U.S. EPA, 1979). Several studies, however, have been
unable to show a correlation between renal levels of cadmium and hyper-
tension (Morgan 1972; Lewis, et al. 1972; Beevers, et al. 1976).
Friberg (1950) and Blejer (1971) have noted abnormal liver function
tests in workers exposed to cadmium; however, these workers were occupa-
tionally exposed to a variety of agents.
The Immunosuppressive effects of cadmium exposure, including an in-
creased susceptibility to various infections, have been reported in several
animal studies (Cook, et al. 1975; Koller, 1973; Exon, et al. 1975).
V. AQUATIC TOXICITY
A. Acute Toxicity
Acute toxicity in freshwater fish has been studied in a number of
96-hour bioassays consisting of one static renewal, 22 static, and 19 flow-
through tests. LC50 values ranged from 1 ug/1 for stripped bass larvae
(Roccus saxatilus) (Hughes, 1973) to 73,500 for the fathead minnow
(Pimephales promelas) (Pickering and Henderson, 1966). Increased resistance
to the toxic action of cadmium in hard waters was observed. The LC5Q
values for freshwater invertebrates ranged from 3.5 for Cladoceran
»
(Simoephalus serrulatus) to 28,000 pg/1 for the mayfly (Ephemerella qrandis
qrandis). Acute LC-- values for marine fish ranged from 1,600 jjg/1 for
-356--
-------
larval Atlantic silversides (Menidla menida) (Middaugh and Dean, 1977) to
114,000 pg/1 for juvenile mummichog (Fundulus heteroclitus) (Voyer, 1975).
Intraspecific and life stage differences have shown that larval stages of
the Atlantic silversides and mummichog are four times more sensitive than
adults under the same test conditions (Middaugh and'Dean, 1977). Marine
invertebrates are more sensitive to cadmium than are marine fishes. LC__
values ranged from 15.5 jjg/1 for the. mysid shrimp (Nimmo, et al. 1977a) to
46,600 for the fiddler crab (Uca puqilator) (O'Hara, 1973).
B. Chronic Toxicity
Chronic values for freshwater fish ranged from 0.9 ug/1 in a brook
trout (Salvelinus fontinalis) embryo larval assay (Sauter, et al. 1976) to
50 jug/1 in a life cycle (or partial life cycle) assay for the bluegill
(Lepomis marcochirus) in hard water (Eaton, 1974). Salmonids were in
general the most sensitive species examined. Data for freshwater inverte-
brates depend on a single jug/1 obtained for Daphnia maqna (Biesinger and
Christensen, 1972). No chronic studies were available for cadmium effects
in marine fishes. The only marine invertebrates data reported was the
chronic value of 5.5 pg/1 for the mysid shrimp, Mysidapsis bahia. In this
animal no measurable effects on brood appearance in the pouch, release,
average number per female, or survival were observed at concentrations of
4.8 jug/1.
C. Plant Effects
Effective concentrations for freshwater plants ranged from 2 jjg/1,
which causes a 10 fold growth rate decrease in the diatom, Asterionella
formosa (Conway, 1978), to 7,400 jug/1, which causes a 50% root weight^inhi-
bition in Eurasian water-milfoil (Myriophyllum spicatum). In marine algae,
-------
96-hour £C5Q growth rate assays yielded values of 160 and 175 jjg/1 for
Cyclotella nana and Skeletonema costatum respectively (Gentile and Johnson,
1974).
0. Residues
Bioconcentration factors ranged from 151 for brook trout to 1,988
for the flagfish (Jordanella floridae). One characteristic of cadmium tox-
icity in aquatic organisms was the possible long half-life of the chemical
in certain tissues of exposed brook trout even after being placed in clean
water for several weeks. Testicular damage to adult mallards was observed
when fed 20 mg/kg cadmium in the diet for 90 days. In marine organisms
bioconcentration values ranged from 37 for the shrimp Crangpn cranqon to
1,230 for the American oyster, Crassostrea virginica (Schuster and Pringle,
1969).
E. Miscellaneous
Several studies on marine organisms have demonstrated significant
reduction in gill-tissues respiratory rates in the cunner, Tautogolabrus
adepersus, the winter flounder, Pseudopleuronectes americanus, and the
stripped bass, Morone saxatilis, at concentrations as low as 0.5 jug/1.
VI. EXISTING GUIDELINES
A. Human
It is not recommended that cadmium be considered a suspect human
carcinogen for purposes of calculating a water quality criterion (U.S. EPA,
1979).
The EPA Primary Drinking Water Standard for protection of human
health is 10 ug/1. This level was also adopted as the draft ambient water
quality criterion (U.S. EPA, 1979).
-------
The OSHA time-weighted average exposure criterion for cadmium is
100 pg/m3.
B. Aquatic
The draft criterion proposed for freshwater organisms to cadmium
has been prepared following the Guidelines, and is listed according to the
following equation:
e(0.867 In-(hardness) - 4.38)
for a 24-hour average and not to exceed the level described by the following
equation:
(1.30 In-(hardness) - 3.92)
The proposed marine criterion derived following the Guidelines is 1.0 ug as
a 24-hour average not to exceed 16 jug/1 at any time (U.S. EPA, 1979).
-------
CADMIUM
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-------
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-------
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-------
Levy, L.S., et al. 1973. Absence of prostatic changes in rts
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-------
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-------
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109
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*
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-------
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-------
No. 32
Carbon Bisulfide
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-uracy.
-------
CARBON DISULFIDE
I. PHYSICAL AND CHEMICAL PROPERTIES
It is soluble in water at 0.294% at 20°C. It can chelate trace metals,
especially Cu and Zn. Its formula weight is 76.14 and it is a colorless,
volatile, and extremely flammable liquid at Rt. No odor when pure. At 27°C,
its vapor pressure is 200 mm Hg.
II.PRODUCTION AND USE
It is produced in pretroleum and coal tar refining. Its principal uses
are in the manufacture of rayon, rubber, chemicals, solvents, and pesticides.
* 0
In 1974, 782 million pounds of CS2 were produced in the United States. In
1971, 53% was used in production of viscose rayon and cellphane and 25% for
manufacture of CC,..
III. EXPOSURE
It was detected in 5 of 10 water supplies surveyed by the EPA.3 NIOSH25
estimates that 20,000 employees are potentially exposed to CS2 fulltime in the
United States.
III. PHARMACOKINETICS
A. Absorption: Absorption differs with species and route of administra-
4
ti on.
B. Distribution: Large concentrations of both free and bound CSp are
found in brain (guinea pig) and peripheral nerves (rats) of exposed animals.
The ratio of bound to free CS- in brain is 3:1. Blood and fatty tissues
contain mainly bound CS- while liver contains mainly free.
C. Metabolism: It is 90% metabolized by the P-450 system to inorganic
sulfate. A portion of the S released by CS2 is thought to react with SH
groups of cysteine residues in the microsomal protein to form hydro-sulfide.
M. Greenberg
ECAO/RTP NC
-------
D. Excretion: Small amounts are excreted (0.5%) as thiourea, 5-mercapto-
thioazolidone, and inorganic constituents in urine. Some portion (8-10%) is
also excreted unchanged in the breath. Inhalation studies have shown that 18%
of the CSj inhaled is exhaled unchanged. Of the remaining inhaled dose, 70%
is excreted as free or bound CS2 and urinary sulfates and 30% is stored in the
body and slowly excreted as CS- and its metabolites.
V. EFFECTS ON MAMMALS
A, Carcinoqem'city: No available data.
B. Mutagem'city: No available data.
8 3
C. Teratogem'city: Bariliah et al. showed that inhalation of 10 rng/m
was lethal to embryos before and after implantation. CS- at 2.2 gm/m for 4
hr/day was embryotoxic if given to female rats during gestation and had no
g
effect on male rats. Inhalation of lower concentrations (0.34 mg/1 for 210
days) caused disturbances of estrus. t In a dominant lethal test, inhalation
3 8
of 10 mg/m by male rats before copulation proved lethal to embryos.
D.
E. Toxicity
1. Humans
The lowest lethal concentration has been reported as 4,000 ppm in 30
minutes. In the same study, a person subjected to a concentration of 50
mg/m for 7 years had CNS effects. Moderate chronic exposure of humans at
3 12
less than 65 mg/m for several years has been reported by Cooper to cause
polyneuropathy. In a study by Baranowska et al. humans have been shown to
absorb 8.8-37.2 mg from an aqueous solution containing 0.33-1.67 q/1. This -
was over a period of 1 hour of hand-soaking.
The most thoroughly documented studies on health effects of C$2 exposure
2g—90 oc 27
have been on cardiovascular system. Heinberg et al. ' reported significantly
-------
elevated rates of coronary heart disease mortality, angina, and high blood
pressure. In a 5-year followup of these vicose rayon workers, he reported
again increased coronary heart disease mortality and higher than expected
incidences of total infarctions, nonfatal infarctions, angina. In an 8-year
followup in 1976, Heinberg found no excess coronary heart disease mortality
during the last 3 years of the followup.
2. Other species.
IP injection of 400 mg/hg was the lowest lethal dose in guinea pigs.
An IV LD50 of 694 mg/kg in mice was reported by Hylen and Chin.15
With SC injection, LD50 was 300 mg/kg in rabbits.16 Toxic effects have
been observed at 1.7 mg/kg in rabbits.17 Rats showed toxic SC effects at 1
mg/kg. Oral doses in rats produced toxic effects at 1 mg/kg.18'19 Vinogradov20
showed that 1 ppm in drinking water was nontoxic to rabbits; 70 ppm was fatal.
In a chronic study, Paterni et al. found that 6 mg/kg/day produced
toxic effects in rabbits. The lowest lethal chronic dose for rabbits was
shown to be 0.1 ml 3 times a week for 7 months.22
Applied topically, it produced a higher incidence of anemia in female than
7"\
in male rats and teratogenic effects were observed. When rats inhaled CS2
at 10 mg/m , abnormalities of genitourinary and skeletal systems were found.
Disturbances of ossification and blood formation and dystrophic changes in
Q
liver and kidney were noted.
-37O-
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VI. EXISTING GUIDELINES AND STANDARDS
The NAS did not recommend limits for drinking water because estimates of
effects of chronic oral exposure cannot be made with any confidence.
24- 3
The NIOSH recommended standard IS 3 mg/m .
Human studies have shown that exposure effects the cardiovascular system,
the nervous system, the eyes, the reproductive organs, and other systems.
25 ^
The current federal standard is 20 ppm (62 mg/m ) with a ceiling
concentration of 30 ppm (93 mg/m ) for an 8-hour day, 5 day work week.
tf
-37/-
-------
REFERENCES
1. U.S. Environmental Protection Agency. Identification of organic compounds
in effluents from industrial sources, 1975.
2. U.S. International Trade Commission, Syn. Org. Chem., 1974.
3. U.S. Environmental Protection Agency. Preliminary Assessment of suspected
carcinogens in drinking water. Report to Congress. EPA 560-14-75-005 PB
260961, 1975.
4. NAS. Drinking Water and Health, 1977.
5. Dalve et al. Chem. Biol. Inter. 10:347-361, 1975.
6. Catiguani and Neal. 8BRC 65(2):629-636, 1975.
.7. Theisinger. Am. Ind. Hyg. Assoc. 35(2):55-61, 1974.
8. Bariliah et al. Anat. Gistol. Embriol. 68(5):77-81, 1975.
9. Sal'nikova and Chirkova. Gig. Tr. Prof. Zabol 12:34-37, 1974.
10. Rozewiski et al. Med. Pr. 24(2):133-139, 1973.
11. Registry of Toxic Effects of Chemical Substances, 1975.
12. Cooper. Food Cosmet. Toxicol. 14:57-59, 1976.
13. Saranowska et al. Ann. Acad. Med. Lodz 8:169-174, 1966. Chem. Abs.
70:31443W, February 24, 1969.
14. Davidson and Feinlab. Am. Heart J. 83(1):100-114, 1972.
15. Hylin and Chen. Bull. Environ. Contam. Toxicol. 3(6):322332, 1968.
16. Merch Index, 1968.
17. Okamoto. Tokyo Jikeikai Ika Daigaku Zasshi 74:1184-1191, 1959.
18. Freun.dt et al. Int. Arch Arfaeitsmed. 32:297-303, 1974.
19. Freundt et al. Arch. Toxicol. 32:233-240, 1974.
20. Vinogradov. Gig. Sanit. 31(1):13-18, 1966.
21. Paterm et al. Folia Med. 41:705-722, 1958.
22. Michalova et al. Arch. Gewerbepth Gewerbehgy 16:653-665, 1959.
-373-
-------
23. Gut. Prac. Lek. 21(10):453-458, 1969.
24. NIOSH. Criteria for a Recommended Standard CS-, May, 1977.
25. 29 CFR 1910, 1000.
26. Hernfaerg. Br. J. Ind. Med. 27:313-325, 1970.
27. Hernberg et al. Work Env. Health 8:11-16, 1971.
28. Hernberg et al. Work Env. Health 10:93-99, 1973.
29. Tolonen et al. Br. J. Ind. Med. 32:1-10, 1975.
30. Heinberg et al. Work Env. Health 2:27-30, 1976.
St
-373-
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No. 33
Carbon Tetrachloride (Tetrachlororaethane)
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.
-37S"-
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (GAG) has evaluated
carbon tetrachloride and has found sufficient evidence to
indicate that this compound is carcinogenic.
-------
CARBON TETRACHLORIDE
Summary
Carbon tetrachloride (CC1,) is a haloalkane with a wide range of in-
dustrial and chemical applications. lexicological data for non-human mam-
mals are extensive and show that CC1. causes liver and kidney damage, bio-
chemical changes in liver function, and neurological damage. CCl^ has
been found to induce liver cancer in rats and mice. Mutagenic effects have
not been observed and teratogenic effects have not been conclusively demon-
strated.
The data base on aquatic toxicity is limited. LC^ (96-hour) values
for bluegill range from 27,300 to 125,000 pg/1 in static tests. For Daphnia
magna, the reported 48-hour EC5Q is 35,200 jjg/1. The 96-hour LC^g fqr
the tidewater silverside is 150,000 pg/1. An embryo-larval test with the
fathead minnow showed no adverse effect from carbon tetrachloride concentra-
tions up to 3,400 jug/1. No plant effect data are available. The bluegill
bioconcentrated carbon tetrachloride to a factor of 30 times within 21 days
exposure. The biological half-life in the bluegill was less than 1 day.
-377-
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CARBON TETRACHLORIDE
I. INTRODUCTION
Carbon tetrachloride (CC1.) is a haloalkane with a wide range of in-
dustrial and chemical applications. Approximately 932.7 million pounds are
produced at 11 plant sites in the U.S. (U.S. EPA, 1977b; Johns, 1976). The
bulk of CCl^ is used in the manufacture of fluorocarbons for aerosol pro-
pellants. Other uses include grain fumigation, a component in fire extin-
guisher solutions, chemical solvent, and a degreaser in the dry cleaning in-
dustry (Johns, 1976).
Carbon tetrachloride is a heavy, colorless liquid at room temperature.
Its physical/chemical properties include: molecular weight, 153.82; melting
point, -22.99°C; solubility in water, 800,000 jug/1 at 25°C; and vapor
pressure, 55.65 mm Hg at 10°C. CC14 is relatively non-polar and misciT
ble with alcohol, acetone and most organic solvents.
Carbon tetrachloride may be quite stable under certain environmental
conditions. The hydrolytic breakdown of CC1, in water is estimated to re-
"quire 70,000 years for 50 percent decomposition (Johns, 1976). This decom-
position is accelerated in the presence of metals such as iron (Pearson and
McConnell, 1975). Hydrolytic decomposition as a means of removal from water
is insignificant when compared with evaporation. In one experiment the
evaporative half-life of CC14 in water at ambient temperatures was found
to be 29 minutes (Oilling, et al. 1975), but this is highly dependent on ex-
perimental conditions, such as surface area to bulk volume ratios. For ad-
ditional information regarding Halomethanes as a class, the reader is refer-
red to the Hazard Profile on Halomethanes (U.S. EPA, 1979b).
-------
II. EXPOSURE
A. Water
CC1, has been found in many water samples including rain, sur-
•face, potable, and sea, in the sub-part per billion range (McConnell, et al.
1975). The National Qrganics Monitoring Survey (NOMS) found CC1. in 10
percent of 113 public water systems sampled, with mean values ranging from
2.4-6.4jjg/l (U.S. EPA, 1977a).
Although CC1, is a chlorinated hydrocarbon, it is not produced in
finished drinking water as a result of the chlorination process (Natl. Res.
Coun., 1977,1978).
B. Food
Carbon tetrachloride has been detected in a variety of foodstuffs
other than fish and shellfish in levels ranging from 1 to 20 jug/kg (McCon-
nell, et al. 1975).
Results of various studies on CC1, fumigant residues in food in-
dicate that the amount of residue is dependent upon fumigant dosage, storage
conditions, length of aeration and the extent of processing (U.S. EPA,
1979a). Usually, proper storage and aeration reduce CC1 residues to
trace amounts.
The U.S. EPA (1979a) has estimated the weighted average bioconcen-
tration factor for carbon tetrachloride to be 69 for the edible portions of
fish and shellfish consumed by Americans. This estimate is based on measur-
ed steady-state bioconcentration studies in bluegills.
C. Inhalation
The occurrence of CC1, in the atmosphere is due largely to the
^ »
volatile nature of the compound. Concentrations of CC1, in continental
and marine air masses range from .00078 - .00091 mq/wP. Although some
-------
higher quantities (.0091 mg/m3) have been measured in urban areas, concen-
trations of CC1, are universally widespread with little geographic varia-
tion (U.S. EPA, 1979a).
III. PHARMACQKINETICS
A. Absorption
CCl^ is readily absorbed through the lungs, and more slowly
through the gastrointestinal tract (Nielsen and Larsen, 1965). It can also
be absorbed through the skin. The rate and amount of absorption are enhanc-
ed with the ingestion of fat and alcohol (Nielson and Larson, 1965; Moon,
1950). Robbins (1929) found that considerable amounts of CCl^ are absorb-
ed from the small intestine, less from the colon, and little from the stom-
ach. Absorption from the gastorintestinal tract appears to vary by species,
i.e., it occurs more rapidly in rabbits than dogs.
B. Distribution
The organ distribution of CCl^ varies with the route of adminis-
tration, its concentration, and the.duration of exposure (U.S. EPA, 1979a).
After oral administration to dogs, Robbins (1929) found the highest
concentrations of CCl^ in the bone marrow. The liver, pancreas and spleen
had one-fifth the amount found in the bone marrow. The highest concentra-
tions of CCl^ after inhalation, however, were found in the brain (Von Oet-
tingen, et al. 1949,1950). After inhalation of CC14 by monkeys, the high-
est levels were detected in fat, followed by liver and bone marrow (McColli-
ster, et al. 1950). McConnell, et al. (1975) found human tissue levels of
CC14 to range as follows: kidney, 1-3 rag/1; liver, 1-5 mg/1 and fat, 1-13
mg/1.
On the cellular level, McClean, et al. (1965) found CC14 in all
cell fractions with higher concentrations in ribosomes.
-3SO-
-------
C. Metabolism
When CC1. is administered to mammals, it is metabolized to a
small extent, the majority being excreted through the lungs. The metabo-
lites include chloroform, hexachloroethane, and carbon dioxide. These meta-
bolites play an important role in the overall toxicity of CC14 (U.S. EPA,
1979a). Some of the CC14 metabolic products are also incorporated into
fatty acids by the liver and into liver microsomal proteins and lipids (Gor-
dis, 1969).
The chemical pathology of liver injury induced by CCl^ is a re-
sult of the initial homolytic cleavage of the C-C1 bond which liberates tri-
chloromethyl- and chlorine-free radicals (Fishbein, 1976). The next step
may be one of two conflicting reactions: direct attack via alkylation on
cellular constituents (especially sulfhydryl groups), or peroxidative decom-
position of lipids of the endoplasmic reticulum as a key link between the
initial bond cleavage and the pathological phenomena characteristic of
CC14 (Butler, 1961; Tracey and Sherlock, 1968).
D. Excretion
The largest portion of absorbed CCl^ is rapidly excreted. Ap-
proximately 50-79 percent of absorbed radioactive CCl^ is eliminated
through the lungs, and the remainder is excreted in the urine and feces. No
CCl^ was detected in the blood or in the expired air, 48 hours and 6 days,
respectively, after CC1. inhalation (Beamer, et al. 1950). CCl^ is ex-
creted as 85 percent parent compound, 10 percent carbon dioxide, and smaller
quantities of other products including chloroform (NRC, 1977).
IV. EFFECTS
*
A. Carcinogenicity
CCl^ has been shown to be carcinogenic in rats, mice, and ham-
sters via subcutaneous injection, intubation, and rectal instillation (U.S.
If
-3SJ-
-------
EPA, 1979). Current knowledge lead to the conclusion that carcinogenesis is
a non-threshold, non-reversible process. However, some scientists do argue
that a threshold may occur.
Rueber and Glover (1970) administered injections of 1.3 ml/kg of
body weight of a 50 percent solution of CC1, in corn oil to rats, two
times per week until death. Carcinoma of the liver were present in 12/15
(80 percent) Japanese male rats, 4/12 (33 percent) Wistar rats, and 8/13 (62
percent) Osborne-Mendel rats, whereas Black Rats or Sprague-Dawley rats did
not develop carcinomas. The incidence of cirrhosis of the liver also dif-
fered with the strain of the rat. Carcinoma of the liver tended to develop
along with mild or moderate, rather than severe cirrhosis of the liver.
When administered with CCl^, methylchplanthrene (a potent enzyme inducer)
was found to increase the incidence of hyperplastic hepatic nodules aqd
early carcinomas in rats (Rueber, 1970). Females were found to be more sus-
ceptible to the development of hyperplastic nodules and carcinomas.
The National Cancer Institute (1976) studied the carcinogenic ef-
fect of CC14 in male and female mice (1,250 mg/kg or 2,500 mg/kg of body
weight, oral gavage 5 times/week/78 weeks). Hepatocellular carcinomas were
found in almost all of the mice receiving CC1.. Andervant and Dunn (1955)
transplanted 30 CCl^-induced tumors into mice. They observed growth in 28
of the hepatomas, through 4 to 6 transplant generations.
B. Mutagenicity
Conclusive evidence on the mutageniciity of CC1. has not been re-
ported. Kraemer, et al. (1976) found negative results using the Ames bac-
terial reversion tests. However, they explain that halogenated hydrocarbons
are usually negative in the Ames test.
-------
C. Teratogenicity
Very little data are available concerning the teratogenic effects
of CC14. Schwetz, et al. (1974) found CC14 to be slightly embryotoxic,
and to a certain degree retarded fetal development, when administered to
rats at 300 or 1,000 mg/1 for 7 hr/day on days 6 through 15 of gestation.
Bhattacharyya (1965) found that subcutaneous injection occasionally gave
rise to changes in fetal liver.
D. Other Reproductive Effects
Pertinent data concerning other reproductive effects of CCl^ were
not encountered in the available literature.
E. Chronic Toxicity
Cases of chronic poisoning have been reported by Sutsch (1932),
Wirtschafter (1933), Strauss (1954), Von Oettingen (1964), and others. The
clinical picture of chronic CC1. poisoning is much less characteristic
than that of acute poisoning. Von Oettingen (1964) has done an excellent
job of reviewing the symptoms. Patients suffering from this condition may
complain of fatigue, lassitude, giddiness, anxiety, and headache. They suf-
fer from paresthesias and muscular twitchings, and show increased reflex ex-
citability. They may be moderately jaundiced, have a tendency to hypogly-
cemia, and biopsy specimens of the liver may show fatty infiltration. Pa-
tients may complain of a lack of appetite, nausea, and occasionally of diar-
rhea. In some instances, the blood pressure is lowered and is accompanied
by pain in the cardiac region and mild anemia. Other patients have develop-
ed pain in the kidney region, dysuria, and slight nocturia, and have had
urine containing small amounts of albumin and a few red blood cells. Burn-
ing of the eyes and, in a few instances, blurred vision are frequerlt com-
plaints of those exposed. If these symptoms are not pronounced, or of long
-383-
-------
standing, recovery usually takes place upon discontinuation of the exposure
if the proper treatment is received (Von Oettingen, 1964).
Reports on pathological changes in fatalities from CCl^ poison-
ings are generally limited to findings in the liver and kidneys. The brain
and lungs may be edematous. The intestines may be hyperemic and covered
with numerous petechial hemorrhages and the spleen may be enlarged and hy-
peremic. Occasionally the adrenal glands may show degenerative changes of
the cortex and the heart may undergo toxic myocarditis (Von Oettingen, 1964).
F. Other Relevant Information
The toxic effects of CC1, are potentiated by both the habitual
and occasional ingestion of alcohol (U.S. EPA, 1979a). Pretreatment of lab-
oratory animals with ethanol, methanolj or isopropanol increases the suscep-
tibility of the liver to CC1A (Wei, et al. 1971; Traiger and Plaa, 1971). ;
Hafeman and Hoekstra (1977) reported that protective effects
against CC1.-induced lipid peroxidation .are exhibited by vitamin E, sele-
nium, and methionine.
According to Davis (1934), very obese or undernourished persons or
those suffering from pulmonary diseases, gastric ulcers or a tendency to
vomiting, liver or kidney diseases, diabetes or glandular disturbances, are
especially sensitive to the toxic effect of CC1, (Von Oettingen, 1964).
V. AQUATIC TOXICITY
A. Acute Toxicity
Two studies have investigated the acute toxicity of carbon tetra-
chloride to bluegills (Lepomis macrochirus) in static tests. The determined
LC5Q varied from 27,300 ug/1 to 125,000 ug/1 (Dawson, et al. 1977; U.S.
EPA, 1978). With Daphnia magna, the reported 48-hr. EC5Q is 35,200 jjg/1
(U.S. EPA, 1978). The 96-hr. LC5Q for the tidewater silversides (Menidia
beryllina) is 150,000 jjg/1 (Oawson, et al. 1977).
-------
B. Chronic Toxicity
An embryo-larval test with the fathead minnow (Pimephaies promelas)
showed no adverse effect from carbon tetrachloride concentrations up to
3,400 jjg/L (U.S. EPA, 1978). Other chronic data are not available.
C. Plant Effects
There are no data in the available literature describing the ef-
fects of carbon tetrachloride on freshwater or saltwater plants.
D. Residues
The bluegill bioconcentrated carbon tetrachloride to a factor of 30
times within 21 days. The biological half-life in these tissues was less
than 1 day. -
VI. EXISTING GUIDELINES AND STANDARDS
Neither the human health nor the aquatic criteria derived by U.S. EPA
(1979a), which are summarized below, have been reviewed; therefore, there is
a possibility that these criteria will be changed.
A. Human
The American Conference of Governmental Industrial Hygienists
(1971) recommends a threshold limit value (TLV) of 10 mg/m for CC1,,
with peak values not to exceed 25 mg/m even for short periods of time.
The Occupational Safety and Health Administration adopted the American Na-
tional Standards Institute (ANSI, 1967) standard Z37.17 - 1967 as the Feder-
al standard for CC14 (29 CFR 1910.1000). This standard is 10 mg/m3 for
an 8-hour TWA, with an acceptable ceiling of 25 mg/m and a maximum peak
for 5 minutes in any 4-hour period of 200 mg/m .
The draft ambient water quality criteria for carbon tetrachloride
has .been set to reduce the human carcinogenic risk levels to 10" •, 10"
or 10~ (U.S. EPA, 1979a). The corresponding criteria are 2.6 jjg/1, 0.26
-------
pg/1, and 0.026 pg/1, respectively. Refer to the Halomethane Hazard Profile
for discussion of criteria derivation (U.S. EPA, 1979b).
8. Aquatic
For carbon tetrachloride, the drafted criteria to protect fresh-
water aquatic life is 620 ug/1 as a 24-hour average and the concentration
should never exceed 1,400 ug/1 at any time. To protect saltwater aquatic
life, the drafted criterion is 2,000 ug/1 as 24-hour average and the concen-
tration should not exceed 4,600 ug/1 at any time (U.S. EPA, 1979a).
-38-6-
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No. 34
Chloral
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.
-------
CHLORAL
Summary
Chloral (trichloroacetaldehyde) is used as an intermediate in the manu-
fscture of DDT, methoxychlor, DDVP, naled, trichlorfon, and TCA. Chloral is
readily soluble in water, forming chloral hydrate. Chloral hydrate decom-
poses to chloroform with a half-life of two days. Chloral hydrate has been
used as a therapeutic agent due to its hypnotic and sedative properties.
Chloral (as chloral hydrate) has been identified in chlorinated water
samples at concentrations as high as 5.0 ug/1. Chloral hydrate is formed
through the chlorination of natural humic substances in the raw water. At-
mospheric chloral concentrations up to 273.5 mg/m3 have been reported from
spraying and pouring of polyurethanes in Soviet factories. Similar data on
exposure levels in U.S. plants were not found in the available literature.
Specific information on the pharmacokinetic behavior, carcinogenicity,
mutagenicity, teratogenicity, and other reproductive effects of chloral was
not found in the available literature. However, the pharmacokinetic be-
havior of chloral may be similar to chloral hydrate where metabolism to tri-
chloroethanol aid trichloroacetic acid and excretion via the urine (and pos-
sibly bile) have been observed. Chloral hydrate produced skin tumors in A
of 20 mice dermally exposed. Information on the chronic or acute effects of
chloral in humans was not found in the available literature. Chronic ef-
fects from respiratory exposure to chloral as indicated in laboratory
animals include reduction of kidney function and serum transaminase activ-
ity, change in central nervous system function (unspecified), decrease in
-------
antitoxic and enzyme-synthesizing function of the liver, and alteration of
morphological characteristics of peripheral blood. Slowed growth rate, leu-
kocytosis and changes in arterial blood pressure were also observed. Acute
oral LD5Q values in rats ranged from 0.05 to 1.34 g/kg.
U.S. standards and guidelines for chloral were not found in the avail-
able literature.
-390-
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CHLORAL
ENVIRONMENTAL FATE
Chloral (trichloroacetaldehyde) is freely soluble in water, forming
chloral hydrate (Windholz, et al. 1576). Chloral hydrate was identified in
drinking water from 6 of 10 cities sampled (Keith, 1976). The author postu-
lated that chloral hydrate was formed by the chlorination of other compounds
during the addition of chlorine to the water supplies. Chloral hydrate was
not identified prior to chlorination. Chloral hydrate may be formed by the
chlorination of ethanol or acetaldehyde and may occur as an intermediate in
the reaction involving the conversion of ethanol to chloroform as follows:
Ethanol - Acetaldehyde - Chloral - Chloral hydrate - Chloroform
Chloral hydrate decomposes to chloroform with a half-life of 2 days at pH 8
and 35°C (Luknitskii, 1975). Rook (1974) demonstrated the formation of
haloforms from the chlorination of natural humic substances in raw water.
Chloral polymerizes under the influence of light and in the presence of
sulfuric acid, forming a white solid trimer called metachloral (Windholz,
1976). Dilling, et al. (1976) studied the effects of chloral on the decom-
position rates of trichloroethylene, NO, and NO in the atmosphere and ob-
served that chloral increases the photodecomposition rate of trichloro-
ethylene to a greater extent than it does NO or NQ2.
-------
CHLORAL
I. INTRODUCTION
This profile is based on literature searches in Biological Abstracts,
Chemical Abstracts, MEDLINE, and TOXLINE.
Chloral [C13CCHO], also referred to as trichloroacetaldehyde, anhy-
drous chloral, and trichloroethanol, is an oily liquid with a pungent, ir-
ritating odor. The physical properties of chloral are: molecular weight,
147.39; melting point, -57.5°C; boiling point, 97.75°C at 760 mm Hg;
density, 1.5121 at 20/4°C (Weast, 1976), The compound is very soluble in
water, forming chloral hydrate, and is soluble in alcohol and ether.
Industrial production of chloral involves direct chlorination of ethyl
alcohol followed by treatment with concentrated sulfuric acid (Stanford
Research Institute, 1976). Production may also occur by direct chlorination
of either acetaldehyde or paraldehyde in the presence of antimony chloride.
Prior to 1972, essentially all chloral produced was used in the manufacture
of DOT. Production of chloral was greatest in 1963 at 79.8 million pounds,
decreasing to 62.4 million pounds in 1969. Production data after 1969 were
not reported. Consumption of chloral for DDT manufacture was estimated at
25 million pounds in 1975, with an additional 500,000 pounds used in the
manufacture of other pesticides, including methoxychlor, DDVP, naled, tri-
chlorfon, and TCA (trichloroacetic acid). Mel'nikov, et al.(1975) identi-
fied chloral as an impurity in chlorofos.
Chloral is also used in the production of chloral hydrate, a thera-
peutic agent with hypnotic and sedative effects used prior to the intro-
duction of barbituates. Production of U.S.P. (pharmaceutical) grade chloral
»
hydrate was estimated to be 300,000 pounds per year in 1975 (Stanford
Research Institute, 1976).
-------
II. EXPOSURE
Boitsov, et al. (1970) noted that chloral is evolved in spraying and
pouring of polyurethane. The authors reported chloral concentrations as
high as 273.5 mg/m-5 in Soviet factories. Similar information on atmos- •
pheric occupational exposure to chloral in Western countries was not found
in the available literature.
Chloral exposure from water occurs as chloral hydrate. Keith (1976)
reported chloral hydrate concentrations ranging from 0.01 ;ug/l to 5.0 ^ig/1
in chlorinated drinking water supplies of six of ten U.S. cities studied.
The mean concentration of chloral hydrate in drinking water for the six
cities was 1.92 u'g/1.
Chloral hydrate has been used as a hypnotic and sedative agent. Alco-
hol synergistically increases the depressant effect of the compound, creat-
ing a potent depressant commonly referred to as "Mickey Finn" or "knockout
drops". Addiction to chloral hydrate through intentional abuse of the com-
pound has been reported (Goodman and Gilman, 1970).
III. PHARMACOKINETICS
A. Absorption
Specific information on the absorption of chloral was not found in
the available literature. Goodman and Gilman (1970) reported that chloral
hydrate readily penetrates diffusion barriers in the body.
B. Distribution
Specific information on the distribution of chloral was not found
in the available literature. Goodman and Gilman (1970), reporting on the
distribution of chloral hydrate from oral administration, noted its presence
in cerebrospinal fluid, milk, aoiniotic fluid, and fetal blood. The auth'ors
-------
noted that other investigators were unable to detect significant amounts of
chloral hydrate in the blood after oral administration (owing probably to
its rapid reduction).
C. Metabolism
Information on the metabolic reaction of chloral is obtained in-
directly through a metabolic study of trichloroethylene (Henschler, 1977).
The author reported that trichloroethylene oxidizes to a chlorinated epoxide
which undergoes molecular rearrangement to chloral, which is further metabo-
lized to either trichloroethanol or trichloroacetic acid. The rearrange-
ment, detected by in vivo studies, is hypothesized to occur by a catalytic
action of the trivalent iron of P-450.
Goodman and Oilman (1970) noted that chloral hydrate is reduced to
trichloroethanol in the liver and other tissues, including whole blood, with
the reaction catalyzed by alcohol dehydrogenase. Additional trichloro-
ethanol is converted to trichloroacetic acid. Chloral hydrate may be di-
rectly oxidized to trichloroacetic acid in the liver and kidney.
D. Excretion
Both chloral and chloral hydrate are metabolized to trichloro-
ethanol or trichloroacetic acid (Goodman and Gilman, 1970; Henschler,
1977). Trichloroethanol is then conjugated and excreted in the urine as a
glucuronide (urochloralic acid) or is converted to trichloroacetic acid and
slowly excreted in the urine. The glucuronide may also be concentrated and
excreted in the bile. The fraction of the total dose excreted as trichloro-
ethanol, glucuronide, and trichloroacetic acid is quite variable, indicating
other possible routes of elimination.
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IV. EFFECTS
A. Carcinogenicity
Specific information on the Carcinogenicity of chloral was not
found in the available literature. However, Keith (1976) reported skin
tumors in 4 of 20 mice dermally exposed to chloral hydrate (4 to 5 percent
solution in acetone). Further interpretation of the results and discussion
of the study methodology were not given.
8. Mutagenicity, Teratogenicity, and Other Reproductive Effects
Specific information on the mutagenicity,'teratogenicity, and re-
productive effects of chloral was not found in the available literature.
C. Chronic Effects
Rats receiving 0.1 mg/kg chloral exhibited a reduction of kidney
function and serum transaminase after seven months' exposure (Kryatov,
1970). No physiological effects were observed in rats receiving 0.01 mg/kg
chloral for periods of seven months. The route of exposure was not reported.
Chronic respiratory exposure of rats and rabbits to chloral at 0.1
mg/1 (100 mg/m-5) produced changes in central nervous system function, de-
creased antitoxic and enzyme synthesizing function of the liver, and altered
morphological characteristics of peripheral blood (Pavlova, 1975). Boitsov,
et al. (1570) reported slowed growth rate, leukocytosis, decreased albumin-
globulin ratio, and changes in arterial blood pressure and central nervous
system responses (unspecified) following prolonged respiratory exposure of
mice to chloral at 60 mg/m3.
Goodman and Gilman (1970) reported gastritis, skin eruptions, and
parenchymatous renal injury in patients suffering from chronic chloral hy-
»
drate intoxication. Habitual use of chloral hydrate may result in the
-3 IS-
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development of tolerance, physical dependence, and addiction. Death may oc-
cur either as a result of an overdose or a failure of the detoxification
mechanism due to hepatic damage.
F. Acute Toxicity
According to Hann and Jensen (1974), the human acute oral LD_0
of chloral is between 50 and 500 mg/kg.
Kryatov (1970) reported the following LD5Q values for chloral:
mice, 0.850 g/kg; rats, 0.725 g/kg; and guinea pigs, 0.940 g/kg. The routes
of exposure were not stated. Verschueren (1977) reported an oral LD^Q for
rats of 0.05 to 0.4 g/kg, while Pavlov (1975) reported an acute oral LD5Q
of 0.94 and 1.34 g/kg for mice and rats, respectively. Pavlov (1975) also
reported inhalation LC50 values of 25.5 g/m3 and 44.5 g/m3 for mice
and rats, respectively. Boitsov, et al. (1970) reported an LD5Q Of 0.710
g/kg in mice. The route of exposure was not stated. Hawley (1971) reported
that chloral is a highly toxic, strong irritant and noted ingestion or in-
halation may be fatal. Information on acute toxic effects from occupational
exposure to chloral was not found in the available literature.
G. Other Relevant Information
Verschueren (1977) reported an odor threshold concentration of
chloral in water of 0.047 ppm. The author also reported an inhibition of
cell multiplication in Pseudomonas sp. at a chloral hydrate concentration of
1.6 mg/1.
V. AQUATIC TOXICITY
A. Acute Toxicity
Verschueren (1977) reported inhibition of cell multiplication in
•
Microcystis sp. at 78 mg/1 chloral hydrate. Hann and Jensen (1974) ranked
the 96-hour TL^ aquatic toxicity of chloral in the range from 1 to 10 ppm.
-------
B. Chronic Toxicity
Information on the chronic aquatic toxicity of chloral was not
found in the available literature.
C. Plant Effects
Shimizu, et al. (1974) reported chloral inhibited the growth of
rice stems by 63.4 percent relative to controls, but slightly stimulated
root growth. The concentration of chloral in water culture was not reported.
D. Residue
Keith (1976) identified chloral hydrate In chlorinated drinking
water in six of ten cities sampled. The sample locations and concentrations
of chloral hydrate identified were: Philadelphia, PA, 5.0 pg/1; Seattle,
WA, 3.5 ijg/1; Cincinnati, OH, 2.0 jug/1; Terrebonne Parish, LA, 1.0 jug/1; New
York City, NY, 0.02 jug/1; Grand Forks, ND, 0.01 jug/1.
E. Other Relevant Information
Hann and Jensen (1974) ranked the aesthetic effect of chloral on
water as very low (zero), noting that the chemical neither pollutes waters
nor causes aesthetic problems.
VI. EXISTING GUIDELINES AND STANDARDS
Boitsov, et al. (1970) reported a maximum recommended chloral concen-
tration in workroom air of 0.22 mg/1 (220 mg/Ftv3) (USSR). Kryatov (1970)
reported a maximum recommended permissible concentration in bodies of water
as 0.2 mg/1 (USSR). Verschueren (1977) reported a maximum allowable chloral
concentration of 0.2 mg/1 in Class I waters used for drinking, but the
nation applying this standard was not identified.
X
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References
Boitsov, A.M., et al. 1970. Toxicological evaluation of chloral in the
process of its liberation during spraying and pouring of polyurethane
fosms. Gig. Tr. Prof. Zabol. 14: 26. (Chemical Abstracts CA 73:96934P).
Dilling, W.L., et al. 1976. Organic photochemistry-simulated atmopsheric
photodecomposition rates of methylene chloride, 1,1,1-trichloroethane, tri-
chloroethylene, tetrachloroethylene, and other compounds. Environ. Sci.
Techno 1. 10: 351.
Goodman, L.S. and A. Gilman. 1970. The Pharmacological Basis of Therapeu-
tics. The MacMillan Co., New York. p. 123.
Hann, R.W. and P.A. Jensen. 1974. Water Quality Characteristics of Hazard-
ous Materials. Texas A and M Univ., College Station, TX.
Haw ley, G.G. 1971. Condensed Chemical Distionary, 8th ed. Von Nostrand
Reinhold Co., New York. p. 195.
Henschler, D. 1977. Metabolism and mutagenicity of halogenated olefins - a
comparison of structure and activity. Environ. Health Perspec. 21: 61.
Keith, L.H. (ed.) 1976. Identification and Analysis of Organic Pollutants
in Water. Ann Arbor Science Publishers, Inc., Ann Arbor, Michigan, p. 351.
Kryatov, I.A. 1970. Hygienic assessment of sodium salts of p-chlorobenzene
sulfate and chloral as contaminating factors in bodies of water. Gig.
Sanit. 35: 14. (Chemical Abstracts CA 73:69048).
Luknitskii, F.I. 1975. The chemistry of chloral. Chem. Rev. 75: 259.
Mel'nikov, N.N., et al. 1975. Identification of impurities in technical
chlorofos. Khim. Sel'sk. Khoz. 13: 142. (Chemical Abstracts CA 82:165838K).
Pavlova, L.P. 1975. Toxicological characteristics of trichloroacetal-
dehyde. Tr. Azerb. Nauchno-Issled. Inst. Gig. Tr. Pro. Zabol. 10: 99.
(Chemical Abstracts CA 87:194996U).
Rook, J.J. 1974. Formation of haloforms during chlorination of natural
waters. Water Treatment Exam. 23: 234.
Shimizu, K., et al. 1974. Haloacetic acid derivatives for controlling
Grsmineae growth. Japan 7432,063 (Cl.A Oln) 27 Aug. 1974, Appl. 70 77, 535,
05 Sep. 1970 (Chemical Abstracts CA 82:81709F).
Stanford Research Institute. 1976. Chemical Economics Handbook. Stanford
Research Institute, Menlo Park, CA. p. 632.2030A.
#
Verschueren, K. 1977. Handbook of Environmental Data on Organic Chem-
icals. Von Nostrand Reinhold Co., New York. p. 170.
-------
Weast, R.C. (ed.) 1576. Handbook of Chemistry aid Physics. CRC Press,
Cleveland, OH. p. C-76.
Windholz, M., et al. 1976. The Merck Index. Merck and Co., Inc., Rahway,
N.J. p. 1,236.
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No. 35
Chlordane
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 (GAG) has evaluated
chlordane and has found sufficient evidence to indicate
that this compound is carcinogenic.
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CHLORDANE
Summary
Chlordane is an organochlorlnated cyclodiene insecticide commonly used
as a formulation consisting of 24% trans-, 19% cis-chlordane, 10% hepta-
chlor, 21,5% chlordenes, 7% nonachlor, and 18.5% of other organochlorinated
material. Since heptachlor is also an insecticide and is more toxic than
chlordane, technical chlordane is generally more toxic than pure chlordane.
Pure chlordane, which is a cis/trans mixture of isomers, induces liver
cancer in mice and is mutagenic in some assays. Chlordane has not been shown
to be teratogenic. Little information is available on chronic mammalian
toxicity. Repeated doses of chlordane produced alterations in brain poten-
tials and changes in some blood parameters. Chlordane is a convulsant.
Chlordane and its toxic metabolite oxychlordane accumulate in adipose tissue.
Ten species of freshwater fish have reported 96-hr LC50 values rang-
ing from 8 to 1160 pg/1. Freshwater invertebrates appear to be more resis-
tant to chlordane, with observed 96-hr LC,-n values ranging from 4 to 40
jjg/1. Five species of saltwater fish have LC^ values of 5.5 to 160 jjg/1,
and marine invertebrate LC5Q values range between 0.4 and 480 pg/1.
Chronic studies involving the bluegill Oaphnia maqna gave an LC^ of 1.6
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CHLORDANE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Chlordane (U.S. EPA, 1979).
Chlordane is a broad spectrum insecticide of the group of organochlori-
nated polycyclic hydrocarbons called cyclodiene insecticides. Chlordane has
been used extensively over the past 30 years for termite control in homes
and gardens, and as a control for soil insects.
Pure Chlordane (1,2,4,5,6,7,8,8-octachloro-2,3,3a,4,7,7a-hexahydro-4,7-
methanoindene) is a pale yellow liquid having the empirical formula C,_-
H^Clg and a "molecular weight of 409.8. It is composed of a mixture of
stereoisomers, with the cis- and trans- forms predominating, commonly refer-
red to as alpha- and gamma-isomers, respectively.) The solubility of pure
Chlordane in water is approximately 9 ;jg/l at 25°C (U.S. EPA, 1979).
Technical grade Chlordane is a mixture of chlorinated hydrocarbons with
a typical composition of approximately 24 percent trans(gamma)-chlordane, 19
percent cis(alpha)-chlordane, 10 percent heptachlor (another insecticidal
ingredient), 21.5 percent chlordene isomers, 7 percent nonachlor, and 18.5
percent closely related chlorinated hydrocarbon compounds. Technical chlor-
dane is a viscous, amber-colored liquid with a cedar-like odor. It has a
vapor pressure of 1 x 10~5 mm Hg -at 25°C. The solubility of technical
Chlordane in water is 150 to 220 jjg/1 at 22°C (U.S. EPA, 1979).
Production of Chlordane was 10,000 metric tons in 1974 (41 FR 7559;
February 19, 1976). Both uses and production volume have declined exten-
sively since the issuance of a registration suspension notice by the U.S.
EPA (40 FR34456; December 24, 1975) for all food, crop, home, and 'garden
-V04-
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uses of chlordane. However, use of chlordane for termite control and limit-
ed usage (through 1980) as an agricultural insecticide are still permitted
(43 FR 12372; March, 1978).
Chlordane persists for prolonged periods in the environment (U.S. EPA,
1979), Photo-cis-chlordane can be produced in water and on plant surfaces
by the action of sunlight (Benson, et al. 1971) and has been found to be
twice as toxic as chlordane to fish and mammals (Ivie, et al. 1972; Podow-
ski, et al. 1979). Photo-cis-chlordane (5 ng/1) is accumulated more (ca.
20%) by goldfish (Carassius auratus) than chlordane (5 ng/1) itself (Ducat
and Khan, 1979).
Air transport of chlordane has been hypothesized to account for resi-
dues in Sweden (Jansson, et al. 1979). Residues in agricultural soils may
be as high as 195 ng/g dry weight of soil (Requejo, et al. 1979).
II EXPOSURE
A. Water
Chlordane has been detected in finished waters at a maximum concen-
tration of 8 jjg/1 (Schafer, et al. 1969) and in rainwater (Bevenue, et al.
1972; U.S. EPA, 1976). There have been reports of individual household
wells becoming contaminated after a house is treated with chlordane for ter-
mite control (U.S. EPA, 1979). A recent contamination of a municipal water
system has been discussed by Harrington, et al. (1978). Chlordane has also
been detected in rainwater (U.S. EPA, 1976).
B. Food
Chlordane has been found infrequently in food supplies since 1965,
when the FDA began systematic monitoring for chlordane (Nisbet, 1976). The
only quantifiable sample collected was 0.059 mg chlordane/kg measured in a
sample of grain in 1972 (Manske and Johnson, 1975). In the most recently
-------
published results (for 1975), chlordane was not detected (Johnson and
Manske, 1977). Fish are thought to represent the most significant dietary
exposure. The average daily uptake from fish is estimated at 1 jjg (Nisbet,
1976).
The U.S. EPA (1979) has estimated the weighted average bioconcen-
tration factor for chlordane to be 5,500 for the edible portions of fish and
shellfish consumed by Americans. This estimate was based on measured steady-
state bioconcentration studies in the sheepshead minnow (Cyprinodon variega-
tus).
Eighty-seven percent of 200 samples of milk collected in Illinois
from 1971 to '1973 were positive for chlordane. The average concentration
was 50 ^ig/1 (Moore, 1975 as reviewed by Nat. Acad. Sci., 1977). Cyclo-
dienes, such as chlordane, apparently are ingested with forage and tend to
concentrate in lipids. Oxychlordane, a metabolite of chlordane and hepta-
chlor, was found in 46 percent of 57 human milk samples collected during
1973-74 in Arkansas and Mississippi. The mean value was 5 jjg/1, and the
maximum was 20 jjg/1 (Strassman and Kutz, 1977).
C. Inhalation
In a survey of the extent of atmospheric contamination by pesti-
cides, air was sampled .at nine localities representative of both urban and
agricultural areas. Chlordane was not detected in any samples (Stanley, et
al. 1971). In a larger survey, 2,479 samples were collected at 45 sites in
16 states. Chlordane was detected in only two samples, with concentrations
of 84 and 204 ng/m (Nisbet, 1976). The vapor concentrations to which
spray operators are exposed have not been estimated.
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0. Dermal Effects
Chlordane can be absorbed through the skin to produce toxic effects
(Gosselin, et al. 1976). Spray operators, chlordane formulators and farmers
may be exposed. Chlordane has been known to persist for as long as two
years on the hands (Kazen, et al. 1974). Dermal LD5Q values in rats range
from 530 to 700 mg/kg (U.S. EPA, 1979).
III. PHARMACOKINETICS
A. Absorption
Gastrointestinal absorption of chlordane- in rats ranged from 6 per-
cent with a single dose to 10-15 percent with smaller daily doses (Barnett
and Dorough, 1974).
8. Distribution
In a study of the distribution of chlordane and its metabolites
using radioactive carbon, the levels of residues in the tissues were low,
except in the fat (Barnett and Dorough, 1974). Rats were fed 1, 5, and 25
mg chlordane/g in food for 56 days. Concentrations of chlordane residues in
fat, liver, kidney, brain, and muscle were 300, 12, 10, 4, and 2 percent,
respectively, of the concentration in the diet. All residues declined
steadily for 4 weeks, at which time concentrations were reduced about 60
percent. During the next four weeks, residues declined only slightly.
C. Metabolism
Mammals metabolize chlordane to oxychlordane, via 1,2-dichloro-"
chlordene which is about twenty times more toxic than the parent compound
and persists in adipose tissue (Polen, et al. 1971; Tashiro and Matsumura,
1978; Street and Blau, 1972). Oxychlordane can degrade to form l-hydroxy-2-
»
cyclochlordenes, and l-hydroxy-2-chloro-2,3-epoxy-chlordenes (Tashiro and
Matsumura, 1978). In general, the metabolism of chlordane takes place via a
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series of oxidative enzyme reactions. None of the metabolic intermediates
(except for oxychlordane) and end products are more toxic than chlordane
(Barnett and Dorough, 1974; Tashiro and Matsumura, 1977; Mastri, et al.
1969). Trans-nonachlor, a major impurity in technical chlordene, is con-
verted to trans-chlordane in rats, but this is not important in humans.
This explains the fact that trans-nonachlor accumulates in humans but not in
rats (Tashiro and Matsumura, 1978). A very small amount of cis- or trans-
chlordane can be converted to heptachlor in rat liver (Tashiro and Matsu-
mura, 1977).
D. Excretion
Chlordane is primarily excreted in the feces of rats, only about
six percent of the total intake being eliminated In the urine. Urinary ex-
cretion of chlordane in rabbits is greater than excretion in the feces (Nye
and Dorough, 1976).
The half-life of chlordane in a young boy was reported to be ap-
proximately 21 days (Curley and Garrettson, 1969), while for rats it was 23
days (Barnett and Oorough, 1974). The half-life of chlordane in the serum
of a young girl was 88 days (Aldrich and Holmes, 1969).
IV. EFFECTS
A. Carcinogenicity
Hepatocellular carcinomas were induced in both sexes of two strains
of mice fed pure (95%) chlordane (56.2 mg/kg) in the diet for 80 weeks (Na-
tional Cancer Institute, 1977; Epstein, 1976). In contrast to findings with
mice, a significantly increased incidence of hepatocellular carcinomas did
not appear in rats administered chlordane. Dosages were near the maximum
permissible (National Cancer Institute, 1977).
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B. Mutagenicity
Pure or technical chlordane Induced unscheduled DMA synthesis in
the SV-40 transformed human fibroblast cell line VA-4. Metabolic activation
eliminated this effect (Ahmed, et al. 1977). Chlordane did not induce muta-
tions in the dominant lethal assay in mice (Arnold, et al. 1977).
While neither pure cis-chlordane nor pure trans-chlordane was muta-
genic in the Ames Salmonella microsome assay, technical grade chlordane was
mutagenic. Microsomal activation did not enhance the mutagenic activity
(Simmon, et al. 1977).
C. Teratogenicity
Chlordane was found not to be teratogenic in rats when fed at con-
centrations of 150 to 300 mg/kg during gestation (Ingle, 1952).
D. Other Reproductive Effects
Pertinent data could not be located in the available literature.
E. Chronic Toxicity
There appears to be little information on chronic mammalian toxi-
city. ^aily injections of 0.15 to 25 mg chlordane/kg in adult rats resulted
in dose-dependent alterations of brain potentials (Hyde and Falkenberg,
1976). As changes were directly related to length of exposure, it was con-
cluded that chlordane may be a cumulative neurotoxin. Length of exposure
was not specified. Repeated doses of chlordane given to gerbils produced
changes in serum proteins, blood glucose, and alkaline and acid phosphatase
activities (Karel and Saxena, 1976). Again, duration of treatment was not
specified.
F. Other Relevant Information
Carbon tetrachloride produced more extensive hepatocellular necro-
sis in chlordane-pretreated rats than in rats which were not pretreated
(Stenger, et al. 1975). Rats suffered greater cirrhosis when chlordane (50
-------
jjg/kg/day) exposure for ten weeks followed prior exposures of ten weeks for
carbon tetrachloride above (110 mg/1) or with chlordane (Mahon and Oloffs,
1979). Quail treated with chlordane followed by endrin had considerably
more chlordane residues in their brains than did quail treated with chlor-
dane alone (Ludke, .1976). Quail pretreated with 10 mg/kg chlordane exhibit-
ed decreased susceptibility to parathion (Ludke, 1977). Chlordane is a con-
vulsant and emetic. It induces twitching, seizures and electroencephalo-
graphic dysrhythmia in humans. Acute symptoms can be alleviated with pheno-
barbital. Acute oral LD5Q values for the rat range from 100 to 112 mg/kg
(U.S. EPA, 1979). The no observable effect level was found to be 2.5 mg/kg/
day over 15 days (Natl. Acad. Sci., 1977).
Chlordane inhibits growth of human viridans streptococci of the
buccal cavity. Complete inhibition of growth occurred at 3 ppm, and about
20 percent inhibition was seen at 1 ppm (Goes, et al. 1978).
V. AQUATIC TOXICITY
A. Acute Toxicity
Ten species of freshwater fish have reported 96-hr. LC_n values
ranging from 8 to 1160 jjg/1 resulting from technical and pure chlordane
exposure with a geometric mean of 16 pg/1. Rainbow trout, Salmo qairdneri
(Mehrle, et al. 1974) 'was the most sensitive species tested, the channel
catfish (Ictalurus punctatus) the least sensitive. The freshwater inverte-
brates were more sensitive to chlordane, with a reported LC5Q value rang-
ing from 4.0 for freshwater shrimp Palaemonetes kadiakensis (Sanders, 1972)
to 40 jjg/1 (Gammarus fasciatus), with a geometric mean of 0.36 ug/1. In
goldfish (Carassius auratus), only 0.13 percent of cis-chlordane is metabo-
lized in ,24 hours. Only 0.61 percent is converted after 25 days. Some
metabolites were chlordene chlorohydrin and monohydroxy derivatives (Feroz
and Khan, 1979).
/
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The LCcn's for four species of saltwater fish, sheepshead minnows
(Cyprinodon verieqatus), striped bass (Morone saxatilis), pinfish (Lagodon
rhomboides), and white mullet (Mugil curema), ranged from 5.5 to 24.5 jug/1.
The three-spine stickleback (Gasterosteus aculeatus) yielded 96-hr LC5Q
values which ranged from 90-160 pg/1 (Katz, 1961). Invertebrate LC5Q val-
ues ranged from 0.4 for the pink shrimp, Penaeus duorarum (Parrish, et al.
1976) to 480 jug/1. The geometric mean of the adjusted LC50 values for in-
vertebrates was 0.18jug/l (U.S. EPA, 1979).
B. Chronic Toxicity
In a life cycle bioassay involving freshwater organisms, the chron-
ic values for" the bluegill Lepomis macrochirus (Cardwell, et al. 1977) was
1.6 jjg/1. In two tests involving the sheepshead minnow, Cyprinodon variega-
tus, the chronic values were 0.63 jjg/1 for the life cycle test (Parrish, et
al. 1978) and 5.49jjg/l for an embryo-level test (Parrish, et al. 1976).
Many blood parameters (clotting time, mean corpuscular hemoglobin
and cholesterol level) are lowered after the teleost, Sacco-branchus fossil-
us, is exposed to 120 jjg/1 of chlordane for 15 to 60 days (Verna, et al.
1979). Similar results were obtained in Labeo rohita at doses — 23 jjg/1
after 30 to 60 day exposures (Bansal, et al. 1979).
C. Plant Effects '
A natural saltwater phytoplankton community suffered a 94 percent
decrease in productivity during a 4-hour exposure at 1,000 pg/1 (Butler,
1963).
0. Residues
In Daphnia magna, chlordane was bioconcentrated 6,000-fold after
»
seven days' exposure and 7,400-fold by scuds (Hyallela azteca) after 65 days
of exposure (Cardwell, et al. 1977). After 33 days' exposure, the fresh-
-HH-
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water alga (Oedegonium sp.) bioconcentrated chlordane 98,000-fold; Physa
sp., a snail, concentrated it 133,000-fold (Sanborn, et al. 1976). Equili-
brium bioconcentration factors for the sheepshead minnow ranged from 6,580
to 16,035 (Goodman, et al. 1978; Parrish, et al. 1976).
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
The issue of the carcinogenicity of chlordane in humans is being
reconsidered; thus, there is a possibility that the criterion for human
health will be changed. Based on the data for qarcinogenicity in mice (Ep-
stein, 1976), and using the "one-hit" model, the U.S. EPA (1979) has esti-
mated levels of chlordane 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 ID-6 ICf5
2 liters of drinking water 0 0.012 ng/1 0.12 ng/1 1.2 ng/1
and consumption of 18.7
grams fish and shellfish.
Consumption of fish and 0 0.013 ng/1 0.13 ng/1 1.3 ng/1
shellfish only.
The ACGIH (1977) adopted a time-weighted average value of 0.5
mg/m for chlordane, with a short-term exposure limit (15 minutes) of 2
mg/m .
A limit of 3 jjg/1 for chlordane in drinking water is suggested
under the proposed Interim Primary Drinking Water Standards (40 FR 11990,
March 14, 1975).
Canadian Drinking Water Standards (Dept. Natl. Health Welfare,
t
1968) limit chlordane to 3 jug/1 in raw water supplies.
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B. Aquatic
For chlordane, the proposed criterion to protect freshwater aquatic
life is 0.024 jug/1 for a 24-hour average, not to exceed 0.36 jug/1 at any
time (U.S. EPA, 1979). For saltwater aquatic species, the draft criterion
is 0.0091 ^jg/1 for a 24-hour average, not to exceed 0.18 jjg/1 at any time
(U.S. EPA, 1979).
Mf
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CHLORDANE
REFERENCES
Ahmed, F.E., et al. 1977. Pesticide induced DNA damage
and its repair in cultured human cells. Mutat. Res. 42:
161.
Aldrich, F.D. , and J.H. Holmes. 1969. Acute chlordane
intoxication in a child. Arch. Environ. Health 19: 129.
ACGIH. 1977. TLVs thresholds limit values for chemical
substances in workroom air adopted by the American Confer-
ence of Governmental Industrial Hygienists for 1977. Cincin-
nati, Ohio.
Arnold, D.W., et al. 1977. Dominant lethal studies with
technical chlordane, HCS-3260, and heptachlor: heptachlor
epoxide. Jour. Toxicol. Environ. Health 2: 547.
Bansal, S.K., et al. 1979. Physiological dysfunction of
the haemopoletic system in a freshwater teleost, Rabeo ro-
hitaf following chronic chlordane exposure. Part 1. Altera-
tions in certain haemotological parameters. Bull. Environ.
Contain. Toxicol. 22: 666.
Barnett, J.R., and H.W. Dorough. 1974. Metabolism of chlor-
dane in rats. Jour. Agric. Food Chem. 22: 612.
Benson, W.R., et al. 1971. Chlordane photoalteration pro-
ducts: Their preparation and identification. Jour. Agric.
Food Chem. 19: 857.
Bevenue, A., et al. 1972. Organochlorine pesticides in-
rainwater Oahu, Hawaii, 1971-72, Bull. Environ. Contam.
Toxicol. 8: 238.
Butler, P.A., et al. 1963. Effects of pesticides on oy-
sters. Proc. Shell Fish. Assoc. 51: 23.
Cardwell, R.D., et al. 1977. Acute and chronic toxicity
of chlordane to fish and invertebrates. EPA Ecol. Res.
Ser., U.S. Environ. Prot. Agency, Duluth, Minn.
Curley, A., and L.K. Garrettson. 1969. Acute chlordane
poisoning. Arch. Environ. Health 18: 211.
Department of National Health and Welfare. 1968. Canadian
drinking water standards and objectives. Ottawa, Canada.
Ducat, D.A- and M.A.Q. Khan. 1974. Absorption and elimina-
tion of C-cis-chlordane and C-photo-cis-chlordane by
goldfish, Carassius auratus. Arch. Enviorn. Contam. 8: 409.
-------
Epstein, S.S. 1976. Carcinogenicity of heptachlor and
chlordane. Sci. Total Environ. 6: 103.
Feroz, M. , and M.A.Q. Khan. 1979. Fate of 14C-cis-chlor-
dane in goldfish, Carassius auratus. Bull. Enviorn. Contam.
Toxicol. 23: 64.
Goes, T.R., et al. 1978. In vitro inhibition of oral Viri-
dous streptococei by chlordane. Arch. Environ. Contam.
Toxicol. 7: 449.
Goodman, L., et al. 1978. Effects of heptachlor and toxa-
phene on laboratory-reared embryos and fry of the sheepshead
minnow, Proc. 30th Annu. Conf. S.E. Assoc. Game Fish Comm.
Gosselin, R.E., et al. 1976. Clinical toxicology of commer-
cial products. 4th ed. Williams and Wilk-ins Co., Baltimore,
Md.
Harrington, J.M., et al. 1978. Chlordane contamination
of a municipal water system. Environ. Res. 15: 155.
Hyde, K.M., and R.L. Falkenberg. 1976. Neuroelectrical
disturbance as indicator of chronic chlordane toxicity.
Toxicol. Appl. Pharmacol. 37: 499,
Ingle, L. 1952. Chronic oral toxicity of chlordane to
rats. Arch. Ind. Hyg. Occup. Med. 6: 357.
Ivie, G.W., et al. 1972. Novel photoproducts of heptachlor
epoxide, trans-chlordane and trans-nonachlor. Bull, Environ.
Contam. Toxicol. 7: 376.
Jansson, B. , et al. 1979. Chlorinated terpenes and chlor-
dane components found in fish, guilleiuot and seal from
Swedish waters. Chemosphere 8: 181.
Johnson, R.D., and D.D. Manske. 1977. Pesticide and other
chemical r-esidues in total diet samples (XI). Pestic. Monitor.
Jour. 11: 116.
Karel, A.K., and S.C. Saxena. 1976. Chronic chlordane
toxicity: effect on blood biochemistry of Meriones hurrianae
Jerdon, the Indian desert gerbil. Pestic. Biocfiem. Physiol7
6: 111.
Katz, 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.
Kazen C. , et al. 1974. Persistence of pesticides on the
hands of some occupationally exposed people. Arch. Environ.
Health 29: 315.
-------
Ludke, J.L. 1976. Organochlorine pesticide residues associ-
ated with mortality: additivity of chlordane and endrin.
Bull. Environ. Contam. Toxicol, 16: 253.
Ludke, J.L. 1977. DDE increases the toxicity of parathion
to coturnix quail. Pestic. Biochem. Physiol. 7: 28.
Mahon, D.C., and P.C. Oloffs. 1979. Effects of subchronic
low-level dietary intake of chlordane on rats with cirrhosis
of the liver. Jour. Environ. Sci. Health B14: 227.
Manske, D.D., and R.D. Johnson. 1975. Pesticide residues
in total diet samples (VIII). Pestic. Monitor. Jour. 9: 94.
Mastri, C., et al. 1969. Unpublished data. In 1970 evalua-
tion of some pesticide residues in food. Foo~d~Agric. Org.
United Nations/World Health Org.
Mehrle, P.M., et al. 1974. Nutritional effects on chlor-
dane toxicity in rainbow trout. Bull. Enviorn. Contam.
Toxicol. 2:' 513
Moore, S., III. 1975. Proc. 27th Illinois Custom Spray
Operators Training School. Urbana.
National Academy Science. 1977. Drinking water and health.
Washington, D.C.
National Cancer Institute. 1977. Bioassay of chlordane
for possible carcinogenicity. NCI-CG-TR-8.
Nisbet, I.C.T. 1976. Human exposure "to chlordane, hepta-
chlor, and their metabolites. Contract WA-7-1319-A. U.S.
Environ. Prot. Agency.
Nye, D.E., and H.W. Dorough. 1976. Fate of insecticides
administered endotracheally to rats. Bull. Environ. Contam.
Toxicol. 15: 291.
Parrish, P.R., et al. 1976. . Chlordane:" effects on several
estuarine organisms. Jour. Toxicol. Environ. Health 1:
485.
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trifluralin and pentachlorophenol to sheepshead minnows
(Cyprinodon variegatus). EPA 600/3-78-010: 1. U.S. Environ.
Prot. Agency.
Podowski, A.A., et al. 1979. Photolysis of heptachlor
and cis-chlordane and toxicity of their photoisomers t°
animals. Arch. Environ. Contam. Toxicol. 8: 509.
Polen, P.B., et al. 1971. Characterization of oxychlor-
dane, animal metabolites of chlordane. Bull. Enviorn. Contam.
Toxicol. 5: 521.
-------
Requejo, A.G., et al. 1979. Polychlorinated biphenyls
and chlorinated pesticides in soils of the Everglades National
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Sanborn, J.R., et al. 1976. The fate of chlordane and
toxaphene in a terrestrial-aquatic model ecosystem. Environ.
Entomol. 5: 533.
Sanders, H.O. 1972. Toxicity of some insecticides to four
species of malacostracan crustaceans. U.S. Dept. Interior.
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Environ. Mutagens, Edinburgh, Scotland, July 1977.
Stanley, C.W., et al. 1971. Measurement of atmospheric
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Stenger, R.J., et al. 1975. Effects of chlordane pretreat-
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Mol. Pathol. 23: 144.
Strassman, S.C., and F.W. Kutz. 1977. Insecticide residues
in human milk from Arkansas and Mississippi, 1973-74. Pestic.
Monitor. Jour. 10: 130.
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lation in rat adipose tissue on feeding chlordane isomers
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Tashiro, S., and F. Matsumura. 1977. Metabolic routes
of cis- and trans-chlordane in rats. Jour. Agric. Food
Chem. 25: 872.
Tashiro, 3., and F. Matsumura. 1978. Metabolism of trans-
nonachlor and related chlordane components in rat and man.
Arch. Enviorn. Contam. Toxicol. 7: 413.
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Bull. Environ. Contam. Toxicol. 22: 467.
-HI7-
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No. 36
Chlorinated Benzenes
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 ac-c-uracy.
-------
CHLORINATED BENZENES
Summary
The chlorinated benzenes are a group of compounds with
a wide variety of physical and chemical characteristics
depending on the degree of chlocination. As chlorination
increases, the persistence of the compound in the environ-
ment increases. On chronic exposure liver and kidney changes
are noted, while the degree of toxicity increases with the
degree of chlorination. The chlorinated benzenes have not
been shown to be teratogens or mutagens. Only hexachloro-
benzene has been demonstrated to be carcinogenic in labora-
tory animals.
Aquatic toxicity data indicate a trend to increasing
toxicity with increasing chlorination for all species tested.
The bluegill for example, has the following 96-hour LC
values; chlorobenzene, 15,900 jjg/1; 1,2,4-trichlorobenzene
3,360 jig; 1,2,3,5-tetrachlorobenzene, 6,420 micrograms/1;
1,2,4,5-tetrachlorbenzene 1,550 pg/1 and pentachlorobenzene,
200 pg/1. Other freshwater and saltwater fish, invertebrates
and plants were generally less sensitive to chlorobenzenes
toxicity than the bluegill. The sheepshead minnow yielded
a chronic value of 14.5 jjg/1 for 1,2,4,5-tetrachlorobenzene
in an embryo-level test. After 28 days exposure, the biocon-
centration factor for the bluegill for pentachlorobenzene
and 1,2,4,5-tetrachlorobenzene were 3,400 and 1,800, respec-
tively.
i
-------
CHLORINATED BENZENES
I. INTRODUCTION
This profile is based on the Ambient Water Quality
Criteria Document for Chlorinated Benzenes (U.S. EPA, 1979}.
This document will summarize the general properties of the
chlorinated benzenes. For further information on monochloro-
benzene, 1,2,4-trichlorobenzene, or hexachlorobenzene, refer
to the specific EPA/ECAO Hazard Profiles for these compounds.
For detailed information on the other chlorinated benzenes
refer to the Ambient Water Quality Document (U.S. EPA, 1979).
The chlorinated benzenes, excluding dichlorobenzenes,
are monochlorobenzene (CgH-jCl) , 1, 2 , 4-tr ichlorobenzene (CgH3C
1,3,5-trichlorobenzene (C,-H-,C1-1) , 1, 2,3 ,4-tetrachlorobenzene
o j j
(CgH2Cl4), 1,2,3,5-tetrachlorobenzene (CgH2Cl4), 1,2,4,5-
Tetrachlorobenzene (CgH2Cl4), pentachlorobenzene (CgHCl^),
and hexachlorobenzene (CgClg). All chlorinated benzenes
are colorless liquids or solids with a pleasant aroma.
The most important properties imparted by chlorine to these
compounds are solvent power, viscosity, and moderate chemi-
cal reactivity. Viscosity and nonflammability tend to in-
crease from chlorobenzene to the more highly chlorinated
benzenes. Vapor pressures and water solubility decrease
progressively with the degree of chlorination (U.S. EPA,
1979).
The current production, based on annual production
r
in the U.S., was 139,105 kkg of monochlorobenzene in 1975,
* t
12,849 kkg of 1,2,4-trichlorobenzene, 8,182 kkg of 1,2,4,5-
',*, * "'
-------
tetrachlorobenzene and 318 kkg of hexachlorobenzene in 1973
(West and Ware, 1977; EPA, 1975a). The remaining chlori-
nated benzenes are produced mainly as by-products from the
production processes for the above four chemicals. Chlori-
nated benzenes have many and diverse uses in industry depend-
ing upon the individual properties of the specific compound.
Some uses are as solvents, chemical intermediates, flame
retardants, and plasticizers.
II. EXPOSURE
A. Water
Mono-, tri-, and hexachlorobenzene have been de-
tected in ambient water. Because of its high volatility,
monochlorobenzene has a short half-life of only 5.8 hours
(Mackay and Leinonen, 1975). However, hexachlorobenzene
has an extremely long residue time in water, appearing to
be ubiquitous in the aqueous environment. Monochloroben-
zene has been detected in "uncontaminated" water at levels
of 4.7 ^ig/1. Both trichlorobenzene and hexachlorobenzene
have been detected in drinking waters at concentrations
of 1.0 pg/1 and 4 to 6 ng/1 respectively {U.S. EPA, 1979).
There is no information available on the concentration of
the other chlorinated benzenes in water.
B. Food
There is little data on the consumption of chlorin-
ated benzenes in food. All the chlorinated benzenes appear
*•
to concentrate in fat, and are capable of being absorbed
-------
by the plants from contaminated soil. Both pentachloroben-
zene and hexachlorobenzene have been detected in meat fat
(e.g. Stijve, 1971; Ushio and Doguchi, 1977). Hexachloro-
benzene, the most extensively studied compound, has been
found in a wide variety of foods from cereals to milk (includ-
ing human breast milk), eggs, and meat. The U.S. EPA (1979)
has estimated the weighted bioconcentration factor of the
following chlorinated benzenes:
Weighted
Chemical bioconcentration factor
monochlorobenzene 13
1,2,4-trichlorobenzene 290
1,2,4,5-tetrachlorobenzene 1,000
pentachlorobenzene 7,800
hexachlorobenzene 12,000
These estimates were based on the octanol/water parti-
tion coefficient of the chlorinated benzenes.
C. Inhalation
There is no available data on the concentration
of chlorinated benzenes in ambient air with the exception
of measurements of aerial fallout of particulate bound 1,2,4-
trichlorobenzene in southern California. Five sampling
sites showed median levels of 1,2,4-trichlorobenzene of
less than 11 ng/m2/day (U.S. EPA, 1979). The primary site
of inhalation exposure to chlorinated benzenes is the work-
place in industries utilizing and/or producing these compounds.
III. PHARMACOKINETICS
A. Absorption
There is little data on the absorption of orally
administered chlorinated benzenes. It is apparent from
-------
the toxicity of orally administered compounds that absorp-
tion does take place, and tetrachlorobenzene has been shown
to be absorbed relatively efficiently by rabbits (Jondorf,
et al. 1958). Pentachlorobenzene was absorbed poorly after
subcutaneous injection {Parke and Williams, I960}. Hexa-
chlorobenzene was absorbed poorly from an orally administered
aqueous solution {Koss and Kornasky, 1975}, but with high
efficiency when administered in oil (Albro and Thomas, 1974).
The more highly chlorinated compounds in food products will
selectively partition into the lipid portion and be absorbed
far better than that in an aqueous medium (U.S. EPA, 1979).
A. Distribution
The chlorinated benzenes are lipophilic, compounds
with greater lipophilic tendencies in the more highly chlor-
inated compounds. The predominant disposition site is either
suspected to be, or shown to be, in the lipid tissues of
the body (Lee and Metcalf, 1975; U.S. EPA, 1979).
C. Metabolism
The chlorinated benzenes are metabolized in the
liver by the NADPH-cytochrome P-448 dependent microsomal
enzyme system {Ariyoshi, et al. 1975; Koss, et al. 1976).
At least for monochlorobenzene, there is evidence that toxic
intermediates are formed during metabolism (Kohli, et al.
1976}. Various conjugates and phenolic derivatives are
the primary excretory end products of chlorinated benzene
metabolism. In the more highly chlorinated compounds, such
as hexachlorobenzene, conjugates are formed to only a limited
extent, and metabolism is relatively slow.
-------
D. Excretion
The less-chlorinated benzenes are excreted as
polar metabolites or conjugates in the urine. An exception
occurs with monochlorobenzene where 27 percent of an admin-
istered dose appeared as unchanged compounds in the expired
air of a rabbit (Williams, 1959). The two highly chlorinated
compounds, pentachlorobenzene and hexachlorobenzene, are
eliminated predominately by fecal excretion as unchanged
compounds (Koss and Koransky, 1975; Rozman, et al. press) .
The biological half-lives of these two compounds are extremely
long in comparison to that of the less-chlorinated compounds
(U.S. EPA, 1979).
IV. EFFECTS
A. Carcinogencity
Mono- and tetrachlorobenzene have not been in-
vestigated for carcinogenic potential (U.S. EPA, 1979) .
In one study, trichlorobenzene was not shown to produce
any significant increase in liver tumors (Gotto, et al.
1972). There is one report, which was not critically evalu-
ated by U.S. EPA (1979) , which alludes to the carcinogencity
of pentachlorobenzene in mice and the absence of this activity
in rats and dogs (Preussman, 1975) . Life-time feeding studies
in hamsters (Cabral, et al. 1977) and mice (Cabral, et al.
1978) have demonstrated the carcinogenic activity of hexa-
chlorobenzene. However, shorter term studies failed to
»
demonstrate an increasd tumor incidence in strain A mice
or ICR mice (Theiss, et al. 1977; Shirai, et al. 1978).
-S25"-
-------
B. Mutagenicity
There are no available studies conducted to evalu-
ate the mutagenic potential of mono-, tri-, tetra-, and
pentachlorobenzene (U.S. EPA, 1979). Hexachlorobenzene
was assayed for mutagenic activity in the dominant lethal
assay, and shown to be inactive (Khera, 1974).
C. Teratogenicity
There are no available studies conducted to evalu-
ate the teratogenic potential of mono-, tri-, tetra-, and
pentachlorobenzene (U.S. EPA, 1979). Khera (1974) concluded
hexachlorobenzene was not a teratogen when given to CD-I
mice at 50 mg/kg/day on gestation days from 7 to 11.
D. Other Reproductive Effects
Hexachlorobenzene can pass through the placenta
and cause fetal toxicity in rats (Grant, et al. 1977).
The distribution of hexachlorobenzene in the fetus appears
to be the same in the adult, with the highest concentration
in fatty tissue.
E. Chronic Toxicity
There is no available data on the chronic effects
of pentachlorobenzene (U.S. EPA, 1979). Mono- and trichloro-
benzene produce histological changes in the liver and kidney
(Irish, 1963; Coate, et al. 1977). There is also some evi-
dence for liver damage occurring with prolonged exposure
of rats and dogs to tetrachlorobenzene (Fomenko, 1965; Braun,
et al. 1978). Hexachlorobenzene has caused histological
changes in the livers of rats (Koss, et al. 1978). In humans
-------
exposed to undefined amounts of hexachlorobenzene for an
undetermined time, porphyrinuria has been shown to occur
(Cam and Nigogosyan, 1963).
F. Other Relevant Information
Chlorinated benzenes appear to increase the activity
of microsomal NADPH-cytochrome P-450 dependent enzyme systems.
Induction of microsomal enzyme activity has been shown to
enhance the metabolism of a wide variety of drugs, pesticides
and other xenobiotics (U.S. EPA, 1979).
V. AQUATIC TOXICITY
A. Acute Toxicity
The dichlorobenzenes are covered in a separate
EPA/ECAO hazard profile and will not be covered in this
discussion on chlorobenzenes.
All data reported for freshwater fish are from
96-hour static toxicity tests. Pickering and Henderson
(1966) reported 96-hour LC values for goldfish, guppys
and bluegills to be 51,620, 45,530, and 24,000 pg/1, respec-
tively, for chlorobenzene. Two 96 hour LCcg values for
chlorobenzene and fathead minnows are 33,930 ;jg/l in salt-
water and 29,120 pg/1 in hard water. Reported 96-hour values
for the bluegill exposed to chlorobenzene, 1,2,4-trichloro-
benzene, 1,2,3,5-tetrachlorobenzene, 1,2,4,5-tetrachloro-
benzene and pentachlorobenzene are 15,900, 3,360, 6,420,
1,550 and 250 ug/1, respectively (U.S. EPA, 1978). These
data indicate a trend to increasing toxicity with chlorina-
tion, except for 1,2,3,5-tetrachlorobenzene (U.S. EPA, 1978).
-------
EC50 (48 hour) values reported for Daphnia magna are: chloro-
benzene 86,000 pg/1, 1,2,4-trichlorobenzene 50,200 ug/1,
1,2,3,5-tetrachlorobenzene 9,710 pg/1, and pentachlorobenzene
5,280 pg/1 (U.S. EPA, 1978).
Toxicity tests with the sheepshead minnow, Cypri-
nodon variegatus, performed with five chlorinated benzenes
under static conditions and yielded the following 96-hour
LC50 values: chlorobenzene 10,500 pg/1, 1,2,4-trichloroben-
zene 21,400 pg/1, 1,2,3,5-tetrachlorobenzene 3,670 pg/1,
1,2,4,5 tetrachlorobenzene 840 pg/1, and pentrachlorobenzene
835 pg/1 (U.S. EPA, 1978). As with sheepshead minnows,
sensitivity of the mysid shrimp, Mysidopsis bahia, to chlori-
nated benzenes generally increases with increasing chlorina-
tion. The reported 96-hour LC5Q values are as follows:
chlorobenzene 16,400 pg/1, 1,2,4-trichlorobenzene 450 pg/1,
1,2,3,5-tetrachlorobenzene 340 pg/1, 1,2,4,5-tetrachloro-
benzene 1,480 pg/1, and 160 pg/1 for pentachlorobenzene
(U.S.' EPA, 1979} .
B. Chronic Toxicity
Chronic toxicity data are not available for fresh-
water fish or invertebrate species. Only one saltwater
species, Cyprinodon veriegatus, has been chronically exposed
to any of the chlorinated benzenes. In an embryo-level
test, the limits for 1,2,4,5-tetrachlorobenzene are 92 to
180 pg/1, with a final chronic value of 64.5 pg/1 (U.S.
*
EPA, 1978).
-------
C. Plant Effects
The green freshwater algae Selenastrum capricornutum
has been exposed to five chlorinated benzenes. Based on
cell number, the 96-hour EC5Q values are as follows: chloro-
benzene 220,000 ^g/1, 1,2,4-trichlorobenzene 36,700 ^ug/1,
1,2,3,5-tetrachlorobenzene 17,700 jag/1, 1,2,4,5-tetrachloro-
benzene 46,800 ^g/1, and pentachlorobenzene 6,780 jjg/1.
D. Residues
No measured bioconcentration factor (BCF) is avail-
able for chlorobenzenes. However, the average weighted
BCF of 13 was calculated from octanol-water partition coeffi-
cient and other factors. (U.S. EPA, 1979).
VI. EXISTING GUIDLINES 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
Monochlorobenzene: The American Conference of
Governmental Industrial Hygienists {ACGIH, 1971) threshold
limit value for monochlorobenzene is 350 mg/m . The U.S.
EPA draft ambient water quality criterion for monochloro-
benzene is 20/^ig/l based on the threshold concentration
for odor and taste (U.S. EPA, 1979).
Trichlorobenzene: The American Conference of
•
Governmental Industrial Hygienists (ACGIH, 1977} threshold
limit value for 1,2,4-trichlorobenzene is 40 mg/m (5 ppm).
-------
The U.S. EPA (1979) draft ambient water quality criterion
for 1,2,4-trichlorobenzene is 13 jjg/1 based on the threshold
concentration for odor and taste.
Tetrachlorobenzene: The U.S. EPA (1979) draft
ambient water quality criterion for tetrachlorobenzene is
17 jig/1.
Pentachlorobenzene: The U.S. EPA (1979) draft
ambient water quality criterion for pentachlorobenzene is
0.5 jjg/1.
Hexachlorobenzene: The value of 0.6 jag/kg/day
hexachlorobenzene was suggested by FAO/WHO as a reasonable
upper limit for residues in food for human consumption (FAO/WHO,
1974) . The Louisiana State Department of Agriculture has
set the tolerated level of hexachlorobenzene in meat fat
a 0.3 rag/kg (U.S. EPA, 1976). The FAO/WHO recommendations
for residues in foodstuffs were 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 cancer bioassy 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^ Correspond ing Criteria
(pec day) _6 _5
2 1Q 10 10 3
2 liters of drinking water 0 0.01-25 ng/1 0.125 ng . 1 1.25 ng/1
and consumption of 18.7
grams fish and shellfish. •
Consumption of fish and 0 0.126 ng/1 0.126 ng/1 1.26 ng/1
shellfish only.
-H30-
-------
B. Aquatic
The drafted criteria to protect freshwater aquatic
life as is follows: (U.S. EPA, 1979)
Compound
Chlorobenzene
1,2,4-trichlorobenzene
1,2,3,5-tetrachlorobenzene
1,2,4,5-tetrachlorobenzene
Pentachlorbenzene
Concentration not to
be exceeded at anytime
3,500
470
390
220
36
The drafted criteria to protect saltwater aquatic
life are as follows: {U.S. EPA, 1979)
Compound
Chlorobenzene
1,2,4-Trichlorobenzene
1,2,3,5-Tetrachlorbenzene
1,2,45-Tetrachlorobenzene
Pentachlorobenzene
24-hr.
Average
120
3.4
2.6
9.6
1.3
Concentration not to
be excee_ded at anytime
,8
,9
280
7.
5.
26
2.9
-H31-
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CHLORINATED BENZENES
REFERENCES
Albro, P.W., and R. Thomas. 1974. Intestinal absorption of hexachloroben-
zene and hexachlorocyclohexane isomers in rats. Bull. Environ. Contam.
Toxicol. 12: 289.
American Conference of Governmental Industrial Hygienists. 1971. Documen-
tation of the threshold limit values for substances in workroom air. 3rd ed.
Ariyoshi, T., et al. 1975a. Relation between chemical structure and acti-
vity. I. Effects of the number of chlorine atoms in chlorinated benzenes on
the components of drug metabolizing systems and hepatic constituents. Chem.
Pharm. Bull. 23: 817.
Braun, W.H., et al. 1978. Pharmacokinetics and toxicological evaluation of
dogs fed 1,2,4,5-tetrachlorobenzene in the diet for two years. Jour. Envi-
ron. Pathol. Toxicol. 2: 225.
Cabral, J.R.P., et al. 1977. -Carcinogenic activity of hexachlorobenzene in
hamsters. Nature (London) 269: 510.
Cabral, J.R.P., et al. 1978. Carcinogenesis study in mice with hexachloro-
benzene. 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.
Coate, W.B., et al. 1977. Chronic inhalation exposure of rats, rabbits and
monkeys to 1,2,4-trichlorobenzene. Arch. Environ. Health. 32: 249.
Fomenko, v.N. 1965. Determination of the maximum permissible concentration
of tetrachlorobenzene in water basins. Gig. Sanit. 30: 8.
Food and Agriculture Organization. 1974. 1973 evaluations of some pesti-
cide residues in food. FAQ/AGP/1973/M/9/1; WHO Pestic. Residue Ser. 3.
World Health Org., Rome, Italy, p. 291.
Gotto, M., et al. 1972. Hepatoma formation in mice after administration of
high doses of hexachlorocyclohexane isomers. Chemosphere 1: 279.
Grant, D.L., et al. 1977. Effect of hexachlorobenzene on reproduction in
the rat. Arch. Environ. Contam. Toxicol. 5: 207.
Irish, D.D. 1963. Halogenated hydrocarbons: II. Cyclic. In Industrial Hy-
giene and Toxicology, Vol. II, 2nd ed., F.A. Patty, (ed.) Interscience, New
York. p. 1333.
Jondorf, W.R., et al. 1958. Studies in detoxication. The metabolism of
halogenobenzenes 1,2,3,4-, 1,2,3,5- and 1,2,4,5-tetrachlorobenzenes. Jour.
Biol. Chem. 69: 189.
-••/aa-
-------
Khera, K.S. 1974. Teratogenicity and dominant lethal studies on hexa-
chlorobenzene in rats. Food Cosmet. Toxicol. 12: 471.
Kohli, I., et al. 1976. The metabolism of higher chlorinated benzene iso-
mers. Can. Jour. Biochem. 54: 203.
Koss, G., and W. Koransky. 1975. Studies on the toxicology of hexachloro-
benzene. I. Pharmacokinetics. Arch. Toxicol. 34: 203.
Koss, G., et al. 1976. Studies on the toxicology of hexachlorobenzene.
II. Identification and determination of metabolites. Arch. Toxicol.
35: 107.
Koss, G., et al. 1978. Studies on the toxicology of hexachlorobenzene.
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.
Mackay, D., and P.J. Leinonen. 1975. Rate of evaporation of low-solubility
contaminants from water bodies to atmosphere. Environ. Sci. Technol.
9: 1178.
Parke, D.V., and R.T. Williams. 1960. Studies in detoxification LXXXI.
Metabolism of halobenzenes: (a) Penta- and hexachlorobenzene: (b) Further
ob- servations of 1,3,5-trichlorobenzene. Biochem. Jour. 74: 1.
Parrish, P.R., et al. 1974. Hexachlorobenzene: effects on several estua-
rine animals. Pages 179-187 In: Proc. 28th Annu. Conf. S.E. Assoc. Game
Fish Comm.
Pickering, Q.H., and C. Henderson. 1966. Acute toxicity of some important
petrochemicals to fish. Jour. Water Pollut. Control Fed. 38: 1419.
Preussmann, R. 1975. Chemical carcinogens in the human environment. Hand.
Allg. Pathol. 6: 421.
Rozman, K., et al. Metabolism and pharmacokinetics of penta- chlorbbenzene
in rhesus monkeys. Bull. Environ. Contam. Toxicol. (in press)
Shirai, T., et al. 1978. Hepatocarcinogenicity of polychlorinated ter-
phenyl (PCT) in ICR mice and its enhancement by hexachlorobenzene (HCB).
Cancer Lett. 4: 271.
Stijve, T. 1971. Determination and occurrence of hexachlorobenzene resi-
dues. Mitt. Geb. Lebenmittelunters. Hyg. 62: 406. :
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A mice. Cancer Res. 37: 2717.
-H33-
-------
U.S. EPA. 1975. Survey of Industrial Processing Data: Task I, Hexachloro-
benzene and nexachlorobutadiene pollution from chlorocarbon processes. Mid.
Res. Inst. EPA, Off. Toxic Subs. Contract, Washington, O.C.
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4646.
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York.
-H3H-
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No. 37
Chlorinated Ethanes
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 (GAG) has evaluated
chlorinated ethanes and has found sufficient evidence to
indicate that this compound is carcinogenic.
-------
CHLORINATED ETHANES
SUMMARY
Four of the chlorinated ethanes have been shown to
produce tumors in experimental animal studies conducted
by the National Cancer Institute (NCI). These four are
1,2-dichloroethane, 1,1,2-trichloroethane, 1,1,2,2-tetra-
chloroethane, and hexachloroethane. Animal tumors were
also produced by administration of 1,1,1-trichloroethane,
»*
but this bioassay is being repeated due to premature deaths
in one initial study.
Two of the chlorinated ethanes, 1,2-dichloroethane
and 1,1,2,2-tetrachloroethane, have shown mutagenic activity
in the Ames Salmonella assay and in E. coli. 1, 2-Dichloroethane
has also shown mutagenic action in pea plants and in Drosophila.
No evidence is available indicating that the chloroethanes
produce teratogenic effects. Some toxic effects on fetal
development have been shown following administration of
1,2-dichloroethane and hexachloroethane.
Symptoms produced by toxic exposure to the chloroethanes
include central nervous system disorders, liver and kidney
damage, and cardiac effects.
Aquatic toxicity data for the effects of chlorinated
ethanes to freshwater and marine life are few. Acute studies
have indicated that hexachloroethane is the most toxic of
the chlorinated ethanes reviewed. Marine organisms tepd
to be more sensitive than freshwater organisms with acute
toxicity values as low as 540 ug/1 being reported.
-------
CHLORINATED ETHANES
I. INTRODUCTION
This profile is based on the draft Ambient Water Quality
Criteria Document for Chlorinated Ethanes (U.S. EPA, 1979).
The chloroethanes (see table 1) are hydrocarbons in
which one or more of the hydrogen atoms have been replaced
by chlorine atoms. Water solubility and vapor pressure
decrease with increasing chlorination, while density and
melting point increase. Monochloroethane is a gas at room
temperature, hexachloroethane is a solid, and the remaining
compounds are liquids. All chloroethanes show some solubility
in water, and all, except monochloroethane, are more dense
than water.
The chloroethanes are used as solvents, cleaning and
degreasing agents, in the manufacture of plastics and textiles,
and in the chemical synthesis of a number of compounds.
Current production:
monochloroethane 335 x 10:;. tons/yr in 1976
1,2-dichloroethane 4,000 x 10 tons/yr in 1976
1,1,1-trichloroethane 215 x 10 tons/yr in 1976
The chlorinated ethanes form azeotropes with water
{Kirk and Othmer, 1963). All, are very soluble in organic
solvents (Lange, 1956). Microbial degradation of the chlorin-
ated ethanes has not been demonstrated (U.S.- EPA, 1979).
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. Water monitoring studies
-------
have shown the following levels of various chloroethanes:
1,2-dichloroethane, 0.2-8 ug/1; 1,1,2-trichloroethane, 0.1-
8.5 ug/1; 1/1,1,2-tetrachloroethane, 0.11 pg/1 (U.S. EPA,
1979) . In general, air levels of chloroethanes are produced
by evaporation of volatile chloroethanes widely used as
degreasing agents and in dry cleaning operations (U.S. EPA,
1979). Industrial monitoring studies have shown air levels
of 1,1,1-tr ichloroethane ranging from 1.5 to 396 ppm (U.S.
EPA, 1979).
TABLE 1
Chloroethanes and Synonyms
Synonyms
Compound Name
Monochloroethane
1,1,-Dichloroethane
1, 2-Dichloroethane
1,1,1-Trichloroethane
1,1,2-Tr ichloroethane
1,1,1,2-Tetrachloroethane
1,1,2,2-Tetrachloroethane
Pentachloroethane
Hexachloroethane
Chloroethane
Ethylidene Bichloride
Ethylene Dichloride
Methyl Chloroform
Ethane Trichloride
Tetrachloroethane
Acetylene Tetrachloride
Pentalin
Perchloroethane
Ethyl chloride
Ethylidene Chloride -
Ethylene Chloride
Chlorothene
Vinyl Trichloride
Sym-Tetrachloroethane
Ethane Pentachloride
Sources of human exposure to chloroethanes include
water, air, contaminated foods and fish,'and dermal absorption.
»
The two most widely used solvents, 1,2-dichloroethane and
1,1,1-trichloroethane, are the compounds most often detected
in foods. Analysis of several foods indicated 1,1,1-trichloro-
-------
ethane levels of 1-10 ug/kg (Walter, et al. 1976), while
levels of 1,2-dichloroethane found in 11 of 17 species have
been reported to be 2-23 ug/g (Page and Kennedy, 1975) .
• Fish and shellfish have shown levels of chloroethanes in
the nanogram range (Dickson and Riley, 1976).
The U.S. EPA (1979) has derived the following weighted
average bioconcentratioa factors for the edible portions
of fish and shellfish consumed by Americans: 1,2-dichloro-
ethane, 4.6; 1,1,1-trichloroethane, 21; 1,1,2,2-tetrachloro-
v
ethane, 18; pentachloroethane, 150; hexachloroethane, 320.
These estimates were based on the measured steady-state
bioconcentration studies in bluegill. Bioconcentration
factors for 1,1, 2-trichloroethane (6.3) and 1,1,1,2-tetrachloro-
ethane (18) were derived by EPA (1979) using octanol-water
partition coefficients.
III. PHARMACOKINETICS
A. Absorption
The chloroethanes are absorbed rapidly following
ingestion or inhalation (U.S. EPA, 1979). Dermal absorption
is thought to be slower in rabbits based on studies by Smyth,
et al. (1969). However, rapid dermal absorption has been
seen in guinea pigs with the same trichloroethane (Jakobson,
et al. 1977).
Human studies on the absorption of inhaled 1,1,2,2-
tetrachloroethane indicate that the compound is completely
absorbed after exposure to trace levels of radiolabeled
vapor (Morgan, et al.r 1970, 1972). At higher exposure
levels absorption is rapid in man and animals, but obviously
not complete.
-------
B. Distribution
Studies on the distribution of 1,1,1-trichloroethane
in mice following inhalation exposure have shown levels
in the liver to be twice that found in the kidney and brain
(Holmberg, et al. 1977). Postmortem examination of human
tissues showed 1,1,1-trichloroethane in body fat (highest
concentration) kidneys, liver, and brain (Walter, et al.
1976). Due to the lipid solubility of chloroethanes, body
distribution may be expected to be widespread. Stahl, et
v
al. (1969) have noted that human tissue samples of liver,
brain, kidney, muscle, lung, and blood contained 1,1,1-tri-
chloroethane' following acute exposure, with the liver contain-
ing the highest concentration.
Passage of 1,1,1,2-tetrachloroethane across the
placenta has been reported by Truhaut, et al. (1974) in
rabbits and rats.
C. Metabolism
The metabolism of chloroethanes involves both
enzymatic dechlorination and hydroxylation to corresponding
alcohols (Monster, 1979; Truhaut, 1972), Oxidation reactions
may produce unsaturated metabolites which are then transformed
to the alcohol and ester (Yllner, 1971 a,b,c,d).
Metabolism appears to involve the activity of
the mixed function oxidase enzyme system (Van Dyke and Wineman,
1971) . Animal experiments by Yllner (197'i a,b,c,d,e) indicated
that the percentage of administered compound metabolized
decreased with increasing dose, suggesting saturation of
metabolic pathways.
if
-443.-
-------
D. Excretion
The chloroethanes are excreted primarily in the
urine and in expired air {U.S. EPA, 1979). As much as 60
to 80 percent of an inhaled dose of 1,1,1-trichloroethane
(70 or 140 ppm for 4 hours) was expired unchanged by human
volunteers (Monster, et al. 1979). Animal studies conducted
by Yllner (1971 a,b,c,d) indicate that largest amount of
chloroethanes, administered by intraperitoneal (i.p.) injec-
tion is excreted in the urine; this is followed by expira-
^
tion (in the changed or unchanged form), with very little
excretion in the feces. Excretion appears to be rapid,
since 90 percent of i.p. administered doses of 1,2-dichloro-
ethane or 1,1,2-trichloroethane were eliminated in the first
24 hours (U.S. EPA, 1979). However, the detection of chloro-
ethanes in postmortem tissue samples indicates that some
portion of these compounds persists in the body (Walter,
et al. 1976).
IV. EFFECTS
A. Carcinogenicity
Several chlorinated ethanes have been shown to
produce a variety of tumors in rats and mice in experiments
utilizing oral administration. Tumor types observed after
compound administration include squamous cell carcinoma
of the stomach, hemangiosarcoma, adenocarcinoma of the mam-
mary gland, and hepatocellular carcinoma (NCI, 1978a,b,c,d).
The four chlorinated ethanes which have been classified
as carcinogens based on animal studies are: 1,2-dichloro-
ethane, 1,1,2-trichloroethane, 1,1,2,2-tetrachloroethane,
-------
and hexachloroethane. Increased tumor production was also
noted in animals treated with 1 , 1, 1-tr ichloroethane, but
high mortality during this study (NCI, 1977) caused retest-
ing of the compound to be initiated. Iri vitro transforma-
tion of rat embryo cells and subsequent f ibrosarcoma produc-
tion by these cells when injected in vivo, indicate that
1, 1,1-tr ichloroethane does have carcinogenic potential (Price,
et al. 1978) .
B. Mutagenicity
«.
Two of the chlorinated ethanes, 1, 2-dichloroethane
and 1,1, 2, 2-tetrachloroethane, have shown mutagenic activity
in the Ames Salmonella assay and for DNA polymerase deficient
strain of E. cgl^ (Brem, et al. 1974) . In these two systems,
1,1, 2, 2-tetrachloroethane showed higher mutagenic activity
than 1, 2-dichloroethane (Rosenkranz, 1977).
Mutagenic effects have been produced by 1, 2-dichloro-
ethane in pea plants {Kirichek, 1974) and in
{Nylander, et al. 1978). Several metabolites of dichloro-
ethane {chloroacetaldehyde, chloroethanol , and S-chloroethyl
cysteine have also been shown to produce mutations in the
Ames assay (U.S. EPA, 1979). .
Testing of hexachloroethane in the Ames Salmonella
assay or in a yeast assay system failed to show any mutagenic
activity (Weeks, et al. 1979) .
C. Teratogenicity
Inhalation exposure of pregnant rats and mice
to 1 , 1, 1-tr ichloroethane was shown to produce some soft
4
-------
tissue and skeletal deformities; this incidence was not
judged statistically significant by the Fisher Exact proba-
bility test (Schwetz, et al. 1975).
Testing of hexachloroethane administered to rats
by intubation or inhalation exposure did not show an increase
in teratogenic effects (Weeks, et al. 1979). Inhalation
exposure of pregnant rats to 1,2-dichloroethane also failed
to demonstrate teratogenic effects (Schwetz, et al. 1974;
Vozovaya, 1974).
D. Other Reproductive Effects
Decreased litter size, reduced fetal weights and
a reduction in live births have been reported in rats exposed
3
to 1,2-dichloroethane (57 mg/m m four hours/day, six days/week)
by inhalation (Vozovaya, 1974). 1,1-Dichloroethane retarded
fetal development at exposures of 6,000 ppm. (Schwetz, et
al. 1974). Higher fetal resorption rates and a decreased
number of live fetuses per litter were observed in rats
following administration of hexachloroethane by intubation
(15, 48 or 260 ppm, 6 hours/day) or inhalation (50, 100
or 500 mg/kg/day) (Weeks, et al. 1979).
E. Chronic Toxicity
Neurologic changes and liver and kidney damage
have been noted following long term human exposure to 1,2-
dichloroethane (NIOSH,1978). Cardiac effects (overstimulation)
have been noted following human exposure to 1,1-dichloroethane
(U.S. EPA, 1979) .
Central nervous system disorders have been reported
in humans exposed to 1,1,1-trichloroethane. Symptoms noted
-------
were altered reaction time, perceptual speed, manual dexterity,
and equilibrium (U.S. EPA, 1979}.
Animal studies indicate that the general symptoms
of toxicity resulting from exposure to the chloroethanes
involve effects in the central nervous system, cardiovascular
system, pulmonary system, and the liver and kidney (U.S.
EPA, 1979). Laboratory animals and humans exposed to chloro-
ethanes show similar symptoms of toxicity (U.S. EPA, 1979).
Based on data derived from animal studies, the
U.S. EPA (1979) has concluded that the relative toxicity
of the chlproethanes is as follows: 1,2-dichloroethane>
1,1,2, 2- tetrachloroethane p>l, 1, 2-tr ichloroethane >hexachloro-
ethane 1,1-dichloroethane>1,1,1-trichloroethane> monochloro-
ethane.
F. Other Relevant Information
The hepatotoxicity of 1,1,2-trichloroethane was
increased in mice following acetone or isopropyl alcohol
pretreatment (Traiger and Plaa, 1974). Similarly, ethanol
pretreatment of mice increased the hepatic effects of 1,1,1-
trichloroethane (Klassen and Plaa, 1966).
Hexobarbital sleeping times in rats were reported
to be decreased following inhalation exposure to 1,1,1-tri-
chloroethane (3,000 ppm), indicating an effect of the compound
on stimulation of hepatic microsomal enzymes (Fuller, et
al. 1970).
V. AQUATIC TOXICITY
A. Acute Toxicity
Acute toxicity studies were conducted on three
species of freshwater organisms and two marine species.
-------
For freshwater fish, 96-hour static ^£50 values for the
bluegill sunfish, Lepomis macrochirus, ranged from 980 ug/1
hexachloroethane to 431,000 ug/1 1,2-dichloroethane, while
the range of 48-hour LCrn values for the freshwater inverte-
brate Daphnia magna was 8,070 ug/1 to 218,000 ug/1 for hexa-
chloroethane and 1,2-dichloroethane respectively. Among
marine organisms, the sheepshead minnow (Cypr inodon vagie-
gatus) produced LC5Q values ranging from 2,400 jjg/1 for
hexachloroethane to 116,000 ug/1 for pentachloroethane.
The marine mysid shrimp (Mysidopsis bahia) produced LC^Q
values ranging from 940 ug/1 for hexachloroethane to 113,000
ug/1 for 1,2-dichloroethane. The general order of acute
toxicities for the chlorinated ethanes reviewed for fresh-
water fish is: hexachloroethane (highest toxicity), 1,1,2,2-
tetrachloroethane, 1,1,2-trichloroethane, pentachloroethane,
and 1,2-dichloroethane (U.S. EPA, 1979).
B. Chronic Toxicity
The only chronic study available for the chlori-
nated ethanes is for pentachloroethane's chronic effects
on the marine shrimp (Mysidopsijj bahia) , which produced
a chronic value of 580 pg/1 {U.S EPA, 1978).
C. Plant Effects
Effective EC5Q concentrations, based on chlorophyll
a and cell numbers for the freshwater'- algae Selenastrum
capeiconutum ranges from 87,000 ug/1' for hexachloroethane
to 146,000 ug/1 for 1,1,2,2-tetrachloroethane, with penta-
chloroethane being intermediate in its phytotoxicity. For
the marine algae Skeletonema costaturn, a greater sensi-
^
-4H7-
-------
tivity was indicated by effective ECgQ_ concentrations based
on cell numbers and chlorophyll a ranging from 6,230 ug/1
for 1,1,2,2-tetrachloroethane and 7,750 ug/1 for hexachloro-
ethane to 58,200 ug/1 for pentachloroethane.
D. Residues
The bioconcentration value was greatest for hexa-
chloroethane with a value of 139 ug/1 being reported for
bluegill. Bioconcentration values of 2, 8, and 9 were obtained
for 1,2-dichloro, 1,1,2,2-tetrachloro, and 1,1,1-trichloro-
».
ethane for bluegills. 1,1,2-Trichloroethane and 1,1,1,2-
tetrachloroethane used the octanol/water coefficients to
derive BCF's of 22 and 62, respectively.
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
Based on the NCI carcinogenesis bioassay data,
and using a linear, non-threshold model, the U.S. EPA (1979)
has estimated levels of four chloroethanes in ambient water
that will result in an additional cancer risk of 10~ : 1,2-
dichloroethane, 7.0 jug/I; 1,1,2-trichloroethane, 2.7 ug/1;
1,1,2,2-tetrachloroethane, 1.8 pg/1; hexachloroethane, 5.9
ug/1. A draft ambient water quality/ criterion to protect
human health has been derived by EPA for 1,-1,1-tr ichloro-
ethane based on mammalian toxicity data at the level of
15.7 mg/1.
-------
Insufficient mammalian toxicological data prevented
derivation of a water criterion for monochloroethane, 1,1-
dichloroethane, 1,1,1,2-tetrachloroethane, or pentachloro-
ethane (U.S. EPA, 1979) .
The following compounds have had eight hour, TWA
exposure standards established by OSHA: monochloroethane,
1,000 ppm; 1,1-dichloroethane, 100 ppm; 1,2-dichloroethane,
50 ppm; 1,1,1-trichloroethane, 350 ppm; 1,1,2-trichloroethane,
10 ppm; 1,1,2,2-tetrachloroethane, 5 ppm; hexachloroethane,
1 ppm.
B. Aquatic
Criteria for protecting freshwater organisms have
been drafted for five of the chlorinated hydrocarbons: 62
pg/1 (average concentation) not to exceed 140 pg/1 for hexa-
chloroethane; 170 pg/1 not to exceed 380 pg/1 for 1,1,2,2-
tetrachloroethane; 440 /ag/1 not to exceed 1,000 jug/1 for
pentachloroethane; 3,900 pg/1 not to exceed 8,800 pg/1 for
1,2-dichloroethane; and 5,300 pg/1 not to exceed 12,000
pg/1 for 1,1,1-trichloroethane. For marine organisms, cri-
teria have been drafted as: 7 pg/1 (average concentration)
not to exceed 16 pg/1 for hexachloroethane; 38 pg/1 not
to exceed 87 pg/1 for pentachloroethane; 70 pg/1 not to
exceed 160 pg/1 for 1,1,2,2-tetrachloroethane; 240 pg/1
not to exceed 540 pg/1 for 1,1,1-trichloroethane; and 880
pg/1 not to exceed 2,000 ug/1 for 1,2-d'i'chloroethane.
-------
CHLORINATED ETHANES
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-------
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National Institute for Occupational Safety and Health. 1978.
Ethylene dichloride (1,2-dichloroethane). Current Intelli-
gence Bull. 25. DHEW (NIOSH) Publ. No. 78-149.
Nylander, P.O., et al. 1978. Mutagenic effects of petrol in
Drosophilia melanoqaster. I. Effects of benzene of and 1,2-
dichloroethane. Mutat. Res. 57: 163.
Page, R.D., and P.P.C. Kennedy. 1975. Determination of
mthylene chloride, ethylene dichloride, and trichloroethylene
as solvent residues in spice oleoresins, using vacuum distil-
lation and electron-capture gas chronatography. Jour.
Assoc. Off. Anal. Chem. 58: 1062.
Price, P.J., et al.' 1978. -Transforming activities of tri-
chloroethylene and proposed industrial alternatives. In
vitro. 14: 290.
Rosenkranz, H.S. 1977. Mutagenicity of halogenated alkanes
and their derivatives. Environ. Health Perspect. 21: 79.
Schwetz, B.A., et al. 1974. Enbryo- and fetotoxicity of in-
haled carbon tetrachloride, 1,1,-dichloroethane, and methyl
ethyl ketone in rats. Toxicol. Appl. Pharmacol. 28: 452.
Schwetz, B.A., et al. 1975. Effect of"maternally inhaled>
trichloroethylene, perchloroethylene, methyl chloroform, and
methylene chloride on embryonal and fetal development in mice
and rats. Toxicol. Appl. Pharmacol. 32: 84.
-------
Smyth, H.F., Jr., et al, 1969. Range-finding toxicity data:
list VII. Am. Ind. Hyg. Assoc. Jour. 30: 470.
Stahl, C.J., et al. 1969. Trichloroethane poisoning: ob-
servations on the pathology and toxicology in six fatal
cases. Jour. Forensic Sci. 14: 393.
Traiger, G.J., and G.L. Plaa. 1974. Chlorinated hydrocarbon
toxicity. Arch. Environ. Health 28: 276.
Truhaut, R. 1972. Metabolic transformations of 1,1,1,2-
tetrachloroethane in animals (rats, rabbits). Chem. Anal.
(Warsaw) 17: 1075.
Truhaut, R., et al. 1974. Toxicological study of 1,1,1,2-
tetrachloroethane. Arch. Mai. Prof. Med. Trav. Secur. Soc.
35: 593.
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. 1979. Chlorinated Ethanes: Ambient Water Quality
Criteria (Draft).
Van Dyke, R.A., and C.G. Wineman. 1971. Enzymatic dechlori-
nation: Dechlorination of chloroethanes and propanes jjn
vitro. Biochem. Pharmacol. 20: 463.
Vozovaya, M.A. 1974. Development of progeny of two genera-
tions obtained from female rats subjected to the action of
dichloroethane. Gig. Sanit. 7: 25.
Walter, P., et al. 1976. Chlorinated hydrocarbon toxicity
(1,1/1-trichloroethane, trichloroethylene, and tetrachloro-
ethylene): a monograph. PE Rep, PB-257185. Natl. Tech.
Inf. Serv., Springfield, Va.
Weeks, M.H., et al.- 1979. The toxicity of hexachloroethane
in laboratory animals. Am."Ind. Hyg. Assoc, Jour. 40: 187.
Yllner, S. 1971a. Metabolism of 1,2-dichloroethane-14c
in the mouse. Acta. Pharmacol. Toxicol. 30: 257.
Yllner, S. 1971b. Metabolism of 1,1,2-trichloroethane-l,2-
14C in the mouse. Acta. Pharmacol. Toxicol. 320: 248.
Yllner, S. 1971c. Metabolism of 1,1,1/2-tetrachloroethane
in the mouse. Acta. Pharmacol. Toxicol. 29: 471.
Yllner, S. 1971d. Metabolism of 1,1,2,2-tetrachloroethane-
C in the mouse. Acta. Pharmacol. Toxicol. 29: 499.
Yllner, S. 1971e. Metabolism of pentachloroethane in the
mouse. Acta. Pharmacol. Toxicol. 29: 481.
-------
No. 38
Chlorinated Naphthalenes
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.
-------
CHLORINATED NAPTHALENE5
Summary
Chlorinated naphthalenes have been used in a variety of industries,
usually as mixtures. Chronic toxicity varies with the degree of chlori-
nation, with the more highly chlorinated species being more toxic. The
clinical signs of toxicity in humans are damage to liver, heart, pancreas,
gall bladder, lungs, adrenal glands, and kidney. No animal or human studies
have been presented on the carcinogenicity, mutagenicity, or teratogenicity
of polychlorinated naphthalenes.
Very little data on aquatic toxicity are available for individual
chlorinated naphthalenes. 48-Hour and 96-hour LC^Q values of octachloro-
naphthalene over 500,000 pg/1 have been reported for Daphnia magna and the
bluegill, respectively. A freshwater alga also resulted in a 96-hour LC5Q
value for octachloronaphthalene of over 500,000 pg/1.
Toxicity studies with aquatic organisms are confined to tests with 1-
chloronaphthalene on one freshwater fish and two algal species (one fresh
and one saltwater). All tests have reported 96-hour LC5Q values of be-
tween 320 and 2,270 jjg/1. Exposure of sheepshead minnow to 1-chloronaphtha-
lene in an embryo-larval study resulted in a chronic value of 328 jjg/1.
-------
CHLORINATED NAPTHALENES
I. INTRODUCTION
This profile is based on the draft Ambient Water Quality Criteria Docu-
ment for Chlorinated Naphthalenes (U.S. EPA, 1979).
Chlorinated naphthalenes consist of two fused six carbon-membered aro-
matic rings where any or all of the eight hydrogen atoms can be replaced
with chlorine. The physical properties of the chlorinated naphthalenes are
generally dependent on the degree of chlorination. Melting points range
from 17°C for 1-chloronaphthalene to 198°C- for 1,2,3,4-tetrachloro-
naphthalene. As the degree of chlorination increases, the specific gravity,
boiling point, fire and flash points all increase, while the vapor pressure
and water solubility decrease. Chlorinated naphthalenes have been used as
the paper impregnant in automobile capacitors (mixtures of tri- and tetra-
chloronaphthalenes), and as oil additives for engine cleaning, and in fabric
dyeing (mixtures of mono- and dichloronaphthalenes). In 1956, the total
U.S. production of chlorinated naphthalenes was approximately 3,175 metric
tons (Hardie, 1964).
II. EXPOSURE
A. Water
To date, polychlorinated naphthalenes have not been identified in
drinking waters (U.S. EPA, 1979).' However, these compounds have been found
in waters or sediments adjacent to point sources or areas of heavy poly-
chlorinated biphenyl contamination.
B. Food
Polychlorinated naphthalenes appear tto be biomagnified in the aqua-
tic ecosystem, with the degree of biomagnification being greater for the
more highly chlorinated polychlorinated compounds (Walsh, et al. 1977).
-------
Erickson, et al. (1978) also noted a higher relative biomagnification of the
lowest chlorinated naphthalenes by the fruit of apple trees grown on contam-
inated soil. The U.S. EPA (1979) has estimated the weighted average biocon-
centration factor for Halowax 1014 (a mixture of chlorinated naphthalenes)
to be A, 800 for the edible portions of fish and shellfish consumed by
Americans. This estimate is based on measured non-steady-state bioconcen-
tration studies in brown shrimp.
C. Inhalation
Erickson, et al. (1978) found ambient -air concentrations of poly-
chlorinated naphthalenes ranging from 0.025 to 2.90 jug/m near a poly-
chlorinated naphthalene plant. Concentrations of trichloronaphthalene were
as high as 0.95 pg/m , while hexachloronaphthalene concentrations never
exceeded 0.007 jug/m .
III. PHARMACOKINETICS
A. Absorption
Pertinent data could not be located in the available literature.
B. Distribution
In the rat fed 1,2-dichloronaphthalene, the chemical and its metab-
olites were found primarily in the intestine, kidney, and adipose tissue
(Chu, et al. 1977).
C. Metabolism
There appears to be appreciable metabolism in mammals of poly-
chlorinated naphthalenes containing four chlorine atoms or less (U.S. EPA,
1979). Cornish and Block (1958) investigated /the excretion of polychlori-
nated naphthalenes in rabbits and found 79 percent of 1-chloronaphthalene,
»
93 percent of dichloronaphthalene, and 45 percent of tetrachloronaphthalene
-------
were excreted in the urine as metabolites .of the parent compounds. Metab-
olism may involve hydroxylation alone or hydroxylation in combination with
dechlorination. In some cases, an arene oxide intermediate may be formed
(Ruzo, et al. 1976).
0. Excretion
In rats fed 1,2-dichloronaphthalene, initially more of the chemical
and its metabolites were found in the urine; however, by the end of seven
days a greater proportion had been excreted in the feces (Chu, et al. 1977).
In the first 24 hours, 62 percent of the dose was excreted in the bile, as
compared to 18.9 percent lost in the feces. This suggests that there is an
appreciable reabsorption and enterohepatic recirculation of this particular
chlorinated naphthalene.
IV. EFFECTS
No animal or human studies have been reported on the carcinogenicity,
mutagenicity, or teratogenicity of chlorinated naphthalenes. No other re-
productive effects were found in the available literature.
A. Chronic Toxicity
Chronic dermal exposure to penta- and hexachlorinated naphthalenes
causes a form of chloracne which, if persistent, can progress to form a cyst
or sterile abcess (Jones, 1941; Shelley and Kligman, 1957; Kleinfeld, et al.
1972). Workers exposed to these two isomers complained of eye irritation,
headaches, fatigue, vertigo, nausea, loss of appetite, and weight loss
(Kleinfeld, et al. 1972). More severe exposure to the fumes of polychlori-
nated naphthalenes has produced severe liver damage, together with damage to
r
the heart, pancreas, gall bladder, lungs, adrenal glands, and kidney tubules
»
(Greenburg, et al. 1939). Chronic toxicity in animals appears to be quali-
tatively the same (U.S. EPA, 1979). Polychlorinated naphthalenes containing
-------
three or fewer chlorine atoms were found to be nontoxic, while tetrachloro-
naphthalene resulted in mild liver disease at levels as high as 0.7 nig/kg/-
day; the higher chlorinated naphthalenes produce more severe disease at
lower doses (Bell, 1953). Because of their insolubility, hepta- and octa-
chloronaphthalene were less toxic when given in suspension than when given
in solution.
8. Other Relevant Information
Drinker, et al. (1937) showed enhancement of hepatoxicity of a mix-
ture of ethanol/carbon tetrachloride in rats pretreated with 1.16 mg/m of
a penta-Xhexachloronaphthalene mixture in air for six weeks. In a similar
study trichloronaphthalene was inactive.
V. AQUATIC TOXICITY
A. Acute Toxiclty
The 96-hour LC^ value reported for the bluegill, Lepomis
-2U
macrochirus, exposed to 1-chloronaphthalene is 2,270 /jg/1 (U.S. EPA, 1978).
With saltwater species, exposure of the sheepshead minnow, Cyprinodon
variqatus, and mysid shrimp, Mysidopsis bahia, to 1-chloronaphthalene pro-
vided 96-hour LC5Q values of 1,290 and 370 jug/1, respectively. Daphnia
magna and the bluegill, Lepomis macrochirus, have a slight sensitivity to
octachloronaphthalene,' with respective 48-hour and 96-hour LC5Q values
greater than 530,000 pg/1 (U.S. EPA, 1978).
B. Chronic Toxicity
In the only chronic study reported (embryo-larval), exposure of
1-chloronaphthalene to the sheepshead minnow resulted in a chronic value of
329 ug/1 (U.S. EPA, 1978). '.''..
-------
C. Plant Effects
A freshwater alga, Selenastrum capricornutum, and a saltwater alga,
Skeletonema costatum , when exposed to 1-chloronaphthalene, both produced 96-
hour EC5g values ranging from 1,000 to 1,300 jjg/1 based on cell numbers.
Octachloronaphthalene exposure to Selenastrum capricornutum re-
sulted in a 96-hour EC50 value of over 500,000 jug/1 based on cell numbers
(U.S. EPA, 1978).
0. Residues
Pertinent data could not be located in the available literature.
VI. EXISTING GUIDELINES AND STANDARDS
A . Human
The only standards for polychlorinated naphthalenes are the ACGIH
Threshold Limit Values (TLV) adopted by the Occupational Safety and Health
Administration and are as follow:
ACGIH (1977)
Threshold Limit Values
Trichloronaphthalene
Tetrachloronaphthalene
Pentachloro naphthalene
Hexachloronaphthalene
Octachloronaphthalene
5
2
0.5
0.2
0.1
mg/m3
. mg/rrP
mg/rfK
mg/m3
There are no state or federal water quality or ambient air quality standards
for chlorinated naphthalenes.
The U.S. EPA is presently evaluating the available data and has
recommended that a single acceptable daily intake (ADI) of 70 pg/man/day be
r
used for the tri-, tetra-, penta-, hexa-, ano\ octa-chlorinated naphthalenes.
»
This ADI will be used to derive the human health criteria for the chlori-
nated naphthalenes.
-'-/(.O-
-------
B. Aquatic
For 1-chloronaphthalene, the draft criterion to protect freshwater
aquatic life is 29 jug/1 as a 24-hour average, not to exceed 67 ;jg/l at any
time. For saltwater aquatic species, the draft criteron is 2.8 pg/1 as a
24-hour average, not to exceed 6.4;jg/l at any time (U.S. EPA, 1979).
-------
CHLORINATED NAPHTHALENE
REFERENCES
American Conference of Governmental Industrial Hygienists.
1977. TLVs Threshold Limit Value for chemical substances and
physical agents in the workroom environment with intended
changes. Cincinnati, Ohio.
Bell, W.S. 1953. The relative toxicity of the chlorinated
naphthalenes in experimentally produced bovine hyperkeratosis
(X-disease). Vet. Met. 48: 135.
Chu, I., et al. 1977. Metabolism and tissue distribution of
(1,4,5,-14c)-l,2-dichloronaphthaline in rats. Bull.
Environ. Contain. Toxicol. 18: 177.
Cornish, H.H., and W.D. Block. 1958. Metabolism of chlori-
nated naphthalenes. Jour. Biol. Chem. 231: 583.
Drinker, C'.K. , et al. 1937. The problem of possible sys-
temic effects from certain chlorinated hydrocarbons. Jour.
Ind. Hyg. Toxicol. 19: 283.
Erickson, M.D., et al. 1978. Sampling and analysis for
polychlorinated naphthalenes in the environment. Jour.
Assoc. Off. Anal. Chem. 61: 1335.
Greenburg, L., et al. 1939. The systemic effects resulting
from exposure to certain chlorinated hydrocarbons. Jour.
Ind. Hyg. Toxicol. 21: 29.
Hardie, D.W.F. 1964. Chlorocarbons and chlorohydrocarbons:
Chlorinated Naphthalenes. pp. 297-303 In: Kirk-Othmer, En-
cyclo. of Chemical Technology. 2nd ed. John Wiley and Sons,
Inc., New York.
Jones, A.T. 1941. The etiology of acne with special -refer-
ence to acne of occupational origin. Jour. Ind. Hyg. Toxi-
col. 23: 290.
Kleinfeld, M., et al. 1972. Clinical effects of chlorinated
naphthalene exposure. Jour. Occup. Med. 14: 377.
Ruzo, L., et al. 1976. Metabolism of chlorinated naphtha-
lenes. Jour. Agric. Food Chem. 24: 581.-
Shelley, W.B., and A.M. Kligman. 195-7'. The experimental
production of acne by penta- and hexachloronaphthalenes.
A.M.A. Arch. Derm. 75: 689.
-------
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, Washington, D.C.
U.S. EPA. 1979. Chlorinated Naphthalenes: Ambient Water
Quality Criteria. (Draft).
Walsh, G.E., et al. 1977. Effects and uptake of chlorinated
naphthalenes in marine unicellular algae. Bull. Environ.
Contam. Toxicol. 18: 297.
-------
No. 39
Chlorinated Phenols
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.
-------
SPECIAL NOTATION
The National Cancer Institute (1979) has recently published the results
of a bioassay indicating that 2,4,6-trichlorophenol induced cancer in rats
and mice. This study was not included in the Ambient Water Quality Criteria
Document (U.S. EPA, 1979} and has not been reviewed for this hazard profile.
-------
CHLORINATED PHENOLS
SUMMARY
Mammalian data supporting chronic effects for most of these compounds
is limited. Insufficient data exist to indicate that any of the chlorinated
phenols are carcinogens. In skin painting studies, 3-chlorophenol and
2,4,5-trichlorophenol promoted papillomas. A lifetime feeding study with
2,4,6-trichlorophenol was inconclusive and only provided weak suspicion of
carcinogenicity. 2,4,6-Trichlorophenol gave some evidence of mutagenicity
in two assays. Tetrachlorophenol was not found to be fetotoxic in animals.
Chronic exposure to 4-chlorophenol produced myoneural disorders in humans
and animals. Adverse health effects in workers exposed to 2,4,5-trichloro-
phenol may have been due to 2,3,7,8-tetrachlorodibenzo-p-dioxin contamina-
tion of the chlorophenol.
Workers chronically exposed to tetrachlorophenol, pentachlorophenol,
and small amounts of chlorodibenzodioxins developed severe skin irritations,
respiratory difficulties, and headaches. Chlorophenols are uncouplers of
oxidative phosphorylation. 2,6-Dichlorophenol and trichlorocresol are con-
vulsants. Chlorocresol has caused several cases of local and generalized
^
reactions.
In acute toxicity tests, 4-chloro-3-methylphenol has been proven toxic
at concentrations as low as 30 ug/1 in freshwater fish, whereas other fresh-
water and marine organisms appear to be more resistant. The tainting of
rainbow trout flesh has been demonstrated at exposures of 15 to 84 ug/1 for
several of the chlorinated phenols.
-------
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Chlorinated Phenols (U.S. EPA, 1979).
The chlorinated phenols represent a group of commercially produced sub-
stituted phenols and cresols also referred to as chlorophenols or chlorocre-
sols. The compounds 2-chlorophenol, 2,4-dichlorophenol, 2,4,6-trichlorophe-
nol, and pentachlorophenol are discussed in separate hazard profiles.
Purified chlorinated phenols are colorless, crystalline solids (with
the exception of 2-chlorophenol which is a liquid), while the technical
grades may be light tan or slightly pink due to impurities. Chlorophenols
have pungent odors. In general, the volatility of chlorinated phenols de-
creases and the melting and boiling points increase as the number of substi-
tuted chlorine atoms increases. Although the solubility of chlorinated phe-
nols in aqueous solutions is relatively low, it increases markedly when the
pH of the solution exceeds the specific pKa. The solubilities of chlori-
nated phenols and chlorocresols (with the exception of 2,4,6-trichloro-m-
cresol) range from soluble to very soluble in relatively non-polar solvents
such as benzene and petroleum ether (U.S. EPA, 1979).
The chlorinated phenols that are most important commercially ..are 4-
chlorophenol, 2,4-dichlorophenol, 2,4,5-trichlorophenol, 2,3,4,6-tetra-
chlorophenol, pentachlorophenol, and 4-chloro-o-cresol. Many of the chloro-
phenols have no commercial application but are produced to some extent as
byproducts of the commercially important compounds. The highly toxic poly-
chlorinated dibenzo-p-dioxins may be formed during'the chemical synthesis of
f
IT
some chlorophenols. During the chlorination of'drinking waters and waste-
>
water effluents, chlorophenols may be inadvertently produced (U.S. EPA,
1979).
-------
Chlorinated phenols are used as intermediates in the synthesis of dyes,
pigments, phenolic resins, pesticides, and herbicides. Certain chlorophe-
nols are used directly as flea repellants, fungicides, wood preservatives,
mold inhibitors, antiseptics, disinfectants, and antigumming agents for
gasoline.
It is generally accepted that chlorinated phenols will undergo photoly-
sis 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 with subsequent polymerization (U.S. EPA, 1979). Micro-
bial degradation of chloropnenols has been reported by numerous investiga-
tors (U.S. EPA, 1979).
3-CHLOROPHENOL and 4-CHLORQPHENOL
II. EXPOSURE
Monochlorophenols have been found in surface waters in the Netherlands
at concentrations of 2 to 20 pg/1 (Piet and DeGrunt, 1975). Ingestion of
chlorobenzene can give rise to internal exposure to 2-, 3-, and 4-chlorophe-
nols, as chlorobenzene is metabolized to monochlprophenols (Lindsay-Smith,
et al. 1972). No data were found demonstrating the presence of monochloro-
*
phenol in food.
For 4-chlorophenol the U.S. EPA has estimated the weighted average bio-
concentration factor for the edible portions of all aquatic organisms con-
sumed by Americans to be 12. This estimate is based on the octanol/water
partition coefficient.
Data were not found in the available literature regarding inhalation
exposure.
-------
III. PHARMACOKINETICS
Systematic studies of the pharmacokinetics of 3- or 4-chlorophenol are
not available. Dogs excreted 87 percent of administered 4-chlorophenol in
the urine as sulfuric and glucuronic conjugates (Karpow, 1893).
IV. EFFECTS
A. Carcinogenicity
Information is not adequate to determine whether 3- or 4-chlorophe-
nol possess carcinogenic properties. A 20 percent solution of 3-chlorophe-
nol promoted papillomas when repeatedly applied to the backs of mice after
initiation with dimethylbenzanthrene (Boutwell and Bosch, 1959).
B. Mutagenicity, Teratogenicity and Other Reproductive Effects
Pertinent data cannot be located in the available literature re-
garding mutagenicity, teratogenicity and other reproductive effects.
C. Chronic Toxicity
Rats exposed 6 hrs/day for four months to 2 mg 4-chlorophenol/m
showed a temporary weight loss and increased myoneural excitability. Body
temperature and hematological parameters were not altered (Gurova, 1964).
In a survey comparing the health of workers, 4-chlorophenol exposed workers
had a significantly higher incidence of neurological disorders compared to1
f^.
unexposed workers in the same plant. Peripheral nerve stimulation studies
showed increased myoneural excitability in exposed workers. The minimum
detection distance in a two-point touch discrimination test was increased
(Gurova, 1964).
D. 'Other Relevant Information
f
3- and 4-Chlorophenol are weak uncouplers of oxidative phosphoryla-
tion (Mitsuda, et al. 1963; Weinback and Garbus, 1965).
-------
2,5-OICHLOROPHENOL, 2,6-OICHLORQPHENOL,
3,4-OICHLOROPHENOL, and 3,5-DICHLQRQPHENOL
II. EXPOSURE
Unspecified dichlorophenol isomers have been detected in concentrations
of 0.01 to 1.5 /jg/1 in Dutch surface waters (Piet and DeGrunt, 1975). Di-
chlorophenols have been found in flue gas condensates from municipal incin-
erators (Olie, et al. 1977). No data on exposure from foods or the dermal
route were found. Exposure to other chemicals can result in exposure to
dichlorophenols (i.e., dichlorobenzenes, lindane, and the alpha and delta
isomers of 1,2,3,4,5,6-hexachlorocyclohexane are metabolized by mammals to
dichlorophenols) (Kohli, et al. 1976; Foster and Saha, 1978).
III. PHARMACOKINETICS
Pharmacokinetic data specific to these dichlorophenol isomers could not
be located in the available literature.
IV. EFFECTS
A. .Carcinogenicity
Pertinent data cannot be located in the available literature.
B. Mutagenicity -
2,3-, 2,4-, 2,5-, 2,6-, 3,4-, and 3,5-Oichlorophenols were found to
be non-mutagenic in the Ames test with or without microsomal activation
(Rasaner and Hattula, 1977).
C. Teratogenicity, Other Reproductive Effects and Chronic Toxicity
Pertinent data cannot be located in the available literature re-
garding teratogenicity, other reproductive effects and chronic toxicity.
0. Other Relevant Information ^
• t.'
2,6-Dichlorophenol is a convulsant (Farquharson, et al. 1958). •
X
-N7/-
-------
TRICHLOROPHENOLS
II. EXPOSURE
Trichlorophenols have been detected in surface waters in Holland at
concentrations ranging from 0.003 to 0.1 >jg/l (Piet and DeGrunt, 1975).
2,4,5-Trichlorophenol can be formed from the chlorination .of phenol in water
(Burttschell, et al. 1959).
One possible source of trichlorophenol exposure for humans is through
the food chain, as a result of the ingestion by grazing animals of the
chlorophenoxy acid herbicides 2,4,5-T (2,4,5-trichlorophenoxyacetic acid)
and silvex (2-(2,4,5-trichlorophenoxy)-propionic acid). Residues of the
herbicides on sprayed forage are estimated to be 100-300 ppm. Studies in
which cattle and sheep were fed these herbicides at 300, 1000 and 2000 ppm
(Clark, et al. 1976) showed the presence of 2,4,5-trichlorophenol in various
tissues. In lactating cows fed 2,4,5-T at 100 ppm, an occasional residue of
0.06 ppm or less of trichlorophenol was detected in milk (Bjerke, et al.
1972).
Exposure to other chemicals such as trichlorobenzenes, lindane, the
alpha and delta isomers of 1,2,3,4,5,6-hexachlorocyclohexane, isomers of
benzene hexachloride, and the insecticide Ronnel can result in exposure to
trichlorophenols via metabolic degradation of the parent compound (U.S. EPA,
1979).
The U.S. EPA (1979) has estimated the weighted average bioconcentration
factors for the edible portions of all aquatic organisms consumed by Ameri-
cans to be 130 for 2,4,5-trichlorophenol and 110; for 2,4,6-trichlorophenol.
F
These estimates are based on the octanol/water. partition coefficients for
these chemicals.
-------
Trichlorophenols are found In flue gas condensates from municipal in-
cinerators (Olie, et al. 1977). 2,4,5-Trichlorophenol was detected in 1.7
percent of urine samples collected from the general population (Kutz, et al.
1978).
III. PHARMACOKINETICS
A. Absorption and Distribution
Information dealing with tissue distribution after administration
of trichlorophenols could not be located in the available literature. Feed-
ing of 2,4,5-T and silvex to sheep and cattle produced high levels of 2,4,5-
trichlorophenol in liver and kidney and low levels in muscle and fat (Clark,
et al. 1976).
B. Metabolism
Pertinent data could not be located in the available literature.
C. 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 5 days after the last dose (Korte, et al. 1978). The ap-
proximate blood half-life for 2,4,5-trichlorophenol is 20 hours, after dos-
ing of sheep with Erbon (an herbicide which is metabolized to 2,4,5-tri-
chlorophenol) (Wright, et al. 1970).
IV. EFFECTS
A. Carcinogenicity
A 21 percent solution of 2,4,5-trichlorophenol in acetone promoted
f
papillomas but not carcinomas in mica after initiation with dimethylbenzan-
-------
threne (Boutwell and Bosch, 1959). 2,4,6-Trichlorophenol showed no promot-
ing activity.
Results from a study of mice receiving 2,4,6-trichlorophenol in the
diet throughout their lifespans (18 months) were inconclusive. The inci-
dence of tumors, while higher than that for compounds classified as noncar-
cinogens, was not significantly increased (Innes, et al. 1969).
B. Mutagenicity
2,4,6-Trichlorophenol (400 mg) increased the mutation rate but not
the rate of intragenic recombination in a strain of Saccharomyces cerevisiae
(Fahrig, et al. 1978). Two of the 340 offspring from mice injected with 50
mg/kg of 2,4,6-trichlorophenol during gestation were reported to have
changes in hair coat color (spots) of genetic significance. At 100 mg/kg, 1
out of 175 offspring had a spot (U.S. EPA, 1979). 2,3,5-, 2,3,6-, 2,4,5-,
and 2,4,6-Trichlorophenol were found to be nonmutagenic in the Ames test
with and without microsomal activation (Rasanen and Hattula, 1977).
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
When rats were fed 2,4,5-trichlorophenol (99 percent pure) for 98
days (McCollister, et al. 1961), levels of 1000 mg trichlorophenol/kg feed
(assumed to be equivalent to 100 mg/kg body weight) or less produced no
adverse effects as judged by behavior, mortality, food consumption, growth,
terminal hematology, body and organ weights, and gross or microscopic patho-
f
logy. At 10,000 mg/kg diet (1000 mg/kg body weight), growth was slowed in
*
females. Histopathologic changes were noted in liver and kidney. There
-------
were no hematologic changes. At 3000 mg/kg feed (300 mg/kg body weight),
milder histopathologic changes in liver and kidney were observed. The his-
topathologic changes were considered to be reversible.
Adverse health effects including chloracne, hyperpigmentation, hir-
sutism and elevated uroporphyrins were described in 29 workers involved in
the manufacture of 2,4-D and 2,4,5-T (Bleiberg, et al. 1964). It is likely
that some of these symptoms represent 2,3,7,8-tetrachlorodibenzo-p-dioxin
toxicosis (U.S. EPA, 1979).
E. Other Relevant Information
Trichlorophenols are uncouplers of oxidative phosphorylation (Wein-
back and Garbus, 1965; Mitsuda, et al. 1963).
TETRACHLOROPHENQL
II. EXPOSURE
There are three isomers of tetrachlorophenol: 2,3,4,5-, 2,3,5,6-, and,
most importantly, 2,3,4,6-tetrachlorophenol. Commercial pentachlorophenol
contains three to 10 percent tetrachlorophenol (Goldstein, et al. 1977;
Schwetz, et al. 1978). Commercial tetrachlorophenol contains pentachloro-
phenol (27 percent) and toxic nonphenolic impurities such as chlorodibenzo-
furans and chlorodioxin isomers (Schwetz, et al. 1974). There are reports
suggesting the presence of lower chlorophenols in drinking water, but the
presence of tetrachlorophenol has not been documented (U.S. EPA, 1979).
Exposure to other chemicals such as tetrachlorobenzenes can result in expo-
sure to tetrachlorophenols via degradation of the'-parent compound (Kohli, et
r
al. 1976). -V:
»
Data could not be located in the available literature on ingestion from
foods. The U.S. EPA (1979) has estimated a weighted average bioconcentra-
i
-W75"-
-------
tion factor for 2,3,4,6-tetrachlorophenol of 320 for the edible portion of
aquatic organisms consumed by Americans. This estimate is based on the
octanol/water partition coefficient of 2,3,4,5-tetrachlorophenol.
Tetrachlorophenols have been found in flue gas condensates from munici-
pal incinerators (Olie, et al. 1977).
II. PHARMACOKINETICS
A. Absorption and Distribution
Pertinent data could not be located in the available literature
regarding absorption and distribution.
8. Metabolism and Excretion
In rats, over 98 percent of an intraperitoneally administered dose
of 2,3,5,6-tetrachlorophenol was recovered in the urine in 24 hours. About
66 percent was excreted as the unchanged compound and 35 percent as tetra-
chloro-p-hydroquinone. About 94 percent of the intraperitoneal dose of
• 2,3,4,6-tetrachlorophenol was recovered in the urine in 24 hours, primarily
as the unchanged compound with trace amounts of trichloro-p-hydroquinone.
Fifty-one percent of the intraperitoneal dose of 2,3,4,5-tetrachlorophenol
was recovered in the urine in 24 hours, followed by an additional seven per-
cent in the second 24 hours, primarily as the unchanged compound with trace
f*.
amounts of trichloro-p-hydroquinone. In these experiments, the urine was
boiled to split any conjugates (Alhborg and Larsson, 1978).
IV. EFFECTS
A. Carcinogenicity
Pertinent data could not be located in the available literature.
B. Mutagenicity
2,3,4,6-Tetrachlorophenol was reported to be nonmutagenic in 'the
Ames test, both with and without microsomal activation (Rasanen, et al.
1977).
-H76-
-------
C. Teratogenicity
Tetrachlorophenol did not induce teratogenic effects in rats at
doses of 10 or 30 mg/kg administered on days six through 15 of gestation
(Schwetz, et al. 1974).
D. Other Reproductive Effects
Tetrachlorophenol produced fetotoxic effects (subcutaneous edema
and delayed ossification of skull bones) in rats at doses of 10 and 30 mg/kg
administered on days six through 15 of gestation (Schwetz, et al. 1974).
E. Chronic Toxicity
Workers exposed to wood dust containing 100-800 ppm 2,3,4,6-tetra-
chlorophenol, 30-40 ppm pentachlorophenol, 10-50 ppm chlorophenoxyphenols,
1-10 ppm chlorodibenzofurans and less than 0.5 ppm chlorodibenzo-p-dioxins
developed severe skin irritations, respiratory difficulties and headaches
(Levin, et al. 1976).
No toxicity studies of 90 days or longer were found in the avail-
able literature.
F. Other Relevant Information
2,3,4,6-Tetrachlorophenol is a strong uncoupler of oxidative phos-
phorylation (Mitsuda, et al. 1963; Weinback and Garbus, 1965).
CHLOROCRESOLS
II. EXPOSURE
There are no published data available for the determination of current
human exposure to chlorocresols (U.S. EPA, 1979).,- p-Chloro-m-cresol (4-
chloro-3-methylphenol) has been detected in chldrinated sewage treatment
* •*-'.
effluent (Jolley, et al. 1975). Another potential source of chlorocresol's
is the herbicide MCPA (4-cnloro-2-methylphenoxyacetate), which (in its tech-
irf -
-------
nical grade) is contaminated with four percent 4-chloro-o-cresol (Rasanen,
et al. 1977) and which can be degraded to 5-chloro-o-cresol (Gaunt and
Evans, 1971).
III. PHARMACOKINETICS
A. Absorption
Chlorocresol (unspecified isomer) permeated human autopsy skin more
readily than either 2- or 4-chlorophenol, but less readily than 2,4,6-tri-
chlorophenol (Roberts, et al. 1977).
B. Distribution and Metabolism
Pertinent data could not be located in the available literature.
C. Excretion
Fifteen to 20 percent of a subcutaneous dose of p-chloro-m-cresol
given to a rabbit was recovered in the urine. The same compound given in-
tramuscularly was not recovered in the urine to any appreciable extent (Zon-
dek and Shapiro, 1943).
IV. EFFECTS
A. Carcinogenicity
Pertinent information could not be located in the available litera-
ture.
B. Mutagenicity
3-Chloro-o-cresol, 4-chloro-o-cresol and 5-chloro-o-cresol were re-
ported to be nonmutagenic in the Ames test, with and without microsomal ac-
tivation (Rasanen, et al. 1977).
C. "Teratogenicity and Other Reproductive Effects
Pertinent data could not be located/ in the available literature
regarding tertogenicity and other reproductive effects.
3d
-W7S1-
-------
0. Chronic Toxicity
No information on chronic toxicity in humans or toxicity studies of
90 days or longer in experimental animals were presented in the Ambient
Water Quality Criteria Document (U.S. EPA, 1979). p-Chloro-m-cresql given
subcutaneously to young rats for 14 days (80 mg/kg/day) produced mild in-
flammation at the injection site but did not affect growth or produce le-
sions in kidney, liver, or spleen (Wien, 1939). Rabbits (weighing 1.5-2.3
kg) injected subcutaneously with 12.5 mg p-chloro-nt-cresol/day suffered no
obvious ill effects (Wien, 1939). Liver and kidney were normal histologic-
ally.
E. Other"Relevant Information
Chlorocresol, a preservative in heparin solutions, caused several
cases of generalized and local reactions (Hancock and Naysmith, 1975; Ain-
ley, et al. 1977). Systemic reactions included collapse, pallor, sweating,
hypotension, tachycardia and rashes. Trichlorocresol is also a convulsant
(Eichholz and Wigand, 1931).
. CHLORINATED PHENOLS
I. AQUATIC TOXICITY
A. Acute Toxicity
*i
The acute toxicity of eight chlorophenols was determined in nine
bioassays. Acute 96-hour LC5Q values for freshwater fish ranged from 30
jug/1 for the fathead minnow, Pimephales prgmelas, for A-chloro-3-methylphe-
nol (U.S. EPA, 1972) to 9,040 jug/1 for the fathead minnow for 2,4,6-tri-
chlorophenol (Phipps, et al. manuscript). Among the freshwater inverte-
brates, Oaphnia magna was assayed with seven chlorophenols in eight 48-hour
static bioassays. Acute LC_Q values ranged from 290 pg/1 for 2,3,4,6-
tetrachlorophenol and 4-chloro-2-methylphenol to 6,040 ug/1 for 2,4,6-tri-
-HI79-
-------
chlorophenol (U.S. EPA, 1978). Acute 96-hour static LC50 values in the
sheepshead minnow ranged from 1,660 ^ig/1 for 2,4,5-trichlorophenol to 5,350
pg/1 for 4-chlorophenol. The only marine invertebrate species acutely
tested has been the mysid shrimp) Mysidopsis bahia , with acute 96-hour
static LC5Q values reported by the U.S. EPA (1978) as: 3,830 jjg/1 for
2,4,5-trichlorophenol; 21,900 ug/1 for 2,3,5,6-tetrachlorophenol, and 29,700
fjg/l for 4-chlorophenol.
B. Chronic Toxicity
No data other than that presented in the specific hazard profile
for 2-chlorophenol, 2,4-dichlorophenol, and pentachlorophenol were available
for freshwater organisms. An embryo-larval study provided a chronic value
of 180 ,ug/l for sheepshead minnows^ Cyprinodon variegatus? exposed to 2,4-
dichloro-6-fnethylphenol (U.S. EPA, 1978).
C. Effects on Plants
Effective concentrations for 15 tests on four species of freshwater
plants ranged from chlorosis LC5Q of 603 pq/1 for 2,3,4,6-tetrachlorophe-
nol to 598,584 /ug/1 for 2-chloro-6-fnethylphenol in the duckweed, lemna minor
(Blackman, et al. 1955). The marine algae^ Skeletonema costatum has been
used to assess the relative toxicities of three chlorinated phenols. Effec-
tive concentrations, based on chlorophyll a content and cell growth, of 440
and 500 jug/1 were obtained for 2,3,5,6-tetrachlorophenol. 2,4,5-Trichloro-
phenol and 4-chlorophenol were roughly two and seven times as potent, re-
spectively, as 2,3,5,6-tetrachlorophenol.
0. Residues
Steady-state bioconcentration factors ..have not been calculated for
if
the chlorinated phenols. However, based upon "octanol/water partition coef-
ficients, the following bioconcentration factors have been estimated for
-------
aquatic organisms with a lipid content of eight percent: 41 for 4-chloro-
phenol; 440 for 2,4,5-trichlorophenol; 380 for 2,4,6-trichlorophenol; 1,100
for 2,3,4,6-tetrachlorophenol; and 470 for 4-chloro-3-methylphenol.
E. Miscellaneous
The tainting of fish flesh by exposure of rainbow trout, .Salmo
qairdneri i to various chlorinated phenols has derived a range of estimated
concentrations not impairing the flavor of cooked fish from 15 ug/1 for
2-chlorophenol to 84 jjg/1 for 2,3-dichlorophenol (Schulze, 1961; Shumway and
Palensky, 1973).
II. 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. Draft criteria recommended for chlorinated phenols by the U.S. EPA
(1979) are given in the following table:
Draft Ambient Water Quality Criteria
Compound
Criterion from
Organoleptic
Effects
Criterion from
Toxicological
Data
Monochlorophenols
3-chlorophenol
4-chlorophenol
Dichlorophenols
2,5-dichlorophenol
2,6-dichlorophenol
Trichlorophenols
2,4,5-trichlorophenol
2,4,6-trichlorophenol
Tetrachlorophenol*
2,3,4,6-tetrachlorophenol
50 jjg/1
30 jug/1
3.0 ug/1
3.0 ug/1
10 ,ug/l
100 ug/1
915 ug/1 •
none
none
none
none
1600 jjg/1
263 jjg/1
-------
Chlorocresol
Insufficient data on which
to base a criterion
*The criterion will be based on toxicological effects (U.S. EPA, 1979).
B. Aquatic
The proposed draft criterion for 2,4,6-trichlorophenol is 52 pg/1,
not to exceed 150 jjg/1 in freshwater environments. No additional criterion
for other chlorinated phenols can presently be derived for either freshwater
or marine organisms because of insufficient data (U.S. EPA, 1979).
-------
CHLORINATED PHENOLS
REFERENCES
Ahlborg, U.G. and K. Larsson. 1978. Metabolism of tetrachlorophenols in
the rat. Arch. Toxicol. 40: 63.
Ainley, E.J., et al. 1977. Adverse reaction to chlorocresol-preserved
heparin. Lancet 1803: 705.
Bjerke, E.L., et al. 1972. Residue study of phenoxy herbicides in milk and
cream. Jour. Agric. Food Chem. 20: 963.
Blackman, G.E., et al. 1955. The physiological activity of substituted
phenols. I. Relationships between chemical structure and physiological
activity. Arch. Biochem. Biophys. 54: 45.
Bleiberg, J., et al. 1964. Industrially acquired porphyria. Arch. Derma-
tol. 89: 793.
Boutwell, R.K. and O.K. Bosch. 1959. The tumor-promoting action of phenol
and related compounds for mouse skin. Cancer REs. 19: 413.
Burttschell, R.H., et al. 1959. Chlorine derivatives of phenol causing
taste and odor. Jour. Am. Water Works Assoc. 51: 205.
Clark, O.E., et al. 1976. Residues of chlorophenoxy acid herbicides and
their phenolic metabolites in tissues of sheep and cattle. Jour. Agric.
Food Chem. 23: 573.
Eichholz, F. and R. Wigand. 1931. Uber die wirkung von darmdesinfektion
smilleln. Eingegangen. 159: 81.
Fahrig, R. et al. 1978. Genetic activity of chlorophenols and chlorophenol
impurities. Pages 325-338 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.
Foster, T.S. and J.G. Saha. 1978. The in vitro metabolism of lindane by an
enzyme preparation from chicken liver. Jour. Environ. Sci. Health 13: 25.
Gaunt, J.K. and W.C. Evans. 1971. Metabolism of 4-chlor-2-methylphenoxy-
acetate by a soil pseudomonad. Biochem. Jour. 122: 519.
f
Goldstein, J.A., et al. 1977. Effects of•'pentachlorophenol on hepatic
drug-metabolizing enzymes and porphyria related" to contamination with ahlor-
inated dibenzo-p-dioxins and dibenzofurans. Biochem. Pharmacol. 26: 1549.
Gurova, A.I. 1964. Hygienic characteristics of p-chlorophenol in the ani-
line dye industry. Hyg. Sanita. 29: 46.
-------
Hancock, 8.W. and A. Naysmith. 1975. Hypersensitivity of chlorocresol
preserved heparin. Br. Med. Jour. 746.
Innes, J.R.M., et al. 1969. Sioassay of pesticides and industrial chemi-
cals for tumorigenicity in mice: A preliminary note. Jour. Natl. Cancer
Inst. 42: 1101.
Jolley, R.L., et al. 1975. Analysis of soluble organic constituents in
natural and process waters by high-pressure liquid chromatography. Trace
Subs. Environ. Hlth. 9: 247.
Karpow, G. 1893. On the antiseptic action of three isomer chlorophenols
and of their salicylate esters and their fate in the metabolism. Arch. Sci.
Bid. St. Petersburg. 2: 304. Cited by W.F. von Oettingen, 1949.
Kohli, J., et al. 1976. The metabolism of higher chlorinated benzene iso-
mer s. Can. Jour. Biochem. 54: 203.
Korte, I., et al. 1976. Studies on the influences of some environmental
chemicals and their metabolites on the content of free adenine nucleotides,
intermediates of glycolysis and on the activities of certain enzymes of
bovine lenses in vitro. Chemosphere 5: 131.
Kutz, F.W., et al. • 1978. Survey of pesticide residues and their metabo-
lites in urine from the general population. Pages 363-369 _In: K. Rango Rao,
ed. Pentachlorophenol: Chemistry, pharmacology and environmental toxico-
logy, Plenum Press, New York.
Levin, J.Q., et al. 1976. Use of chlorophenols as fungicides in sawmills.
Scand. Jour. Work Environ. Health 2: 71.
Lindsay-Smith, Jr., et al. 1972. Mechanisms of mammalian hydroxylation:
Some novel metabolites of chlorobenzenes. Xenobiotica 2: 215.
McCollister, D.O., et al. 1961. Toxicologic information on 2,4,5-tri-
chlorophenol. Toxicol. Appl. Pharmacol. 3: 63.
Mitsuda, H., et al. 1963. Effect of chlorophenol analogues on tFie oxida-
tive phosphorylation in rat liver mitochondria. Agric. Biol. Chem. 27: 366.
Olie, K., et al. 1977. Chlorodibenzo-p-dioxins and chlorodibenzoflurans
are trace components of fly ash and flue gas - of some municipal incinerators
in the Netherlands. Chemosphere 8: 445.
Phipps, G.L., et al. The acute toxicity of phenol and substituted phenols
to the fathead minnow. (Manuscript)
Piet, G.J. and F. DeGrunt. 1975. Organic chi'oro 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, Luxem-
bourg.
Rasanen, L. and M.L. Hattula. 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.
-------
Rasanen, L., et al. 1977. The mutagenicity of MCPA and its soil metabo-
lites, chlorinated phenols, catechols and some widely used slimicides in
Finland. Bull. Environ. Contam. Toxicol. 18: 565.
Roberts, M.S., et al. 1977. Permeability of human epidermis to phenolic
compounds. Jour. Pharm. Pharmac. 29: 677.
Schulze, E. 1961. The effect of phenol-containing waste on the taste of
fish. Int. Revue Ges. Hydrobiol. 46, No. 1, p. 81.
Schwetz, B.A., et al. 1974. Effect of purified and commercial grade tetra-
chlorophenol on rat embryonal and fetal development. Toxicol. Appl. Pharma-
col. 28: 146.
Shumway, D.L. and J.R. Palensky. 1973. Impairment of the flavor of fish by
water pollutants. EPA-R3-73-010. U.S. Environ. Prot. Agency, U.S. gov-
ernment Printing Office, Washington, D.C.
U.S. EPA. 1972. The effect of chlorination on selected organic chemicals.
Water Pollut. Control Res. Ser. 12020.
U.S. EPA. 1978. In-depth studies on health and environmental impacts on
selected water pollutants. Contract No. 68-01-4646.
U.S. EPA. 1979. Chlorinated Phenols: Ambient Water Quality Criteria.
(Draft)
Weinbach, E.C. and J. Garbus. 1965. The interaction of uncoupling phenols
with mitochondria and with mitochondrial protein. Jour. Biol. Chem.
210: 1811.
Wien, R. 1939. The toxicity of parachlorometacresol and of phenylmercuric
nitrate. Quarterly Jour, and Yearbook of Pharmacy. 12: 212.
Wright, F.C., et al. 1970. Metabolic and residue studies with 2-(2,4,5-
trichlorophenoxy)-ethyl 2,2-dichloropropionate. Jour. Agric. Food Chem.
18: 845.
Zondek, B. and B. Shapiro. 1943. Fate of halogenated phenols in the or-
ganism. Biochem. Jour. 37: 592.
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No. 40
Chloroacetaldehyde
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.
-------
CHLOROACETALDEHYDE
Summary
No carcinogenic effects were observed in female ICR Ha Swiss mice follow-
ing administration of chloroacetaldehyde via dermal application or subcutaneous
injection. Mutagenic effects, varying from weak to strong, have been reported
in the yeasts Schizosaccharomyces pombe and Saccharomyces cerivisiae and in
certain Salmonella bacterial tester strains. There is no evidence in the
available literature to indicate that chloroacetaldehyde produces teratogenic
».
effects. Occupational exposure studies have shown chloroacetaldehyde to be a
severe irritant of the eyes, mucous membranes and skin.
Data concerning the effects of chloroacetaldehyde on aquatic organisms
were not found in the available literature.
-------
CHLOROACETALDEHYOE
I. INTRODUCTION
Chloroacetaldehyde (C^H^CIO) is a clear, colorless liquid with a pungent
odor. Its physical properties include: boiling point, 90.0-100.1°C (40 per-
cent sol.); freezing point, -16.3°C (40 percent sol.); and vapor pressure, 100
mm at 45°C (40 percent sol.). . Synonyms for Chloroacetaldehyde are:
monochloroacetaldehyde, 2-chloroacetaldehyde and chloroaldehyde. It is soluable
in water, acetone and methanol. Primary uses of Chloroacetaldehyde include:
*.
use as a fungicide, use in the manufacture of 2-aminothiazole, and use in the
removal of bark from tree trunks.
II. EXPOSURE
No monitoring data are available to indicate ambient air or water levels
of Chloroacetaldehyde, nor is any information available on possible exposure
from food.
Occupational routes of human exposure to Chloroacetaldehyde are primarily
through inhalation and skin absorption.
Bioaccumulation data on Chloroacetaldehyde were not found in the available
literature. However, 2-chloroacetaldehyde is known to be a chemically reactive
compound and its half-life in aqueous solution has been reported as slightly
greater than 24 hours (Van Duuren et al., 1972).
III. PHARMACOKINETICS
A. Absorption
Exposure to Chloroacetaldehyde is primarily through inhalation and
skin absorption.
Chloroacetaldehyde proved to be very lethal by inhalation. In an ijihalation
study conducted by Lawrence et al. (1972), mice were placed in a chloroacetaldehyde-
free chamber and air containing Chloroacetaldehyde vapor was then passed
-------
through the chamber. The time of exposure required to kill 50% of the animals,
LT^g, was 2.57 min. (the chamber atmosphere was calculated to have reached 45%
equilibrium within that time.)
In comparison studies conducted on chloroacetaldehyde and 2-chloroethanol,
chloroacetaldehyde was reported as exhibiting greater irritant activity, but
having lesser penetrant capacity (Lawrence et al., 1972).
B. Distribution
Information on the distribution of. chloroacetaldehyde was not found
in the available literature.
C. Metabolism
Chloroacetaldehyde appears to be a metabolite of a number of compounds
including 1,2-dichloroethane, chloroethanol and vinyl chloride (McCann et al.,
1975).
Johnson (1967) conducted in vitro studies on rat livers, the results of
which indicated that S-carboxymethylglutathione was probably formed via
chloroacetaldehyde metabolic action. Based upon these studies, Johnson suggested
that the same metabolic mechanism was operative in the i_n vivo conversion of
chloroethanol to S-carboxymethylglutathione.
In recent studies, Watanabe et al. (1976a,b) reported that chloro-
acetaldehyde would conjugate with glutathione and cysteine leading ultimately
to the types of urinary metabolites found in animals exposed to vinyl chloride.
The authors reported that as nonprotein free sulfhydral concentrations are
depleted, the alkylating metabolites, one of which is chloroacetaldehyde, are
likely to react with protein, DMA and RNA, eliciting proportionally greater
toxicity. This is in agreement with other studies conducted on vinyl chloride
•
metabolism (Hefner et al., 1975; Bolt et al., 1977).
-------
Chloroacetaldehyde was shown to cause the destruction of lung hemoprotein,
cytochrome P450, as well as liver microsomal cytochrome P450, with no requirement
for NADPH (Harper and Patel, 1978). The results suggested that the aldehydes
tested, one of which was Chloroacetaldehyde, were the toxic intermediates
which inactivated pulmonary enzymes following exposure to some environmental
agents.
D. Excretion
Information specifically on the rates and routes of Chloroacetaldehyde
elimination was not found in the available literature. Studies on vinyl
chloride and ethylene dichloride, however, indicate that Chloroacetaldehyde,
as an intermediate metabolite, may ultimately convert to a number of urinary
metabolites—including chloroacetic acid, S-carboxymethylcysteine and thiodiacetic
acid—depending on the particular metabolic pathway involved in the biotransforma-
tion of the parent compound (Johnson, 1967; Yllner, 1971; Watanabe, 1976a,b).
IV. EFFECTS
A. Carcinogenicity
In a study on the carcinogenic activity of alkylating agents, Van
Duuren et al. (1974) exposed female ICR Ha Swiss mice to 2-chloroacetaldehyde
(assayed as diethylacetal). The routes of administration were via skin and
subcutaneous injection. The authors reported no significant tumor induction.
Later studies confirmed these findings (Goldschmidt, personal communication,
1977). However, in a report by McCann et al. (1975), the authors stated that
previous reports of changes of respiratory epithelium in lungs of rats exposed
to Chloroacetaldehyde were suggestive of premalignant conditions.
B. Mutagenicity
Many studies have been reported which show that Chloroacetaldehyde
exhibits varying degrees of mutagenic activity (Huberman et al., 1975; Border
-------
and Webster, 1976; Elmore et al., 1976; Rosenkranz, 1977). Loprieno et al,
(1977) reported that 2-chloroacetaldehyde showed only feeble genetic activity
when tested in the yeasts Schizosaccharomyces pombe and Saccharomyces cerevisiae.
However, McCann et al. (1975) reported that chloroacetaldehyde was quite
effective in reverting Salmonella bacterial tester strain TA 100, but did not
revert TA 1535. In a later study, Rosenkranz (1977) found that
2-chloroacetaldehyde did display some mutagenic activity towards TA 1535.
In a study conducted by Elmore et al. (1976) the authors reported that
•_
the chloroacetaldehyde monomer and monomer hydrate were more mutagenically
active that the dimer hydrate and the trimer.
Rannug et al. (1976) reported that the mutagenic effectiveness of
chloroacetaldehyde is about 10 times higher than expected from kinetic data.
C. Teratogenicity and Other Reproductive Effects
Pertinent information could not be found in the available literature.
D. Chronic Toxicity
No chronic information could be found in the available literature.
However, extensive toxicity studies conducted by Lawrence et al. (1972) revealed
some subacute effects of chloroacetaldehyde on Sprague-Dawley and Black Bethesda
rats. Groups of rats received .001879 and ,003758 ml/kg of chloroacetaldehyde
(representing 0.3 and 0.6 of the acute LD5Q dose, respectively) daily for 30
consecutive days. Hematologic tests at the end of 30 days showed that there
was a significant decrease in hemoglobin, hematocrit, and erthrocytes in the
high dose group; the low dose group showed an increase in monocytes accompanied
by a decrease in lymphocytes. The animals were sacrificed and organ-to-body
weight ratios were calculated. Ratios for both brain and lungs were sig/iificantly
greater in the low dose group, while the high dose group showed a significant
increase in the brain, gonads, heart, kidneys, liver, lungs and spleen.
-------
Histologies! examination did not reveal any abnormalities attributable to
chloroacetaldehyde except for the lungs which showed more severe bronchitis,
bronchiolitis and bronchopneumonia than were seen in controls.
In another subacute (subchronic) study, chloroacetaldehyde was administered
to rats in doses of .00032, .00080, .00160 and .00320 ml/kg, three times a
week for 12 weeks. Hematologic determinations showed no significant differences
between controls and the two lower dose groups, while animals administered
.0016 ml/kg showed a decrease in red cell count and lymphocytes and an increase
V
in segmented neutrophiles; the highest dose group showed a significant decrease
in red blood cells and hemoglobin with an increase in clotting time and segmented
neutrophiles. Organ-to-body weight ratios were determined for several organs
and, although there were some significant differences from controls, there
were no apparent dose-related responses.
D. Acute Toxicity
Lawrence et al. (1972) conducted a series of acute toxicity tests on
ICR mice, Sprague-Dawley and Black Bethesda rats, New Zealand albino rabbits
and Hartlez strain guinea pigs. The results were reported as follows: the
LD5Qs (ml/kg) for chloroacetaldehyde administered intraperitoneally ranged
from .00598 in mice to .00464 in rabbits; the LD50s (ml/kg) for chloroacetaldehyde
administered intragastrically were reported as .06918 in male mice, .07507 in
female rats and .08665 in male rats; the dermal LD5Q (ml/kg) in rabbits was
reported as .2243; and the inhalation U~5Q in mice was reported as 2.57 min.
E. Other Relevent Information
Case studies show that contact with a strong solution of chloroacetaldehyde
in the human eye will likely result in permanent impairment of vision and skin
contact with a potent solution will result in burns (Proctor and Hughes,
1978).
-493-
-------
V. AQUATIC TOXICITY
Data concerning the effects of chloroacetaldehyde on aquatic organisms
were not found in the available literature.
VI. EXISTING GUIDELINES
The 8-hour, TWA occupational exposure limit established for chloroacetaldehyde
is 1 ppm. This TLV of 1 ppm was set to prevent irritation (ACGIH, 1976).
si
-------
CHLOROACETALDEHYDE
References
1. American Industrial Hygiene Association. 1976. Threshold limit values
for substances in workroom air. 3rd ed. p. 48. Cincinnati. Cited in
Proctor and Hughes, 1978.
2. Bolt, H. M. et al . 1977. Pharmacokinetics of vinyl chloride in the rat.
Toxicology. 7:179.
3. Border, E. A. , and I. Webster. 1977. The effect of vinyl chloride
monomer, chloroethylene oxide and chloroacetaldehyde on ONA synthesis in
regenerating rat liver. Chem. Biol. Interact.-- 17:239.
4. Elmore, J. 0. et al . 1976. Vinyl chloride mutagenicity via the metabolites
chlorooxirane and chloroacetaldehyde monomer hydrate. Biochim. Biophys.
Acta. 442:405.
5. Harper, C. , and J. M. Patel. 1978. Inactivation of pulmonary cytochrome
P 450 by aldehydes. Fed. Proc. 37:767.
6. Hefner, R. E. , Jr. et al. 1975. Preliminary studies of the fate of inhaled
vinyl chloride monomer in rats. Ann. N.Y. Acad. Sci. 246:135.
7. Huberman, E. et al. 1975. Mutation induction in Chinese hamster V79
cells by two vinyl chloride metabolites, chloroethylene oxide and
. 2-chloroacetaldehyde. Int. J. Cancer. 16:639.
8. Johnson, M. K. 1967. Metabolism of chloroethanol in the rat. Biochem.
Pharmacol. 16:185.
9. Lawrence W. H. et al . 1972. Toxicity profile of chloroacetaldehyde. J.
Pharm. Sci. 61:19.
10. Loprieno, N. et al . 1977. Induction of gene mutations and gene conversions
by vinyl chloride metabolites in yeast. Cancer Res. 36:253.
11. McCann, J. et al . 1975. Mutagenicity of chloroacetaldehyde, a possible
metabolic product of 1,2-dichloroethane (ethylene dichloride), chloroethanol
(ethylene chlorohydrin) , vinyl chloride, and cyclophosphamide. Proc.
Nat. Acad. Sci. 72:3190.
12. Proctor, N. H. , and J. P. Hughes. 1978. Chemical hazards of the workplace.
p. 160. Lippincott Co., Philadelphia.
13. Rannug, U. et al. 1976. Mutagenicity of chloroethylene oxide,
chloroacetaldehyde, 2-chloroethanol and chloroacetic acid, conceivable
metabolites of vinyl chloride. Chem. Biol. Interact. 12:251,
14. Rosenkranz, H. S. 1977. Mutagenicity of halogenated alkanes and their
derivatives. Environ. Health Perspect. 21:79.
-H9S"-
-------
15. Van Quuren, B. L. et al. 1972. Carcinogen!city of halo-ethers. II.
Structure-activity relationships of analogs of bis-(chloromethyl) ether.
J. Nat. Cancer Inst. 48:1431.
16. Van Ouuren, B. L. et al. 1974. Carcinogenic activity of alkylating
agents. J. Nat. Cancer Inst. 53:695
17. Watanabe, P. G. et al. 1976a. Fate of 14C vinyl chloride after single
oral administration in rats. Toxicol. Appl. Pharmacol. 35:339.
Watanabe, P. G. et al. 1976b. Fate of 14C vinyl chloride folli
inhalation exposure in rats. Toxicol. App. Pharmacol. .37:49.
Yllner, S. 1970. Metaboli;
Pharmacol. Toxicol. 30:69.
Yllner, S. 1971. Metaboli;
Pharmacol. Toxicol. 30:257.
18. Watanabe, P. G. et al. 1976b. Fate of 14C vinyl chloride following
icol
19. Yllner, S. 1970. Metabolism of chloroacetate -1- C in the mouse. Acta
20. Yllner, S. 1971. Metabolism of l,2-dichloroethane-14C in the mouse. Acta
jef
-------
No. 41
Chloroalkyl Ethers
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.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (GAG) has evaluated
chloroalkyl ethers and has found sufficient evidence to
indicate that this compound is carcinogenic.
-------
CHLOROALKYL ETHERS
SUMMARY
Bis(chloromethyl)ether {BCME), chloromethyl methyl ether
(CMME), and bis{2-chloroethylJether (BCEE) have shown carcin-
ogenic effects in animal studies following administration by
various routes. Epidemiological studies in the United States,
Germany, and Japan have indicated that workers exposed to
BCME and CMME developed an increased incidence of respiratory
tract tumors.
Testing of BCME, CMME, BCEE, and bis(2-chloroisopropyl)-
ether (BCIE) in the Ames Salmonella assay and in E. coli have
indicated that these compounds have mutagenic activity. Cy-
togenetic studies of lymphocytes from workers exposed to BCME
and CMME have reported an increased frequency of aberrations,
which appear to be reversible.
There is no available evidence to indicate chloroalkyl
ethers produce adverse reproductive or teratogenic effects.
The information base for freshwater organisms and chloro-
alkyl ethers is limited to a few toxicity tests of 2-chloro-
ethyl vinyl ether and bis(2-chloroethyl)ether. The reported
96-hour LC5Q value for bis(2-chloroethyl)ether in the
bluegill is greater than 600,000 ug/1. A "no effect" value
of 19,000 ug/1 was observed using the fathead minnow in an
embryo-larval test. Bis{2-chloroethyl)ether has a reported
bioconcentration factor of 11 in a 14-day exposure to blue-
gills. The half-life is from four to seven days. The re-
ported 96-hour LC50 value for the bluegill and 2-chloro-
ethyl vinyl ether is 194,000 ug/1.
-------
CHLOROALKYL ETHERS
I. INTRODUCTION
This profile is based on the Ambient Water Quality
Criteria Document for chloroalkyl ethers (U.S. EPA, 1979).
The chloroalkyl ethers are compounds with a hydrogen
atom in one or both of the aliphatic ether chains substituted
by a chlorine atom. The chemical reactivity of these com-
pounds varies greatly, depending on the nature of the ali-
phatic groups and the placement of the chlorine atoms. The
most reactive compounds are those with short aliphatic groups
and those in which chlorine substitution is closest to the
ether oxygen (alpha-chloro) (U.S. EPA, 1979).
As an indication of their high reactivity, chloromethyl
methyl ether (CMME), bis(chloromethyl)ether (BCME), 1-chloro-
ethyl ethyl ester, and 1-chloroethyl methyl ether decompose
rapidly in water. The beta-chloroethers, bis(2-chloroethyl}-
ether (BCEE) and bis(2-chloroisopropyl}ether (BCIE) are more
stable in aqueous systems; they are practically insoluble in
water but miscible with most organic solvents (U.S. EPA,
1979).
The chloroalkyl ethers have a wide variety of industrial
and laboratory uses in organic synthesis, textile treatment,
the manufacture of polymers and insecticides, in the prepara-
tion of ion exchange resins, and as degreasing agents (U.S.
EPA, 1979).
While the short chain alpha-chloroalkyl ethers (BCME,.
CMME) are very unstable in aqueous systems, they appear to be
relatively stable in the atmosphere (Tou and Kallos, 1974).
Bis(chloromethyl)ether will form spontaneously in the pres-
-------
ence of both hydrogen chloride and formaldehyde (Frankel, et
al. 1974).
II. EXPOSURE
The beta-chloroalkyl ethers have been monitored in
water. Industrial effluents from chemical plants involved in
the manufacture of glycol products, rubber, and insecticides
may contain high levels of these ethers (U.S. EPA, 1979).
The highest concentrations in drinking water of bis{2-chloro-
ethyljether, bis(2-chloroisopropyl)ether, and bis-l,2-(2-
chloroethoxy)ethane (BCEXE) reported by the U.S. EPA (1975)
are 0.5, 1.58, and 0.03 ug/1, respectively. The average con-
centration of these compounds in drinking water is in the
nanogram range (U.S. EPA, 1979). Chloroalkyl ethers have
been detected in the atmosphere, and human inhalation expo-
sure appears to be limited to occupational settings.
The chloroalkyl ethers have not been monitored in food
(U.S. EPA, 1979). The betachloroalkyl ethers, because of
their relative stability and low water solubility, may have a
tendency to be bioaccumulated. The U.S. EPA (1979) has esti-
mated the weighted bioconcentration factor to be 25 for the
edible portions of fish and shellfish consumed by Americans.
This is based on the measured steady-state bioconcentration
studies in bluegills. Bioconcentration factors for BCME (31)
and BCIE (106) have been derived using a proportionality con-
stant related to octanol/water partition coefficients (U.S.
»
EPA, 1979). Dermal exposure for the chloroalkyl ethers has
not been determined (U.S. EPA, 1979).
t
-------
III. PHARMACOKINETICS
A. Absorption
Experiments with radio-labelled BCIE and BCEE in
female rats and monkeys have indicated that both compounds
are readily absorbed in the blood following oral administra-
tion (Smith, et al.f 1977; Lingg, et al., 1978). Pertinent
data could not be located in the available literature re-
trieved on dermal or inhalation absorption of the alkyl
ethers.
B. Distribution
Species differences in the distribution of radio-
labelled BCIE have been reported by Smith, et al. (1977).
Monkeys, as compared to rats, retain higher amounts of radio-
activity in the liver, muscle, and brain. Urine and expired
air from the rat contained higher levels of radioactivity
than those found in the monkey. Blood levels of BCIE in mon-
keys reached a peak within two hours following oral adminis-
tration and then declined in a biphasic manner (t^/2*s
= 5 hours and 2 days for the first and second phases, respec-
tively) .
C. Metabolism
The biotransformation of BCEE in rats following
oral administration appears to involve cleavage of the ether
linkage and subsequent conjugation (Lingg, et al., 1978).
Thiodiglycolic acid and chloroethanol-D-glucuronide were
identified as urinary metabolites of BCEE. Metabolites of
BCIS identified in the rat included l-chloro-2-propanol, pro-
pylene oxide, 2-(l-methyl-2-chloroethoxy)-propionic acid, and
carbon dioxide (Smith, et al., 1977).
-------
D. Excretion
BCEE administered orally to rats was excreted
rapidly, with more than 60 percent of the compound excreted
within 24 hours. Virtually all of this elimination was via
the urine (Lingg, et al., 1978).
IV. EFFECTS
A. Carcinogenicity
There are several studies with bis(chloromethyl)-
ether (BCME), chloromethyl methyl ether (CMME), and bis{2-
chloroethyl)ether (BCEE) that show carcinogenic effects.
BCME induced malignant tumors of the male rat respiratory
tract following inhalation exposure (Kuschner, et al.,
1975). Application of BCME and BCEXE to the skin of mice
produced skin tumors (Van Duuren, et al., 1968), while subcu-
taneous injection of BCME to newborn mice induced pulmonary
tumors (Gargus, et al., 1969).
Oral administration of bis(2-chloroethylJether (BCEE) to
mice has been shown to increase the incidence of hepatocellu-
lar carcinomas in males (Innes, et al., 1969).
Epidemiological studies of workers in the United States,
Germany, and Japan who were occupationally exposed to BCME
and CMME have indicated these compounds are human respiratory
carcinogens (U.S. EPA, 1979).
Both BCME and CMME have been shown to accelerate the
rate of lung tumor formation in Strain A mice following inha-
r
lation exposure (Leong, et al., 1971). BCME and BCEE have
shown tumor initiating activity for mouse skin, while CMME
showed only weak initiating activity (U.S. EPA, 1979).
-------
Preliminary results of a National Cancer Institute
study indicate that oral administration of BCIE does not pro-
duce an increase in tumor incidence (U.S. EPA, 1979).
B. Mutagenicity
Testing of the chloroalkyl ethers in the Ames Sal-
monella assay on E. coll have indicated that BCME, CMME,
BCIE, and BCEE all produced mutagenic effects (U.S. EPA,
1979). BCEE has also been reported to induce mutations in
Saccharomyces cerevisiae (U.S. EPA, 1979). Neither BCEE nor
BCIE showed mutagenic effects in the heritable translocation
test in mice (Jorganson, et al. 1977). An increase in cyto-
genetic aberrations in the lymphocytes of workers exposed to
BCME and CMME was reported by Zudova and Landa (1977); the
frequency of aberrations decreased following the removal of
workers from exposure.
C. Teratogenicity and Other Reproductive Effects
Pertinent data could not be located in the avail-
able literature.
D, Chronic Toxicity
Chronic occupational exposure to CMME contaminated
with BCME has produced bronchitis in workers (U.S. EPA,
1979). Cigarette smoking has been found to act synergisti-
cally with CMME exposure to produce bronchitis (Weiss, 1976,
1977) .
Animal studies have indicated that chronic exposure
to BCIE produces liver necrosis in mice. Exposure in rats'
causes major effects on the lungs, including congestion and
pneumonia (U.S. EPA, 1979).
-------
E. Other Relevant Information
The initiating activity of several chloroalkyl
ethers indicates that these compounds may interact with other
agents to produce skin papillomas (Van Duuren, et al., 1969,
1972}.
V. AQUATIC TOXICITY
A. Acute Toxicity
The reported static 96-hour LC50 value for the
bluegill (Lepomis macrochirus) with 2-chloroethyl vinyl ether
(concentration unmeasured) is 194,000 ug/1 (U.S. EPA, 1978).
The 96-hour LC5Q values for the bluegill could not be de-
termined in a static test for bis(2-chloroethyl)ether with
exposure concentrations as high as 600,000 ug/1- The concen-
tration of the ether was not monitored during the bioassay.
Pertinent data could not be located in the available litera-
ture on saltwater species.
B. Chronic Toxicity
An embryo-larval test was conducted with bis(2-
chloroethyDether and the fathead minnow, (Pimephales prome-
las). Adverse effects were not observed at test concentra-
tions as high as 19,000 ug/1.
C. Plant Effects
Pertinent data could not be located in the avail-
able literature.
D. Residues
Using bis(2-chloroethylJether, a bioconcentration
factor of 11 was determined during a 14-day exposure of blue-
gills (U.S. EPA, 1979). The half-life was observed to be
between four and seven days.
-SO4,-
-------
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
Based on animal carcinogenesis bioassays, and using
a linear, nonthreshold model, the U.S. EPA (1979) has esti-
mated the following ambient water levels of chloroalkyl
ethers which will produce an increased cancer risk of
10~5: BCIE, ll.Sug/lf BCEE, 0.42 ug/1? and BCME 0.02
ng/1.
Eight-hour TWA exposure values (TLV) for the fol-
lowing chloroalkyl ethers have been recommended by the Ameri-
can Conference of Governmental and Industrial Hygienists
(ACGIH, 1978): BCME, 1 ppb; BCEE, 5 ppm.
B. Aquatic
Freshwater and saltwater drafted criteria have not
been derived for any chloroalkyl ethers because of insuffi-
cient data (U.S. EPA, 1979).
-537-
-------
CHLOROALKYL ETHERS
REFERENCES
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.
Frankel, L.S., et al. 1974. Formation of bis (chloromethyl)
ether from formaldehyde and hydrogen chloride. Environ. Sci.
Technol. 8: 356.
Gargus, J.L., et al. 1969. Induction of lung adenomas in
newborn mice by bis{chloromethyl) ether. Toxicol. Appl.
Pharmacol. 15: 92.
Innes, J.R.M., et al. 1969. Bioassay of pesticides and in-
dustrial chemicals for tumorigenicity in mice: A preliminary
note. Jour. Natl. Cancer Inst. 42: 1101.
Jorgenson, T.A., et al. 1977. Study of the mutagenic poten-
tial of bis(2-chloroethyl) and bis (2-chloroisopropyl) ethers
in mice by the heritable translocation test. Toxicol. Appl.
Pharmacol. 41: 196.
Kuschner, M., et al. 1975. Inhalation carcinogenicity of
alpha halo esthers. III. Lifetime and limited period inhala-
tion studies with bis(chloromethyl)ether at 0.1 ppm. Arch
Environ. Health 30: 73.
Leong, B.K.J., et al. 1971. Induction of lung adenomas by
chronic inhalation of bis(chloromethylJether. Arch. Environ.
Health 22: 663.
Lingg, R.D., et al. 1978. Fate of bis(2-chloroethyl}ether
in rats after acute oral administration. Toxicol. Appl.
Pharmacol. 45: 248.
Smith, C.C., et al. 1977. Comparative metabolism of halo-
ethers. Ann. N.Y. Acad. Sci. 298: 111.
Tou, J.C., and G.J. Kallos. 1974. Kinetic study of the sta-
bilities of chloromethyl methyl ether and bis(chloromethyl)-
ether in humid air. Anal. Chem. 46: 1866.
U.S. EPA. 1975. Preliminary assessment of suspected carcin-
ogens in drinking water. Rep. Cong. U.S. Environ. Prot.
Agency, Washington, D.C.
U.S. EPA. 1978. In-depth studies on health and environmen-
tal impacts of selected water pollutants. U.S. Environ.
Prot. Agency, Contract No. 68-01-4646.
-------
U.S. EPA. 1979. Chloroalkyl Ethers: Ambient Water Quality
Criteria. {Draft}.
Van Duuren, B.L., et al. 1968. Alpha-haloethers: A new type
of alkylating carcinogen. Arch. Environ. Health 16: 472.
Van Duuren, B.L., et al. 1969. Carcinogenicity of halo-
ethers. Jour. Natl. Cancer Inst. 43: 481.
Van Duurenj B.L., et al. 1972. Carcinogenicity of halo-
ethers. II. Structure-activity relationships of analogs of
bis(chloromethyl)ether. Jour. Natl. Cancer Inst. 48: 1431.
Weiss, W. 1976. Chloromethyl ethers, cigarettes, cough and
cancer. Jour. Occup. Med. 18: 194.
Weiss, W. 1977. The forced end-expiratory flow rate in
chloromethyl ether workers. Jour. Occup. Med. 19: 611.
Zudova, Z., and K. Landa. 1977. Genetic risk of occupation-
al exposures to haloethers. Mutat. Res. 46: 242.
-509-
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No. 42
Chlorobenzene
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
-SVO-
-------
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.
-S//-
-------
CHLOROBENZENE
Summary
There is little data on the quantities of chlorobenzene in air, water
and food, although this compound has been identified in these media. Chron-
ic exposure to chlorobenzene appears to cause a variety of pathologies under
different experimental regimens; however, the liver and kidney appear to be
affected in a number of species. There have been no studies conducted to
evaluate the mutagenic, teratogenic, or carcinogenic potential of chloro-
benzene.
Four species of freshwater fish have 96-hour LC5Q values ranging from
24,000 to 51,620 ;jg/l. Hardness does not significantly affect the values.
In saltwater, a fish and shrimp had reported 96-hour LC5Q values of 10,500
jug/1 and 6,400 pg/1, respectively. No chronic data involving chlorobenzene
are available. Algae, both fresh and saltwater, are considerably less sen-
sitive to chlorobenzene toxicity than fish and invertebrates.
-------
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Chlorinated Benzenes (U.S. EPA, 1979).
Chlorobenzene, most often referred to as monochlorobenzene (MCB;
CgHjjCl; molecular weight 112.56), is a colorless liquid with a pleasant
aroma. Monochlorobenzene has a melting point of -45.6°c, a boiling point
of 131-132°c, a water solubility of 488 mg/1 at 25°C, and a density of
1.107 g/ml. Monochlorobenzene has been used as a synthetic intermediate in
the production of phenol, DDT, and aniline. It is also used as a solvent in
the manufacture of adhesives, paints, polishes, waxes, diisocyanates,
Pharmaceuticals and natural rubber (U.S. EPA, 1979).
Data on current production derived from U.S. International Trade Com-
mission reports show that between 1969 and 1975, the U.S. annual production
of monochlorobenzene decreased by 50 percent, from approximately 600 million
pounds to approximately 300 million pounds (U.S. EPA, 1977).
II. EXPOSURE
A. Water
Based on the vapor pressure, water solubility, and molecular weight
of Chlorobenzene, Mackay and Leinonen (1975) estimated the half-life of
evaporation from water to be 5.8 hours. Monochlorobenzene has been detected
in ground water, "uncontaminated" upland water, and in waters contaminated
either by industrial, municipal or agricultural waste. The concentrations
ranged from 0.1 to 27 jug/1, with raw waters having the lowest concentration
and municipal waste the highest (U.S. EPA, 1975, 1977). These estimates
should be considered as gross estimates of exposure, due to the volatile
nature of monochlorobenzene.
-SV3-
-------
8. Food
The U.S. EPA (1979) has estimated the weighted average bioconcen-
tration factor of monochlorobenzene to be 13 for the edible portions of fish
and shellfish consumed by Americans. This estimate was based on octanol/-
water partition coefficients.
C. Inhalation
Data have not been found in the available literature which deal
with exposure to chlorobenzene outside of the industrial working environment.
III. PHARMACOKINETICS
A. Absorption
There is little question, based on human effects and mammalian
toxicity studies, that chlorobenzene is absorbed through the lungs and from
the gastrointestinal tract (U.S. EPA, 1977).
8. Distribution
Because chlorobenzene is highly lipophilic and hydrophobic, it
would be expected that it would be distributed throughout total body water
space, with body lipid providing a deposition site (U.S. EPA, 1979).
C. Metabolism
Chlorobenzene is metabolised via an NADPH-cytochrome P-448 depen-
dent microsomal enzyme system. The first product, and rate limiting step,
is a epoxidation; this is followed by formation of diphenolic and monophe-
nolic compounds (U.S. EPA, 1979). Various conjugates of these phenolic
derivatives are the primary excretory products (Lu, et al. 1974). Evidence
indicates that the metabolism of monochlorobenzene results in the formation
of toxic intermediates (Kohli, et al. 1976). Brodie, et al. (1971) induced
*
microsomal enzymes with phenobarbital and showed a potentiationin in the
toxicity of monochlorobenzene. However, the • use of 3-methylcho-
-------
lanthrene to induce microsomal enzymes provided protection for rats (Oesch,
et al. 1973). The metabolism of chlorobenzene may also lead to the forma-
tion of carcinogenic active intermediates (Kohli, et al. 1976).
D. Excretion
The predominant route of elimination is through the formation of
conjugates of the metabolites of monochlorobenzene and elimination of these
conjugates by the urine (U.S. EPA, 1979). The types of conjugates formed
vary with species (Williams, et al. 1975). In the rabbit, 27 percent of an
administered dose appeared unchanged in the expired air (Williams, 1959).
IV. EFFECTS
Pertinent data could not be located in the available literature on the
carcinogenicity, mutagenicity, teratogenicity, or other reproductive effects
of chlorobenzene.
A. Chronic Toxicity
Data on the chronic toxicity of chlorobenzene is sparse and some-
what contradictory. "Histopathological changes" have been noted in lungs,
liver and kidneys following inhalation of monochlorobenzene (200, 475, and
1,000 ppm) in rats, rabbits and guinea pigs (Irish, 1963). Oral administra-
tion of doses of 12.5, 50 and 250 mg/kg/day to rats produced little patholo-
gical change, except for growth retardation in males (Knapp, et al. 1971).
B. Other Relevant Information
Chlorobenzene appears to increase the activity of microsomal NADPH-
cytochrome P-450 dependent enzyme systems. Induction of microsomal enzyme
activity has been shown to enhance the metabolism of a wide variety of
drugs, pesticides and other xenobiotics (U.S. EPA, 1979).
- SIS-
-------
V. AQUATIC TOXICITY
A. Acute Toxicity
Pickering and Henderson (1966) reported observed 96-hour LC5Q
values for goldfish, Carassius auratus, guppy, Poecilia reticulatus, and
bluegill, Lepomis macrochirus, to be 51,620, 45,530, and 24,000 jug/1, re-
spectively, for chlorobenzene. Two 96-hour LC^g values for chlorobenzene
and fathead minnows, Pimephales promelas, are 33,930 ug/1 in soft water (20
mg/1) and 29,120 jug/1 in hard water (360 mg/1), indicating that hardness
does not significantly affect the acute toxicity -of chlorobenzene (U.S. EPA,
1978). With uaphnia magna, an observed 48-hour EC^ value of 86,000 ug/1
was reported/ In saltwater studies, sheepshead minnow had a reported un-
adjusted LC5Q (96-hour) value of 10,500 jug/1, with a 96-hour EC5Q of
16,400 jjg/1 for mysid shrimp (U.S. EPA, 1978).
B. Chronic Toxicity
No chronic toxicity studies have been reported on the chronic
toxicity of chlorobenzene and any salt or freshwater species.
C. Plant Effects
The freshwater alga Selenastrum capricornutum is considerably less
sensitive than fish and Daphnia magna. Based on cell numbers, the species
has a reported 96-hour EC5Q value of 224,000 /jg/1. The saltwater alga,
Skeletonema costatum, had a 96-hour ECgQ, based on cell numbers of 341,000
Ajg/1.
D. Residues
A bioconcentration factor of 44 was obtained assuming an 8 percent
lipid content of fish.
4
-------
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 American Conference of Governmental Industrial Hygienists
(ACGIH, 1971) threshold limit value for chlorobenzene is 350 mg/rn . The
acceptable daily intake (ADI) was calculated to be-.1.003 mg/day. The U.S.
EPA (1979) draft water criterion for chlorobenzene is 20 pg/1, based on
threshold concentration for odor and taste.
B. Aquatic
For chlorobenzene, the drafted criterion to protect freshwater
aquatic life is 1,500 jug/1 as a 24-hour average; the concentration should
not exceed 3,500 ;jg/l at any time. To protect saltwater aquatic life, a
draft criterion of 120 jug/1 as a 24-hour average with a concentration not
exceeding 280 jug/1 at any time has been recommended (U.S. EPA, 1979).
-------
CHLOROBENZENE
REFERENCES
American Conference of Governmental Industrial Hygienists.
1971. Documentation of the threshold limit values for sub-
stances in workroom air. 3rd. Ed.
Brodie, B.B., et al. 1971. Possible mechanism of liver ne-
crosis caused by aromatic organic compounds. Proc. Natl.
Acad. Sci. 68: 160.
Irish, D.D. 1963. Halogenated hydrocarbons: II. Cyclic.
Tn"f Industrial Hygiene and Toxicology, Vol. II, 2nd Ed., .ed.
F.A. Patty , Interscience, New York. p. 1333.
Knapp, W.K., Jr., et al. 1971. Subacute oral toxicity of
monochlorobenzene in dogs and rats. Tofpxicol. Appl. Pharma-
col. 19: 393.
Kohli, I., et al. 1976. The metabolism of higher chlori-
nated benzene isomers. Can. Jour. Biochem. 54: 203.
Lu, A.Y.H., et al. 1974. Liver microsomal electron trans-
port systems . III. Involvement of cytochrome bg in the
NADH-supported cytochrome p^-450 dependent hydroxylation of
chlorobenzene. Biochem. Biphys. Res. Comm. 61: 1348.
Mackay, D., and P.J. Leinonen. 1975. Rate of evaporation of
•low-solubility contaminants from water bodies to atmosphere.
Environ. Sci. Technol. 9: 1178.
Oesch, F., et al. 1973. Induction activation, and inhibition
of epoxide hydrase. Anomalous prevention of chlorobenzene-
induced hepatotoxicity by an inhibitor of epoxide hydrase.
Chem. Biol. interact. 6: 189.
Pickering, Q.H., and C. Henderson. 1966. Acute toxicity of
some important petrochemicals to fish. Jour. Water Pollut.
Control Fed. 38: 1419.
U.S. EPA. 1975. Preliminary assessment of suspected carcin-
ogens in drinking water. Report to Congress. Environ.
Prot. Agency, Washington, D.C.
U.S. EPA. 1977. Investigation of selected potential envi-
ronmental contaminants: Halogenated benzenes. EPA 560/2-77-
004.
U.S. EPA. 1978. In-depth studies on health and environmen-
tal impacts of selected water pollutants. U.S. Environ.
Prot. Agency, Contract No. 68-01-4646.
-------
U.S. EPA. 1979. Chlorinated Benzenes: Ambient Water Quality
Criteria (Draft).
Williams, R.T. 1959. The metabolism of halogenated aromatic
hydrocarbons. Page 237 in Detoxication mechanisms. 2nd ed.
John Wiley and Sons, New York.
Williams, R.T., et al. 1975. Species variation in_the meta-
bolism of some organic halogen compounds. Page 91 An* A.D.
Mclntyre and C.F. Mills, eds. Ecological and toxicological
research. Plenum Press, New York.
-------
No, A3
p-Chloro-m-cresol
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.
-------
9
p-CHLORO-m-CRESOL
I ' '
SUMMARY .
p-Chloro-m-cresol has been found to be susceptible to biodegradation
under aerobic conditions in a synthetic sewage sludge. It has been found
to be formed by the chlorination of waters receiving effluents from electric
power-generating plants and by the chlorination of the effluent from a
domestic sewage treatment facility.
Very little, information on the health effects of p-chloro-m-cresol
was located. p-Chloro-m-cresol has been characterized as very toxic
in humans, although support for this statement is 'limited. In rats, a
a subcutaneous U>50 of 400 mg/kg and an oral LDLo of 500 mg/kg have been
J reported.
I. INTRODUCTION
I
';? p-Chloro-m-cresol (4-chloro-3-methylphenol; C H CIO; molecular
[j weight 142.58) is a solid (dimorphous crystals) at room temperature. The
pure compound is odorless, but it has a phenolic odor in its most common, impure
form. Its melting point is 55.5 C and its boiling point is 235°C.
It is soluble in water and many organic solvents (Windholz 1976).
A review of the production range (includes importation) statistics
for p-chloro-m-cresol (CAS No. 59-50-7) as listed in the initial TSCA
Inventory (U.S. EPA 1979) shows that between 10,000 and 90,000 pounds of
*
this chemical were produced/imported in 1977.
p-Chloro-m-cresol is used as an external germicide and as a preserva-
tive for glues, gums, paints, inks, textiles and leather goods (Hawley 1971).
It is also used as a preservative in cosmetics (Wilson 1975, Liem 1977).
EPA (1973) indicates that p-chloro-m-cresol is "cleared for use in adhesives
used in food packaging."
.""This production range information does not include any production/importation
j data claimed as confidential by the person(s) reporting for the TSCA
J Inventory, nor does it include any information which would compromise Con-
fidential Business Information. The data submitted for the TSCA Inventory,
including production range information, are subject to the limitations con-
tained in the Inventory Reporting Regulations (40 CFR 710).
*
-------
II. EXPOSURE
A. Environmental Fate
Voets et al. (1976) reported that p-chloro-m-cresol was quite susceptible
to microbial breakdown under aerobic conditions in an organic medium
(synthetic sewage sludge), while degradation under aerobic conditions in a
mineral solution (simulating oligotrophic aquatic systems) was relatively
difficult. No degradation was observed in either system under anaerobic
conditions.
B. Bioconcentration
No studies on the bioconcentration potential of this compound were
found. Based on its solubility, p-chloro-m-cresol would not be expected
to have a high bioconcentration potential.
C. Exposure
Human exposure to p-chloro-m-cresol occurs through its presence in
certain cosmetics and in a variety of other consumer products in which
it is used as a preservative (Wilson 1975, Liem 1977).
p-Chloro-m-cresol has been found to be formed by the chlorination
of water from a lake and a river receiving cooling waters from electric
power-generating plants, at concentrations of 0.2 ug/1 and 0.7 ug/1, res-
pectively. It has also been found to be formed by the chlorination of the
effluent from a domestic sewage treatment facility at a concentration of
1.5 ug/1 (Jolley et al. 1975).
III. PHARMACOKINETICS
No information was found.
IV. HEALTH EFFECTS
Very little toxicological data for p-chloro-m-cresol was available. The
subcutaneous LD-- for p-chloro-m-cresol in rats is 400 mg/kg (NIOSH 1975).
The oral LD for p-chloro-ra-cresol in rats is 500 mg/kg. In mice the
J_iO
intraperitoneal LDT is 30 mg/kg and the subcutaneous LD is 200 mg/kg
Lo LO
-------
(U.S. DREW 1978). One author has rated p-chloro-m-cresol as very toxic,
with a probable lethal dose to humans of 50-500 mg/kg. (Von Oettingen
as quoted in Gosselin et al. 1976). p-Chloro-m-cresol was also reported
as non-irritating to skin in concentrations of 0.5 to) 1.0% in alcohol.
V. AQUATIC TOXICITY
A. Acute
The only information available is that for Dapjinia. pulex._ The
96-hour t-C5Q for p-chloro-m-cresol exposure is 3.1 rag/L (Jolley et al. 1977),
VI. GUIDELINES
No guidelines for exposure to p-chloro-m-cresol were located.
-------
References
Gosselin RE et al. 1976. Clinical Toxicology of Commercial Products.
Fourth Edition.
Hawley GG (Ed.) 1971. Condensed Chemical Dictionary, 8th Edition. Van
Nostrand Reinhold Co.
Jolley RL., Jones G, Pitt WW, and Thompson JE. 1975. Chlorination of
Organics in Cooling Waters and Process Effluents, In Proceedings of the
Conference on the Environmental Impact of Water Chlorination, Oak Ridge,
Tennessee, Oct. 22-24, 1975, published July 1976.
Jolley RL, Gorchev H, Hamilton DH. 1978. Water Chlorination Environmental
Impact and Health Effects In Proceedings of the Second Conference on the
Environmental Impact of Water Chlorination, Gatlinburg, Tenn. 1977.
Liem DH. 1977. Analysis of antimicrobial compounds in cosmetics, Cosmetics
and Toiletries, 92: 59-72.
National Institute of Occupational Safety and Health. 1975. Registry of
Toxic Effects of Chemcial Substances. 1978 Edition. DHEW (NIOSH) Publication
79-100, Rockville, MD.
U.S. EPA. 1973. EPA Compendium of Registered Pesticides, Vol. II, Part I,
Page P-01-00.01.
U.S. EPA. 1979. Toxic Substances Control Act Chemical Substance Inventory,
Production Statistics for Chemicals on the Non-Confidential TSCA Inventory.
Voets JP, Pipyn P, Van Lancker P, and Verstraete W. 1976. Degradation of
microbicides under different environmental conditions. J. Appl. Bact.
40:67-72.
Wilson, CH. 1975. Identification of preservatives in cosmetic products by
thin layer chromatography. J. Soc. Cosmet. Chem., 26:75-81.
Windholz M. ed. 1976. The Merck Index, Merck & Co., Inc., Rahway, New Jersey.
-------
No. 44
Chloroethane
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, B.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.
-------
CHLOROETHANE
SUMMARY
There is no available evidence which indicates that
monochloroethane produces carcinogenic, mutagenic, or terato-
genic effects. Symptoms produced by human poisoning with
monochloroethane include central nervous system depression,
respiratory failure, and cardiac arrhythmias. The results of
animal studies indicate that liver, kidney, and cardiac tox i-
city may be produced by monochloroethane.
V
Data examining the toxic effects of chloroethane on
aquatic organisms were not available.
-------
CHLOROETHANE
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 have been replaced by chlorine atoms.
Water solubility and vapor pressure decrease with increasing
chlorination, while density and melting point increase.
Monochloroethane (chloroethane, M.W. 64.52) is a gas at room
temperature. The compound has a boiling point of 13.1°C, a
melting point of -138.7°C, a specific gravity of 0.9214, and
a solubility of 5.74 g/1 in water (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 1976 production of monochloroethane was 335 x 10^
tons/year (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 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 volatile compounds widely
used as degreasing agents and in dry cleaning operations
(U.S. EPA, 1979a).
Sources of human exposure to chloroethanes include
water, air, contaminated foods and fish, and dermal absorp-
tion. Fish and shellfish have shown levels of chloroethanes
in the nanogram range (Dickson and Riley, 1976). Data on the
levels of monochloroethanes in foods is not available.
An average bioconcentration factor for monochloroethane
in fish and'shellfish has not been derived by the EPA.
III. PHARMACOKINETICS
Pertinent data could not be located in the available
literature on monochloroethane 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
following oral or inhalation exposure; widespread distribu-
tion of the chloroethanes throughout the body; enzymatic de-
chlorination and oxidation to the alcohol and ester forms;
and excretion of the chloromethanes primarily in the urine
and expired air. Specifically for monochloroethane, absorp-
tion following dernal application is minor; and excretion
appears to be rapid, with the major portion of the injected
compound excreted in the first 24 hours (U.S. EPA, 1979a) .
-------
IV. EFFECTS
Pertinent data could not be located in the available
literature on monochloroethane for carcinogenicity, mutageni-
city, teratogenicity and other reproductive effects.
A. Chronic Toxicity
Human symptons of monochloroethane poisoning indi-
cate central nervous system depression, respiratory failure,
and cardivascular symptoms, including cardiac arrhythmias
(U.S. EPA, 1979a). Animal toxicity has indicated kidney dam-
age and fatty infiltration of the liver, kidney, and heart
(U.S. EPA, 1979a).
V. AQUATIC TOXICITY
Pertinent data could not be located in the available
1iterature.
VI. EXISTING GUIDLINES AND STANDARDS
A. Human
The eight-hour TWA standard prepared by OSHA for
monochloroethane is 1,000 ppm.
Sufficient data are not available to derive a cri-
terion to protect human health from exposure to monochloro-
ethane in ambient water.
B. Aquatic
There are not sufficient toxicological data to cal-
culate exposure criteria.
-------
CHLOROETHANE
REFERENCES
Dickson, 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.
Kirk, R., and Qthmer, D. 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.
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.F. V/ineman. 1971. Enzymatic dechlorination:
Dechlorination of chloroethanes and propanes iin vitro. Biochem.
Pharmacol. 20: 463.
-------
No. 45
Chloroethene
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.
-------
CHLOROETHENE
(VINYL CHLORIDE)
Summary
Vinyl chloride has been used for over 40 years in the produc-
tion of polyvinyl chloride. Animal studies indicate that vinyl
chloride is not teratogenic, but it has been found to be mutagenic
in several biologic test systems. Vinyl chloride has been found to
be carcinogenic in laboratory animals and has been positively asso-
ciated with angiosarcoma of the liver in humans. Recently "vinyl
chloride disease", a multisystem disorder, has been described in
workers exposed to vinyl chloride.
Data are lacking concerning the effects of vinyl chloride
in freshwater and saltwater aquatic life.
-------
CHLOROETHENE
(VINYL CHLORIDE)
I. INTRODUCTION
Vinyl chloride {CH2CHC1; molecular weight 62.5) is a highly
flammable chloro-olefinic hydrocarbon which emits a sweet or
pleasant odor, and has a vapor density slightly more than twice
that of air. Its physical properties include: melting point,
-153.8°C; and solubility in water, O.llg/100 g at 28°C. It is
soluble in alcohol and very soluble in ether and carbon tetra-
chloride (Weast, 1972). Many salts of metals (including silver,
copper, iron, platinum, iridium) have the ability to complex
with vinyl chloride resulting in its increased solubility in
water. Conversely, alkali metal salts, such as sodium or potas-
sium chloride, may decrease the solubility of vinyl chloride
in aqueous solutions (Fox, 1978).
Vinyl chloride has been used for over 40 years in the produc-
tion of polyvinyl chloride (PVC), which in turn is the most widely
used material in the manufacture of plastics. Production of vinyl
chloride in the U.S. reached slightly over 5 billion pounds in 1977
(U.S. Int. Trade Comm, 1978).
Vinyl chloride and polyvinyl chloride are used in the manufac-
ture of numerous products in building and construction, the automo-
tive industry, for electrical wire insulation and cables, piping,
industrial and household equipment, packaging for food products,
medical supplies, and are depended upon heavily by the rubber,
paper and glass industries (Maltoni, 1976a).
In the U.S. about 1500 workers were employed in monomer syn-
thesis and an additional 5000 in polymerization operations {Falk,
-------
et al. 1974). As many as 350,000 workers were estimated to be asso-
ciated with fabricating plants {U.S. EPA, 1974). By 1976, it was
estimated that worldwide nearly one million persons were associated
with manufacturing goods derived from PVC (Maltoni, 1976a).
Potential sources of population exposure to vinyl chloride are
emissions from PVC fabricating plants, release of monomers from
various plastic products, and emissions from the incineration of
PVC products (U.S. EPA, 1975).
II. EXPOSURE
A. Water
Small amounts of vinyl chloride may be present in public
water supplies as a result of industrial waste water discharges.
The levels of vinyl chloride in effluents vary considerably de-
pending on the extent of in-plant treatment of waste water. Vinyl
chloride in samples of waste water from seven areas ranged from
0.05 ppm to 20 ppm, typical levels being 2-3 ppm (U.S. EPA, 1974).
The low solubility and high volatility of vinyl chloride tend to
limit the amounts found in water; however, the presence of certain
salts may increase the solubility and therefore could create situa-
tions of concern (U.S. EPA, 1975).
Polyvinyl chloride pipe.used in water distribution sys-
tems provides another source of low levels of vinyl chloride in
drinking water. In a study by the U.S. EPA of five water distribu-
tion systems which used PVC pipes, water from the newest, longest
pipe system had the highest vinyl chloride concentration (1.4 ug/1)
while the two oldest systems only had traces of vinyl chloride (0.3
*
jig/I and 0.6 pg/1) (Dressman and LMcFarren, 1978). The National
-537-
-------
Science Foundation (NSF) has adopted a voluntary standard of 10 ppra
or less of residual monomer in finished pipe and fittings. Three
times a year NSF samples water supplies in several cities. In
1977, more than 95 percent of the samples conformed to the stan-
dard; however, levels of 5.6 ug/1 and 0.27 ug/1 vinyl chloride have
been detected in at least two cities.
B. Food
Small quantities of vinyl chloride are ingested by humans
when the entrained monomer migrates into foods packaged in PVC
wrappings and containers. The solubility of vinyl chloride in
foods packaged in water is low (0.11 percent); however, the monomer
is soluble in alcohols and mineral oil. In 1973, the U.S. Treasury
Department banned the use of vinyl chloride polymers for packaging
alcoholic beverages (Int. Agency Res. Cancer, 1974). The FDA anal-
yzed a number of PVC packaged products in 1974. The concentrations
ranged from "not detectable" to 9,000 ppb.
The U.S. EPA (1979) has estimated the weighted average
bioconcentration factor of vinyl chloride to be 1.9 for the edible
portions of fresh and shellfish consumed by Americans. This esti-
mate was based on the octanol/water coefficient of vinyl chloride.
C. Inhalation
Inhalation of vinyl chloride is the principal route of
exposure to people working in or living near vinyl chloride indus-
tries. After 1960, Dow Chemical Co. was successful in reducing ex-
posures to workers to about 25 ppm level, though levels up to^500
ppm still occurred. Inhalation exposures drastically dropped after
appropriate controls were instituted following case reports of
vinyl chloride induced angiosarcoma of the liver in workers and ex-
perimental animals (U.S. EPA, 1979).
-------
III. PHARMACOKINETICS
A. Absorption
Vinyl chloride is rapidly absorbed through the lungs and
enters the blood stream (Duprat, et al. 1977).
B. Distribution
The liver of rats accumulates the greatest percentage
of vinyl chloride and/or metabolites of vinyl chloride 72 hours
after a single oral dose (Watanabe, et al. 1976) . Ten minutes
after a 5-minute inhalation exposure to vinyl chloride at 10,000
ppm, the compound was found in the liver, bile duct, stomach,
and kidney of - rats (Duprat, et al. 1977). Immediately after
exposure by inhalation to C-vinyl chloride at 50 ppm for 5
14
hours, the percent incorporated as C/radioactivity per gram
of tissue was highest for kidney (2.13), liver (1.86), and spleen
(0.73). Forty-eight hours after the beginning of exposure, labeled
material could still be detected in these tissues.
C. Metabolism
Detoxification of vinyl chloride takes place primarily in
the liver by oxidation to polar compounds which can be conjugated
to glutathione and/or cysteine (Hefner, et al. 1975). These cova-
lently bond metabolites are then excreted in the urine.
Vinyl chloride is metabolized extensively by rats _in vivo
and the metabolic pathways appear to be saturable. The postulated
primary metabolic pathway involves alcohol dehydrogenase and, for
rats, appears to be saturated by exposures to concentrations ex-
ceeding 220 to 250 ppm. In rats exposed to higher concentratibns,
metabolism of vinyl chloride is postulated to occur via a secondary
X
-------
pathway involving epoxidation and/or peroxidation. Present data
indicates that vinyl chloride is metabolized to- an activated car-
cinogen electrophile and is capable of covalent reaction with
nucleophilic groups or cellular macromolecules (U.S. EPAr 1979).
There is ample evidence that the mixed function oxidase
(MFC) system may be involved in the metabolism of vinyl chloride.
Rat liver microsomes catalyze the covalent binding of vinyl chlor-
ide metabolites to protein and nucleic acids; chloroethylene
oxide is thought to be the primary microsomsl metabolite capable
of alkylating these cellular macromolecules (Kappus, et al. 1975;
1976; Laib and Bolt, 1977). Hathway (1977) reports _in vitro
depurination of calf thymus DNA by chloroacetaldehyde identical
to that observed in hepatocyte DNA following the administration
of vinyl chloride to rats ir\ vitro.
D. Excretion
Watanabe, et al. (1976) monitored the elimination of
vinyl chloride for 72 hours following a single oral dose adminis-
tered to rats. The total 14c-activity recovered at each dose level
ranged from 82-92 percent. At a dose level of 1 mg/kg, 2 percent
was exhaled as vinyl chloride, 13 percent was exhaled as carbon
dioxide, 59 percent was eliminated in the urine and 2 percent in
the feces. Excretion of vinyl chloride at a dose level of 100 mg/kg
was 66 percent exhaled as vinyl chloride, 2.5 percent as carbon
dioxide, 11 percent in the urine and 0.5 percent in the feces. Ad-
ministration by inhalation produced almost the same results.
Green and Hathway (1975) found that more than 96 percent
of 250 jjg C-vinyl chloride administered via intragastric, intra-
-------
venous or intraperitoneal routes was excreted within 24 hours. The
rats given vinyl chloride by the intragastric -route exhaled 3.7
percent as vinyl chloride, 12.6 percent as CO-; 71.5 percent
of the labeled material was in the urine and 2.8 percent in the
feces. Intravenous injections resulted in 9.9 percent exhaled
as vinyl chloride, 10.3 percent as CG>2; 41.5 percent in the urine
and 1.6 percent in the feces.
IV. EFFECTS
A. Carcinogenicity
The carcinogenicity of vinyl chloride has been investi-
gated in several animal studies. Viola, et al. (1971) induced skin
epidermoid carcinomas, lung carcinomas or bone steochrondromas in
24/25 male rats exposed to 30,000 ppm vinyl chloride intermittently
for 12 months. Tumors appeared between 10 and 11 months. Caputo,
et al. (1974) observed carcinomas and sarcomas in all groups of
male and female rats inhaling various concentrations of vinyl
chloride except those exposed to 50 ppm.
Maltoni and Lefemine (1974a,b; 1975} reported on a
series of experiments concerning the effects on rats, mice, and
hamsters of inhalation exposure to vinyl chloride at concentra-
tions ranging from 50 to 10,000 ppm for varying periods of time.
The animals were observed for their entire lifetime. Angiosar-
comas of the liver occurred in all three species, as well as
tumors at several other sites. A differential response between
the sexes was not reported,
Maltoni (1976b) observed four subcutaneous angiosar-
comas, four zymbal gland carcinomas, and one nephroblastoma in
-------
66 offspring of rats exposed by inhalation 4 hours/day to 10,000
or 6,000 ppra vinyl chloride from the 12th to 18th day of gesta-
tion. Liver angiosarcomas were also observed in rats administered
vinyl chloride via stomach tube for 52 weeks.
Recent experiments by Lee, et al. (1977) with rats
and mice confirm the carcinogenicity of vinyl chloride. Each
species was exposed to 50,250 or 1000 ppm vinyl chloride or 55
ppm vinylene chloride 6 hr/day, 5 days/week for 1-12 months.
After 12 months, bronchioalveolar adenomas, jnammary gland tumors,
and angiosarcomas in the liver and other sites developed in mice
exposed to all three dose levels of vinyl chloride. Rats exposed
to 250 ppm or 100 ppm vinyl chloride developed angiosarcoma in
the liver, lung and other sites (Lee, et al. 1978).
The primary effect associated with vinyl chloride expo-
sure in man is an increased risk of cancer in several organs in-
cluding angiosarcoma of the liver. Liver angiosarcoma is an ex-
tremely rare liver cancer in humans, with 26 cases reported annual-
ly in the U.S. (Natl. Cancer Inst., 1975). Human data on the car-
cinogenic effects of vinyl chloride have been obtained primarily
from cases of occupational exposures of workers. The latent period
has been estimated to be 15-20 years; however, recent case reports
indicate a longer average latent period (Spirtas and Kaminski,
1978).
A number of epidemiological studies of vinyl chloride
have been reported (U.S. EPA, 1979). Tabershaw/Cooper Associates
*
(1974) found no increase in the overall mortality rate for vinyl
chloride workers nor significant increases in standard mortality
-------
rates (SMR's) for malignant neoplasms. Reexamination of this
data by Ott, et al. (1975) including more clearly defined expo-
sure levels confirmed the previous findings: no increase over
that expected for malignant neoplasms in the low exposure group
(TWA 10-100 ppm vinyl chloride) and a non-significant increase
in deaths due to malignant neoplasms in the high exposure group
(TWA, greater than 200 ppm).
However, liver cancer death were twelve-fold, and brain
cancer deaths were five-fold greater than, that expected in a
study by Wagoner (1974). Likewise, Monson, et al. (1974) found
death due to cancer to be 50 percent higher than expected in
vinyl chloride workers who died from 1947-1973, including a 900
percent increase in cancers of the liver and biliary tract.
In the most recent update of the NIOSH register, a total
of 64 cases of hepatic angiosarcoma have been identified worldwide
among vinyl chloride exposed industrial workers (Spirtas and Kamin-
ski, 1978). Twenty-three of these cases were reported in the
United States. Six cases were documented since 1975.
B. Mutagenicity
Vinyl chloride has been found to be mutagenic in a number
of biological systems including: metabolically activated systems
using Salmonella typhimurium; back mutation systems using
Escherichia coli; forward mutation and gene coversion in yeast; and
germ cells of Drosophila and Chinese hamster V79 cells (U.S. EPA,
1979).
*
The dominant lethal assay was used to -test the mutageni-
city of inhaled vinyl chloride in mice. Levels as high as 30,000
-------
ppm (6 hours/day for 5 days) yielded negative results (Anderson, et
al. 1976).
Several investigators have observed a significantly
higher incidence of chromosomal aberrations in the lymphocytes of
workers chronically exposed to high levels of vinyl chloride
(Ducatman, et al. 1975; Purchase, et al. 1975; Funes-Cravioto, et
al. 1975).
C. Teratogenicity
Animal studies using mice, rats and rabbits, indicate
that inhalation of vinyl chloride does not induce gross teratogenic
abnormalities in offspring of mothers exposed 7 hours daily to con-
centrations ranging from 50 to 2,500 ppm (John, et al. 1977); how-
ever, excess occurrences of minor skeletal abnormalities were
noted. Increased fetal death was noted at the higher exposure
levels. These findings were confirmed by Radike, et al. (1977a)
who exposed rats to 600-6,000 ppm vinyl chloride, 4 hours daily on
the 9th to the 21st day of gestation.
Further examination is needed of reported high rates of
congenital defects in three small communities in which vinyl chlor-
ide polymerization plants are located (U.S. EPA, 1979).
D. Other Reproductive Effects
No effect on fertility in mice was noted in a dominant
lethal assay conducted by Anderson, et al. (1976).
E. Chronic Toxicity
There are numerous clinical indications that chronic
exposure to vinyl chloride is toxic to humans (U.S. EPA, 1979).
Hepatitis-like changes, angioneurosis, Raynaud's syndrome, derma-
-------
titis, acroosteolysis, thyroid insufficiency, and hepatomegaly
have been reported around the world. Other long term effects
include functional disturbances of the central nervous system
with adrenergic sensory polyneuritis (Smirnova and Granik, 1970);
thrombocytopenia, splenomegaly, liver malfunction with fibrosis,
pulmonary changes (Lange, et al. 1974) ; and alterations in serum
enzyme levels (Makk, et al. 1976) .
F. Other Relevant Information
Pretreatment of rats with pyrazole'- (an alcohol dehydro-
genose inhibitor) and ethanol inhibits the metabolism of vinyl
chloride (Hefner, et al. 1975). This indicates the involvement
of alcohol dehydrogenose in the metabolism of vinyl chloride.
The chronic ingestion of alcohol was found to increase
the incidence of liver tumors and tumors in other sites in in-
dividuals exposed to vinyl chloride (Radike, 1977b).
Jaeger (1975) conducted experiments to determine the
interaction between vinylidene chloride (1,1-DCE) and vinyl chloride.
In this experiment, the effects of 4-hour exposures to 200 ppm
of 1,1-DCE and 1,000 ppm vinyl chloride were less than if 1,1-
DCE was given alone.
V. AQUATIC TOXICITY
A. Pertinent information relevant to acute and chronic toxi-
city, plant effects and residues for vinyl chloride were not found
in the available literature.
-------
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
The current federal OSHA standard for vinyl chloride is 1
ppm (TWA) with a maximum of 5 ppra for a period of no longer than 15
minutes in 1 day. (39 PR 35890 (Oct. 4, 1979)).
In 1974, a notice to cancel registrations of pesticide
spray products containing vinyl chloride as a propellant was issued
(39 FR 14753 (April 26, 1974)). Other aerosol products, such as
hair spray, utilizing vinyl chloride as a propellant were banned
from the market in the U.S. and other countries (Int. Agency Res.
Cancer, 1974). The U.S. EPA proposed in 1975 and 1976 an emission
standard of 10 ppm vinyl chloride at the stack for industry.
The draft ambient water quality criterion for vinyl
chloride has been set to reduce the human lifetime cancer risk
level to 10~5, 10~6 and 10~7 (U.S. EPA, 1979). The corresponding
criteria are 517 jig/1, 51.7 jug/1 and 5.17 pg/1/ respectively. The
data base from which this criterion has been derived is currently
being reviewed, therefore, this criteria to protect human health
may change.
B. Aquatic
Fresh or salt water criteria could not be derived because
of insufficient data (U.S. EPA, 1979).
-------
CHLOROETHENE
(VINYL CHLORIDE)
REFERENCES
Anderson, 0., et al. 1976. Vinyl chloride: dominant lethal studies in male
CD-I mice. Mutat. Red. 40: 359.
Caputo, A., et al. 1974. Oncogenicity of vinyl chloride at low concentra-
tions in rats and rabbits. IRCS 2: 1582.
Dressman, R.C. and E.F. McFarren. 1978. Determination of vinyl chloride
migration from polyvinyl chloride pipe into water. Am. Water Works Assoc.
Jour. 70: 29.
Ducatman, A., et al. 1975. Vinyl chloride exposure and human chromosome
aberrations. Mutat. Rec. 31: 163.
Duprat, P., et al. 1977. Metabolic approach to industrial poisoning:
blood kinetics and distribution of ^^C-vinyl chloride monomer (VCM).
Toxicol Pharmacol. Suppl. 142.
Falk, H., et al. 1974. Hepatic disease among workers at a vinyl chloride
polymerication plant. Jour. Am. Med. Assoc. 230: 59.
Fox, C.R. 1978. Plant uses prove phenol recovery with resins. Hydrocarbon
processing. November, 269.
Funes-Cravioto, F., et al. 1975. Chromosome aberrations in workers exposed
to vinyl chloride. Lancet 1: 459.
Green, T. and D.E. Hathway. 1975. The biological fate in rats of vinyl
chloride in relation to its oncogenicity. Chem. Biol. Interactions.
11: 545.
Hathway, O.E. 1977. Comparative mammalian metabolism of vinyl chloride and
vinylidene chloride in relation to oncogenic potential. Environ. Health
Perspect. 21: 55.
Hefner, R.E., Jr., et al. 1975. Preliminary studies of the fate of inhaled
vinyl chloride monomer in rats. Ann. N.Y. Acad. Sci. 246: 135.
International Agency for Research on Cancer. 1974. Monograph on the evalu-
ation of carcinogenic risk of chemicals to man. Vol. 7. Lyon, France.
Jaeger, R.J. 1975. Vinyl chloride monomer: comments on its hepatotoxicity
and interaction with 1,1-dichloroethylene. Ann. N.Y. Acad. Sci. 246: 150.
John, J.A., et al. 1977. The effects of maternally inhaled vinyl chloride
on embryonal and fetal development in mice, rats and rabbits. Toxicol.
Appl. Pharmacol. 39: 497.
-------
Kappus, H, , et al. 1975. Rat liver microsomes catalyse covalent binding of
chloride to macromolecules. Nature 257: 134.
Kappus, H., et al. 1976. Liver microsomal uptake of (^C) vinyl chloride
and transformation to protein alkylating metabolites in vitro. Toxicol.
Appl. Pharmacol. 37: 461.
Laib, R.J. and H.M. Bolt. 1977. Alkylation of RNA by vinyl chloride meta-
bolites in vitro and in vivo: formation of l-N6-ethenoadenosine. Toxico-
logy 8: 185.
Lange, C.E., et al. 1974. The so-called vinyl chloride sickness-and-occu-
pationally-related systemic sclerosis? Int. Arch. Arbeitsmed. 32: 1.
Lee, C.C., et al. 1977. Inhalation toxicity of vinyl chloride and vinyli-
dene chloride. Environ. Health Perspect. 21: 25.
Lee, C.C., et al. 1978. Carcinogenicity of vinyl chloride and vinylidene
chloride. Jour. Toxicol. Environ. Health 4: 15.
Makk, L., et al. 1976. Clinical and morphologic features of hepatic angio-
sarcoma in vinyl chloride workers. Cancer 37:149.
Maltoni, C. 1976a. Carcinogenicity of vinyl chloride: Current results.
Experimental evidence. Proc. 6th Int. Symp. Biological Characterization of
Human Tomours, Copenhagen May 13-15, 1975. Vol. 3 Biological characteriza-
tion of human tumours, 1976. American Elsevier Publishing Co., Inc., New
York.
Maltoni, C. 1976b. Predictive value of carcinogenesis bioassays. Ann.
N.Y. Acad. Sci. 271: 431.
Maltoni, C. and G. Lefemine. 1974a. Carcinogenicity bioassays of vinyl
chloride. I. Research plan and early results. Environ. Res. 7:387.
Maltoni, C. and G. Lefemine. 1974b. La potentiality dei saggi sperimentali
mella predizion; dei rischi oncogeni ambiental: Un esemplo: 11 chlorure di
vinile. Acad. Natl. Lincei. 56: 1.
Maltoni, C. and G. Lefemine. 1975. Carcinogenicity assays of vinyl chlor-
ide: Current results. Ann. N.Y. Acad. Sci. 246: 195.
Monson, R.R., et al. 1974. Mortality among vinyl chloride workers. Pre-
sented at Natl. Inst. Environ. Health Sci. Conf., Pinehurst, N.C., July
29-31.
National Cancer Institute Monograph 41. 1975. Third national cancer sur-
vey: incidence data.
Ott, M.G., et al. 1975. Vinyl chloride exposure in a controlled industrial
environment: a long-term mortality experience in 595 employees. Arch. Envi-
ron. Health 30: 333.
-------
Purchase, I.F.H., et al. 1975. Chromosomal and dominant lethal effects of
vinyl chloride. Lancet 28: 410.
Radike, M., et al. 1977a. Transplacental effects of vinyl chloride in
rats. Annual Report. Center for the Study of the Human Environment. USPHS-
ES-00159. Dept. Environ. Health, Med. College, University of Cincinnati.
Radike, M.J., et al. 1977b. Effect of ethanol and vinyl chloride on the
induction of liver tumors: preliminary report. Environ. Health Perspect.
21: 153.
Smirnova, N.A. and N.P. Granik. 1970. Long-term side effects of acute
occupational poisoning by certain hydrocarbons and their derivatives. Gig.
Tr. Prof. Zabol. 14: 50.
Spirtas, R. and R. Kaminski. 1978. Angiosarcoma of the liver in vinyl
chloride/polyvinyl chloride workers. Update of the NIOSH Register. Jour.
Occup. Med. 20: 427.
Tabershaw/Cooper Assoc., Inc. 1974. Epidemiologic study of vinyl chloride
workers. Final report submitted to Manufacturing Chemists Assoc., Washing-
ton, O.C. Berkeley, Calif.
U.S. EPA. 1974. Preliminary assessment of the environmental problems asso-
ciated with vinyl chloride and poly vinyl chloride. EPA 560/4-74-001. Natl.
Tech. Inf. Serv., Springfield, Va.
U.S. EPA. 1975. A scientific and technical assessment report on vinyl
chloride and polyvinyl chloride. EPA-600/6-75-004. Off. Res. Dev., U.S.
Environ. Prot. Agency, Washington, D.C.
U.S. EPA. 1979. Vinyl Chloride: Ambient Water Quality Criteria. (Draft).
U.S. International Trade Commission. 1978. Synthetic organic chemicals.
U.S. Production and Sales, 1977. Publ. 920. U.S. Government Printing Of-
fice, Washington, D.C.
Viola, P.L., et al. 1971. Oncogenic response of rat skin, lungs, bones to
vinyl chloride. Cancer Res. 31: 516.
Wagoner, J.E. 1974. NIOSH presented before the environment. Commerce Comm.
U.S. Senate, Washington, D.C.
Watanabe, P.G., et al. 1976. Fate of (l^C) vinyl chloride after single
oral administration in rats. Toxicol. Appl. Pharmacol. 36: 339.
Weast, R.C. (ed.) 1972. Handbook of chemistry and physics. CRC Press,
Cleveland, Ohio.
-------
No. 46
2-Chloroethyl Vinyl Ether
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
-550-
-------
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 ac-c-uracy.
-SSI-
-------
2-CHLOROETHYL VINYL ETHER
SUMMARY
Very little Information is available for 2-chloroethyl vinyl ether. It
appears to be relatively stable except, under acidic conditions. There is some
potential for bioconcentration of the compound in exposed organisms. No expo-
sure data were available, although 2-chloroethyl vinyl ether has been identified
in industrial effluent discharges.
The acute toxicity of 2-chloroethyl vinyl ether is relatively low: oral
LD50: 25° raS/kS; dermal LT>5Q 3.2 mL/kg; LC^: 250 ppm (A hrs). Eye irrita-
tion has been reported following exposure to 2-chloroethyl vinyl ether. No
other data on health effects were available.
I. INTRODUCTION
2-Chloroethyl vinyl ether (C1CH2CH2OCH=CH2; molecular weight 106.55) is a
liquid having the following physical/chemical properties (Windholz, 1976; Weast,
1972; U.S. EPA, 1979c):
Boiling point (760 mm Hg): 109°C
Melting point: -70°C
20
Density: 1.0475
Solubility: Soluble in water to the extent
of 6g/L; very soluble in
alcohol and ether
The compound finds use in the manufacture of anesthetics, sedatives, and
cellulose ethers (Windholz, 1976).
A review of the production range (includes importation) statistics for 2-
chloroethyl vinyl ether (CAS No. 110-75-8) which is listed in the initial XSCA
inventory (1979a) has shown that none of this chemical was produced or imported
in 1977*.
*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 (40CFR710).
-------
II. EXPOSURE
A. Environmental Rate
The $-chloroalkyl ethers have been shown to be quite stable to hydrolysis
and to persist for extended periods without biodegradation (U.S. EPA, 1979b).
2-Chloroethyl ethyl ether (a fj-chloroalkyl other) is stable to sodium hydroxide
solutions but will undergo hydrolysis in the presence of dilute acids to acet-
aldehyde and 2-chloroethanol (Windholz 1976). Conventional treatment systems
may be inadequate to sufficiently remove the g-chloroalkyl ethers once present
in water supplies (U.S. EPA 1979b; U.S. EPA 1975).
B. Bioconcentration
A calculated bioconcentration factor of 34.2 (U.S. EPA, 1979b) points to
some potential for 2-chloroethyl vinyl ether accumulation in exposed organisms.
C. Environmental Occurrence
There is no specific information available on general population exposure
to 2-chloroethyl vinyl ether. The compound has been identified three times in
the water of Louisville, Kentucky (3/74): twice in effluent
facturing plants and once in the effluent from a latex plant (U.S. EPA 1976). No
concentration levels were given.
NIOSH, utilizing data from the National Occupational Hazards Survey
(NOHS 1977) has compiled a listing summarizing occupational exposure to 2-
chloroethyl vinyl ether (Table 1). As shown, NIOSH estimates 23,473 people
are exposed annually to the compound. The number of potentially exposed indi-1
viduals is greatest for the following areas: fabricated metal products; whole-
sale trade; leather, rubber and plastic, and chemical products.
III. PHARMACOK1NETICS
»
Vinyl ethers readily undergo acid catalysed hydrolysis to give alcohols and
aldehydes, e.g., 2-chloroethyl vinyl ether is hydrolyzed to 2-chloroethanol and
acetaldehyde (Salomaa et al. 1966).
-------
TABLE 1
PROJECTED NUMBERS BY INDUSTRY
HAZARD DESCRIPTION
84673 Chloroethyl Vinyl Ether, 2-
SIC,
CODE DESCRIPTION
25 Furniture and fixtures
28 Chemicals and allied products
30 Rubber and plastic products-
31 Leather and leather products
34 Fabricated metal products
35 Machinery, except electrical
36 Electrical equipment and supplies
37 Transportation equipment
38 Instruments and related products
39 Miscellaneous manufacturing industries
50 Wholesale trade
73 Miscellaneous business services
ESTIMATED
PLANTS
ESTIMATED
PEOPLE
ESTIMATED
EXPOSURES
I
T
V1
In
I
TOTAL
2,059
23,473
23,473
-------
IV. HEALTH EFFECTS
A. Mutag enicity^
Although no information on the mutagenicity of 2-chloroethyl vinyl ether was
available, its hydrolysis product, 2-chloroethanol, has been shown to be muta-
genie in Salmonella typhimurium TA 1535 (Rannug et al. 1976), TA100 and TA98
(McCann et al. 1976), as well as Klebsiella pneumonia (Voogd et al. 1972).
B. Cither Toxicity
Very little toxicological data for 2-chloroethyl vinyl ether is available.
The oral LD5Q for 2-chloroethyl vinyl ether in rats is 250 mg/kg (U.S. EPA, 1975,
Patty 1963). Dermal exposure to the shaven skin of rabbits for 24 hours resulted
in an LD,-0 of 3.2 mL/kg (U.S. EPA, 1976). The acute inhalation toxicity of
2-chloroethyl vinyl ether in rats was determined following single four-hour
exposures. The lowest lethal concentration was 250 ppm (U.S. EPA, 1975). In a
similar inhalation study, 1/6 rats exposed by inhalation to 500 ppm died during
the 14-day observation period (U.S. EPA, 1975).
Primary skin irritation and eye irritation studies have also been conducted
for 2-chloroethyl vinyl ether. Dermal exposure to undiluted 2-chloroethyl vinyl
ether did not cause even slight erythema. Application of 0.5 mL undiluted 2-
chloroethyl vinyl ether to the eyes of rabbits resulted in severe eye injury
(U.S. EPA, 1975).
V. AQUATIC TOXICITY
A. Acute
The adjusted 96-hour LC „ for blue gill exposure to 2-chloroethyl vinyl
ether is 194,000 ug/L (U.S. EPA, 1979b). Dividing by the species sensitivity
factor (3.9), a Final Fish Acute Value of 50,000 ug/L is obtained (Table 1).
There is no data on invertebrate or plant exposure.
VI. EXISTING GUIDELINES
No guidelines were located.
-------
Table 2. Freshwater fish acute values (U.S. EPA, 1979b)
Adjusted
Bioassay Test Chemical Time LC5Q LC5Q
Organism Method Cone.** Description. (hrs) (ug/L) (ug/L)
Bluegill, S U 2-chloroethyl 96 354,000 194,000
Lepomis macrochirus vinyl ether
* S = static
** U = unmeasured
Geometric mean of adjusted values: 2-chloroethyl vinyl ether = 194,000 ug/L
. 50j000 ug/L
-------
References
Lange NA (ed.). 1967. Lange's Handbook of Chemistry, rev. 10th ed. , New York:
McGraw-Hill Book Co.
McCann J, Simmon V., Streitwieser D, Ames EN. 1975. Mutagenicity of chloro-
acetaldehyde, a possible metabolic product of 1,2-dichloroethane (ethylene
dichloride), chloroethanol (ethylene chlorohydrin), vinyl chloride and cyclo-
phosphamide. Proc. Nat. Acad. Sci. 72:3190-3193.
National Occupational Hazard Survey (NOHS) 1977 Vol. Ill, U.S. DREW, NIOSH,
Cincinnati, Ohio (Special request for computer printout: 2-chloroethyl vinyl
ether Dec. 1979)
Rannug U., Gothe R. Wachtmeister CA. 1976. The mutagenicity of chloroethylene
oxide, chloroacetaldehyde, 2-chloroethanol and chloroacetic acid, conceivable
metabolites of vinyl chloride. Chem-Biol. Interactions 12:251-263.
Saloraaa P, Kankaanpera A. Lajunen M. 1966. Protolytic cleavage of vinyl
ethers, general acid catalysis, structural effects and deuterium solvent isotope
effects. Acta Chemica Scand. 20:1790-1801.
U.S. EPA, 1975. Investigation of selected potential environmental
contaminants: Haloethers. EPA 560/2-75-006.
U.S. EPA, 1976. Frequency of organic compounds identified in water. EPA 600/4-
76-062.
U.S. EPA, 1979a. Toxic Substance Control Act, Chemical Substance Inventory,.
Production Statistics for Chemicals on the Non-Confidential Initial TSCA Inventory.
U.S. EPA, 1979b. Ambient Water Quality Criteria Document on Chloroalkyl Ethers.
PB 297-921.
U.S. EPA, I979c. Ambient Water Quality Criteria Document on Haloethers. PB 296-796.
Voogd CE, Jacobs JJJAA, van der Stel JJ. 1972. On the mutagenic action of
dichlorvos. Mutat. Res. 16:413-416.
Weast RC (ed.). 1972. Handbook of Chemistry and Physics, 53rd ed. The Chemical
Rubber Co., Cleveland, OH.
Windholz M. (ed.). 1976. The Merck Index, 9th ed. Merck & Co. Inc., Rahway, NJ.
-------
No. 47
Chloroform (Carbon Trichlororaethane)
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.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated
chloroform and has found sufficient evidence to indicate
that this compound is carcinogenic.
-560-
-------
CHLOROFORM
SUMMARY
Chloroform has been found to induce hepatocellular
carcinomas in mice and kidney epithelial tumors in rats.
Hepatomas have also been induced in mice, but necrosis may
be a prerequisite to tumor formation. Bacterial assays
involving chloroform have yielded no mutagenic effects.
Chloroform has produced teratogenic effects when administered
to pregnant "rats.
Reported 96-hour LCcn values for two common freshwater
fish range from 43,800 to 115,000 ug/1 in static tests.
A 48-hour static test with Daphnia magna yielded an LC^g
of 28,900 pg/1. The observed 96-hour LC5Q for the saltwater
pink shrimp is 81,500 |jg/l. In a life cycle chronic test,
the chronic value was 2,546 }ag/l for Uaphn_ia_ rpagnja. Per-
tinent information on chloroform toxicity to plants could
not be located in the available literature. In the only
residue study reported, the bluegill concentrated chloroform
six times after a 14-day exposure. The tissue half-life
was less than one day suggesting that residues of chloroform
would not be an environmental hazard to aquatic life.
-5C.J-
-------
CHLOROFORM
I. INTRODUCTION
This profile is based on the Ambient Water Quality
Criteria Document for Chloroform (U.S. EPA, 1979a).
Chloroform (CHC13; molecular weight 119.39) is a clear,
colorless liquid with a pleasant, etheric, non-irritating
odor and taste (Hardie, 1964; windholz, 1976). It has the
following physical/chemical properties (Hardie, 1964; Irish,
1972; Windholz, 1976):
Boiling Point: 61-62°C
Melting Point: -63.5°C
Flash Point: none (none-flammable)
Solubility: Water - 7.42 x 10 pq/1 at 25°C
Miscible with alcohol, benzene,
ether, petroleum ether, carbon
tetrachloride, carbon disulfide,
and oils.
Vapor Pressure: 200 mm Hg at 25 C
Current Production: 1.2 x 105 metric tons/year (U.S.
EPA, 1978a).
Chloroform is currently used either as a solvent or
as an intermediate in the production of refrigerants (prin-
cipleus), plastics, and Pharmaceuticals (U.S. EPA, 1975).
Chloroform is relatively stable under normal environ-
mental conditions. When exposed to sunlight, it decomposes
slowly in air but is relatively stable in water. The mea-
sured half-life for hydrolyis was found to be 15 months
(Natl. Acad. Sci., 1978a). Degradation in water can occur
in the presence of metals and is accelerated by aeration •
(Hardie, 1964).
-------
For additional information regarding halomethanes as
a class the reader is referred to the Hazard Profile on
halomethanes (U.S. EPA, 1979b).
II. EXPOSURE
Chloroform appears to be ubiquitous in the environment.
A major source of chloroform contamination is from the chlor-
ination of water and wastewater (U.S. EPA, 1975; Bellar,
et al., 1974). Industrial spills may occasionally be a
pulse source of transient high level contamination (Nat.
Acad. Sci., 1978a; Neely, et al., 1976; Brass and Thomas,
1978).
Based on available monitoring data including informa-
tion from the National Organics Monitoring Survey (MOMS) ,
the U.S. EPA (1978b) has estimated the uptake of chloroform
by adult humans from air, water, and food:
Source
Atmosphere
Water
Food Supply
Total
Atmosphere
Water
Food Supply
Total
Atmosphere
Water
Food Supply
Total
AduTt
mg/yr
Maximum Conditions
204
343
16
563
Minimum Conditions
0.41
0.73
2.00
3.14
Mean Conditions
20.0
64.0
9.00
93
Percent
uptake
36
61
3
100.00
13
23
64
100.00
22
69
10
iob.oo
-------
A similar estimate, not using NOMS data, has been made by
the National Academy of Sciences (Nat. Acad. Sci., 1978a).
The U.S. EPA (1979a) has estimated the bioconcentration
factor for chloroform to be 14 for the edible portions of
fish and shellfish consumed by Americans. This estimate'
is based on measured steady-state bioconcentration studies
in bluegills.
III. PHARMACOKINETICS
A. Absorption
The efficiency of chloroform absorption by the
gastrointestinal tract is virtually 100 percent in humans
(Fry, et al., 1972). The corresponding value by inhalation
is 49 to 77 percent (Lehmann and Hassegawa, 1910). Quantita-
tive estimates of dermal absorption efficiency were not
encountered. Since chloroform was used as an anesthetic
via dermal administration, some dermal absorption by humans
can be assumed (U.S. EPA, 1979a).
B. Distribution
Chloroform is transported to all mammalian body
organs and is also transported across the placenta. Strain
differences for chloroform distribution in mice have been
documented by Vessell, et al., (1976).
C. Metabolism
Most absorbed chloroform is not metabolized by
mammals. Toxication, rather than detoxication, appears
f
to be the major consequence of metabolism and probably involves
mixed-function oxidase (MFO) enzyme systems. This observa-
-------
tion is based on enhancement of chloroform toxicity by MFO
inducers and the diminution of toxicity by MFO inhibitors
(Ilett, et al., 1973, McLean, 1970). At least in the liver,
covalent binding of a metabolite to tissue is associated
with tissue damage (Lavigne and Marchand, 1974). Limited
human data (two people) suggest that about 50 percent of
absorbed chloroform is metabolized to CO2 (Fry, et al./
1972; Chiou, 1975).
D. Excretion
In humans, the half-life of chloroform in the
blood and expired air is 1.5 hours (Chiou, 1975). Most
unchanged chloroform and C02 generated from chloroform are
eliminated via the lungs. Chlorine generated from chloroform
metabolism is eliminated via the urine (Taylor, et al.,
1974; Fry, et al., 1972).
IV. EFFECTS
A. Carcinogenicity
Eschenbrenner and Miller (1945) demonstrated that
oral doses of chloroform administered over a 16-month period
induced hepatomas in strain A mice. Based on variations
in dosing schedules, these researchers concluded that necro-
sis was prerequisite to tumor induction.
In the National Cancer Institute bioassay of chloro-
form (NCI, 1976), hepatocellular carcinomas were induced
in mice (Table 1) and kidney epithelial tumors were induced
in male rats (Table 2), following oral doses over extended ,
periods of time.
-------
Ten epidemiologic studies have been conducted
on the association of human exposure to chloroform and/or
other trihalomethanes with cancer. A review of these studies
by the National Academy of Sciences (NAS, 1978b) indicated
that these studies suggest that higher concentrations of
trihalomethanes in drinking water may be associated with
an increased frequency of cancer of the bladder. One of
these studies {McCabe, 1975) claimed to demonstrate a statis-
tically significant correlation between age, sex, race,
adjusted death rate for total cancer, and chloroform levels.
B. Mutagenicity
Chloroform yielded negative results in the Ames
assay (Simmon, et al. 1977).
C. Teratogenicity
At oral dose levels causing signs of maternal
toxicity, chloroform had fetotoxic effects on rabbits (100
mg/kg/day) and rats (316 mg/kg/day) (Thompson, et al., 1974).
Fetal abnormalities (acaudia, imperforate anus, subcutaneous
edema, missing ribs, and delayed ossification) were induced
when pregnant rats were exposed to airborne chloroform at
489 and 1,466 mg/m , 7 hrs/day, on days 6 to 15 of gestation.
At 147 mg/m , the only effects were significant increases
in delayed skull ossification and wavy ribs (Schwetz, et
al., 1974).
-------
Table 1. Hepatocellular Carcinoma Incidence in Mice*
Male
Female
Controls Low
Colony Matched 22,§£
1777""^ 1/15 138 mg/kg 187"50
(6%)
1/80
(1%)
(6%) (30%)
0/20 238 mg/kg 36/45
(0%) (80%)
High
Dose _
277 mg/kg34745
(98%)
477 mg/kg 39/41
(95%)
Table 2. Statistically Significant Tumor Incidence in Rats'
Controls
Colony Matched
Males
Low
Dose
High
Dose
Kidney 0/99
epithelial
tumors/animals
P value 0.0000
_-
0/19
90 mg/kg
(8%)
4/50 180/mg/kg 12/50
(24%)
0.0016
Source: National Cancer Institute, 1976.
D. Other Reproductive Effects
Pertinent data could not be located in the avail-
able literature.
E. Chronic Toxicity
The NIOSH Criteria Document (1974) tabulates data
on the effect of chronic chloroform exposure in humans.
The primary target organs appear to be the liver and kidneys,
with some signs of neurological disorders. These effects
have been documented only with occupational exposures.
-5-47-
-------
With the exception of the possible relationship to cancer
(Section IV.A), chronic toxic effects in humans, attribut-
able to ambient levels of chloroform, have not been documented.
The chronic effects of chloroform in experimental
mammals is similar to the effects seen in humans: liver
necrosis and kidney degeneration (Torkelson, et al., 1976;
U.S. EPA, 1979a).
F. Other Relevant Information
Ethanol pretreatment of mice reportedly enhances
the toxic effects of chloroform on the liver (Kutob and
Plaa, 1961), as does high fat and low protein diets (Van
Oettingen, 1964; McLean, 1970). These data were generated
using experimental mammals.
V. AQUATIC TOXICITY
A. Acute Toxicity
Bentley, et al. (1975) observed the 96-hour LC5Q
values for rainbow trout, (Salmo gairdneri), of 43,800 and
66,800 |ig/l and for bluegills (Lepomis macrochirus) , 100,000
to 115,000 pg/1, all in static tests. A 48-hour static
test with Djtphnia magna resulted in an LCen of 28,900 pg/1
(U.S. EPA 1979a). The observed 96-hour LC5Q for the pink
shrimp (Panaeus duorarum) is 81,500 pg/1. (Bentley, et
al., 1975).
B. Chronic Toxicity
The chronic effects of chloroform on Daphruja magna
were determined using flow-through methods with measured *
concentrations. The chronic effect level was 2,546 ^g/1
{U.S. EPA, 1979a). No other chronic data were available.
-------
C. Plant Effects
Pertinent information could not be located in
the available literature concerning acute chronic toxicity
of chloroform to plants.
D. Residues
In the only residue study reported, the bluegill
(Lepomis macrochirus] bioconcentrated chloroform six times
after a 14-day exposure (U.S. EPA, 1979a). The tissue half-
life was less than one day.
VI. EXISTING GUIDELINES AND STANDARDS
Both the human health and aquatic criteria derived
by U.S. EPA (1979a), which are summarized below, are being
reviewed; therefore, there is a possibility that these crite-
ria may be changed.
A. Human
Based on the NCI mice data, and using the "one-
hit" model, the EPA (1979a) has estimated levels of chloro-
form in ambient water which will result in specified risk
levels of human cancer:
Exposure Assumption Risk Levels and Corresponding Criteria
(per day") ™ "' ~~ _5
0 10 ' 10 ° 10
2 liters of drinking 0 0.021 pg/1 0.21 pg/1 2.1 pg/1
water and consumption
of 18.7 grams fish and
shellfish.
*
Consumption of fish 0 0.175 pg/1 1.75 jjg/1 17.5
shellfish only.
-------
The above risks assume that drinking water treatment
and distribution will have no impact on the chloroform con-
centration .
The NIOSH time-weighted average exposure criterion
for chloroform is 2 ppm or 9,8 mg/m .
The FDA prohibits the use of chloroform in drugs, cos-
metics, or food contact material (14 PR 15026, 15029 April
9, 1976).
Refer to the Halomethane Hazard Profile for discussion
of criterion derivation (U.S. EPA, 1979b).
B. Aquatic
For chloroform, the draft criterion to protect
freshwater aquatic life, based on chronic invertebrate toxi-
city, is 500 ^jg/1 as a .24-hour average and the concentration
should not {based on acute effects) exceed 1,200 ^ig/1 at
any time (U.S. EPA, 1979a). To protect saltwater aquatic
life, the concentration of chloroform should not exceed
620 |ig/l as a 24-hour average and the concentration should
not exceed 1,400 ^ig/l'at anytime (U.S. EPA, 1979a) . These
were calculated from an experiment on a marine invertebrate.
Sf
-S7Q-
-------
CHLOROFORM
REFERENCES
Bellar, T.A., et al. 1974. The occurrence of organohalides in chlorinated
drinking water. Jour. Am. Water Works Assoc. 66: 703.
Bentley, R.E., et al. 1975. Acute toxicity of chloroform to bluegill
(Lepomis macrochirus), rainbow trout, (Salmo gairdneri), and pink shrimp
(Penaeus duorarum)7' Contract No. WA-6-99-1414-B. U.S. Environ. Prot.
Agency.
Brass, H.J. and R.F. Thomas. 1978. Correspondence with Region III. Tech.
Support Div., U.S. Environ. Prot. Agency, Washington, O.C.
Chiou, W.L. 1975. Quantitation of hepatic and pulmonary first-pass, effect
and its implications in pharmacokinetic study. I. Pharmacokinetics of
chloroform in man. Jour. Pharmacokin. Siopharmaceu. 3: 193.
Eschenbrenner,' A.B. and E. Miller. 1945. Induction of hepatomas in mice by
repeated oral administration of chloroform, with observations on sex dif-
ferences. Jour. Natl. Cancer Inst. 5: 251.
Fry, B.J., et al. 1972. Pulmonary elimination of chloroform and its meta-
bolites in man. Arch. Int. Pharmacodyn. 196: 98.
Hardie, D.W.F. 1964. Chlorocarbons and chlorohydrocarbons: chloroform.
In: Kirk-Othmer encyclopedia of chemical technology. 2nd ed. John Wiley
and Sons, Inc., New York.
Ilett, K.F., et al. 1973. Chloroform toxicity in mice: Correlation of
renal and hepatic necrosis with covalent binding of metabolites to tissue
macromolecules. Exp. Mol. Pathol. 19: 215.
Irish, O.D. 1972. Aliphatic halogenated hydrocarbons. In: Industrial
hygiene and toxicology. 2nd ed. John Wiley and Sons, Inc., New York.
Kutob, S.D. and G.L. Plaa. 1961. The effect of acute ethanol intoxication
on chloroform-induced liver damage. Jour. Pharmacol. Exp. Ther. 135: 245.
Lavigne, J.G. and C. Marchand. 1974. The role of metabolism in chloroform
hepatotoxicity. Toxicol. Appl. Pharmacol. 29: 312.
Lehmann, K.B. and Hassegawa. 1910. Studies of the absorption of chlori-
nated hyrocarbons in animals and humans. Archiv. fuer Hygiene. 72: 327.
McCabe, L.J. 1975. Association between trihalomethanes in drinking water
(NORS data) and mortality. Draft report. U.S. Environ. Prot. Agency.
»
McLean, A.E.M. 1970. The effect of protein deficiency and microsomal en-
zyme induction by DDT and phenobarbitone on the acute toxicity of chloroform
and pyrrolizidine alkaloid retrorsine. Brit. Jour. Exp. Pathol. 51: 317.
-S-71-
-------
National Academy of Sciences, 1978a. Nonfluorinated halomethanes in the
environment. Environ. Studies Board, Natl. Res. Council, Washington, D.C.
National Academy of Sciences/National Research Council. 1978b. Epidemiolo-
gical studies of cancer frequency and certain organic constituents of drink-
ing water - A review of recent literature for U.S. Environ. Prot. Agency.
National Cancer Institute. 1976. Report on carcinogenesis bioassay of
chloroform. Natl. Tech. Inf. Serv. PB-264018. Springfield, Va.
National Institute for Occupational Safety and Health. 1974. Criteria for
a recommended standard...Occupational exposure to chloroform. NIQSH Publ.
No. 75-114. Dept. Health Educ. Welfare, Washington, D.C.
Neely, W.B., et al. 1976. Mathematical models predict concentration-time
profiles resulting from chemical spill in river. Environ. Sci. Technol.
10: 72.
Schwetz, B.A., et al. 1974. Embryo and fetotoxicity of inhaled chloroform
in rats. Toxicol. Appl. Pharmacol. 28: 442.
Simmon, J.M., et al. 1977. Mutagenic activity of chemicals identified in
drinking water. In: D. Scott, et al., (ed.) Progress in genetic toxico-
logy. Elsevier/North Holland Biomedical Press, New York.
Taylor, D.C., et al. 1974. Metabolism of chloroform. II. A sex difference
in the metabolism of (14C)-chloroform in mice. Xenobiotica 4: 165.
Thompson, D.J., et al. 1974. Teratology studies on orally administered
chloroform in the rat and rabbit. Toxicol. Appl. Pharmacol. 29: 348.
Torkelson, T.R., et al. 1976. The toxicity of chloroform as determined by
single and repeated exposure of laboratory animals. Am. Ind. Hyg. Assoc.
Jour. 37: 697.
U.S. EPA. 1975. Development document for interim final effluent limita-
tions guidelines and new source performance standards for the significant
organic products segment of the organic chemical manufacturing point source
category. EPA-440/1-75/045. U.S. Environ. Prot. Agency, Washington, O.C.
U.S. EPA. 1978a. 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. 1978b. Office of Water Supply. Statement of basis and purpose
for an amendment to the national interim primary drinking water regulations
on trihalomethanes. Washington, D.C.
U.S. EPA. 1979a. Chloroform: Ambient Water Quality Criteria Document.
(Draft)
»
U.S. EPA. 1979b. Environmental Criteria and Assessment Office. Chloro-
form: Hazard Profile. (Draft)
-------
Van Dettingen, W.F. 1964. The hydrocarbons of industrial and toxicalogical
importance. Elsevier Publishing Co., New York.
Vessell, E.S., et al. 1976. Environmental and genetic factors affecting
the response of laboratory animals to drugs. Fed. Am. Soc. Exp. Biol. Proc.
35: 1125.
Windholz, M., ed. 1976. The Merck Index. 9th ed. Merck and Co., Inc.,
Rahway, N.j.
- 5-73-
-------
No. 48
Chloromethane
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.
-------
CHLOROMETHANE
SU14MARY
Chloromethane is toxic to humans by its action on the
central nervous system. In acute toxicity, symptoms consist
of blurring vision, headache, vertigo, loss of coordination,
slurring of speech, staggering, mental confusion, nausea,
and vomiting. Information is not available on chronic toxicity,
teratogenicity, or carcinogenicity. Chloromethane is highly
mutagenic to the bacteria, Salmonella tyghi.murj.urn.
Only three toxicity tests have been conducted on three
species of fish yielding acute values ranging from 147,000
to 300,000 )jg/l. Tests on aquatic invertebrates or plants
have not been conducted.
-------
CHLOROMETHANE
I. INTRODUCTION
This profile is based on the Ambient Water Quality
Criteria Document for Halomethanes (U.S. EPA, 1979a).
Chloromethane (CH.,C1; methyl chloride; molecular weight
50.49) is a colorless, flammable, almost odorless gas at
room temperature and pressure (Windholz, 1976}. Chloromethane
has a melting point of -97.7°C, a boiling point of -24.2°C,
a specific gravity of 0.973 g/ml at -10°C, and a water solubi-
-_
lity of 5.38 x 10 }ig/l. It is used as a refrigerant,
a methylating agent, a dewaxing agent, and catalytic solvent
in synthetic rubber production (MacDonald, 1964). However,
its primary use is as a chemical intermediate (Natl. Acad.
Sci., 1978). Chloromethane is released to the environment
by manufacturing and use emissions, by synthesis during
chlorination of drinking water and municipal sewage, and
by natural synthesis, with the oceans as the primary site
(Lovelock, 1975). 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 Chloromethane
qualitatively in finished drinking waters in the U.S. How-
ever, there are no data on its concentration in drinking
water, raw water, or waste-water (U.S. EPA, 1979a), probably
because it is more reactive than other chlorinated methanes
{Natl. Acad. Sci., 1978).
/
-S77-
-------
B. Food
There is no information on the presence of chloro-
methane in food. There is no bioconcentration factor for
chloromethane (U.S. EPA, 1979a) .
C. Inhalation
Saltwater atmospheric background concentrations
3
of chloromethane averaging about 0.0025 mg/m have been
reported (Grimsrud and Rasmussen, 1975; Singh, et al. 1977).
This is higher than reported average continental background
and urban levels and suggests that the oceans are a major
source of global chloromethane (National Acad. Sci., 1978).
Localized sources, such as burning of tobacco or other com-
bustion processes, may produce high indoor-air concentra-
tions of chloromethane (up to 0.04 mg/m ) (Natl. Acad. Sci.,
1978) . Chloromethane is the predominant halomethane in
indoor air, and is generally in concentrations two to ten
times ambient background levels.
III. PHARMACOKINETICS
A. Absorption
Chloromethane is absorbed readily via the lungs,
and to a less significant extent via the skin. Poisonings
involving gastrointestinal absorption have not been reported
{Natl. Acad. Sci., 1977; Davis, et al. , 1977).
B. Distribution
Uptake of chloromethane by the blood is rapid
but results in only moderate blood levels with continued
exposure. Signs and pathology of intoxications suggest
-578-
-------
wide tissue (blood, nervous tissue, liver, and kidney) distri-
bution of absorbed chloromethane (Natl. Acad . Sci., 1978).
C. Metabolism
Decomposition and sequestration of chloromethane
result primarily by reaction with sulfhydryl groups in intra-
cellular enzymes and proteins (Natl. Acad. Sci., 1977).
IV. EFFECTS
A. Carcinogenicity
Pertinent information could not be located in
the available literature.
B. Mutagenicity
Simmon and coworkers (1977) reported that chloro-
methane was mutagenic to Salmonella tryphimurium strain
TA 100 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. Also, Andrews, et al. (1976) have demonstrated
that chloromethane was mutagenic to S_._ typhimur ium strain
TA1535 in the presence and absence of added liver homogenate
preparations .
C. Teratogenicity and Other Reproductive Effects
Information on positive evidence of teratogenisis
or other reproductive effects was not available in the literature.
D. Chronic Toxicity
Under prolonged exposures to chloromethane (dura-
»
tion not specified) increased mucous flow and reduced mucosta-
-579-
-------
tic effect of other agents (e.g., nitrogen oxides) were
noted in cats (Weissbecker , et al., 1971).
E. Other Relevant Information
In acute human intoxication, chloromethane pro-
duces central nervous system depression, and systemic poison-
ing cases have also involved hepatic and renal injury (Hansen,
er al., 1953; Spevac, et al . , 1976).
V. AQUATIC TOXICITY
A. Acute Toxic ity
A single 96-hour static renewal test serves as
the only acute study for freshwater providing an adjusted
LC5Q value of 550,000 ug/1 for the bluegill sunfish (Lepomij;
macrochirus) . (Dawson, et al., 1977). Studies on fresh-
water invertebrates were not found. For the marine fish,
the tidewater silversides (Menidia bejryljlina) , a 96-hour
static renewal assayed provided an LC5Q value of 270,000
ug/1 (Dawson, et al., 1977). Acute studies on marine inverte-
brates were not found.
B. Chronic Toxicity
In a review of the available literature, chronic
testing with chloromethane has not been reported.
C. Plant Effects
Pertinent information could not be located in the
available literature.
VI. EXISTING GUIDELINES AND STANDARDS
Neither the human nor the aquatic criteria derived
by U.S. EPA, 1979a, which are summarized below, have gone
-580-
-------
through the process of public review; .therefore, there is
a possibly that these criteria may be changed.
A. Human
OSHA (1976) has established the maximum acceptable
time-weighted average air concentrations for daily eight-
hour occupational exposure at 210 mg/m . The U.S. EPA (1979a)
Draft Water Quality Criteria for Chloromethane is 2 ug/1.
Refer to the Halomethanes Hazard Profile for discussion
of criteria derivation {U.S. EPA, 1979b) . •-
B. Aquatic
Criterion recommended to protect freshwater or-
ganisms have been drafted as 7,000 pg/1, not to exceed 16,000
pg/1 for a 24-hour average concentration. For marine life,
the criterion has been drafted as 3,700 ^ug/1, not to exceed
8,400 jjg/1 as a 24-hour average concentration.
-------
CHLOROMETHANE
REFERENCES
Andrews, A.W., et al. 1976. A comparison of the mutagenic
properties of vinyl chloride and methyl chloride. Mutat.
Res. 40: 273.
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, of Contract
No. 68-01-4315. Off. Toxic Subst., U.S. Environ. Prot.
Agency, Washington, D.C.
Dawson, G.W. , et al. 1977. The acute tox-icity of 47 indus-
trial chemicals to fresh and saltwater fishes. Jour. Hazard.
Mater. 1: 303.
Grimsrud, E.P., and R.A. Rasmussen. 1975. Survey and an-
alysis of halocarbons in the atmosphere by gas chromatography-
mass spectrometry. Atmos. Environ. 9: 1014.
Hansen, H., et al. 1953. Methyl chloride intoxification:
Report of 15 cases. AMA Arch. Ind. Hyg. Occup. Med. 8:
328.
Lovelock, J.E. 1975. Natural halocarbons in the air and
in the sea. Nature 256: 193.
MacDonald, J.D.C. 1964. Methyl chloride intoxication.
Jour. Occup. Med. 6: 81.
National Academny of Sciences. 1977'. Drinking water and
health. Washington, D.C.
Nation-al Academy of Sciences. 1978. Nonfluorinated halo-
methanes in the environment. Washington, D.C.
Occcupational Safety and Health Administration. 1976.
General industry standards. OSHA 2206, revised January,
1976. U.S. Dep. Labor., Washington, D.C.
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, N^ and other atraOspheric constituents.'
Atmos. Environ. 11: 819.
-------
Spevac, L., et al. 1976. Methyl chloride poisoning in
four members of a family. Br. Jour. Ind. Med. 33: 272.
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. 1979a. Halomethanes: Ambient Water Quality Cri-
teria (Draft) .
U.S. EPA. 1979b. Environmental Criteria and Assessment
Office. Haloraethanes: Hazard Profile (Draft).
Weissbecker, L., et al. 1971. Cigarette smoke and tracheal
mucus transport rate: Isolation of effect of components
of smoke. Am. Rev. Resp. Dis. 104: 182.
Windholz, M., (ed.) 1976. The Merck Index. Merck and Co.,
Rahway, N.J.
-------
No. 49
2-Chloronaphthalene
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, B.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.
-------
2-CHLORONAPHTHALENE
SUMMARY
Monochlorinated naphthalenes are relatively insoluble in
water. They can be slowly degraded by bacteria and are subject
to photochemical decomposition. Monochlorinated naphthalenes
appear to bioconcentrate in plants and animals exposed to the
substances. 2-Chloronaphthalene has been identified as a pol-
lutant in a variety of industries.
No information was located on the carcinogenicity, mutagen-
icity, or teratogenicity of 2-chloronaphthalene or other mono-
chlorinated naphthalenes. The metabolism of some chlorinated
naphthalenes, however, proceeds through an epoxide mechanism. If
an epoxide is formed as an intermediate in the metabolism of 2-
chloronaphthalene, it could react with cellular macromolecules
possibly resulting in cytotoxicity, mutagenicity, oncogenicity,
or other effects.
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria
Document for Chlorinated Naphthalenes (U.S. EPA, 1979b).
2-Chloronaphthalene (C^H^Cl; molecular weight 162) is a
crystalline solid with a melting point of 61°C and a boiling
point of 256"C. Its density at 168C is 1.27. It is insoluble in
*
water and soluble in many organic solvents (Weast; 1972 and
Hardie, 1964).
-------
A review of the production range (includes importation)
statistics for 2-chloronaphthalene (CAS. No, 91-58-7) which is
listed in the inital TSCA Inventory (1979a) has shown that
between 1,000 and 9,000 pounds of this chemical were
produced/imported in 1977 ,_V
Monochloronaphthalenes and mixtures of mono- and dichloro-
naphthalenes have been used for chemical-resistant gauge fluids
and instrument seals, as heat exchange fluids, high-boiling
specialty solvents {e.g., for solution polymerization), color
dispersions, engine crankcase additives to dissolve sludges and
gums, and as ingredients in motor tuneup compounds. Monochloro-
naphthalene was formerly used as a wood preservative (Dressier,
1979).
II. EXPOSURE
A. Environmental Fate
Polychlorinated naphthalenes do not occur naturally in the
environment. Potential environmental accumulation can occur
around points of manufacture of the compounds or products
containing them, near sites of disposal of polychorinated
naphthalene-containing wastes, and, because polychlorinated
_V 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).
-------
biphenyls (PCBs) are to some extent contaminated by polychlori-
nated naphthalenes (Vos et._ a±. 1970; Bowes et_ ai_. 1975) near
sites of heavy PCB contamination.
Because polychlorinated naphthalenes are relatively insol-
uble in water, they are not expected to migrate far from their
point of disposition. The use of mono- and dichlorinated naphtha-
lenes as an engine oil additive and as a polymerization solvent
in the fabric industry suggests possible contamination of soil or
water.
Walker and Wiltshire (1955) found that soil bacteria when
first grown on naphthalene could also grow on 1-chloronaph-
thalene, producing a diol and chlorosalicylic acid. Canonica et
al. (1957) found similar results for 2-chloronaphthalene. Okey
and Bogan (1965) studied the utilization of chlorinated sub-
strates by activated sludge and found that naphthalene was
degraded at a fairly rapid rate, while 1-and 2-chloronaphthalenes
were handled more slowly.
Ruzo _et_ _a3L. (1975) studied the photodegradation of 2-chloro-
naphthalene in methanol. The major reaction pathways seen were
dechlorination and dimerization. Jaffe and Orchin (1966) indi-
cated that any 2-chloronaphthalene present at the surface of
water could be degraded by sunlight to naphthalene. In the
aquatic environment, 2-chloronaphthalene can exist as a surface
film, be adsorbed by sediments, or accumulated by biota.
-------
B. Bioconcentration
Monochlorinated naphthalenes appear to bioconcentrate in the
aquatic environment. Adult grass shrimp (Palaemonetes pugio)
were exposed to a mixture of mono- and dichloro naphthalenes for
15 days. The concentration of chloronaphthalenes detected in the
shrimp was 63 times that of the experimental environment. When
removed from the contaminated environment, however, the concen-
tration in the shrimp returned to virtually zero within 5 days
(Green and Neff, 1977).
Erickson et_ ^1_. (1978a) reported a higher relative biocon-
centration of the lower chlorinated naphthalenes in the fruit of
apple trees grown on contaminated soil. The soil was found to
have a polychlorinated naphthalene level of 190 ug/kg of which
1.6 ug/kg consisted of monochloronaphthalenes. While the apples
grown on this soil had only 90 ug/kg of polychlorinated naphtha-
lenes, the level of monochloronaphthalene was &2 ug/kg.
C. Environmental Occurrence
2-Chloronaphthalene has been identified as a pollutant in a
variety of industries, e.g. organic chemical, rubber, power
generation, and foundries (U.S. EPA, 1979c).
Chlorinated naphthalenes have been found more consistently
in air and soil samples than in associated rivers and streams
(Erickson _et_ ajU, 1978b) . The air samples contained mainly the
mono-, di- and trichlorinated naphthalenes, while soil contained
t
mostly the tri-, tetra- and pentachlorinated derivatives.
To date polychlorinated naphthalenes have not been identi-
fied in either drinking water or market basket food. The Food
and Drug Administration has had polychlorinated naphthalene
-------
monitoring capability for foods since 1970, but has not reported
their occurrence in food (U.S. EPA, 1975).
III. PHARMACOKINETICS
Ruzo et_ al_. (1976b) reported the presence of 2-chloronaph-
thalene in the brain, kidney, and liver of pigs six hours after
injection. Small concentrations of 3-chloro-2-naphthol, a
metabolite , were seen in the kidney and liver with large amounts
occurring in the urine and bile. The metabolism of some chlori-
nated napthalenes proceeds through an epoxide mechanism (Ruzo et
al. 1975, 1976ab; Chu &t_ al_. , 1977ab) .
IV. HEALTH EFFECTS
A. Teratogenicity, Mutagenicity, and Carcinogenicity
No information was located on the carcinogenicity, muta-
genicity, or teratogenicity of polychlorinated naphthalenes.
If an epoxide is formed as an intermediate in the metabolism
of 2-chloronaphthalene, it could react with cellular macromole-
cules. Binding might occur with protein, RNA, and DNA resulting
in possible cytotoxicity, mutagenicity, oncogenicity, or other
effects (Garner, 1976; Heidelberger, 1973; Wyndham and Safe,
1978).
B. Other Toxity
In man, the first disease recognized as being associated
f
with occupational exposure to higher polychlorinated naphthalenes
was chloracne. Occurrence of this disease was associated with
the manufacture or use of polychloronaphthalene-treated electri-
cal cables. Kleinfeld et^ al_. (1972) noted that workers at
-590-
-------
an electric coil manufacturing plant had no cases of chloracne
while using a mono- and dichloronaphthalene mixture. When a
tetra-/pentachlorinated naphthalene mixture was substituted for
the original mixture, 56 of the 59 potentially exposed workers
developed chloracne within a "short" time.
The lower chlorinated naphthalenes appear to have low acute
toxicity. Mixtures of mono-/dichloronaphthalenes and tri-/tetra-
chloronaphthalenes at 500 mg/g in a mineral oil suspension
applied to the skin of the human ear caused no response over a
30-day period. A mixture of penta-/hexachloronaphthalenes given
under the sane conditions caused chloroacne (Shelley and Kligman,
1957}.
The oral LD50 for rats and mice is 2078 mg/kg and 886 mg/kg
respectively (NTIOSH, 1978). No mortality or illness was reported
in rabbits given 500 mg/kg orally (Cornish and Block, 1958).
V. AQUATIC EFFECTS
The LC50 (ppb) of a mixture of 60% mono- and 40% dichloro-
naphthalenes in grass shrimp (Palaemonetes pugio)is as follows:
72-hr 96-hr
post larval stage - 449
adult 370 325
(Green and Neff, 1977)
VI. EXISTING GUIDELINES
r
There are no existing guidelines for 2-chloronaphthalene.
-------
BIBLIOGRAPHY
Bowes, G. W. et al. 1975. Identification of chlorinated diben-
zofurans in American polychlorinated biphenyls. Nature 256, 305.
(as cited in U.S. EPA, 1979b).
Canonica, L. et^ _al_. 1957. Products of microbial oxidation of
some substituted naphthalenes. Rend. 1st. Lombardo Sci. 91, 119-
129 (Abstract).
Cornish H.H., and W.D. Block. 1958. Metabolism of chlorinated
naphthalenes. J. Biol. Chem. 231, 583. (as cited in U.S. EPA,
1979b).
Chu, I., et al. 1977a. Metabolism and tissue distribution of
(1,4,5,8-^cT-l* 2-dichloronaphthalene in rats. Bull. Environ.
Contain. Toxicol. 18, 177. {as cited in U.S. EPA, 1979b) .
Chu, I., et al. 1977b. Metabolism of chloronaphthalenes. J.
Agric. Pood Chem. 25, 881. (as cited in U.S. EPA, 1979b).
Dressier, H. 1979. Chlorocarbons and chlorohydrocarbons:
chlorinated naphthalenes. In. Standen A. ed. Kirk-Othmer
Encyclopedia of Chemical Technology, 3rd ed. New York: John Wiley
and Sons, Inc.
Erickson, M.D., ^t^ ^1_. 1978a. Sampling and analysis for
polychlorinated naphthalenes in the environment J. Assoc. Off.
Anal. Chem. 61, 1335. (as cited in U.S. EPA, 1979b).
Erickson, M.D., et al. 1978b. Development of methods for
sampling and analysis of polychlorinated naphthalenes in ambient
air. Environ. Sci. Tech. 12(8), 927-931.
Garner, R.C. 1976. The role of epoxides in bioactivation and
carcinogenesis. In: Bridges, J. W. and L. F. Chasseaud, eds.
Progress in drug metabolism, Vol. 1. New York: John Wiley and
Sons. pp. 77-128.
Green, F. A., Jr. and J. M. Neff. 1977. Toxicity, accumulation,
and release of three polychlorinated naphthalenes (Halowax 1000,
1013, and 1099) in postlarval and adult grass shrimp,
Palaemonetes pugio. Bull. Environ. Contam. Toxicol. 14, 399.
Hardie, D.W.F. 1964. Chlorocarbons and chlorohydrocarbons:
chlorinated naphthalenes. In: Kirk-Othmer Encyclopedia of ,
.Chemical Technology. 2nd ed. John Wiley and Sons. Inc., New York.
Heidelberger, C. 1973. Current trends in carcinogenesis. Proc.
Fed. Am. Soc. Exp. Biol. 32,2154-2161.
Jaffe, H. H. and M. Orchin. 1966. Theory and aplication of
ultraviolet spectroscopy^ Wiley Pub. New York, 624pp.
-------
Kleinfeld, M. , et_ _a^. 1972. Clinical effects of chlorinated
naphthalene exposure. J. Occup. Med. 14_, 377-379, (as cited in
U.S. EPA, 1979b).
National Institute of Occupational Safety and Health. 1978.
Registry of Toxic Effects of Chemical Substances. DREW Publ. No.
79-100.
Okey, R. W. and R. H. Bogan. 1965. Apparent involvement of
electronic mechanisms in limiting microbial metabolism of
pesticides. J. Water Pollution Contr. Fedr. 37, 692.
Ruzo, L.O., £t_ _al_. 1975. Hydroxylated metabolites of chlo-
rinated naphthalenes (Halowax 1031) in pig urine. Chemosphere _3_,
121-123.
Ruzo, L. O., et al. 1976a. Metabolism of chlorinated
naphthalenes. . J. Agric. Food Chem. 24, 581-583.
Ruzo, L.O., et al. 1976b. Uptake and distribution of
chloronaphthalenes and their metabolities in pigs. Bull.
Environ. Contain. Toxicol. 16(2), 233-239.
Shelley, W. B., and A. M. Kligman. 1957. The experimental
production of acne by penta-and hexachloronaphthalenes. A.M.A.
Arch. Dermatol. 75, 639-695. (as cited in U.S. EPA, 1979b).
U.S. EPA. 1975. Environmental Hazard Assessment Report:
Chlorinated Naphthalenes. (EPA 560/8-75-001).
U.S. EPA. 1979a. Toxic Substances Control Act Chemical
Substance Inventory, Production Statistics for Chemicals on the
Non-Confidential Initial TSCA Inventory, .
U.S. EPA. 1979b. Ambient Water Quality Criteria: Chlorinated
Naphthalenes, PB-292-426.
U.S. EPA. Unpublished data obtained from the U.S. EPA
Environmental Research Laboratory, Athens, Georgia, February 22,
I979c.
Vos, J.G., et_ al_. 1970. Identification and toxicological evalu-
ation of chlorinated dibenzofurans and chlorinated naphthalenes
in two commercial polychlorinated biphenyls. Food Cosmet.
Toxicol. 8_, 625. (as cited in U.S. EPA, 1979b)
Walker, N. and G.H. Wiltshire. 1955. The decomposition of 1-
chloro- and 1-bromonaphthalene by soil bacteria. J. Gen.
Microbiol. 12, 478-483.
jar
-------
Weast, R.C., ed. 1972. CRC Handbook of Chemistry and Physics.
CRC Press, Inc., Cleveland, Ohio.
Wyndham, D. , and S. Safe. 1978. In vitro metabolism of 4-
Sf
-------
No. 50
2-Chlorophenol
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
-S'lS'-
-------
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.
-------
2-CHLOROPHENOL
SUMMARY
Insufficient data exist to indicate that 2-chlorophenol
is a carcinogenic agent. 2-Chlorophenol appears to act as a
nonspecific irritant in promoting tumors in skin painting
studies. No information is available on mutagenicity, tera-
togenicity, or subacute and chronic toxicity. 2-Chlorophenol
is a weak uncoupler of oxidative phosphorylation and a con-
vulsant.
2-Chlorophenol is acutely toxic to freshwater fish at
"cbncen"tFatTons~rang ing from 6 ,590 ~to~207r70~ug7r. ~No~mar"ine
studies are available. Concentrations greater than 60 ug/1 .
have been reported to taint cooked rainbow trout flesh in
flavor impairment studies.
-S97-
-------
I. INTRODUCTION
This profile is based primarily on the Ambient Water
Quality Criteria Document for 2-Chlorophenol (U.S. EPA,
1979).
2-Chlorophenol (ortho-chlorophenol) is a liquid having
the empirical formula CgHgCl (molecular weight: 128.56).
It has the following physical/chemical properties (Rodd,
1954; Judson and Kilpatrick, 1949; Sax, 1975; Stecher, 1968;
Henshaw, 1971):
Melting Point: 8.7°C
Boiling Point Range: 175-176°C
Vapor Pressure: 1 mm Hg at 12,1°C
Solubility: Slightly soluble (lg/1)
in water at 25°C and
neutral pH
2-Chlorophenol is a commercially produced chemical used
as an intermediate in the production of higher chlorophenols
and phenolic resins and has been utilized in a process for
extracting sulfur and nitrogen compounds from coal (U.S. EPA,
1979).
2-Chlorophenol undergoes photolysis in aqueous solutions
as a result of UV irradiaton (Omura and Matsuura, 1971;
Joschek and Miller, 1966). Laboratory studies suggest that
microbial oxidation could be a degradation route for 2-chlo-
rophenol (Loos, et al., 1966; Sidwell, 1971; Nachtigall and
Butler, 1974). However, studies performed by Ettinger and
Ruchhoft (1950) on the persistency of 2-chlorophenol in sew-
age and polluted river water indicated that the removal of
f
monochlorophenols requires the presence of an adapted micro-
flora.
-------
II. EXPOSURE
A. Water
The generation of waste from the commercial produc-
tion and use of 2-chlorophenol (U.S. EPA, 1979} and the inad-
vertent synthesis of 2-chlorophenol due to chlorination of
water contaminated with phenol (Aly, 1968: Barnhart and Camp-
bell, 1972; Jolley, 1973; Jolley, et al., 1975) are potential
sources of contamination of water with 2-chlorophenol. How-
ever, no data regarding 2-chlorophenol concentrations in fin-
ished drinking water are available (U.S. EPA, 1979).
B. Food
Information on levels of 2-chlorophenol in foods is
not available. Any contamination of foods is probably indi-
rect as a result of use and subsequent metabolism of phenoxy-
alkanoic herbicides (U.S. EPA, 1979). Although residues of
2,4-dichlorophenol were found in tissues of animals fed 2,4-D
and nemacide containing food (Clark, et al. 1975); Sherman,
et al. 1972) , no evidences were cited to indicate the pres-
ence of 2-chlorophenol; moreover, there was no contamination
of 2-chlorophenol in milk and cream obtained from cows fed
2,4-D treated food (Bjerke, et al. 1972).
The potential for airborne exposure to 2-chloro-
phenol in the general environment, excluding occupational ex-
posure, has not been reported (U.S. EPA, 1979).
The U.S. EPA (1979) has estimated the weighted
average bioconcentration factor for 2-chlorophenol and the*
edible portion of fish and shellfish consumed by Americans at
-------
490. This estimate is based on measured steady state biocon-
centration studies in bluegills. -
C. Inhalation
Pertinent data regarding concentrations of 2-chloro-
phenol in ambient air could not be found in the available
literature.
III. PHARMACOKIN ETICS
A. Absorption
Data dealing directly with the absorption of 2-
chlorophenpl by humans and experimental animals has not been
found. Chlorophenol compounds are generally considered to be
readily absorbed, as would be expected from their high lipid
solubility and low degree of ionization at physiological pH
(Doedens, 1963; Farquharso'n, et al., 1958). Toxicity studies
indicate that 2-chlorophenol is absorbed through the skin.
B. Distribution
Pertinent data regarding tissue distribution of 2-
chlorophenol was not located in the available literature.
C. Metabolism
Data regarding the metabolism of 2-chlorophenol in
humans was not available (U.S. EPA, 1979). Based on experi-
mental work in two species, it appears that the metabolism of
2-chlorophenol in mammals is similar to that of phenol in
regard to the formation and excretion of sulfate and glucur-
onide conjugates (Von Oettingen, 1949; Lindsay-Smith, et al.
»
1972) Conversion of chlorobenzene to monochlorophenols,
including 2-chlorophenol, has been shown _in vitro with rat
-(,00-
-------
liver (Selander, et al. 1975) and in vivo, w-ith rabbits
(Lindsay-Smith, et al. 1972).
D. Excretion
Studies on rate and route of excretion for
2-chlorophenol in humans were not available. Dogs excreted
87 percent of administered 2-chlorophenol in the urine as
sulfate and glucuronide conjugates (Von Oettingen, 1949).
The same metabolites were found in the urine of rabbits after
administration of chlorobenzene (Lindsay-Smith, et al. 1972);
however, out of the total free and conjugated chlorophenols
only 6 percent were present as 2-chlorophenol.
IV. EFFECTS
A. Carcinogenicity
Insufficient data exist to indicate that 2-chloro-
phenol is a carcinogen. In the only study found (Boutwell
and Bosch, 1959), 2-chlorophenol promoted skin cancer in mice
after initiation with dimethylbenzanthracene and when repeat-
edly applied at a concentrations high enough to damage the
skin. 2-Chlorophenol was not carcinogenic when applied re-
peatedly without initiation with dimethylbenzanthracene, but
did induce a high incidence of papillomas and no carcinomas.
Information regarding mutagenicity, teratogenicity,
other reproductive effects and chronic toxicity could not be
found in the available literature.
F. Other Relevant Information
2-Chlorophenol is a weak uncoupler of oxidative
phosphorylation (Mitsuda, et al., 1963) and a convulsant
(Farquharson, et al., 1958; Angel and Rogers, 1972).
-------
V. AQUATIC TOXICITY
A. Acute Toxicity
Acute studies on four species of fish have produced
96-hour static LC5Q values ranging from 6,590 ug/1 i° the
bluegill (Lepomis macrochirus) (U.S. EPA, 1978) to 20,170
ug/1 to the guppy (Poecilia reticulatus). Juvenile bluegills
were more sensitive in a static renewal assay with an LC50
value of 8,400 ug/1. The fathead minnow (Pimephales prome-
las) was the only freshwater fish tested in a flow through
system and gave an LC50 value of 12,380 ug/1. Daphnia
magna has been found to have 48-hour static LC50 values
of 2,580 y.g/1 and 7,430 ug/1. No data concerning the effects
of 2-chlorophenol to marine fish or invertebrates are avail-
able.
B. Chronic Toxicity
Effects were not obtained in a chronic embryo-
larval test of 2-chlorophenol at concentrations as high as
1,950 ug/1 for the freshwater fathead minnow. Additional
chronic studies are not available.
C. Plant Effects
The only plant assay available provides an effec-
tive concentration of 500,000 ug/1 in chlorophyll reduction
in the algae, Chlorella pyrenoidosa.
D. Residues
A measured bioconcentration factor of 214 has been
obtained for the bluegill. The half-life was less than one"
day, indicating a rapid depuration rate for 2-chlorophenol.
-------
E. Miscellaneous
Flavor impairment of the edible portion of fish
exposed to 2-chlorophenol has been reported. The highest
concentration of 2-chlorophenol in the exposure water which
would not impair the flavor of cooked rainbow trout (Salmo
gairdneri) has been estimated at 60 ug/1 Shumway and
Palensky, 1973).
VI. EXISTING GUIDLINES 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 may be changed.
A. Human
Based on the prevention of adverse organoleptic ef-
fects, the U.S. EPA (1979) draft interim criterion recommend-
ed for 2-chlorophenol in ambient water is 0.3 ug/1- There
are no other standards or guidelines for exposure to 2-chlo-
rophenol.
B, Aquatic
Based on the tainting of fish, the draft criterion
to protect freshwater organisms from 2-chlorophenol is 60
ug/1 as a 24-hour average, not to exceed 180 ug/1 at any
time. No criterion was derived for marine life (U.S. EPA,
1979).
if
-------
2-CHLOROPHENOL
REFERENCES
Aly, O.M. 1968. Separation of phenols in waters by thin
layer chromatography. Water Res. 2: 587.
Angel, A., and K.J. Rogers. 1972. An analysis of the con-
vulsant activity of substituted benzenes in the mouse. Tox i-
col. Appl. Pharmacol. 21: 214.
Barnhart, E.L., and G.R. Campbell. 1972. The effect of
chlorination on selected organic chemicals. U.S. Government
Printing Office, Washington, D.C.
Bjerke, E.L., et al. 1972. Residue study of phenoxy herbi-
cides in milk and cream. Jour. Agric. Food Chem. 20: 963.
Boutwell, R.K., and O.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.
Doedens, J.D. 1963. Chlorophenols. Page 325 _in Kirk-Othmer
encyclopedia of chemical technology. John Wiley and Sons,
Inc., New York.
Ettinger, M.B., and C.C. Ruchhoft. 1950. Persistence of
monochlorophenols in polluted river water and sewage dilu-
tion. U.S. Pub. Health Serv., Environ. Health Center, Cin-
cinnati, Ohio.
Farquharson, M.E., et al. 1958. The biological action of
chlorophenols. Br. Jour. Pharmacol. 13: 20.
Henshaw, T.B. 1971. Adsorption/filtration plant cuts
phenols from effluent. Chem. Eng. 76: 47.
Jolley, R.L. 1973. Chlorination effects on organic
constituents in effluents from domestic sanitary sewage
treatment plants. Ph.D. dissertation. University of
Tennessee.
Jolley, R.L., et al. 1975. Chlorination of cooling water: A
source of environmentally significant chlorine-containing
organic compounds. Proc. 4th Natl. Symp. Radioecology.
Corvallis, Ore. •
Joschek, H.I., and S.I. Miller. 1966. Photocleavage of
phenoxyphenols and bromophenols. Jour. Am. Chem. Soc. 88:
3269.
-------
Judson, C.M., and M. Kilpatrick. 1949. The effects of sub-
stituents on the dissociation constants o.f substituted
phenols. I. Experimental measurements in aqueous solutions,
Jour. Am. Chem. Soc. 74: 3110.
Lindsay-Smith, J.R., et al. 1972. Mechanisms of mammalian
hydroxylation: Some novel metabolites of chlorobenzene.
Xenobiotica 2: 215.
Loos, M.A., et al. 1966. Formation of 2,4-dichlorophenol
and 2,4-dichlorophenoxyacetate by Arthrobacter Sp. Can.
Jour. Microbiol. 13: 691.
Mitsuda, H., et al. 1963. Effect of chlorophenol analogues
on the oxidative phosphorylation in rat liver mitochondria.
Agric. Biol. Chem. 27: 366.
Nachtigall, H., and R.G. Butler. 1974. Metabolism of
phenols and chlorophenols by activated sludge microorganisms.
Abstr. Annu. Meet. Am. Soc. Microbiol. 74: 184.
Omura, K., and T. Matsuura. 1971. Photoinduced reactions -
L Photolysis of halogenophenols in aqueous alkali and in
aqueous cyanide. Tetrahedron 27: 3101.
Rodd, E.H. 1954. Chemistry of carbon compounds. III-A.
Aromatics. Elsevier Publishing Co., Amsterdam.
Sax, H.I. 1975. Dangerous properties of industrial mate-
rials. 4th ed. Van Nostrand Reinhold Co., New York.
Selander, H.G., et al. 1975. Metabolism of chlorobenzene
with hepatic microsomes and soluble cytochrome ?45o Sys-
tem. Arch. Biochem. Biophys. 168: 309
Sherman, J., et al. 1972. Chronic toxicity and residues
from feeding nemacide 0(2,4-dichlorophenol) 0, 0-diethylphos-
phorothioate to laying hens. Jour. Agric. Food Chem. 23:
617.
Shumway, D.L., and J.R. Palensky. 1973. Impairment of the
flavor of fish by water pollutants. EPA-R3-73-010. U.S.
Environ. Prot. Agency, U.S. Government Printing Office,
Washington, D.C.
Sidwell, A.E. 1971. Biological treatment of chlorophenolic
wastes - the demonstration of a facility for the biological
treatment of a complex chlorophenolic waste. Water Pollut.
Control Res. Ser. 12130 EKG.
»
Stecher, P.G., ed. 1968. The Merck Index. 8th ed. Merck
and Co., Rahway, N.J.
-------
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. 2-Chlorophenol: Ambient Water Quality Cri-
teria (Draft).
Von Oettingen, W.F. 1949. Phenol and its derivatives: the
relation between their chemical constitution and their effect
on the organism. Natl. Inst. Health Bull. 190: 193.
-------
No. 51
Chromium
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCT?
WASHINGTON, D.C. 20460
APRIL 30, 1980
-£,07-
-------
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.
-------
CHROMIUM
Summary
Hexavalent chromium, at low concentrations in water, has
a deleterious effect on the growth of fishes, aquatic inver-
tebrates, and certain species of algae. For the most sensi-
tive aquatic species, Daphnia magna, a final chronic no-ef-
fect level of less than 10 yg/1 has been derived by the U.S.
EPA. For trivalent chromium, toxic effects are more pro-
nounced in soft than in hard water; chronic no-effect levels
are derived as a function of water hardness.
Several hexavalent chromium compounds have produced tu-
mors at site of administration in animal studies. Human epi-
demiology studies indicate a possible etiology of chromium
exposure in the production of lung tumors in occupationally
exposed workers. Trivalent chromium has not shown carcino-
genic effects.
Mutagenic effects, including cytogenetic effects in ex-
posed workers, have been reported for hexavalent chromium
compounds. Trivalent chromium compounds were not mutagenic
in the Ames bacterial assay. Teratogenic effects of chromium
have been reported in a single study and have not been con-
firmed .
Impairment of pulmonary function has been reported in
chrome electroplating workers subject to chronic chromium ex-
posure. However, exposure to multiple agents complicates the
»
interpretation of this finding.
-------
CHROMIUM
I. INTRODUCTION
This profile is based on the Ambient Water Quality Cri-
teria Document for Chromium (U.S. EPA, 1979).
Chromium (Cr) is a steel gray, lustrous, hard metal that
melts at 1857 + 20°C, boils at 2672°C, and has a specific
gravity of 7.18 to 7.20 at 20°C {Weast, 1974). Chromium com-
pounds exist in a variety of oxidation states; the most com-
monly occurring are those of the trivalent and hexavalent
states. Physical properties of some chromium compounds are
summarized in Table 1.
Chromium compounds are utilized in the paint and dye in-
dustries as pigments and mordants, in metallurgy for the pro-
duction of stainless steel and other alloys, in the chrome
tanning of leather goods, in the production of high melting
refractory materials, and for chrome plating.
Hexavalent chromium compounds are relatively water sol-
uble and are readily reduced to more stable and insoluble tri
valent forms by reactions with organic reducing matter. Tri-
valent chromium forms stable hex accordinate complexes with a
great variety of ligands (water, ammonia, urea, halides, sul-
fates, ethylene diamine, organic acids) (U.S. EPA, 1978). In
neutral and basic solutions, trivalent chromium may form poly
nuclear bridge compounds that eventually precipitate (U.S.
EPA, 1978). Hexavalent chromium exists in solution as a com-
ponent of an an ion (hydrochromate, chromate, or dichromatej
and does not precipitate from alkaline solution. The anionic
form of hexavalent chromium is dependent on pH - in the acid
/
-t/O-
-------
environmental pH range hydrochromate predominates, while in
the alkaline range the predominant form is chromate (Trama and
Benoit, 1960).
Since chromium is an element, it will persist indefinite-
ly in the environment in some form. Trivalent chromium com-
pounds are more likely to accumulate in sediments, while hexa-
valent forms would remain soluble and dissipate with the water
flow (U.S. EPA, 1979).
-------
Table 1. Physical Properties of Typical Chromium Compounds
Compound
UxiJation state 0
Chrciuiura carbonyl
l)ib«n»en«
chromium(O)
tiHlduLlon atute t 1
UI3( Ll|)t;2Cr Broun
crystals
(CfiH C6Hs)2CrJ Orange plates
Cr-(C,,H,0-), -2H,0 Red crystals
c 4 j £ 4 C
CrCl2 White
crystals
CrSO,'(NII,),S011.6H 0 Blue crystals
H Ht1! c.
CrCl, Bright purple
3 plates
CrfCIUCQCHCOCK ), Red-violet
* crystals
KCr(SO,J2' 12HpO Deep purple
crystals
(Cr(H 0),.C12)C1'2H-0 Bright green
crystals
(Cr(H20)6)Cl Violet
crystals
Crystal ayatem Density
and apace group (g/cm )
Orthorhombio, c' '-^ift
Cubic, Paa . 1.519
j
l.617lfi
MonocllnlO) C2/o 1.79
Tetragonal, D,1^ 2.93
C
Honocllnlo, C^.
to
Hexagonal > D^ 2,87
j • *-3
Honocllnlo 1.31
t
Cubic, A° llB26i5
Trlolinicor 1.8352j
monocllnlo
Rhoabohedral, D
Halting Boiling
point point
/ °r \ / r \
1 C/ \ v)
150 151
(decomposed) (decomposed)
(sealed tube)
2611-285 Sublimes 150
(vacuum)
178 Decomposes
815 1120
Sublimes 885
208 . 315
89
( inaongruent )
95
90
Solubility
Slightly soluble in
CC1U; insoluble in
H207 (C^I^O,
Insoluble In HJ);
soluble in CfaH6
Soluble in
C2H5°H> C5H5N
Slightly soluble In
tip; soluble in
solids
Soluble in H,,0 to blue
solution, absorbs 0 '
Soluble in H^O,
absorbs 0
Insoluble in H^O,
soluble in presence
of Cr
Insoluble in H^O;
soluble in ^tt\._
Soluble in H.,0
Soluble in HgO, green
solution turning
green-violet
Soluble In 11,0, violet
solution tDrnlng
green-violet
-------
Compound Formula
Chromic oxide CrJ3~
c J
Oxidation state t '1
Chroaluu <1V) oxide CrO-
Chromium (JV) CrCl^
chloride
Oxidation state - 5
barium chroraateUY) Ba,(CrOh),
3 12
I
£N Oxidution state * 6
— X ChromluotWl) CrO,
f oxide
1
Cdromyl chloride CrO-Cl.
Ammonium (NH^-jCr.^
ditflrouate
Potassium K Cr?0
dichromute
Sodium dlchromate ILCr^O, • 21^0
Potassium cliromate Kr.CrO.,
d "
Sodium chromate Na?CrO^
* ^
Potassium ctilor^)- KCrO-,Cl
chruuiale
Sllvur clironute Ag-.CrO,,
°2 1
Barium chraoiijte BaCrO^
Appearance
Green powder
or crystals
Dark Brown or
black powder
Black-green
crystals
Ruby-red
crystals
Cherry-red
liquid
Red-orange
crystals
Orange-red
crystals
Orange-red
crystals
Yellow
crystals
Yellow
crystals
Orange
crystals
Maroon
crystals
Haln yellow
solid
Crystal syaten Density
and space group (g/cm )
Hhombohodral, D, 5.22._
JO tj
til
Tetragonal, D^ 1.98
(calculated)
Stable only at
high temp.
Same aa
Ca (PO, )
3 *l 2
*
ifi
Orthorhombic, c'° 2.?
1.91152S
Monoo linlo 2.1552c
Triolinio . 2.fi7625
Konoclinlo I.Blftp';
3
Orthorhonbio 2'732)fl
l*t
Orthorhonblo, ft' 2.72325
Honoollnlo ' 2.1973g
Honocllnlc 5-6252c
•*
Orthorhombio 1.H982g
. Melting Boiling
point point
2135 oa. 3000
Decomposes
to Cr20
630
197 Decomposes
-96.5 US. 8
Decomposes
398 Decomposes
81. 6 Decomposes
(incongruent)
971
792
Decomposes
Decomposes
Solubility
Insoluble
Soluble in acids to
Cr and Cr
Slightly decomposes
in H.O; soluble in
dilute acida to
^3* j r. 0 +
Cr and Cr
Very soluble in H?0;
soluble in CH
COCH. (CH2CO)|5
Insoluble in 11^6;
hydrolyzesj soluble
in CS , CC1, . '
Soluble in H,,O
Soluble In H£0
Very soluble in HgO
Soluble in H.,0
Soluble in K20
Soluble in H^O,
hydrolyzea
Very slightly soluble
In 11 0; soluble in
dilute acids
Very slightly soluble
in H-,0; soluble in
-------
Melting Boiling
Compound Formula Appearance Crystal system Density point point Solubility
and space group (g/cra ) (°C) (°C)
Strontium chroraate SrCHL Yellou solid Monoclinia, c!j 3,895.,. Decomposes Slightly soluble in
3 HO; soluble in
dilute acids
Lead chronate PbCrOjj Yellow solid Orthorhombio 5
Orange solid Monoclinio, C? 6.12 , 811 Practically insoluble
5 in ll.,0j soluble in
strong acids
Red solid Tetragonal
1
£
X
1
Source: Adapted from U.S. EPA, 1978.
-------
II. EXPOSURE
Large amounts of hexavalent chromium are produced and
utilized in industry, primarily as chromates and dichromates
{U.S. EPA, 1979). Industrial processes consumed 320,000
metric tons of chromium metal alone in 1972.
Much of the detectable chromium in air and water is pre-
sumably derived from industrial processes. Levels of total Cr
in the air exceeding 0.010 mg/m^ were reported from 59 of
186 urban areas examined (U.S. EPA, 1973). Air levels in non-
urban areas generally fall below detection limits. Mean con-
centration of Cr in 1577 samples of surface water was deter-
mined as 9.7 ug/1 (Kopp, 1969). Cr is also naturally distrib-
uted in the continental crust at an average concentration of
125 mg/kg (U.S. EPA, 1978).
Based on available monitoring data, the U.S. EPA (1979)
has estimated the uptake of Cr by adult humans from air, water
and food:
Source Uptake ug/day
Atmosphere 1
Water 50-100
Food supply 2Q
121
The amount of chromium entering the blood is a function
i
of the extent of fractional absorption occurring in the intes-
tine; this in turn is influenced by the chemical form in which
the compound is presented, the presence of other dietary con-
stituents, and poorly understood intestinal epithelial bar-.
riers (U.S. EPA, 1979).
-------
The U.S. EPA (1979) has derived a bioconcentration factor
(BCFJ of 11 for chromium.
III. PHARMACOKIN ETICS
A. Absorption
The efficiency of chromium absorption by the gastro-
intestinal tract is a function of the oxidation state of the
compound and the presence of other dietary constituents.
-Jlertz (1969) has reported 25 percent absorption of glucose
tolerance factor, a chromium complex, from the digestive
tract. In general, trivalent chromium may be expected to bind
with epithelial components, retarding absorption (U.S. EPA,
1979). Dermal absorption of chromium has not been estimated
to contribute greatly to total body load, except in situations
where toxic external concentrations have produced ulceration
(U.S. EPA, 1979). Pulmonary exposure to chromium leads to
prolonged retention at this site (Baetjer, et al. 1959); the
contribution of the inhalation route to total absorbed chro-
mium is probably not major (U.S. EPA, 1979).
B. Distribution
Distribution of administered chromium depends on its
chemical state and the amount given. Chromium has an affinity
for the reticuloendothelial system and for the spleen, liver,
and -bone marrow. This may reflect uptake of chromium by red
cells and phagocytosis of chromium containing colloidal par-
ticles (National Academy of Sciences, 1974). Chromium levels
in tissues other than the lungs decline with age (Schroeder,
et al. 1962). Chromium has been demonstrated to cross the
placental barrier in certain forms (National Academy of
Sciences, 1974).
-------
IV. EFFECTS
A. Carcinogenicity
The hexavalent chromium compounds have been grouped
by NIOSH (1975) into "carcinogenic" and "non-carcinogenic"
categories, based primarily on extensive animal studies.
"Non-carcinogenic chromium" includes the mono- and dichromates
of hydrogen, lithium, sodium, potassium, rubidium, cesium and
ammonium, and also chromic oxide. The "carcinogenic chromium"
group includes hexavalent chromium compounds not listed in the
first category. A large proportion of the tumors produced by
hexavalent chromium in animal studies were injection site
specific; the role of foreign body carcinogenesis ("Oppen-
heimer effect") should be considered in evaluating these re-
sults.
The relatively high incidence of lung cancer in
workers employed in the chromate industry has been detailed in
numerous studies (National Academy of Sciences, 1974). This
tumor type has been produced in animal studies using intra-
bronchial implantation of calcium chromate (Laskin, et al.
1970). Taylor (1966) determined an 8-fold increase in lung
tumor mortality for a large group of chromate workers relative
to the expected incidence. interpretation is complicated by
the finding that chromium carbonyl has some cocarcinogenic
effects in combination with benzofa]pyrene (Lane and Mass,
1977).
Taylor's cohort also exhibited a statistically sig-
nificant increase in digestive cancer. Although the precise
exposure level of Taylor's subjects is not known, some conser-
-US-
-------
C. Metabolism
Analysis of chromium metabolism is complicated by
the extensive binding of chromium to tissue components (en-
zymes, proteins, nucleic acids) and by the inability of analy-
tical methods to distinguish between the different forms of
chromium (U.S. EPA, 1978).
Studies of the kinetics of radiochromium distribu-
tion in humans indicated three major accumulation and clear-
ance components (Lim, 1978). Animal studies with radioactive
chromium trichloride injected intravenously showed that heart,
lung, pancreas, and brain retained 10 to 31 percent of their
initial radioactivity after four days, while spleen, kidney,
testis, and epididymis concentrated chromium (Hopkins, 1965).
D. Elimination
Chromium turnover in humans appears to be very slow
(National Academy of Sciences, 1974). A long component of
chromium elimination has been calculated with a half-life of
616 days (Taylor, 1975). In rats, triple compartment half-
lives for trivalent chromium have been estimated to be 0.5,
5.9, and 83.4 days (Hertz, et al. 1965).
Chromium is excreted in both the urine and the
feces. Urinary excretion is the major route of elimination,
accounting for recovery of 80 percent of injected chromium
(Mertz, 1969). Up to 20 percent of intravenously injected
trivalent chromium was found in the feces of rats (Visek, et
al. 1.953).
ft
-------
vative assumptions lead to a calculated level of 8 nanogram
CrVl/liter of drinking water.
B. Hutagenicity
Kexavalent chromium has been shown to produce muta-
genic effects in several systems. Chromates and dichromates
have produced mutations in E. coli (Venitt and Levy, 1974),
chromosomal aberrations in cultured fetal mouse cells {Rafetto,
et al. 1977), and cytogenetic effects in mouse (Wild, 1978)
and rat (Bigalief, et al. 1977) bone marrow cells. Trivalent
chromium compounds have not shown mutagenic effects.
Cytogenetic effects in workers exposed to welding fumes have
been attributed to chromium aerosol (Hedenstedt, et al. 1977) .
These effects have also been reported in chromate production
workers (Bigalief, et al. 1977).
Testing of hexavalent chromium in the Ames assay has
shown positive results without metabolic activation; trivalent
chromium compounds were not mutagenic (Petrilli and DeFlora,
1977).
C. Teratogenicity
Embryonic abnormalities have been produced in the
developing chicken by direct injection of trivalent or hexa-
valent chromium into the yolk sac or onto the chorioallantoic
membrane (Ridgway and Karnofsky, 1952). The specificity of
this type of aberration production is not clear, since other
metals will produce positive effects in this system.
D. Other Reproductive Effects
Pertinent information could not be located in the
available literature.
-------
E. Chronic Toxicity
Dermal effects of chromium compounds include ulcera-
tive changes and allergic contact dermatitis, generally after
exposure to high concentrations of compound (NIOSH, 1975).
In one instance a correlation has been drawn between
hepatic lesions and worker exposure to chromium. Biopsy speci-
mens showed changes although these workers did not display
clinical symptoms (U.S. EPA, 1979).
Pulmonary dynamics have been reported to change in
chrome electroplating workers (Bovett, et al. 1977). However,
exposure of these workers is to multiple chemical agents.
V. AQUATIC TOXICITY (from U.S. EPA, 1979)
A. Chronic Toxicity
No chronic data for toxicity of trivalent chromium
for freshwater fishes is available. The geometric mean of
chronic toxicity values for the freshwater invertebrate Daph-
nia magna is based on data from a single study, and is re-
ported as 445 ug/1. No chronic data for trivalent chromium
for freshwater algae are available.
Chronic embryo-larval tests on six species of fresh
water fish exposed to hexavalent resulted in chronic values
ranging from 37 to 72 ug/1 for rainbow trout (Salmo gairdneri)
and lake trout, Salvelinus mamaycush. White suckers, Catos-
totnus commersoni, and channel catfish, Ictalurus punctatus,
were intermediate in sensitivity and northern pike, Esox lu-
cius, and bluegills, Lepomis macrochirus, were least sensitive
with chronic values of 360 and 368 ug/1 respectively. In life
cycle or partial life cycle tests both the rainbow trout and
-------
snook trout, Salvelinuja fontinalis, were sensitive with chron-
ic values of. 265 ug/1. Chronic testing of hexavalent chromium
in Daphnia magna found significant survival and fecundity
changes at concentrations as low as 10 ug/1- The effects of
hexavalent chromium on the freshwater alga, Chlamydomonas
reinhardi, were recorded at levels as low as 10 ug/1. The
Eurasian watermilfoil displayed the greatest resistance to
hexavalent chromium at levels as high as 9,900 ug/1.
There are no chronic toxicity data available for
trivalent chromium compounds in marine fish or marine inver-
tebrates. Data for trivalent chromium effects to marine algae
are not available.
The only available bioconcentration data for fresh-
water is from studies on rainbow trout, and indicates a bio-
concentration factor of 1 for potassium chromate. The only
marine bioconcentration factors result from three species of
bivalve molluscs, Mytilus edulj.s, 34; Crassostrea Virginia,
166; and Mya arenaria, 152.
B. Acute Toxicity
The acute toxicity of trivalent chromium compounds
has been examined more intensely. The 96-hour LC50 values
for 14 tests ranged from 3,330 to 71,900 ug/1 and correlated
with the hardness of water over a range of 20 to 360 ug/1 as
CaCC>3 in 11 species of freshwater fish. The guppy Poecilia
reticulata was most sensitive and the bluegill the most resis-
»
tant. Among eight species of freshwater invertebrates, acute
96-hour LC50 values ranged from 2,000 to 64,000 ug/1.
Vt
-------
For hexavalent chromium 96-hour LC.50 values
ranged from 17,600 ug/1 in the fathead minnow, Pimephales pro-
melas, in soft water to 195,000 ug/1 for large mouth bass, Mi-
cropteus salmoider, in hard water. The 96-hour LC5Q values
for freshwater invertebrates exposed to hexavalent chromium
ranged from 3,100 ug/1 in the rotifer, Philodina acuticornis,
to 12,000 ug/1 in the rotifer, Philodina roseola.
There are no pertinent acute toxicity data available
for trivalent chromium compounds to marine species.
The acute toxicity data for hexavalent chromium to
marine fishes resulted in 96-hour LC5g values of 30,000 to
30,000 ug/1 for the speckled sanddab, Citharichthys stigmaeus,
and 91,000 ug/1 for the mummichog, Fundulus heteroclitus. In-
vertebrates appeared more sensitive to hexavalent chromium
than marine fish. The 96 hour LC^Q values for hexavalent
chromium ranged from 2,000 ug/1 for the polychaete worm,
Nereis vinens, to 105,000 ug/1 for the mud snail, Nassarius
obsoleutus, in static bioassays.
The U.S. EPA (1978) offers an extensive review of
the environmental effects of chromium compounds in freshwater
and marine organisms.
VI. EXISTING 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 change. •
Based on animal data indicating carcinogenic effects of
chromium VI and estimates of lifetime exposures from consump-
-------
tions of both drinking water and aquatic life forms, the U.S.
EPA (1979) has estimated levels of hexavalent chromium in
ambient water which will result in specified risk levels of
human cancer:
Exposure Assumptions (per day) Risk Levels and Corresponding Criteria
0 10~7 IP"6 10~5
2 liters of drinking water and 0 0.08 ng/1 0.8 ng/1 8 ng/1
consumption of 18.7 grams fish
and shellfish
Consumption of fish and shell 0 8.63 ng/1 86.3 ng/1 863 ng/1
fish alone
The OSHA time-weighted average exposure criterion for
chromium (carcinogenic compounds) is 1 ug/m^; for the "non-
carcinogenic" classification of chromium compounds the cri-
terion is 25 ug/3 TWA {U.S. EPA, 1979).
For the protection of aquatic species, proposed water
criteria for both trivalent and hexavalent chromium in fresh-
water and marine environments have been prepared in accordance
with the Guidelines for Deriving Water Quality Criteria
(Federal Register J_3:21506, May 18, 1975 and Federal Register
J_3:29028, July 5, 1978). In freshwater environments the pro-
posed criterion for hexavalent chromium is 10 ug/1, not to ex-
ceed 110 ug/1, and the proposed criterion for trivalent cr is
given a Chronic Final Value represented by the following
equation:
C.F.V. = e(°*83 In (water hardness) = 2.94)
The proposed criterion for trivalent chromium in marine
*
environments could not be determined by criteria established
in the Guidelines.
-------
CHROMIUM
REFERENCES
Baejter, A., et al. 1959. The distribution and retention of
chromium in men and animals. AMA Arch. Ind. Health 20: 126.
Bigalief, A., et al. 1977. Evaluation of the mutagenous
activity of chromium compounds. Gig. Tr. Prof. Zabol. 6: 37.
Bovett, P., et al. 1977. Spirometric alterations in workers
in the chromium electroplating industry. Int. Arch Occup.
Environ. Health 40: 25.
Hedenstedt, A., et al. 1977. Mutagenicity of fume particles
from stainless steel welding. Scand. J. Work. Environ. Health
3: 203.
Hopkins, L. 1965. Distribution in the rat of physiological
amounts of .injected Cr-51 (III.) with time. Am. J. Physiol.
209: 731.
Kopp, J. 1969. The occurrence of trace elements in water in:
Trace Substances in Environmental Health III, University of
Missouri, Columbia, Mo. p. 59.
Lane, B. and M. Mass. 1977. Carcinogenicity and cocarcino-
genicity of chromium carbonyl in heterotropic tracheal grafts.
Cancer Res. 37: 1476.
Laskin, S., et al. 1970. Studies in pulmonary carcino-
genesis in: Inhalation Carcinogenesis, M. Hanna, P. Nettle-
sheim, J. Gilbert (eds.). U.S. Atomic Energy Commission. p.
321.
Lim, T. 1978. The kinetics of the trace element chromium
(III) in the human body. Paper presented at 2nd International
Congress of Nuclear Medicine and Biology, Washington, D.C.
Mertz, W. 1969. Chromium occurrence and function in bio-
logical systems. Physiol. Rev. 49: 163.
Mertz, W., et al. 1965. Biological activity and fate of
trace quantities of intravenous chromium (III) in the rat.
Amer. J. Physiol. 209: 484.
National Academy of Sciences. 1974. Medical and biological
effects of environmental pollutants: Chromium. Washington,
D.C.
»
National Institute for Occupational Safety and Health. 1975.
Criteria for a recommended standard - occupational exposure to
chromium (VI). U.S.D.H.E.W. Publications #76-129.
-------
Petrilli, F. , and S. DeFlora. 1977. Toxici-ty and mutageni-
city of hexavalent chromium on Salmonella typhimurium. Appl.
Environ. Microbiol. 33: 805.
Rafetto, G., et al. 1977. Direct interaction with cellular
targets as the mechanism for chromium carcinogenesis. Tumor
63: 503.
Ridgway, L., and D. Karnofsky. 1952. Effects of metals in
the chick embryo - toxicity and production of abnormalities in
development. Ann. N.Y. Acad. Sci. 55: 203.
Schroeder, H., et al. 1962. Abnormal trace metals in man-
chronium. J. Chronic Dis. 15: 941.
Taylor, F. 1966. The relationship of mortality and duration
of employment as reflected by a cohort of chromate workers.
Am. J. Pub. Health 56: 218.
Taylor, F. 1975. Distribution and retention of chromium in
small mammals from cooling tower drift. Presented at Fourth
National Symposium on Radioecology, Corvallis, Ore. May 12-14,
1975.
Trama, F., and R. Benoit. 1960. Toxicity of hexavalent chro-
mium to bluegills. J. Water Pollut. Control Fed. 37: 868.
U.S. EPA. 1973. Air quality data for metals - 1968 and 1969.
EPA document #APTD 1467.
U.S. EPA. 1978. Reviews of the environmental effects of pol-
lutants: Chromium. EPA document #600/1-78-023.
U.S. EPA. 1979. Chromium: Ambient Water Quality Criteria.
Venitt, S., and L. Levy. 1974. Mutagenicity of chromates in
bacteria and its relevance to chromate carcinogenesis. Nature
250: 493.
Visek, W., et al. 1953. Metabolism of Cr-51 by animals as
influenced by chemical state. Proc. Soc. Exp. Biol. Med. 84:
610.
Weast, R. 1974. Handbook of Chemistry and Physics, 55th ed.,
CRC Press, Cleveland, Ohio p. 2216.
Wild, .D. 1978. Cytogenetic effect in the mouse of 17 chemi-
cal mutagens and carcinogens evaluated by the micronucleus
test.
-------
No. 52
Chrysetie
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRI!, 30, 1980
-(02<0~
-------
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
chrysene and has found sufficient evidence to indicate that
this compound is carcinogenic.
-------
CHRY5ENE
Summary
Chrysene is a member of the polynuclear aromatic hydrocarbons (PAH)
class. Numerous compounds in the PAH class are well-known as potent animal
carcinogens. However, chrysene is generally regarded as only a weak carcin-
ogen to animals. There are no reports available concerning the chronic
toxicity of chrysene. Although exposure to chrysene in the environment oc-
curs in conjunction with exposure to other PAH, it is not known how these
compounds may interact in human systems.
No standard toxicity data for chrysene are available for freshwater or
marine organisms.
-------
I. INTRODUCTION
This profile is based on 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).
Chrysene ^cinH12^ ^s one °^ tne ?am^y °? polynuclear aromatic
hydrocarbons (PAH) formed as a result of incomplete combustion of organic
material. Its physical/chemical properties have not been well-character-
ized, other than a reported melting point of 254°C and a boiling point of
448°C (U.S. EPA, 1979b).
PAH, including chrysene, are ubiquitous in the environment, being found
in ambient air, food, water, soils, and sediment (U.S. EPA, 1979b). The PAH
class contains a number of potent carcinogens (e.g., benzo(a)pyrene), moder-
ately active carcinogens (e.g., benzo(b)fluoranthene), weak carcinogens
(e.g., chrysene), and cocarcinogens (e.g., fluoranthene), as well as numer-
ous non-carcinogens (U.S. EPA, 1979b).
PAH which contain more than three rings (such as chrysene) are rela-
tively stable in the environment. and may be transported in air and water by
adsorption to participate 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 more general dis-
cussion of PAH (U.S. EPA, 1979c).
II. EXPOSURE
A. Water
Levels of chrysene are not routinely monitored in water. However,
the concentration of six representative PAH (benzo(a)pyrene, fluoranthene,
benzo(k)fluoranthene, benzo(j)fluoranthene, benzo(g,h,i)perylene, and 'inde-
no(l,3-cd)pyrene) in U.S. drinking water averaged 13.5 ng/1 (Basu and Sax-
ena, 1977,1978).
-------
8. Food
Chrysene has been detected In a wide variety of foods such as coco-
nut oil (12 ppb), and smoked or cooked meats (up to 66 ppb) (U.S. EPA,
1979b). Although it is not possible to accurately estimate the human diet-
ary intake of chrysene, it has been concluded (U.S. EPA, 1979b) that the
daily dietary intake for all types of PAH is in the range of 1.6 to 16 ug.
The U.S. EPA (1979a) has estimated the weighted average bioconcentration
factor for chrysene to be 3,100 for the edible portion of fish and shellfish
consumed by Americans. This estimate is based on the octanol/water parti-
tion coefficient for chrysene.
C. Inhalation
Chrysene is commonly found in ambient air. Measured concentrations
of chrysene have reportedly been in the range of 0.6 to 4.8 ng/m (Gordon,
1976; Fox and Staley, 1976). Thus, the human daily intake of chrysene by
inhalation of ambient air may be in the range of 11.4 to 91.2 ng, assuming
that a human breathes 19 m of air per day.
III. PHARMACOKINETICS
Pertinent data could not be located in the available literature con-
cerning the pharmacokinetics of chrysene or other PAH in humans. Neverthe-
less, it is possible to make limited assumptions based on the results of
animal research conducted with several PAH, particularly benzo(a)pyrene.
A. Absorption
The absorption of chrysene in humans has not been studied. How-
ever, it' is known (U.S. EPA, 1979a) that, as a class, PAH are well-absorbed
across the respiratory and gastrointestinal epithelia. In particular, chry-
r
sene was reported to be readily transported across the gastrointestinal
mucosa (Rees, et al. 1971). The high lipid solubility of compounds in the
PAH class supports this observation.
-------
8. Distribution
The distribution of chrysene in mammals has not been studied. How-
ever, 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 tis-
sues (e.g., breast).
C. Metabolism
Limited work on the metabolism of chrysene has been conducted, as
part of an investigation into the mechanism of its bioactivation to a muta-
gen/carcinogen (Wood, et al. 1977).
Chrysene, like other PAH, is apparently metabolized by the microso-
mal mixed-function oxidase enzyme system in mammals (U.S. EPA, 1979b).
Metabolic attack on one or more of the aromatic double bonds leads to the
formation of phenols, and isomeric dihydrodiols by the intermediate forma-
tion of reactive epoxides. Dihydrodiols are further metabolized by microso-
mal 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 chrysene by mammals has not been studied. How-
ever, 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 sig-
nificant percentage of administered PAH. However, the rate of disappearance
•
of various PAH from the body, and the principal routes of excretion, are in-
-------
fluenced both by the structure of the parent compound and the route of admi-
nistration (U.S. EPA, 1979b). It is unlikely that PAH will accumulate in
the body with chronic low-level exposures.
IV. EFFECTS
A. Carcinogenicity
Chrysene is regarded as a weak animal carcinogen (U.S. EPA, 1979b).
LaVoie and coworkers (1979) reported that chrysene can act as both a tumor
initiator and as a complete carcinogen on the skin of mice.
B. Mutagenicity
Chrysene is positive in the Ames Salmonella assay in the presence
of a metabolizing enzyme system (LaVoie, et al. 1979; Wood, et al. 1977).
Chrysene is also positive in the induction of sister-chromatid exchanges in
Chinese hamster cells (Roszinsky-Kocher, et al. 1979).
C. Teratogenicity
Pertinent data could not be located in the available literature
concerning the possible teratogenicity of chrysene. Other related PAH are
apparently not significantly teratogenic in mammals (U.S. EPA, 1979a).
D. Other Reproductive Effects and Chronic Toxicity
Pertinent data could not be located in the available literature re-
garding other reproductive effects and chronic toxicity.
V. AQUATIC TOXICITY
The only data concerning the effects of chrysene to aquatic organisms
is a single bioconcentration factor of 8.2 (24-hour) for the marine clam
(Rangia cuneata) (Neff, et al. 1976). No standard aquatic toxicity data for
chrysene either in acute or chronic studies are available for freshwater or
*
marine species.
-------
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
There are no established exposure criteria for chrysene. However,
PAH as a class are regulated by several authorities. The World Health Or-
ganization has recommended that the concentration of PAH in drinking water
(measured as the total of fluoranthene, benzo(g,h,i)perylene, benzo(b)fluor-
anthene, benzo(h)fluoranthene, indeno(l,2,3-cd)pyrene, and benzo(a)pyrene)
not exceed 0.2 jug/1. Occupational exposure criteria have been established
for coke oven emissions, coal tar products, and coal tar pitch volatiles,
all of which contain large amounts of PAH including chrysene (U.S. EPA,
1979a).
The U.S. EPA (1979a) draft recommended criteria for PAH in water
are based upon the extrapolation of animal carcinogenicity data 'for benzo-
(a)pyrene and dibenz(a,h)anthracene.
B. Aquatic
Data is insufficient for drafting freshwater or marine criterion.
-------
CHRYSENE
REFERENCES
Basu, D.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, Expo. Evalu. Branch, HERL, Cincin-
nati.
Basu, O.K., and J. Saxena. 1978. Polynuclear aromatic hydrocar-
bons in selected U.S. drinking waters and their raw water sources.
Environ. Sci. Technol. 12: 795.
Fox, M.A., and S.W. Staley. 1976. Determination of polycyclic
aromatic hydrocarbons in atmosphere particulate matter by high
pressure liquid chromatography coupled with flourescence tech-
niques. Anal. Chem. 48: 992.
Gordon, R.J., 1976. Distribution of airborne polycyclic aromatic
hydrocarbons throughout Los Angeles. Environ. Sci. Technol.
10: 370.
Lasnitzki, A., and Woodhouse, D.C. 1944. The effect of 1:2:5:6-
Dibenzanthracene on the lymph-nodes of the rat. Jour. Anat.
78: 121.
LaVoie, E., et al. 1979. A comparison of the mutagenicity tumor-
initiating activity and complete carcinogenicity of polychlorin-
ated aromatic hydrocarbons In: Polynuclear Aromatic Hydrocarbons,
P.W. Jones and P. Leber (eds~.). Ann Arbor Science Publishers.
Neff, J.M., et al. 1976. Accumulation and release of petroleum-
derived aromatic hydrocarbons by four species of marine animals.
Mar. Biol. 38: 279.
Rees, E. 0., et al. 1971. A study of the mechanism of intestinal
absorption of benzo(a)pyrene. Biochem. Biophys. Act. 225: 96.
Roszinsky - Kocher, et al. 1979. Mutagenicity of PAH's. Induc-
tion of cister-chromatied exchanges ir\ vivo. Mutation Research.
66: 65.
U.S. EPA. 1979a. Polynuclear Aromatic Hydrocarbons: Ambient
Water Quality Criteria (Draft).
U.S. EPA. 1979b. Multimedia Health Assessment Document for Poly-
cyclic Organic Matter. Environmental Criteria and Assessment
Office, Research Triangle Park, N.C. Prepared by Syracuse Research
Corporation.
U.S. EPA. 1979c. Environmental Criteria and Assessment Office.
Hazard Profile: Chlorinated Ethanes (Draft).
-------
Wood A.W., et al. 1977. Metabolic Activation of Libenzo(ah)an-
thracene and its Dihydrodiols to Bacterial Mutagens. Cancer Res.
38: 1967.
World Health Organization. 1970. European standards for drink-
ing waters. 2nd edition. Revised. Geneva.
-------
No. 53
Creosote
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.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated
creosote and has found sufficient evidence to indicate that
this compound is carcinogenic.
-------
CREOSOTE
I. INTRODUCTION
Creosote is a coal-tar distillate used mainly as a wood preservative.
It is highly toxic to wood-destroying organisms and has a low evaporation
rate (Farm Chemicals Handbook, 1977). In 1972, an estimated 521,000 tonnes
(575,000 tons) were produced by six companies at 25 sites in the United
States (von Rumker, et al. 1974). About 90 percent of the creosote is sold
to the wood-preservation industry; the remainder is burned as fuel (von Rum-
ker, et al. 1974).
Creosote's other pesticidal uses are as an herbicide, an insecticide,
an acaricide," an arachnicide, a fungicfde, a tree dressing, a disinfectant,
and a horse repellent (Table 1).
TABLE 1.
USES AND SITES FOR CREOSOTE
(Cummings, 1977)
Use
Preservative
Insecticide
(screwworm)
Acaricide (mites)
Arachnicide (ticks)
Herbicide
Fungicide
Insecticide
(Certain insects, worms,
moths and borers)
Horse repellent
Disinfectant
Site
Wood
Horses and mules
Poultry and horses
Poultry and horses
Along roads, highways, and fences;
farms; flower beds
Rope, canvas, tarpaulins, tree wounds
Tree dressing
Wood stalls, mangers, gates, fence
rails, posts, trees, trailer sites
Outhouses, water closets, garbage
cans, feeding and watering equipment
-------
Creosote is produced by the distillation of coal tar obtained from the
coking of coal. The composition of creosote is highly variable and depends
on the composition of the coal used to make the tar, the design and operat-
ing conditions of the coke oven (e.g., gas collection system, temperature,
coking time), and the design and operating condition of the still (e.g.,
feed rate, temperature, and blending of tar distillation fractions) (43 PR
48154, 1978).
Continuous tar distillation at temperatures of up to 400°C produces
fractions typically ranging from crude benzene to residue pitches (von Rum-
ker, et al. 1974). A common distillation temperature for creosote is about
200 to 400°C (Hawley, 1977; von Rumker, et al. 1974). The creosote frac-
tion is a mixture of organic compounds, mainly liquid and solid cyclic hy-'
drocarbons, including two-ring and polynuclear aromatic hydrocarbons (PAH)
•(Table 2). Among the PAH, phenanthrene represents 12 to 14 percent of the
composition of creosote (Considine, 1976). Benzo(a)pyrene (BaP) is present
at a concentration of about 200 ppm (Guerin, 1977).
II. EXPOSURE
A. Water
Each year an estimated 60 to 115 million pounds (27,000-52,000
tonnes) of creosote are discharged in wastewater treatment sludges by creo-
sote producers. At large tar distillation plants, wastewater streams con-
taining creosote are treated on-site and/or conveyed to public sewage treat-
ment facilities. Wastewater sludges treated on-site are transferred to
landfill or burial sites (von Rumker, et al. 1974),. The estimated flux of
2
creosote from these disposal sites ranges from 0.75 kg/m /hr to 11.0
kg/m2/yr (U.S. EPA, 1980). In 1972, about one billion pounds ('455,000
-------
TABLE 2.
PHYSICAL AND CHEMICAL PROPERTIES OF CREOSOTE
Synonyms: Brick oil, coal tar oil, creosote oil, creosotum, cresylic creo-
sote, dead oil, heavy oil, liquid pitch oil, naphthalene oil, tar
oil, wash oil
Structural and Empirical Formula: Consists principally of liquid and solid
cyclic hydrocarbons; contains substantial amounts of naphthalene
and anthracene; 12-14 percent phenanthrene; 200 ppm benz(a)pyrene
Molecular Weight: —
Description: Dark brown green, yellowish or colorless above 38°C, naph-
thenic odor; soluble in alcohol, benzene toluene; immiscible
• with water
Specific Gravity and/or Density: d^5 more than 1.076
25
Melting and/or Boiling Points: Common distillation range 200 to 400°C
Stability: Overall degradation rate (0.48/day) = same as microbial degra-
dation
Solubility (water): approx. 5 g/1; sed . . 2
H^O . 1
Vapor Pressure: —
Bioconcentration Factor (BCF) and/or
Octanol/water partition coefficient (KQW): BCF =0.6
Kow = 1.0
Source: Hawley, G.G., 1977; Windholz, 1976; U.S. EPA, 1980; Lopedes, 1978
-------
tonnes) of creosote were used to preserve railroad ties, marine pilings and
utility poles (NIOSH, 1977a).
Some of the organics present in creosote are moderately soluble.
Creosote partitions between sediments and water in a ratio of 1:5. It is
considered stable in groundwater, but decomposes at an estimated rate of 90
percent in five days in river water flowing 50-250 miles. About 99 percent
decomposed in a lake environment in one year (U.S. EPA, 1980).
Creosote migrates from treated wood into the environment, but the
impact of this migration is unknown. Creosote-was found to have a vapor
loss of 27.5 and 15.2 percent from the outer two inches of seasoned and
green poles,- respectively; high residue creosote was estimated to have a
10.3 and 4.4 vapor loss, respectively. Creosote losses to the aquatic envi-
ronment are the greatest during the first years after installation. One
eight-year study is summarized below (43 FR 48154, 1978).
Creosote Loss
Year pounds/linear foot
1 0.31
2 0.05
3 0.06
4 0.22
4-8 0.15 (average)
B. Food
Naiussat and Auger (1970) found that PAHs in a contaminated lagoon
accumulated to the greatest extent in species near the top of the food
chain. One of these compounds, BaP, has been reported to accumulate in mus-
sels (about 50 pg/kg; 20 times background) taken from creosote-treated pil-
ings (43 FR 48154, 1978). Elevated levels of BaP in mussels growing near
creosoted timbers or pilings suggest that creosote is a significant* source
of BaP in the marine environment. This suggestion was supported by compari-
-------
sons of gas chromatography profiles of polycyclic aromatic hydrocarbons iso-
lated from mussels and creosoted wood (Dunn and Stich, 1976).
High levels of PAH have been found in commercial seafoods grown in
impoundments constructed of creosoted wood. Commercial samples of oysters,
clams, and mussels were found to contain BaP at concentrations generally
less than 10 ng/g (wet weight). PAHs were also found in cockels, abalone,
scallops, lobster, and shrimp. Levels of BaP and other related PAHs were
found to be inversely related to the ability of the species to metabolize
PAH, except in the case of lobster. Unexpectedly high levels were found in
all edible meat of lobsters maintained in commercial tidal compounds con-
structed of creosoted timber: up to 281 ng/g BaP, 303 ng/g chrysene, 222
ng/g benzo(a)anthracene, 261 ng/g benzo(b)fluoranthene, 153 ng/g dibenz-
(a,h)anthracene, and 137 ng/g indeno(l,2,3-cd)pyrene (Dunn and Fee, 1979).
III. PHARMACOKINETICS
A. Absorption
Creosote is (readily) absorbed through the skin and mucous mem-
branes (NIQSH, 1977b).
IV. EFFECTS
A. Carcinogenicity
Creosote has been associated with several occupational cases of
skin cancer over a 50-year period (Farm Chemicals Handbook, 1977); its role
in human cancer is still not clearly understood (NIOSH, 1977b).
Henry (1947), Lenson (1956), 0'Donovan (1920), Cookson (1924), and
Mackenzie (1898) described various kinds of workers who were occupationally
exposed to creosote and developed skin tumors. Dermal application of creo-
sote produced skin tumors in mice (Woodhouse, 1950; Poel and Kammer, 1957;
Lijinsky, et al. 1956; Boutwell and Bosch, 1958; Roe, et al. 1958). Roe, et
-O.HH-
-------
al. (1958) also found that dermal application of creosote to mice produced
lung tumors. Soutwell and Bosch (1958) found that creosote had the ability
to initiate tumor formation when applied for a limited period prior to
treatment with croton oil. Sail and Shear (1940) found that the number of
skin tumors was increased by dermal treatment with creosote and benzo(a)py-
rene over the number of tumors produced by benzo(a)pyrene or creosote alone.
There is considerable evidence to show that creosote produces tumors in
mice; that creosote, when applied dermally, is a tumor-initiating agent when
followed by dermal treatment with croton oil (Boutwell and Bosch, 1958);
that creosote accelerates the tumor production caused by benzo(a)pyrene
(Sail and Shear,- 1940); and that workers occupationally exposed to creosote
developed tumors (Table 3). These studies have not yet demonstrated a cor-
relation between the carcinogenic potency of, creosote oils and the content
of benzpyrene (Patty, 1963).
Results from dose response studies are summarized below (NIOSH,
1977a).
Concentration
and duration Effects
100% 3x/wk Skin carcinomas in 82%,
28 wk tumors in 92%
20-80% 3x/wk Skin carcinomas in 88%,
6-44 wk tumors in 100%
100% 2x/wk Skin and lung tumors
21 wk in 74%
100% 3x/wk Skin tumors in 50%
70 wk
10-100% 2x/wk* Skin tumors''in 38-74%
70 wk
2% 2x/wk* No tumors
70 wk
*Creosote plus 1 percent 7,12-dimethylbenz(a)anthracene.
-------
TABLE 3.
SUMMARY TABLE ON ONCOGENICITY OF CREOSOTE
A. Human Case Reports
Substance
and Type
Authors Year of Exposure
Occupation
of Exposed
Individual(s)
Type of Tumor
Response
Mackenzie 1896 Handling of
Creosote
O1 Donovan 1920 Handling of
Creosote
Cookson 1524 Handling of
Creosote
Henry 1947 Handling of
Creosote
Lenson 1956 Painting of
Creosote
Worker who dipped Warty elevation on arms;
railway ties in papillomatous swellings
creosote on scrotum
Workers who creo- Skin cancer
soted timbers
Creosote factory
worker
Squamous epitheliomata
on hand; epitheliomatous
deposits in liver, lungs,
kidneys and heart walls
37 men of various Cutaneous epitheliomata
occupations
Shipyard worker
Malignant cutaneous
tumors of the face
B. Animal Studies
Dermal Exposure
Authors
Sail and
Shear
Year
1940
Substance
Tested
Creosote and
benzo(a)pyrene
Animal and
Strain
Mice (Strain A)
Type of Tumor
Response
Accelerated tumor forma-
tion
Woodhouse 1950 Creosote oil
Lijinsky,
et al.
Poel and
Kanmer
Boutwe11
and Bosch
Roe,
et al.
1956 #1 creosote
oil
1557 Blended creo-
sote oils;
Mice (Albino;
Undefined strain)
Mice - Swiss
Mice (C57L
Strain)
Light creosote Mice (C57L
oil Strain)
1958 Creosote
(Carbasota)
1958 Creosote oil
(Carbasota)
Mice (Albino
random bred)
Mice (Strain
Undefined)
Papillomas and carcinomas
Papillamas and carcinomas
Papillomas and carcinomas
metastatic growths in
lungs and lymph nodes
Papillomas
*
Papillomas and carcinomas
Skin and lung tumors
-------
B. Mutagenicity
Simmon and Poole (1978) found that, following metabolic activation
by Arochlor 1254-stimulated rat liver homogenate, both the creosote PI and
the coal tar-creosote P2 produced a mutagenic dose-response and a doubling
above background mutation rate with Salmonella typhimurium strains TA 1537,
TA 98, and TA 100. Mitchell and Tajiri (1978) found that, following meta-
bolic activation by Arochlor 1254-stimulated rat liver homogenate, creosote
PI and coal tar creosote P2 increased the number of forward mutations at the
thymidine kinase locus of L5178Y mouse lymphoma cells in a dose-related man-
ner. There is considerable evidence which proves that creosote PI and P2
cause mutations in Salmonella typhimurium strains TA 1537, TA 98 and TA 100,
and in L5178Y mouse lymphoma cells.
C. Teratogenicity and Other Reproductive Effects
Investigations utilizing pregnant swine indicate that direct con-
tact with lumber freshly treated with creosote would produce acute toxico-
sis, resulting in extensive mortality in newborn swine. The direct contact
of the pregnant sow with lumber freshly treated with creosote provides suf-
ficient dermal absorption to cause fetal deaths and weak pigs at birth.
Creosote is extremely toxic to young swine; the degree of toxicity lessens
as the pigs become older (Schipper, 1961).
D. Chronic and Acute Toxicity
Skin contact with creosote or exposure to its vapors may cuase
burning, itching, papular and vasicular eruptions, or gangrene. Eye injur-
ies can include keratitis, conjunctivitis, and -corneal abrasion (Patty,
1963). Exposed skin shows increased susceptibility to sunburn, an effect
•
attributed to photo-toxic substances usually present in commercial grades of
creosote. Eventually, exposed skin areas become hyperpigmented (NIOSH,
1977b).
-------
Serious systemic effects, including cardiovascular collapse and
death, have been observed only after ingestion (NIOSH, 1977b). Fatalities
have occurred within 14 to 36 hours after ingestion of 7 grams by adults or
1 to 2 grams by children. Symptoms of systemic illness include salivation,
vomiting, respiratory difficulties, vertigo, hypothermia, cyanosis, and mild
convulsion (Patty, 1963). Once widely used in medicine, occasional in-
stances of self-medication are still reported and sometimes lead to chronic
visual disturbances, hypertension, and gastrointestinal bleeding (NIOSH,
1977b).
The oral LD^g in rats is estimated at 725 mg creosote per kilo-
gram body weight (mg/kg). The reported LD, for dogs, cats, and rabbits
is 600 mg/kg (Fairchild, 1977).
V. AQUATIC TOXICITY
Ellis (1943) found fish kills occurring at creosote concentrations as
low as 6.0 mg/1 in less than 10 hours. Applegate, et al. (1957), using
small numbers of subjects, found that concentrations of 5.0 mg/1 produced no
mortalities in rainbow trout (Salmo gainneri), bluegill (Lepomi s macrg-
chirus), or lamprey larvae (Petromyzon marinus).
The 8-day LD~n of a 60:40 mixture of creosote and coal tar in bob-
white quail (Colinus virginianus) was reported to be about 1,260 ppm; in the
mallard duck (An a s pla ty rhynchos), 10,388 ppm. The 24-hour 50 percent medi-
um tolerance limit (TL^) of the creosote/coal tar mixture was 3.72 ppm in
rainbow trout (Salmo gainneri) and 4.42 ppm in the bluegill (Lepomis macro-
chirus). The 24-hour Tl_5Q concentrations in goldfish (Carrasius auratus)
and rainbow trout were 3.51 and 2.6 ppm, respectively (Webb, 1975).
-------
VI. EXISTING GUIDELINES AND STANDARDS
The Office of Toxic Substances of EPA has issued RPAR on creosote and
is continuing preregulatory assessment under Section 6 of the Federal Insec-
ticide, Fungicide and Rodenticide Act.
A time-weighted average creosote concentration of 0.1 mg/m^ has been
recommended for occupational air exposure.
The aquatic toxicity rating for creosote is reported as Tl_mgg = io-i
ppm (Fairchild, 1577).
-------
REFERENCES
Applegate, V.C., et al. 1957. Toxicity of 4,346 chemicals to larval lam-
preys and fishes. Dept. of Interior, Special Sci. Rept. No. 207.
Boutwell, R.K. and O.K. Bosch. 1958. The carcinogenicity of creosote oil:
its role in the induction of skin tumors in mice. Cancer Res. 18: 1171.
Considine, D.M. (ed.) 1976. Van Nostrand's Scientific Encyclopedia, 5th
ed. Van Nostrand Reinhold Co., New York.
Cookson, H.A. 1924. Epithelioma of the skin after prolonged exposure to
creosote. Brit. Med. Jour. 68: 368. •
Cummings, W. 1977. Use of profile for coal tar derivatives (exclusive of
wood preservatives). Mentioned in 43 FR 48211, 1978.
Dunn, B.P. and J. Fee. 1979. Polycyclic aromatic hydrocarbon carcinogens
in commercial seafoods. Jour. Fish Res. Board Can. 36: 1469.
Dunn, B.P. and H.F. Stich. 1976. Monitoring procedures for chemical car-
cinogens in coastal waters. Jour. Fish Res. Board Can. 33: 2040.
Ellis, M.M. 1943. Stream pollution studies in the State of Mississippi.
U.S. Dept. of Interior, Special Sci. Rept. No. 3.
Fairchild, E.J. 1977. Agricultural chemicals and pesticides: A subfile of
the NIOSH registry of toxic effects of chemical substances. U.S. Dept. of
HEW, July.
Farm Chemicals Handbook. 1977. Meister Publishing Company, Willoughby,
Ohio.
Guerin, M.R. 1977. Energy sources of polycyclic aromatic hydrocarbons.
Oak Ridge National Laboratory.
Haw ley, G.G. 1977. The Condensed Chemical Dictionary, 9th ed. Van Nos-
trand Reinhold Co., New York.
Henry, S.A. 1947. Occupational cutaneous cancer attributable to certain
chemicals in industry. Brit. Med. Bull. 4: 398.
Lenson, N. 1956. Multiple cutaneous carcinoma after creosote exposure.
New Engl. Jour. Med. 254: 520.
Lijinsky, W., et al. 1956. A study of the chemical-constitution and car-
cinogenic action of creosote oil. Jour. Natl. Cancer Inst. 18: 687.
Lopedes, D.N. (ed.) 1978. Dictionary of Scientific and Technical Tesms,
2nd ed.
Mackenzie, S. 1898. Yellow pigmentary strains of haemorrhagic origin and a
class of tar eruption. Brit. Jour. Derm. 10: 417.
-------
Mitchell, A.O. and O.T. Tajiri. 1978. In vitro mammalian mutagenicity as-
says of creosote PI and P2. SRI International. EPA Contract No. 68-01-2458.
Naiussat, P. and C. Auger. 1970. Distribution of benzo(a)pyrene and pery-
lene in various organisms of the Cliperton Lagon ecosystem. C.R. Acad.
Siv., Ser. 0. 270: 2702.
National Institute for Occupational Safety and Health. 1977a. Criteria for
a Recommended'Standard: Occupational exposure to coal tar products. DHEW
(NIOSH) Publ. No. 78-107.
National Institute for Occupational Safety and Health. 1977b. Health Haz-
ard Evaluation Determination. DHEW (NIOSH) Publ. No. 75-117-372.
0'Donovan, W.J. 1920. epitheliomatous ulceration among tar workers. Brit.
Jour. Derm. Syphilis. 32: 215.
Patty, F.A. 1963. Industrial Hygiene and Toxicology, Vol. 2, 2nd ed.
Interscience, New York.
Poel, W.E. and A.G. Kammer. 1957. Experimental carcinogenicity of coal-tar
fractions. The carcinogenicity of creosote oils. Jour. Natl. Cancer Inst.
18: 41.
Roe, F.J.C., et al. 1958. The carcinogenicity of creosote oil. The induc-
tion of lung tumors in mice. Cancer Res. 18: 1176.
Sail, R.D. and M.J. Shear. 1940. Studies in carcinogenesis. XII. Effect
of the basic fraction of creosote oil on the production of tumors in mice by
chemical carcinogens. Jour. Natl. Cancer Inst. 1: 45.
Schipper, I.S. 1961. The toxicity of wood preservatives for swine. Am.
Jour. Vet. Res. 22: 401.
Simmon, V.F. and D.C. Poole. 1978. In vitro microbiological mutagenicity
assay of creosote PI and creosote P2. SRI International. EPA Contract No.
68-10-2458.
U.S. EPA. 1980. Aquatic fate and transport estimates for hazardous chemi-
cal exposure assessments. Environmental Research Laboratory, Athens, Geor-
gia.
von Rumker, R., et al. 1974. Production, distribution, use and environ-
mental impact potential of selected pesticides. Report No. EPA 540/1-74-
001. U.S. Environ. Prot. Agency, Office of Water and Hazardous Materials,
Office of Pesticide Programs.
Webb, D.A. 1975. Environmental aspects of creosote. Proceedings American
Wood-Preservation Association. 7: 176. (
Windholz, M. 1976. The Merck Index, 9th ed. Merck and Co., Inc., Rahway,
New Jersey.
-------
Woodhouse, O.L. 1950. The carcinogenic activity of some petroleum frac-
tions and extracts; comparative results in tests on mice repeated after an
interval of eighteen months. Jour. Hygiene. 48: 121.
-------
No. 54
Cresols and Cresylic Acid
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.
-------
SJ-27-12
Cresol and Cresylic Acid
I. INTRODUCTION
Cresols are methyl phenols with methyl group at the
o-, p-, m- position. It has a molecular weight of 108, a
melting point of between 11-35'C and a boiling point of
between 191-203°C. It is slightly soluble in water, but
soluble in alcohols, glycols, dilute alkalis, ether and
chloroform. Cresylic acid is the refined product from coal
tar and contains the three isomers of cresol (the crude
product from coal tar is creosote).
Cresols are quite stable in soil due to their antimicro-
bial properties. o-cresol is degraded in air to quinones and
dihydroxybenzene by 03 with an estimated half-life of 1 day.
The m- and p- isomers are expected to behave similarly.
Cresols are used as disinfectants, agricultural chemicals,
solvents, chemical intermediates, metal cleaners, and motor
oil additives. p-cresol is permitted in U.S. as a food
flavoring and for fragrance in soaps, lotions and perfumes.
Annual production if 150 million pounds. NIOSH1 estimates
that the annual environmental release of the mixed isomers is
30 million pounds.
II. PHARMAC0 KIN E TIC S
»•
Cresols are rapidly metabolized and thus unlikely to bio-
accumulate in mammals.2
•
III. EFFECTS ON MAMMALS
A. Carcinogenicity: CAG2 concluded that the data base
-------
for this chemical is weak. No data exist on which to determine
carcinogenesis in mice. The literature cites three case
reports of cancer in humans occupationaly exposed.
B. Mutagenicity: CAG^ concluded that cresols cause
chromosome fragmentation in plants. No other mutagenicity
studies have been done.^
C. Toxicity: They are corrosive to the skin and
mucuous membranes and moderately toxic by ingestion and dermal
exposure. The organs affected are CNS, liver, lung, kidneys,
stomach, eyes, and heart. No epidemiological studies of
workers have been done.-'-
IV. EXISTING GUIDELINES
The current occupational standard (TWA) is 5 ppm. NIOSH1
recommends a lowering to 2.3 ppm.
-------
DOSSIER
ON
CSESOLS"
BY
Clement Associates, Inc.
1055 Tnomas Jefferson Street, NW
Washington,- D.C. 20007
December, 1977
Contract No. NSF-C-ENV77-15417
Prepared for
TSCA Interagency Testing Committee
Washington, D.C.
-------
FOREWORD
This dossier has been prepared bv Clement Associates, Inc.
(Clement), in partial fulfillment of Contract NSF-C-ENV77-15417,
sponsored by the National Science Foundation, to provide techni-
cal support to the Toxic Substances Control Act (TSCA) Interagency'
Testing Committee. The Committee is charged with the responsibility
for making recommendations to the Administrator of the Environmental
Protection Agency (EPA) regarding chemical substances or mixtures
which should be given"priority by EPA for testing to determine ad-
verse effects on man or the environment.
The dossier was designed to provide the Committee with infor-
mation on the chemical's physical.and chemical properties, exposure
characteristics, and biological properties in sufficient detail to
support an, .informed'judgment on whether the substance should be given
priority for testing.. The dossier is not intended to represent a com-
prehensive critical review. 'Such a.review could not be performed with
the constraints imposed upon the Committee (and, therefore, the con-
tractor) by the statutory deadlines of TSCA.
Faced with the task of preparing dossiers which could be quick-
ly assembled and yet contain sufficient information for the Commit-
tee's purposes, Clement proceeded along the following lines.
Literature searches were conducted using the National Library
of Medicine's TOXLINE and the Environmental Mutagen Information
Center (EMIC) automated data banks. Each reference on a list of
sources of general information (see "General References" in biblio-
graphy) was reviewed. Further references and information were
obtained from monographs, criteria documents, reviews, and reports
available from government agency files and trade association li-?
braries. Information received in response to the Committee's July
1977 Federal Register notice requesting information on certain
substances was reviewed. Clement scientists relied upon'their own
knowledge of the literature to augment the data sources.
In general, secondary sources were relied upon in preparing
the dossiers. When an article was judged to contain information
of major significance or to require a critical.*review the primary
source was consulted. The text makes clear whether a primary or
secondary source of information was used.
-------
KEY TO ABBREVIATIONS
TCLo - Lowest published toxic concentration
- the concentration of a substance in air which has
been reported to produce any toxic effect in animals
or humans over any given exposure time.
TDLo - Lowest published toxic dose
- the lowest dose of a substance introduced by any
route othet"than inhalation over any given period
of time that has been reported to produce any toxic
effect in animals or humans.
LCLo - Lowest published lethal concentration-
- the lowest concentration of a substance, other than
an LC50, in air that has been reported to have
caused, death in humans"or animals over any given
exposure time.
LDLo - Lowest published lethal dose
- the lowest dose of a substance other than LD50
introduced by any route other than inhalation over
any given period of time that has been reported to
have caused death in humans or animals.
LC50 - Median lethal concentration
- the concentration of a test material that kills 50
per cent of an experimental animal population
within a given time period.
LD50 - Median lethal dose
- the dose of a test material, introduced by any route
other than inhalation, that kills 50 percent of an
experimental animal population within a- given time
period.
LT50 - Median Lethal Response Time
-Statistical estimate of the time from dosage to the
death of 50 percent of the organisms in the population
subjected to a toxicant under specified conditions.
TLni - Median tolerance limit
- the concentration of a test material at which 50 per
cent of an experimental animal population are able
to survive for a specified time period.
TLV®- Threshold limit value
- the airborne concentration of a substance to which
nearly all workers nay be repeatedly exposed day
after day without adverse effect.
-------
TLV-TWA - Threshold limit value - tine, weighted average
- the time-weighted average concentration of a
substance for an 8-hour workday or 40-hour
workweek, to which nearly all workers may be
repeatedly exposed, day after day, without
adverse effect.
TLv^-STEL- Threshold limit value - short term exposure limit-
- the.maximal concentration of a substance to which
workers can be exposed for up to 15 minutes
without suffering acute or chronic toxic effects.
. No more than four excursions per day are per- •
mitted. There must be at least 60 minutes
between•exposure periods. The daily TLV-TWA
must not be exceeded.
i- - •
BOD - Biochemical oxygen demand
- a measure of the presence of organic materials
which will be oxidized biologically in bodies
of water.
NOHS Occupational Exposure:
- Rank
- an ordering of the approximately 7000 hazards
occurring in the workplace from most common to
least common .
- Estimated number of persons exposed
- includes full- and part—time workers.. For hazards
ranked 1 through 200, the figure projected to
national statistics by NIOSH is given; for the re-
maining hazards the number of people exposed given
in the survey was multiplied'by a-fixed-number to
give a rough estimate of national exposure. The
fixed number used, —30—, is derived from the sta-
tistical sampling technique used in this survey.
i - insoluble
ss — slightly soluble -*
s - soluble
vs - very soluble
- soluble in all proportions
bz - benzene
chl - chloroform
-------
eth - ether
peth - petroleum ether'
ace - acetone
lig - ligroin
ale - alcphol•
CCl, - carbon tetrachloride
dil. alk. - dilute alkalis
CS2 - carbon disulfide
os — organic solvents
oos - ordinary organic solvents
-------
CRESOLS '
AN OVERVIEW
There are three isomers of cresol: o_-cresol, m-cresol,
and p_-cresol. ,+All three isomers as well as mixtures are art-
icles of commerce.' Cresols are solid or liquid at room tem-
perature (melting points 11-35°C>. They are slightly soluble
in water and soluble in'orgasic solvents.
Total annual production of cresols in the United States is
probably in excess of 100 million pounds. They are used for
a wide variety of purposes, including uses as disinfectants,
solvents, in ore flotation, and as intermediates in the pro-
duction of phosphate esters and phenolic resins. The number
of persons occupationally exposed to cresols is estimated to
be two million. They are also present in a number, of con-
sumer products, including disinfectants, metal cleaners, and
motor oil additives.
Cresols are manufactured both from petroleum"and from
"coal. The composition of the commercial products depends
on the method of production and upon the degree of refining.
Cresols are sold in a wide variety of grades, varying in com-
position, color, and boiling ranger Technical grade cresols
••
commonly contain xylenols and phenol. A less refined pro-
duct called creosote oil contains 10-20% by volume of tar from
the coking process.
-------
Cresols are relatively easily metabolized by mammals 'and
micro-organisms and are unlikely to undergo significant bio-
accumulation. They are moderately toxic to mammals by ingestion
and dermal exposure, and are corrosive to skin and other tissues.
No data are available on their toxicity by inhalation. Little
information is available on effects of chronic exposure.
In one experiment all 'three isomers of cresol were re-
ported to promote the carcinogenicity of dimethylbenzanthracene
*
on mouse skin. m-Cresol caused developmental abnormalities in-
toad embryos. Otherwise/ no significant information is avail-
able on the potential carcinogenicity, mutagenicity, or terato-
genicity of cresols.
Cresols have a broad spectrum of toxicity to micro-organisms
and are used as disinfectants and fungicides. There is little
other information on their potential toxicity to wildlife.
-------
CPESQLS
PART I
GENERAL .
I* Cresol (mixed isomers}
1.1 Identification CAS No. 001319773
' -NIOSHNo. G059500
1.2 Synonyms and'-Trade Names
Cresyiic acid; methyl phenol; hydroxytoluene;
tricresol; cresylol
(G23,G21,G16)
.1..3 Chemical Formula and Molecular Weight
C?H80 Mol. Wt. 108.15
(G23)
1.4 Chemical and Physical Properties
1.4.1 Description: A .mixture of isomers in which
m-isomer predominates , obtained
f~rom coal tar or petroleum;
colorless, yellow or pinkish
liquid; 'phenolic odor;' combustible;
becomes darker with age and on
exposure to light.
(G21,G23)
1.4.2 Boiling Point: 191 - 203° C (G21)
1.4. -3 Melting Point; 11 - 35° C (G21)
1.4.4 Absorption Spectrometry:
*"* k
T
No information found in sources searched
1.4.5 Vapor Pressure; No information found in sources searched
1.4.6 Solubility; Soluble in alcohol, glycol,
dilute alkalis, ether, chloro-
form;
Slightly soluble in water
(G21,G25)
1.4,7 -Cctanol/Water Partition Coefficient:
Log Poct = 2.70' (estimate)
-------
1.5 Production ana Use
1-5.1 Production:
60 Million Ibs (1968)
80 Million Ibs (1973)
(G25)
1*5.2 Use:
As a disinfectant;- intermediate in manufactur-
ing of phenolic resins, tricresyl phosphate,
salicylaldehyde, coumarin, and herbicides; as
an ore flotation agent; as a textile scouring__.^_,
agent; as an organic intermediate; as a- sur-..:T~
factant
Quantitative Distribution -cf Uses:
Phosphate esters
Magnet wire
Antioxidants
Resins
Exports
Cleaning and disinfectant
compounds
Ore flotation
Miscellaneous
Percent
22
15
15
15
10
6
6
11
100
Consumer Product Information;
. Cresol is present in:
automotive parts cleaner
metal cleaner, stripper, degreaser
disinfectant
motor oil additive
carbon remover
embalming supplies
i Estimates
I.S.J. Helease Rate;
30.4 Million Ibs
270HS Occupational Exposure:
(G21)
(G25)
(G35)
(G28)
Rank: 105
Estimates no. of persons exposed: 1,9 14, -000
t
(G29)
-------
1.7 Manufacturers
American Cyanamid Co.
Amoco Oil Co.
Crowley Tar Products Co., Inc,
Frefese Chemicals,
Koppers. Co., Inc.
Merichem Co.,
Mobil Oil Corp.
Northwest Petrochemical Corp.
Pitt-Consol Chemicals
Productol Chemical Co.
Sherwin-Williams Co.
United States Steel Corp.
(G25)
-------
CPESCIS
n. nt-Cresol
1.1 Identification CAS No.: 000108394
NIOSH No. : G061250
1.2 Synonyms and Trice Names
»
nj-CresyliC acid; m-methyiphenol; 3-nethyiphenol; l^ydrcay-3-nethyl-
benzene; ift-kresolT m-oxytoluexie
~ - (CIS)
*
1.3 Chemical FoEnula and Molecular Weight
CH--
C_HQ0 Mai. wt. 108. IS
t '
i-s
J • (G22)
1.4 QTemical and Physical Properties
1.4.1 EteseriPtion; Colorless to yellowish liquid; phenol-like
ry^f ' ~
' CG21)
1,4.2 Boiling Point: 202.2° C . (G22)
1.4.3 Melting Point: 11.5° C (G22)
1.4.4 Absorption Spectramtry:
-214,271,277,
log ^ - 3.79, 3.20, 3.27 (G22)
1.4.5 Vapor Pressure; 1 ma at 52.0° C (G22)
1.4.6 Solubility: Slightly soluble in water; *
• Soluble in hot water, organic solvents;
,.'•"" Soluble in all proportions in alcohol, ether,
acetone, benzene and carbon tetrachloride
1.4.7' Cctanol/Water Partition Coefficient
?oct
log P_. = 2.37
-------
'1.5 Production and Use
1.-5.1 Production.:
No information found in sources searched
1,5.2 Use: In disinfectants and fumigants; in photographic
developers, explosives (G23)
1.6 Exposure Sstiisates
t
1.6.1 Release Rate:
No information found in sources searched
. 1.6.2 NOHS Occupational Exposure;
Rank: 2731 *
Estimated no. of persons exposed: 9,000*
*rough estiicate (G29)
1.7 Manufacturers
Kcppers Co., Inc. • (G24)
-------
CRESOLS
HI. £-Cresol
1.1 Identification
CAS NO. .000095487
NIOSH NO, G063000
1.2 Synonyms and Trade Names
o_-Cresylic acid'; o-methyl phenol; 2-nethyl phenol?
orthocresol; l-hydroxy-Z-nethylbenzene; o-hydroxy-
toluene; o-methylphenol; o-oxytoluene; 2-nydroxy-
toluene ~
1.3 Chemical Formula and Molecular Weicht
(G16)
C?H80
1.4 Chemical and Physical Properties
1.4.1 Description:
Mol. Wt. 108.15
(G22)
1.4.2 Boiling Point;-
1.4.3 Melting Point;
1.4.4 Absorption Spectrometry;
Water
White crystals; phenol-like odor;
combustible; becomes dark with age
and exposure to air and light.
(G23,G21)
190.95° C (G22)
(G22)
30.94° C
Max
log £
= 219, 275 nia
3.71, 3.22
1.4.5 Vapor Pressure;
1.4.6 Solubility;
1 rnm at 38.2 C
(G22)
(G22)
Soluble in water and ordinary
organic solvents;
Very soluble in alcohol and ether;
Soluble in" all proportions in
acetone, benzene, caraon tetrachloride
(G22)
-------
1.4.7 Octanol/Water Partition Coefficient:
1.5 Production and Use
1.5.1 Production:
Log P ^ = 3.40
' oct
49..700 Million Ibs
20.481 Million Ibs
22.187 Million Ibs
(1972)
(1975)
(1976)
1.5.2 Use:
Disinfectant; solvent
1.6 Exposure Estimates
l.b.l Release Rate: 15.b Million Ibs
1.6.2 NOHS Occupational Exposure;
Rank: 1480
Estimates no. of persons exposed:
*rough estimate
1.7 Manufacturers
from coal tar:
Koppers Co., Inc.
Ferro Corp.
from petroleum:
Mericnem Co.
Ferro Corp.
Sherwin-Williams Co.
(G15)
(G28)
(G24)
(G24)
(G23)
(G28)
52,000*
(G29)
(G24)
-------
CRESOLS
IV. £-Cresol
1.1 Identification . CAS No.: 000106445
NIOSH NO. : G064750
1.2 Synonyms and frrade Names
*
4-Cresol; p_-cresylic acid; l-hydroxy-4-methylbenzene; p_-
hydroxytoluene; 4-hydroxy toluene; p_-Kresol; l-methyl-4-
hydroxybenzene; p_-methylphenol; 4-methylphenol; p_-oxyto-
luene; para-cresol; peyramethyl phenol
CG16)
1 . 3 Chemical. Formula and Molecular Weight
OH
K)l " C H 0 Mol. wt. 108.15
XTX 7 8
CH
3 (G22)
1.4 'Chemical and Physical Properties
1.4.1 Description; ' Crystalline mass; phenol-like
odor
(G21)
1.4.2 Boiling Point; 201.9* C ' (G22)
1.4.3 Melting Point: 34.8* C (G22)
• 1.4.4 Absorption Spectrometry:
« 280 nm
log 6 =3.23 (G22)
>
1.4.5 Vapor Pressure: 1 mm at 53.0° C (G22)
1.4.6 Solubility: Slightly soluble in water;
Soluble In hot water, organic solvents;
Soluble in all proportions in alcohol,
ether, acetone, benzene and carbon
tetrachloride
(G22)
1.4.7 Octanol/Water Partition Coefficient
Poct ** 2'35 <
-------
1.5 Production and Use
1.5.1 Production:
No information found in sources searched -
1.5.2 Use; As a chemical intermediate (G24)
1.6 Exposure Estimate
1.6.1 Re 1 e a s e .._Ra_t_e:
No information found in sources searched
1.S.2 NOHS Occupational Exposure
Rank: 2466
Estimated no. of persons exposed: 14/000*
*rough estimate
CG29)
1.7 Manufacturers
Sherwin-Williams Co.
(G24)
-------
CPFSOLS
StMIAHY OF aUUVCTERlSTICS
tfaroe
Cresol
(mixed isaners)
o-Cresol
m-Cresol
pj-Cresol
u)
i
Solubility
s in ale, glycol,
dil. alk, eth,
chl.
ss in HJD
IxxjP
oct
2.70
s in »20 and COS.
vs in ale and eth.
oo in ace, bz, CC1..
3.40
BS in H2O; s in hot 2.37
lUO, os;00 in ale,
eth, bz, ace, OC1.
S3 in H2Oj B in 2.35
hot H2O, ba; v° in
ale, eth, bz, ace,
CC1,,
Estimated
Environnvental
Release
(Million Ibs)
30.4
Production
(Million Ibs)
-GO (1968)
^-00 (1973)
15.6
49.7 (1972)
20,481(1975)
22.187(1976)
Estimated no.
of persons
exposed
(occupational)
1,914,000
52,000
9,000
14,000
Use
Disinfectant; phenolic
resins; tricresyl phos-
phate; ore flotation;
textile scouring agent;
organic intermediate;
nfg. of salicylaldehyde
coumarin, and herbicides
surfactant
Disinfectant, solvent
In disinfectants, funii-
gants, pliotographic
developers, explosives
cyclic intermediate
* No information found in sources searched.
-------
CRESOLS
PART II.
BIOLOGICAL PROPERTIES
2.1 Bioaccumulation
Log octahol/water partition coefficients are 3.40, 2.37, and
2.35 for the Q-, m-, and p_-isomers, respectively CG15) . The high
partition coefficient of the o-isomer is due to the steric effect
of the methyl group on the hydroxyl group. The high octanol/
water partition coefficients of the cresols indicate that bio-
accumulation in aquatic organisms is a possibility, but specific
data on such bioaccumulation are not available. By analogy with
phenol, which appears to be completely eliminated from the body
within 24 hours (G19), it is expected that cresols would not be
bioaccumulated in mammals '. Cresols in waste waters near indust-'
rial plants are reported to undergo rapid biodegradation (G14),
which indicate-s that cresols, like phenol, are relatively easily
metabolized.
2-2 Contaminants and Environmental Degradation or Conversion
Products
Cresols are sold in a wide variety of technical and special
grades, classified by color and distillation range (G25). The
composition of the various materials depends upon the starting
material and the method of production. A major source of cresols
is the tar-acid oil obtained as a by-product of coking of coal (G25)
Cresols (boiling above 204°C), available as a mixture of o-,
m-, and p_-isomers from tar acids- are called cresylic acid. A less
refined product called creosote oil contains 10-20% by volume of
the tar from the coking process; it is used as a wood preservative
(G25). Creosote oil may contain polynuclear aromatic hydrocarbons.
Xylenols and phenol are common impurities (or ingredients) of tech-
nical grade cresols (G25).
-------
The high environmental stability of the cresols in soils
(owing to their antimicrobial properties) contributes to their
widespread use as wood perservatives. p_-Cresol is degraded by
the hydroxyl radical and ozone in air and by organic peroxide
radicals in water; half life estimates are less than 1 day in
air and 10 days in water (G14) . The m- and p_-isomers are ex-
pected to behave similarly. Environmental degradation is likely
to be by air oxidation to give quinones and dihydroxybenzenes (G14)
Biodegfadation'products of cresols by sewage microorganisms
include carbon dioxide, methane, 3-methylcatechol, 2-hydroxy-6-
oxahepta-2,4-dienoic acid, oxalic acid, pyrocatechol,carboxylic
acid, and salicylic acid (G14). By analogy with phenol, cresols
may be methylated in th.e environment to form the corresponding
anisoles.
2.3 Acute Toxicity
The NIOSH Registry of Toxic Effects of Chemical Substances
(G16) reports the acute toxicity of cresols as follows:
Substance Parameter
Dosaae
Animal
Cresol LD50
LD50
O-Cresol LD50
"~ LD50
LD50 *
LDLo
LDLo-
LDLo
LD50
LOLo
LDLo
LDLo
LDLo
m-Cresol LD50
~ LD50 -
LD50
LD50
LDLo
LDLo
LDLo
LD50
LDLo
LDLo
LDLo
LDLO
1454 mg/kg
861 mg/kg
121 mg/kg
1100 mg/kg
344 mg/kg
410 mg/kg
55 mg/kg
940 mg/kg
1380 mg/kg
450 mg/kg
180 mg/kg
360 mg/kg
200 mg/kg
242 mgAg
620 mg/kg
350 mg/kg
828 mg/kg
450 mg/kg
180 rng/kg
1400 mgAg
2050 mg/kg
500 mg/kg
280 mg/kg
100 mg/kg
• 250 mg/kg
rat
mouse
rat
rat
mouse
mouse
cat
rabbit
rabbit
rabbit
rabbit
guinea pig
frog *
rat '
rat
rat
mouse
mouse
cat
rabbit
rabbit
rabbit
rabbit
guinea pig
frog
Route
oral
oral
oral
skin
oral
subcutaneous
subcutaneous
oral
skin *
subcutaneous
intravenous
intraperitoneal
subcutaneous
oral
skin
unknown
oral '
subcutaneous
subcutaneous
oral
skin
subcutaneous
intravenous
intraperitoneal
subcutaneous
-6,75"-
-------
(continued)
Substance Parameter Dosage Animal Route
£-Cresol LD50 207 mg/kg rat oral
LDSO 705 mg/kg rat ' skin
. LDSO 344 mg/kg mouse . oral
LDLo • 150 mg/kg mouse subcutaneous
LDSO 160 mg/kg mouse unknown
LDLo ' . 80 mg/kg cat subcutaneous
LDLo ' 620 mg/kg rabbit oral
LDSO • 301 mg/kg rabbit skin
LDLo 300 mg/kg rabbit subcutaneous
LDLo 180 mg/kg rabbit intravenous
LDLo 100 mg/kg ' .guinea pig intraperitoneal
LDLo • - 150 mg/kg fxog subcutaneous
Cresols are rated as moderately toxic to humans (G4) . Acute
exposures can cause muscular weakness, gastroenteric disturbances,
severe depression,"collapse, _and death (G38) . Organs attacked by
cresols include the central nervous system, liver, kidneys, lungs,
pancreas, spleen, eyes, heart, and skin (G38).. The type of exposure
to cresols determines, in part,, the toxic effects. .Cresols are highly
corrosive to any tissues they contact (G5) and are 'readily absorbed
by skin and mucous membranes. Systemic effects, including death,
occur after dermal exposure. Because their vapor pressure is low
at 25°C, cresols do .not usually constitute an acute inhalation
hazard. No data are available on the toxicity of cresol vapors to
humans (G39).
In animals, cresol toxicity varies with the isomer, the species
ajid the route of exposure. Reported LDSOs' vary from a low of 121
mg/kg in the rat (oral, pj-cresol) to a high of 2050 mgAg in the
rabbit (skin, m-cresol) (G16). Evidence for different biological
effects of the three isomers includes the observation that the ratios
between the LDSOs of the least toxic and most "toxic isomer^ vary from
as low as 1.8 (cutaneous, rat) to as high as 6.8 (cutaneous,'rabbit).
Furthermore, £-cresol, but neither o- nor m-cresol, produced*
permanent pigment loss in the hair of mice (1) .
2.4 Other Toxic Effects
Chronic poisoning from absorption of cresols through the skir.,
-------
mucous membranes or respiratory tract has not been well studied.
Campbell (2) presented incomplete studies showing that exposure
of mice to an atmosphere saturated with cresylic acid vapors for
1 hr/day on consecutive days caused irritation of the nose and
eyes., and death in some animals. Uzhdavini et al. (3) performed
poorly documented studies on the chronic effects of p^-cresol in-
halation. In mice, £hey found evidence for: tail necrosis; slowed
weight gain; cellular degeneration of the CMS; respiratory tract
hyperemia, edema, proliferation of cellular elements, and hemor-
rhagic patche's; myocardial fiber degeneration; and protein deposits
in liver and kidney cells. In rats, they reported alterations in
a conditioned reflex, and alterations in both peripheral blood and
bone marrow elements.
The Threshold Limit Value established by the ACGIH for eresols
is 5 ppm (Gil).
2.5 Care inogeni city
o, m, and p_-Cresol have been reported to promote the carcino-
genicity of dimethylbenzanthracene (DMBA) in skin tests with mice
(4). They were slightly less active as promoters than phenol in
this experiment (see table below).
No. mice Avg. no. % survivors
survivors/ papillomas with
Promoter* original no. per survivor papilloma
Benzene Control 12/12 0 0
20% phenol 22/27 1.50 64
20% o-cresol 17/27 1.35 59
20% m-cresol 14/29 0.93 50 ^
20% £-cresbl 20/28 0.55 35
* Initiator: 0.3% DMBA in acetone. Promoter in benzene.
Data at 12 weeks.
No carcinogenicity tests conducted with cresols alone have been
found in the searched literature.
-6,77-
-------
2.6 Mutagenicity
In onion root tips/ however, m- and p_-cresol produced cyto-
logical abnormalities including stickiness, erosion, pycnosis,
C-mitosis, polyphoidy, and chromosome fragmentation(5). o-Cresol
» .
did not appear as active (5) . These chromosomal effects do not
necessarily imply that the cresols will have genetic activity in
mammals. No other mutagenicity studies were found in the searched
literature.
2.7 Teratogenicity
No systematic studies of the teratogenic potential of the
cresols have been found. The only information available is
on the effect of m-cresol on embryos of a toad (Xenopus laevis)
at the neural tube stage.of development (6). Concentrations of
20 to 80 ppm, m-cresol caused two developmental- abnormalities:
edema and tail flexion..
2.8 Metabolic Inforroatioji
Very little is known about the metabolic fate of cresols
in mammals. One study showed that the cresols are excreted in
rabbit urine primarily as oxygen conjugates: 60-72% as
ether glucuronides and 10-15% as ethereal sulphates (7).
Paper chromatography showed that oj- and m-cresol are
hydroxylated and that p_-cresol forms p_-hydroxybenzoic acid (7) .
£-Cresol glucuronide was isolated from the urine of rabbits >
closed by stomach tube with p_-cresol, whereas' o- and m-cresol
were metabolized to 2,5-dihydroxytoluene (7) . No studies
have been traced of the biological effect of these and othar
possible metabolites of the cresols.
-------
2.9 Ecological Effects
The 96-hour LC50 of o-cresol to channel catfish (Ictalurus
punctatus) is reported to be 67 mg/1 (8). In tests with
perch and sunf^ish, .lethal concentrations (not LCSOs) were
determined in 1 hqur exposures. In perch (Perca fluviatilis),
lethal concentrations for o-, m- and p_-cresols were in the
range 10-20 ppm (9) . The Aquatic .Toxicity Rating (96-hour
TLm, species unspecified) for cresols is listed as 1,0-1 ?pm
(G16). Although o-cresol is less toxic to juvenile Atlantic
salmon (Salmo salar) than p_-cresol, the salmon avoided
o-cresol more efficiently (10).
Cresols have a broad spectrum of toxicity to microorganisms.
They are used as disinfectants and as fungicides to protect
materials such as wood. They are also reported to be active
against mycoplasmas (11), viruses (12), and-plant galls (13).
2.10 Current Testing ana Evaluation
A criteria document on cresols is planned for completion
in 1977 by NIOSH.
-------
REFERENCES
1. Shelley, W. B. o-Cresol: cause of ink-induced hair depitiiient-
ation in mice. Brit. J. Dermatol. 90:169-174 (1974).
2. Campbell, J. Petroleum cresylic acids - a study of their toxi-
city and th^ toxicity of cresylic disinfectants. Soap Sanit.
Chera. 17:103-111 (1941).
3. Uzhdavini,1 E.R. , Astafyeva, I.K. , Mamayeva, A.A. and Bakhtizina,
G.2. Inhalation toxicity of o-cresol. Tr. Ufim. Nauchno-Issled
Inst. Gig. Profzabol. 7:115-119 (1972). (Russian)
4. Bontwell, R.K., and Bosch, D.K. The tumor-promoting action of
phenol" and related comnounds for mouse skin. Cancer Res.
19:413-424 -(1959).
5. Sharma, A.K. and Ghosh, S. Chemical basis of the action of
cresols and nitrophenols on chromosomes. The Nucleus 8:183-
190 (1965).
6. Johnson, D.A. The effects' of meta-eresol on the embryonic
development of the African Clawed Toad, Xenopus laevis.
J. Ala. Acad. Sci. 44:1-77 (1973).
7. Bray, H.G., Thorpe, W.V., and White, K. Metabolism of
- derivatives of toluene.. 4. Cresols.. Biochem. J. 46:275-
278 (1950). V
8. Clemens, H..P. , and Sneed, K.E. Lethal dose of several com-
mercial chemicals for fingerling channel catfish. U.S. Fish.
Wildlife Serv. Spec. Sci. Rep. Fisheries 316 (1959). .
9-. Jones/ J.R.E. Fish >.and River Pollution. Butter-worths, London
(1964). Pp 118-153.
10. Zitko, V., and Carson, W.G. Avoidance of organic solvents
and substituted phenols by juvenile Atlantic salmon. Fish-
eries Res. Board Can. MS. Rep. 1327 (1974).
• >
11. Kihara, K., Sasaki, T., and Arima, S. Efeect of antiseptics
and detergents on Mycoplasma. Igakii 7.Q Seibutsugaku, 83:5-8
(1971).
12. Sellers, R. F. The inactivation of foot-and-mouth disease
virus by chemicals and disinfectants. Vet. Rec., 83:504-506
(1963).
13. Schroth, M.N. and Hildebrand, D.C. A chemotherapeutic treatment
for selectively eradicating crown gall and olive knot neoplasms.
Phytopath. 58:848-854 (1954).
-------
GENERAL REFERENCES
Gl. Browning, E. Toxicity anrf Metabolism of Industrial Solvents.
Elsevier, Amsterdam (1965).
'«
G2. Browning, E. Toxicity of Industrial Metals, 2nd ed. Appleton-
Century-Crofts, New York (1969).
G3. Fairhall, L.T. Industrial' Toxicology, 2nd ed. Williams •
& Wilkins Co. (1969).
» +
G4. Sax, N.I. Dangerous Properties of Industrial Materials,
3rd--ed., Reinhold Publishing Corp., New York (1975).
G5. Chemical Safety Data Sheets. Manufacturing Chemists Asso-
ciation, Washington, D.C.
G6. Industrial Safety Data Sheets. National Safety Council,
Chicago, Illinois.
G7. Shepard, T.H. Catalog of Teratogenic Agents. Johns Hopkins
University Press, Baltimore (1973).
G8. Thienes, C.L. & Haley, T.J. Clinical Toxicology. Lea &
Febiger, Philadelphia (1972).
G9. IARC Monographs on the Evaluation of Carcinogenic Risk of
Chemicals to Man. Lyon, France. WHO, International Agency
for Research on Cancer.
G10. Debruin, A. Biochemical Toxicology of Environmental Agents.
Elsevier/North-Holland, Inc., New York (1975).
Gil. Threshold Limit Values for Chemical Substances and Physical
Agents in the Workroom Environment with Intended Changes
for 1976. American Conference of Government Industrial
Hygienists.
>
G12. Chemicals Being Tested for Carcinogenicity by the Bioassay
Program, DCCP. National Cancer Institute (1977).
G13. Information Bulletin on the Survey of Chemicals Being Tested
For Carcinogenicity, No. 6. WHO, Lyon, France (1976),
G14. Brown, S.L., et_ al_. Research Program on Hazard Priority
Ranking of Manufactured Chemicals, Phase II - Final Report
..to National' Science Foundation. Stanford Research Institute,
Menlo Park, California (1975) .
-------
G15. Dorigan, J. , et aL. Scoring of Organic Air Pollutants,
Chemistry, Production and Toxicity of Selected Synthetic
Organic Chemicals. MITRE, MTR-724S (1976),
.G16. NIOSH Registry of Toxic Effects of Chemical Substances (1976).
G17. Kirk-Othzner Encyclopedia of Chemical Technology. Edited
Standen,A(ed.),Interscience Publishers, New York (1963, 1972). .
GIB. Survey of Compounds Which Have Been Tested for Carcinogenic
Activity Through 1972-1973 Volume. .DHEW Publication No.
NIH73-453, National Cancer Institute, Rockville, Maryland.
G19. Criteria for a Recommended Standard - Occupational Exposure
to .... , prepared ny NIOSH .
G20. Suspected Carcinogens - A subfile of the NIOSH Toxic Sub-
stance" List (1375).
G21. The Condensed Chemical Dictionary, 9th ed. Van Nostrand
Reinhold Co., New York (1977).
G22. Handbook of Chemistry and Physics , 57th ed. The Chemical
Rubber Company, Cleveland, Ohio (1976) .
G23. The Merck Index, 9th ed. Merck & Co., Inc., Rahway, N.J.
(1976) .
G24. Synthetic Organic Chemicals, United States Production and
Sales. 1966-76. U.S. International Trade Commission, U.S. •
Government Printing Office, Washington, D.C.
G25. Lowenheim, F.A. & Moran, M.K. Faith. Keyes, and Clark's
Industrial Chemicals, 4th ed. John wiley & Sons, New York
(1975) . .
G26. Gosselin, Hodge, Smith & Gleason. ..Clinical Toxicology of
Commercial Products, 4th ed. The Williams and Wilkins Co.,
Baltimore (1975).
G27. Chemical Consumer Hazard Information System. Consumer Product
Safety Commission, Washington, D.C. (1977)^
G2S. A Study of Industrial Data on Candidate Chemicals for Test-
ing. Stanford Research Institute, Palo Alto, California (1976,7).
G29. National Occupational Hazards Survey (NOHS). National
Institute for Occupational Safety and Health, Cincinati •>
Ohio (1976).
G30. The Aldrich Catalog/Handbook of Organic and Biochemicals.
.Aldriuh Chemical Co., Inc. (1977-78).
-------
G31. McCutcheon's Functional Materials 1977 Annual. McCutcheon
Division, MC Publishing Co. (1977).
G32. Hampel & Hawley. The Encyclopedia of Chenistry, 3rd ed.
Van Nostrand Reinhold Co., New YorK 11973).
G33. Casarett, L. J. & Doull, J. Toxicology, the Basic Science
of Poisons., Macmillan Publishing Co." Inc., New York (1975).
G34. 'EPA/Office of Research and Development, Chemical Production.
G35. CTCP/Rochester Computer Service. (See Reference No. G26.)
G36. Leo, A., Hansch, C. & Elkins, D. Partition coefficients.
and their uses. . Chem. Rev. 71:525-616 (1571).
G37. 1977-78 OPp Chemical Buyers Directory.
G38. Patty, F.A, Industrial Hygiene and Toxicology.. Vol. 2, 2nd ed.
Wiley Interscience, New York (1963)..
G39. Directory of Chemical Producers. Stanford Research Institute,
Menlo Park, California (1977).
-------
No. 55
Crotonaldehyde
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.
-------
CROTONALDEHYDE
SUMMARY
Crotonaldehyde is not expected to be overly persistent in
water or the atmosphere. It is not expected to bioconcentrate.
It has been detected in finished drinking water and in sewage
treatment plant effluents.
An increased incidence of malignant neoplasms has been
observed in workers at an aldehyde factory who were exposed to
crotonaldehyde, among other substances. There is, however, no
indication that crotonaldehyde was the causative factor in the
excess incidence of cancer.
Pathologic change was observed in the testes of mice receiv-
ing crotonaldehyde in the drinking water (0.2 g/1) for one month.
I. INTRODUCTION
Crotonaldehyde (CH3CH=CHCHO; molecular weight 70.1) is a
water-white, mobile liquid with a pungent, suffocating odor
(Hawley, 1977). It has the following physical/chemical
properties (U. S. EPA, 1979a; Hawley, 1977):
Boiling Point: 102°C
Melting Point: -60°C
Vapor Pressure: 19 mm Hg at 20°C
Solubility: very soluble in water;
also soluble in many
organic solvents.
-------
A review of the production range (includes importation)
statistics for crotonaldehyde (CAS No. 4170-30-3) which was
listed in the initial TSCA Inventory (1979b) has shown that
between 1 million and 8 million pounds of this chemical were
produced/imported in 1977. _/
Crotonaldehyde is used as an intermediate in the manufacture
of n-butanol and crotonic and sorbic acids; solvent in the
purification of mineral oil; intermediate in resin and rubber
antioxidant manufacture; and in organic syntheses (NCI, 1978).
Other uses are as a warning agent in fuel-gas, insecticides,
leather tanning, production of rubber accelerators, and as an
alcohol denaturant (Hawley, 1977).
II. ENVIRONMENTAL FATE
Formaldehyde, the simplest aldehyde, is almost entirely
hydrated in water, thus it is nonvolatile and is inactive toward
photochemical dissociation. Higher aldehydes, such as crotonal-
dehyde, are less hydrated in water, more volatile, and somewhat
active toward photochemical degradation (Calvert and Pitts,
1966). Crotonaldehyde is expected to be oxidized in water at the
double bond to form keto aldehydes and cleavage products (U.S.
EPA, 1977). Crotonaldehyde biodegrades at a slow to moderate
This production range information does not include any pro-
duction/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 Inven-
tory, including production range information, are subject to
the limitations contained in the Inventory Reporting Regula-
tions (40 CFR 710).
-------
rate; acclimated bacteria can speed the degradation rate (U.S.
EPA, 1979a). In general, neither crotonaldehyde nor its
oxidation products are expected to be overly persistent in water
(U.S. EPA, 1977).
In air, aldehydes are expected to photodissociate to RCO and
H atoms rapidly and competitively with their oxidation by HO
radical. The projected half-life is on the order of 2 to 3 hours
{Calvert and Pitts, 1966). Oxidation of crotonaldehyde by HO
radical should result in addition at the double bond to form a
keto aldehyde (U.S. EPA, 1977). Crotonaldehyde is a reactive
component of auto exhaust and may contribute to smog (Dimitriades
and Wesson, 1972).
B. Bioconcentration
Crotonaldehyde is not expected to bioconcentrate (based on
its similarity to acrolein) (U.S. EPA, 1977).
C. Environmental Occurrence
Crotonaldehyde has been detected in finished drinking water.,
sewage treatment plant effluents (U.S. EPA, 1976), in wastewater
used for irrigation of potatoes (Dodolina _et^ _al_., 1976), and the
atmosphere (IARC, 1976).
Crotonaldehyde occurs naturally in essential oils extracted
from the wood of oak trees (Egorov, 1976). It has also been
found in the volatiles from cooking mutton (Nixon _et_ _al_. , 1979)
and in tobacco and tobacco smoke constituents (Pilott, 1975).
-------
III. PHARMACOKINETICS
Although no information was found specifically on the metab-
olism of crotonaldehyde, it is probably oxidized to an acid and
subsequently to CCu in the same manner as other small aliphatic
aldehydes. Crotonaldehyde is a potential alkylating agent by the
metabolic formation of an activated epoxy derivative at the
double bond and via reaction with amino groups of cellular
macromolecules (NCI, 1978).
IV. HEALTH EFFECTS
A. Carcinogenicity
An increased incidence of malignant neoplasms has been
observed in workers at an aldehyde factory who were exposed to
acetaldehyde, butyraldehyde, crotonaldehyde, aldol, several
alcohols, and longer chain aldehydes. Crotonaldehyde was found
in concentrations of 1-7 mg/m3. Of the 220 people employed in
this factory, 150 had been exposed for more than 20 years. Dur-
ing the period 1967 to 1972, tumors were observed in nine males
(all of whom were smokers). The tumor incidences observed in the
workers exceeded incidences of carcinomas of the oral cavity and
bronchogenic lung cancer expected in the general population and,
for the age group 55-59 years, the incidence of all cancers in
chemical plant workers. There is no indication that crotonalde-
hyde was the causative factor in the excess incidence of cancer
(Bittersohl, 1974, 1975).
-------
B. Mutagenicity
Schubert (1972) reported chromosome breakage in human lymph-
ocytes exposed to crotonaldehyde in vitro. When tested in
Salmonella typhimurium (tester strains TA1535, TA1537, TA1538,
TA100, and TA98) both -in the presence and absence of a metabolic
activation system, crotonaldehyde was nonmutagenic. It also
failed to increase the incidence of mitotic recombination in
Saccharomyces cerevisiae D3 in the presence and absence of a
metabolic activation system (NCI, 1978).
C. Reproductive Effects
Pathologic change was observed in the testes of mice one
month following a single intraperitoneal injection of crotonalde-
hyde (1 mg/mouse). In a related study, similar changes were
observed in the testes of mice receiving crotonaldehyde in the
drinking water (0.2 g/1) for one month (Auerbach £t_ ai. , 1977;
Moutschen-Dahmen _et_ _al_. , 1975; Moutschen-Dahmen et_ SL!. , 1976).
D. Other Toxicity
Skog (1950) studied the effects of lower aliphatic aldehydes
in rats and mice. When administered subcutaneously or by
inhalation, crotonaldehyde caused lung edema and mild narcosis.
Death was delayed and probably resulted from the lung damage.
With cats, similar effects were seen, with death due to lung
edema or bronchial pneumonia occurring within 24 hours for injec-
tion and between 6 and 48 hours for inhalation studies (Skog,
1950).
*
The oral LDg0 for crotonaldehyde in the rat is 300 mg/kg;
the 30-minute LC50 in the rat is 4000 mg/kg. The rabbit dermal
LD50 is 380 mg/kg (NIOSH, 1979).
-------
E. Other Relevant Information
A case of apparent sensitization to crotonaldehyde has been
reported in a laboratory worker who handled "small" amounts of
the material (ACGIH, 1971).
Crotonaldehyde is a strong mucous membrane irritant (NIOSH,
1978).
V. AQUATIC EFFECTS
The 96-hour LC5Q (partial flow-through system) for crotonal-
dehyde in bluegill sunfish is 3.5 ppm; in tidewater silversides
the 96-hour LC5Q is 1.3 ppm (Dawson, 1975/1977).
VI. EXISTING GUIDELINES
The OSHA standard for crotonaldehyde in air is a time
weighted average (TWA) of 2 ppm (39CFR23540).
-------
References
ACGIH. American Conference of Governmental and Industrial
Hygienists, Documentation of the threshold limit values,
Cincinnati, Ohio. 1971.
Auerbach, C. et al. Genetic and cytogenetical effects of
formaldehyde and related compounds. Mut. Res. 39, 317-362, 1977.
Bittershol, G. Epideraiological investigations on cancer in
workers exposed to aldol and other aliphatic aldehydes. Arch.
Geschwalstforsch. 43, 172-176, 1964.
Bittersohl, G. Env. Qual. Safety 4, 235-238, 1975. {as cited in
NCI, 1978).
Calvert, J. G. and J.N. Pitts. Photochemistry. Wiley and Sons,
New York, 899 pp. 1966. (as cited in U.S. EPA, 1977).
Dawson, G.W. et al. The acute toxicity of 47 industrial chemi-
cals to fresh and salt water fishes. J. Hazardous Materials l^,
303-318, 1975/1977.
Dimitriades, B. and T.C. Wesson. Reactivities of exhaust
aldehydes. J. Air Poll. Contr. Assoc. 2(1), 33-38, 1972.
Dodolina, V. T. _et_ al. Vestn. S-Kh. Nauki (Moscow) _6_, 110-113,
1976. (as cited in NCI, 1978).
Egorov, I. A. et_ al. Prikl. Biokhim. Mikrobiol. 12(1), 108-112,
1976. (as cited in NCI, 1978).
Hawley, G.G. 1977. Condensed Chemical Dictionary, 9th edition.
Van Nostrand Reinhold Co.
IARC (International Agency for Research on Cancer). IARC mono-
graphs on the evaluation of carcinogenic risk of chemicals to
man. 13, 311, 1976.
Moutschen-Dahmen, J. et al. Genetical hazards of aldehydes from
mouse experiments. Mut. Res. 29(2), 205, 1975.
Moutschen-Dahmen, J. et al. Cytotoxicity and mutagenicity of two
aldehydes: Crotonaldehyde and butyraldehyde in the mouse. Bull.
Soc. R. Sci. , Liege _45_, 58-72, 1976. (as cited in NCI, 1978).
NCI (National Cancer Institute). Chemical Selection Worki/ig
Group. September 28, 1978.
NIOSH (National Institute for Occupational Safety and Health).
Information Profiles on Potential Occupational Hazards-Classes of
Chemical. 1978
-------
NIOSH (National Institute for Occupational Safety and Health).
Registry of Toxic Effects of Chemical Substances. 1979.
Nixon, L. N. _et^ _al_- Nonacidic constituents of volatiles from
cooked mutton. J. Agric. Food Chem. 27(2), 355-359, 1979.
Pilott, A. et al. Toxicology _5_, 49-62, 1975. (as cited in NCI,
1978).
Schubert, J. _et_ _al. EMS Newsletter _6_, 17, 1972. (as cited in NCI,
1978).
Skog, E. A toxicological investigation of lower aliphatic alde-
hydes I. Toxicity of formaldehyde, acetaldehyde, propionaldehyde,
and butyraldehyde; as well as of acrolein and crotonaldehyde.
Acta Pharmacol. £, 299-318, 1950. (as cited in NIOSH, 1978).
U.S. EPA. Frequency of organic compounds identified in water.
PB-265 470, 1976.
U.S. EPA. Review of the Environmental Fate of Selected Chemi-
cals. EPA-560/5-77-003, 1977.
U.S, EPA. Oil and Hazardous Materials. Technical Assistance Data
System (OHMTADS DATA BASE), 1979a.
U.S. EPA. Toxic Substances Control Act Chemical Substances Inven-
tory, Produciton Statistics for Chemicals Listed on the Non-
Confidential Initial TSCA Inventory, 1979b.
Sf
-------
No. 56
Cyanides
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.
-------
CYANIDES
SUiMMARY
Cyanide is well-known as an acute, rapidly acting poison
which has caused numerous deaths, primarily in occupational
situations. The mechanism of cyanide intoxication is attrib-
utable to the biochemical inhibition of cellular respiration,
which produces a condition resembling acute hypoxia. De-
spite the considerable potency of cyanide as an acute poison,
repeated sub'lethal exposures do not result in cumulative ad-
verse effects in animals or man. In a chronic feeding study
in rats, a no observable adverse effect level (NOAEL) was
found to be 12 mg/kg/day. Extrapolation of this value to
humans, using the application of a safety factor of 100,
results in an acceptable daily intake for man (ADI) of 8.4 rag.
Cyanide exists in water in the free form (CN~ and HCN),
which is extremely toxic, or in a form bound to organic
or inorganic moieties which is less toxic. Cyanide is lethal
to freshwater fishes at concentrations near 50 jjg/1 and
has been shown to adversely affect invertebrates and fishes
at concentrations near 10 jug/1. Very few saltwater data
have been generated. Cyanide affects fish and invertebrates
by inhibiting utilization of available oxygen for metabolism
at the cellular level of respiration.
-------
CYANIDES
I. INTRODUCTION
This profile is based primarily upon the Ambient Water
Quality Criteria Document for Cyanides (U.S. EPA, 1979).
The National Institute for Occupational Safety and Health
(NIOSH, 1976) has also prepared a recent comprehensive review
of health hazards associated with hydrogen cyanides (HCN)
and commercially important cyanide salts (NaCN, KCN, and
Ca(CN)2).
The toxicologic effects of cyanides are based upon
their potential for rapid conversion by mammals to HCN.
Cyanide production in the United States is now over 700
million pounds per year and appears to be increasing steadily
(Towill, et al. 1978). The major industrial users of cyanide
in the United States are the producers of steel, plastics,
synthetic fibers and chemicals, and the electroplating and
metallurgical industries (NIOSH, 1976; Towill, et al. 1978).
II. EXPOSURE
A. • Water
Cyanide exists in water in the free form (CN~
and HCN), or bound to organic or inorganic moieties. Cya-
nide is not commonly found in United States water supplies.
Among 2,595 water samples tested, the highest cyanide con-
centration found was 8 ppb (Towill, et al. 1978). The vola-
*
tility of HCN, the predominant form in water, accounts in
part for the low levels usually measured. The U.S. EPA
(1979) has estimated the bioconcentration factor of cyanide
at 2.3.
t
-------
B. Food
Except for certain naturally occurring organoni-
triles in plants (e.g., cyanogenic glycosides, such as amyg-
dalin), it is uncommon to find cyanide in foods.
C. Ambient Air
There is insufficient information available to
estimate population exposures to cyanide via ambient air
(U.S. EPA, 1979).
Ill PHARMACOKINETICS
A. Absorption
The common inorganic cyanides are rapidly absorbed
across the skin (Drinker, 1932; Potter, 1950; Tovo, 1955;
Walton and Witherspoon, 1926), stomach and duodenum, and
lungs (Goesselin, et al. 1976). Quantitative estimates
of the rate of penetration by various routes of exposure
are unavailable, however. The rapid absorption of cyanide
is evidenced by the fact that death may be produced within
a matter of minutes following inhalation or ingestion.
B. Distribution
Cyanide is distributed to all organs and tissues
via the blood, where its concentration in red cells is greater
than that in plasma by a factor of two or three. This may
be due, at least in part, to a preferential binding of cya-
nide to methemoglobin (Smith and Olson, 1973'j . Although
quantitative data are lacking, it is predicted that cyanide
*
may readily cross the placenta.
-------
C. Metabolism
By far, the major pathway for the metabolic detoxi-
fication of cyanide involves its conversion to thiocyanate
via the enzyme rhodanase (deDuve, et al. 1955). A minor
pathway for cyanide metabolism involves nonenzymatic conjuga-
tion with cysteine, a reaction which accounts for no more
than 15 percent of cyanaide metabolism in the rat (Wood
and Cooley, 1956). Very small amounts of cyanide can be
excreted unchanged (as HCN) or converted to carbon dioxide
{Friedberg and Schwarzkopf, 1969).
D. Excretion
Among rats given 30 mg of sodium cyanide intra-
peritoneally over eight days, it was estimated that 80 per-
cent of the total dose was excreted in the urine in the
form of thiocyanate (Wood and Cooley, 1956). Cyanide does
not appear to accumulate significantly in any body compart-
ment with chronic exposures.
IV. EFFECTS
A. Carcinogenicity
Pertinent data confirming the carcinogenicity
of cyanide were not found in the available literature.
B. Mutagenicity
Pertinent data concerning the mutagenicity of
cyanide were not found in the available literature.
C. Teratogenicity
Cyanide is not known to be teratogenic. However,,
thiocyanate, the major metabolic product of cyanide rn vivo.
-------
has produced developmental abnormalities in the chick (Nowinski
and Pandra, 1946) and ascidian embryo (Ortolani, 1969) at
high concentrations.
D. Other Reproductive Effects
Pertinent information regarding the possible ef-
fect of cyanide on fertility or reproductive success was
not found in the available literature.
E. Chronic Toxicity
Human inhalation of 270 ppm HCN brings almost
immediate death, while 135 ppm is fatal after 30 minutes
of exposure (Dudley, et al. 1942). The mean lethal dose
of HCN and its alkali metal salts by ingestion in humans
is in the range of 50 to 200 mg (1 to 3 rag/kg), with death
coming in less than one hour (Gosselin, et al. 1976). In
non-fatal poisonings, recovery is generally rapid and com-
plete. The mechanism of acute cyanide intoxication can
be attributed to the biochemical inhibition of cytochrome
c oxidase, the terminal enzyme complex in the respiratory
electron transport chain of mitochondria (Gosselin, et al.
1976) . The major feature of cyanide poisoning resembles
the effects of acute hypoxia, which results in a decreased
utilization of oxygen by the tissues. Cyanide poisoning
differs from other types of hypoxia in that the oxygen ten-
sion in peripheral tissues usually remains normal or may
even be elevated {Brobeck, 1973).
Despite the high lethality of large single doses
or acute inhalation exposures to high vapor concentrations
t
-70O-
-------
of cyanide, repeated sublethal doses do not result in cumula-
tive adverse effects (Hertting, et al. 1960; Hayes, 1967;
American Cyanamid, 1959).
F. Other Relevant Information
Since cyanide acts by inhibiting cytochrome c
oxidase, it is reasonable to assume that any other inhibitor
of the same enzyme (e.g. sulfide or azide) would have toxic
effects synergistic with (or additive to) those of cyanide.
This has not been demonstrated experimentally, however.
Cyanide poisoning is specifically antagonized
by any chemical agent capable of rapidly generating methemo-
globin _in v_ivo, such as sodium nitrite, or other aromatic
nitro and amino compounds (Smith and Olson, 1973).
V. AQUATIC TOXICITY
A. Acute Toxicity
There have been numerous studies investigating
the toxicity of cyanide in freshwater fish. Free cyanide
concentrations in the range of about 50 to 200 ug/1 have
eventually proven fatal to most species. Certain life stages
and species of fish appear to be more sensitive to cyanide
than others. Eggs, sac fry, and warmwater species tended
to be the most resistant.
Several authors have reported increased cyanide
j-
toxicity with the reduction of dissolved oxygen or with
a rise in water temperature. However, water alkalinity,
hardness, and pH below 8.3 have not been shown to have a
pronounced effect on the acute toxicity of cyanide to fish.
The reported range for LC5Q values for freshwater fish is
*
-101-
-------
from 52 }ig/i, for juvenile brook trout, to 507 pg/1, for
sac fry brook trout, Salvelinus fontinalis. For the fresh-
water invertebrates, the results from 11 acute tests on
6 species have shown a range of LCj-Q values from 83 pg/1
for cladoceran, Daphnia pulex to 760,000 )ag/l for snail,
Goniobasis livescens.
The only saltwater species to be studied is the
oyster. A short exposure of an oyster to cyanide resulted
in supression of activity after 10 minutes of exposure to
150 ^ug/1 {U.S. EPA, 1979).
B. Chronic Toxicity
Based on long-term tests with bluegills (embryo-
larval) and reproduction by brook trout and fatheads, the
geometric mean of the chronic effect level concentrations
is 9.6 ;jg/l (Koenst, et al. 1977; Lind, et al. 1977; Kimball,
et al. 1978) . Life cycle tests on the scud, Gammarus pseudo-
limnaeus, and the isopod, Ascellus communis, show the chronic
values to be 18.3 and 34.1 ug/1, respectively {U.S. EPA,
1979). The chronic toxicity of cyanide in marine species
has not been reported.
C. Plant Effects
In the only plant test reported, 90 percent of
the blue-green alga, Microcystis aerusinoss, was killed
when exposed to a free cyanide concentration'of 7,790
{Fitzgerald, et al. 1952).
There was an inhibition of respiration in the
marine alga, Prototheca jzopf_!, at 3,000 ^ig/1 and enzyme
g
-7O2-
-------
inhibition in Chlorella sg. at 30,000 jig/1 (Webster and
Hackett, 1965; Nelson and Tolbert, 1970).
D. Residue
No residue data is available for cyanide toxicity
in either salt or freshwater species. The U.S. EPA (1979)
has estimated the bioconcentration factor of cyanide to
be 2.3.
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 change.
A. Human
The U.S. Public Health Service Drinking Water
Standards of 1962 established 0.2 mg CN~/1 as the acceptable
level for water supplies. A similar criterion has been
adopted by the Canadian government (Health and Welfare Canada,
1977). In addition to defining the 0.2 mg CN~/1 criterion,
the U.S. Public Health Service (1962) has set forth an "objec-
tive" of 0.01 mg CN~/1 in water, "because proper treatment
will reduce cyanide levels to 0.01 mg/1 or less."
The U.S. Occupational Safety and Health Administra-
tion (OSHA) has established a permissible exposure limit
for KCN'and NaCN at 5 mg/m as an eight-hour'time-weighted
average. The National Institute for Occupational Safety
*
and Health (NIOSH) recommends 5 mg/m as a ten minute ceil-
ing for occupational exposure to KCN and NaCN.
-------
The OSHA permissible limit for exposure to HCN
is 10 ppm (11 mg/m ) as an eight-hour time-weighted average.
NIOSH recommends 5 mg/m as a ten minute ceiling level for
exposure to HCN.
Based upon the results of a two year chronic feed-
ing study in rats, the U.'S. EPA (1979) has calculated an
acceptable daily intake (ADI) of cyanide for man to be 8.4
mg/kg. This value was derived from the no observable adverse
effect level (NOAEL) for rats of 12 mg/kg/day and the applica-
tion of a safety factor of 100. The corresponding draft
water quality criterion derived from these data is 4.11
mg/1. However, the U.S. EPA (1979) has recommended that
the U.S. Public Health Service Drinking Water Standard
of 200 jug/1 be retained as a safe level for man.
B. Aquatic
For free cyanide (expressed as CN), the draft
criterion to protect freshwater aquatic life is 1.4 ug/1
as a 24-hour average, and the concentration should not exceed
38 ^ig/1 at any time (U.S. EPA, 1979).
Draft saltwater criterion is not available for
cyanide toxicity, because of the paucity of valid data (U.S.
EPA, 1979).
Sf
-704-/-
-------
CYANIDES
REFERENCES
American Cyanamid Co. 1959. Report on sodium cyanide:
30-day repeated feeding to dogs. Central Med. Dep.
Brobeck, T.R. 1973. Best and Taylor's physiological basis
of medical practice. 9th ed. Williams and Wilkins Co.,
Baltimore.
de Duve, C., et al. 1955. Tissue fractionation studies:
6. Intracellular distribution patterns of enzymes in rat-
liver tissue. Biochem. Jour. 60: 604.
Drinker, P. 1932. Hydrocyanic acid gas poisoning by absorp-
tion through the skin. Jour. Ind. Hyg. 14: 1.
Dudley, H.C., et al. 1942. Toxicology of acrylonitrile
(vinyl cyanide): II. Studies of effects of daily inhalation.
Jour. Ind. Hyg. Toxicol. 24: 255.
Fitzgerald, G.P., et al. 1952. Studies on chemicals with
selective toxicity to blue-green algae. Sewage Ind. Wastes
24: 888.
Friedberg, K.D., and H.A. Schwarzkopf. 1969. Blausaureexhala-
tion bei der Cyanidvergiftung (The exhalation of hydrocyanic
acid in cyanide poisoning). Arch Toxicol. 24: 235.
Gosselin, R.E., et al. 1976. Clincial toxicology of com-
merical products. 4th ed. Williams and Wilkins Co., Baltimore,
Hayes, W.T. Jr. 1967. The 90-dose kE>5Q and a chronicity
factor as measures of toxicity. Toxicol. Appl. Pharmacol.
11: 327.
Hertting, G. , et al. 1960. Untersuchungen uber die Folgen
einer chronischen Verabreichung akut toxicher Dosen von
Natriumcyanid an Hunden. Acta Pharmacol. Toxicol. 17: 27.
Kimball, G. , et al. 1978. Chronic toxicity of hydrogen
cyanide to bluegills. Trans. Am. Fish. Soc. 107: 341.
Koenst, W., et al. 1977. Effect of chronic'exposure of
brook trout to sublethal concentrations of hydrogen cyanide.
Environ. Sci. Technol. 11: 883.
f
Lind, D., et al. 1977. Chronic effects of hydrogen cyanide
on the fathead minnow. Jour. Water Pollut. Control Fed.
49: 262.
-------
National Institute for Occupational Safety and Health. 1976.
Criteria for recommended standard occupational exposure
to hydrogen cyanide and cyanide salts (NaCN, KCN and Ca(CN)2),
NIOSH Publ. No. 77-108. Dep. Health Edu. Welfare. U.S. Govern-
ment Printing Office, Washington, D.C.
Nelson, E.B., and N.E. Tolbert. 1970. Clycolate dihydro-
genase in green algae. Arch. Biochem. Biophys. 141: 102.
Nowinski, W.W., and J. Pandra. 1946. Influence of sodium
thiocyanate on the development of the chick embryo. Nature
157: 414.
Ortolani, G. 1969. The action of sodium thiocyanate (NaSCN)
on the embryonic development of the ascidians. Acta Embryol.
Exp. 27-34.
Potter, A.L. 1950. The successful treatment of two recent
cases of cyanide poisoning. Br. Jour. Ind. Med. 7: 125.
Smith, R.P., and M.V. Olson. 1973. Drug-induced methemo-
globinemia. Semin. Hematol. 10: 253.
Tovo, S. 1955. Poisoning due to KCM absorbed through skin.
Mineria Med. 75: 158.
Towill, L.E., et al. 1978. Reviews of the environmental
effects of pollutants: V. Cyanide. Inf. Div. Oak Ridge
Natl. Lab. Oak Ridge, Tenn.
U.S. EPA. 1979. Cyanides: ambient water quality criteria.
(Draft) EPA PB296792. National Technical Information Ser-
vice, Springfield, VA.
Walton, D.C., and M.G. Witherspoon. 1926. Skin absorption
of certain gases. Jour. Pharmacol. Exp. Ther . 26: 315,
Webster, D.A. , and D.P. Hackett. 1965. Respiratory chain
of colorless algae. I. Chlorophyta and Euglenophyta .
Plant Physiol. Lancaster ^D~7~'"
Wood, J.L., and S.L. Cooley. 1956. Detoxication of cyanide
by cystine. Jour. Biol. Chem. 218: 449.
JXi
-------
No. 57
Cyanogen Chloride
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
-7O7-
-------
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.
-------
CYANOGEN CHLORIDE
I. INTRODUCTION
Cyanogen chloride is a colorless gas at room temperature with a molec-
ular weight of 61.48, a melting point of -6.5Qc, a boiling point of
13.8°C, and a specific gravity of 1.186.' It is soluble in alcohol or
ether, and slightly soluble in water. (Int. Teh. Inf. Inst., 1978).
Cyanogen chloride is used as a fumigant, metal cleaner, in ore refin-
ing, production of synthetic rubber and in chemical synthesis (Arena,
1974). Cyanogen chloride can be used in the military as a poison gas.
II. EXPOSURE
The major sources of exposure to cyanogen chloride are in the above
mentioned industrial uses. The potentiality of cyanogen chloride as a water
pollutant has not been described in the available literature.
III. PHARMACOKINETICS
The toxicity of cyanogen chloride resides very largely on its pharmaco-
kinetic property of yielding readily to hydrocyanic acid (also called hydro-
gen cyanide or prussic acid) HI vivo. The red cells of whole blood rapidly
convert cyanogen chloride to cyanide, while serum destroys cyanogen chloride
without forming hydrocyanic acid (Aldridge and Evans, 1946).
Reference should be made to the EPA/ECAO Hazard Profile for cyanides
(U.S. EPA, 1979) for a general discussion on absorption, distribution,
metabolism and excretion. Cyanogen chloride, like HCN, is metabolically
converted to thiocyanate (HCNS) (Aldridge and Evans, 19.46).
-------
IV. EFFECTS
A. Carcinogenicity, Mutagenicity, Teratogenicity,-and Other
Reproductive Effects
Pertinent information could not be located in the available liter-
ature.
B. Chronic Toxicity
Inhaling small amounts of cyanogen chloride causes dizziness,
weakness, congestion of the lungs, hoarseness, conjunctivitis, loss of appe-
tite, weight loss, and mental deterioration. These effects are similar to
those found from inhalation of cyanide (Dreisbach, 1977). Cyanogen chloride
is an irritant to both eyes and throat (Int. Tech. Inf. Inst., 1978).
Cyanogen chloride acts as a chemical asphyxiant, releasing cyanide
which causes internal asphyxia by inhibiting cellular respiration. Cyano-
hemoglcbin may also be formed slowly, but the toxicity is mainly due to the
inhibition of cytochrome oxidase, an enzyme which utilizes molecular oxygen
for cell respiration (Oreisback, 1977).
C. Acute Toxicity
Ingestion or inhalation of a lethal dose of cyanogen chloride
CLD^p = 13 mg/kg), as for cyanide or other cyanogenic compounds, causes
dizziness, rapid respiration, vomiting, flushing, headache, drowsiness, drop
in blood pressure, rapid pulse, unconsciousness, convulsions with death oc-
curring within 4 hours (Dreisbach, 1977).
By subcutaneous route, the LDLQ for cyanogen chloride are as
follows: mouse, 39 mg/kg; dog, 5 mg/kg; and rabbit, 20 mg/kg. By inhala-
-•
tion, an LCLO j.n the dog was found to be 79 ppm/8 hours. Also by inhala-
tion, the LC5Qfs in terms of ppm for 30 minute exposures are: rat, 118;
mouse, 177; rabbit, 207; and guinea pig, 207 (Int. Teh. Inf. Inst., 1978).
-7/0-
-------
V. AQUATIC TOXICITY
Pertinent information could not be found in the .available literature
pertaining to the toxic effects of cyanogen chloride to aquatic organisms.
The reader is referred to EPA/ECAO Hazard Profile for cyanides (U.S. EPA,
1979).
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
Threshold limit values for cyanogen chloride have been set at 0.3
ppm and 0.6 mg/m3 for an 8-hour workday. (ACGIH, 1979).
1
111-
-------
CYANOGEN CHLORIDE
REFERENCES
Aldridge, W.N. and Evans, C.L, 1946. The physiological effects and fate of
cyanogen chloride. Quart. Jour. Expl. Physiol. 33: 241.
American Conference of Governmental Industrial Hygienists. 1979. Thres-
hold-limit-values for chemical substances and physical agents in the work-
room environment for 1979. Cincinnati, Ohio.
Arena, J.M. 1974. Poisoning. Clark C. Thomas Company. Springfield,
Illinois, p. 210.
Deischmann, W.B. and Gerarde, H.W. 1969. Toxicology of drugs and chem-
icals. Academic Press, New York, p. 641.
Dreisbach, R.H. - 1974. Handbook of Poisoning, IX edition. Lange Medical
Publications, Los Altos, California, p. 221.
International Technical Information Institute. 1978. Toxic and hazardous
chemicals safety manual. Tokyo, Japan, p. 142.
U.S. EPA. 1979. Environmental Criteria and Assessment Office. Cyanides:
Hazard profile. (Draft).
-7/2-
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No. 58
ODD
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
-7/S-
-------
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/H-
-------
Disclaimer Notice
Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
-7/5"-
-------
ODD
Summary
ODD can exist in two forms, the o,p'- or the p,p'-isomers. p,p'-DDD
[l,l-(2,2-dichloroethylidlene)-bis-4-chlorobenzene] is a contaminant
(^0.3%) of commerical preparations of DDT [l,l'-(2,2,2-trichloroethyli-
dene)-bis-A-chlorobenzene] as well as being a metabolite of DDT. It has
also been used as an insecticide in its own right under the names TDE or
Rhothane. p,p'-ODO is the first metabolite of p,p'-ODT leading to the even-
tual elimination of p,p'-DDT from the body as p,p'-DDA [2,2-bisU-chlorophe-
nyl) acetic acid]. The residency time of ODD in the body is relatively
short. There is some evidence that ODD is carcinogenic in mice; however, in
other species, it appears to be non-carcinogenic. p,p'-DDD has been shown
to be mutagenic in Drgggpjula, but not in yeast or bacteria. In cell cul-
ture, p,p'-DDD causes chromosomal breaks.
The only available p,p'-ODD toxicity data involves saltwater fish and
invertebrates and freshwater invertebrates. The 96-hour LC^n value for
two invertebrates and three fish range from 1.6 to 42.0 ug/1. p,p'-ODD
appears to be one-fifth to one-seventh as acutely toxic as p,p'-DDT.
-------
ODD
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for DOT (U.S. EPA, 1979a).
ODD is a contaminant of technical p,p'-DDT [l,l'-(2,2,2-trichloroethyl-
idene)-bis-4-chlorobenzene]. It has also been utilized as an insecticide in
its own right under the common names TOE or Rhothane. Its two isomers,
p,p'-DDO [l,l'-(2,2-dichloroethylidene)-bis-4-chlorobenzene] and o,p'-ODD,
make up approximately 0.3 and 0.1 percent, respectively, of technical DDT.
Between 1970 and 1973 (the EPA banned DDT in 1972), a significant drop in
residues of DOO and DDT occurred in the U.S.A., constituting decreases of 89
and 86 percent, respectively.
II. EXPOSURE
Little information is available on exposure to ODD, although the gener-
al exposure pattern probably follows that of DDT, as outlined in DDT: Hazard
Profile (U.S. EPA, 1979b). DOD appears to be disappearing from the U.S.
environment at approximately the same rate as DDT as a result of the 1972
ban on DDT (U.S. EPA, 1975). Wessel (1972) calculated the daily intake of
*
p,p'-ODD to be 0.012 mg/man/day; this was about half the daily intake of
p,p'-DDT.
III. PHARMACOKINETICS
A. Absorption
Pertinent data could not be located in the available literature.
B. - Distribution
The distribution of ODD is the same as that described for DDT in
OPT: Hazard Profile (U.S. EPA, 1979b). The human adipose storage of ODD is
less than that of either DDT or ODE [l,l'-(2,2-dichloroethenylidene)-bis-4-
chlorobenzene].
-------
C. Metabolism
p,p''-DDD is the first metabolite in the multistep pathway of con-
verting p,p'-DDT to p,p'-DDA [2,2-bis(4-chlorophenyl)-acetic acid], the
metabolite which is eventually excreted by rats and by man (Peterson and
Robinson (1964). Urinary p,p'-DDA excretion and serum ODD concentrations
showed increases with DDT dosage in man and declined after dosing ended
(Morgan and Roan, 1977). The enzymes for converting p,p'-DDT to p,p'-DDD
are present in all tissues, while the enzymes for further metabolism of ODD
appear to be absent in brain, heart, pancreas, and muscle of rats (Fang, et
al. 1977).
D. Excretion
Doses of o,p'-ODD yield o,p'-DDA and ring hydroxylation products of
o,p'-DDA in the urine and feces of rats in addition to serine and glycine
conjugates in urine (Reif and Sinsheimer, 1975),
DDO is further metabolized to ODA, which is excreted in the urine
(U.S. EPA, 1979a).
IV. EFFECTS
A. Carcinogenicity
Only two studies have been performed to assess the carcinogenicity
of p,p'-ODD. In a lifespan study, CF1 mice were fed 37.5 mg/kg/day ODD in
their diet (Tomatis, et al. 1974). ODD-exposed animals showed slight in-
creases in liver tumors in males only, but lung adenomas were markedly in-
creased in both sexes. In a National Cancer Institute study (1978), Osborne-
Mendel rats and B6C3F1 mice were dosed with p,p'-DDQ. for 78 weeks. In rats,
DDD had no carcinogenic effects in the females, (43 or 35 mg/kg/day), but
caused a significant increase of follicular cell adenomas in the low dose
males (82 mg/kg/day). Carcinomas of the thyroid were also observed. Be-
t
-71%-
-------
cause of high variation of thyroid lesions in control male rats, these find-
ings are considered only suggestive of a chemical-related effect. In mice,
p,p'-ODD was not carcinogenic.
8. Mutagenicity
p,p'-DDD has been shown to be non-mutagenic in E. coli Pol-A
strains (Fluck, et al. 1976) and Escherichia marcescens (Fahrig,'1974). The
only positive result found in any of the bacterial test systems was reported
by Buselmaier, et al. (1972) upon the administration of p,p'-ODO to mice and
assaying for back mutation of Salmonella typhiniurium and E. marcescens fol-
lowing incubation in the peritoneum in the host-mediated assay. Yeast host
mediated assays using Saccharomyces cerevisiae were negative (Fahrig, 1974),
along with an X-linked recessive lethal assay in Draspnila melanogaster
(Vogel, 1972). In mammalian systems, the mutagenic activity of p,p'-ODD is
relatively weak. This is evidenced by the fact that, depending upon the
dose and route of administration and the species sensitivity of the test
organism, reported studies are negative or marginally positive (U.S. EPA,
1979a). Some chromosomal aberrations and inhibition of proliferation have
been observed with p,p'-000 in cell culture (Palmer, et al. 1972; Mahr and
Miltenburger, 1976). The o,p'-isomer is less active with regard to chromo-
some damage (Palmer, et al. 1972).
C. Teratogenicity, Other Reproductive Effects, and Chronic Toxicity
Pertinent data could not be located in the available literature.
0. Other Relevant Information
Since ODD is a metabolite of DDT, as well as a contaminant of com-
mercial preparations of DOT, many of the effects of DDT could be mediated
»
through ODD. Information on DDT is presented in DDT: Hazard Profile (U.S.
EPA, 1979b).
-7/9-
-------
V. AQUATIC TOXICITY
A. Acute Toxicity
The most insensitive freshwater invertebrate was the scud, Gammarus
lacustris, with a 96-hr. LC5Q static value of 0.60 pg/L (Sanders, 1969).
Of the Cladoceran, the Daphnia pulex species was the most sensitive with a
static LC50 of 3.2 pg/1, while the Simocephalus serrulatus was the least
sensitive with a LC5Q of 5.2 pg/1 (Sanders and Cope, 1966). p,p'-DDD
toxicity has been investigated for several saltwater species. LC5Q values
for two invertebrates, the Eastern oyster, Crassostrea virqinica, and the
Korean shrimp, Palaemon macrodactylus (Schoettger, 1970),. are 25 pg/1 and
1.6 jug/1, respectively, in 96-hr flow-through exposures. In flow-through
exposures to three species of saltwater fish, 96-hr LC^g values range from
2.5 to 42 pg/1 for the stripped bass, Morone saxatilis, Korn and Earnest,
1974). Two species, Morone saxatilis (Korn and Earnest, 1974) and Fundulus
similis (U.S. EPA, 1979a), were exposed to both p,p'-DDD and p,p'-DDT under
similar conditions. A comparison of the results indicates that p,p'-ODD is
one-fifth to one-seventh as acutely toxic to these species as is p,p'-DDT.
However, four to five week old tadpoles of the freshwater toad (Bufo wood-
huusei fowleri) were much more sensitive, having 96-hr. LC5g values of
160 pg/1 compared with 1,000 jug/1 for p,p'-DOT, The DOT sensitivity in-
creased with age (Sanders, 1970).
B. Chronic Toxicity, Plant Effects and Residues
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.
ff
-72O-
-------
A. Human
In 1972, the U.S. EPA banned the agricultural use of DDT in the
United States. There are no other specific guidelines or standards for ODD.
However, for the protection of human health with respect to ODD, criteria of
0.98, 0.098, and 0.0098 ng/1 have been proposed for DDT corresponding to
risk levels of 10~5, 10~fi, and 10~7, respectively. If water alone is
consumed, the water concentration should be less than 0,36 pg/1 to keep the
lifetime cancer risk below 10 .
8. Aquatic
The criteria for DDT and its metabolites are proposed for the pro-
tection of aquatic life from the effects of ODD. The 24-hour average for
the protection of freshwater aquatic life is 0.00023 ug/1, not to exceed
0.41 pg/1 at any time. For saltwater aquatic life, the 24-hour average is
0.0067 pg/1, not to exceed 0.021 ug/1 at any time.
-72J-
-------
ODD
REFERENCES
Buselmaier, W., et al. 1972. Comparative investigations on the muta-
genicity of pesticides in mammalian test systems. Eur. Environ. Mutagen
Sec. 2nd Ann. Meet., Ziukovy Castle, 25.
Fahrig, R. 1974. Comparative mutagenicity studies with pesticides. Page
161 J.n: R. Montesano and L. Tomatis, eds. Chemical carcinogenesis essays,
WHO. IARC Sci. Publ. No. 10.
Fang, S.C., et al. 1977. Maternal transfer of l^C-p.p'-DDT via placenta
and milk and its metabolism in infant rats. Arch. Environ. Contant. Toxicol.
5: 427.
Fluck, E.R., et al. 1976. Evaluation of a DMA polymerase-deficient mutant
of E. coli for the rapid detection of carcinogens. Chem. Biol. Inter-
actions 15: 219.
Korn, S. and Earnest, R. 1974. Acute toxicity of twenty insecticides to
striped bass, Mqrone saxatilis. Calif. Fish and Game 60: 128.
Mahr, U. and H.G. Miltenburger. 1976. The effect of insecticides on
Chinese hamster cell cultures. Mutat. Res. 40: 107.
Morgan, O.P. and C.C. Roan. 1977. The metabolism of DDT in man. Essays
Toxicol. 5: 39.
National Cancer Institute. 1978. Bioassays of DOT, TDE and p,p'-DDE for
possible carcinogenicity. Cas No. 50-29-3, 72-54-8, 72-55-9, NCI-CG-TR-131.
U.S. Dept. Health Edu. Welfare.
Palmer, K.A., et al. 1972. Cytogenetic effects of DDT and derivatives of
DDT in a cultured mammalian cell line. Toxicol. Appl. Phamacol. 22: 355.
Peterson, J.E. and W.H. Robison. 1964. Metabolic products of p,p'-DDT in
the rat. Toxicol. Appl. Pharmacol. 6: 321.
Reif, V.D. and J.E. Sinsheimer. 1975. Metabolism of 1-10-chloro-
phenyl)-l-(p-chlorophenyl)-2,2-dichloroethane (o,pl-000) in rats. Drug.
Metals. Oisp. 15.
Sanders, H.O. 1969. Toxicity of Pesticides to .the Crustacean gammarus
lacustris. Bur. Sport Fish Wildl. Tech. Paper. 25: 18.
Sanders, H.O. 1970. .Pesticide toxicities to tadpoles of the western^chorus
frog. Pseudocris triseriata and Fowler's toad, Bufg woodhousei fdwleri.
Copeia No. 2: 246.
Sanders, H.O. and O.B. Cope. 1966. Toxicities of several pesticides to two
species of cladocerans. Trans. Am. Fish Soc. 95: 165.
-------
Schoettger, R.A. 1970. Progress in sport fishery research 1970. Research
Publ. NO. 106. U.S. Oept. Interior.
Tomatis, L., et al. 1974. Effect of long-term exposure to 1,1-dichloro-
2,2-bis(p-chlorophenyl) ethylene, to l,l-dichloro-2,2-bis (p-chlorophenyl)
ethane, and to the two chemicals combined on CF-1 mice. Jour. Natl. Cancer
Inst. 52: 883.
U.S. EPA. 1975. Preliminary assessment of suspected carcinogens in drink-
ing water. Interim report to Congress, U.S. Environ. Prot. Agency, Washing-
ton, o.c.
U.S. EPA. 1979a. DOT: Ambient Water Quality Criteria. (Draft).
U.S. EPA. 1979b. Environmental Criteria and Assessment Office. DOT:
Hazard Profile. (Draft).
Vogel, E. 1972. Mutagenitatsuntersuchungen mit DDT und den DDT-metaboliten
ODE, ODD, DOOM and DDA an Drosphila melanogaster. Mutat. Res. 16: 157.
Wessel, J.R. 1972. Pesticide residues in foods. Environmental contami-
nants in foods. Spec. rep. No. 9. N.Y. State Agric. Exp. Sta., Geneva.
-------
No. 59
DDE
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.
-------
DDE
Summary
DOE exists as two isomers, o,p'- and p,p'-DDE [!,!'-(2,2-dichloroeth-
enylidene)-bis-4-ehlorobenzene] is the major contaminant (ca. 4 percent) of
commercial preparations of p,p'-DDT [!,!'-(2,2,2-trichloroethylidene)- bis-
4-chlorobenzene], as well as being a metabolite of p,p'-DDT. p,p'-DDE is a
highly lipophilic compound which undergoes no further metabolism. Its resi-
dency time in the body is extremely long. p,p'-DDE has been shown to be
carcinogenic in mice but not in rats. In cell culture it causes chromosomal
breaks.
The only aquatic toxicity data available on p,p'-DDE involve acute tox-
ic flow-through exposures to two saltwater invertebrates. The 48-hr. LC5g
for a shrimp is 28 jjg/1; the 96-hr. LCL- for the Eastern oyster is 14 pg/1.
-------
DDE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for DDT and metabolites (U.S. EPA, 1979a).
DDE is a contaminant of technical l,l'-(2,2,2-trichloroethyiidene)-bis-
4-chlorobenzene (DDT). Its two isomers, p,p'-DDE [l,lt-(2,2-dichioroetheny-
lidene)-bis-4-chlorobenzene] and o,p'-DDE make up approximately 4.0 and 0.1
percent, respectively, of technical grade DDT. Between 1970 and 1973 (the
EPA banned DDT in 1972), a significant drop in the residues of DDT in the
U.S. occurred, constituting a .decrease of 86 percent. However, during this
time period, residues of DDE decreased only 27 percent. In fact, p,p'-DDE
residues comprise most of the biological residues (ca. 71 percent) arising
from DDT application (U.S. EPA, 1979a; Kveseth, et al. 1979).
II. EXPOSURE
Little information is available on exposure to DDE, although the gener-
al exposure pattern probably follows that of DDT, as outlined in DDT: Haz-
ard Profile (U.S. EPA, 1979b). DDE residues appear to be disappearing from
the environment at a slower rate than DDT following the 1972 ban on DDT
(U.S. EPA, 1975). Wessel (1972) calculated the daily dietary intake of
p,p'-DDE to be 0.018 mg/man/day, as compared with a value of 0.027 mg/man/
day for DDT. A recent study by de Campos and Olszyne-Marzys (1979) based on
studies in Latin American countries still using DDT indicates that human
milk contains more p,p'-DDE than p,p'-ODT (up to 3 pg/1 whole milk) in every
sample taken.
-72.7-
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III. PHARMACOKINETICS
A. Absorption
DDE is absorbed from the gastrointestinal tract with high efficien-
cy characteristic of dietary fat. Maximum lipid solubilities reach 100,000
ppm.
B. Distribution
The distribution of DDE is similar to that described for DDT in the
EPA/ECAO Hazard Profile on DDT (U.S. EPA, 1979b). Serum and adipose concen-
trations of p,p'-DDE rise slower than DDT, with the'' peak some months in fol-
lowing termination of dosing. The human adipose storage of p,p'-DDE is
greater than that for DDT, and p,p'-DDE is eliminated from the body very
slowly. This is also true for the Rhesus monkey (Durham, et al. 1963).
Storage loss data predict that, if dietary intake were eliminated, it would
take an entire lifespan to eliminate the average human body burden .of p,p'-
DDE. It has been shown that tissue storages of p,p'-DDE in the general pop-
ulation originate almost entirely from dietary p,p'-DDE rather than DDT con-
version (U.S. EPA, 1979a). However, this may not be the case for p,p'-DDE
residues in human milk (de Campos and Olszyne-Marzys, 1979).
C. Metabolism
The end product of the metabolism of DDT which proceeds via reduc-
tive dehydrochlorination is p,p'-DDE. In addition, p,p'-DDE is the major
storage product of DDT in animals [apart from hamsters (Agthe, et al. 1970)]
and humans. The enzymes for metabolizing DDT to p,p'-ODE are present in all
tissues (Fang, et al. 1977).
In humans given p,p'-DDT orally, no more than one-fifth of.the
absorbed DDT ultimately undergoes conversion to p,p'-DDE (Morgan and Roan,
1977). p,p'-DOE does not undergo further metabolism to 2,2-bis(4-chloro-
phenyD-acetic acid (DDA), the urinary excretion product of DDT.
t
-------
D, Excretion
Excretion of p,p'-DDE has not been demonstrated in man. In mice,
p,p'-DDE is excreted in the urine (Wallcave, et al. 1974). The o,p'-isomer
is more easily excreted than the p,p'-isomer (Morgan and Roan, 1977).
IV. EFFECTS
A. Carcinogenicity
Only two studies have been performed to assess the carcinogenicity
of p,p'-ODE. In a lifespan study, CF-1 mice were fed 37.5 mg/kg/day p,p'-
DOE in their diet (Tomatis, et al. 1974). p,p^-DDE increased liver tumor
incidence from 1 percent in controls to 90 percent in treated female
animals, and from 34 to 74 percent in male animals. The combination
p,p'-DDE/ODD produced more tumors than either constituent alone at the same
concentration in the combination. In a National Cancer Institute study
(1978), Qsborne-Mendel rats and B6C3F1 mice were dosed with p,p'-DDE for 78
weeks. In rats, p,p'- DDE had no carcinogenic effect on either females (22
mg/kg/day) or males (42 mg/kg/ day), although hepatotoxicity was evident.
In mice, hepatocellular carcinomas were significantly increased in the
animals fed p,p'-ODE (22 and 39 mg/kg/day for females and males,
respectively).
B. Mutagenicity .
p,p'-DDE has been shown to be nonmutagenic in §_._ coli Pol-A strains
(Fluck, et al. 1976), Escherichia marcescerjs (Fahrig, 1974), and in the host
mediated assay using Salmonella typhimurium and §_._ marcescens (Buselmaier,
et al. 1972) and Saccharomyces cerevisiae (Fahrig, -1974). Vogel (1972) mea-
sured X-linked recessive lethal mutations in Drosophila melanogaster and
»
found no activity for p,p'-DDE. In mammalian systems, the mutagenic activi-
ty of p,p'-DOE is relatively weak. This is evidenced by the fact that, de-
-------
pending upon the dose and route of administration and the species sensitivi-
ty of the test organism, reported studies are negative or marginally posi-
tive (U.S. EPA, 1979a). Some chromosomal aberrations and inhibition of pro-
liferation have been observed with p,p'-DDE in cell culture (Palmer, et al.
1972; Mahr and Miltenburger, 1976). The o,p'-isomer causes fewer chromosom-
al aberrations (Palmer, et al. 1972).
C. Teratogenicity, Other Reproductive Effects and Chronic Toxicity
Pertinent information could not be located in the available litera-
ture.
D. Other Relevant Information
Since p,p'-DDE is a metabolite of DDT, as well as a contaminant of
commercial preparations of DDT, many of the effects of DDT could be mediated
through p,p'-DDE. Information on DOT is presented in DDT: Hazard Profile
(U.S. EPA, 1979b). Oral acute LD values for p,p'-DDE in rat are 380
mg/kg for males but 1,240 mg/kg for females (Hayes, et al. 1965).
V. AQUATIC TOXICITY
A. Acute Toxicity
The 96-hr. LC^ value- for p,p'-DOE for the comparatively resis-
tant freshwater planarian (Polycelis felina) was 1,050 pg/1 (Kouyoumjian and
Uglow, 1974). The acute toxicity of p,p'-DDE has also been investigated in
two saltwater invertebrates. The 48-hr. l_C,-n for the brown shrimp, Penae-
us aztecus, was 28 jjg/1; the 96-hr. LCg_ for the Eastern oyster, Crassos-
trea virqinica. was 14 jug/1 (U.S. EPA, 1979a). Both studies were flow-
through exposures.
B. Chronic Toxicity and Plant Effects
»
Pertinent data could not be located in the available literature.
X
-730-
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C. Residues
p,p'-DDE is a major metabolite of DDT in aquatic ecosystems. One
study involving bird eggshells and DDT showed p,p'-DDE to comprise 62 per-
cent of the DDT metabolites (U.S. EPA, 1979a). Average residues in egg-
shells of the great black-backed gull ranged from 14 to 68 ng/g of lipid
(Cooke, 1979). p,p'-ODE in fat and muscle of the white-faced ibis in 1974/
75 were as high as 65 ng/g lipid (Capen and Leiker, 1979). No studies are
available, however, involving p,p'-DDE specifically.
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
In 1972, the U.S. EPA banned the agricultural use of DDT in the
United States. There are no other specific guidelines or standards for
DDE. However, for the protection of human health with respect to ODE,
criteria of 0.98, 0.098 and 0.0098 ng/1 have been proposed for DDT corres-
ponding to risk levels of 10~ , 10~ , and 10" , respectively. If
water alone is consumed, the water concentration should be less than 0.36
pg/1 to keep the lifetime cancer risk-below 10~ .
B. Aquatic
The criteria for DDT and its metabolites are proposed for the
protection of aquatic life from the effects of DDE. -• The 24-hour average for
the protection of freshwater aquatic life is 0.00023 jjg/1, not to exceed
*
0.41 ug/1 at any time. For saltwater aquatic life, the 24-hour average is
0.0067 ug/1, not to exceed 0.021 ;jg/l at any time.
-------
DDE
REFERENCES
Agthe, C., et al. 1970. Study of the potential carcinogenicity of DDT in
the Syrian Golden Hamster. Proc. Soc. Biol. Med. 134: 113.
Buselmaier, W., et al. 1972. Comparative investigations on the mutagenici-
ty of pesticides in mammalian test systems. Eur. Environ. Mutagen Soc. 2nd
Ann. Meet., Ziukovy Castle, 25. .
de Campos, M. and D.E. Qlszyna-Marzys. 1979. Contamination of human milk
with chlorinated pesticides in Guatamala and in El Salvador. Arch. Environ.
Contam. Toxicol. 8: 43.
Capen, D.E. and T.J. Leiker. 1979. DDE residues in blood and other tissues
of white-faced ivis. Environ. Pollut. 19: 163.
Cooke, A.S. • 1979. Eggshell characteristics of gannets (Sula bassoud),
shaps (Phalacroc_o_r_ax aristotelis) and great black-packed gulls (Carus marin-
us) exposed to DDE and other environmental pollutants. Environ. Pollut.
19: 47.
Durham, W.F., et al. 1963, The effect of various dietary levels of DDT on
liver function, cell morphology and DDT storage in the Rhesus monkey. Arch.
Int. Pharmacodyn. Ther. 141: ill.
Fahrig, R. 1974. Comparative mutagenicity studies with pesticides. Page
161 .In: Montesano and L. Tomatis, (eds). Chemical carcinogenesis essays,
WHO. IARC Sci. Publ. No. 10.
Fang, S.C., et al. 1977. Maternal transfer of 1AC-p,p'-DDT via placenta
and milk and its metabolism in infant rats. Arch. Environ. Contam. Toxicol.
5: 427.
Fluck, E.R., et al. 1976. Evaluation of a DNA polymerase-deficient mutant
°f §• coli for the rapid detection of carcinogens. Chem. Biol. Interac-
tions. 15: 219.
Hayes, W.J., Jr., et al. 1965. Chlorinated hydrocarbon pesticides in the
fat of people in New Orleans. Life Sci. 4: 1611,
Kouyoumjian, H.H. and R.F. Uglow. 1974. Some aspects of the toxicity of
pjpi-DDT, p,pl-ODE and p,pl-DDD to the freshwater planarian Polycelis
felina (Tricladida). Environ. Pollut. 7: 103.
•^^^^^ s
Kveseth, N.3., et al. 1979. Residues of DDT in a Norwegian fruit growing
district two and four years after termination of DDT usage. Arch. Environ.
(Contam. Toxicol.). 8: 201.
Mahr, U. and H.G. Miltenburger. 1976. The effect of insecticides on Chi-
nese hamster cell cultures. Mutat. Res. 40: 107.
-732.-
-------
Morgan, D.P. and C.C. Roan. 1977. The metabolism of DDT in man. Essays
Toxicol. 5: 39.
National Cancer Institute. 1978. Bioassays of DDT, TDE and p,p'-QDE for
possible carcinogenicity. NCI-CG-TR-131. U.S. Dep. Health Edu. Welfare.
Palmer, K.A., et al. 1972. Cytogenetic effects of DDT and derivatives of
DDT in a cultured mammalian cell line. Toxicol. Appl. Pharmacol. 22: 355.
Tomatis, L., et al. 1974. Effect of long-term exposure to 1,1-dichloro-
2,2-bis(p-chlorophenyl) ethylene, to l,l-dichloro-2,2-bis (p-chlorophenyl)
ethane, and to the two chemicals combined on CF-1 mice. Jour. Natl. Cancer
Inst. 52: 883.
U.S. EPA. 1975. DDT. A review of scientific and economic aspects of the
decision to ban its use as a pesticide. EPA52Q/1-75-Q22. U.S. Environ.
Prot. Agency, Washington, D.C.
U.S. EPA. 1979a. DDT: Ambient Water Quality Criteria. (Draft).
U.S. EPA. 19795. Environmental Criteria and Assessment Office. DDT: Haz-
ard Profile (Draft).
Vogel, E. 1972. Mutagenitatsuntersuchungen mit DDT und den DDT-metaboliten
DDE, ODD, QDOM and DDA. an Drosphila melanogaster. Mutat. Res. 16: 157.
Wallcave, L., et al. 1974. Excreted metabolites of l,l,l-trichloro-2,2-bis
(p-chlorophenyl) ethane in the mouse and hamster. Agric. Food Chem.
22: 904.
Wessel, J.R. 1972. Pesticide residues in foods. Environmental contami-
nants in foods. Spec. Rep. No. 9. N.Y. State Agric. Exp. Sta., Geneva.
-733-
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No. 60
DDT
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-uracy.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (GAG) has evaluated
DDT and has found sufficient evidence to indicate that this
compound is carcinogenic.
-------
DDT
Summary
.The most commonly used DDT was a technical formulation and usually con-
sisted of a mixture of p,p'-DDT (77.1 percent), o,p'-DDT (14 percent), p,p'-
DDD (0.3 percent), o,p'-DDD (0.1 percent), p,p'-DDE (4 percent), c,p'-DDE
(0.1 percent and 3.5 percent unidentified compounds. Pure DDT is the p,p'-
isomer [l,l'-(2,2,2-trichloraethylidene)-bis-4-chlorobenzene]. Unless spe-
cifically identified, the term DDT will refer tcr- the pure form. Prior to
being banned in the U.S. in 1972, DDT was used extensively as a pesticide.
Due to the high lipid solubility of DDT, it has a long residency time
in the body. DDT has produced adverse reproductive effects in rodents and
birds, but adverse effects have not been noted in man. The lowest acute
oral LD5Q value was found for the dog (60-75 mg/kg). There is suggestive
evidence that DDT might be a carcinogen, and weak evidence that it might be
a teratogen. Chromosomal breaks have been observed with DDT exposure ir\_
vitro and in vivo.
DDT is acutely toxic to freshwater fish at concentrations as low as 0.8
;ug/l and to invertebrates at 0.18 ;ug/l. Chronic toxicity has been manifest-
ed in the fathead minnow in the range of 0.37 to 1.48 pg/1. A weighted
average bioconcentration factor of 39,000 has been estimated for DDT for
consumed fish and shellfish. For saltwater fish and invertebrates, DDT con-
centrations as low as 0.2 jug/1 and 0.14 /jg/1, respectively, have been re-
ported to be acutely toxic. Chronic toxicity data--for saltwater organisms
are not available.
-737-
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DDT
I. INTRODUCTION
This profile is based primarily on the Ambient Water Quality Criteria
Document for DDT (U.S. EPA, 1979a).
DDT has been used extensively world-wide for public health and agricul-
tural programs as a broad spectrum insecticide. It has played a large role
in the world-wide control of the malaria mosquito. In 1972, following an
extensive review of health and environmental hazards of the use of DDT, the
U.S. EPA decided to ban any further use of DDT. .. Prior to this, technical
grade DDT had been widely used in the U.S., with a peak usage in 1959 of 80
million pounds. This amount decreased steadily to less than 12 million
pounds by 1972. Since the 1972 ban, the use of DDT in the U.S. has been ef-
fectively discontinued. However, technical grade DDT is still used in many
other countries and widespread contamination still occurs. Since ODD and
DDE are also metabolites of DDT, it is sometimes difficult to separate con-
tamination from metabolic accumulation. The compounds of DDT are extremely
persistent and are so widespread that levels as high as 15 ppb have been de-
tected in feed for laboratory animals (Coleman and Tardiff, 1979).
II. EXPOSURE
The primary route of human .exposure to DDT is from ingestion of small
amounts in the diet. Biological magnification of DDT in the food chains oc-
curs by two routes: (1) direct absorption from contaminated water by aquat-
ic organisms; (2) transfer of residues through sequential predator feeding.
Meats, fish, poultry, and dairy products are the primary sources of DDT
residues in the human diet. The U.S. EPA (1979a) has estimated the weighted
•
average bioconcentration factor of DDT at 39,000 for consumed fish and
shellfish. Due to the banned usage of DDT in the U.S., there has been a
-73S-
-------
continual decline in the DDT residue in food. These decreases are reflected
in the changing amounts of estimated dietary intake: 1965 - 0.062 rag/man/
day; 1970 - 0.024 mg/man/day; 1973 - 0.008 mg/man/day (U.S. EPA, 1975).
Levels of DDT- found in the air are far below levels that add significantly
to total human intake. Stanley, et al. (1971) sampled air in nine locali-
ties, and found DDT in the ranges of 1 ng/m to 2520 ng/m of air.
Wolfe and Armstrong (1971) showed that industrial workers not wearing respi-
rators could be exposed to significant levels of DDT in the air (up to 34
mg/man/hour}, particularly in the formulating plants. Exposure for agricul-
tural spray operators may be as high as 0.2 mg/man/hour (Wolfe, 1967). Der-
mal exposure for formulators was estimated to range from 5 to 993 mg/man/
hour (Wolfe and Armstrong, 1971). Little DDT was found in the urine, how-
ever. Dermal absorption of DDT is minimal.
Dermal toxicity in rats occurs at 3,000 mg/kg (U.S. EPA, 1979a). Hayes
(1966) estimated the intake of DDT to be in the following proportions: food
- 0.04 iTig/man/day; water - 4.6 x 10" tng/man/day; and air - 9 x 1Q~
mg/man/day. The actual dose for the average man is now estimated to be 0.01
mg/man/day (U.S. EPA, 1979a).
III. PHARMACOKINETICS
A. Absorption
DDT is absorbed from the gastrointestinal tract with efficiency ap-
proaching 95 percent when ingested with dietary fat. In humans, Morgan and
Roan (1971) showed that absorption of an oral dose of 20 mg DDT proceeded
faster than transport out of the vascular compartment into tissue storage.
Studies concerning the kinetics of absorption of DDT via inhalation or der-
*
mal routes were not found in the available literature.
-------
8. Distribution
DDT has been found in virtually all body tissues, approximately in
proportion to respective tissue content of extractable lipid. The adipose/
blood ratios of DDT have been recently estimated to be approximately 280:1
(Morgan and Roan, 1977). DOT concentrations in body tissues were highest
for fat tissue, followed by reproductive organs, the liver and kidney to-
gether, with lowest concentrations found in the brain (Tomatis, et al.
1971). Elimination of very low levels of DDT from storage proceeds much
more slowly than that of the large stores of DDT ""accumulated by occupation-
ally exposed workers or dosed volunteers (Morgan and Roan, 1971). The aver-
age North American adult, with 17 kg of body fat, contains approximately 25
mg of DDT. It is predicted from storage loss data that, if dietary -intake
were eliminated, most of the DDT would be lost within one or two decades
(U.S. EPA, 1979a). Trace metals in the diet, particularly cadmium, may af-
fect the mobilization of DDT in tissues (Ando, 1979).
C. Metabolism
The metabolism of DDT in man appears to be the same as the pathways
reported by Peterson and Robison (1964) for the mouse. Generally, two sepa-
rate reductive pathways produce the primary endpoint metabolites, p,p'-DDE
and p,p'-DDA. The predominant conversion is of DDT to p,p'-DDD via dechlor-
ination. This is the first product in a series which results in metabolites
which are later excreted. The other primary pathway proceeds via reductive
dehydrochlorination which results in the formation of p,p'-DDE the major
storage product in animals and humans. Fant, et al. (1977) suggest that
enzymatic activity for the dehydrochlorination and reductive dechlorination
reactions transforming DDT to ODD and DDE is present in all tissues, whereas
the enzymes involved in the hydrogenation and hydroxylation steps changing
-7VO-
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ODD to ODA are absent in the brain, heart, pancreas, and muscle of the rat.
Metabolic conversion of DDT to DDA proceeds more rapidly than conversion to
the storage metabolite of DDE. For additional information regarding the DDT
metabolites ODD and ODE, the reader is referred to the Hazard Profile for
those chemicals (U.S. EPA, 1979b,c).
D. Excretion
The excretion of DDT was investigated in human volunteer studies of
Hayes, et al. (1971) and Roan, et al. (1971). Urinary excretion predominat-
*.
ed, with 13 to 16 percent of the daily dose being excreted as p,p'-DDA, and
was shown to correlate with exposure levels of individuals working in a for-
mulating plant (Ortelee, 1953). p,p'-DDE and DDT are the predominant com-
pounds excreted and p,p'-DDD and p,p'-DDA are excreted in the least amounts
(Morgan and Roan, 1977). p,p'-DDE was found in slightly higher concentra-
tions in exposed workers versus the general population. Gut microorganisms
have demonstrated a capacity for degradation of DDT to p,pf-DDD and p,p'-DDA.
IV. EFFECTS
A. Carcinogeniity
Lifetime and multigeneration exposures to DDT in the diet of rats,
mice, and fish have produced significant increases in the formation of a
number of tumor types (U.S. EPA, 1979a). The predominant lesion appears to
be hepatoma. Also, Tomatis, et al. (1974) demonstrated that short-term ex-
posure to technical grade DDT (37.5 mg/kg/day for 15 or 30 weeks), using
CF-1 mice, resulted in an increased incidence and early appearance of hepa-
tomas, similar to that caused by lifespan exposure. Mice appear much more
susceptible than rats (U.S. EPA, 1979a) and the use of the mouse as an ani-
mal model for humans has been criticized (Deichmann, 1972). In these stud-
ies contaminants p,p'-DDD and p,p'-DDE were present, both of which have "pro-
-7N1-
-------
duced liver tumors in CF-1 mice (Tomatis, et al. 1974). Also, the combina-
tion of p,p'-ODD/DDE was found to produce more tumors than and equal concen-
tration of either compound alone. Tarjan and Kemeny (1969) noted leukemias
and pulmonary carcinomas in Bald-C mice fed 3 ppm DDT in the diet. Hepato-
mas have been observed in rainbow trout (Halver, et al. 1962).
A number of other studies have shown no significant increase in
tumor formation following DDT exposure. Lifetime feeding studies with Syri-
an Golden Hamsters (Agthe, et al. 1970) and a number of long term feeding
studies with various strains of rats have shown -no significant increase in
tumor incidence (Cameron and Cheng, 1951; Fitzhugh and Nelson, 1947; Radom-
ski, et al. 1965; Deichmann, et al. 1967). In a 78-week National Cancer In-
stitute study (1978), Osborne-Mendel rats given 16 and 32 mg/kg/day (males)
or 11 and 21 mg/kg/day (females) showed no tumors. B6C3F1 mice given 3.3
and 6.6 mg/kg/day (males) or 13 and 26 mg/kg/day (females) also showed no
tumor development. Durham, et al. (1963) found no liver pathology in Rhesus
monkeys fed 100 mg/kg/day or less DDT for up to 7.5 years. At the present
time, no evidence of neoplasia has been found in the studies performed in
occupationally exposed or dosed volunteer subjects (U.S. EPA, 1979a).
B. Mutagenicity
DDT has not shown mutagenic activity in any of the bacterial test
systems thus far studied: Salmonella typhimurium (McCann, et al. 1975; Mar-
shall, et al. 1976); §_._ coli Pol-A strains (Fluck, et al. 1976); Bacillus
subtilis (Shirasu, et al. 1976). Tests on eukaryotic yeast cells have been
uniformly negative, with Fahrig (1974) using Saccharomyces cerevisiae and
Clark (1974) using Neurospora crassa. Vogel (1972) and Clark (1974) found
»
positive mutagenic activity in Prosophila melanoqaster by measuring x-linked
recessive lethal mutations. In mammalian systems, the mutagenic activity of
-742.-
-------
DDT is relatively weak. This is evidenced by the fact that, depending upon
the dose and route of administration and the species sensitivity of the test
organisms, reported studies are negative or only marginally positive (U.S.
EPA, 1979a). In vivo and in vitro cytogenetic studies seem to indicate that
DDT is a clastogenic (chromosome breaking) substance. The metabolites p,p'-
DDE, p,p'-DDD, p,p'-DDA and p,p'-DDOH were also non-mutagenic except possi-
bly for p,p'-DDD (U.S. EPA, 1979a). Chromosomal aberrations in cell lines
of the kangaroo rat occurred more often with p,p'-isomers than o,p'-isomers
(Palmer, et al. 1972).
C. Teratogenicity
Only minimal teratogenic effects have been reported following high
dosages of DDT. Sprague-Dawley rats receiving 200 ppm DDT in their diet
showed a significant increase in ring tail, a constriction of the tail fol-
lowed by amputation, in the offspring (Ottoboni, 1969).
D. Other Reproductive Effects
Hart, et al. (1971) showed that DDT has an effect on prematurity
and causes an increase in the number of fetal resorptions in rabbits given
50 mg/kg on days 7, 8, and 9 of gestation. Chronic exposure (less than 200
mg/kg) of rats and mice produced no adverse effects on survival of the off-
spring (Ware and Good, 1967; Ottoboni, 1969). Krause, et al. (1975) noted a
damaging effect on spermatogenesis in rats following acute exposure to DDT
(7,200 mg/kg). Also, DDT has been shown to possess estrogenic activity in
rodents and birds (Welch, et al. 1969; Bittman, et al. 1968).
E. Chronic Toxicity
A number of pathological changes have been noted in rodents; the
t
most consistent finding in lifetime feeding studies has been an increase in
the size of liver, kidneys, and spleen; extensive degenerative changes in
-------
the liver; and an increased mortality rate .(U.S. EPA, 1979a). In contrast
to the rodent models, Rhesus monkeys fed diets with up to 200 ppm DDT did
not show liver histopathology, decrease in weight gain or food consumption,
or clinical signs of illness (Durham, et al. 1963),
F. Other Relevant Information
DDT is a strong inducer of the mixed function oxidase system; this
could potentially enhance the biological effects of other chemicals by acti-
vation, or diminish their activities through detoxification mechanisms (U.S.
EPA, 1979a). Exposure to DDT has caused enhanced tumor incidence in N-fluor-
enacetamide-treated rats (Weisburger and Weisburger, 1968) and decreased
phenobarbital-induced sleeping times (Conney, 1967). Acute oral LDgQ val-
ues in rats typically range from 100 to 400 mg/kg and 40 to 60 mg/kg i.v.
The oral LD^Q values in other animals are: 60 to 75 mg/kg (dogs); 250 to
400 mg/kg (rabbits); approximately 200 mg/kg (mice). For p,p'-DDE, the val-
ues are 380 and 1,240 mg/kg in male and female rats, respectively; for p,p'-
DDA in rats, the values are 740 and 600 mg/kg, respectively (U.S. EPA,
1979a). Symptoms of DOT poisoning in humans include the following: convul-
sions, parasthesia of extremities and vomiting (at high doses), convulsions
and nausea (less than 16 mg/kg), dizziness, confusion and most characteris-
tically, tremors (Hayes,' 1963). In rats, the liver shews changes at dietary
doses less than 5 ppm (Laug, et al. 1950). No permanent injury to .man from
DDT has been recorded (U.S. EPA, 1979a).
V. AQUATIC TOXICITY
A. Acute Toxicity
The acute toxicity of DDT to freshwater organisms has been well
»
documented. Data are available for 25 species of fish. The 96-hour LC5Q
values are available for the following freshwater fish: rainbow trout (Sal-
-------
mo qairdneri), 1.7 to 42 ug/1; fathead minnow (Pimephal_es promelas), 7.4 to
58 jug/1; channel catfish (Ictalurus punctatus), 16 to 17.5 ug/1; bluegill
(j-egomis macrochirus), 1.2 to 210 ug/1. The most sensitive of fish was the
yellow perch (Perca flavesceus) with a 96-hour LC,-n of 0.6 pg/1 (Marking,
1966). Invertebrate freshwater species are more sensitive than fish. For
Daphnia maqna, 48-hour LC5Q values of 1.48 jug/1 have been reported (Pries-
ter, 1965). One week old crayfish (Oreoneetus nais) had a 96-hour LC50
value of 0.18 pg/1 (Saunders, 1972). LC5Q values for nine saltwater fish
species range from 0.2 to 4.2 jjg/1. Saltwater invertebrates were slightly
more sensitive, with LC5_ values ranging from 0.14 to 10.0 pg/1 (U.S. EPA,
1979a).
Concentrations as low as 8 ug/1 elicited hyperactive locomotor re-
sponses in bluegill (Lepomis macrochirus) over 16 days old (Ellgaard, et al.
1977). The acute LDcn in adult summer frogs (Rana jgmppraria) was only
7.6 mg/kg. Though adipose tissues contained most of the DDT, the ovaries of
females contained as much of the compound as did bones and spleen (Harri, et
al. 1979).
B. Chronic Toxicity
Only one chronic freshwater fish value is available (Pimephales
promelas). indicating that the chronic toxicity value is 0.74 jug/1 (Jarvi-
nen, et al., 1977). Freshwater invertebrate chronic toxicity data are not
available. Concentration of DDT affecting three saltwater invertebrate spe-
cies in chronic studies are similar in LC5Q values (U.S. EPA, 1979a).
C. Plant Effects
Four species of freshwater algae (Calovella sp.) have evidenced a
^«^^^___ ^
wide range of sensitivities, 0.3 to 800 ^jg/1 (Sodergren, 1968). Wurster
(1968) investigated the effects of DDT on four species of marine algae. The
-------
data showed reduced rates of photosynthesis at 10,ug/l, indicating that al-
gae are much less sensitive to DDT than are fish and.invertebrates.
D. Residues
DDT is bioconcentrated to a very high degree in aquatic organisms.
An average bioconcentration factor (BCF) of 640,000 has been calculated from
31 experimental measurements of bioconcentration done on 26 species of
freshwater fish. Individual BCF's ranged from 490 .to 2,236,666. In the
field, BCF factors have been observed which are seven times higher than the
*.
average values derived from laboratory data. This discrepancy may be due to
the many additional trophic levels involved and the possibly higher lip'id
content of the organisms in the field. In saltwater species, the BCF for
DDT ranges from 800 to 76,300 times for fish and shellfish (U.S. EPA,
1979a). The lowest observed allowable maximum tissue concentration was 0.5
>ug/kg for domestic animals in animal feed (U.S. FDA, 1977) and in the brown
pelican (Peiecanus occidentalis) for eggshell thinning (Blus, et al. 1972,
1974).
VI. EXISTING GUIDELINES AND STANDARDS
Neither the human health nor the aquatic criteria derived by U.S. EPA
(1979c), 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 existing guidelines and standards for DDT are:
V
YEAR AGENCY/ORG. STANDARD REMARKS
1971 WHO 0.005 mg/kg Maximum Acceptable Daily
body weight Intake in food
1976 U.S. EPA 0.001 jjg/1 Ambient Water Quality
Criteria
-7V6-
-------
1977 Natl. Acad. Sci.,
"Natl. Res. Counc.
1978 Occup. Safety
Health Admin.
1978 U.S. EPA
0.41 jug/1
0.00023 jug/1
In light of carcinogenic
risk projection, suggested
strict criteria for DDT
and DDE in drinking water
Skin exposure
Final acute and chronic
values for water quality
criteria for protection of
aquatic life (freshwater)
The U.S. EPA (1979a) is in the process of establishing ambient
*.
water quality criteria. Based on the potential carcinogenicity of DDT, cur-
rent draft criteria are calculated on the estimate that 0.98 jug/man/day
would result in an increased additional lifetime cancer risk of no more than
1/100,000. Since man and the rat appear to be less sensitive than mice,
greater levels may be tolerable.
B. Aquatic
For DDT, the proposed draft criterion to protect freshwater aquatic
life is 0.00023jug/1 as a 24-hour average; the concentration should not ex-
ceed 0.41 /jg/1 at any time. For saltwater aquatic species, the concentra-
tion is 0.0067 jug/1 as a 24-hour average and should not exceed 0.021yug/l at
any time (U.S. EPA, 1979a).
-------
DDT
REFERENCES
Agthe, C., et al. 1970. Study of the potential carcinogenicity of DDT in
the Syrian Golden hamster. Proc. Soc. Exp. Biol. Med. 134: 113.
Ando, M. 1978. Transfer of 2,4,5,2',A1,5'-hexachlorobiphenyl and 2,2,-bis-
(p-chlorophenyl)-l,l,l-trichloroethane(p,p'-DDT) from maternal to newborn
and suckling rats. Arch. Toxicol. 41: 179.
Bittman, J., et al. 1968. Estrogenic activity of o,p'-DDT in the mammalian
uterus and avian oviduct. Science 162: 371.
Blus, L.J., et al. 1972. Logarithmic relationship of DDE residues to egg-
shell thinning. Nature 235: 376.
Blus, L.J., et al. 1974. Relations of the brown pelican to certain envi-
ronmental pollutants. Pestic. Monit. Jour. 7: 181.
Cameron, G.R., and K. Cheng. 1951. Failure of oral DDT to induce toxic
changes in rats. Br. Med. Jour. 819.
Clark, J.M. 1974. Mutagenicity of DDT in mice, Drosophila melanoqaster and
Neurospora crssa. Aust. Jour. Biol. Sci. 27: 427.
Coleman, W.E. and R.G. Tardiff. 1979. Contaminant levels in animal feeds
used for toxicity studies. Arch. Environ. Contam. Toxicol. 8: 693.
Conney, A.M. 1967. Pharmacological implications of microsomal enzyme in-
duction. Pharmacol. Rev. 19: 317.
Deichmann, W.8. 1972. The debate on DOT. Arch. Toxicol. 29: 1.
Deicnmann, W.B., et al. 1967. Synergism among oral carcinogens. IV. The
simultaneous feeding of four tumorigens to rats. Toxicol. Appl. Pharmacol.
11: 88.
Durham, W.F., et al. 1963. The effect of various dietary levels of DDT on
liver function, cell morphology and DDT storage in the Rhesus monkey. Arch.
Int. Pharmacokyn. Ther. 141: 111.
Ellgaard, E.G., et al. 1977. Locomotor hyperactivity induced in the blue-
gill sunfish, Lepomis macrgchirus, by sublethal corrections of DDT. Jour.
Zool. 55: 1077.
*'
Fahrig, R. 1974, Comparative mutagenicity studies with pesticides. Page
161 _in R. Montesano and L. Tomatis, eds. Chemical carcinogenesis essays,
WHO. IARC Sci. Publ. No. 10.
Fang, S.C., et al. 1977. Maternal transfer of ^C-p-p'-DDT via placenta
and milk and its metabolism in infant rats. Arch. Environ. Contam. Toxicol.
5: 427.
-------
Palmer, K.A. 1972. Cytogenic effects of DDT and derivatives of DDT in a
cultured mammalian cell line. Toxicol. Appl. 'Pharmacol. 22: 355.
Peterson, J.E. and W.H. Robison. 1964. Metabolic products of p,p'-DDT in
the rat. Toxicol. Appl. Pharmacol. 6: 321.
Priester,' E.L., Jr. 1965. The accumulation and metabolism of DDT, para-
thion, and endrin by aquatic food-chain organisms. Ph.D. Thesis. Clemson
Univ. Clemson, S.C. 74 p.
Radomski, J.L., et al. 1965. Synergism among oral carcinogens. I. Results
of the simultaneous feeding of four tumorigens to rats. Toxicol. Appl.
Pharmacol. 7: 652.
Roan, C., et al. 1971. Urinary excretion of DDA following ingestion of DDT
and DDT metabolites in man. Arch. Environ. Health 22: 309.
Shirasu, V., et al. 1976. Mutagenicity screening of pesticides in the
microbial system. Mutat. Res. 40: 19.
Sodergren, A. 1968. Uptake and accumulation of C14-DDT by Chlorella sp.
(Chlorophyceae) Oikos 19: 126.
Stanley, C.W., et al. 1971. Measurement of atmospheric levels of pesti-
cides. Environ. Sci. Technol. 5: 430.
Tarjan, R. and T. Kemeny. 1969. Multigeneration studies on DDT in mice.
Food Cosmet. Toxicol. 7: 215.
Tomatis, L., et al. 1971. Storage levels of DDT metabolites in mouse tis-
sues following long-term exposure to technical DDT. Tumori 57: 377.
Tomatis, L., et al. 1974. Effect of long-term exposure to l,l-dichlor-2,2-
bis(p-chlorophenyl) ethylene, to l,l-dichloro-2,2-bis (p-chlorophenyl) eth-
ane, and to the two chemicals combined on CF-l mice. Jour. Natl. Cancer
Inst. 52: 883.
U.S. EPA. 1979a. DOT: Ambient Water Quality.Criteria. (Draft).
U.S. EPA. 1979b. Environmental Criteria and Assessment Office. DDE: Haz-
ard Profile. (Draft).
U.S. EPA. 1979c. Environmental Criteria and Assessment Office. ODD: Haz-
ard Profile. (Draft)
U.S. FDA. 1977. Administrative Guidelines Manual 7426-04, Attachment E.
Vogel, E. 1972. Mutagenitatsuntersuchungen mit DDT und den DDT-metaboliten
DDE, ODD, DOOM und DDA. an Drosphila melanogaster. Mutat. Res. 16: 157.
Ware, G.W. and E.E. Good. 1967. Effects of insecticides on reproduction in
the laboratory mouse. II. Mirex, Telodrin and DDT. Toxicol. Appl. Phar-
macol. 10: 54.
-74**-
-------
Weisburger, J.H. and E.K. Weisburger. 1968. Food additives and chemical
carcinogens: on the concept of zero tolerance. Food Cosmet. Toxicol.
6: 235.
Welch, R.M., et al. 1969. Estrogenic action of DDT and its analogs. Toxi-
col. Appl. Pharmacol. 14: 358.
Wolfe, H.R. and J.F. Armstrong. 1971. Exposure of formulating plant work-
ers to DOT. Arch. Environ. Health 23: 169.
Wolfe, H.R., et al. 1967. Exposure of workers to pesticides. Arch. Envi-
ron. Hlth. 14: 622.
Wurster, C.F., Jr. 1968. DDT reduces photosynthesis by marine phytoplank-
ton. Science. 159: 1474.
--7 50-
-------
No. 61
Dibromochloromethane
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
-75*)-
-------
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.
-------
DIBROMOCHLORCMETHANE
SUMMARY
Dibromochlorome thane has been detected in drinking water in
the United States. It is believed to be formed by the haloform
reaction that may occur during water chlorination. Dibromochlo-
romethane can be removed from drinking water via treatment with
activated carbon. There is a potential for dibromochlorome thane
to accumulate in the aquatic evironment because of its resistance
to degradation. Volatilization is likely to be an important
means of environmental transport.
Very little toxicity information is available. Dibromochlo-
romethane gave positive results in mutagenicity tests with
Salmonella typhiirmrium TA100. It is currently under test by the
National Cancer Institute.
I. INTRODUCTION
Dibromochlorome thane (CHB^Cl, molecular weight 208.29) is a
clear, colorless liquid. It is insoluble in water, but is solu-
ble in a number of organic solvents. Its boiling point is 119-
120°C and its density is 2.45 at 20°C (Wsast, 1972). At 10.5°C,
its vapor pressure is 15 torr (Dreisbach, 1952).
A review of the production range {includes importation)
statistics for dibromochlorome thane (CAS No.--124-48-1) which is
listed in the initial TSCA Inventory (1979) has shown that
t
-75*3-
-------
between 0 and 900 pounds of this chemical were produced/imported
in 1977.V
Dibromochlorome thane is used as a chemical intermediate in
the manufacture of fire extinguishing agents, aerosol propel-
lants, refrigerants, and pesticides (Verschueren, 1977).
II. EXPOSURE
A. Environmental Fate
No information was found pertaining to the rate of oxidation
of dibromochloromethane in either the aquatic or atmospheric
environments. Dibromochlorome thane is probably like other halo-
genated aliphatics in that it is not easily oxidized in aquatic
systems because there are no functional groups which react
strongly with HO radical. A maximum hydrolytic half-life of 274
years has been reported for dibromochlorome thane at pH 7 and 25°C
(Mabey and Mill, 1978).
The vapor pressure of dibromochlorome thane, while lower than
that for chloroform and other chloroalkanes, is, nonetheless,
sufficient to ensure that volatilization will be an important
means of environmental transport. The concentration of dibromo-
chlorome thane present in water supplies has been reported to
JV This production range information does no't 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) .
-------
decrease as a result of volatilization while flowing through open
channels {Rook, 1974).
B. Bio accumulation
The log of the octanol/wa ter partition coefficient (log P)
as calculated by the method of Hansch is 2.09 (Tute, 1971) indi-
cating that dibromochlororaethane is somewhat lipophilic. As a
result, dibromochloromethane may exhibit a tendency to bioac-
cumulate in organisms. No experimental data were found to
confirm this.
C. Environmental Occurrence
Dibromochloromethane has been detected in finished drinking
water (Kleoper and Fairless, 1972; U.S. EPA, 1975), in drinking
water supplies (U.S. EPA, 1975), and in wastewater effluents
(Glaze and Henderson, 1975). Dibromochlorcme thane is hypothe-
sized to be present in water supplies as a result of the haloform
reaction which takes place during the chlorination of such water
(Rook, 1974; U.S. EPA, 1975; Glaze and Henderson 1975).
III. HEALTH EFFECTS
A. Carcinogenicity
Dibromochloromethane is currently under test for
carcinogenicity by the National Cancer Institute. No results are
available.
*•
B. flutagenicity
Dibromochloromethane was found mutagenic in Salmonella,
typhimurium TA100 in the absence of metabolic activation (Simmon,
1977) .
-------
C. Other Toxicity
A long-term test conducted by administration of high doses
of the chemical by gavage in mice showed a dose-dependent
decrease in the activity of liver and spleen phagocytes (Munson
e_t _a]L. , 1978).
The oral LD^Q of dibromochloromethane in mice is 800 mg/kg
and 1200 mg/kg for males and females respectively. Sedation and
anesthesia occurred within 30 minutes of administration of the
compound and lasted 4 hours. Necropsies were performed on ani-
mals that died. Hemorrhaging was observed in the adrenals, the
kidneys were pale, and the liver appeared to have fatty infiltra-
tion (Bowman, 1978).
IV. AQUATIC EFFECTS
No information was found.
V. EXISTING GUIDELINES
The Maximum Contaminant Level (MCL) for total trihalometh-
anes (including dibromochloromethane) in drinking water has been
set by the U.S. EPA at 0.10 mg/1 (44 FR 68624). The concentra-
tion of dibromochloromethane produced by chlorination can be
reduced by treatment of drinking water with powdered activated
carbon (Rook, 1974). This is the technology that has been pro-
j
posed by the EPA to meet this standard.
-------
REFERENCES
Bowman, F.J. e_t_ _al_. The Toxicity of Some Halomethanes in Mice.
Toxicology and Applied Pharmacology 44, 213-215, 1978.
Dreisbach, R. R. Pressure-Volume-Temperature Relationships of
Organic Compounds, Handbook Publishers, Inc. Sand us ky/ Ohio
1952.
Glaze, W.H. and J.E. Henderson, IV. Formation of Organochlorine
Compounds from the Chlorination of a Municipal Secondary Efflu-
ent. Journal Water Pollution Cont. Fed. 47, 2511-2515, 1975.
Kleopfer, R. D. and B.J. Fairless. Characterization of Organic
Components in a Municipal Water Supply. Environ. Sci. Technol.
6(12}, 1036-1037, 1972.
tlabey, W. and T. Mill. Critical Review of Hydrolysis of Organic
Compounds, in Water Under Environmental Conditions J. Phys. Chen.
Ref. Data 7, 103, 1978.
Munson, A.S. e_t al_. Retoculoendothelial System Function in Mice
Exposed to Four Haloalkanes: Drinking Water Contaminants.
Toxicology and Applied Pharmacology 45(1), 329-330, 1978.
Rook, J.J. Formation of haloforms during chlorination of natural
waters. Journal of the Society of Water Treatment and Examina-
tion 23(Part 2), 234-243, 1974.
Rook, J.J. Chlorination Reactions of Fulvic Acids in Natural
Waters. Environ. Sci. Technol. 11(5), 473-432, 1977.
Simmon, v.F. Structural Correlations of Carcinogenic and
ilutagenic Alky lhal ides, Proc. 2nd FDA Office of Science Summer
Sym. 163-171, 1977.
Tute, M.S. Principles and Practices of Hansch Analysis. A Guide
to Structure-Activity Correlation for the Medicinal Chemist.
Advances in Drug Research 6, 1-77, 1971.
U.S. EPA. Preliminary Assessment of Suspected Carcinogens in
Drinking Water. EPA 560/4-75-003, 1975.
U.S. EPA. Toxic Substances Control Act Chemical Substance
Inventory, production Statistics for Chemicals on the Non-
Confidential Initial TSCA Inventory, 1979.
*-
Verschueren, K. Handbook of Environmental Data on Organic
Chemicals. Van Nostrand Reinhold Co., New York. 1977.
t
Weast, R. C. , ed . 1972. CRC Handbook of Chemistry and Physics.
CRC Press, Inc., Cleveland, Ohio.
-757-
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No. 62
Di-n-butyl 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 accuracy.
-------
DI-n-BUTYL PHTHALATE
Summary
Teratogenie effects in rats have been reported in testing of di-
n-butyl phthalate following i.p. administration, but not after oral
administration at high doses (0.600 g/kg/day). Other reproductive
effects in rats following i.p. administration include impaired implantation
and parturition. Rats fed di-n-butyl phthalate or its monoester metabolite
have developed testicular damage and atrophy.
Mutagenic or carcinogenic effects of di-n-butyl phthalate have
not been reported.
One clinical study has indicated that workers exposed primarily,
but not exclusively, to di-n-butyl phthalate showed a higher incidence
of toxic polyneuritis.
The only toxicity data available for review demonstrate that di-
n-butyl phthalate is acutely toxic to freshwater organisms at concentrations
as low as 730 ^ug/1.
X
-7(oO-
-------
DI-n-BUTYL PHTHALATE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Phthalate Esters (U.S. EPA, 1979a).
Di-n-Butyl phthalate (DBF) is a diester of the ortho form of
benzene dicarboxylic acid. The compound has a molecular weight of 278.31*,
specific gravity of 1.0465, boiling point of 340°C and a solubility of
0.45 gms per 100 ml of water at 25°C (U.S. EPA, 1979a).
DBF is used as a plasticizer in polyvinyl acetate emulsions and
as an insect repellent.
Current Production: 8.3 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 DBP 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 of phthalate 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
r
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, 1979a).
-------
Monitoring studies have indicated that most water phthalate concen-
trations are in the ppra range, or 1-2 jug/liter (U.S. EPA, 1979a). Industrial
air monitoring studies have measured air levels of phthalates from 1.7
to 66 mg/m3 (Milkov, et al. 1973). Levels of DBF in foods have ranged
from not detectable to 60 ppra (Toraita, et al. 1977). Cheese, milk,
fish and shellfish present potential sources of high phthalate intake
(U.S. EPA, 1979a). The U.S. EPA (1979a) has estimated the weighted
average bioconcentration factor for DBF to be 26 for the edible portions
of fish and shellfish consumed by Americans. This estimate was based
on the octanol/water partition coefficient.
III. PHARMACOKINETICS
A. Absorption
A human 'study in which subjects ate food containing DBP
leached from plastic containers shows significantly higher levels of
DBF found in the blood (Tomita, et al. 1977).
3. Distribution
Pertinent data could not be located in the available literature.
C. Metabolism
Monobutyl phthalate has been identified as a urinary metabolite
in rabbits administered DBP (Ariyoshi, et al. 1976). This metabolite
has also been detected in the urine of rats, hamsters, and guinea pigs,
as well as other metabolites with side chain oxidation, and phthalic
acid (Tanaka, et al. 1978).
D. Excretion
Pertinent data could not be located in the available literature.
-76, A-
-------
IV. EFFECTS
A. Carcinogenicity
Pertinent data could not be located In the available literature.
B. Mutagenicity
Mutagenic effects of DBF were not observed in the Ames
Salmonella assay (Rubin, et al. 1979) or in a yeast (Saccharomyces)
assay system (Shahin and VonBorstel, 1977).
C. Teratogenicity
Teratogenic effects were not produced by DBF, (0.600 g/kg/day),
following oral administration to pregnant rats (Hikonorow, et al. 1973)
while Singly et al. (1972) reported teratogenic effects of DBF following
i.p. injection of pregnant rats.
D. Other Reproductive Effects
Intraperitoneal injection of DBF to pregnant rats showed
that adverse effects prior to gestation day six were primarily on implanta-
tion, while after this day the effect was primarily on parturition
(Peters and Cook 1973).
Testicular damage has been reported in rats fed DBF or its monoester
metabolite (Carter, 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.
Lesser levels of exposure to dioctyl, diisooctyl, and benzylbutyl phthalates,
and to tricresyl phosphate were also noted in these workers.
-------
V. AQUATIC TOXICITY
A. Acute Toxicity
Acute toxicity for di-n-butyl phthalate ranged from a 96-
'hour static LC$Q of 730 ;ug/l for the bluegill sunfish (Lepomis macrochirus)
to 6,470^/1 for the rainbow trout (Salmo gairdneri) (Mayer and Sanders,
1973). The freshwater scud (Qammarus pseudolimnaeus) was shown to
provide a *(8-hour static 1050 value of 2,100 ug/1 di-n-butyl phthalate.
Marine data were not available for review.
B. Chronic
Pertinent data could not be located in the available literature.
C. Plants
Pertinent data could not be located in the available literature.
D. Residues
Bioconcentration factors ranging from 400 to 1400 have been obtained
for the aquatic invertebrates Daphnla magna and Gamnarus pseudolimnaeus.
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 review; therefore,
there is a possibility that these criteria may be changed.
A. Human
Based on "no effect" levels observed in chronic feeding studies
in rats or dogs, the U.S. EPA (1979a) has calculated an acceptable daily
intake (ADI) level of 12.6 rag/day.
The recommended water quality criterion level for protection
of human health is 5 mg/liter for DBF (U.S. EPA, 1979a).
B. Aquatic
t-
The data base for toxic effects in both freshwater and marine
environments was insufficient for the drafting of a water quality criterion
to protect aquatic organisms.
-76H-
-------
DI-n-BUTYL PHTHALATE
REFERENCES
Aciyoshi, T., et ai. 1975. Metabolism of dibutyl phthalats
and the effects of its metabolites on animals. Kyushu Yaku-
gakkai Kaiho 30; 17.
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.
Engelhardt, G. , et al. 1975. The microbioal metabolism
of di-n-butyl phthalate and related dialkyi phthalates.
Bull. Environ. Contain. Toxicol. 17: 342.
Mayer, F.L., Jr., and H.O. Sanders. 1973. Toxicology of
phthalic ac-id esters in aquatic organisms. Environ. Health
Perspect. 3: 153.
Milkov, L.E., 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.
Nikonorow, M. , et al. 1973. Effect of orally administered
plasticizers and polyvinyl chloride stabilisers in the rat.
Toxicol. Appl. PhariTiacol. 26: 253.
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 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.
Shahin, M., and R. Von Borstel. 1977. Mutagenic and lethal
effects of a-benzene hexachloride, dibutyl phthalate and
trichloroethylene in Saccharomyces cerevisiae. Mutat. Res.
48: 173.
Singh, A., et al. 1972. Teratogenicity of phthalate esters
in rats. Jour. Pharm. Sci. 61: 51.
Tanaka, A., et al. 1978. Biochemical studies on phthalic
esters. III. Metabolism of dibutyl phthalate (DBP) 'in
animals. Toxicology 1: 109.
-76,5-
-------
Tomita, I., et al. 1977. Phthalic acid esters in various
foodstuffs and biological materials. Ecotoxicology and
Environmental Safety. 1: 275.
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).
-------
No. 63
DIbenzo(a,h)antnracene
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.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (GAG) has evaluated
dibenzo(a,h)anthracene and has found sufficient evidence to
indicate that this compound is carcinogenic.
-------
DIBENZO(a,h)ANTHRACENE
Summary
Dibenzo(a,h)anthracene (DBA) is a member of the polycyclic
aromatic hydrocarbon (PAH) class. DBA was the first pure chemi-
cal shown to produce tumors in animals. It is carcinogenic by
skin application, by injection, and by oral administration to
rodents. Since humans are not exposed to only DBA in the environ-
ment, it is not possible to attribute human cancers solely to
exposure to DBA. Furthermore, it is not known how DBA may inter-
act with other carcinogenic and non-carcinogenic PAH in human
systems.
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DIBENZO (a, h) ANTHRACENE
I . INTRODUCTION
This profile is based primarily on the Ambient Water Quality
Criteria Document for Polynuclear Aromatic Hydrocarbons (U.S. EPA,
1979a) and the iMultimedia Health Assessment Document for Polycyclic
Organic Matter {U.S. EPA. 1979b) .
Dibenzo{a,h) anthracene (DBA; C22H14) is one of the family of
polycyclic aromatic hydrocarbons (PAH) formed, as a result of incom-
plete combustion of organic material. Other than a reported
melting point of 266-266. 5°C (U.S. EPA. 1979b) , its physical and
chemical properties have not been well-characterized.
PAH, including DBA are ubiquitous in the environment, being
found in ambient air, food, water, soils and sediment (U.S. EPA.
1979b) . The PAH class contains a number of potent carcinogens
(e.g., benzo(a) pyrene) , moderately active carcinogens (e.g.,
benzo (b) f luor anthene) , weak carcinogens (benz (a) anthracene) , and
cocarcinogens (e.g., fluoranthene) , as well as numerous non-carcin-
ogens (U.S. EPA. 1979b) .
PAH which contain .more than three rings (such as DBA) are re-
latively stable in the environment, and may be transported in air
and water by adsorption to particulate matter. However, biodegrad-
ation and chemical treatment are effective in eliminating most PAH
in the environment.
II. EXPOSURE
*
A. Water
Levels of DBA in water have not been reported. However,
the concentration of six representative PAH (benzo (a) pyrene , fluor-
-77/-
-------
anthene, benzo ( j) fluoranthene, benzo ( k).f luoranthene, benzo(ghi)-
perylene, indeno (1 , 2, 3-cd-pyrene) in United States drinking water
averaged 13.5 nanograms/liter (Basu and Sacena, 1977, 1978).
B . Food
Based on limited monitoring studies, DBA has been de-
tected in various foods, such as, butter and smoked fish. Al-
though, it is not possible to estimate the human dietary intake
of DBA, it has been concluded (U.S. EPA. 1979b) that the daily
dietary intake of all types of PAH is about 1.6 to 16 yg per
day. The U.S. EPA (1979a) has estimated the weighted average
bioconcentration factor of DBA to be 24,000 for the edible por-
tions of fish and shellfish consumed by Americans. This estimate
is based on the octanol/water partition coefficient for DBA.
C. Inhalation
Levels of DBA have not been monitored in ambient air.
However, it has been estimated that the average total PAH level in
ambient air is about 10.9 nanograms/m (U.S. EPA, 1979a) . Thus the
total daily intake of PAH by inhalation of ambient air may be about
207 nanograms, assuming that a human breathes 19 ra of air per day.
III. PHARMACOKINETICS '
There are no data, available concerning the pharmacokinetics of
DBA, or other PAH, in humans. Nevertheless, it is possible to make
limited assumptions based on the results of animal research con-
ducted with several PAH, particularly benzo(a) p'yrene .
A. Absorption
f
The absorption of DBA in humans or other animals has not
been thoroughly studied. However, it is known (U.S. EPA, 1979a)
that, as a class, PAH are well-absorbed across the respiratory and
-17Z-
-------
gastrointestinal epithelia. The high lipics soluoility of compounds
in the PAH class supports this observation.
B. Distribution
Only limited work on distribution of DBA in mammals
has been performed (Heidelberger and Weiss, 1959). However,
it is known (U.S. EPA, 197ya) that other PAH become localized
in a wide variety of body tissues 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 mammalian metabolism of DBA has been well-character-
ized (Sims, 1976). DBA, like other PAH, is 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 dihydro-
diols by the intermediate formation of reactive epoxides. Dihydro-
diols are further metabolized by microsomal mixed function oxi-
dases to yield diol epoxides, compounds which are known to be
ultimate carcinogens for certain PAH. Removal of activated inter-
mediates by conjugation with glutathione or glucuronic acid,
or oy further metabolism to tetrahydrotetrols, is a key step
in protecting the organism from toxic interaction with cell macro-
molecules .
D. Excretion
There is no direct information available concerning the
*
excretion of PAH in man. The excretion of DBA however, by mice was
studied by Heidelberger and Weiss (195y). The excretion of DBA was
-773-
-------
rapid and occurred mainly via the feces. Elimination in the bile
accounts for a significant percentage of all administered PAH (U.S.
EPA, 1979a). It is unlikely that PAH will accumulate in the body
with chronic low-level exposures.
IV. EFFECTS
A. Carcinogenicity
DBA was the first pure chemical ever shown to produce
tumors in animals. DBA has considerable carcinogenic potency
when applied to the skin of mice (Iball, 19.39; U.S. EPA. 1979b} ,
injected subcutaneously in mice (U.S. EPA. 1979b), injected
into newborn mice (Beuning, et al. 1979), injected into Strain
A mice (Shimkin and Stoner, 1975} or administered orally to mice
(Snell and Stewart, 1962).
B. Mutagenicity
DBA is a mutagenic in the Ames Salmonella assay (Andrews,
et al. 1978; Wood, et al. 1973) in cultured hamster cells (Huberman
and Sacks, 1974), and is positive in the rn y_^vo sister-chromatid
exchange assay in Chinese hamsters (Roszinsky-Kocher, et al.
1979) .
C. Teratogenicity
There are no data available concerning the possible tera-
togenicity of DBA in man. Other related PAH apparently are not
significantly teratogenic in mammals (U.S. EPA, 1979a).
D. Other Reprodutive Effects
Pertinent information could not be located in the avail-
able literature.
-774-
-------
E. Chronic Toxicity
As long ago as 1937, investigators knew that carcinogenic
PAH, including DBA, could inhibit growth in rats and mice (Haddow,
et al. 1937). In early studies, DBA was administered to mice in
weekly subcutaneous injections for 40 weeks, which produced in-
creased reticulum (stem) cells, dilation of lymph sinuses, and de-
creased spleen weights in comparison to controls (Hoch-Ligeti ,
1941) .
A more detailed study of subchrofiic effects of DBA on
lymph nodes of male rats was reported in 1944 (Lasnitzki and Wood-
house, 1944). -Subcutaneous injections given five times weekly for
several weeks caused normal lymph nodes to undergo hemolymphatic
changes .
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 yet gone
through the process of public review; therefore, there is a possi-
bility that these criteria may be changed.
A. Human
There are no established exposure criteria for DBA. How-
ever, PAH as a class are regulated by several authorities. The
World Health Organization recommends that the concentration of PAH
in drinking water (measured as the total of f luoranthene, benzo-
(g,h, i)perylene, benzo (b) f luoranthene , benzo ( k} f luoranthene , in
-775"-
-------
deno(l,2,3-cd)pyrene, and benzo(a) pyrene) not exceed 0.2 ug/1.
Occupational exposure criteria have been established for coke
oven emissions, coal tar products, and coal tar pitch volatiles,
all of which contain large amounts of PAH including DBA (U.S.
EPA, 1979a).
The U.S. EPA (1979a) draft recommended criteria for
PAH in water are based upon the extrapolation of animal carcinogenicity
data for benzo(a)pyrene and DBA. Levels for each compound are de-
rived which will result in specified risk levels of human cancer as
shown in the table below.
BaP
Exposure Assumptions Risk Levels and_Cgrresponding Criteria
(per day) ng/1
£ 12lZ. l£lf 10~5
2 liters of drinking water
and consumption of 18.7
grams fish and shellfish 0 0.097 0.97 9.7
Consumption of fish and
shellfish only 0.44 4.45 44.46
DBA
2 liters of drinking water 0 0.43 4.3 43
and consumption of 18.7
grams fish and shellfish
Consumption of fish and 1.96 19.6 196
shellfish only.
B. Aquatic
f
The criterion for freshwater and marine life have not
been derived (U.S. EPA, 1979a) .
-776-
-------
DIBENZO(a,h)ANTHRACENE
REFERENCES
Andrews, A.W., et al. 1978. The relationship between carcinogeni-
city and mutagenicity of some polynuclear hydrocarbons. Mutation
Research 51: 311.
Basu and Saxena, 1977, 1978. Polynuclear aromatic hydrocarbons in
selected U.S. drinking waters and their raw water sources.
Environ. Sci. Technol. 12: 795.
Beuning, M.K., et al. 1979. Tumorigenicity of the dihydrodiols of
dibenzo(a,h)anthracene on mouse skin and in newborn mice. Cancer
Res. 39: 1310.
*.
Haddow, A., et al. 1937. The influence of certain carcinogenic
and other hydrocarbons on body growth in the rat. Proc. Royal Soc.
B. 122: 477.
Heidelberger, C., and S.M. Weiss. 1959. The distribution of
radioactivity in mice following administration of 3,4-benzopyrene-
5C and 1,2,5,6-dibenzanthracene-9, IOC . Cancer Res. 11: 885.
Hoch-Ligeti, C. 1941. Studies on the changes in the lymphoid
tissues of mice treated with carcinogenic and non-carcinogenic
hydrocarbons. Cancer Res. 1: 484.
Huberman, E., and L. Sachs. 1974. Cell-mediated mutagenesis of
mammalian cells with chemical carcinogens. Int. Jour. Cancer.
13: 326.
Iball, J. 1939. The relative potency of carcinogenic compounds.
Am. Jour. Cancer. 35: 188.
Lasnitzki, A., and Woodhouse, D.C. 1944. The Effect of 1:2:5:6-
Dibenzanthracene on the lymph-nodes of the rat. J. Anat. 78: 121.
Roszinsky - Kocker, et al. 1979. Mutagenicity of PAH's. Induc-
tion of sister-chromatid exchanges in vivo. Mutation Research.
66: 65.
Shimkin, M.B., and G.D. Stoner. 1975. Lung tumors in mice: appli-
cation to carcinogenesis bioassay. In: G. Klein and S. Wein-
house, (eds.) Advances in Cancer Research, Vol. 12 Raven Press,
New York.
Sims, P. 1976. The metabolism of polycyclic hydrocarbons totdi-
hydrodiols and diol epoxides oy human and animal tissues. Pages
211-224 in R. Montesano, et al. eds. screening tests in chemical
carcinogenesis. IARC Publ. No. 12. Lyon, France.
-777-
-------
Snell, K.C., and H.L. Stewart. 1962.. Induction of pulmonary
adenomatosis in DBA/2 mice by the oral administration of dibenzo-
{a,h)anthracene. Acta. Vn. Int. Cone. 19: 692.
U.S. EPA. 1979a. Polynuclear aromatic hydrocarbons: ambient
water quality criteria. (Draft).
U.S. EPA. 1979b. Health Effects Research Laboratory, Environment-
al Criteria and Assessment Office Research Triangle Park, N.C.
Wood, A.W., et al. 1978. Metabolic activation of dibenzo(a,h)-
anthracene and its dibydiodiols to bacterial mutageh's. Cancer
Res. 38: 1967.
World Health Organization. 1970. European Standards for drinking
water, 2nd ed. Geneva.
-------
No. 64
1,2-Dichlorben3ene
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-uracy.
-------
1,2-DICHLOROBENZENE
SUMMARY
1,2-Dichlorobenzene is a lipophilic compound which
upon absorption into the body, deposits in the fatty tissues.
This compound is detoxified by the liver microsomal enzymes.
On chronic exposure to 0.1 mg 1,2-dichlorobenzene/kg, rats
developed anemia, liver damage, and central nervous system
depression. There have not been studies available to deter-
mine the carcinogenic or teratogenic potential of 1,2-di-
chlorobenzene. 1,2-Dichlorobenzene was mutagenic when tested
with the mold Aspergillis nidulans and negative when tested
with the bacteria Salmonella typhimurium in the Ames assay.
The toxicity of 1,2-dichlorobenzene appears to be simi-
lar for freshwater and marine organisms with reported LCcn
values ranging between 1,970 and 27,000 ug/1.
X
-7SJ-
-------
1,2-DICHLOROBEN2ENE
INTRODUCTION
This profile is based on the Ambient Water Quality
Criteria Document for Dichlorobenzenes (U.S. EPA, 1979a) .
1,2-Dichlorobenzene {1,2-DCB or ODCB; CgH4Cl2; molecular
weight 147.01) is a liquid at normal environmental tempera-
tures. 1,2-Dichlorobenzene has a melting point of -17.6°C,
a boiling point of 179°C, a density of 1.30 g/ml at 20°C,
a water solubility of 145,000 ug/1 at 25°C,V and a vapor
pressure of 1 mm Hg at 20°C (Weast, 1975). The major uses
of 1,2-dichlorobenzene are as a process solvent in the manu-
facturing o'f toluene diisocyanate and as an intermediate
in the synthesis of dyestuffs, herbicides, and degreasers
{West and Ware, 1977).
II. EXPOSURE
A. Water
1,2-Dichlorobenzene has been detected in rivers,
groundwater, municipal and industrial discharges, and drink-
ing water. 1,2-Dichlorobenzene has been reported entering
water systems at average levels of 2 mg/1 as a result of
its use by industrial wastewater treatment plants for odor .
control (Ware and West, 1977). In 4 out of 110 drinking
waters, 1,2-dichlorobenzene was detected at an average con-
centration of 2.5 pg/1 (U.S. EPA, 1979a). Also, 1,2-dichloro-
benzene may be formed during chlorination of water contain-
*
ing organic precursor material (Glaze, et al. 1976).
-------
B. Food
There are not enough data to state quantitatively
the degree of 1,2-dichlorobenzene exposure through total
diet (U.S. EPA, 1979a). The U.S. EPA (1979a) has estimated
the weighted average bioconcentration factor of 1,2-dichloro-
benzene to be 200 for the edible portion..of aquatic organisms
consumed by Americans. This estimate is based on measured
steady-state bioconcentration studies in bluegill.
C. Inhalation v
1,2-Dichlorobenzene has been detected on airborne
particulate matter in California at concentrations between
8 and 53 ng/m2 (Ware and West, 1977). There is no other
available information on the concentration of this compound
in ambient air (U.S. EPA, 1979a).
III. PHARMACOKINETICS
A. Absorption
There is little information provided in U.S. EPA
(1979a) on the absorption specifically of 1,2-dichloroben-
zene. General information on the absorption of dichloro-
benzenes can be found in the Hazard Profile for Dichloro-
benzenes (U.S. EPA, 1979b) . Reidel (1941) has reported
absorption of 1,2-dichlorobenzene through the skin of rats
in lethal amounts after five dermal applications under severe
test conditions (painting twice daily directly on a 10 cm
area of abdominal skin). Also, 1,2-dichlorobenzene fed to
*
rats at less than 0.4 to 2 mg/kg/day was absorbed and accu-
-------
mulated in various tissues indicating significant absorption
by the gastrointestinal tract even at low levels of exposure
(Jacobs, et al. 1974a,b}.
B. Distribution
After feeding rats low levels of 1,2-dichloroben-
zene, in combination with other trace pollutants found in
the Rhine River, tissue accumulation was greater in fat
than in the liver, kidney, heart, and blood (Jacobs, et
al. 1974a).
C. Metabolism
The metabolism of 1,2-dichlorobenzene was studied
by Azouz, et al. (1955) in rabbits. 1,2-Dichlorobenzene
was mainly metabolized by oxidation to 3,4-dichlorcphenol
followed by the formation of conjugates with glucuronic
and sulfuric acids. Minor oxidative metabolites and their
conjugates were also detected.
D. Excretion
Excretion of the metabolic products of 1,2-dichloro-
benzene in the rabbit was mainly through the urine (Azouz,
et al. 1955).
IV. EFFECTS
A. Carcinogenicity
Specific positive evidence of the carcinogenicity
of DCB's is lacking. However, a sufficient collection of
varied data exist to suggest prudent regard of DCB as a
potential carcinogen (U.S. EPA, 1979a}.
-------
B. Mutagenicity
Treatment of the soil mold Aspergillus nj.duj.ans
for one hour in an ether solution of 1,2-dichlorobenzene
increased the frequency of back-mutations (Prasad, 1970).
In the Ames assay, 1,2-dichlorobenzene did not increase
the mutational rate of the histidine-requiring strains of
Salmonella typhimurium (Andersen, et al. 1972).
C. Teratogenicity
Studies of the teratogenicity of'•!, 2-dichloroben-
zene could not be located in the available literature.
D. Other Reproductive Effects
Information is not available.
E. Chronic Toxicity
In an inhalation study, Hollingsworth, et al.
(1958) exposed groups of 20 rats, 8 guinea pigs, 4 rabbits,
and 2 monkeys to the vapor of 1,2-dichlorobenzene seven
hours per day, five days per week for six to seven months
at an average concentration of 560 mg/m . No adverse effects
were noted in behavior, growth, organ weights, heraatology,
or upon gross and microscopic examination of tissues. In
a nine month chronic toxicity study, Varshavskaya (1967)
gave rats 1,2-dichlorobenzene at daily doses of 0.001, 0.01,
and 0.1 mg/kg. The toxicological observations in the highest
dose group were anemia and other blood changes, liver damage,
and central nervous system depression. The highest no-observ-
able-adverse-effect level for 1,2-dichlorobenzene by Var-
shavskaya (1967) was 0.001 mg/kg/day, whereas the compar-
-------
able level in the rat study by Hollingswor th , et al. (1958)
was 18.8 rag/kg/day.
F. Other Relevant Information
I/ 2-Dichlorobenzene can induce microsomal drug
metabolizing enzymes (Ware and West, 1977).
V. AQUATIC TOXICITY
A. Acute Toxicity
For freshwater fish, two 96-hour static bioassays
have produced LC5Q values of 5,590 and 27,0-00 ug/1 for the
bluegill (Lepgmis macrocjiirus) (U.S. EPA, 1978; Dawson,
et al. 1977). A single 96-hour static assay for the fresh-
water invertebrate Daphnia magna provided an LC5Q value
of 2,440 ug/1. In marine fish, LC5Q values reported were
7,300 pg/1 for the tidewater silverside {Men_id_j.a be r y 11 i n a )
and 9,660 ug/1 for the sheepshead minnow (Cyprj.no_don yar iega-
tus;) (U.S. EPA, 1978). An adjusted LC5Q value of 1,970
pg/1 was obtained for the marine invertebrate (Mysidgpsis
bah_ia) .
B. Chronic
The only freshwater organisms tested were embryo-
larval stages of the fathead minnow (Pimephales prprne la s ) ,
which produced a chronic value of 1,000 pg/1 for 1,2-dichloro-
benzene. No chronic data for marine organisms were avail-
able for evaluation.
C. Plants
+
The freshwater algae Selenastrum capr icpr;nuj:um
has been tested for the effects of 1, 2-dichlorobenzene on
-736-
-------
chlorophyll a and cell numbers. The EC^0 values were 91,600
and 98,000 pg/1, respectively, while comparable values of
44,200 to 44,100 pg/1 were reported for the marine algae
Skeletonema cgstatum (U.S. EPA, 1978).
D. Residues
A bioconcentration of 89 was obtained for the
bluegill.
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
The Occupational Safety and Health Administration
(OSHA, 1976), and the American Conference of Governmental
Industrial Hygienists (ACGIH, 1977) threshold limit value
is 300 mg/m for 1,2-dichlorobenzene. The U.S. EPA (1979a)
draft water quality criterion for total dichlorobenzene
(all three isomers) is 160 ug/1.
3. Aquatic
Criteria have been drafted for freshwater organisms
as 44 pg/1 for the 24-hour average concentration, not to
exceed 99 pg/1. The marine draft criterion is 15 pg/1 not
to exceed 34 pg/1 (U.S. EPA, 1979a).
-737-
-------
1,2-DICHLOROBENZENE
REFERENCES
American Conference of Governmental Industrial Hygienists. 1977. Documen-
tation of the threshold limit values for substances in workroom air (with
supplements' for those substances added or changed since 1971). 3rd sd.
Cincinnati, Ohio.
^
Andersen, K.J., et al. 1972. Evaluation of herbicides for possible muta-
genic properties. Jour. Agric. Food Chem. 20: 649.
Azouz, W.M., et al. 1955. Studies in detoxication, 62: The metabolism of
halogenobenzenes. Orthoand paradichlorobenzenes. Biochem. Jour. 59: 410.
Dawson, G.W., et al. 1977. The toxicity of 47. industrial chemicals to
fresh and saltwater fishes. Jour. Hazard Mater. 1: 303.
Glaze, W.H., et al. 1976. Analysis of new chlorinated organic compounds
formed by chlorination of municipal wastewater. In: Proc. Conf. Environ.
Impact Water Chlorination. Iss. Conf.-751096, pages 153-75. (Abstract)
Hollingsworth, R.L., et al. 1958. Toxicity of o-dichlorobenzene. Studies
on animals and industrial experience. AMA Arch. Ind. health 17: 180.
Jacobs, A., et al. 1974a. Accumulation of noxious chlorinated substances
from Rhine River water in the fatty tissue of rats. Votn Wasser (German)
43: 259. (Abstract)
Jacobs, A., et al. 1974b. Accumulation of organic compounds, identified as
harmful substances in Rhine water, in the fatty tissues of rats. Kern-
forschungszentrum Karlsruhe (Ber.). KFK 1969 UF, pp. 1. (Abstract)
National Academy of Sciences. 1977. Drinking water and health. U.S. EPA
Contract No. 68-01-3169. Washington, O.C.
Occupational Safety and Health Administration. 1976. General industry
standards. 29 CFR 1910, -July 1, 1975; OSHA 2206, revised Jan. 1976. U.S.
Dep. Labor, Washington, O.C.
Prasad, I. 1970. Mutagenic effects of the herbicide 3',4'-dichloropropio-
nanilide and its degradation products. Can. Jour. Microbiol. 16: 369.
Riedel, H. 1941. Einige beobachtungen uber orrhodichlorbenzol. Arch.
Gewerbepath. u Gewerbehyg. 10: 546. (German)
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. Oichlorobenzenes: Ambient Water Quality Criteria.
(Draft).
-------
U.S. EPA. 1979b. Environmental Criteria and Assessment Office. Dichloro-
benzenes: Hazard Profile. (Draft)
Varshavskaya, S.P. 1967. Comparative toxicological characteristics of
chlorobenzene and dichlorobenzene (ortho- and para-isomers) in relation to
the sanitary protection of water bodies. Gig, Sanit. (Russian) 33: 17.
Ware, S., and W.L. West. 1977. Investigation of selected potential envi-
ronmental contaminants: halogenated benzenes. EPA-560/2-77-004. Rep. EPA
Contract No. 68-01-4183. Off. Toxic Subst. U.S. Environ. Prot. Agency,
Washington, D.C,
Weast, R.C., et al. 1975. Handbook of chemistry and physics. 56th ed.
CRC Press, Cleveland, Ohio.
-------
No. 65
1,3-Dichlorobenzene
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
-no-
-------
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.
-------
1,3-DICHLOROBENZENE
Summary
1,3-Dichlorobenzene is not used commercially and is produced only as a
by-product in the manufacture of chlorinated benzenes. This compound is
metabolized by the liver mixed function oxidase system. Little is known of
the toxicological, teratogenic, or carcinogenic properties of this compound.
1,3-Dichlorobenzene has been shown to be mutagenic to the soil mold Asper-
gillus nidulans. Since 1,3-dichlorobenzene may be a contaminant of the
other dichlorobenzenes, some of the toxicologic properties ascribed to these
isomers may be due to the 1,3-isomer.
For freshwater and marine fish and invertebrates, acute toxicity values
ranged from 2,414 to 4,248 pg/1, but the freshwater invertebrate, Daphnia
magna, was more resistant to 1,3-dichlorobenzene with an acute value of
23,800 jug/1.
-------
1.3-DICHLOROBENZENE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Dichlorobenzenes (U.S. EPA, 1979a).
1,3-Oichlorobenzene (1.3-DCB; MDCB; C,H.C10; molecular weight
O 4 i.
147.01) is a liquid at normal environmental temperatures, has a melting
point of -24.2°C, a boiling point of 172°C, a density of 1.29 g/ml at
20°C, a water solubility of 123,000 jjg/1 at 25°C, and a vapor pressure
of 5 mm Hg at 39°C (Weast, 1975). 1,3-Dichlorobenzene may occur as a con-
taminant of 1,2- or 1,4-dichlorobenzene formulations (U.S. EPA, 1979a).
II. EXPOSURE
A. Water
1,3-Dichlorobenzene has been detected or quantified in groundwater,
raw water, and drinking water. In two of 110 drinking water samples, 1,3-
dichlorobenzene was detected at an average concentration of 0.1 jug/1 (U.S.
EPA, 1979a). Also, 1,3-dichlorobenzene may be formed during chlorination of
raw and waste water containing organic precursor material (Glaze, et al.
1976).
B. Food
The data are insufficient to state quantitatively the degree of
1,3-dichlorobenzene exposure through total diet (U.S. EPA, 1979a). 1,3-Di-
chlorobenzene is reported to be among several metabolites of gamma-penta-
chloro-1-cyclohexane found in corn and pea seedlings (Mostafa and Moza,
1973). The U.S. EPA (1979a) has estimated' the weighted average
bioconcentration factor to be 150 for 1,3-dichlorobenzene for the edible
»
portions of fish and shellfish consumed by Americans. This estimate is
based on measured steady- state bioconcentration studies in bluegill.
-------
C. Inhalation
Pertinent data could not be located in the available literature.
III. PHARMACOKINETICS
A. Absorption
Specific information on the absorption of 1,3-dichlorobenzene was
not found in the available literature. General information on the absorp-
tion of the dichlorobenzenes can be found in the Hazard Profile for Oichlor-
obenzenes (U.S. EPA,- 1979b).
B, Distribution
Specific information on the distribution of 1,3-dichlorobenzene was
not found in the available literature. Reference may be made to the Hazard
Profile for Oichlorobenzene (U.S. EPA, 1979b) and the 1,2-isomer (U.S. EPA,
1979c).
C. Metabolism
The metabolism of 1,3-dichlorobenzene in rabbits was studied by
Parke and Williams (1955). 1,3-Dichlorobenzene v/as mainly metabolized by
oxidation to 2,4-dichlorophenol followed by the formation of the glucuro-
nides and ethereal sulfates. Minor oxidative metabolites and their conju-
gates were also detected.
D. Excretion
Excretion of the metabolic products of 1,3-dichlorobenzene in the
rabbit is mainly through the urine with excretion being essentially complete
within five days (Parke and Williams, 1955).
IV. EFFECTS
A. Carcinogenicity
»
Reports of specific carcinogenicity tests of 1,3-dichlorobenzene in
animals or of pertinent epidemiologic studies in humans were not found in
the available literature (U.S. EPA, 1979a).
-------
B. Mutageniclty
Treatment of the soil mold Asperqillus nidulans for one hour in an
ether solution of 1,3-dichlorobenzene increased the frequency of back muta-
tions (Prasad, 1970).
C. Teratogenicity and Other Reproductive Effects
Studies of the teratogenicity and other reproductive effects of
1,3-dichlorobenzene were not found in the available literature.
0. Chronic Toxicity
Specific information on the chronic toxic'ity of 1,3-dichlorobenzene
was not found in the available literature. However, 1,3-dichlorobenzene may
have been a contaminant of the 1,2- and 1,4-dichlorobenzenes used in toxico-
logical studies. For further information on the general toxicologic proper-
ties of the dichlorobenzenes, refer to the Hazard Profile for Oichloroben-
zenes (U.S. EPA, 1979b).
E. Other Relevant Information
1,3-Dichlorobenzene can induce microsomal drug metabolizing en-
zymes. Changes in the levels of microsomal enzymes can affect the metabo-
lism and biological activity of a wide variety of xenobiotics (Ware and
West, 1977).
V. AQUATIC TOXICITY
A. Acute Toxicity
For the bluegill (Lepomis macrochirys), a 96-hour static LC5Q of
5,020 jjg/1 has been obtained. The freshwater invertebrate, Daphnia maqna,
has a much higher LC5Q of 28,100 jjg/1 for a 48-hour' static assay. For the
sheepshead minnow, an acute LC5Q of 7,770 jjg/1 has been obtained. A value
of 2,850 jug/1 has been obtained for the marine mysid shrimp (Mysidopsis
bahia) (U.S. EPA, 1978).
-------
B. Chronic
Chronic studies with either freshwater or marine species are not
available.
C. Plant Effects
The freshwater alga Selenastrum capricornutum was tested for the
effects of 1,3-dichlorobenzene on chlorophyll a and cell numbers. The
EC5Q values ranged from 149,000-179,000 pg/1. For the marine alga Skele-
tonema costatum, the EC5Q values for cell number and chlorphyll a ranged
from 49,600-52,800 pg/1 (U.S. EPA, 1979a).
0. Residues
A bioconcentration factor of 66 was obtained for the bluegill (U.S.
EPA, 1979a).
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
There are no existing standards for 1,3-dichlorobenzene. The U.S.
EPA (1979a) draft water quality criterion for total dichlorobenzene (all
three isomers) is 160 pg/1.
B. Aquatic
i
A criterion for the protection of freshwater organisms has been
drafted as 310 jug/1 for a 24-hour average concentration not to exceed 700
fjg/1. For marine life, the criterion has been proposed as 22 ug/1 for 24-
hour average not to exceed 49 jjg/1.
it
-------
1,3-DICHLOROBENZENE
REFERENCES
Glaze, W.H., et al. 1976. Analysis of new chlorinated organic compounds
formed by chlorination of municipal wastewater. In Proc. Conf. Environ.
Impact Water Chlorination. Iss. Conf.-751096, pages 153-75. (Abstract)
Mostafa, I.Y. and P.N. Moza. 1973. Degradation of gamma-pentachloro-1-
cyclohexane (gamma-PCCH) in corn and pea seedlings. Egypt. Jour. Chem. Iss.
Spec.: 235. (Abstract)
Parke, D.V. and R.T. Williams. 1955. Studies in detoxication: The metabo-
lism of halogenobenzenes. (a) Metadichlorobenzene (b) Further observations
on the metabolism of chlorobenzene. Biochem. Jour. 59: 415.
Prasad, I. 1970. Mutagenic effects of the herbicide 3',4'-dichloropropio-
nanilide and its degradation products. Can. Jour. Microbiol. 16: 369.
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. Dichlorobenzenes: Ambient Water Quality Criteria Docu-
ment. (Draft)
U.S. EPA. 1979b. Environmental Criteria and Assessment Office. Dichloro-
benzenes: Hazard Profile. (Draft)
U.S. EPA. 1979c. Environmental Criteria and Assessment Office. 1,2-Di-
chlorobenzene: Hazard Profile. (Draft)
Ware, S. and W. L. West. 1977. Investigation of selected potential envi-
ronmental contaminants: halogenated benzenes. EPA 560/2-77-004. Rep. EPA
Contract No. 68-01-4183. Off. Toxic Subst. U.S. Environ. Prot. Agency,
Washington, D.C.
Weast, R.C., et al. 1975. Handbook of chemistry and physics. 56th ed.
CRC Press, Cleveland, Ohio.
-------
No. 66
1,4-Dichlorobenzene
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.
-------
1,4-DICHLOROBEN ZENE
SUMMARY
1,4-Dichlorobenzene is a lipophilic compound which, upon
absorption into the body, deposits in the fatty tissues.
This compound is detoxified by the liver microsomal enzymes.
Chronic intoxication produces increased liver and kidney
weights' and abnormal liver pathology. Studies to determine
the carcinogenic or teratogenic potential of 1,4-dichloroben-
zene could not be located in the available literature. 1,4-
Dichlorobenzene produces chromosomal aberrations in root tips
and has been shown to increase the mutation rate in the mold
Aspergillus nidulans.
Acute values for freshwater and marine organisms ranged
from 1,990 to 11,000 ug/1 for 1,4-dichlorobenzene. Marine in
vertebrates were most sensitive and freshwater invertebrates
were most resistant to the effects of 1,4-dichlorobenzene.
-"3OO-
-------
1,4-DICHLOROBENZENE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Cri-
teria Document for Dichlorobenzene (U.S. EPA, 1979a).
1,4-Dichlorobenzene (CgH4Cl2; molecular weight 147.01)
is a solid at normal environmental temperatures. 1,4-Di-
chlorobenzene has a melting point of 53.0°C, a boiling point
of 174°C, a density of 1.25 g/ml at 20°C, a water solubility
of.80,000 ug/1 at 25°C, and a vapor pressure of 0.4 mm Hg at
25°C (Weast, et al. 1975). The primary use of 1,4-dichloro-
benzene is 'as an air deodorant and insecticide. This com-
pound is produced almost entirely as a byproduct during the
manufacture of monochlorobenzene (Ware and West, 1977).
For a more general discussion of dichlorobenzene, the
reader is referred to the Hazard Profile for Dichlorobenzene
(U.S. EPA, 1979b).
II. EXPOSURE
A. Water
1,4-Dichlorobenzene has been detected or quantified
in rivers, groundwater, municipal and industrial discharge,
and drinking water. 1,4-Dichlorobenzene enters wastewater
systems because of its use in toilet blocks (Ware and West,
1977). 1,4-Dichlorobenzene may also be formed during chlori-
nation of raw and waste water containing organic percursor
material {Glaze, et al. 1976). In 20 of 113 drinking water
samples, 1,4-dichlorobenzene was detected at an average coh-
centration of 0.14 ug/1 (U.S. EPA, 1979a).
-------
B. Food
There are not enough data available to quantita-
tively state the degree of 1,4-dichlorobenzene exposure
through total diet (U.S. EPA, 1979a). Schmidt {1971} report-
ed the tainting of pork as a result of the use of an odor
control agent containing 1,4-dichlorobenzene in pig stalls.
Also, Morita, et al. (1975) reported 0.05 mg/kg 1,4-dichloro-
benzene in fish from Japanese coastal waters. The U.S EPA
(1979a) has estimated the weighted bioconcentration factor of
1,4-dichlorobenzene to be 140 for the edible portion of fish
and shellfish consumed by Americans. This estimate is based
on measured steady-state bioconcentration studies in blue-
gills.
C. Inhalation
Morita and Ohi (1975) measured 1,4-dichlorobenzene
in the vapor phase, in and around Tokyo, by use of a cold
solvent trap. Urban levels were found to range from 2.7 to
4.2 ug/m-^, while suburban levels were lower, ranging from
1.5 to 2.4 ug/ra ; indoor levels were considerably higher,
ranging 0.105 to 1.7 mg/m^_ NO other information was found
regarding the concentration of this compound in ambient air
*
(U.S. EPA, 1979a).
III. PHARMAKIN ETICS
A. Absorption
In humans, toxic effects following accidentally or
t
deliberately ingested 1,4-dichlorobenzene clearly indicate
significant absorption by the gastrointestinal route (Camp-
bell and Davidson, 1970; Frank and Cohen, 1961; Hallowell,
-------
1959). Also, Azouz, et al. (1955} detected no 1,4-dichloro-
benzene in the feces of rabbits dosed intragastrically with
the compound in oil. This suggests virtually complete ab-
sorption under these conditions.
B. Distribution
The studies of Morita and Ohi (1975) and Morita, et
al. (1975) have shown 1,4-dichlorobenzene in adipose tissue
(mean about 2 mg/kg) and blood (about 0.01 mg/1) of humans
exposed to ambient pollution levels in the Tokyo area.
C. Metabolism
The metabolism of 1,4-dichlorobenzene in rabbits
was studied by Azouz, et al. (1955). 1,4-Dichlorobenzene was
primarily metabolized by oxidation to 2,5-dichlorophenol,
followed by the formation of the glucuronides and ethereal
sulfates. Minor oxidative metabolites and their conjugates
were also detected. Pagnatto and Walkley (1966) indicated
that 2,5-dichlorophenol was also the principal metabolite of
1,4- dichlorobenzene in humans.
D. Excretion
Excretion of the metabolic products of 1,4-di-
chlorobenzene in the rabbit occurs mainly through the urine
(Azouz, et al. 1955), with no mention made of fecal excre-
tion.
IV. EFFECTS
A. Carcinogenicity
*
No reports of specific carcinogenicity tests of
1,4-dichlorobenzene in animals or of pertinent epidemiologic
studies in humans were available. A few inconclusive experi-
-------
ments which indicate further investigation of the carcino-
genic potential of 1,4-dichlorobenzene is warranted are re-
viewed in U.S EPA (1979a).
B. Mutagenicity
Various mitotic anomalies were observed in cells
and somatic, chromosomes of 1,4-dichlorobenzene treated root
tips (Carey and McDonough, 1943; Sharma and Sarkar, 1957;
Srivastava, 1966). Treatment of Aspergillus nidulans (a soil
mold organism) for one hour in an ether solution of 1,4-di-
chlorobenzene increased the frequency of back-mutations
(Prasad, 1970) .
C. Teratogenicity and Other Reproductive Effects
Pertinent data could not be located in the avail-
able literature.
D. Chronic Toxicity
Effects observed in rats and guinea pigs exposed to
a concentration of 2,050 mg/m3 1,4-dichlorobenzene for six
months included: growth depression (guinea pigs); increased
liver and kidney weights (rats); abnormal.liver pathology
(cloudy swelling, fatty degeneration, focal necrosis, cirrho-
sis) (Hollingsworth, et al. 1956). In animals exposed to
4,800 mg/m3 1,4-dichlorobenzene, up to 25 percent deaths
were noted; and in survivors, symptoms were noted that were
similar to those observed at the lower dose. Similar pathol-
ogy was also observed in female rats, who received 376 mg/kg
»
dose of 1,4-dichlorobenzene by stomach tube 5 days a week for
a total of 138 doses.
-------
E. Other Relevant Information
1,4-Dichlorobenzene can induce microsomal drug-
metabolizing enzymes. Changes in the levels of microsomal
enzymes can affect the metabolism and biological activity of
a wide variety of xenobiotics (Ware and West, 1977).
V. AQUATIC TOXICITY
A. Acute Toxicity
Acute 96-hour LCcQ values for all aquatic species
tested were relatively similar. For the freshwater fish, the
bluegill (Lepomis macrochirus), a LC5Q of 4,280 ug/1 was
obtained, while the freshwater invertebrate Daphnia magna was
more resistant, with a LC50 value of 11,000. An LC50
value of 7,400 ug/1 was obtained for the marine fish, the
sheepshead minnow (Cyprinodon yariegatus); and the myrid
shrimp (Mysid op sis bahia) had an LC50 value of 1,990 ug/1
{U.S. EPA, 1978).
B. Chronic
Pertinent data could not be located in the avail-
able literature.
C. Plants
The freshwater alga, Selenastrum capricornutum,
when tested for the effects of 1,4-dichlorobenzene on chloro-
phyll _a and cell numbers, was shown to have had a range of
effective concentration of 96,700 to 98,100"' ug/1, while the
marine alga Skeletonema costatum was more sensitive, with an
• • ™~™" ^
effective concentration range of 54,800 to 59,100 ug/1.
-------
D. Residues
A bioconcentration factor of 60 was obtained for
the freshwater bluegill.
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
The Occupational Safety and Health Administration
Standard (OSHA, 1976), and the American Conference of Govern-
mental Industrial Hygienists (ACGIH, 1977) threshold limit
value are 450 mg/m3 for 1,4-dichlorobenzene. The
acceptable daily intake. (ADI) of 1,4-dichlorobenzene is 0.94
ing/day (Natl. Acad. Sci., 1977). The U.S. EPA (1979a) draft
water quality criterion for total dichlorobenzene (all three
isomers) is 0.16 mg/1.
B. Aquatic
A criterion for the protection of freshwater aqua-
tic life has been drafted as a 190 ug/1 24-hour average con-
centration, not to exceed 440 ug/1 at any time. For the pro-
tection of marine life, the criterion is 15 u.g/1 as a 24-hour
average, not to exceed 34 u.g/1 at any time.
-------
1, 4-DICHLOROBENZENE
REFERENCES
American Conference of Governmental Industrial Hygienists.
1977. Documentation of the threshold limit values for sub-
stances in workroom air (with supplements for those sub-
stances added or changed since 1971). 3rd ed. Cincinnati,
Ohio.
Azouz, W.M., et al. 1955. Studies in detoxication, 62: The
metabolism of halogenobenzenes. Ortho- and paradichloro-
benzenes. Biochem. Jour. 59: 410.
Campbell, D.M., and R.J.L. Davidson. 1970. Toxic haemolytic
anaemia in pregnancy due to a pica for paradichlorobenzene.
Jour. Obstet. Gynaec. Br. Cmnwlth. 77: 657.
Carey, M.A., and E.S. McDonough. 1943. On the production of
polyploidy .in Allium with paradichlorobenzene.
Frank, S.B., and H.J. Cohen. 1961. Fixed drug eruption due
to paradichlorobenzene. N.Y. Jour. Med. 61: 4079.
Glaze, W.H., et al. 1976. Analysis of new chlorinated
organic compounds formed by chlorination of municipal waste-
water. In: Proc. Conf. Environ, impact Water Chlorination.
Iss. Conf.-751096, pages 143-75. (Abstract).
Hallowell, M. 1959. Acute haemolytic anemia following the
ingestion of paradichlorobenzene. Arch. Dis. Child. 34:
74.
Hollingsworth, R.L., et al. 1956. Toxicity of paradichloro-
benzene. Determinations on experimental animals and human
subjects. AMA Arch. Ind. Health 14: 138.
Morita, M., et al. 1975. A systematic determination of
chlorinated benzenes in human adipose tissue. Environ.
Pollut. 9: 175 (Abstract).
Morita, M., and G. Ohi. 1975. Para-dichlorobenzene in human
tissue and atmosphere in Tokyo metropolitan area. Environ.
Pollut. 8: 269.
National Academy of Sciences. 1977. Drinking water and
health. U.S. EPA Contract No. 68-01-3169. -Washington, D.C.
Occupational Safety and Health Administration. 1976. Gener-
al industry standards. 29 CFR 1910, July 1, 1975; OSHA 2206,
revised Jan. 1976. U.S. Dep. Labor, Washington, D.C.
/ .
-•307-
-------
Pagnotto, L.D. , and J.E. Walkley. 1966. Urinary dichloro-
phenol as an index of paradichlorobenzene exposure. Ind.
Hyg. Assoc. Jour. 26: 137. (Rev. in Food Cosmet. Toxicol.
4: 109. (Abstract).
Prasad, I. 1970. Mutagenic effects of the herbicide 3',4'-
dichloropropionanilide andd its degradation products. Can.
Jour. Microbiol. 16: 369.
Schmidt, G.E. 1971. Abnormal odor and taste due to p-di-
chlorobenzene. Arch. Lebensmittelhyg. (German) 22: 43.
(Abstract).
Sharma, A.K., and S.K. Sarkar. 1957. A study on the compar-
ative effect of chemicals on chromosomes of roots, pollen
mother cells and pollen grains. Proc. Indian Acad. Sci.
Sect. B. 45: 288.
Srivastava,, L.M. 1966. Induction of mitotic abnormalities
in certain genera of tribe vicieae by paradichlorobenzene.
Cytologia 31: 166.
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. 1979a. Dichlorobenzenes: Ambient Water Quality
Criteria. (Draft).
U.S. EPA. 1979b. Environmental Criteria Assessment Office.
Dichlorobenzene: Hazard Profile (Draft).
Ware, S.A., and W.L. West. 1977. Investigation of selected
potential environmental contaminants: halogenated benzenes.
EPA 560/2-77-004. Rep. EPA Contract No. 68-01-4183. Off.
Toxic Subst. U.S. Environ. Prot. Agency, Washington, D.C.
Weast, R.C., et al. 1975. Handbook of chemistry and
physics. 56th ed. CRC Press, Cleveland, Ohio.
-------
No. 67
Dichlorobenzenes
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.
-------
DICHLORQBENZENE5
Summary
Dichlorobenzenes are lipophilic compounds which, upon absorption into
the body, deposit in the fatty tissues. These compounds are metabolized by
the liver microsomal enzyme system to water soluble compounds. Chronic ex-
posure to any of the three isomers produces effects on the liver, blood,
central nervous system and respiratory tract. Studies to determine the car-
cinogenic or teratogenic potential of the dichlorobenzenes were not located
in. the .available., literature.. .In one study--these-compounds-have increased
the mutational rate of soil mold.
The position of the chlorine atoms on the benzene ring appears to have
little significant effect on the toxicity of the 1,2-, 1,3-, or
1,4-dichlorobenzene isomers to fish and invertebrates, except for the appar-
ent resistance of the freshwater invertebrate Daohnia macna to 1,3-chloro-
benzene. Marine fish tend to be slightly more resistant than freshwater
fish, although the inverse is true for freshwater and marine invertebrates.
-------
DICHLOROBENZENES
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Dichlorobenzenes (U.S. EPA, 1979).
The dichlorobenzenes (CgH^Cl-; molecular weight 147.01) are a
class of halogenated aromatic compounds represented by three structurally
similar isomers: 1,2-dichloro-, 1,3-dichloro-, and 1,4-dichlorobenzenes
(Weast, et al. 1975). 1,2-Dichloro- and 1,3-dichlorobenzene are liquids at
normal environmental temperatures while 1,4-dichlorobenzene is a solid. All
the dichlorobenzenes boil at approximately 175°C and have a density close
to 1.28 g/ml. The solubilities in water of the 1,2-, 1,3-, and
1,,4-dichlorobenzens isomers at 25°C are 145,000 jLig/1, 123,000 jug/1, and
80,000 jjg/1, respectively (Jacobs, 1957). The vapor pressure of
1,2-dichlorobenzene at 20°C is 1 mm Hg; the vapor pressure of
1,3-dichlorobenzene at 39°C is 5 mm Hg; and the vapor pressure of
1,4-dichlorobenzene at 25°C is 0.4 mm Hg (Jordan, 1954; Kirk and Gthrr.er,
1963).
The major uses of 1,2-dichlorobenzene are as a process solvent in the
manufacturing of toluene diisocyanate and as an intermediate in the syn-
thesis of dyestuffs, ' herbicides, and degreasers. 1,4-Dichlorobenzene is
used as an air deodorant and an insecticide. 1,3-Dichlorobenzene is found
as a contaminant of the other two isomers. The combined annual production
of 1,2-, and 1,4-dichlorobenzene in the United States approaches 50,000
metric tons (Ware and West, 1977).
-------
II. EXPOSURE
A. Water
Dichlorobenzenes have been detected or quantified in rivers, ground
water, municipal and industrial discharges, and drinking water. Dichloro-
benzenes enter the water systems from the use of 1,2-dichlorobenzene as a
deodorant in industrial wastewater treatment and from the use of
1,4-dichlorobenzene toilet blocks (Ware and West, 1977). Chlorinated ben-
zenes may also be formed during chlorination of raw and wastewater con-
taining organic precursor material (Glaze, et al. 1976). In two case
studies the concentration of dichlorobenzene in finishedI water. was higher
than in the raw water supply (Gaffney, 1976).
B. Food
There are not enough data to state quantitatively the degree of
dichlorobenzene exposure 'through total diet. Tainting of pork has been
reported due to the use of an odor control product containing 1,4-cichioro-
benzene in pig stalls (Schmidt, 1971). Also, low levels of contamination of
plant products have been noted from the metabolism of lindane and gamma-
pentachlor-1-cyclohexane (Balba and Saha, 1974; Mostafa and Moza, 1973).
Morita, et al. (1975) reported detectable levels 'of 1,4-dichlorobenzene in
fish of the Japanese coastal waters; the concentration was 0.05 my/kg. The
U.S. EPA (1979) has estimated the weighted average bioconcentration factors
for the edible portion of fish and shellfish consumed by Americans for 1,2-
dichloro-, 1,3-dichloro-, and 1,4-dichlorobenzene to be 200, 150, and 140,
respectively. These estimates are based on measured steady-state biocon-
centration studies in bluegills.
-------
C. Inhalation
1,2-Dichlorobenzene has been detected in airborne particulate
o
matter in California at concentrations between 8 and 53 ng/m (Ware and
West, 1977). Morita and Ohi (1975) measured 1,4-dichlorobenzene in the
vapor phase, by the use of a cold solvent trap, in and around Tokyo. Urban
levels were 2.7 to 4.2 /jg/m3; suburban levels were lower at 1.5 to 2.4
/jg/m ; however, indoor levels were considerably higher at 0.105 to 1.7
mg/m3.
III. PHARMACQKINETICS
A. Absorption
The dichlorcbsnzenes may be absorbed through the lungs, gastro-
intestinal tract, and intact skin (Ware and West, 1977). There is no data
on the quantitative efficiency of absorption of dichlorobenzenes; however,
as indicated from the appearance of metabolites in the urine, respiratory
absorption during inhalation exposure is rapid (Pagnatto and Walkley,
1966). In humans, toxic effects following accidentally or deliberately in-
gested 1,4-dichlorobenzene clearly indicate significant absorption by the
gastrointestinal route (Campbell and Davidson, 1970; Frank and Cohen, 1961;
Hallowell, 1959). Also, 1,2-dichlorobenzene fed to rats at less than 0.4 to
2 mg/kg/day was absorbed and accumulated in various tissues, indicating
significant absorption by the gastrointestinal tract even at low levels of
exposure by ingestion (Jacobs, et al. 1974a,b).
B. Distribution
*•
After feeding rats low levels of 1,2-dichlorobenzene in combination
with other trace pollutants found in the Rhine River, tissue accumulation
was greater in fat than in the liver, kidney, heart, and blood (Jacobs, et
al. 1974a). Studies of Morita and Ohi (1975) and Morita, et al. (1975) have
-------
shown 1,4-dlchlorobenzene in adipose tissue (mean about 2 mg/kg) and blood
(about 0.01 mg/1) of humans exposed to ambient pollution levels in the Tokyo
area.
C. Metabolism
Metabolism of the 1,2- and 1,4-dichiorobenzenes was studied by
Azouz, et al. (1955), and 1,3-dichlorobenzene was studied by Parke and
Williams (1955) in rabbits. These compounds are mainly metabolized by oxi-
dation to 3,4-dichlorophenol, 2,5-dichlorophenol, and 2,4-dichlorophenol
respectively, which are subsequently conjugated.. Other oxidation products
are formed to a lesser extent, followed again by conjugation. Pagnatto and
Walkley (1966) indicated that 2,5-dichlorophenol was also the principal
metabolite of 1,4-dichlorobenzene in humans.
D. Excretion
In studies of rabbits, Azouz, et ai. (1955) and Parke and Williams
(1955) reported the excretion of metabolic products of the dichlorobenzenes
in the urine.
IV. EFFECTS
A. Carcinogenicity
No reports of carcinogenicity testing of specific dichlorobenzenes
could be located in .the available literature. Inconclusive experiments
reviewed in U.S. EPA (1979) indicate that further investigation of the car-
cinogenic potential of the dichlorobenzenes is warranted.
B. Mutagenicity
Various mitotic anomalies were observed in cells and somatic
chromosomes of 1,4-dichlorobenzene-treated root tips (Srivastava, 1966;
Sharma and Sarkar, 1957; Carey and McDonough, 1943). Treatment of
• Asperqillus nidulans (a soil mold organism) for one hour in an ether
-------
solution .of any of the three isomers of dichlorobenzene increased the
frequency of back-mutations (Prasad, 1970). -In the Ames assay,
1,2-dichlorobenzene did not increase the mutational rate of the
histidine-requiring strains of Salmonella typhircurlum (Andersen, et al.
1972).
C. Teratogenicity and Other Reproductive Effects
Pertinent data could not be located in the available literature.
Campbell and Davidson (1970) reported the history of a woman who
was eating p-OCB during her pregnancy, and which had no apparent effect on
the offspring.
D. Chronic Toxicity
In humans, chronic occupational exposure by inhalation has occurred
mainly from 1,4-dichlorobenzene and to a lesser extent 1,2-dichlorobenzene.
Toxicity has involved the following organs and tissues: liver, blood (or
reticuloendothelial system, including bone marrow and/or immune components),
central nervous system, respiratory tract, and integument (U.S. EPA, 1979).
In an inhalation study, Hollingsworth, et al. (1958) exposed groups of 20
rats, eight guinea pigs, four rabbits, and two monkeys to vapor of
1,2-dichlorobenzene for seven hours per day, five days per week for six to
seven months at an average concentration of 560 mg/m . NO adverse effects
were noted in behavior, growth, organ weights, hematology, or gross and
microscopic examination of tissues. In a nine-month chronic toxicity study
Varshavskaya (1967), gave rats 1,2-dichlorobenzene at daily doses of 0.001,
0.01, and 0.1 mg/kg. The toxicological observations in the highest dose
group was anemia and other blood changes, liver damage, and central nervous
system depression. Liver damage has also been observed with rats and guinea
pigs exposed to 1,4-dichlorobenzene at a concentration of 2,050 mg/m for
-------
six months (Hollingsworth, et al. 1956)". There have been no specific
studies on the chronic effects of 1,3-dichiorobenzene, although this com-
pound may have been a contaminant in the preparations of the other two iso-
mers used for toxicoiogical testing (U.S. EPA, 1979).
E. Other Relevant Information
Dichlorobenzenes can induce the microsomal drug metabolizing en-
zymes. Changes in the levels of microsomal enzymes can affect the metab-
olism and biological activity of a wide variety of xenobiotics (Ware and
West, 1977).
V. AQUATIC-TGXICITY
A. Acute Toxicity
Acute studies have indicated that the position of the chlorine
atoms on the benzene ring do not dramatically influence the toxicity of
dichlorobenzenes for freshwater fish. In 96-hour static bioassays with
bluegills, . Lepcmis macrohirus, LC5Q values were 4,280, 5,590 and 5,020
jug/1 for 1,4-, 1,2, and 1,3-dichlorobenzene, respectively (U.S. EPA, 1978).
However, Dawson, et al. (1977) has provided a 96-hour static LC,-,, value of
27,000 pg/1 for 1,2-dichlorobenzene for the same species. A greater range
of toxicities was obtained for the freshwater invertebrate Daphnia rnaqna
tested in 96-hour static bioassays. LC^ values were: 2,440; 11,000; and
28,100 /jg/1 for the 1,2-, 1,4-, and 1,3-dichlorobenzene isomers, respec-
tively (U.S. EPA, 1978). Marine fish were slightly more resistant than
freshwater fish in 96-hour static assays with l,C50 values ranging from
17,400 to 9,660 pg/1 for 1,4- and 1,2-dichlorobenzene, respectively, for the
sheepshead minnow. Marine invertebrates were the most sensitive organisms
-------
tested with LC^ values of 1,970, 1,990, and 2,850 /ug/1 obtained for 1,2-,
1,4-, and 1,3- dichlorobenzenes respectively in mysid shrimp (Mysidopsis
bahia) (U.S. EPA, 1978).
B. Chronic Toxicity
The only chronic study performed was an embryo-larval test .of the
freshwater fish, the fathead minnow (Pimephales promelas), that produced a
chronic value of 1,000 ug/1. No other chronic studies were available.
C. Plant Effects
The freshwater algae Selenastrum capricbrnutum, when tested for the
effects of dichlorobenzenes on chlorophyll a and cell numbers, had effective
concentrations ranging from 91,600 to 98,000; 149,000 to 179.000; and 96,700
to 98,100 jjg/1 for 1,2-, 1,3-, and 1,4-dichlorobenzene, respectively.
Similar studies in the marine algae Skeletonema costatum revealed effective
concentrations of 44,100 to 44,200; 49,600 to 52,800; and 54,300 to 59,100
for 1,2-, 1,3-, and 1,4-dichlorobenzenes.
0. Residues
Bioconcentration factors of 89, 66, snd 60 were obtained for 1,2-,
1.3-, and 1,4-dichlorobenzenes in the bluegill. Data on marine biocon-
centration factors are not available.
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
The Occupational Safety and Health Administration, (OSHA, 1976),
and the American Conference of Governmental Industrial Hygienists (ACGIH,
1977) threshold limit value is 300 mg/m3 for 1,2-dichlorobenzene and 450
mg/m for 1,4-dichlorobenzene. The acceptable daily intake (ADI) of 1,2-
*
or 1,4-dichlorobenzene is 1.316 mg/day (Natl. Acad. Sci., 1977). There are
-------
no standards for 1,3-dichlorobenzene. The U.S. EPA (1979) draft water
quality criterion for total dichlorobenzene (all three isomers) is 0.16 tng/1.
B. Aquatic
The draft criteria for the protection of freshwater organisms are
44 pg/1 not to exceed 99 jug/1 for 1,2-dichlorobenzene; 310 /jg/1 not to ex-
ceed 700 ug/1 for 1,3-dichlorobenzene; and 190 pg/1 not to exceed 440 /jg/1
for 1,4-dichlorobenzene. For marine organisms criteria have been drafted as
15 jjg/1 not to exceed 34 jjg/1 for 1,2-dichlorobenzsne; 22 /jg/1 not to exceed
49 jjg/1 for 1,3-dichlorobenzene; and 15 jjg/1 '"not to exceed' 34 jjg/1 for
1,4-dichlorobenzene.
-------
DICHLOROBENZENE5
References
American Conference of Governmental Industrial Kygienists. 1977. Docu-
mentation of the threshold limit values for substances in workroom air (with
supplements for those substances added or changed since 1971). 3rd ed.
Cincinnati, Ohio.
Anderson, K.J., et al. 1972. Evaluation of herbicides for possible muta-
genic properties. Jour. Agric. Food Chem. 20:- 649.
Azouz, W.M., et al. 1955. Studies in detoxication, 62: The metabolism of
halogenobenzenes. Orthoand paradichlorobenzenes. Biochem. Jour. 59: 410.
Balba, M.H. and J.G. Sana. 1974. Metabolism'- of lindane-14C by wheat
plants frown from treated seed. Environ. Latt. 7: 181 (Abstract).
Campbell, D.M. and R.J.L. Davidson. 1970. Toxic haemolytic anaemia in
pregnancy due to a pica for paradichlorobenzene. Jour. Obstet. Gynaec. Br,
Cmnwlth. 77: 657.
Carey, M.A. and E.S. McDonough. 1943. On the production of polyploidy in
Allium with paradichlorobenzene.
Dawson, G.W., et al. 1977. The toxicity of 47 industrial chemicals to
fresh and saltwater fishes. Jour. Hazard. Mater. 1: 303.
Frank, S.8. and H.J. Cohen. 1961. Fixed drug eruption cue to para-
dichlorobenzene. N.Y. Jour. Med. 61: 4079.
Gaffney, P.E. 1976. Carpet and rug industry case study. I. Water and
wastewater treatment plant operation. Jour. Water Pollut. Control Fed.
43: 2590.
Glaze, W.H., et al. 1976. Analysis of new chlorinated organic compounds
formed by chlorination of municipal wastewater. In: Proc. Conf. Environ.
Impact Water Chlorination. Iss. Conf.-751096, pages 153-75. (Abstract).
Hallowell, M. 1959. Acute haemolytic anemia following the ingestion of
paradichlorobenzene. Arch. Dis. Child. 34: 74.
Hollingsworth, R.L., et al. 1956. Toxicity of paradichlorobenzene. Deter-
minations on experimental animals and human subjects. AMA Arch. Ind.
Health 14: 138.
s
Hollingsworth, R.L., et al. 1958. Toxicity of o-dichlorobenzene. Studies
on animals and industrial experience. AMA Arch. Ind. Health 17: 180.
t
Jacobs, S. 1957. The handbook of solvents. 0. Van Nostrand Co., Inc., New
York.
-------
Jacobs, A., et al. 1974a. Accumulation o-f noxious chlorinated substances
from Rhine River water in the fatty tissue of rats. Vom Wasser (German)
43: 259. (Abstract). ,
Jacobs, A., et al. 1974b. Accumulation of organic compounds, identified as
harmful substances in Rhine water, in the fatty tissues of rats.
Kernforschungszentrum Karlsruhe (Ber.) KF',< 1969 UF, pp. 1 (Abstract).
Jordan, I.E., 1954. Vapor pressure of organic compounds. Interscience
Publishers, Inc., New York.
Kirk, R.E. and D.E. Othmer. 1963. Kirk-Othmer encyclopedia of chemical
technology. 8th ed. John Wiley and Sons, Inc. New York.
Morita, M., et al. 1975. A systematic determination of chlorinated ben-
zenes in human adipose tissue. Environ. Pollut. 9: 175. (Abstract).
Morita, M. and G. Ohi. 1975. Para-dichlorobenzene in human tissue and
atmosphere in Tokyo metropolitan area. Environ. Pollut. 8: 269.
Mostafa, I.Y. and P.M. Moza. 1973. Degradation of ganiiTia-penta-
chloro-1-cyclohexane (gamma-PCCH) in corn and pea seedlings. Egypt. Jour.
Chem. Iss. Spec.: 235. (Abstract).
National Academy of Sciences. 1977. Drinking water and health. U.S. EPA
Contract No. 68-01-3169. Washington. D.C.
Occupational Safety and Health Administration. 1S76. General industry
standards. 19 CFR 1910, July 1, 1975; OSEA 2206, revised Jan. 1976. U.S.
Dep. Labor. Washington, D.C.
Pagnotto, L.D. and J.E. Walkley. 1966. Urinary dichlorcpnenol as an index
of paradichlorobenzene exposure. Ind. Eyg. Assoc. Jour. 26: 137. (Rev. in
Food Cosmet. Toxicol. 4: 109. (Abstract).
Parke, D.V. and R.T. Williams. 1955. Studies in detoxication: The
metabolism of halogenobenzenes. (a) Metadichlorobenzene (b) Further obser-
vation on the metabolism of chlorobenzene. Siochem. Jour. 59: 415.
Prasad, I. 1970. Mutagenic effects of the herbicide 3',4'-dichloropropio-
nanilide and its degradation products. Can. Jour. Microbiol. 16: 369.
Schmidt, G.E. 1971. Abnormal odor and taste due to p-dichlorobenzene.
Arch. Lebensmittelhyg. (German) 22: 43. (Abstract).
Sharma, A.K. and S.K. Sarkar. 1957. A study on the comparative effect of
chemicals on chromosomes of roots, pollen mother cells and pollen grains.
Proc. Indian Acad. Sci. Sect. B. 45: 288.
Srivastava, L.M. 1966. Induction of mitotic abnormalities in certain
genera of tribe Vicieae by paradichlorobenzene. Cytologia 31: 166.
-------
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. Dicnlorobenzenes: Ambient Water Quality Criteria.
(Draft).
Varshavskaya, S.P. 1967. Comparative toxicological characteristics of
chlorobenzene and dichlorobenzene (orthoand para-isomers) in relation to the
sanitary protection of water bodies. Gig. Sanit. (Russian) 33: 17.
Ware, S.A. and W.L. West. 1977. Investigation of selected potential envi-
ronmmental contaminants: halogenated benzenes. U.S. Environ. Prot. Agency,
Washington, D.C,
Weast, R.C., et al. 1975. Handbook of chemistry and physics. 56th ed.
CRC Press, Cleveland, Ohio.
-------
No. 68
3,3'-Dichlorobenzidine
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-uracy.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated
3 , 3'-dichlorobenzidine and has found sufficient evidence to
indicate that this compound is carcinogenic.
-------
3,3'-PICHLOROBEN 21DIN E
SUMMARY
The adverse health effects associated with 3,3r-dichloro-
benzidine include the elevated risk of carcinogenicity based
upon data from several experimental bioassays. Animals ex-
posed to dust containing dichlorobenzidine were found to have
a slight to moderate pulmonary congestion.
One aquatic toxicity test has been performed for di-
chlorobenzidine, yielding results indicating that concentra-
tions of 0.5 ug/1 were acutely toxic to a freshwater fish
species.
-------
3,3'-DICHLOROBEN ZIDIN E
I. INTRODUCTION
This profile is based primarily on the Ambient Water
Quality Criteria Document for Dichlorobenzidine (U.S. EPA,
1979). The molecular formula of 3,3'-dichlorobenzidine
(4,4'-diamino-3,3'-dichlorobiphenyl} is C12H10C12N2'
and has a molecular weight of 253.13. The chemical is spar-
ingly soluble in water (0.7 g/1 at 15°C), but readily soluble
in organic solvents. Because of the fact that 3,3'-dichloro-
benzidine is an organic base, it may be fairly tightly bound
to humic materials, causing long-term storage in soils.
3, 3*-Dichlorobenzidine has been demonstrated to be a
carcinogen in experimental animals. Various types of sar-
comas and adenocarcinomas have been induced at injection
sites, and in specific organ systems upon dosage by gavage.
No evidence is available implicating 3,3'-dichlorobenzidine
as a human carcinogen.
II. EXPOSURE
A. Water
3,3'-Dichlorodibenzidine has been detected in water
near a waste disposal lagoon ranging from 0.13 to 0.27 mg/1,
as have benzidine concentrations up to 2.5 mg/1 (Sikka, et
al. 1978). In water of the Sumida River in Tokyo receiving
effluents of dye and pigment factories (Takemura, et al.
1965) total aromatic amines including 3,3'-dichlorobenzidine
were reported as high as 0.562 mg/1. The literature tends, to
support the possibility that the use of storage lagoons to
handle 3,3'-dichlorobenzidine wastes may pose a threat to
persons relying on nearby wells for drinking water.
-------
B. Food
Data quantifying levels of 3,31-dichlorobenzidine
in foods have not been reported. It was suggested that con-
sumption of fish would serve as the major dietary intake of
3,3'-dichlorobenzidine. No measurable levels of 3 , 3'-dichloro-
benzidine were detected (<10 ug/D in fish sampled near a
contaminated waste-lagoon (Diachenko, 1978).
The U.S. EPA (1979) has estimated the weighted
average bioconcentration factor to be 1,150 for 3,3'-dichloro-
benzidine for the edible portions of fish and shellfish con-
sumed by Americans. This estimate is based on the octanol/
water partition coefficient.
C. Inhalation
The low volatility and large crystal structure of
3,3'-dichlorobenzidine would tend to minimize the risk of ex-
posure to the chemical in ambient air. However, inhalation
may be a major source of exposure to those individuals occu-
pationally exposed to 3,3'-dichlorobenzidine. Concentrations
as high as 2.5 mg/100 m3 have been reported in one Japanese
pigment factory (Akiyama, 1970).
D. Dermal
Under specific conditions of moist skin and high
atmospheric humidity and temperature dermal absorption of
3,3'-dichlorobenzidine may be possible.
III. PHARMACOKINETICS
A. Absorption ••
Data concerning the rates and degree of absorption
of dichlorobenzidine have not be quantitated.
-------
B. Distribution
One study administering C^-^C)-3 , 3 '-dichloroben-
zidine at doses of 0.2 mg/kg intravenously in rats, monkeys,
and dogs revealed a general distribution of radioactivity
after 14 days. The highest {l^cj-3,3'-dichlorobenzidine
levels were found in the livers of all three species, in the
bile of monkeys and in lungs of dogs (Kellner, et al. 1973).
C. Metabolism
Following the intravenous injection of 0.2 mg/kg
(14C)-3,3'-dichlorobenzidine, the total urinary radioac-
tivity was recovered as one-third unchanged (14C)-3,3'-
dichlorobenzidine, one-third as the mono-N-acetyl derivative
of the parent compound, and the remainder not recoverable
(Kellner, et al., 1973). Chronic ingestion of small doses of
3,3'-dichlorobenzidine lead to the appearance of four meta-
bolic products including benzidine (U.S. EPA, 1979), however,
the results may be questionable due to the analytical methods
employed in the study. No metabolites of 3 ,3'-dichlorobenzi-
dine have been detected in the excreta of dogs experimentally
administered the parent compound (U.S. EPA, 1979), nor the
urine of human subjects experimentally administered the chem-
ical (Gerarde and Gerarde, 1974).
E. Excretion
Several studies have indicated that fecal elimina-
tion may be a major route of excretion in animals and humans
(U.S. EPA, 1979). One study (Meigs, et al. 1954) detected'
unspecified amounts of 3,3'-dichlorobenzidine in the urine of
occupationally exposed workers.
-------
IV. EFFECTS ON MAMMALS
A. Carcinogenicity
A number of investigations have reported the car-
cinogenic potential of 3,3'-dichlorobenzidine. Dietary 3,3'-
dichlorobenzidine at 1,000 mg/kg have been associated with
the significant occurrence of mammary adenocarcinomas, granu-
locytic leukemia, and zymbal gland carcinomas in male rats
and mammary adenocarcinomas in female rats (Stula, et al.
1975). In dogs, oral doses of 100 mg/kg were associated with
the significant occurrence of hepatic and urinary bladder
carcinomas (Stula, et al. 1975). Levels'of 0.5 and 1.0 mis
of a 4.4 percent suspension of 3,3'-dichlorobenzidine in rat
feed, resulting in a 4.53 g total dose of the chemical, pro-
duced an increase of cancers of the mammary gland, 2ymbal
gland, urinary bladder, skin, snail intestine, liver, thyroid
gland, kidney, hematopoietic system and salivary glands
(Pliss, 1959). Hepatic tumors and sebaceous gland carcinoma
were observed in mice exposed to a total dose of 127.5 to 135
mg over a ten month period of time (Pliss, 1959). 3,3"-Di-
chlorobenzidine was administered at levels of 30 mg every 3
days for 30 days by gavage. Observations over nine months
demonstrated that DCB is ineffective as a mammary carcinogen
(Griswold, et al. 1968). A diet of 0.3 percent 3,3'-dichloro-
benzidine was marginally carcinogenetic and tumorigenic to
hamsters {U.S. EPA, 1979). 3,3'-Dichlorobenzidine has also
found to produce transformation in cultured rat embryo celis
(Freeman, et al. 1973). Epidemiology studies in the United
States, Great Britian, and Japan have not provided evidence
-------
that 3,3'-dichlorobenzidine by itself induces bladder cancer
in workers occupationally exposed to the chemical. For some
studies, though, the latent period for tumor formation might
not have elapsed.
B. Mutagenicity
3,3'-Dichlorobenzidine has been shown to induce
frame shift mutations in Salmonella typhimurium tester strain
TA1598 in the presence of the S9 NADPH-fortified rat liver
enzyme preparation (Garner, et al. 1975). Similar results
with tester strain TA98 indicating frame shift mutations and
tester strain 1000 indicating base-pair substitutions were
observed by prior metabolic activation with a male mouse
enzyme system (Lazear and Louis, 1977).
C. Teratogenicity
Information relative to the teratogenic effects of
3,3'-dichlorcbenzidine was not found in the available
literature. Document (U.S. EPA, 1979). The chemical has
been shown to cross the placental barrier and increase the
incidence of leukemia in the offspring of pregnant mice given
doses of 8-10 mg of 3,3'-dichlorobenzidine subcutaneously
during the last week of pregnancy, but these results may
represent tox ic effects on neonates through suckling milk
from dosed mothers (Golub, et al. 1969, 1974). Altered
growth and morphology of cultured kidney tissue obtained from
prenatally exposed mouse embryos has been observed (Shabad,
t
et al. 1972; Golub, et al. 1969).
-------
D. Toxic ity
An acute oral LD^Q for DCB in mice, given to
mice for seven consecutive days was 352 mg/kg/day for females
and 386 mg/kg/day for males. Single-dose LVc,Q values
were reported as 488 and 676 mg/kg for female and male mice,
respectively. Rats exposed to atmospheric dust containing
unspecified amounts of 3,3'-dichlorobenzidine for 14 days
showed no increased mortalities. Upon autopsy slight to
moderate pulmonary congestion and one pulmonary abcess were
observed.
V. AQUATIC TOXICITY
The only aquatic species tested for the toxic effects of
3,3'-dichlorobenzidine was the bluegill, Lepomis macrochirus.
It was found to be acutely toxic at concentrations of 0.5
mg/1 or greater (Sikka, et al. 1978).
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 possibil-
ity that these criteria will be changed.
A. Human
The American Conference of Governmental Industrial
Hygienists has recommended that exposure to 3,3'-dichloroben-
zidine be reduced to zero, based on the demonstrated carcino-
genicity of the chemical in experimental animals. Occupa-
tional standards have not been placed on 3,3'-dichlorobenzi-
dine and standards regulating levels of the chemical in the
environment or in food have not been proposed.
if
-------
A recommended draft criterion of 1.69 x 10~2 U9/1
has been established, corresponding to a lifetime cancer risk
of 10~-*. This value was derived from data relating 3,3'-
dichlorobenzidine to the daily consumption of two liters of
water and 18.7 g of fish and shellfish.
B. Aquatic
Data were insufficient to draft criteria for either
freshwater or marine life.
-------
3,3'-DICHLOROBEN ZIDINE
REFERENCES
Akiyama, T. 1970. The investigation on the manufacturing
plant of organic pigment. Jikei. Med. jour. 17: 1.
Diachenko, G. 1978. Personal communication, U.S. Food and
Drug Administration.
Freeman, A.E., et al. 1973. Transformation of cell cultures
as an indication of the carcinogenic potential of chemicals.
Jour. Natl. Cancer Inst. 51: 799.
Garner, R.C., et al. 1975. Testing of some benzidine ana-
logues for microsomal activation to bacterial mutagens.
Cancer Lett. 1: 39.
Gerarde, H.W. , and D.F. Gerarde. 1974. Industrial experi-
ence with 3,3'-dichlorobenzidine: an epidemiological study of
a chemical manufacturing plant. Jour. Occup. Med. 16: 322.
Golub, N.I. 1969. Transplacental action of 3,3'-dichloro-
benzidine and orthotolidine on organ cultures of embryonic
mouse kidney tissue. Bull. Exp. BioJ.. Med. (U.S.S.R.) 68:
1280.
Golub, N.I., et al. 1974. Oncogenic action of some nitrogen
compounds on the procjeny of experimental mice. Bull. Exp.
Biol. Med. (U.S.S.R'.) 78: 62.
Griswold, D.P., et al. 1968. The carcinogenicity of multi-
ple intragastric doses of aromatic and heterocyclic nitro or
amino derivatives in young female Sprague-Dawley rats.
Cancer Res. 28: 924.
Kellner, H.M., et al. 1973. Animal studies on the kinetics
of bensidine and 3 ,3'-dichlorobenzidine. Arch. Toxicol. 31:
61.
Lazear, E.J., and S.C. Louis. 1977. Mutagenicity of some
congeners of benzidine in the Salmonella typhimurium assay
system. Cancer Lett. 4: 21.
Meigs, J.W., et al. 1954. Skin penetration by diamines of
the benzidine group. Arch. Ind. Hyg. Occup. Med. 9: 122.
Pliss, G.B. 1959. Dichlorobenzidine as a blastomogenic
agent. Vopr. Onkol. 5: 524.
Sf
-------
Shabad, L.M., et al. 1972. Transplacental effects of some
chemical compounds on organ cultures of embryonic kidney
tissue. Cancer Res. 32: 617.
Sikka, H.C., et al. 1978. Fate of 3,3'-dichlorobenzidine in
aquatic environments. U.S. Environ. Prot. Agency 600/3-8-
068.
Stula, E.F.r et al. 1975. Experimental neoplasia in rats
from oral administration of 3,3'-dichlorobenzidine , 4,4'-
methylene-bis(2-chloroaniline) , and 4,4'-methylene-bis(2-
methylaniline). Toxicol. Appl. Pharmacol. 31: 159.
Takemura, N., et al. 1965". A survey of the pollution of the
Sumida River, especially on the aromatic amines in the water.
Internat. Jour. Air Water Pollut. 9: 665.
U.S. EPA. 1979. Dichlorobenzidine: Ambient Water Quality
Criteria. (Draft).
Sf
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No. 69
1,1-Dichloroethane
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.
-------
1.1-OICHLORQETHANE
Summary
There is no available evidence to Indicate that 1,1-dichloroethane pro-
duces carcinogenic or mutagenic effects. A single study in rats failed to
show teratogenic effects following inhalation exposure.
Symptoms produced by human poisoning include respiratory tract irrita-
tion, central nervous system depression, and marked cardiac excitation. An-
imal studies indicate that 1,1-dichloroethane may produce liver damage.
Sufficient toxicological data are not available to calculate aquatic
exposure criteria.
-------
1,1-DICHLOROETHANE
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 have been replaced by chlorine atoms. Water solubility and vapor
pressure decrease with increasing chlorination, while density and melting
point increase. 1,1-Dichloroethane (ethylidene dichloride; ethylidene
chloride; molecular weight 98.96) is a liquid at room temperature with a
boiling point of 57.3°C, a melting point of -98°C, a specific gravity of
1.1776, and a solubility in water of 5 g/liter (U.S. EPA, 1979a).
The chloroethanes are used as solvents, cleaning and degreasing agents,
and in the chemical synthesis of a number of compounds. No commercial pro-
duction of 1,1-dichlcroethane has been reported in the United States (NIOSH,
1973).
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 chlqroethanes may be formed
by chlorination of drinking water or treatment of sewage. Air levels of
these volatile compounds are produced by evaporation during use as degreas-
ing agents and in dry-cleaning operations (U.S. EPA, 1979a).
-•339-
-------
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).
No information on levels of 1,1-dichloroethane in foods was found in
the available literature. Sufficient data is not available to estimate a
steady-state bioconcentration factor for 1,1-dichloroethane.
III. PHARMACOKINETICS
Pertinent data could not be located in the available literature on
1,1-dichloroethane 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 chloroeth-
anes throughout the body; enzymatic dechlorination and oxidation to the al-
cohol and ester forms; and excretion of the chloroethanes primarily in the
urine and in expired air.
Additionally, it has been indicated that the absorption of 1,1-dichlor-
oethane is most similar to that of the 1,2-isomer (indicating significant
dermal absorption as well as rapid oral or inhalation absorption).
IV. EFFECTS
A. Carcinogenicity and Mutagenicity
Pertinent data could not be located in the available literature.
B. Teratogenicity
An inhalation study in rats has indicated no major teratogenic ef-
fects of 1,1-dichloroethane (Schwetz, et al. 1974).
C. Other Reproductive Effects
»
Inhalation of 1,1-dichloroethane by pregnant rats produced delayed
ossification of sternebrae in fetuses, indicating an effect of the compound
in retarding fetal development (Schwetz, et al. 1974).
-------
D. Chronic Toxicity
Use of 1,1-dichloroethane as an anesthetic was discontinued because
of marked excitation of the heart (Browning, 1965). Poisoning cases have
shown respiratory tract irritation and central nervous system depression
(U.S. EPA, 1979a). Animal studies indicate that inhalation of 1,1-dichloro-
ethane may produce liver damage (Sax, 1975).
V. AQUATIC TOXICITY
Pertinent aquatic toxicity data could not be located in the available
literature.
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
The current promulgated Occupational Safety and Health Administra-
tion exposure standard for 1,1-dichloroethane is 100 ppm, time-weighted
average for up to a 10-hour work day, 40-hour work week.
Sufficient data are not available to derive a criterion to protect
human health from exposure to 1,1-dichloroethane from ambient water.
B. Aquatic
Sufficient toxicologic data are not available to calculate aquatic
exposure criteria.
-------
1,1-DICHLOROETHANE
REFERENCES
Browning, E. 1965. Toxicity and metabolism of industrial solvents.
Elsevier Publishing Co., Amsterdam.
Dickson, 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.
Kirk, R., and D. Othmer. 1963. Encyclopedia of Chemical Technology. 2nd
ed. John Wiley and Sons Inc. New York.
Lange, N.f ed. 1956. Handbook of Chemistry. 9th ed. Handbook Publishers,
Inc. Sandusky, Ohio.
National Institute for Occupational Safety and Health. 1978. Ethylene
dichloride (1,2-dichloroethane). Current Intelligence Bull. 25. DHEW
(NIOSH) Publ. No. 78-149.
Sax, N.I., ed. 1975. Dangerous properties of industrial materials. 4th
ed. Reinhold Publishing Corp. New York.
Schwetz, 3.A., et al. 1974. Embryo and fetotcxicity of inhaled carbon
tetrachloride, 1,1-dichloroethane, and methyl ethyl ketone in rats.
Toxicolo. Appl. Pharnscol. 28: 452.
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).
-------
No. 70
1,2-DIchloroethane
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-uracy.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated
1,2-dichloroethane and has found sufficient evidence to
indicate that this compound is carcinogenic.
-------
1.2-OICHLOROETHANE
Summary
Results of an NCI carcinogenesis bioassay in rats and mice have shown
that 1,2-dichloroethane may produce a wide variety of tumors, including
squamous cell carcinomas, hemangiosarcomas, mammary adenocarcinomas, and
hepatocellular carcinomas. Mutagenic effects have been shown in the Ames
Salmonella system and in E. coli; metabolites of 1,2-dichloroethane have
also shown mutagenic effects in the Ames assay. v
One study has failed to indicate teratogenic effects following inhala-
tion exposure to 1,2-dichloroethane although reproductive toxicity was
demonstrated. Chronic human exposure to 1,2-dichloroethane has produced
neurological symptoms and liver and kidney damage. Poisoning victims have
shown diffuse dystrcphic changes in the brain and spinal cord.
Acute toxicity values for freshwater organisms ranged from 431,000 to
550,000 jjg/1. Marine invertebrates appeared to be somewhat more sensitive
to 1,2-dichloroethane with an LC50 value of 113,000 jug/1 reported.
-w-
-------
1,2-DICHLOROETHANE
I. INTRODUCTION
This profile is based on the draft Ambient Water Quality Criteria Docu-
ment 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 density and melting
point increase. 1,2-Oichloroethane (molecular weight 98.96) is a liquid at
room temperature with a boiling point of 83.4°C, a melting point of
-35.4°C, a specific gravity of 1.253, and a solubility of 8.1 g/1 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. A large portion of
1,2-dichloroethane is used in the production of vinyl chloride and chlori-
nated chemicals, and as an ingredient, along with tetraetnyi lead, in anti-
knock mixtures (U.S. EPA, 1979a).
1,2-Dichloroethane production in 1976 was 4,000 x 103 tons (U.S. EPA,
1979a). The chlorinated ethanes form azeotropes with water (Kirk and Qthmer,
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
j-
The chloroethanes present in raw and finished waters are due primarily
to industrial discharges. Small amounts of the chloroethanes may be* formed
-w-
-------
by chlorination of drinking water or treatment of sewage. Of 80 water
samples tested, 27 contained 1,2-dichloroethane at concentrations of 0.2 to
8 jug/1 (U.S. EPA, 1974).
Sources of human exposure to chloroethanes not only include water, but
also air, contaminated foods and fish, and dermal absorption. For example,
1,2-dichloroethane has been detected in 11 of 17 spices in concentrations
ranging from 2 to 23 jjg/g of spice (Page and Kennedy, 1975). In fish and
shellfish, levels of chloroethanes in the nanogram range have been found
(Dickson and Riley, 1976).
The U.S. EPA (1979a) has estimated the weighted average bioconcen-
tration factor for 1,2-dichloroethane to be 4.6 for the edible portions of
fish and shellfish consumed by Americans. This estimate was based on the
measured steady-state bioconcentration studies in bluegills.
III. FHARMACOKINETICS
A. Absorption
The chloroethanes are absorbed rapidly following oral or inhalation
routes of exposure (U.S. EPA, 1979a). Animal studies indicate that signif-
icant absorption of 1,2-dichloroethane may occur .following dermal appli-
cation (Smyth, et al. 1969).
B. Distribution
Pertinent information could not be located in the available litera-
ture on 1,2-dichloroethane. The reader is referred to more general treat-
ment of the chloroethanes (U.S. EPA, 1979b) which indicates a widespread
distribution of chloroethanes through the body.
-m-
-------
C. Metabolism
In general, the metabolism of chloroethanes- involves both enzymatic .
dechlorination and hydroxylation to corresponding alcohols (U.S. EPA,
1979a). Metabolism of 1,2-dichloroethane produces a variety of metabolites
in the urine. The main two are: s-carboxymethylcysteine and thiodiacetic
acid (Yllner, 1971a,b,c,d). Yllner (1971a,b,c,d) also stated that the
percentage of 1, 2-dichloroethane metabolized decreased with increasing
dose, suggesting saturation of metabolic pathways.
0. Excretion
The chloroethanes are excreted primarily in the urine and in ex-
pired air (U.S. EPA, 1979a). Animal studies conducted by Yllner (1971a,b,
c,d) indicate that large amounts of chlorinated ethanes administered by i.p.
injection are excreted in the urine, with very little excretion in the
feces. Excretion appears tc be rapid, since 90 percent of an i.p. adminis-
tered dose of 1,2-dichloroethane was eliminated in the first 24 hours (U.S.
EPA, 1979a).
IV. EFFECTS
A. Carcinogenicity
Results of the NCI bioassay for carcinogenicity (NCI, 1978) have
indicated that 1,2-dichloroethane administration produced an increase in
several types of tumors. Squamous cell carcinomas and hemangiosarcomas were
produced in male rats, and mammary adenocarcinomas in female rats, following
feeding of 1,2-dichloroethane. In mice, hepatocellular carcinomas in males
and mammary adenocarcinomas in females were both increased after oral treat-
ment with 1,2-dichloroethane.
-------
B. Mutagenicity
Testing of 1,2-dichloroethane in the Ames -Salmonella assay and an
E. coli assay system have indicated mutagenic activity of this compound
(Brem, et al. 1974). Metabolites of 1,2-dichloroethane (S-chloroethyl
cysteine, chloroethanol, and chloroacetaldehyde) have shown positive muta-
genic effects in the Ames system (U.S. EPA, 1979a). 1,2-Dichloroethane has
also been reported to increase .mutation frequencies in pea plants
(Kiricheck, 1974) and Orosophila (Nylander, et al. 1978).
C. Teratogenicity
Inhalation studies with 1,2-dichloroethane in pregnant rats did not
indicate teratogenic effects (Vozovaya, 1974).
D. Other Reproductive Effects
Rats exposed' to 1,2-dichloroethane by inhalation showed reduced
litter sizes, decreased live births, and decreased fetal weights (Vozovaya,
1974).
E. Chronic Toxicity
Patients suffering from 1,2-dichloroethane poisoning have shown
diffuse dystrophic changes in the brain and spinal cord (Akimov, et al.
1978). Chronic exposures have produced neurologic changes and liver and
kidney impairment (NIOSH, 1978a).
Animal studies with 1,2-dichloroethane toxicity have shown liver
and kidney damage and fatty infiltration, and some bone marrow effects (U.S.
EPA, 1979a).
F. Acute Toxicity
Oral human I_DLQ (lowest dose which has caused death) values have
r
been estimated at 500 and 810 mg/kg in two studies (NIOSH, 1978b). Other
species show a similar sensitivity to 1,2-dichloroethane, except for the
-------
rat. An LD5Q value for this species has been estimated to be 12 ug/kg
(NIOSH, 197Sb).
V. AQUATIC TOXICITY
A. Acute Toxicity
Acute 96-hour static LC5Q values ranged from 431,000 to 550,000
pg/1 for the bluegill (Lepomis macrochirus), while a single 48- hour static
LC5Q value of 218,000- jug/1 was obtained for the freshwater cladoceran
Daphnia maqna (U.S. EPA, 1978). A single acute marine invertebrate study
was available, reporting a 96-hour static LC^" value of 113,000 ;jg/l for
the mysid shrimp (Mysidopsis bahia) (U.S. EPA, 1978).
B. Chronic Toxicity and Plant Effects
Pertinent information could not be located in the available litera-
ture on chronic toxicity and plant effects.
C. Residues
A bioconcentration factor of 2 has been reported for the bluegill
(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
Based on the NCI carcinogenesis bioassay data, and using a linear,
nonthreshold model, the U.S. EPA (1979a) has' estimated a level of
1,2-dichloroethane in ambient water that will result in an additional cancer
»
risk of 10~3 to be 7 ug/1.-
-------
The 8-hour TWA exposure standard developed by OSHA for
1,2-dichloroethane is 50 ppm.
B. Aquatic
In freshwater environment a criterion has been drafted for
1,2-dichloroethane as 3,900 jjg/1 as a 24-hour average, not to exceed 8800
jug/1. For marine life, the criterion has been drafted as 880 jjg/1, not to
exceed 2000
-------
1,2-DICHLOROETHANE
REFERENCES
Akinov, G.A., et al. 1978. Neurologic disorders in acute
dichloroethane poisoning. Zh. Nerropatol. Psikhiatr. 78:
687.
Brem, S.L., et al. 1974. The mutagenicity and DMA-Modifying
effect of haloalkanes. Cancer Res. 34: 2576.
Dickson, A.G., and J.P. Riley. 1976. The distribution of
short-chain halogenated aliphatic hydrocarbons in some marine
organisms. Mar. Pollut. Bull. 79: 167.
Kirk, R., and D. Othmer. 1963. Encyclopedia of chemical
technology. 2nd ed. John Wiley and Sons,"Inc., New York.
Kiricheck, Y.F. 1974. Effect of 1,2-dichloroethane on muta-
tions in peas. Usp. Khim. Mutageneza Se. 232.
Lange, N. (ed.) 1956. Handbook of chemistry. 9th ed.
Handbook Publishers, Inc., Sandusky, Ohio.
National Cancer Institute. 1978. Bioassay of 1,2-dichloro-
ethane for possible careinogenicity. Natl. Inst. Health,
ilatl. Cancer Inst. Carcinogenesis Testing Program. DHE?J
Publ. No. (NIH) 78-1305. Pub. Health Serv. U.S. Dep. Health
Edu. Welfare.
National Institute for Occupational Safety and Health. 1978a.
Ethylene dichloride (1,2-dichloroethane). Current Intelli-
gence Bull. 25. DHEW (NIOSH) Publ. No. 78-149.
National Institute for Occupational Safety and Health. 1978b.
Registry of toxic effects of chemical substances, DHEW {NIOSH)
Publ. No. 79-100.
Mylander, P.O.., et'al. 1978. Mutagenic effects of petrol in
Prosoph ila melanoaaster. I. Effects of benzene of and 1,2-
dichloroethane. Mutat. Res. 57: 163.
Page, B.D., and B.P.C. Kennedy. 1975. Determination of
methylene chloride, ethylene dichloride, and trichloroethy-
lene as solvent residues in spice oleoresins, using vacuum
distillation and electron-capture gas chromatography. Jour.
Assoc. Off. Anal. Chem. 58: 1062.
Smyth, H.F. Jr., et al. 1969. Range-finding toxicity data:
List VII. Am. Ind. Hyg. Assoc. Jour. 30: 470.
U.S. EPA. 1974. "Draft analytical report-Mew Orleans area
water supply study," EPA 906/10-74-002. Lower Mississippi
River Facility, Slidell, La., Surveill. Anal. Div. Region VI,
Dallas, Tex.
-------
U.S. EPA. 1978. In-depth studies on health and environmen-
tal impacts of selected water pollutants. U.S. Environ.
Prot. Agency. Contract No. 68-01-4646.
U.S. EPA. 1979a. Chlorinated Ethanes: Ambient Water Quality
Criteria. (Draft).
U.S. EPA. 1979b. Environmental Criteria and Assessment Of-
fice. Chlorinated Ethanes: Hazard Profile (Draft).
Van Dyke, R.A., and C.G. Wineman. 1971. Enzymatic
dechlorination: Dechlorination of chloroethanes and propanes
in vitro. Biochem. Pharmacol. 20: 463.
Vozovaya, M.A. 1974. Development of progeny of two genera-
tions obtained from female rats subjected to the action of
dichloroethane. Gig. Sanit. 7: 25.
Yllner, S. 1971a. Metabolism of 1,2-dichloroethane -14c
in the mouse. Acta. Pharmacol. Toxicol. 30: 257.
Yllner, S. 1971b. Metabolism of 1,1,2-trichloroethane-1-2-
- cin the mouse. Acta. Pharmacol. Toxicol. 30: 248.
Yllner, S. 1971c. Metabolism of 1,1,1,2-tetrachloroethane
in the mouse. Acta. Pharmacol. Toxicol. 29: 471.
Yllner,. S. 1971d. Metabolism of .1,1,2 ,2-tetrachloroethane-
- 4C in the mouse. Acta. Pharmacol. Toxicol. 29: 499.
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No. 71
1,1-Dlchlorethvlena
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.
-855? •
-------
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-dichlorc-
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-Qichloroethylene 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-hcur LC5Q values range from
73,900 to 108,000 jjg/1 1,1-dichloroethylene. Reported 48-hour EC5Q values
for Daphnia maqna range from 11,600 to 79,000 ug/1. 96-Hour LC5Q values
of over 224,QCO 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.
-------
1,1-OICHLOROETHYLENE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Oichloroethylenes (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
(PVCOs). Prior to 1976, annual production of 1,1-dichloroethylene was ap-
proximately 120,000 metric tons (Arthur D. Little, Inc., 1976). 1,1-Di-
chloroethylene has the following physical/chemical properties: water solu-
bility of 2,500 pg/ml, vapor pressure 591 mm Hg, and a melting point of
-122.1°C. For more general information regarding the dichloroethylenes,
the reader is referred to the EPA/ECAO Hazard Profile on Dichicroatnylenes
(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.
B. 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.
-------
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 on 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 dose is absorbed at two
dose levels: 1 and 50 mg/kg.
B. Distribution
Distribution of 1,1-dichloroethylene was studied in rats follovdng
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 macromolecules of the
microsomes and mitochondria (Jaeger, et ai. 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-dichlordethylene is metabolically converted, with 36 percent appearing
in the urine of rats within 26 hours (Jaeger, et al. 1977). The essential
-------
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
are found in the kidney and liver (McKenna, et al. 1978b). Interaction of
1,1- dichloroethylene with the microsomal mixed function oxidasa 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 (McKenna, et al. 1978a).
D. Excretion
It is- speculated that 1,1-dichloroethylene has a rapid rate of
elimination, since a substantial fraction of the total absorbed dose may be
recovered in the urine within 26 to 72 hours (Jaeger, et al. 1977; McKenna,
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-Dichloroethylene has been shown to produce kidney adenocarci-
nomas in male mice and mammary adenocarcinomas in female mice upon inhala-
3 *
tion of 100 mg/m (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 £he same
-860-
-------
conditions as the mice failed to exhibit an increased tumor incidence (Mal-
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 £_._ 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; Anderson, et al. 1977).
C. Teratcgenicity
A study by Murrary, et al. (1979) failed to shew teratcgenic ef-
fects in rats or rabbits inhaling concentrations of 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.
D. 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 pq/m or 0.79 jug/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
-------
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, et
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, Pimeghales grome-
las, 96-hour LC-0 values of 169,000 ,ug/l using static techniques and
103,000 ug/1 using flow-through tests with measured concentrations. The re-
ported 96-hour LCcQ value for the bluegill, Lepomis macrochirus, is 73,900
jug/1 in a static test (U.S. EPA, 1978b). Two 48-hour tests with Oaphnia
maqna resulted in EC5Q values of 11,600 and 79,000 jug/1, respectively
(Dill, et al.; U.S. EPA, 1978b). The 96-hour LC5Q values for the sheeps-
heaa minnow, Cyprinodon variegatus, and the tidewater silverside, 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, Mysidopsis
bahia, is reported to be 224,000 jjg/1 (U.S. EPA, 1978b).
8. Chronic Toxicity
An embryo-larval test with the fathead minnow resulted in no ad-
»
verse effects occurring at 2,800 pg/1, the highest test concentration (U.S.
EPA, 1978b).
-------
C. Plant Effects
The 96-hour EC5Q 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 712,000 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 1,1-dichloroethylans in
ambient water which will result in specified risk levels of human cancer:
Exposure Assumptions Risk S_evels with Corresponding Draft Criteria
(per"day)'
10-7 10-6 io-5
2 liters of drinking water 0.013 jug/1 0.13 ug/1 1.3 ug/1
and consumption of 18.7
grams fish and shellfish.
Consumption of fish and 0.21 ug/1 2.1 jug/1 21 jug/1
Shellfish only.
B. 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
»
ug/1 at any time. No saltwater criterion has been proposed because of in-
sufficient data.
-------
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 0. Little, Inc. April, 1976. Vinylidene 'chloride monomer emissions
from the monomer, polymer, and polymer processing industries, Arthur D. Lit-
tle, Inc., for the U.S. Environ. Prot. Agency, Research Triangle Park, N.C.
Sartsch, H., et al. 1975. Tissue-mediated rnutagenicity of vinylidene
chloride and 2-chlorobutadiene in Salmonella tyohimurium. Nature. 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-dichloroetnyl-
ene. Res. Comm. Cnem. Pathol. Fharmacol. 4: 553.
Dawson, G.W., et al. 1977. The acute toxicity of 47 industrial chemicals
to fresh and saltwater fishes. Jour. Hazard Mater. 1: 3G3.
Diil, O.C., et ai. Toxicity of 1,1-dichloroethylene (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.
Biochem. Pharmacol. 24: 2013.
Henschler, D. 1977. Metabolism and mutagenicity of halogenated olefins - A
comparison of structure and activity. Environ. Health Perspect. 21: 61.
Hushon, '3. and M. Kornreich. 1978. Air pollution assessment of vinylidene
chloride. EPA-450/3-78-015. U.S. Environ. Prot. Agency, Washington, D.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. Pharmacol. 6: 465..
Jaeger, R.L., et al. 1977. 1,1-Dichloroethylene hepatotoxicity: Proposed
mechanism of action of distribution and binding of l^C-radioactivity^ fol-
lowing inhalation exposure in rats. Environ. Health Perspect. 21: 113!
Jenkins, L.I., et al. 1972. Biochemical effects of 1,1-dichloroethylene in
rats: Comparison with carbon tetrachloride and 1,2-dichloroethylene. Toxi-
col. Appl. Pharmacol. 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 chlor-
ide. Research plan and e'arly results. Med. Lav. 68: 241.
McKenna, M.J., et al. 1978a. The pharmacokinetics of (l^c) vinylidene
chloride in rats following inhalation exposure. Toxicol. Appl. Fharmacol.
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, dichlorodifluoromethane, 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-
chloroethylene. 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 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 the
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)
-------
No. 72
trans—1,2-DIchloroethylene
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.
-------
TRANS-1,2-DICHLORETHYLENE
SUMMARY
There is little specific information available on trans-
1,2-dichloroethylene. This compound is quantitatively less
toxic than the 1,1-dichloroethylene isoraer; however, the
t
toxicity appears qualitatively the same with depression
of the central nervous system as well as liver and kidney
damage. Trans-1,2-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 LC5Q value for the bluegill is 135,000 ^ig/1 in a static
bioassay.
-------
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; C2H2C12;
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-1,2-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-1,2-dichloroethylene was found at a concen-
tration of 1 pg/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.
-------
III. PHARMACOKINETICS
A. Absorption
Animal or human studies do not appear to exist
which specifically document the degree of systemic absorp-
tion of trans-1,2-dichloroethylene by any route.
B. Distribution
Pertinent data could not be located in the avail-
able literature.
C. Metabolism
Trans-1,2-dichloroethylene is metabolized through
an epoxide intermediate to either a dichloroacetaldehyds
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-isoraer relative to the amount taken up
by the liver was much greater than the trans-isomer (McKenna,
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-1,2-dichloroethylene has been shown to be
negative in the £_._ coli K12 and Salmonella mutagenicity
assays (Greim, et al. 1975; Cerna and Kypenova, 1977).
-------
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. (1972) found trans-1,2-dichloroethylene to be consider-
ably less potent than 1,1-dichloroethylene.
V. AQUATIC TOXICITY
A. Acute Toxicity
The reported 96-hour LC50 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.
-------
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 ^jg/1 as a 24-hour aver-
age and not to exceed 1200 ^g/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.
Re s. 4 6: 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 pharmacokinetics of [14C]
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).
-873-
-------
No. 73
Mchloroethylenes
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.
-------
DICHLOROETHYLE'NES
Summary
Of the three dichloroethylene isomers, cis 1,2-dichloroethylene,
trans 1,2-dichloroethylene, and 1,1-dichlcrcethylene, 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-
dichloroethyiene are mutagenic 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-
dichloroethylene. Reported 96-hour LC50 values for the bluegill are 73,900
and 135,500 ug/1, respectively, for 1,1-dichloroehtylene and 1,2-di-
chloroethylene. Two observed 48-hour LC50 values for Daphnia exposed to
1,1-dichloroethylene range were 11,600 and 79,000 ug/1. All saltwater
fish and invertebrates tested with 1,1-dichloroethylene showed 96-hour
LC50 values over 224,000 ug/1, and all algae tested both in fresh and
saltwater, had 96-hour EC50 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.
-------
DICHLOROETHYLENES
I. INTRODUCTION
This profile is based on the draft Ambient Water Quality Criteria
Document for Dichloroethylenes (U.S. EPA, 1979).
The dichloroethylenes (C2H2C12; 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 jig/1, vapor pressures
».
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 ir. industry, with annual production prior tc 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
and in the production of polyvinylidene chloride copolymers (PVDCs).
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. Both cis and trans-
1,2-dichloroethylene 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-
j-
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
-277-
-------
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 isomers.
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 McKenna, et al. (1978b) suggests that in rats most,
if not all, of the orally administered dose is absorbed at two dose levels: 1
and 50 rag/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 liver, spleen, heart, and 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
-878-
-------
bonds with tissue macromolecules (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 microsoroal mixed function oxidase system is not clear, since
both inhibitors Cdithiocarbamate) 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-dichloroethyiene iscmer. It is postulated that the 1,1-dichlorc-
ethylene isomer 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
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 1,1-dichloroethylene by gavage 4 to 5
days a week for 52 weeks. The incidence was 42 percent in the treated
».
animals and 3*J 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. typhimurium
(Bartsch, et al. 1975) and E. ooli K12 (Greim,"et al. 1975); however,
both the cis and trans isomers of "i ,2-dichloroethylene were non-mutagenic
when assayed with E_._ coli K12. In order to demonstrate mutagenic activity,
1 ,1-dic'nloroethylene needed rnicrosomal activation. In addition, cis
1,2-dichloroethylene was mutagenic 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 al. 19Y7b; 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.
-------
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 glutathicne
associated with greater toxicity and elevated gluthathione associated
with decreased toxicity (Jaeger, et al. 1973, 1977).
V. AQUATIC TOXICITY
A. Acute Toxicity
All of the available data for dichloroethylene,. 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: 4^,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.
-------
The 96-hour LC50 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 >Jg/I (U.S. EPA, 1979).
C. Plant Effects
The 96-hour ECgo values based on cell numbers of the freshwater
algae, Salenestruni capricornvitim and the saltwater algae, Skeletonema
costatum, are 798,000 and 712,000 ug/1, respectively, fcr exposure to
1,1-dichloroethylene (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/m3 (1,2-dichloroethylene). These values allow daily
exposures of 286 ing 1,1-dichloroethylene per day and 5,6U3 mg 1,2-di-
chloroethylene per day. The U.S. EPA (1979) draft water criteria document
M-
for dichloroethylene states that no human health criterion could be derived
for cis- and trans-1,2-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
grains fish and shellfish
Consumption of fish and
shellfish only.
Risk levels and Correscondine Draft Criteria
10'
10
0.013 ug/1 0.13 ug/1
0.11 ug/1 2.1 ug/1
21
1.3 ug/1
ug/1
B. Aquatic
The -proposed draft criterion to protect freshwater species
from dichloroethylene toxicity are as follows (U.S. EPA, 1979):
Concentration not to be
exceeded at anytime
Compound
1,1-dichloroethylene
1,2-dichloroethylene
For saltwater species:
1,1-dichloroethylene
1,2-dichloroethylene
24-hr. Average
530 ug/1
620 ug/1
1,700 ug/1
Not available
1.200 ug/1
1,400 ug/1
3,900 ug/1
Not available
-m-
-------
DICHLOROETHYLENES
References
American Conference of Governmental Industrial Hygenists. 1977. Docu-
mentation of the threshold limit values. 3rd ed.
Anderson, D., et al. 1977. Dominant lethal studies with the halogenated
olefins vinyl chloride and vinylidene dichloride in male CD-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. '
Sartsch, H., "et al., 1975. Tissue-mediated mutagenicity of vinylidene
chloride and 2-chlorobutadisne in Salmonella typhimuriu;n. Nature 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-dichlcro-
ethylene. Res. Comm. Chem. Pathol. Pharmacol. 4: 553.
Carna, M., and H. Kypenova. 1977. The acute toxicity of 47 industrial
chemicals to fresh and saltwater fishes. Jour. Hazard. Mater. 1: 303.
Dill, D.C., et al. Toxicity of 1,l-dichloroethylene (vinylidene chloride)
to aquatic organisms. Dow Chemical Co. (Manuscript).
Greim, H., et al. 1975. Mutagenicity i_n vitro and potential carcino-
genicity of chlorinated ethylenes as a function of metabolic oxirane forma-
tion. Biochem. Pharmacol. 24: 2013.
Henschler, 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-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,l-dichloroethylene inhalation toxicity in rats.
Res. Comm. Chem. Pathol. Pharmacol. 6: 465.
*•
Jaeger, R.L., et al. 1977. 1,1-Oichloroethylene hepatotoxicity: Proposed
mechanism of action of distribution and binding of l^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. Pharmacol. 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. 68: 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: 189.
Prendergast, J.A., et al. 1967. Effects on experimental animals of long-
term inhalation of trichlorcethylsne, carbon tetrachloride, 1,1,1-trichloro-
ethane, dichlorodifluoromethane, and 1,1-dichloroethylene. Toxicol. Appl.
Pharmacol. 10: 270.
Rampy, L.W.,' et al. 1977. Interim results of a tv/o-year toxicolcgical
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-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.D., et al. 1977b. A dominant lethal study in male rats after
repeated exposure to vinyl chloride or vinylidene chloride. Jour. Toxicol.
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 Tor 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, D.C.
U.S. EPA. 1978b. List of organic compounds identified in U.S. drinking
water. Health Effects Res. Lab. U.S. Environ. Prot. Agency, Cincinnati,
Ohio.
-8S5-
-------
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).
-------
No. 74
Dichloromethane
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.
-m-
-------
DICHLOROMETHANE
SUMMARY
In humans, dichloromethane is a central nervous system
depressant resulting in narcosis at high concentrations.
Dichloromethane is metabolized to carbon monoxide and causes
an increase in carboxyhemoglobin. There is no information on
the human chronic toxicity or teratogenicity of dichloro-
methane. Dichloromethane was not shown to be a carcinogen in
the strain A mouse bioassay, although there was a significant
increase in tumor response. Dichloromethane has been shown
to be mutagenic to Salmonella, but not to _§. cerevis ia and
Drosophiia.
Aquatic organisms tend to be fairly resistant to
dichloronethane, with acute toxicity values ranging from
193,000 to 331,000 ug/1.
-------
DICHLOROMETHANE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Cri-
teria Document for Halomethanes (U.S. EPA, 1979a).
Dichloromethane (C^Cl^i niethylene chloride, methy-
lene dichloride, and methylene bichloride; molecular weight
84.93) is a colorless liquid with a melting point of -95.1°C,
a boiling point of 40°C, a specific gravity of 1.327 g/ml at '
20°C, a vapor pressure of 362.4 mm Hg at 20°C, and a solubil-
ity in water of 13.2 g/1 at 25°C. Dichloromethane is a com-
mon industrial solvent found in insecticides, metal cleaners,
paints, and paint and varnish removers (Balmer, et al.,
1976). In 1976, 244,129 metric tons were produced in the
U.S. with an additional 19,128 metric tons imported (U.S.
EPA, 1977). For additional information regarding the halo-
methanes as a class, the reader is referred to the Hazard
Profile on Haiomethanes (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 8 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, 19-75; 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) .
-------
B. Food
Pertinent information could not be located in the
available literature.
C. Inhalation
Reported background concentrations of dichlorometh-
ane in both continental and saltwater atmospheres were about
0.00012 mg/m3, and urban air concentrations ranged from
less than 0.00007 to 0.00005 mg/m3. Local indoor concen-
trations can be high due to the use of aerosol sprays or sol-
vents (Natl. Acad. Sci., 1978).
III. PHARMACOKIN ETICS
A. Absorption
Efficiencies of absorption of dichloromethane by
the lungs are between 31 to 75 percent, depending on length
of exposure, concentration, and activity level {Natl. Acad.
Sci.., 1978; Natl. Inst. Occup. Safety and Health, 1976).
B. Distribution
Upon inhalation and absorption, dichloromethane
levels increase rapidly in the bleed 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 con-
centrations 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 is involved in the formation of carboxyhemoglobin
(Natl. Inst. Occup. Safety and Health, 1976). Cardiorespira-
tory 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 car-
bon 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
excreted 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
Theiss and coworkers (1977) examined the tumori-
genic activity of dichloromethane in strain A mice. Dichloro-
methane at the low dose (1:5 dilution of the maximum toler-
ated dose) produced marginally significant increases in tumor
response, Shimkin and Stoner (1975) did not report a posi-
tive carcinogenic response for the strain A mouse bioassay
system.
B. Mutagenicity
Simmon, et al. (1977) reported that dichloromethane
was mutagenic to Salmonella typhimurium strain TA100 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.
Dichloromethane did not increase mitotic recombination in J3.
cerevisia D3 (Simmon, et al., 1977), and it was reported neg-
ative 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
Pertinent information could not be located in the
available literature.
D. Other Reproductive Effects
Gynecologic problems in female workers exposed for
long periods to gasoline and dichlororaethane vapors were re-
ported by Vosovaya (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 embryo-toxicity (Schwetz, et al., 1975; Natl. Inst.
Occup. Safety and Health, 1976).
E. Chronic Toxicity
Pertinent information could not be located in the
available literature.
F. Other Relevant Information
»
Acute exposures to dichloromethane produce central
nervous system disfunction, are irritating to mucous mem-
branes, and increase the level of carboxyhemoglobin (Natl.
-m-
-------
Acad. Sci., 1978). Price, et al. (1978) reported that Fis-
cher rat embryo cells (F1706) were transformed by dichloro-
methane at high concentrations (1.6 x 10~^M) in the growth
medium. However, Sivak (1978) indicated the presence of car-
cinogenic contaminants in the dichloromethane and could not
demonstrate transformation in the BALD/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 in-
vertebrates. LC50 values for th_e fathead minnow (Pime-
phales 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 (Lepom,is_ mac-
rochirus) in a static assay. Daphnia magna were reported as
having an LC5Q value of 224,000 ug/1 (U.S. EPA, 1979a).
For the marine fish, the sheepshead minnow (Cyprinodon
varlegatus) , an LC^Q of 331,000 ug/1 was obtained. The
marine mysid shrimp was reported as having an LC^Q value
of 256,000 ug/1.
8. Chronic Toxicity
Chronic tests for neither freshwater nor marine
species could not be located in the available literature.
C. Plant Effects
Both species of freshwater algae, Selenastrum cap-
ricornutum and marine algae, Skeletonema cornujiuiri, were
-------
equally resistant to dichloromethane, with ECcQ values in
excess of 662,000 ug/1.
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 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/m3; how-
ever, NIOSH (1976) has recommended a ten-hour, time-weighted
average exposure limit of 261 rng/m3. The U.S. EPA (I979a)
draft water quality criterion for dichloromethane is 2 ug/1.
The reader is referred to the Hale-methanes Hazard Profile foi
discussion of criteria derivation (U.S. EPA, 1979b).
B. Aquatic
Criterion for protecting freshwater aquatic life
have been drafted as 4,000 ug/1/ not to exceed 9,000 ug/1/
while marine criterion have been drafted as 1,900 ug/1/ not
co exceed 4,400 u,g/l.
-------
DICHLOROMETHANE
References
Salmer, M.F., et al. 1976. Effects in the liver of methylene chloride in-
haled 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 l^C-labeled methylene dichloride in rat. Scand. Jour.
Work Environ. Health 1: 104.
DiVincenzo, G.D., et al. 1972. Human and canine exposures to methylene
chloride vapor. Am. Ind. Hyg. Assoc. Jour. 33: 125.
Filippova, L.M., et al. 1967. Chemical mutagens. IV. Mutagenic activity
of geminal system. Genetika 8: 134.
Jongen, W.M.F., et al. 1978. Mutagenic effect of dichloromethane on
Salmonella typhlmurium. 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. 1578. Nonfluorinated halomethanes in the
environment. Washington, D.C.
National Institute for Occupational Safety and Health. 1976a. Criteria for
a recommenced standard: Occupational exposure to methylene chloride. HEW
Pub. No.-76-133. U.S. Dep. Health Edu. Welfare, Cincinnati, Ohio.
Occupational Safety and Health Administration. 1976. General industry
standards. OSHA 2206, revised January 1976. U.S. Oep. Labor. Washington,
D.C.
Price, P.J., et al. 1978. Transforming activities of trichloroethylene and
proposed Ind. alternatives. In Vitro 14: 290.
Schwetz, B.A., et al. 1975. The effect of maternally inhaled trichloro-
ethylene, perchloroethylene, methyl chloroform, and methylene chloride on
embryonal and fetal development in mice and rats. Toxicol. Appl.
Pharmacol. 32: 84.
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. In: S. Scott, et al., eds. Progress in genetic toxicology.
Sivak, A. 1978. BALB flash 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 contam-
inants 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, 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 Criteria, (Draft).
U.S. EPA 1979b. Environmental Criteria and Assessment Office. Halo-
methanes: Hazard Profile. (Draft).
Vozovaya, M.A. 1974. Gynecological illnesses in workers of major indus-
trial rubber products plants occupations. Gig. Tr. Sostoyanie Spetsifich-
eskikh Funkts. Tab. Neftekhim. Khim. Prom-sti. (Russian)56. (Abstract)
-------
No. 75
2,4-Dichlornphenol
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.
-------
2.4-DICHLOROPHENOL
Summary
Insufficient data exist to indicate that 2,4-dichlorophenol is a car-
cinogenic agent. 2,4-Dichlorophenol appears to act as a nonspecific irri-
tant in promoting tumors in skin painting studies. No information on muta-
genicity, teratogenicity, or chronic toxicity is available. In a subacute
study, the only adverse effect noted in mice was microscopic nonspecific
liver changes. 2,4-Oichlorophenol appears to be a weak uncoupler of oxida-
tive phosphorylation,
Acute toxic effects of 2,4-dichlorophenol have been observed at a con-
centrations ranging between 2,020 and 8,230 ug/1/animal species. Freshwater
plants seem to be more resistant. 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 jug/1 de-
pending on the species of fish consumed.
-too-
-------
2,4-DICHLOROPHENQL
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for 2,4-dichlorophenol (U.S. EPA, 1979).
2,4-Dichlorophenol is a colorless, crystalline solid having the empiri-
cal formula C^H^Cl^Q 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-Dichlorophenol is a commercially produced, substituted phenol used
entirely as an intermediate in the manufacture of industrial and agricultur-
al products such as the herbicide 2,4-dichlorophenoxyacetic acid (2,4-0),
germicides, and miticides.
Little data exists regarding the persistence of 2,4-dichlorophenol in
the environment. Its low vapor pressure and non-volatility from aqueous
alkaline solutions would cause it to be only slowly removed from surface
water via volatilization (U.S. EPA, 1979). Studies have indicated low ab-
sorption of 2,4-dichlorophenol from natural surface waters by. various clays
(Aly and Faust, 1964). 2,4-Dichlorophenol is photolabile in aqueous solu-
tions (Aly and Faust, 1964; Crosby and Tutass, 1966) and can be degraded
microbially to succinic acid in soils and aquatic environments (Alexander
and Aleem, 1961; Ingols, et al., 1966; Loos, et al., 1967).
-------
II. EXPOSURE
A. Water
Sources of 2,4-dichlorophenol in water are agricultural run-off (as
a contaminant and metabolic breakdown product of biocides) and manufacturing
waste discharges (U.S. EPA, 1979), Recent experiments under conditions sim-
ulating the natural environment have not demonstrated that 2,4-dichlorophe—
nol is a significant product resulting from chlorination of phenol-contain-
ing wastes (Glaze, et al. 1978; Jolley, et al. 1978).
B. Food
Contamination of food with 2,4-dichlorophenol would probably result
from use of the herbicide 2,4-D (U.S. EPA, 1979).
The U.S. EPA (1979) has estimated the weighted average bioconcen-
tration factor for 2,4-dichlorophenol to be 37 for the edible portions of
fish and shellfish consumed by Americans. This estimate is based on the
octanol/water partition coefficient.
C. Inhalation
Pertinent information regarding direct evidence indicating that
humans are exposed to significant amounts of 2,4-dichlorophenol through
inhalation has not been found in the available literature.
III. PHARMACOKINETICS
A. Absorption
Pertinent information regarding the absorption of 2,4-dichlorophe-
nol 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 2,4-0 (300 - 2000 jjg/g feed) to cattle and sheep (Clark, et al.
1975) and Nemacide (50 - 800 pg/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 admini-
stered 2,4-dichlorophenol was not found in the available literature. In
mice, urinary metabolites of C-labelled gamma or beta benzene hexachlor-
ide (hexachlorocylohexane) included 2,4-dichlorophenol 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 admi-
nistration 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. Carcinogenic!ty
t
Insufficient data exist to indicate that 2,4-dichlorophenol is a
carcinogenic agent. The only study performed (Boutwell and Bosch, 1959)
suggested that 2,4-dichlorophenol may promote skin cancer in mice after ini-
-703-
-------
tiation with dimethylbenzanthracene and when repeatedly applied at a concen-
tration high enough to damage the skin. An analysis of the data of Boutwell
and Bosch using the Fisher Exact Test indicated that the incidence of papil-
lomas in 2,4-dichlorophenol-treated groups was significantly elevated over
controls, while the incidence of carcinomas was not (U.S. EPA, 1979).
B. Mutagenicity, Teratogenicity and Other Reproductive Effects
No studies addressing the mutagenicity, teratogenicity or other
reproductive effects of 2,4-dichlorophenol in mammalian systems were found
in the available literature. However, genotoxic effects of 2,4-dichlorophe-
nol have been.reported in plants. Exposure of flower buds or root cells of
vetch, (Vicia fabia) to solutions of 2,4-dichlorcphenol, 0.1M 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, 1979).
C. Chronic Toxicity
One report (Sleiberg, et al. 1564) suggested that 2,4-dichlorophe-
nol was involved in inducing chloracne and porphyria in workers manufactur-
ing 2,4-dichlorophenol and 2,4,5-trichlorophenol and exposed to acetic acid,
phenol, monochloroacetic acid, and sodium hydroxide. Since various dioxins
(including one associated with chloracne) have been implicated as contami-
nants of 2,4,5-trichlorophenol, the 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 nonspeci-
-------
fie 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-Dichlorophenol appears to be a weak uncoupler of oxidative
phosphorylation (Farquharson, et al. 1958; Mitsuda, et al. 1963). Values on
odor threshold for 2,4-dichlorophenol in water range from 0.65 to 6.5 jug/1,
depending on the temperature of water (Hoak, 1957).
V. AQUATIC TOXICITY
A. Acute Toxicity
Two 96-hour assays have been performed examining the acute effects
of 2,4-dichlorophenol in freshwater fish. An LC5Q value of 2,020 /jg/1 for
the bluegill, Lepomis macrocharus, (U.S. EPA, 1978), and an LC5Q value of
3,230 ug/1 for the juvenile fathead minnow, Pimephales promsias, (Phipps, et
al. manuscript), have been reported. Two studies on the freshwater clado-
ceran, Daphnia magna, have produced 48-hour static LCcn values of 2,610
J3U
and 2,600 ^ug/1 (Kopperman, et al. 1974; U.S. EPA, 1978).
Only one marine fish or invertebrate species has been tested for
the acute effects cf 2,4-dichlorophenol. niatt, et al. (1553) observed only
a moderate reaction to a concentration of 20,000 ug/1 in mountain bass, a
species endemic to Hawaii.
B. Chronic Toxicity
Data for the chronic effects of 2,4-dichlorophenol for either
freshwater or marine organisms were not located in the available literature.
C. Plant Effects
Concentrations of 2,4-dichlorophenol that caused a 56 percent re-
duction in photosynthetic oxygen production or a complete destruction of
-------
chlorophyll were 50,000 or 100,000 pg/1, respectively, in algal assays with
Chlorella pyrenoidosa (Huang and Gloyna, 1968). An earlier study by Black-
man, et al. (1955) reported a concentration of 2,4-dichlorophenol that
caused a 50 percent reduction in chlorophyll to be 58,320 ug/1 in the duck-
weed, Lemna minor. No marine plant species have been examined.
D. Residues
A bioconcentration factor of 130 has been estimated from the octa-
nol-water partition coefficient of 2,4-dichlorophenol for aquatic organisms
having a lipid content of eight percent. The estimated weighted average
bioconcentration factor for the edible portion of aquatic organisms is 37.
E. Miscellaneous
Flavor impairment studies indicated that the highest concentration
of 2,4-dichlorophenol in the exposure water which would not cause tainting
of the edible portion of fish ranged from 0.4 jug/1 for the largemouth bass
(Micropterus salmoides), to 14 jug/I for the bluegill (Lepomis macrochirus).
The value for the rainbow trout (Salmo gairdneri) was 1 fig/1 (Shumway and
Palensky, 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.
A. Human
Based upon the prevention of adverse organoleptic effects, the
t
draft interim criterion for 2,4-dichlorophenol in water recommended by the
U.S. EPA (1979) is 0.5 ug/1, although the recommended draft interim criter-
ion could be 371 /ug/1 based on calculations by the U.S. EPA (1979) from sub-
acute toxicity data in mice.
-------
8. Aquatic
The draft criterion for protecting freshwater organisms is 0.4 jug/1
as a 24-hour average concentration, not to exceed 110 pg/1. No criterion
was derived for marine organisms (U.S. EPA, 1979).
-------
2.4-OICHLOROPHENOL
REFERENCES
Alexander, M. and M.I.H. Aleem. 1961. Effect of chemical structure on
microbial decomposition of aromatic herbicides. Jour. Agric. Food Chem.
9: 44.
Aly, O.M. and S.D. Faust. 1964. Studies on the fate of 2,4-0 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.
Arner, S.M. and E.M. Ali. 1969. Cytological effects of pesticides. IV.
Mitotic effects of some phenols. Cytologia 34: 533.
Amer, S.M. and E.M. Ali. 1974. Cytological effects of pesticides. V. Ef-
fect of some herbicides on Specia faba. Cytologia 33: 633.
Blackman, G.E., et al. 1955. The physiological activity of substituted
phenols. I. Relationships between chemical structure and physiological
activity. Arch. Biochem. Biophys. 54: 45.
Bleiberg, J.M., et al. 1964. Industrially acquired porphyria. Arch. Der-
matol. 89: 793.
Boutwell, R.K. and O.K. Bosch. 1959, The tumor-promoting action of phenol
and related compounds for mouse skin. Cancer Res. 19: 413.
Clark, O.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. Photodecompcsition of 2,4-dichlorophe-
noxyacetic 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: 20.
Glaze, W.H., et al. 1978. Analysis of new chlorinated organic compounds
formed by chlorination of municipal wastewater. Page 139 In: R.L. Jolley,
Ced.) Water chlorinaticn - environmental impact and health effects. Ann
Arbor Science Publishers.
Hiatt, R.W., et al. 1953. Effects of chemicals on schooling fish, Kuhlia
sandvicensis. Biol. Bull. 104: 28.
Hoak, R.D. 1957. The causes of tastes and odors in drinking water. Water
and Sew. Works. 104: 243.
-------
Huang, J. and E.F. Gloyna. 1968. Effect of organic compounds on photosyn-
thetic oxygenation. I. Chlorophyll destruction and suppression of photosyn-
thetic oxygen production. Water Res. 2: 347.
Huff, J.E. and J.S. Wassom. 1974. Health hazards from chemical impurities:
chlorinated dibenzodioxins and chlorinated dibenzofurans. Int. Jour. Envi-
ron. 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. Md. Soc. Toho, Japan. 19: 356.
Kopperman, H.L., et al. 1974. Aqueous Chlorination and ozonation studies.
I. Structure-toxicity correlations of phenolic compounds to Daphnia maqna.
Chem. Biol. Interact. 9: 245.
Kurihara, N. 1975. Urinary metabolites from and B-Bh'C in the mouse:
chlorophenolic conjugates. Environ. Qual. Saf. 4: 56.
Loos, M.H., et al. I967b. Phenoxyacetate herbicide detoxication by bacter-
ial enzymes. Jour. Agric. Food Chem. 15: 358.
Mitsuda, W., et al. 1963. Effect of chlorophenol analogues on the oxida-
tive phosphorylation in rat liver mitochondria. Agric. Biol. Chem. 27: 366.
Phipps, G.L., et al. The acute toxicity of phenol and substituted phenols
to the fathead minnow. (Manuscript)
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 haloand nitrophe-
nols. 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 nema-
cide [o-(2,4-dichlorophenol)-o,o-diethylphosphorothipate] to laying hens.
Jour/ Agric. Food Chem. 20: 617.
Shumway, D.L. and J.R. Palensky. 1973. Impairment of the flavor of fish by
water pollutants. EPA-R3-73-010. U.S. Environ. Prot. Agency.
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.
-909-
-------
U.S. EPA. 1979. 2,4-Dichlorophenol: Ambient Water Quality Criteria.
(Draft).
Weast, R.C., ed. 1975. Handbook of chemistry and physics. 55th ed. CRC
Press, Cleveland, Ohio.
-------
No. 76
2,6-uichlorophenol
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.
-------
2,6-DICHLGROPHENOL
Summary
There is no available information on the possible carcinogenic, terato-
genic, or adverse reproductive effects of 2,6-dichlorophenol.
The compound did not show mutagenic activity in the Ames assay. A sin-
gle report has indicated that 2,6-dichlorophenol produced chromosome aberra-
tions in rat bone marrow cells; details of this study were not available for
evaluation.
Prolonged administration of 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 meta-
chlorophenol, 2,^, 5-trichlorophenol, and 2,3,4,6-tetrachlorophenol.
-9/3-
-------
I. INTRODUCTION
2,6-Oichlorophenol, CAS registry number 87-65-0, exists as white nee-
dles and has a strong penetrating odor resembling o-chlorophenol. It has
the following physical and chemical constants (Weast, 1972; Hawley, 1971):
Formula: C6H4Ci20
Molecular Weight: 163
Melting Point: 68°C - 69°C
Boiling Point: 219°C - 220°C (740 torr)
Vapor Pressure: 1 torr @ 59.5°C
pH: 6.79
^. -• *
Production: unknown "'
2,6-Oichlorophenol 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-Oichloro-
phenol has a taste and odor threshold of 0.002 mg/1 and 0.003 mg/1, respec-
tively (McKee and Wolf, 1963). Piet and OeGrunt (1975) found unspecified
dichlorophenols in Dutch surface waters at 0.01 to 1.5 ug/1, and Burtt-
schell, et al. (1959) demonstrated that chlorination of phenol-containing
water produced, among other products, 2,6-dichlorophenol in a 25-percent
yield after 18 hours of reaction.
8. Food
p
Pertinent data could not be located in the available literature.
-------
C. Inhalation
Olie, et al. (1977) reported finding dichlorophenols in flje gas
condensates from municipal incinerators. The levels were not quantified.
0. 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, J56A).
III. PHARMACOKINETICS
A. Absorption
Pertinent data could not be located in the available literature.
Sy comparison with other chlorophenols, it is expected that 2,6-dichlorophe-
nol will be absorbed through the skin and from the gastrointestinal tract
(U.S. EPA, 1979).
B. Distribution
Pertinent data could not be located in the available literature.
The high lipid solubility of the compound would suggest that unexcreted com-
pound 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 expected that 2,6-dichlorophe-
nol is rapidly eliminated from the body, primarily as urinary sulfate and
glucuronide conjugates (U.S. EPA, 1979).
IV. EFFECTS
A. Carcinogenicity ,
Pertinent data could not be located in the available literature.
-------
B. Mutagenicity
2,6-Oichlorophenol 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 Toxicity
Administration of 2,6-dichlorophenol to rats (route and dosage not
specified) has been reported to produce hepatic degeneration (Chung, 1978).
E. Other Relevant Information
tests have indicated that 2,6-dichlorophenol will inhibit
liver mitochondria! respiration (level not specified) (Chung, 1978).
V. AQUATIC TOXICITY
A. Acute
McLeese, et al. (1979) reported a 52-hour lethal threshold limit
of 19,100 ug/1 for marine shrimp ( Cranggn septemsainosa) exposed to 2,6-di-
chlorophenol.
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-dichlorophenol, a
s
water quality criterion of 3.0 ug/1 has been recommended by the U.S. EPA
( 1979) .
B. Aquatic
No existing criteria to protect fresh and saltwater organisms were
found in the available literature.
-------
REFERENCES
Burttschell, R.H., et al. 1959. Chlorine derivatives of phenol causing
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.
Ooldens, J.D. 1964. Chlorophenols. Ir\i Kirk-Othmer Encyclopedia of Chemi-
cal 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.
Hoakt 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.
Mcteese, D.W., V. Zitko and M.R. Peterson. 1979. Structure-lethality rela-
tionships 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 incerators in the
Netherlands. Chemosphere 8: 445.
Piet, G.J. and F. OeGrunt. 1975. Organic chloro compounds in surface and
drinking water of the Netherlands in problems raised by the contamination of
man and his environment. Comm. Eur. Communities, Luxemborg, p. 81.
Rasanen, (_., 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. Toxicol. 18: 565.
U.S. EPA. 1979. Chlorinated phenols: Ambient water quality criteria.
Washington, O.C. U.S. Environmental Protection Agency. (Draft)
Weast, R.C. 1972. Handbook of Chemistry and Physics, 53rd ed. Chemical
Rubber Co., Cleveland, Ohio.
-------
No. 77
2,4-Oichlorophenoxyacetic Acid (2,4-D)
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.
-9/9-
-------
2,4-DICHLOROPHENOXYACETIC ACID
Summary
Oral administration of 2,4-Dichlorophenoxyacetic acid (2,4-D) failed to
produce carcinogenic effects in mice or dogs; however, feeding technical
grade 2,4-D did produce tumors in a study with rats. Subcutaneous adminis-
tration of the isooctyl ester of 2,4-0 has been reported to produce r'eticu-
lum cell sarcomas 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, Drosophila, Saccharomyces, or the dominant
lethal assay with mice.
2,4-D and several of its esters failed to show teratogenic effects in
mice; the oropylene glycol butyl ether ester of the compound produced an in-
crease in cleft palates in this study. Studies in hamsters orally adminis-
tered 2,4-D 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-D to rats found teratogenic ef-
fects. A three-generation feeding study of 2,4-D to rats indicated feta-
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-D are more toxic than the 2,4-D acid, di-
methyl amine, or sodium salt. Freshwater trout and bluegill sunfish were
adversely affected by the propylene glycol butylether (PGBE) ester at con-
centrations of 900 to 2,000 jjg/1. Daphnids and freshwater seed shrimp were
sensitive to the PGBE ester at concentrations of 100 to 300 ug/1. Chrpnic
exposure of several species of fish to concentrations up to 310 pg/1 has not
demonstrated any toxic effect.
-------
2,4-OICHLOROPHENOXYACETIC ACID
I. INTRODUCTION
2,4-Dichlorophenoxyacetic acid, CAS Registry number 94-75-7, commonly
known as 2,4-D, is a white or slightly yellow crystalline compound which is
odorless when pure. 2,4-D has the following physical and chemical proper-
ties (Herbicide Handbook, 1979):
Formula:
Molecular Weight:
Melting Point:
Soiling Point:
Density:
Vapor Pressure:
Solubility;
221.0
135°C-138GC (technical);
140°C-141°C (pure)
160°C © 0.4 torr
1.56530
0.4 torr H 160°c
Acetone, alcohol, dioxane ether,
isopropyl alcohol; slightly
soluble in benzene, solubility in
water 0.09g/lQOg, H20
Production: Unknown
2,4-D 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-D from water are
*
available; however, after treatment for water milfoil in reservoirs in
-------
Alabama and Tennessee, the Tennessee Valley Authority found the concentra-
tion at downstream monitoring stations to be 2 ppb. 2,4-D 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-D 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 2,4-0 ester. The highest tissue concentration reached was 0.24 ppm
eight days after application. Subsequently, the herbicide or its metabolite
was eliminated rapidly. Clams and .oysters accumulate more 2,4-0 than co
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-D esters which are much more volatile than the parent com-
pound have been monitored in air up 'to 0.13 ug/m5 (Farwell, et al. 1976;
Stanley, et al. 1971).
D. Dermal
Pertinent data were not found in the available literature.
-------
III. PHARMACOKINETICS
A. Absorption
Human absorption of 2,4-D 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-D esters may be less efficient than
that of the free acid or salt form of the compound (NRCC, 1978).
B. 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 blood (MRCC, 1978). Withdrawal cf dietary compound produced almost
complete tissue loss of residues in seven days (Clark, et al. 1975).
Small amounts of phenoxy herbicides are passed to the young
through the mother's milk (Sjerke, et al. 1972). Transplacental transfer cf
2,4-D has been reported in mice (Lindquist and Ullberg, 1971).
C. Metabolism
Sauerhoff, et al. (1976) determined that following oral adminis-
tration of 2,4-D 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-D 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-D 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, iscpropyl, and isccctyl esters in a long-term carcinogenicity study.
Carcinogenic effects were seen after subcutaneous administration of the iso-
octyl ester (reticulum cell sarcomas) (MCI, 196S).
8. Mutagenicity
No mutagenic effects of 2,4-D in tests with Salmonella,
Saccharomyces, or Drosophila were observed (Fahrig, 1974). Siebert and
Lemperle (1974) have reported mutsgsnic effects following treatment of
Saccharomyces cerevisiae strain 04 with aqueous 2,4-D solution (1,000 mg/1).
Gavage or intraperitoneal administration of 2,4-D to mice failed
to show mutagenic effects in the dominant lethal assay (Epstein, et al.
1972).
,*
C. Teratogenicity
Testing of 2,4-D and its n-butyl, isopropyl, and isooctyl esters
in pregnant mice produced no significant teratogenic effects. There was a
-------
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-D in pregnant mice has been reported to produce terato-
genic effects (Caujolle, et al. 1967), although the DMSO vehicle used is,
itself, a teratogen. Bage, et al. (1973) have also reported teratogenic ef-
fects in mice following injection of 2,4-D.
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 McKinley, 1972). However, Schwetz, et al. (1971) were un-
able to show teratogenic effects in rats following the oral administration
of 2,4-0 or its iscocytol or propylene glycol butyl ether esters.
D. Other Reproductive Effects
Embryotoxic effects following subcutaneous administration of 2,4-0
to pregnant mice have been reported (Caujolle, et al. 1967; Bage, et al.
1973).
Fetotoxic effects of the compound and its esters have been report-
ed after oral administration of maximally tolerated doses (Schwetz, et al.
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 (1966) 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 and Erne,
1971; Bjorn 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 (Dow Chem-
ical, 1962).
V. AQUATIC TOXICITY
A. Acute Toxicity
V
The National Research Council of Canada (1978) has reviewed the
toxic effects of 2,4-0 to fish. For the bluegill sunfish (Lepomis
macrochirus), 2,4-D acid and 2,4-0 dimethyl amine produced toxic effects at
concentrations greater than 100,000 pg/1. At 2,4-D concentrations of 50,000
ug/1 or less, no increased mortalities were reported except in pink salmon.
The isopropyl, butyl, ethyl, butoxy ethanoi, and PGBE esters produced
48-hour LC50 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. et al.
(1974) conducted tests of various formulations of 2,4-D on 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 pg/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 LC50 value of 1,100 jug/i on exposure to
-------
the PGBE ester of 2,4-D. Harlequin fish (Rasbora heteromorpha) showed a
48-hour LC value of 1,000 pg/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 LC5Q values of 26,700; 40,000; 70,100; 70,700;
94,600; 96,500; and 300,600 jug/1 for banded killifish (Funduius diaphanus),
white perch (Roccus americanus), stripped bass (Morone sazatilis), guppies
(Libistes reticulatus), pumpkinseed sunfish (Lepomis gibbosus), carp
(Cyprinus carpio), and American eel (Anguilla rostrata), respectively,
exposed to commercial technical grade 2,4-D.
Sanders (1970) conducted a comparative study on the toxicities of
various formulations of 2,4-D for six species cf freshwater crustaceans.
The PGBE ester was generally most toxic, while the dimethylamine salt was
least toxic. The crayfish (Orconectes nails) was the most resistant species
tested, with 48-hour static LC5Q values greater than 100,000 ^jg/1 for all
formulations tested. The waterflea (Daphnia magna) and seed shrimp
(Cypridopsis vidua) were most sensitive to the PGBE ester, with 43-hour
LC5g values of 100 and 320 jug/1, respectively. Scuds (Gammarus
fasciatus), sowbugs (Ascellus brevicaudus), and freshwater grass shrimp
(Palaemonetes kadiakensis) were also moderately sensitive, with 48-hour
LC5Q values ranging from 2,200 to 2,700 jug/1. Sanders and Cope (1968)
reported a 96-hour LC5Q value of 1,600 jug/1 for stonefly naiads
(Pteronarcy californica) exposed to the butoxyethanol ester of 2,4-D. Tech-
nical grade 2,4-D produced a 96-hour LC50 value of 14,000 jug/1. Robertson
and Bunting (1976) reported 96-hour LC5Q values ranging from 5,320 to
ll,570jug/l for copepods (Cyclops vernalis) nauplli exposed to 2,4-D as free
acid. The range of 96-hour LC5Q values for nauplli exposed to 2,4-D alko-
nolamine salt was 120,000 to 167,000 jug/1.
-------
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 jjg/1 for the PGBE ester of 2,4-D. The 2,4-0 acid
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-D
acid at a concentration of 2,000 jug/1 for 48-hours. Sudak and Claff (1960)
found a 96-hour LC5Q vaiue Of 5,000,000 jug/1 for fiddler crabs (Uca
pugmax) exposed to 2,4-D.
McKee and Wolf (1963) have reviewed the toxic effects of 2,4-D to
aquatic organisms. Toxic concentrations as lev/ as 1,000 ug/1 produced a 40
percent mortality for fingerling bluegills exposed to 2,4-0 butyl ester. In
general, esters of 2,4-D were reported to be more toxic than sodium salts of
2,4-0.
B. Chronic Toxicity
Rehwoldt, et al. (1970) exposed several species of fish to 100
iug/1 2,4-0 for ten months and observed no overt effects to any tested
species. The percent reduction of brain acetylcolinesterase ranged from 16
percent in white perch to 35 percent in American eels. In breeding experi-
ments with guppies, a 100 jug/1 concentration of 2,4-D had no significant ef-
fect on the reproductive process of the species under experimental condi-
tions. Cope, et al. (1970) examined the chronic effects of PGBE ester of
2,4-0 to bluegill sunfish. Fish were exposed to the herbicide in one-eighth
f
acre ponds containing initial concentrations of up to 10,000 jug/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
-------
days. Mount and Stephan (1967) exposed 1-inch fathead minnows (Pimephales
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 jjg/1 1,4-0 (acid equivalent in Weedone
LV-4 formulation).
*.
C. Plant Effects
The genera Microcystis, 5cened85rnus_, Chlorella, and Nitzschia
showed no toxic response when exposed to 2,000 jug/1 2,4-D Lawrence (1962).
Poorman (1973) treated cultures of Euglena qracilis with concentrations of
50,000 jug/1 2,4-D for 24 hours and ooserved depressed growth rates.
Valentine and Bingham (1974) demonstrated that at 100,000 pg/1, 2,4-D re-
duced the cell numbers of Scenedesrcus to one percent of control levels,
Chlamydomonas to 48 percent of control levels, Chlorells to 66 percent of
control levels, and Euglena to 90 percent of control levels within 4 to 12
days. The bluegreen algae (Nostoc muscorum) displayed a 68-percent reduc-
tion in growth when exposed to 100 pg/1 2,4-D (Cenci and Cavazzir.i, 1973).
Singh (1974) exposed Cylindrospermum to 2,4-D sodium salt at concentrations
ranging from 100,000 to 1,200,000 pg/1 and reported that concentrations
above 800,000 pg/1 caused growth to cease completely. McKee and Wolf (1963)
reviewed the effectiveness of 2,4-D in control of emergent aquatic plants
and reported that concentrations ranging from 6,000 to 100,000 /jg/1 have
been effective in controlling a number of species.
-------
D. Residue
Cope, et al. (1970) examined residues of the PGBE ester of 2,4-D
in the freshwater vascular plant, Potamoqeten nodosus, in a one-eighth acre
pond treated with single 100 to 10,000 pg/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-D 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 pg/1 level
realized in the water two weeks after application.
The National Research Council of Canada (NRCC) (1973) has reviewed
the bioconcentration data and associated residues of 2,4-D 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 cf 100 to 200
jug/1, the bioconcentration of 2,^-D various aquatic invertebrates v/as one to
two orders of magnitude greater than in the water. Oysters (Crassostica
virqinica) were reported to have a bioconcentration factor of 180 when ex-
posed to the butoxyethanol ester of 2,4-0. The freshwater bluegili and mos-
quito fish (Gambusia affinis) had bioconcentration factors ranging frcm 7 to
55, respective to water concentrations. Fish fed a diet containing 2,4-0
bioconcentrated the 2,4-0 acid by less than 0.2.
VI. EXISTING GUIDELINES
A. Human
The acceptable daily intake of 2,4-D for humans has been estab-
lished at 0.3 mg/kg (FAQ, 1969).
B. Aquatic
Pertinent data were not found in the available literature.
-------
2.4-DICHLOROPHENOXYACETIC ACID
References
Bage, et al. 1973. Teratogenic and embryotoxic effects of herbicides diand
trichlorophenoxyacetic acids (2,4-0 and 2,4,5-T). Acta Pharmacol. Toxicol.
32: 408.
Bjerke, 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. lexicological 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 Fishary Investigations. U.S. Dept. Interior
U.S. Fish and Wildlife Service Circ. 'l67: 11.
Sutler, P.A. 1965. Effects of herbicides on estuarine fauna. Proc.
Southern Weed Conference 18: 576.
Caujolle, F., et al. 1967. Limits of toxic and tsrstogenic tolerance of
dirr.sthyl sulfc-xids. Ann. N.Y. Acad. Sci. 141: 110.
Cenci, P. and G. Cavazzini. 1973. Interaction between environmental micro-
flora and three herbicidal phenoxy derivatives. Ig. Mod. 66: 451.
Clark, D., et al. 1975. Residues of chlorophenoxy acid herbicides and
their phenolic metabolites in tissues of sheep .and cattle. Jour. Agric.
Food Chem. 22: 573.
Clerk, D., et al. 1964. The fate of 2,4-dichlorophenoxyacetic acid in
sheep. Jour. Agric. Food Chem. 12: 43.
Collins, T., and C. Williams. 1971. Teratogenic studies with 2,4,5-T and
2,4-D in the hamster. Bull. Environ. Contamin. Toxicol. 6: 559.
Cope, O.B., et al. 1970. Some chronic effects of 2,4-D in the bluegill
(Leporciis macrochirus) Trans. Am. Fish Sec. 99: 1.
Courtney, K. 1974. In: The herbicide 2,4-D. U.S. Environmental Protec-
tion Agency, Office of Pesticides Programs, Washington,' DC. 207 pp.
Day, B.E., et al. 1978. The phenoxy herbicides. Council for Agricultural
Science and Technology, Report 77. '
-------
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. Dow Chemical Co., Midland, MI.
Epstein, S., et al. 1972. Detection of chemical mutagens by the dominant ,
lethal assay in the mouse. Toxicol. Appl. Pharmacol. 23: 288.
Food and Agriculture Organization of the United Nations (FAQ). 1569. Work-
ing party of experts on pesticide residues. Evaluations at seme pesticide
residues in food, the monographs. FAO/V/HQ PL 1968/m/9/l.
Fahrig, R. 1974. Comparative mutagenicity studies with pesticides. Chem-
ical Carcinogenesis Assays. IARC Scientific Publication 10: 161. Lyon.
Farwell, S.O. et al. 1976. Survey of airborne 2,4-D in south central
Washington. Jour. Air Pollut. Control Assoc. 26: 224.
Gangst, E.O. 1979. Herbicide Residue of 2,4-D, Office of Chief of Engine-
ers, Washington, D.C. NTIS AD-67160.
Hansen, W. et al. 1971. Chronic toxicity of 2,4-dichlorophenoxyacetic acid
in rats and dogs. Toxicol. Apl. Pharmacol. 20: 122.
Herbicide Handbook. 1979. 4th ed. Weed Science Society of America,
Champaign, IL. p. 129.
Innes, J., et al. 1569. Bioassay of pesticides and industrial chemicals
for tumoriaencity in mice: A preliminary note. Jour. Natl. Cancer Instit.
42: 1101.
Fhera, K. and W. McKinley. 1972. Prs and postnatal studies on 2,4,5-tri-
cnlorophenoxyacetic acid, 2,4-dichlorophenoxyacetic acid and their deriva-
tives in rats. Toxicol. Appl. Pharmacol. 22: 14.
Kohli, J., et al. 1974. Absorption and excretion of 2,4-dichlorophenoxy-
acetic acid in man. Xenobictica, 4: 97.
Lawrence, J.M. 1962. Aquatic Herbicide Data. U.S. Dept. of Agriculture,
Agriculture Handbook No. 231, 133 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-D to
salmonids in southeast Alaska. Jour. Fish Red. Bd. Canada 31: 480.
Mount, D.I. and C.E. Stephen, 1967. A method for establishing acceptable
toxicant limits for fish - malathion and the butoxy ethanol ester of 2,4-D.
Trans. Am. Fish Soc. 96: 185.
-------
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 mutagenic 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. Contam. 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-
water crustaceans. Int. Jcur. Water Pcllut. Control Fed. 42: 1544.
Sanders, H.O. and O.B. Cope. 1968. The relative toxicities of several pes-
ticides to naiads of three species of stoneflies. Limnol and Oceanogr.
13: 112.
Sauernoff, M., et al. 1976. The fate of 2,4-dichlorcphenoxyacetic acid
(2,4-D) fcllov/ing oral administration to man. Toxicol. Appl. Fharmacol.
37: 136.
Scnwetz, B., et al. 1971. The effect of 2,4-dichlorcphenoxyacetic acid
(2,4-0) and esters of 2,4-D on rat embryonal, foetal, and neonatal growth
and development. Food Cosmet. Toxicol. 9: 801.
Schultz, O.P. and E.O. Gangstad. 1976. Dissipation of residues of 2,4-D in
water, hydrosoil, and fish. Jour. Aquae. Plant Managa. 14: 43
Siebert, D. and E. Lemperle. 1974. Genetic effects of herbicides: induc-
tion of mitotic gene conversion in Saccharcmyres cerevisiae. Mut. Res.
22: 111.
Sigmon, C.F. 1979. Influence of 2,4-D 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.
-933-
-------
Sudak, F.N. and C.L. Claff. 1960. Survival of Uca pugnax in sand, water,
and vegetation contaminated with 2,4-dichlorophenoxyacetic acid. Proc.
Northeast Weed Cont. Conf. 14: 508.
Valentine, J.P. and S.W. Bingham. 1974. Influence of several algae on
2,4-D residues in water. Weed Sci. 22: 358.
-------
No. 78
1,2-Dichloropropane
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
froir. 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.
-------
1,2-DICHLOROPROPANE
Summary
The major environmental source of dichloropropane is from the use of a
mixture of dichloropropanes 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. Dichloropropanes 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_n values for the bluegill are
280,000 and 320,000 jjg/1 and the 48-hour LC5Q value for Daphnia maqna is
52,500 jjg/1. A saltwater fish has a reported 96-hour LC5Q value of
240,000 jug/1.
-937- ^
-------
1,2-OICHLOROPROPANE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Dichloropropanes/Dichloropropenes (U.S. EPA, 1979).
1,2-Dichloropropane (1,2-FDC, 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-I,3-dichloropropene are used
as soil fumigants. For the purposes of discussion in this hazard profile
document, dichloropropane refers to the 1,2-dichloropropane isomer. When
heated to decomposition temperatures, 1,2-dichioropropane emits highly toxic
fumes of phosgene (Sax, 1975).
II. EXPOSURE
A. Water
Dichloropropane can enter the aquatic environment as discharges
from industrial and manufacturing processes, as run-off from agricultural
land, and from municipal effluents. This compound was identified but not
quantified in New Orleans drinking water (Dowty, et al. 1975).
B. 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-
»
propane. The weighted average BCF for edible portions of all aquatic organ-
isms consumed by Americans is calculated to be 5.8.
-------
C. Inhalation
Atmospheric levels of dichloropropane 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 (Thomas
and McKeury, 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 urine in 24 hours (Hutson, et al. 1971).
IV. EFFECTS
A. Carcinogenicity
Only one study is reported on the carcinogenicity 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. Oichloropropane has also been shown to cause mutations in A. nidulans
-93?"
-------
(Bignami, et al. (1977), and to cause chromosomal aberrations in rat bone
marrow (Dragusanu and Goldstein, 1975).
C. Teratogenicity
Pertinent information could not be located in available literature
searches regarding teratogenicity.
D. 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 dichlorcpropane 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 (givsn five days per week). The only effect observed was a decreased
weight in rats.
V. AQUATIC TOXICITY .
A. Acute Toxicity
Two observed 96-hour LC-n values for the bluegill, Lepomis
macrochirus, upon exposure to 1,2-dichloropropane were 280,000 and 320,000
pg/1 (Dawscn, et al. 1977; U.S. EPA, 1978). In the only freshwater inverte-
brate study reported, the 48-hour LC5Q for Oaphnia magna is 52,500 ug/1
(U.S. EPA, 1979). Tidewater silverside, (Menidia bevyllina), has an
observed 96- hour LC5Q of 240,000/jg/l (Dawson, et al. 1977).
^
B. Chronic Toxicity
Chronic data are not available for any saltwater or freshwater
species.
-------
C. Plant Effects
The phytotoxicity of 1,2-dichloropropane has not been investigated.
D. 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 ppm (350 mg/m ) (Am. Ccnf. Gov.
Ind. Hyg., 1977). The draft water criteria for dichloropropane is 203 ug/1
(U.S. EPA, 1979).
3. Aquatic
Fcr 1,2-dichloropropane, the proposed draft criteria to protect
freshwater aquatic life are 920 pg/1 a 24-hour average and the concentration
should not exceed 2,100 ug/1 at any time. Criteria are not available for
saltwater species (U.S. EPA, 1979).
-9*1-
-------
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 sone pesticides: The use
of Salmonella typhimurium and Aspergillus nidulans. Mutag.
" 4~6~: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.
Dowtyf B., et al. 1975. Halogenated hydrocarbons in Mew
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. Epideniol. Pneumofitziol. Ig 24: 37.
Ueppel, L.A., et al. 1943. Toxicology of 1,2-diohloropro-
pane (propylene dichloride) IV. Effect of repeated exposures
to a low concentration of the vapor. Jour. Ind. Kyg. Toxi-
col. 30: 189
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, Acric. 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.
-------
U.S. EPA. 1979a. Dichloropropenes/Dichloropropanes: Ambient
Water Quality Criteria. (Draft).
U.S. EPA. 1979b. Dichloropropenes/Dichloropropanes: Hazard
Profile.
-------
No. 79
DIchloropropane/Dichloropropenes
A
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.
-------
DICHLOROPROPANES/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 dichlocopropane 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
LC-Q values for the bluegill, Lepomis macrochirus, for 1,1-,
i,2-7 and 1,3-cichlorcpropane are 97,900, 280,000, and greater
than 520,000 ug/1, respectively. For Daphnjia magna, the corres-
ponding reported 48-hour LC--, values are 23,000, 52,000, and
282,000 |ig/l, respectively. Similar results have been obtained
with marine organisms.
The dichlorcpropenes are considerably more toxic in acute
exposure than the dichloropropanes. For 1,3-dichlorpropene,
the 96-hour LC5Q value for the bluegill is 6,060 ;jg/l compared
to 520,000 ug/1 for 1,3-dichloropropane. For Daphnia magna,
the corresponding values are 6,150 and 282,000 jjg/1, respectively.
The SC50, based on chlorophyll a for a freshwater alga, is 4,950
,ug/l 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.
-------
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-
pene gives off toxic fumes of chlorides (Sax, 1975). Production
of mixtures of dichloropropanes/dichloropropenes approached 60
million pounds per year prior to 1975 (U.S. EPA, 1979).
II. EXPOSURE
A. Water
Dichloropropanes 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
-------
(1979) has estimated the weighted average bioconcentration fac-
tors (BCFs) of dichloropropanes and dichloropropenes to range
between 2.9 and 5.3 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 dichloro-
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)..
III. PHARiMACOKINETICS ;;
A. Absorption, Distribution and Metabolism
Pertinent information regarding the absorption, dis-
tribution, and metabdlisrr. of the dichioroprcpanes and dichlcroprc-
penes could not be located in the available information.
3. Excretion
No human data are available on the excretion of dichlor-
opropanes or dichloropropenes. In the rat, 80 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 dichloropropenes could not be located
-------
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 40Q ppm of 1,2-dichloropropane for 128 to 140 daily 7-hour
periods (5 days per week) decreased normal weight gain in rats
{Heppel, 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 for 1,3-dichloropropene is
6,060 pg/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
pg/1, respectively, (U.S. SPA, 1978) for the only freshwater
invertebrate species tested, Dapfania magna. The 48-hour LC5Q
value for 1,3-dichloropropene and Daphnia magna under static
conditions is 6,150 ug/1 (U.S. SPA, 1978).
The 96-hour LC50 values for the saltwater sheepshead
minnow, Cyprinpdgn yariegatus, exposed to 1,3-dichloropropane ^
and 1,3-dichloropropene were 36,700 ^ug/1 and 1,770 pg/1, respec-
tively (U.S. SPA, 1978). Dawson, et al. (1977) obtained a 96-
hour LC5Q of 240,000 pg/2. for the tidewater silversida, Menidia
beryllina, for exposure to i,2-dichloropropane.
For Mysidopsis 5ahia, the 96-hour LCqQ for 1,3-dichloro-
propene was one-thirteenth that for 1,3-dichloropropane, i.e.,
790 jig/1 and 10,300 pg/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 pg/1 (U.S. EPA, 1978). The chronic value
*
for mysid shrimp, Mysidopsis bahia, was 3,040 jag/1 for 1,3-di-
chloropropane in a life cycle study (U.S. EPA, 1978) .
-------
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 capr icojrnutum, were 4,950 ug/1 and
4,960 jug/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.
D. Residues
Measured steady-state bioconcentration factors (BCF)
are not available for any dichloropropane or dichicropropene
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) .
t
The distribution of 1,3-dichloropropene within soils depends
on soil conditions. For example, cis-1,3-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 ppm (350 mg/m )
{Am. Conf. Gov. Ind. Hyg . , 1977). The draft water criterion
(U.S. EPA, 1979) for dichloropropane is 203 ug/1, The draft
water criterion for dichloropropenes is 0.63 ug/1 (U.S. EPA,
1979).
3. Aquatic
The draft criteria for the dichloropropanes and ci-
chlcroprcper.es to protect freshwater aquatic life are as follows
(U.S. EPA, 1979) :
Concentration not
to be exceeded
Compound 2 j -go u r A v e r a q e at any time
1,1-dichloropropane 410 ug/1 930 ug/1
1,2-dichloropropane 920 ug/1 2,100 ug/1
l,3-dichlOE:opropane 4,800 pg/1 11,000 ug/1
1,3-dichloropropene 18 ^g/1 250 pg/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 pg/1 910 ug/1
1,3-dichloropropane 79 pg/1 180 pg/1
1,3-dichloropropene 5.5 pg/1 14 ;ag/l
-------
OICHLORDPROPANES/DICHLOROPRQPENE5
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.
Dawson, G.W., et al. 1977. The acute toxicity of 47 industrial chemicals
to fresh and saltwater fishes. Jour. Hazard. Mater. 1: 303.
Oowty, B., et al. 1975. Halogenated hydrocarbons in New Orleans drinking
water and blood plasma. Science 87: 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.
Kutscn, Q.H., et al^. 1971. Excretion and retention of components of the
soil fumigant D-CnR> and their metabolites in the rat. Food' Cosmet. Toxi-
col. 9: 677.
Leistra, M. 1970. Distribution of 1,3-dichloropropene over the phase in
soil. Jour. Agric. Food Chern. IS: 1124.
Roberts, R.T. and G. Stoydin. 1976. The degradation of (2)- and
(E)-l,3-di- chloropropenes and 1,2-dichioropropsnes in soil. Psstic. Sci.
7: 325.
Sax, N.I. 1975. Dangerous properties of industrial materials. Reinhold
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 Duuren, B.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.
-------
No. 80
Dichloropropanol
t
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.
-------
OICHLOROPROPANOL
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.
A. -957-
-------
I. INTRODUCTION
This profile is based on computerized searches of Toxline, Biosis, and
Chemical Abstracts, and review of other appropriate information sources as
available. Oichloropropancl (molecular weight 128.9), a colorless, viscous
liquid with a chloroform-like odor, refers to four isomers with the mole-
cular formula C3H6oci2. The physical properties of each isomer are
given below.
Boiling Point Density Solubility (Weast, 1976)
Water Alcohol Ether
2,3-Oichloro-l-propanol 182°c 1.368 slight miscible miscible
l,3-Oichloro-2-propanol 1740C 1.367 very very miscible
3,3-Dichloro-l-propanol 82-83°C 1.316 not listed
1, l-Dichloro-2-propanol 146-1^8^0 1.3334 sliaht verv very
Additional physical data and synonyms of the above isomers are avail-
able in Heilbron (1965), Fairchild (1979), Sax (1979), Windholz (1976), and
Verschueren (1977).
Cichloropropanol is prepared from glycerol, acetic acid, 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 be a moderate fire hazard when exposed to heat, 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 Belyakov, 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.
Bioa'ccumulation 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. Carcinogenlcity
Pertinent data could not be located in the available literature.
B. Mutagenicity
2,3-Oichloropropanol and 1,3-dichlcroprcpanol were evaluated for
mutagenicity by a modified Ames assay using S_._ typhimurium strains. Some
evidence of mutagenie activity was seen, but the authors felt that further
evidence and clarification of the metabolic activation pathway to mutagens
via halcalkar.cls were necessary (Nakarnura, et al. 1979).
C. Teratogenicity, Other Reproductive Effects and Chronic Toxicity
Pertinent data could not be located in the available literature.
0. Acute Toxicity
2,3-Qichloropropanol was found to have an oral LD^g j_n the rat
of 90 mg/kg. The lowest published lethal concentration (LC. ) j_n 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-Oichloropropanol was
found to have an oral L050 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 LD5Q. 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 renal toxicity, including fatty
degeneration and necrosis of the renal tubular epithelium (Sax, 1979; Cos-
selin, et al. 1976).
V. AQUATIC TOXICITY
Data concerning the effects of dichloropropanol to aquatic organisms
were not found 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 sng/m^ (Lipina and Belyakov,
1975).
The maximum allowable concentration in Class I waters for the pro-
duction of drinking water is i mg/i (Verschueren, 1977).
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 halohydrin
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 far 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-
ly 1 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 tvohimurium, Mutat. Res.
66: 373.
Sax, N.I. 1979. Dangerous Properties of Industrial Materials. Van Nos-
trand Reinhold Co., New York.
Verschueren, K. 1977. Handbook of Environmental Data on Organic Chemicals.
Van Nostrand Rsinhold Co., New York, p. 659.
V/east, R.C. (sd.) 1976. Handbook of Chemistry and Physics. CRC Press,
Cleveland, Ohio, p. c-454.
Windholz, M. (ed.) 1976. The Merck Index. 9th ed. Merck and Co., Rahway,
New Jersey.
-------
No. 81
1, 3-Dichloroproper.e
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 fumigant. 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.
The bluegill (Leoomis macrochirus) has a reported 96-hr LC5g value of
6060 JJQ/1: papnnia ma ana has a reported 48-hr LC.~ of 6150 Jjg/i. For the
•^ — ——i—i - ~ - - " ^U
saltwater invertebrate, Mysidoosis bahia, a reported 96-hr LC5Q value is
790 ^g/1. In the only long-term study available, the value obtained for
1,3-dichloropropene toxicity to fathead minnows (Pimephales promeles) in an
embryo-larval test is 122 jug/1. Based on chlorophyll a concentrations and
cell numbers, the 96-hr EC50 values for the freshwater alga Selenastrum
cspricgrnutum are 4,950 and 4,960 Jjg/1, respectively; for the marine alga
Skeletonema costatum, the respective values are 1,000 and 1,040 jug/1.
-964-
-------
1,3-OICHLOROPROPENE
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-l,3-dichloropropene. For
more information regarding the dichloropropenes, ths reader is referred to
the EPA/ECAO Hazard Profile on Dichloropropanes/Dichloropropenes (U.S. EPA,
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
P
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
-9(f-
-------
Americans. This estimate is based on the octanol/water partition coeffi-
cient of dichloropropene.
C. Inhalation
Atmospheric levels of dichlcropropene 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. PHARMACCKINETICS
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 acniinisterea a'ose of dichloropropene was eliminated in
the urine within 24 hours (Hutson, et al. 1971).
TW F~r~ r-i— f~i-r- j—
IV. hrrcuii
A. Carcinogenicity
Van Duuren, 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.
j-
8. Mutagenicity
DeLorenzo, et al. (1977) and Neudecker, et al.' (1977) reported £hat
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
-------
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 3.
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, Lecctnis macrcchirus, yialded a 36-hr LC
value of 6060 jjg/1 for 1,3-dichloropropene exposure. For Daphnia maqna, the
48-hr LC5Q value is 6,150 jug/1 (U.S. EPA, 1978). The observed 96-hr
LC^Q for the saltwater myrid shrimp, Mysidopsis pahia, is 790 jug/1 (U.S.
EPA, 1978).
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 }jg/l (U.S. EPA, 1979a).
C. Plant Effects
»
Based on chlorophyll a concentrations and cell numbers, the 96-hr
EC50 vaiues for tne freshwater alga, Selenestrum caoricornutum, are 4,950
-967-
-------
and 4,960 pg/1, respectively (U.S. EPA, 1978). The respective values for
the saltwater alga Skeletonema costaturn were 1,QOO and 1,040 pg/1 (U.S. EPA,
1978).
D. 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, movement'-of 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 hydrolyzed in moist soils to the cor-
responding cis-3-chloroalkyl alcohol, which can be microbially degraded to
carbon dioxide and water by Pseudo^onas sp. (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 jug/1 as a
24-hr average not to exceed 250 ,ug/l at any time. For marine species, the
•
value is 5.57jg/l as a 24-hr average not to exceed 14 ug/1 at any time (U.S.
EPA, 1979).
-------
1,3-DICHLOROPROPENE
REFERENCES
DeLorenzo, F., et al. 1977. Mutagenicity of pesticides containing 1,3-di-
chloropropene. Cancer Res. 37: 6
Dowty, 3., 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 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.
Neudecker, T., et al. 1977. In vitro mutagenicity of the soil nematocide,
1,3-dichloropropene. Experientia 33: 8.
Sax, N.I. 1975. Dangerous Properties of Industrial Materials. Reinhold
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. Hour. Am. Ind. Hyg.
Assoc. 38: 217.
U.S. EPA. 1973. In-depth studies on health and environmental impacts of
selected water pollutants. Contract NO. 68-01-4646.
U.S. EPA. 1979a. Oichloropropanes/Dichloropropenes: 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).
-------
No. 82
Dieldrin
Health and Environaental 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-uracy.
-------
SPECIAL NOTATION
U.S. EPA1s Carcinogen Assessment Group (CAG) has evaluated
dieldrin and has found sufficient evidence to indicate that
this compound is carcinogenic.
-973.-
-------
DIELDRIN
SUMMARY
Dieldrin is a compound belonging to the group or 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 likely to be a human carcinogen.
Dieldrin has been found to be non-mucagenic in several test sys-
tems. The WHO'S acceptable daily intake for dieldrin is 0.0001
mg/kg/day.
The toxicity of dieldrin to aquatic organisms has been
investigated in numerous studies. The 96-hour LC,-0 values for
the common freshwater fish range from 1.1 to 360 pg/1. The acute
toxicity is considerably more varied for freshwater invertebrates,
with yo-hour ^^Q values ranging from 0.5 jjg/1 for the stonefly
to 7^0 jag/I for the crayfish. Acute !>CrQ values for eight salt-
water fish species range from 0.6b to 24.0 pg/1 in flow-through
tests; LC values for estuarine invertebrates range from 0.70
to 2vQ Jjg/l. The only reported cnronic values are 0.11 ug/1
for steel head trout (Salmo guirdnes) in an embryolarval study
and 0.4 pg/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 and inverteorates.
Bioconcentration factors were 128 for a freshwater alga, 1395
£or Daphnia magna, 2y93 for the channel catfish, ana 8000 for
i^—^.^_^_ ^
the edible tissues of the Eastern oyster.
-------
DIELDRIN
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.
Dieldrin is extremely stable and persistant in the environ-
ment. Its pe'rsistance is due to its extremely low volatility
(1.78 x 10~ 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, and 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-
f
tion of dieldrin is restricted in the United States, formulated
products containing dieldrin are imported from Europe (U.S.
EPA, 1979).
-------
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 (Harris, et al. 1977) . The
half-life of dieldrin in water, 1 meter in depth, has been esti-
mated to be 723 days (MacKay and Wolkoff, 1973).
B. Food
Dieldrin is one of the most stable and persistant
organochlorine pesticides (Nash and Woolscn, 1967), and because
it is lipophilic, it accumulates in the food chain (Wurster,
1971). its persistance 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
-9 "75-
-------
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 other parts of the world.
D. Dermal
Dermal exposure to dieldrin is limited to those in-
volved in its manufacture or application as a pesticide. Wolfe,
et al.' (1972) reported 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
risk o£ exposure.
III. PHARMACOKINETICS
A. Absorption
The absorption o-f 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, 1964). These authors also demonstrated
that absorption takes place via the portal vein, and that dieldrin
-976-
-------
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 a1- 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 cose related
and may vary with the sex of the animal {Walker, et al. 1969).
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 predorainatly 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 ug/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 mean 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,
1968). Correlations between the dose and length of exposure
to dieldrin and the concentration of dieldrin in the blood and
-977-
-------
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 feces in the
form of several compounds more polar than the parent compound
{u.S. EPA, 1979). Bedford and Hut son (1976) summarized the four
known metabolic products of dieldrin in rodents as 5,7-trans-
dihydroxy-dihydro-aldr in (trans-diol) and tri-cyclic dicarborylic
acid (both of which are products of the transformation of the
epoxy group), the syn-12-hydroxy-dieldr in (a mono-hydro deriva-
tive) , and the pentachloroketone. Male rats have been found
to ir.etabolize dieldrin more rapidly chan females (U.S. EPA, 1979),
and differences in the metabolism of dieldrin have been found
between species (Baldwin, et al. 1972) .
D. Excretion
Dieldrin is excreted mainly in the-- feces and, to some
extent, in the urine in the form of several polar metabolites
_ . *
(U.S. EPA, 1979). However, rabbits fed 14C-dieldrin over a 21
week period excreted 42 percent of the radioactivity by the end
of 22 weeks, with 2 to 3 times as much excreted in the urine
-97$-
-------
as in the feces. Robinson, et al. (1969) found that 99 percent
of the dieldcin 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 (Geuto and Biros, 1967}. The half-life
for dieldrin 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-life tc be 265 days. Because there is a rela-
tionship between the concentration of dieldrin in the blood and
that in adipose and other tissues, it seems likely that the half-
life in the blood may reflect the over-all half-life in other
tissues (U.S. EPA, 1979).
IV EFFECTS
A, Careinogenicitv
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;
-------
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 (1973). 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 (Fabric, 1973; Bidwell, et al.
1975; Marshall, et al. 1976) . A host-mediated assay and a domi-
nant lethal test, also yielded negative results (Bidwell, et
al. 1975). Majumdar, et al. (1977), however, found dieldrin
to be mutagenic in 3. typh iph imu r_ i^um, although these positive
results were 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
with mouse bcne marrow ceils and j.n v_it£O with human lung cells
treated with 1 mg/kg and 1 ^ig/ml dieldrin, respectively (Majumdar,
et al. 1976).
D. Teratogenicity
In 1967, Hathaway, et al. established that 14C-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)
-------
in hamsters and mice when administered, in single oral doses dur-
ing gestation (hamsters 50, 30, 5 rug/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 teratogenic 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 dielcrin for day 6 through 14 of gestation
at doses, already mentioned in this report (Ottolenghi, et al.
1974) .
E. Chronic Toxicity
The other effects produced by chronic administratior.
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. Era, 1979).
F. Other Relevant Information
Since aldrin and dieldrin are metabolized by way of
the mixed function oxidase (MFO) 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 doses prior
*
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, et al. 1972).
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 jjg/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; T-arzwell and Henderson,
1957) ; and 8 to 32 jjg/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 /jg/1 for the stone fly (Sanders and Ccpe..
1968) to 740 ,ug/l for the crayfish (Sanders, 1972).
The acute LC-,, values for eiaht saltwater fish scecies
Ju - L
range from 0.65 to 24.0 jag/I in flow-through tests (Butler, 1963;
Earnest and Benville, 1972; Korn and Earnest, 1974; Parrish,
et al. 1973; Schoettger, 1970; and Lowe, undated). LC5Q values
ranging from C. / to 240. u ^ug/1 have been reoorted for estuarian
invertebrates species, with the 'most sensitive species tested
being the commercially important pink shrimp, Penaeuji 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
JS -
-------
and Shumway, 1969). For the guppy, Poecilia rgticulata, in a
life-cycle test, the chronic value is 0.4 jag/1 (Roelofs, 1971).
C. Plant Effects
Freshwater plants are less sensitive to dieldrin than
freshwater fish or invertebrates. For example, a concentration
of 100 pg/1 caused a 22 percent reduction in the biomass of the
alga Scenedlesmus gu ad r i c a u d a ta {Stadnyk and Campbell, 1971),
and 12,800 ^g/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 ^ig/1 (Batterton, et al. 1971).
D. Residues
Bioconcentration factors (BCF) have been determined
for 9 freshwater species (U.S. EPA, 1978). Representative BCF
values are 128 for the alga, Sj^jncJessnius obl_iguus (Reinert, 1972,
1395 for Daphn^ji magna (Reinert, 1972), 2385-2993 for the channel
catfish, l£ta_lu_rus punctatus (Shannon, lS77a; 1977b) and 68,268
for the yearling lake trout, Saj. v el _i n u g_.._nj-_rna_y_c us h (Reinert, et
al. 1974). The edible tissue of the Eastern oyster, Crassistrea
Zi£SAHi££' 'Uia^ a 3CF v^iuS o£ 8000 after 392 days of exposure
(Parrish, 1974). Spot, Leiostonms 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.
-------
A. Human
The current exposure level for dieldrin set by OSHA
is an air time-weighted average of 250 ug/m for skin absorption
(37 FR 22139). In 1969, 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 mg/kg/day (Mrak, 1959).
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 cancec risk for humans below 1Q~ .
B. Aquatic
The draft criterion to protect freshwater life is
0.0019 ug/1 as a 24-hour average; the concentration should not
exceed 1.2 ug at any time. To protect saltwater aquatic life,
the draft criterion is 0.0069 ug/1 as a 24-hour average; the
concentration should not exceed 0.16 pg/1 at any time.
-------
DIELDRIN
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-??7-
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cides. Arch. Environ. Health 25: 29.
Wong, D.T., and L.C. Terriere. 1965. Epoxidation of aldrin,
isodrin, and heptachlor by rat liver microsomes. Biochem.
Pharmacol. 14: 375.
Wurster, C.F. 1971. Aldrin and dieldrin. Environment 13:
33.
'990-
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Mo. 83
o,o-Diethyl Dithiophosphoric Acid
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
99t-
-------
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.
-993-
-------
o.o-OIETHYL 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 Drbs'ophila, 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 malathion is suggested.
-------
I. INTRODUCTION
o,o-Diethyl 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: azinphosmethyl, carbophenothion, dialifor, dioxathion, disulfoton,
ethion, phorate, phosalone and terbufos. It is made from phosphorus penta-
'siilfide (SRI, 1976).
II. EXPOSURE
A. Water
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. 1979), or as a contaminant
of any of the above pesticides for which it is a starting compound.
B. Food
Pertinent data v/ere not found in the available literature; how-
ever, if present in feed, the compound would probably originate from the
same sources discussed above. Organophosphorus pesticide residues have been
found in food (Vettcrazzi, 1976).
C. Inhalation
Pertinent data were not found in the available literature; how-
ever, major exposure could come from fugitive emissions in manufacturing
facilities.
D. Dermal
Pertinent data were not found in the available literature.
-------
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 thiojmethyl 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, 1253).
C. Metabolism
Pertinent data were not found in the available literature. Metab-
olism studies with disulfoton (Bull, 1965) and phorate (Bowman and Casida,
1958) indicate that both compounds are converted to diethyl phosphorcdithio-
ate, diethyl phorphorothioate, and diethyl phosphate.
D. Excretion
Pertinent data were not found in the available literature. Based
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
-------
long-term feeding. No carcinogenic effects were noted in either species
(NCI, 1978).
3. Mutagenicity
Diethyl phosphorothioate, a possible metabolite of the parent com-
pound, did not show mutagenic activity in Drosophila, £. 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 phorate (0.5 to 3.0 ppm) indicated
that the highest level of compound did produce seme adverse reproductive
effects (American Cyanamid, 1966). Chronic feeding of mice with technical
dioxathion at levels of 45G to 600 ppm produced seme testiscular atrophy
(NCI, 1978).
E. Chronic Toxicity
Chronic feeding of technical dioxathion produced hyperplastic
nodules in livers of male mice. c,o-Diethyl dithicphosphoric acid, like
other organophosphates, is expected to produce cholinesterase inhibition
(NAS, 1977).
V. AQUATIC TOXICITY
A. Acute
v
Marking (1977) reports on LC^Q value of 47.2 >ug/l for rainbow
trout (Salmo gairdneri) exposed to the dithiodioxane analogue • of
bis(o,o-diethyl dithiophosphoric acid), dioxathion, and an LCrg value of
-------
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.
-997-
-------
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: 333.
Bull, D. 1965. Metabolism of di-systox by insects, isolated cotton leaves,
and rats. J. Econ. Entomol. 58: 249. •
Oaughton, 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.
v
Fahrig, R. 1974. Comparative mutagenicity studies with pesticides.
Chemical Carcinogenesis Assays, IARC Scientific Publication #10, p. 161.
Gaines, T. 1569. Acute toxicity of pesticides. Toxicol. Appl. Pharmacol.
14: 515.
Marking, L.L. 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. Toxicolcay of Insecticides. New York: Plenum Press,
p. 223.
National Academy of Sciences 1977. Drinking Water and Health, National
Research Council, Washington, p. 615.
National Cancer Institute 1973. Bicassay of Dioxathion for Possible
Carcinogenicity. U.S. OHEW, NCI Carcinoaenesis Technical Report Series
#125, 44 pp.
Richert, E. and K. Prshlad 1972. Effect of the organophosphste o,o-diethyl
s-[(etnylthio)methyi] pnosphorooithioate on the chick. Poult. Sci. 51: 613.
SRI 1976. Chemical Economics Handbook. Stanford Research Institute.
Pesticides, July 1976.
Vettorazzi, G. 1976. State of the art on the toxicological 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.
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No. 84
o,o-Diethyl-^-raethyl Phosphorodlthloate
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.
-1000-
-------
0,o-OIETHYL-S-METHYL PHOSPHORODITHIOATE
Summary
There is no available information on the possible carcinogenic, muta-
genic, teratogsnic 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 Drgsppjiila, E. coii, or Saccharomyces. o.o-Diethyl-S-methyl
phosphorodithioate, like other organophosphate compounds, is expected to
produce cholinestarass inhibition in humans.
There is no available data on the aquatic toxicity of this compound.
-1601-
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0,0-DIETHYL-S-METHYL PHQSPHORQDITHIOATE
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:ll5324p). 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: c^13
Molecular Weight: 200
Boiling Point: lOOoc to 1Q2°C (4 torr)
(CA 55:8335h)
Density: 1.192420
(CA 55:8335h)
Pertinent data were not found in the available literature with respect
to production, consumption or the current use of this compound.
II. EXPOSURE
Pertinent data were not found in the available literature.
III. PHARMACOKINETICS
A. Absorption
Information relating specifically to the absorption of o,o-di-
ethyl-S-methyl phosphorodithioate was not found in the available liter-
ature. Oral administration of the S-ethylthio derivative of this compound,
the insecticide phorate, indicates that this derivative is absorbed from the
gastrointestinal tract (Bowman and Casida, 1958).
f
6. Distribution
Pertinent data were not found in the available literature.
Studies with 32p radiolabelled phorate in the cow indicated that following
oral administration, residues were found in the liver, kidney, lung,
-IQ03L-
-------
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 organophosphates in mammals, o,c-diethyl-
S-methyl phosphorodithioate may be expected to undergo hydrolysis to diethyl
phosphorodithioic acid, diethyl phosphorotnioic acid, and diethyl phosphoric
acid (Matsumura, 1975).
D. 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 phorate
(Bowman and Casida, 1958).
IV. EFFECTS
A. Carcinogenicity
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 were noted in either species (NCI, 1978).
3. Mutagenicity
Pertinent data were not found in the available literature.
Oiethyl phosphorothioate, a possible metabolite of the parent compound, did
not show mutagenic activity in Orosphila, £. coli, or Saccharomyces (Fahrig,
1974).
-1003 -
-------
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 pom) 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 male mice. o,o-Qiethyl-S-methyl phosphoroaithioate, like other
organophosphates, is expected to produce cholinesterase inhibition (MAS,
1977).
V. AQUATIC TOXICITY
Pertinent data were not found in the available literature.
VI. EXISTING GUIDELINES AND STANDARDS
Existing guidelines and standards were not found in the available
literature.
-------
0,o-OIETHYL-S-METHYL PHOSPHORQDITHIOATE
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, 0. 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. 1973. Bioassay of Dioxathion for Possible Car-
cinogenicity. DHEW. National Cancer. Institute. Carcinogenesis Technical
Report Series No. 125: 44. • -
Richert, E.P. and K.V. Prahlad. 1972. Effect of the organcphospnate
o,G-diethyl-5-i(ethylthio)metnyl] phosphorodithiate on the chick. Poult.
Sci. 51: 613.
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No. 85
Diethyl Phthalate
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D-C. 20460
APRIL 30, 1980
1006-
-------
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.
-/007-
-------
DIETHYL PHTHALATE
SUMMARY
Diethyl phthalate has been shown to produce mutagenic
effects in the Ames Salmonella assay.
Teratogenic effects were reported following i.p. admin-
istration of diethyl phthalate to pregnant rats. This same
study has also indicated fetal toxicity and increased resorp-
tions after i.p. administration of DEP.
Evidence that diethyl phthalate produces carcinogenic
effects has not been found.
A single clinical report indicates that the development
of hepatitis in several hemodialysis patients may have been
related to leaching of diethyl phthalate from the plastic
tubings utilized.
Diethyl phthalate appears to be more toxic for marine
species acutely tested, with a concentration of 7,590 ug/1
being reported as the LCcn in marine invertebrates. The
data base for the toxic effects of diethyl phthalates to
aquatic organisms is insufficient to draft criterion for
their protection.
-------
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 diethyl phthalate was:
3.75 x 103 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 ppm 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 rag/m (Milkov,
et 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. PHARMACOKINETICS
Specific information is not available on the absorp-
tion, metabolism, distribution, or excretion of DEP. 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 information could not be located in
the available literature.
B. Mutagenicity
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.
f
injection has been reported to produce teratogenic effects
(Singh, et al. 1972).
-------
D. Other Reproductive Effects
Fetal toxicity and increased resorptions were
produced following i.p. injection of pregnant rats with
DEP (Singh, et al. 1972).
S. 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 macrochirus, has been shown to be acutely sensitive
to diethvl ohthalate; a 96-hour static LCqn of 98,200 ug/1
~ j U '
is reported (U.S. EPA, 1978). For the freshwater inverte-
brate, Daphnia magna, a 48-hour static IjC5Q of 51,100 pg/1
was obtained. Marine organisms proved to be more sensitive,
with the sheepshead minnow, Cyprincdon var iegatjjjs, showing
a 96-hour static LC5Q of 29,600 ug/1, while the mysid shrimp,
Mysidopsis bahia, showed an 96-hour static LCcQ of 7,590
ug/1 (U.S. EPA, 1978).
B. Chronic Toxicity
Pertinent information could' not be located in
the available literature.
9
C. Plant Effects
Effective concentrations based on chlorophyl a
content and cell number for the freshwater alga, Selena-
'/Off'
-------
strum capricornutum, ranged from 85,600 to 90,300 pg/1,
while the marine alga, Skej.etoneina cos tat urn, was more sensi-
tive, with effective concentrations ranging from 65,500
to 85,000 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
Sased on "no effect" levels observed in chronic
feeding studies with rats or dogs, the U.S. EPA has calcu-
lated an acceptable daily intake (ADI) level of 438 mg/day
for DEP.
The recommended water quality criterion level
for protection of human health is 60 mg/1 for D£? (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).
-------
DIETHYL PHTHALATES
REFERENCES
Engelhardt, G. et al. 1975. The microbial metabolism of
di-n-butyl phthalate and related dialkyl phthalates. Bull.
Environ. Contam. Toxicoi. 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. 86
Dimethylnitrosamine
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-
-------
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 pg dimethylnitrosamine in
their diet.
-------
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).
Diraethylnitrosamine 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 been detected at a concen-
tration of 3 to 4 pg/1 in v;astewater 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. Food
Dimethyinitrosamine was found to be present in
a variety of foods (including smoked, dried or salted fish,
cheese, salami, frankfurters, and cured meats} in the 1
to 10 u/kg range and occasionally at levels up to 100 ug/kg
(Montesano and Bartsch, 1976).
The U.S. EPA (1979a) has estimated the weighted
f
average bioconcentration factor for dimethylnitrosamine
for the edible portions of fish and shellfish consumed by
'/OI7
-------
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. PHARMACOKINETIC S
A. Absorption ..
Pertinent data could not be located in the avail-
able literature.
B. Distribution
Following intravenous injection into rats, dimethyl-
nitrosamine is rapidly and rather uniformly distributed
throughout the body (Mages, 1972).
C. Metabolism and Excretion
In vitrx) studies have demonstrated that the organs
in the rat with the major capacity for metabolism of dimethyl-
nitrosamine are the liver and kidney (Montesano and Magee,
1974). After administration of C-labeled-dimethylnitro-
samine to rats or mice, about 60 percent of the isotope
appears as CO- within 12 hours, while 4 percent is excreted
-------
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. Mutagenicity
Dimethylnitrosamine and diethylnitrosaraine have
been reported to induce forward and reverse mutations in
S. fcyphimur ium, E. coli, Neurospora crassa and other organisms;
gene recombination and conversion in Saccharomyces cerevisiae;
"recessive lethal mutation" in Drosophila; and chromosome
aberracions in mammalian cells (Montesano and Bamsch, 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 dimethvlnitrosamine.
-IOU-
-------
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 dimethyl-
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, no mention was made in any reports about plant effects
or residues.
One study reported that Shasta strain rainbow trout
(Saimo ga irdneri) , 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 800 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
JQ3L 0
-------
through the process of public review; therefore, there is
a possibility that these criteria may be changed,
A. Human
The U.S. SPA (1979a) has estimated that the water
concentrations of dimethylnitrosamine corresponding to life-
—5 —6 —7
time cancer risks for humans of 10 , 10 , or 10 are
0.026 ug/1, 0.0026 ug/1, and 0.00026 ug/1, respectively.
B. 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. Cold 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. In: W. Nakahara, et al., eds.
Topics in chemical carcinogenelTs. University of Tokyo
Press, Tokyo.
Magee, P.M., and J.M. Barnes. 1-956. The production of ma-
lignant primary hepatic tumors in the rat by feeding dimethyl-
nitrosamine. Sr, Jour. Cancer 10: 114.
Magee, P.N., and J.M. Barnes. 1959. The experimental pro-
duction of tumors in the rat by dimethylnitrosamine (N-nitro-
sod i.methylamine) . Acta. Un. Int. Cancer 15: 137.
Magee, P.M., et ai.. 1976. N-N itroso 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. Sartsch. 1976. Mutagenic and carcino-
genic N-nitroso compounds: possible environmental hazards.
Mutat. Res. 32: 179.
Montesano, R. , and P.N. Magee. 1974. Comparative metabolism
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 152;, 83.
Schoental, 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., 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. I979b. Environmental Criteria and Assessment Of-
fice. Nitrosamines; Hazard Profile,
-------
No. 87
2,4-Dlmethylphenol
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.
-------
2.4-DIMETHYLPHENQL
Summary
2,4-Dimethylphenol (2,4-QMP) 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 LC5Q value for fathead minnows is 16,750 jug/1;
chronic value using embryo-larval stages of the same species is 1,100 ug/1.
Daohnia maqna has an observed 48-hour LC_Q value of 2,120 jug/1. In
limited testing, one aquatic alga ana duckweed are over 100 times less
sensitive than the Daphnia in acute exposures. The bioconcentration factor
for 2,4- diiTiethylphenoi is 150 for the bluegill. From half-life studies,
residues of the chemical are not a potential hazard for aquatic species.
'/03L6
-------
I. INTRODUCTION
This profile is based primarily on the Ambient Water Quality Criteria
Document for 2,4-Oimethylphenol (U.S. EPA, 1979).
2,4-Dimethylphenol (2,M3MP) is derived from coal and petroleum sources
and occurs naturally in some plants. 2,4-QMP (CgH^O) 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
21Q°C (at 760 mm Hg), a vapor pressure of 1 mm Hg at 52.8°C, and a dens-
ity of 0.0965 g/ml at 20°C (U.S. EPA, 1979).
2,4-OMP is a weak acid (pka-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-OMP 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-
OMP. 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
f
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
-102?-
/
-------
and on the efficiency of water treatment systems in removing phenolic com-
pounds. In the U.S., the gross annual discharge of 2,4-DMP into waters was
estimated to be 100 tons in 1975 (Versar, 1975). Manufacturing' was the lar-
gest source of the discharge. Lsachates 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-
DMP was 55.5 jug/1.
B. Food
DMP's occur naturally in tea (Kaiser, 1967), tobacco (Saggett and
Morie, 1973; Spears, 1963), marijuana (Hoffmann, et al. 1975), and a conifer
(Gcrnostaeva, 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. EFA (1975) has estimated the weighted average biocon-
csntration factor for 2,4-DMP to be 340 for the edible portions of fish and
.shellfish consumed by Americans. This estimate is based on the measured
steady-state bioconcentration studies in the bluegill.
C. Inhalation
2,4-Qimethylphenol has been found in commercial degreasing agents
(NIOSH, 1978), cresol vapors (Corcos, 1939), cigarette smoke condensates
(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 20.8 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
dimethylphenol exposure per se (NIOSH, 1978).
D. Dermal .....
Absorption through the skin is thought to be the primary route of
human exposure to complex mixtures containing 2,4-DMP (U.S. EPA, 1979).
III. PHARMACQKINETICS
A. Absorption
2,4-DMP is readily absorbed through the skin (U.S. EPA, 1579). The
dermal LD5Q for molten 2,4-DMP 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-DMP given
orally to rats for eight months caused damage to the liver, spleen, kidneys,
and heart (Maazik, 1968).
C. Metabolism
Urinary metabolites, resulting from oral administration of 850 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
j
dimethylphenol positional isomers was reported.
D. Excretion
»
A study done on rabbits by Bray, et al. (1950) indicates rapid
metabolism and excretion of 2,4-OMP.
-------
IV. EFFECTS
A. Carcinogenic!ty
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 (Boutwell and Bosch,
1959). In a related study, Boutwell and Bosch (1959) applied 25 jjl of 20
percent 2,4-OMP in benzene to the skin of female Sutter mice twice a week
for 23 weeks 'following a single application of a subcsrcincgenic dose (75
^g) of DM8A. 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 shown to promote carcinogenesis in mouse skin
bioassays (Lazar, et ai. 1966; 3ock, et ai. 1971; Roe, et al. 1959).
3, 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 literatureKU.S. EPA, 1979); however, data
was available on other positional isomers. Examination of rats treated
f
orally with 6 mg/kg of 2,6-dimethylphenol or 14 mg/kg of 3,4-dimethylphenol
for eight months revealed fatty dystrophy and atrophy of the hepatic cey.s,
yojo-
-------
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-DMP may inhibit vaso-
constriction, most likely due to its ability to block ATP (Lunde,. et al.
1968). Because of 2,4-DMP'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 could net be located in the available literature re-
garding any saltwater species.
A. Acute Toxicity
A reported 96-hour LC^ value for juvenile fathead minnows is
16,750 jug/1 (U.S. EPA, 1979). For the freshwater invertebrate Daphnia
maqna, the observed 48-hour LC5Q is 2,120 jjg/1 (U.S. EPA, 1979).
8. Chronic Toxicity
Based on an embryo-larval test with the fathead minnow, Pimephales
promelas, the derived chronic value is 1,100 ug/1 (U.S. EPA, 1978). No
chronic values are available for invertebrate species.
C. Plant Effects
Based on chlorosis effects, the reported LC50 for duckweed, Lemna
minor, is 292,800 jLig/1 for 2,4-dimethylphenol exposure (Slackman, 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
703/-
-------
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-DMP for any sector of the
environment or workplace.
A. Human
The draft criterion for 2,4-dimethylpnenol 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-QMP.
B. 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 jjg/1 at any time. No criterion exists for saltwater species (U.S.
EPA, 1979).
-------
2,4-OIMETHYLPHENOL
References
Baggett, M.S., and G.P. Morie. 1973. Quantitative determination of phenol
and alkylphenols in ciaarette 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 xylenes. Biochem. Jour._ 47: 395.
Corcos, A. 1939. Contribution to the study of occupational poisoning ;by
cresols. Dissertation. Vigot Freres Editeurs. (Fre).
Gornostasva, L.!., st al. 1977. Phenols from abies sibirica sssentaial
oil. Khim. Pirir. Scedin; ISS 3, 417-418.
Hcak, P..D. 1957. The causes of tastes and odors in drinking water. Free.
llth Ind. Waste Conf. Purdue Univ. Eng. Bull. 41: 229.
Hoffmann, D., et al. 1975. On the carcinogenicity of marijuana smoke.
Recent Adv. Phytochsm. 9: 63.
Hoffmann, D., and E.L. Wynder. 1363. Filtration of phenols from cigarette
smoke. Jour. Natl. Cancar Inst. 30: 67.
Kaiser, H.E. 1967. Cancer-promoting effects of phenols in tea. Cancer
20: 614.
Lazar, P., et al. 1966. Benzo(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
ATP-induced vasoconstriction in isolated perfused rabbit lungs. Acta.
Physiol. Scand. 72: 331.
»
Maazik, I.K. 1968. Dimethylphenol (xylenol) isomers and their standard
contents in water bodies. Gig. Sanit. 9: 18.
-/OSS-
-------
National Institute of Occupational Safety and Health. 1973. Occupational
exposure to cresol. DHEW (NIOSH) Publ. No. 78-133. U.S. Oep. Health Edu.
Welfare, Pub. Health Ser., Center for Dis. Control.
Roe, F.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.J. 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., Div. 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, D.C.
U.S. EPA. 1979. 2,4-Dimethylphenol: Ambient Water Quality Criteris
(Draft).
Uzhdovini, E.R., et al. 1974. Acute toxicity of lower phenols. Gig. fr.
Prof. Zabol. (2): 53.
Versar, Inc. 1975. Identification of organic compounds in effluents from
industrial sources. EFA-560/3-75-002. U.S. Environ. Prot. Agency.
Weast, R.C. 1576. Handbook of chemistry and physics. 57th ed. CRC Press,
Cleveland, Ohio.
~/03 V"
-------
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-uracy.
1036-
-------
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 phthalate 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 jug/1 were obtained for freshwater fish.
Criterion could not be drafted because of insufficient data
concerning the toxic effects of dimethyl phthalates to aquatic
organisms.
/ -/037-
-------
DIMETHYL PHTHALATE
I. INTRODUCTION
This profile is based on the Ambient Water Quality
Criteria Document for Phthalate Esters (U.S. EPA, 1979a).
Dimethyl phthalate {DMP} 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 gins 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 _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 vegeta-
tion. Evidence from in vitro studies indicates that certain
bacterial flora may be capable of metabolizing DMP to the
raonoester form (Englehardt, et al. 1975).
For additional information regarding the phthaiate
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 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
'/OS fr-
-------
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. EPA (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-
tration studies in biuegills.
III. PKARMACOKINETICS
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-
j-
able literature.
B. Mutagenicity ,
Dimethyl phthalate has been shown to produce muta-
genic effects in the Ames Salmonella assay (Rubin, et al.
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).
V. AQUATIC TOXICITY
A. Acute Toxicity
Two freshwater species were examined for acute
toxicity from dimethyl phthalate exposure. The 48-hour
static LC5Q for the Cladoceran, Daphnia ruag na, was 33,000
ug/1 (U.S. EPA, 1978; . The 96-hour scatic LC5Q value for
the bluegill, Lepomis macrochirus, was 49,500 ^ig/1. For
marine species, 96-hour static LC5Q values for the sheeps-
head minnow, Cyprinodon variegatus, and mysid shrimp, Mysid-
opsis bahi a, 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 39,300 to 42,700 pg/1 and 26,100 to 29,300
pg/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. 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 "no effect" levels observed in chronic
feeding studies in rats and dogs, the U.S. EPA (1979a) has
calculated an acceptable daily intake (ADI) level of 700
ma/day for DMP.
The recommended water quality criteria level for
protection of human health is 160 nig/liter for DMP (U.S.
EPA, 1979a).
B. Aquatic
The data base for toxicity of dimethyl phthalate
was insufficient for drafting criterion for either fresh-
water or marine organisms (U.S. EPA, 1979a).
-------
DIMETHYL PHTHALATES
REFERENCES
Draize, J.H., 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 microbial metabolism
of di-n-butyl phthalate and related dialkyl phthalates.
Bull. Environ. Contain. Toxicol. 13: 342.
Milkov, L.E., 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 mutagenic assay of a series
of phtnalic acid esters: positive response of the dimethyl
and diethyl esters in TA 100. Abstract. Soc. Tcxicol. Annu.
Meet. New Orleans, March 11.
Singh, A., et al. 1972. Teratogenicity of phthalate esters
in rats. Jour. Pharm. Sci. 61: 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. 1979b. Environmental Criteria and Assessment
Office. Hazard Profile: Phthalate Esters (Draft).
-------
No. 89
Dinitrohenzenes
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.
-------
OINITR08ENZENES
Summary
Due to the lack of available information, no assessment of the poten-
tial of dinitrobenzenes to produce carcinogenic effects, mutagenic effects,
teratogenic effects, or adverse reproductive effects can be made.
Dinitrobenzene is the most potent methemoglobin-forming agent of the
nitroaromatics and rapidly produces cyanosis in exposed populations.
v
Fish have been acutely affected by exposure to non-specified isomers of
dinitrobenzene at concentrations ranging from 2,000 to 12,000 ug/1.
/
-------
DINITROBENZENE-
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 89-9Q°C, boils at 300-303°C, and has a density of
1.55. Para-dinitrcbenzene (1,4-dinitrobenzene) is a white, crystalline
solid with a boiling point of 299°C, a melting point of 173-174°C, and a
density of 1.63 (Windholz, 1976). The dinitrobenzenes have low aqueous
solubility and ars soluble in alcohol.
The dinitrobenzenes are used in organic 'synthesis, the production of
dyes, and as a camphor substitute in celluloid production.
The domestic production 'volume of meta-dinitrobenzene in 1972 was
approximately 6 x 103 tons (U.S. EPA, 1976).
Dinitrcbenzenes are generally stable in neutral aqueous solutions; as
the medium becomes 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
Methemoglobin 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
administration.
B. Distribution
Pertinent information on distribution of dinitrobenzenes was not
found in the available literature.
C. Metabolism
Dinitrobenzene undergoes both metabolic reduction and oxidation.
Animal studies indicate that the major reduction productions following oral
dinitrobenzene administration were nitroaniline and phenylene diamine (35%
of the administered dose) (Parke, 1961). The major oxidative metabolites of
meta-dinitrobenzene were 2,4-diaminophenol (31% of initial dose) and
2-amino-4-nitrophenol (14% of initial dose). The phenols are further con-
jugated as glucuronides or etheral sulfates (Parke, 1961).
-------
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 determinsd in the feces (Parks, 1961).
IV. EFFECTS
A. Carcinogenic!ty
Information on the carcinogenicity of the dinitrobenzenes was not
found in the available literature. •-
B. Mutagenicity
Information on the rnutagenicity of the dinitrobenzenes was not
found in the available literature. The possible dinitrobenzene metabolite,
dinitrophenol (U.S. EPA, 1979), has been reported to induce chromatid breaks
in bone marrow cells of injected mice (Micra and Manna. 1971).
C. Teratogenicity
Information on the teratogenicity of the dinitrobenzenes was not
found in the available literature. The possible dinitrobenzene metabolite,
dinitrophenol (U.S. EPA, 1979), has produced developmental abnormalities in
the sea urchin (Hagstrom and Lonning, 1966). No effects were seen following
injection cr oral administration of dinitrophenol to mice (Gibson, 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).
'/of*-
-------
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 LeClerc (1950)
reported lethal doses of non-specific isomers of dinitrobenzene for minnows
(unspecified) at concentrations of 10,000 to 12,000 (jg/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 jjg/1 for unspecified dinitrobenzenes for an unspeci-
fied fish species.
B. Chronic Toxicity
Pertinent data could not be found in the available literature
regarding aquatic toxicity.
C. Plant Effects
Howard et al. (1976) report that the algae Chlorella sp. displayed
inhibited photosynthetic activity upon exposure to m-dinitrobenzene at a
concentration of 10" M.
VI. EXISTING GUIDELINES
The 8-hour time-weighted-average (TWA) occupational exposure limit for
dinitrobenzenes is 0.15 pptn(ACGIH, 1974).
-------
DINITROBENZENE5
References
ACGIH. 1974. Committee on threshold limit values: Documentation of the
threshold limit values for substances in the workroom air. Cincinnati, Ohio.
Bringmann, 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.E. 1973. Teratology studies in mice with 2-_sec-Butyl-4,- 6-dini-
trophenol (Dinoseb). Fd. Cosmet. Toxicol. 11: 31...
Hagstrom, B.E. and S. Lonning. 1966. Analysis of the effect of -Oinitro-
phenol on cleavaae and development of the sea urchin embryo. Protoplasma.
42(2-3): 246. •
Hashimoto, S. and K. Kano. 1972. Photochemical reduction of nitrobenzene
and reduction intermediates. X. Photochemical reduction of the mono-
substituted nitrobenzanes in 2-propanol. Bull. Chem. Soc. Jap. 45(2): 549.
Howard, P.M., et al. 1976. Investigation of selected potential environ-
mental contaminants: Nitroaromatics. Syracuse, N.Y.: Syracuse Research
Corporation, TR 76-573.
LeClerc, E. 1960. Sslf purification of streams and the relationship be-
tween chemical and 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, F., et al. 1956. Industrial waste water. 2nd ed. Gustav Fisher
Verlag Stuttgart, p. 536.
Micra, A.8. 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.
j
Neely, W.B., et al. 1974. Partition coefficient to measure bioconcsn-
tration potential of organic chemicals in fish. Environ. Sci. Technol.
8: 1113.
Parke, O.W, 1961. Detoxication. LXXXV. The metabolism of m-dinitro-
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.) 1976. The Merck Index. 9th ed. Merck and Co., Inc.,
Rahway, N.J. p. 3269.
•IDS I-
-------
No. 90
4,6-Dinltro-o-cresol
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCV
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 ac-c-uracy.
-------
4,6-DINITRO-0-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 El. 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.
-------
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 18 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.
-------
III. PHARMACOKINETICS
A. Absorption
DNOC 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 nitrophenols, both conjugation of the compound
and reduction of the nitro groups to amino groups occurs.
The metabolism of DNOC to 4-amino-4-nitro-o-cresol is a de-
toxification mechanism that is effective only when toxic
doses of DNOC are administered (U.S. E?A, 1979a). The
metabolism of DNOC is very slow in nan 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. '
-------
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 CNA damage in Proteus mirabilis.
Testing of this compound in the Ames Salmonella system
(Anderson, et al., 1972) or in J3. coli (Nagy, et al., 1975)
failed to show any mutagenic effects.
C. Teratoqenicity and Other Reproductive Effects
Pertinent data could not be located in the
available literature regarding teratogenicity and other
reproductive effects.
D. Chronic Toxicity
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 (MIOSH, 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.
-/of7-
-------
VT. EXISTING GUIDELINES AND STANDARDS
A. An eight-hour TLV exposure limit of 0.2 mg/m3 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
coxicological evidence (U.S. EPA, 1979a) .
-------
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 jig/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
toxicological evidence (U.S. EPA, 1979a).
-------
No. 91
2,4-Dinitrophenol
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.
-------
2,4-DINITRQPHENOL
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. lr\ vitro cell culture assays
failed to show the potential for mutagenic activity of 2,4-dinitro-
phenol as measured by unscheduled DNA 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 teratogenic effects.
At the levels of compound used in these mammalian studies, embryo-
toxic effects were observed.
Human use of 2,4-dinitrophenol as a dieting aid has produced
some cases of agranulocytosis, neuritis, functional heart damage,
and cataract development.
For aquatic organisms LC™ values ranged from 620 jug/1 for
the bluegill to 16,700 pg/1 for the fathead minnow.
<* -/Of 2-
-------
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, photochemicals, pest control agents, wood
preservatives, and explosives (U.S. SPA, 1979a) . The. 1968 pro-
duction of 2,4-dinitrophenol was 4.3 x 10"1 tons/yr. (U.S. EPA,
1979a).
For additional information regarding the nitrophenols 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 to assess exposure from water, inhalation, and foods.
Nitrophenols have been detected in effluents from chemical plants
(U.S. EPA, 1979a) . Dermal absorption of the dinitrophenols has
been reported {U.S. EPA, 1979a).
Exposure to nitrophenols appears to be primarily through
occupational contact (chemical plants, pesticide application).
r
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 on the octanol/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).
a. 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 conjugaj-
tion and reduction of nitro-groups to amino-groups, or oxidation to
dihydric-nitrophenols (U.S. EPA, I979a).
D. Excretion
Experiments with several animal species indicate that
urinary clearance of dinitrophenols is rapid (Harvey, 1959) .
VI. EFFECTS
A. Carcinogenicity
2,4-Dinitrophenol has been found not to promote skin
tumor formation in mice following DMBA initiation (Bautwell and
Bosch, 1959).
B. Mutagenicity
Testing of 2,4-dinitrophenol has indicated rautagenic
#
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 on maternal health may have contributed to these effects.
E. Chronic Toxicity
Use of 2,4-dinitrophenol =s a human dieting aid has pro-
duced some cases of agranulocytosis, neuritis, functional heart
damage, and a large number of patients suffering from cataracts
(Homer, 1942) .
F. 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_ of 620 pg/1 in a static,
96-hour assay (U.S. -EPA, 1978). Juvenile fathead minnows (Pime-
phales p_romela.s) were more resistant in flow through tests, with
an LCj0 of 16,720 ug/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-Q values ranging from 5,500 to 29,400 jjg/1 (Rosenthal
and Stelzer, 1970). The marine mysid shrimp tMysiaopsi^ bahia)
had an LC5Q of 4,850 ug/1 (U.S. EPA, 1978). ;
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 yjg/1 for
the alga (Chlorella pyrsnoidosa) (U.S. EPA, 1979a). The marine
alga (Skeletonema costatum) was more resistant with a reported
96-hour EC5Q value based on cell numbers of 98,700 ^g/1.
D. Residues
Based on the octanol/water partition' coefficient, a bio-
concentration factor of 8.1 has been estimated for 2,4-dinitro-
t
phenol for aquatic organisms with a lipid content of 8 percent.
-------
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 pg/1.
B. Aquatic
For protecting freshwater aquatic life, the draft cri-
terion is 79 pg/1 as a 24-hour avetage concentration not to exceed
180 pg/1. The marine criterion has been proposed as 37 pg/1
as a 24-hour average not to exceed 34 pg/1 at any time {U.S.
EPA, 1979a).
To protect saltwater life, the draft criterion is 37
pg/1 as a 24-hour average not to exceed 84 pg/1 at any time (U.S.
EPA, 1979a) .
''1067-
-------
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 mutagenic 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,
6-dinitrophenol (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'1-
al temperatures. Jour. Pharm. Pharmacoi. 11: 452.
Homer, W.D. 1942. Dinitrophenol and its 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., et al. 1975. Deficient myelination by 2, 4-dinitro-
phenol administration in early stage of development. Teratology
12: 204.
Phipps, G.L., et al. The acute toxicity of phenol and substituted
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. 1979b. Nitrophenols: Hazard Profile. Environmental
Criteria and Assessment Office {Draft).
-------
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.
-------
No. 92
Dinltrotoluene
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.
-------
DINITROTOLtJENE
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, methemo-
globinemia and anemia in humans and animals.
Acute studies in freshwater fish and invertebrates
suggest that 2,3-dinifcrotoluene is much more toxic than
2,4-dinitrotoluene.
-------
DINITROTOLQENE
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-jCgH^ (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 103 tons per year for the 2,4- and 2,6- isomers com-
bined (U.S. SPA, 1979}.
Dinitrotoluene is an ingredient of explosives for commer-
cial and 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 Sinitrotoluene
were not available. Dinitrotoluene waste products are dumped
»
into surface water or sewage by industries that manufacture
dyes, isocyanates, polyurethanes and munitions (U.S. EPA,
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 octanoi/water
partition coefficient.
C. Inhalation
Exposure to dinitrotoluene by inhalation is most
likely to occur occupationaiiy (U.S. EPA, 1979). However,
pertinent data could not be located in the available litera-
ture en atmospheric concentrations of dinitrctoluene and,
thus, possible human exposure cannon be estimated.
III. PHA.RMACOKZNETIGS
A. Absorption
14
The absorption of C-lsbeled isomers of dinitrotol-
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 2,6-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 DNT 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 2,4,6-trinitrotoluene in microbial systems, and the
known metabolism or 2,4,o-trinitrotoluene in mammals, the
U.S. SPA (1979) speculated that the metabolites of 2,4-di-
nitrotoluene in mammals would be either toxic and/or car-
cinogenic.
D. Excretion
i 4
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.
-------
IV. EFFECTS
A. Carcinogenicity
2,4-Dinitrotoluene fed to rats and mice for two
years produced dose-related increases in fibrorcas 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 cf hepatccsllular carcinomas and
neoplastic nodules in the livers.of females, a significant
increase of mammary gland tumors in females, and a suspicious ;
increase of hepatoceiiuiar carcinomas of the liver in males.
Male mice had a highly significant increase of kidney tumors
(Lee, et ai., 1373).
3. Mutagenicity
2,4-Dinitrotoluene was mutagenic in the dominant
lethal assay and in Salmonella typhimurium strain TA1535
(Hodgson, et ai. 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 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,
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, 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-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 LC5Q value of 330 jjg/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 LC5Q values of 660 jug/l(U.S. .EPA, 1978)
and 35,000 jjg/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 LC50 value of 2,280 ;ig/l
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.
B. 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
-------
a in the freshwater algae, Selenastrum capricornutum, were
1,370 or 1,620 pg/1/ respectively. These same effects mea-
sured in the marine algae, Skeletonema costatum, showed
it to be more sensitive. ECcQ 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 f ibroadenomas 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 Criteria
tper day' o IcT7 Itr* I
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 ^ig/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 .
B. Aquatic
A criterion to protect freshwater life has been
drafted as 620 ug/1 for a 24-hour average not to exceed
1,400 jag/1 for 2.4-dinitrotoluene and 12 pg/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 jig/1. Data
was insufficient to draft a criterion for 2,4-dinitrotoluene
for marine environments.
* Y0?o
-------
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, D.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. DAMD-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.
f
McGee, L.C., et al. 1942. Metabolic distrubances in workers exposed to
dinitrotoluene. Am. Jour. Dig. Dis. 9: 329.
9
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 (^-2,4-0^) in the rat. Radioisotopes 26: 780.
-------
National Cancer Institute. M978. Bioassay of 2,4-dinitrotoluene for possi-
ble carcinogenicity. Carcinogenesis Tech. Rep. Ser. No. 54. USDHEW (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.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-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 OAMD17-75-
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) <
-------
No. 93
2,4-Dlnltrotoluene
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.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (GAG) has evaluated
2,4-dinitrotoluene and has found sufficient evidence to
indicate that this compound is carcinogenic.
-jo* 5'
-------
2.4-OINITRQTOUJENE
Summary
2,4-Oinitrotoluene 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 ^ig/1. NO marine data are available.
-------
2,4-OINITROTOLUENE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Oinitrotoluene (U.S. EPA, 1979a).
2,4-Oinitrotoluene (2,4-DNT) has a melting point of 71°C, a boiling
point of 3QO°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-Dinitrotoluene 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 ran
material for the manufacture of dye-stuffs. Dinitrotoluenes are relatively
stable at ambient temperatures (U.S. EPA, 1979a). For additional infor-
mation regarding the dinitrotoluenes in general, the reader is referred to
v
the EPA/ECAO Hazard Profile on Dinitrotoluenes (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.
10*7-
-------
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. PHARMACOKINETICS
A. Absorption
The absorption of 14C-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-Dinitrotoluene is known to be ab-
sorbed through the respiratory tract and skin (U.S. EPA, 1979a).
B. Distribution
Tissue/plasma ratios of radioactivity after administration of
14C-labeled dinitrotoluene (DNT) to rats indicated retention of 14C
2,4-ONT in both liver and kidneys but not in other tissues (Hodgson, et al.
1977). A similar experiment with tritium-labeled 2,4-dinitrotoluene
( H-2,4-ONT) 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 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
In studies involving oral administration of 14C-dinitrotoluene or
3H-2,4-dinitrotoluene to rats (Hodgson, et al. 1377; 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, Carcinogenicity
2,4-Oinitrotoluene 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).
-------
8. Mutagenicity
2,4-Dinitrotoluene was mutagenic in the dominant lethal assay and
in Salmonella typhimurium strain TA 1535 (Hodgson, et ai. 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.
D. Chronic Toxicity -
Chronic exposure to 2,4-dinitrotoluene may produce liver damage,
jaundice, methemoglobinemia 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
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
f
(Friedlander, 1900; McGee, et al. 1942).
7d9o
-------
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 LC5Q value for the fat-
head minnow (Pitnepjhales prgmelas) was reported as 31,000 pg/1 and a 48-hour
static LC-Q value for the cladoceran, Daphnia maqna, was reported as
35,000 jjg/1.
B. 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
*r
(1979a) has estimated levels of 2,4-dinitrotoluene in ambient water which
will result in specified risk levels of human cancer:
10? i
-------
Exposure Assumptions Risk Levels and Corresponding Criteria
(per day)
0 10-7 10-6 iQ-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 pg/l 1.56 jug/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 .
B. 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 jug/1, has been proposed. Data are insuffi-
cient for drafting a marine criterion.
7092-
-------
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 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.
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
(TOT) and isomers of dinitrotoluene (DNT) in rats. Fed.
Proc. 36: 996.
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. 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.
-------
Linen, A.L. 1974. Biological monitoring for industrial ex-
posure to cyanogenic aromatic nitro and amino compounds. Am.
Ind. Hyg. Assoe. 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 toxic-
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) Publ. 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.
U.S. EPA. 1979a. Dinitrotoluene: Ambient Water Quality Cri-
teria. (Draft).
U.S. EPA. 1979b. Dinitrotoluene: Hazard Profile. Environ-
mental Criteria and Assessment Office.
-------
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.
-------
lit
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 ^yjgh imur iurn 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; C7HgN2O4; molecular weight
182.14) is a solid at room temperature. It i's in the shape of
rhombic needles and is soluble in ethanol. Its melting point is
»
66°C and its density is 1.28 at 111°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
-ion-
-------
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 &t_ Q. (1977)
traced the pathway of 14C 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
f
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 CPR 710).
-/099-
-------
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 ^t_ ^1_. (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 mufecu-
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 LD50'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 FR 23540).
-JlOl-
-------
BIBLIOGRAPHY
Alexander, M. and B.K. Lustigmann. Effect of chemical structure
on microbial degradation of substituted benzenes. J. Agr. 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 Agotobacite r Agi1is. Gesundh.-Ing.,
92(9), 273-276, 1971. (As cited in U.S. EPA, 1976).
Burlinson, N.E. ^t_ a^_. 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 jit_ ^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. 8. 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. ^t_ 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. £t_ _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-275-073, 1976.
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 a^. Acute toxicity and' skin corrosion data for
some organic and inorganic compounds and aqueous solutions.
Toxicol. 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.
-------
No. 95
Di-n-octyl 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 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.
-no i-
-------
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 (DOP) 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 iti
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, 1979a).
-------
Monitoring studies have indicated that most water phthalate concentrations
are in the ppm range, or 1-2 jug/liter (U.S. EPA, 1979a). Industrial
air monitoring studies have measured air levels of phthalates from 1.7
to 66 mg/m3 (Milkov, et 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. Caroinogenicity
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.
-------
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 (I979a), 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 eriterion 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.
'I/O?'
-------
DI-N-OCTYL PHTHALATE
REFERENCES
Engelhardt, G., et al. 1975. The microbial metabolism of di-n-butyl phtha-
late and related dialkyl phthalates. Bull. Environ. Contain. Toxicol.
13: 342.
Milkov, L.E., 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.
Singh, A.R., et 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)
-------
No. 96
1,2-Dlphenylhydrazlne
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 II3L-
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (GAG) has evaluated
1,2-diphenylhydrazine and has found sufficient evidence to
indicate that this compound is carcinogenic.
-11/3-
-------
1,2-DIPHENYLHYDRAZINE
Summary
The adverse effects of exposure to 1,2-diphenylhydrazine in-
clude damage to both the kidney and liver. Acute LDen values have
ranged from 300 to 960 mg/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/^
were reported for bluegill and Daphnia magna, respectively, and a
single chronic value of 251 ^ig/1 was reported for Daphnia magna.
-------
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 li"kely an underestimation of the
total amount of diphenylhydrazine available. Diphenylhydrazine i"s
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 (5,-10 rag/kg) and intravenously (4,8
nig/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 18.8 ug/kg/day in both sexes of rats; Zymbal's gland squamous-
cell tumors in male rats at 18.8 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 with'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).
-------
C. Teratogenicity
Pertinent information could not be located in the avail-
able literature.
D. Toxicity
One study reported an LD5Q 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 I^Q is listed as
301 rag/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 rag/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 LC=Q values for freshwater organisms
have been reported as 270 pg/1 for the bluegill, Lepomis macro-
chirus, and the 48-hour LC50 for the cladoceran, Dapjinia roagna,
is 4,100 pg/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 ;ig/l has been obtained for the
freshwater cladoceran, Daphnia Magna (U.S. EPA, 1978). No chronic
*
tests of diphenylhydrazine are available for marine organisms.
-------
C. Plants
Pertinent data could not be located in the available
literature.
D. 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; therefore, 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 Assuroptions Risk Levels and Corresponding Criteria
O 10_~7 10~6 ICT5
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 O .019 pg/1 0/19 pg 1.9
shellfish only.
-------
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 jag/1 and not to exceed 38 jjg/1 at any
time.
-------
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 mutagens and carcinogens. Preliminary results in the
validation of a novel short term test. Mutat. Res. 46: 305.
U.S. EPA. 1975. 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
Disulfoton is a highly toxic organophosphorous insecticide used on many
agricultural crops. The human oral LDLo is estimated at 5 mg/kg body
weight. Exposure results in central nervous system toxicity. The LD50
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/m-5. 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.
X
-------
I. INTRODUCTION
Oisulfoton 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, Mis-
souri (Stanford Research Institute (SRI), 1977). An estimated 4500 tonnes
were produced in 1974 (SRI, 1977). Oisulfoton is made by interaction of
0,0-diethyl hydrogen phosphorodithioate and 2-(2-ethylthio)ethylchloride
*
(Martin and Worthing, 1974). Oisulfoton 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
-// 2 5-
-------
TABLE 1. PHYSICAL AND CHEMICAL PROPERTIES OF DISULFOTON
Synonyms: 0,0-Dlethyl S-(2-(ethylthio)ethyl) phosphorodithioate;
0,0-Oiethyl S-(2-(ethylthio)ethyl) dithiophosphate; Thiodemeton;
Frumin; Glebofos; Ethylthiometon B; VUAgT 1964; Oi-Syston G;
Disipton; ENT-23437; Ethyl thiometon; VUAgT 1-4; Bay 19639; M 74
[pesticide]; Ekatin TD; CAS Reg. No. 298-04-4; M 74 (VAN); Bayer
19639; Di-System; Dlthiodemeton; Dithiosystox; 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
20
Specific Gravity and/or Density: d^ = 1.144
Melting and/or Boiling Points: bp 62OQ 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 20°C
Bioconcentration Factor (BCF) and/or
Octanol/water partition coefficient (Kow): KOW = 1.91
BCF = 1.0
Source: Martin and Worthing, 1974; Fairchild, 1977; Windholz, 1976;
U.S. EPA, 1980; Berg, et al. 1977.
-------
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).
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 reacned 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-
-------
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
Disulfoton 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. Carcinogenicity, Mutagenicity and Teratogenicity
Pertinent data could not be located in the available literature.
B. Chronic Toxicity and Other Relevant Information
Disulfoton is highly toxic to all terrestrial and aquatic fauna.
Human oral LDLo is estimated to be 5 mg disulfoton per kilogram body
weight (5 mg/kg). The symptoms produced by sub-lethal doses are typical of
central and peripheral nervous-system toxicity (Gleason, et al. 1969). The
reported l-u^g concentrations for other species are summarized below (Fair-
child, 1977).
-------
Species Exposure Route LD50 prig/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 TLm (equivalent to a 96Jnour LC50) for fathead
minnows was found to be 2.6 mg/1 in hard water and 3.7 mg/1 in soft water.
-------
Both tests were conducted at 25°C. The corresponding value for bluegilis
is estimated to be 0.07 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/m3. 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.
//30
-------
REFERENCES
Belanger, A. and H.A. Hamilton. 1979. Determination of disulfoton and per-
methrin residues in an organic soil and their translocation into lettuce,
onion and carrot. Jour. Environ. Sci. Health. B14: 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. Lipid and growth changes in organs of
chicks Gallus domesticus during acute a"nd 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.
-I/3/-
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No. 98
Endosulfan
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
-H33L-
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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.
-11-33-
-------
END05ULFAN
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 humans, 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 a
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 jjg/1 for five
freshwater fish; from 0.09 to 0.6 pg/1 for five saltwater fish in 48- or 96-
hour tests; from 0.04 to 380 ug/1 (EC50 and LC5Q) for seven saltwater
invertebrate species; and from 62 to 166 pg/1 for Oaphnia magna (48-hour
LC50). In the only chronic aquatic study involving endosulfan, no adverse
effects on fathead minnows were observed at 0.20 pg/1.
-------
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; C^lgHgO^; 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-
endosulfan or I and II. These isomers are present in- a ratio of 70 parts
alpha-endosulfan to 30 parts beta-endosulf an . 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 Drumrcond, 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
-113 f-
-------
and Mahfouz, 1977; Gorbach, et al. 1968; Miles and Moy, 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 jjg/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 Corneliussen, 1972). The U.S. EPA
(1979) has estimated the weighted average bioconcentration factor for endo-
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 (Corneliussen, 1970).
-------
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; Dorough 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
^ig/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 jjg/kg (Keil, 1972).
D. 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 nig/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 Dobos, 1969; Maier-
Bode, 1968). The beta-isomer is more readily absorbed than the alphaisomer.
Alcohols, oils, and emulsifiers accelerate the absorption of endosulfan by
f
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).
-------
8. Distribution
After ingestion by experimental animals, endosulfan 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 Dobos, 1969; Maier-Bode,
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 arid 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; Demma, 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
j-
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, endosulfan hy-
droxyether, and endosulfan ether (Knowles, 1974; Menzie, 1974). These meta-
bolites have also been found in microorganisms and plants (U.S. EPA, 1979).
-------
0. 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 (Gorbachr 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 ai. 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), Saccharorcyces cerevisiae, Esch-
ericia coli. and Serratia marcescens (Fahrig, 1974) indicate that endosulfan
is not mutagenic.
-ii3f-
-------
C. Teratogenicity
Endosulfan did not produce teratogenic effects in rats (Gupta,
1978).
D. 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 Coturnix
japonica, were sterile (U.S. EPA, 1979).
E. Chronic Toxicity
In the NCI bioassays (Kotin, 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 mo/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 LD5Q for technical endosulfan in rats is — 22 to 46
mg/kg and 6.9 to 7.5 mg/kg in mice (Gupta, 1976). Reagent grade a- and 0-
endosulfan are less toxic to rats (76 and 240 mg/kg, respectively; Hoechst,
-------
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 Dobos, 1969;
Boyd, et al. 1970).
.V. AQUATIC TOXICITY
A. Acute Toxicity
Ninety-six hour LC^n values, using technical grade endosulfan,
•JU
for five species of freshwater fish range from 0.3 pg/1 for the rainbow
trout, Salmo gairdneri, (Macek, et al. 1969) to 11.0 jjg/1 for carp finger-
-------
lings, Cyprinus carpio (Macek, et al. 1969; Schoettger, 1970; Ludemann and.
Neumann, 1960; Pickering and Henderson, 1966). Among freshwater inverte-
brates, Daphnia maqna is reported to have 48-hour LC5g values ranging from
62 to 166 ug/1 (Macek, et al. 1976; Schoettger, 1970), with three other in-
vertebrates yielding 96-hour LC50 values of 2.3 (Sanders and Cope, 1968)
to 107 jjg/l (Sanders, 1969; Schoettger, 1970). Levels of 400 and 800 ng/1
of technical endosulfan damaged the kidney, liver, stomach and intestine of
Gymongcorymbus ternetzi. The 96-hour LCgQ value was 1.6-jjg/l- (Amminikutty
and Rege, 1977,1978).
Of the five saltwater fish species tested, the reported 48- or 96-
hour LC5Q values ranged from 0.09 (Schimmel, et al. 1977) to 0.6 pg/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 ECcg and LC50 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).
B. Chronic Toxicity
Macek, et al. (1976) provided the only aquatic chronic study in-
volving endosulfan. No adverse effects on fathead minnow, Pimephales prome-
las, parents or offspring were observed at 0.20 /jg/1. Gymonocorymbus 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
*
tittle data is available concerning the effects of endosulfan on
aquatic micro/macrophytes. Growth of Chlorella vulaaris was inhibited
>2000jug/l (Knauf and Schulze, 1973).
-I 111-
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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 concen-
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 jjg/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.
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.
-------
The American Conference of Governmental Industrial Hygienists
(ACGIH, 1977) TLV time weighted average for endosulfan is 0.1 mg/nr3. 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 pg/1 at any
time. Saltwater criteria cannot be developed because of insufficient data
(U.S. EPA, 1979).
-1111-
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ENDOSULFAN
REFERENCES
ACGIH. 1977. Threshold limit values for chemical substances and physical
agents in the workroom environment with intended changes for 1977. 1977 TLV
Ariborne Contaminants Committee, American Conference of Government Indus-
trial Hygienists, Cincinnati, Ohio.
Agarwal, O.K., et al. 1978. Effect of endosulfan on drug metabolizing en-
zymes and lipid peroxidation in rat. Jour. Environ. Sci. Health C13: 49.
Aleksandrowicz, O.R. 1979. Endosulfan poisoning and chronic brain syn-
drome. Arch. Toxicol. 43: 65.
Amminikutty, C.K. and M.S. Rege. 1977. Effects of acute and chronic ex-
posure to pesticides, Thioden 35 E.G. 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 ternetzi). Boulenger-Indiana 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.
Sutler, P.A. 1963. Commercial fisheries investigations, pesticide-wildlife
studies. A review of Fish and Wildlife Service Investigations during 1961
and 1962. U.S. Dept. 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. Drummond. 1965. A plant surface oxidation product of
endosulfan. Jour. Econ. Entomol. 58: 356.
Chopra, N. and A. Mahfouz. 1977. Metabolism of endosulfan I, endosulfan
II, and endosulfan sulfate in tobacco leaf. Jour. Agric^ Food Chem. 25: 32.
Corneliussen, 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.
-------
Deema, P., et al. 1966. Metabolism, storage, and excretion of
sulfan in the mouse. Jour. Econ. Entomol. 59: 546.
Demeter, J. 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. Contam. Toxicol. 18: 110.
Domanski, J.J., et al. 1973. Insecticide residues on 1971 U.S. tobacoo
products. Tobacco Sci, 17: 80.
Oomanski, J.J., et al. 1974. Insecticide residues on 1973 U.S. tobacco
products. Tobacco Sci. 18: 111.
Dorough, H.W. and J.R. Givson. 1972. Chlorinated insecticide residues in
cigarettes purchases in 1970-72. Environ. Entomol. 1: 739.
Oorough, H.W., et- al. 1978. Fate of endosulfan in rats and toxicological
considerations of apolar metabolites. Pestic. Biochem. Physiol. 8: 241.
Duggan, R.E. and P.E. Corneliussen. 1972. Dietary intake of pesticide
chemicals in the United States (III), June 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. Occup. Med. 9: 36.
Estesen, B.J., et al. 1979. Dislodgable insecticide residues on cotton
foliage: Permethrin, Curocron, Fenvalarate, Sulprotos, Oecis and Endosulfan.
Bull. Environ. Contam. Toxicol. 22: 245.
Fahrig, R. 1974. comparative mutagenicity studies with pesticides. Int.
Agency Res. Cancer Sci. Publ. 10: 161.
FAD. 1975. Pesticide residues in food: report of the 1974 Joint Meeting of
the FAO 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: Maier-Bode, 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 LDgn, 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
occupationaily 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 Cl^-niarkierten Thiodan.
Unpublished. In: Maier-Bode, 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-
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Wirkung neuzeitlicher Kontaktinsektizide auf einsommerige Karfen (Cyprinum
carpioL.) Z. Angew. Zool. 47: 11.
V//7-
-------
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 to water fleas and
fathead minnows. EPA-600/3-76-099. U.S. Environ. Prot. Agency.
Maier-Bode, 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.
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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.Q. 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.
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Publ. 106.
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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
-------
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.
-1150'
-------
ENDRIN
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 LC50 value of 0.73 ug/lf 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 LC50 values of 3.80 and
58.91 ug/1 for marine and freshwater invertebrates, respec-
tively.
-------
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 ingredient. 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-
-------
1965 to zero in 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 ug/k<3 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 max imun level of 25.6 ug/^3 in
1971 to a maximum level of 0.5 ug/m3 in 1975 (U.S. EPA,
1979).
Tobacco products are contaminated with endrin resi-
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/g (Bowery, et al., 1959; Domanski and Guthrie, 1974).
r
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 rig/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?
f
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
7
-------
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 toxic 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
i
its metabolites may persist for longer periods of time (U.S.
EPA, 1979) .
-------
IV. EFFECTS . .
A. Carcinogenicity
In lifetime feeding studies with Osborne-Mendel
rats, endrin was neither tumorigenic nor carcinogenic (Deich-
raann, et al., 1970; Deichmann and MacDonald, 1971; Deichmann,
1972). A recent NCI bioassay concluded that endrin was not
carcinogenic for Osborne-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 rag/kg} administered intratesticularly
caused chromosomal aberrations in germinal tissues of rats,
including stickiness, bizarre configurations, and abnormal
disjunction (Dikshith and Datta, 1972, 1973).
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^Q
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 CMS
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).
F. Other Relevant Information
Endrin is more toxic, in both acute and chronic
studies, than other cyclodiene 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 LD^Q values, sug-
gestive of antagonism. Endrin given in equitoxic doses with
aldrin (a closely related compound) or chlordane gave higher
than expected LD5Q 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. LCgo values for static
bioassays ranged from 0.046 ug/1 for carp fry (Cyprinus
carpio) fry to 140.00 ug/1 for adult carp {lyatomi, et al.,
-------
1958). Excluding the results of age factor differences for
this species, adjusted static LC50 values ranrged from
0.27 ug/1 for large mouth bass (Microptecus salmoides)
(Fabacler, 1976) to 8.25 ug/1 for the bluegill (Lepomis
ma.crochirus) (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 LC50 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, LC50 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 LCjQ value of 0.037
ug/1, while the blue crab (Callinectejs sapidus) was the most
resistant, with an LC5Q 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 flagfish (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
-------
grass shrimp (Palaemonetes pug i o) 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 quadruplicatum {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 affinis (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
-------
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, discharge is not to ex-
ceed 0.0006 kg per 1,000 kg of production.
XT
-------
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.
-------
ENDRIN
REFERENCES
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fingerling channel catfish, lctalura_s punctatus. Jour.
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Batterton, J.C., et al. 1971. Growth response of bluegreen
algae to aldrin, dieldrin, endrin and their metabolites.
Bull. Environ. Contain. Toxicol. 6: 589.
Bedford, C.T., et al. 1975. The acute toxicity of endrin
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Bowery, T.G., et al. 1959. Insecticide r-esidues on tobacco.
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-------
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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.
Kevin, 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. Contain. 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 Chera. 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.
(Draft).
1979. Endrin: Ambient Water Quality Criteria.
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
Eplchlorohydrln (l-chloro-2,3-epoxypropane)
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.
-------
SPECIAL NOTATION
U.S. EPA's Carcinogen Assessment Group (CAG) has evaluated
epichiorohydrin and has found sufficient evidence to in-
dicate that this compound is carcinogenic.
-------
l-CHLORO-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.
-------
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 (CH2OCHCH2C1; 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
-------
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
(1AC)epichlorohydrin at doses of 1 and 100 ppm in air revealed 0.46+0.19
and 27.8+4.7 jjg epichlorohydrin per ml of plasma, respectively. The rates
-II73L-
-------
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
(•^O-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 ^4C02 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 Duuren (1977) has suggested a metabolite pathway of epi-
chlorohydrin to include glycidol, glycidaldehyde and epoxy-propionic acid.
D. 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
'1173-
3
-------
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 Duuren, 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 turbinates 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.
-mi-
-------
8. 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
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
9
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.
•Jt7f-
7
-------
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.
I979a). 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-
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 LD50 values in experimental animals have ranged from
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 epicniorohydrin in rats at
doses of 150 or 180 mg/kg have resulted in severe injury to the kidney
(Rotara and Pallade, 1966).
-------
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/m^ 10-hour time-weighted average and a
19 mg/m-3 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
-V/77-
-------
1-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. Fert. 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 epichlorohydrin. New
York University Medical Center. Letter dated June 23, 1978.
MIOSH. 1976. NIOSH criteria for a recommended standard:
Occupational exposure to epichlorohydrin. U.S. DHEW. Na-
tional Institute for Occupational Safety and Health.
-------
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.
Santodonato, 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 (-^ci-epichlorohydrin in male and female rats.
Res. Comm. Chem. Pathol. Pharmacol. 20: 275.
-im-
-------
Weil, C.S. 1964. Experimental carcinogenicity and acute
toxicity of representative epoxides. Amer, Ind. Hyg. Jour,
24: 305.
-------
Ho. 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.
-------
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.
-------
ETHYL METHACRYLATE
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 U7°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 LD5Q 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, Deichmann 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 (Deichmann, 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).
-------
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 LD5Q 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). Oeichmann 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.
-------
ETHYL METHACRYLATE
References
Austian, J. 1975. Structure-toxicity relationships of acrylic monomers.
Environ. Health Perspect. 19: 141.
Deichmann, w. 1941. Toxicity of methyl, ethyl, and n-butyl 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.
-------
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, [Fe(CN)6;p~( 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).jt
corresponds actually to one of the ferricyanide compounds, the ferric ferri-
cyanide with the actual formula Fe[Fe(CN)6]; aiso known as Berlin green.
The acid from which these salts are derived is called ferricyanic acid,
^[FefCNjg] (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]; ancj ferric ferrocyanide,
F^fFeCCNjgl-j, 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 Othmer, 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
-lift-
-------
available literature. Prussian blue, potassium ferric hexacyanoferrate
(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.
B. 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.
-------
VI. EXISTING GUIDELINES AND STANDARDS
Pertinent data could not be located in the available literature.
-in 3-
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REFERENCES
Dreisbach, 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. Interscience 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.
•lit1/'
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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.
-iw-
-------
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 mutagenic, 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 96-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/1. Observed 96-hour EC5Q values based on cell
numbers for fresh and saltwater algae are over 45,000 ug/1.
-1197-
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FLUORANTHENE
I. INTRODUCTION
This profile is based on the Ambient Water Quality
Criteria Document for Fluoranthene (U.S. EPA, 1979).
Fluoranthene (1,2-benzacenapthene, M.W. 202) is a poly-
nuclear aromatic hydrocarbon of molecular formula ^ig^lO"
Its physical properties include: melting point, 111°C; boil-
ing point, 375°C; water solubility, 265 jug/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
P
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 pg/day
Food 1.6 - 16 ug/day
Air 0.040 - 0.080 pg/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).
-------
G. Metabolism
Pertinent information could not be located in
the available literature. By 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 -
feces; urinary excretion varies with the degree of formation
of conjugated metabolites (U.S. EPA, 1979).
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 Duuren
-------
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 Antes 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 bluegil!9 Lepomis
mac£ochj.r^j s s is reported to be 3,980 ^ig/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 I-C value (U.S. EPA, 1978) . The fresh-
I&OI-
-------
water invertebrate Daphnia magna appears to have a low
sensitivity to fluoranthene with a reported 48-hour ECcn
value of 325,000 jig/1. The 96-hour LC5Q value for the salt-
water rays id shrirapj 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 for the raysid shrimp is 16 ^ig/1.
C. Plant Effects
The freshwater alga, Selenastrum capricornutum,
when exposed to fluoranthene resulted in a 96-hour EC5Q
value for cell number of 54,400 ^jg/1. On the same criterion,
the 96-hour EC5Q value for the marine alga, S k e le t on ema
costatum, is 45,600 pg/1 (U.S. EPA, 1979).
D. Residues
No measured steady-state bioconcentration factor
(BCF) is available for fluoranthene. A BCF 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 PAH compounds
in drinking water.
Based on the no-effect level determined in a single
t
animal study (Hoffman, et al. 1972), the U.S. EPA (1979)
has estimated a draft ambient water criterion of 200 ^ig/1
for fluoranthene. However, the lower level derived for
A
-------
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 ;ag/l as a 24-hour average,
not to exceed 560 jig/1 at any time. For saltwater life,
the criterion is 0.30 ug/1 as a 24-hour average, not to
exceed 0.69 jag/1 at any time.
-------
FLUOROANTHENE
REFERENCES
Barry, G., et al. 1935. The production of cancer by pure
hydrocarbons-Part III. Proc. Royal Soc., London. 117: 318.
Basu, 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-l, 2-bis-{3 pyridyl)-1-propanone(metopirone ciba)
and/S -diethylaminoethyldiphenyl-n-propyl acetate (SKR 525-A).
Br. Jour. Cancer 24: 548.
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 hydroxyraethyl 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, £., et al. 1979. A comparison of the mutagenicity,
tumor initiating activity and complete carcinogenicity of
polynuclear aromatic hydrocarbons. _T_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 and 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 Cri-
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 Duurenr 8.L., and B.H. 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
-------
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.
-J2.07-
-------
FORMALDEHYDE
SUMMARY
The major 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. EPA's 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
mutagenic 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;
i/t* •
-------
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. U
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 paraformaldehyde 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
NIOSH (1976) estimates that 1,750,000 workers are poten-
.tially exposed to formaldehyde in the workplace.
A. Environmental Pate
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). Hydrogen peroxide can
also be formed during photodecomposition of formaldehyde (Purcell
and Cohen, 1967; Bufalini _et> al^. , 1972). The atmospheric half-
»
life of formaldehyde is less than one hour in sunlight {Bufalini
.et_a., 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
s
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
-/an-
-------
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
•v
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 blood. 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_ _a.L. , 1965).
-------
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 CC>2 and H^O. In rats fed radiolabeled formaldehyde,
40% of the radiolabel was recovered as respiratory CC>2 (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,
1976r Malorny^t^., 1965).
D. Excretion
Some of the formic acid metabolite is excreted in the urine
as the sodium salt; most, however, is oxidized to CC>2 and
eliminated via the lungs (U.S. EPA, 1976).
IV. HEALTH EFFECTS
A. Carcinogenicity
Watanabe ^t^ _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
f!
-------
exposed for an additional 29 weeks at 122 ppm. No tumors or
metaplasias were found, although numerous changes were observed
in respiratory tissues (Horton ^t^ al_., 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 15 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
Xl0 . This estimate may change when the final results of the
CUT study become available.
-------
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 et aj._. (1978)
in a study with S^. cerevisiae r Wilkens and MacLeod (1976) in
E. coli; Martin _et_ _al_. (1978) in an unscheduled DNA synthesis
test in human HeLa cells; Obe and Seek (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/m3 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/m of formaldehyde
for 90 days. Hematologic values were normal, however, some
-------
interstitial inflammation occurred in the lungs of all species
(Coon et al., 1970).
F. 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 LC-Q of 15 to 35 ppm.
The LC^Q 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).
-------
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/m3 (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., Kuschinaky, K., Kewitz, H. and Koran sky, 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., 34(5), 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_0_(1), 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., Uj 3), 459 (as cited in U.S. EPA, 1976).
Kamata, E. 1966. Aldehyde in Lake and Sea Water. Bull. Chem.
Soc. Japan, 39_(6) , 1227 (as cited in U.S. EPA, 1976)
Magana-Schwencke, N. , B. Ekert, and E. Moustacchi. 1978. Bio-
chemical analysis of damage induced in yeast by formaldehyde. I.
-------
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 Formaldeshyds zu Ameiscansaure im Blat. ein Beitrag Zum
Stoffwechsel des Foonaldehyds. 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. Beek. 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 ToxicT 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)
-------
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. ~~
-------
No. 105
Formic Acid
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
-/3L3U-
-------
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.
-------
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 maqna 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.
-------
FORMIC ACID
I. INTRODUCTION
Formic acid (CAS registry number 64-18-6) is a colorless, clear, fuming
liquid with a pungent odor (Hawley, 1971; Windholz, 1976; Walker, 1966). It
is a naturally formed product, produced by bees, wasps, and ants (Casarett
and Doull, 1975). 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 Pure 90% 85%
Formula: CH2°2 — —
Molecular Weight: 46.02 — —
Melting Point: 8.4°C -4°C -12°C
Boiling Point: 100.5°C
Density: 1.22020 1.202g 1.194^
Vapor Pressure: 33.1 torr d 20°C
Solubility: Miscible in water,alcohol,
and ether; soluble in
acetone,benzene, and toluene
Demand (1979): 67.5 million IDS. (CMR, 1979)
-------
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 raw 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, 1976).
B. Food
A large variety of plants contain free formic acid; it has been
detected in pine needles, stinging nettle, and fruits (Walker, 1566). It
has been identified in a number of essential oils, including petitgrain
lemon and bitter orange (Furia and Bellanca, 1971). 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, 1575).
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
-------
has also been identified in the liquid condensate from the pyrolysis of
solid municipal waste (Orphey and Jerman, 1970),
D. 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, 1963; NIOSH, 1977)'.
8. 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 14C02 were detected in these samples (Palese and
Tephly, 1975).
,-
IV. EFFECTS
A. Pertinent data could not be located in the available literature/
-------
8. Chronic Toxicity
Chronic human exposure to formic acid has been reported to produce
albuminuria and nematuria (Windholz, 1976).
C. Other Pertinent Information
Formic acid is severely irritating to th skin, eyes, and respira-
tory tract (NIOSH, 1577). Gleason (1569) 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 LC50 vaiue of
175,000 >jg/l for bluegill sunfish (Lepomis macrochirus) exposed to formic
acid. Bringmann and Kuhn (1955) observed a 48-hour LC5Q value of 120,000
/jg/1 for waterfleas (Daphnia maqna) exposed to formic acid.
Verschueren (1579) 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 50,000 jug/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 100,000>ug/l was toxic to the freshwater algae, Scenedesmus sp.
—i—^^_^^^^^_— f
D. Residue
Pertinent data were not found in the available literature.
-------
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
The eight-hour, TWA exposure limit for occupational exposure to
formic acid is 5 ppm (ACGIH, 1977).
8, Aquatic
Hahn and Jensen (1577) have suggested an aquatic toxicity rating
range of 100,000 to 1,000,000 yug/1 based on 96-hour LC^ values for aqua-
tic organisms exposed to formic acid.
-------
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 L. 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, 8.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.) 1571. Fenaroli's Handbook of Flavor In-
gredients. The Chemical Rubber Company, Cleveland, 0.
Gleason, M. 1969. Clinical Toxicology of Commercial Products, 3rd ed.
Williams and Wilkins, Baltimore, MD.
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 Cceanographic
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: Al.
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. I960. Gas Chromatographic 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: D.
-------
Patty, F. 1563. Industrial Hygiene and Toxicology, Vol. II. 2nd ed.
Interscience, New York.
Portmann, J.E. and K.W. Wilson. 1971. The toxicity of 140 substances to
the brown shrimp and 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-6QOM-76-Q62.
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 Technology, 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.
-------
No. 106
Fumaronitrile
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
-/A 31-
-------
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.
-------
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.
-------
njMARONITRILE
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. Fumaronitrile 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.
Bioaccumulation data on fumaronitrile were not found in the avail-
able literature.
X
-------
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.
B. Acute Toxicity
LD5n values for injected mice and orally dosed rats were 38 and
*.
50 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.
-------
REFERENCES
Campagna, F., et al. 1977. A convenient synthesis of nitriles from
primary amides under mild conditions. Tetrahendron Letters. 21: 1813.
Law, A. 1968. Fumaronitrile as a bactericide. Chen, Abst. 68: 1135.
Ono, T. 1979. Maleonitrile, a bactericide and fungicide. Chem. Abst.
82: 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.
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No. 107
Ralomethanes
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.
-------
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.
-------
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
dichlorodifluoromethane (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
jjg/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:
-------
Halomethanes in the U.S. EPA Region V
Organics Survey (83 Sites)
Compound
Bromodichloromethane
Tribromomethane
Dichlorome thane
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
-------
and urban levels suggesting that the oceans may be a major source of global
chloromethane and bromomethane. Outdoor bromomethane concentrations as high
as 0.00085 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 ever 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 10 mg/m3, and concen-
trations for dichlorofluoromethane ranged from 3.5 x 10~3 to 2.9 x 10"2
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 fluorocarbons (Natl. Acad. Sci.,
1978; Davis, et al. 1977; U.S. EPA, 1976; Howard, et al. 1974).
-------
8. Distribution
Halomethanes are distributed rapidly to various tissues after ab-
sorption into the blood. Preferential distribution usually occurs to
tissues with high 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.
Oichloromethane 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.
0. 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
-------
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 cnloromethane, bromomethane,
and dichloromethane were all mutagenic to Salmonella tvphimurium strain
TA1QO 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. Dichloromethane
was negative in mitotic recombination in S^ cerevisiae 03 (Simmon, et al.
1977) and in mutagenicity tests in Drosophila (Filippova, et al. 1967).
Trichlorofluoromethane and dichlorofluoromethane were negative in the Ames
assay (Uehleke, et 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 o^f feto-
»
embryotoxicity has been noted in rats and mice exposed to dichloromethane
-------
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 monocytogenes) 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 LC5g
values for the bluegill sunfish (Lepomis machrochirus) of 11,000 ug/1 for
methylbromide, 29,300 ;jg/l 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 pg/1 methylene chloride for the fathead minnow
(Pimephales promelas) while a flow-through assay produced an LC5Q value of
193,000 jug/1. In freshwater invertebrates two acute studies with Daphnia
-------
magna resulted in LC__ values of 46,500 pg/1 for bromoform, and 224,000
pg/1 for methylene chloride. In marine fish, LC^ values for the sheeps-
head minnow (Cyprinodon variegatys) were 17,900 pg/1 for bromoform and
180,958 pg/1 for methylene chloride. For the tidewater silversides (Menidia
beryllina) LC5Q values of 12,000 pg/1 for methylbromide and 147,610 pg/1
for methylene chloride were obtained. Adjusted LC^ values for the marine
mysid shrimp (Mysidopsis bahia) were 24,400 pg/1 for bromoform and 256,000
pg/1 for methylene chloride (U.S. EPA, 1979).
8. Chronic Toxicity v *
The only chronic value for an aquatic species was 9,165 /jg/1 for
the sheepshead minnow.
C. Plant Effects
Effective concentrations for chlorophyll a and cell numbers in
freshwater algae Selenastrum capricornutum ranged from 112,000 to 116,000
pg/1 for bromoform and 662,000 pg/1 for methylene chloride, while effective
concentrations for the marine algae (Sketonema cpstatum) were reported as
11,500 to 12,300 pg/1 for bromoform and 662,000 pg/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
-------
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. Oichlorodifluoromethane 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 bromodiehloromethane are the same as that
for chloroform, 2 pg/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/m .
Dichloromethane: OSHA (1976a,b) has established an 8-hour time-
weighted average for dichloromethane of 1,737 mg/rt , 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.
-------
Tribromomethane: QSHA (1976a,b) has established an 8-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/m3,
respectively. The U.S. EPA {1979} draft water quality criteria for tri-
chlorofluoromethane and dichlorodifluoromethane -are 32,000 and 3,000 pg/1,
respectively.
B. Aquatic
Draft criteria for the protection of freshwater life have been
derived as 24-hour average concentrations for the following halomethanes:
methylbromide - 140 pg/1 not to exceed 320 pg/1; bromoform - 840 pg/1 not to
exceed 1,900 pg/1; methylene chloride - 4,000 pg/1 not to exceed 9,000 pg/1;
and methyl chloride - 7,000 jug/1 not to exceed 16,000 pg/1.
Draft criteria for the protection of marine life have been derived
as 24- hour average concentrations for the following halomethanes: methyl-
bromide 170 pg/1 not to exceed 380 pg/1; bromoform - 180 pg/1 not to exceed
420 pg/1; methylene chloride - 1,900 pg/1 not to exceed 4,400 pg/1; and
methyl chloride - 3,700 pg/1 not to exceed 8,400 pg/1.
-------
HALOMETHANES
REFERENCES
Ahmed, A.E., et al. 1977. Metabolism of haloforms to carbon
monoxide, I. In. 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 biotransformation and elimination of '(I4c)-trichloro-
fluoromethane and (^'*c)'-dichlorodifluoromethane 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.
-------
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. Jj_n
Progress in genetic toxicology.
Singh, H.B., et al. 1977. Urban-non-urban relationships of
halocarbons, SFg, N2O 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-
monary tumor response in strain A mice. Cancer Res. 37:
2717.
-------
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
-------
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
heptachlor and has found sufficient evidence to indicate
that this compound is carcinogenic.
-------
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 epoxide, 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 LC5(, values for fresh-
water fish range from 7.0 jjg/1 to 320 pg/1 and 24 to 96-hour LC5n 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.
-------
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 (chemical name l,4,5,6,7,8,8-heptachloro-3a,4,7,7a-
tetrahydro-4,7-methanoindene; cHC1! molecular, weight 373.35) is a
white crystalline solid with a camphor-like odor. It has. a vapor pressure
of 3 x 10~* mm Hg at 25°C, a solubility in water of 0.056 rag/1 at 25 to
29°C, and is readily soluble in relatively nonpolar solvents (U.S. EPA,
1979).
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
t
(Lichtenstein, 1960; Lichtenstein, et al. 1970, 1971; Nash and Harris,
1972), in plants (Gannon and Decker, 1958), and in mammals (Davidow and
-------
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
(Lepotnis 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 jjg/1 (U.S. EPA, 1976). Levels of heptachlor ranged from .001
jjg/1 to 0.035 ug/1 and heptachlor/heptachlor epoxide 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).
-------
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 pg/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/m .
Thus, sprayers, farmers and nearby residents of sprayed fields may receive
significant exposures (Nisbet, 1977).
0. Dermal
Gaines (1960) found rat dermal LD5Q values of 195 and 250 mg/kg
for males and females, respectively, compared with oral Ln50's Of 100 and
162 mg/kg, respectively, for technical heptachlor. Thus, dermal exposures
*
may be important in humans under the right exposure conditions.
-------
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).
B. 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 heg-
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).
-------
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, et 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.
B. 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 (Cerey, et al. 1973). Bone marrow cells of the
treated animals showed increases in the incidence of abnormal mitoses, chro-
»
matid abnormalities, pulverization, and translocation. Both heptachlor and
heptachlor epoxide induced unscheduled DNA synthesis in SV-40 transformed
•ilto '
-------
human cells (VA-4) in culture with metabolic activation (Ahmed, et al.
1977). Neither heptachlor nor heptachlor epoxide was mutagenic for Salmo-
nella typhimurium 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). Phenobarbital 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, et 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, et al..
1978).
Fish are less sensitive to heptachlor than are invertebrate spe.-
cies. Ninety-six hour LC5Q values for fish range from 7.0 ug/1 for the
rainbow trout, Salmo qairdneri, (Macek, et al. 1969) to 320 pg/1 for the
14
goldfish (Carassius auratus). Ten days after a dose of 0.868 pg/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 ug/1 for the stone-
fly, Pteronarcella badia, (Sanders and Cope, 1968) to 80 jjg/1 for the clado-
ceran (SJ.mgceghajLus 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 jug/1 to 120 ug/1 as determined in a 26-hour
LC5Q Oaohnia magna bioassay (Frear and Boyd, 1967).
'/3.6SU-
-------
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 pg/1
(Schimmel, et al. 1976). Other species such as the blue crab, Callinecte.s
sapidus, and American oyster, Crassostrea virginica, are 2,100 and 950 times
less sensitive, respectively, than the pink shrimp (Butler, 1963).
8. Chronic Toxicity
In a 40-week life cycle test with fathead minnows (Pimephales prom-
elas), the determined no-adverse-effect concentration was 0.86 ug/1. All
•^^**™*"""* «
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-
ngdon variegatus) growth of fry was significantly reduced at 2.04 ug/1, 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 Aig/1 caused a
94.4 percent decrease in productivity of a natural- saltwater phytoplankton
community after a 4-hour exposure to heptachlor (Butler, 1963).
D. Residues
The amount of total residues, heptachlor and heptachlor epoxide,
accumulated by fathead minnows after 276 days of exposure was found to be
z
-------
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
7,518 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 andCorresponding Draft Criteria
(per day)
0 10-7 10-6 10-5
2 liters of drinking water 0 0.0023 ng/1 0.023 ng/1 0.23 ng/1
and consumption of 18.7
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 Standards
Agency Published Standard ' Reference
Occup. Safety 500 ug/m^* on skin from air Natl. Inst. Occgp.
Health Admin. Safety Health, 1977
Am. Conf. Gov. 500 ug/m3 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
-------
U.S. Publ. Health Recommended drinking water Natl. Acad. Sci., 1977
Serv. Adv. Comm. standard (1968) 18 jug/1 of
heptachlor and 18 jjg/1 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 pg/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 pg/1 at any time (U.S. EPA, 1979).
Vf
-------
HEPTACHLOR
REFERENCES
Ahmed, F.E., et al. 1977. Pesticide-induced DNA damage
and its repair in cultured human cells. Mutat. Res. 42:
161.
American Conference of Governmental Industrial Hygienists.
1971. Documentation of the threshold limit values for sub-
stances in workroom air. 3rd. ed..
Benson, W.R., et al. 1971. Photolysis of solid and dis-
solved dieldrin. Jour. Agric. Food Chem. 19: 66.
Breidenbach, A.W., et al. 1967. Chlorinated hydrocarbon
pesticides in major river basins, 1957-65. Pub. Health
Rep. 82: 139.
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 and Wildl. Circ. 167: 11.
Cerey, K., et al. 1973. Effect of heptachlor on dominant
lethality and bone marrow in rats. Mutat. Res. 21: 26.
Davidow, B. and J.L. Radomski. 1953. Isolation of an epox-
ide metabolite from fat tissues of dogs fed heptachlor.
Jour. Pharmacol. Exp. Ther. 107: 259.
Davis, K.J. 1965. Pathology report on mice fed aldrin,
dieldrin, heptachlor, or heptachlor epoxide for two years.
Internal Memorandum to Dr. A.J. Lehman. U.S. Food Drug
Admin.
Donovan, M.P., et al. 1978. Effects of pesticides on metabo-
lism of steroid hormone by rodent liver microsomes. Jour.
Environ. Pathol. Toxicol, 2: 447.
Eisler, R. 1970. Factors affecting pesticide-induced
toxicity in an estuarine fish. Bur. Sport Fish. Wildl.
Tech. Paper 45. U.S. Dept. Inter, p. 20.
Epstein, S.S. 1976. Carcinogenicity of heptachlor and
chlordane. Sci. Total Environ. 6: 103.
14
Feroz, M. , and M.A.Q. Khan. 1979. Metabolism of C-hepta-
chlor in goldfish (Carassius auratus). Arch Environ. Contam,
Toxicol. 8: 519.
-------
Frear, D.E.H., and J.E. Boyd. 1967. Use of Daphnia magna
for the microbioassay of pesticides. I. Development of
standardized techniques for rearing Daphnia 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.
Gannon, N., and G.C. Decker. 1958. The conversion of aldrin
to dieldrin on plants. Jour. Econ. Entomol. 51: 8.
Georgackakis, E., and M.A.Q.. Khan. 1971. Toxicity of the
photoisomers of cyclodiene insecticides to freshwater animals.
Nature 233: 120.
Goodman, L.R., et al. 1978. Effects of heptachlor and
toxaphene on Laboratory-reared embryos and fry of the sheeps-
head minnow. Proc. 30th Annu. Conf. S.E. Assoc. Game Fish
Comm. p. 192.
Hansen, D.J., and P.R. Parrish. 1977. Suitability of sheeps-
head minnows (Cyprinodon variegatus) for life-cycle toxicity
tests. Pages 117-126 In: F.L. Mayer and J.L. Hamelink,
eds. Toxicology and hazard evaluation. ASTM STP 634, Am.
Soc. Test. Mater.
Harbison, R.D. 1975. Comparative toxicity of selected
pesticides in neonatal and adult rats. Toxicol. Appl.
Pharmacol. 32: 443.
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 hepta-
chlor expoxide, 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.
Jonsson, V., et al. 1977. Chlorohydrocarbon pesticide
residues in human milk in greater St. Louis, Missouri, 1977.
Am. Jour. Clin. Nutr. 30: 1106.
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 cyclase
system. Life Sci. 13: 1363.
-a 6 7-
-------
Kacew, S., and R.L. Singhal. 1974. Effect of certain-, halo-
genated hydrocarbon insecticides on cyclic adenosine 3 ,5 -
monophosphate- H formation by rat kidney cortex. Jour.
Pharmacol. Exp. Ther. 188: 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. Contam. Toxicol.
1: 159.
Korn, S., and R. Earnest. 1974. Acute toxicity of twenty
insecticides to the striped bass, Mgrone sajttjLlis. 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
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 potential using Salmonella typhimurium mutants.
Jour. Agric. Food Chem. 24:
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.
-------
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
chemicals on pesticide toxicity in rats. Philipp. Jour.
Nutr. 27: 30.
Miranda, C.L., et al. 1973. Effect of dietary protein
quality, phenobarbital, and SKF 525-A on heptachlor metabo-
lism in the rat. Pestic. Biochem. Physiol. 3: 456.
Mizyukova, I.G., and G.V. Kurchatav. 1970. Metabolism
of heptachlor. Russian Pharmacol. Toxicol. 33: 212.
Nash, R.G., and W.G. Harris. 1972. Chlorinated hydrocarbon
insecticide residues in crops and soil. Jour. Environ.
Qual.
National Academy of Sciences. 1977. Drinking water and
health. Washington, D.C.
National Institute for Occupational Safety and Health.
1977. Agricultural chemicals and pesticides: a subfile
of the registry of toxic effects of chemical substances.
Nisbet, I.C.T. 1977. Human exposure to chlordane, hepta-
chlor and their metabolites. Unpubl. rev. prepared for
Cancer Assessment Group, U.S. Environ. Prot. Agency, Wash-
ington, D.C.
Podowski, A.A., et al. 1979. Photolysis of heptachlor
and cis-chlordane and toxicity of their photoisomers to
animals. Arch. Enviorn. Contain. Toxicol. 8: 509.
Radomski, J.L., and B. Davidow. 1953. The metabolite of
heptachlor, its estimation, storage, and toxicity. Jour.
Pharmacol. Exp. Ther. 107: 266.
Ritcey, W.R., et al. 1972. Organochlorine pesticide resi-
dues in human milk, evaporated milk, and some milk substi-
tutes in Canada. Can. Jour. Publ. Health 63: 125.
St. Omer, V.. 1971. Investigations into mechanisms respon-
sible for seizures induced by chlorinated hydrocarbon insecti-
cides: The role of brain ammonia and glutamine in convul-
sions in the rat and cockerel. Jour. Neurochem. 18: 365.
Sanders, H.O., and O.B. Cope. 1968. The relative toxicities
of several pesticides to naiads of three species of stone-^
flies. Limnol. Oceanogr. 13: 112.
-------
Savage, E.P. 1976. National study to determine levels
of chlorinated hydrocarbon insecticides in human milk.
Unpubl. rep. submitted to U.S. Environ. Prot. Agency.
Schimmel, S.C., et al. 1976. Heptachlor: Toxicity to
and uptake by several estuarine organisms. Jour. Toxicol.
Environ. Health 1: 955.
Singhal, R.L., and S. Kacew. 1976. The role of cyclic
AMP in chlorinated hydrocarbon-induced toxicity. Federation
Proc. 35: 2618.
Sovocool, G.W., and R.G. Lewis. 1975. The identification
of trace levels of organic pollutants in humans: compounds
related to chlordane heptachlor exposure. Trace Subst.
Environ. Health 9: 265.
Tashiro, S., and F. Matsumura. 1978. Metabolism of trans-
monachlor and related chlordane components in rats and man.
Arch. Environ. Contam. Toxicol. 7: 113.
U.S. EPA. 1976. Chlordane and heptachlor in relation to
man and the environment. EPA 540/476005.
U.S. EPA. 1977. Risk assessment of chlordane and hepta-
chlor. Carcinogen Assessment Group. U.S. Environ. Prot.
Agency, Washington, D.C. Unpubl. rep.
U.S. EPA. 1979. Heptachlor: Ambient Water Quality Cri-
teria (Draft).
Webb, R.E., and C.L. Miranda. 1973. Effect of the quality
of dietary protein on heptachlor toxicity. Food Cosmet.
Toxicol. 11: 63.
-------
No. 109
Heptachlor Epoxide
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.
-12 73.-
-------
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.
-------
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 FDA showed in their market basket survey (1974-
1975) of 20 different cities that 3 of 12 food classes con-
-1 2 7*/'
-------
tained residues of heptachlor epoxide ranging from 0.0006
to 0.003 ppm {Johnson and Manske, 1977). Heptachlor epoxide
residues greater than 0.03 rag/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 epoxide; 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.
j-
C. Inhalation
Heptachlor epoxide is present in ambient air ,to
a lesser extent than heptachlor and is not thought to con-
2
127S-
-------
tribute substantially to human exposure except in areas
near sprayed fields, where concentrations of up to 9.3 pg/m
may be encountered {Nisbet, 1977).
D. Dermal
Gaines (1960) found rat dermal LD5Q values of
195 and 200 mg/kg for males and females, respectively, com-
pared with oral LD5Q'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. PHARMACOKINETICS
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 butterfat, and in the fat
of poultry eggs and livestock as compared to residue levels
found in their normal food rations. "Heptachlor residues"
-------
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 Matsumura, 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 cultdre
when metabolically activated (Ahmed, et al. 1977), but was
-------
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.
D. Other Relevant Information
Heptachlor epoxide is mote 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.
B. Chronic Toxicity
In the only reported chronic study, the 26-hour
LC50 for heptachlor epoxide in Daphnia magna was 120 yig/1
(Frear and Boyd, 1967). In the same test, the corresponding
value for heptachlor was. 52 jig/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-
-------
tration factor for heptachlor 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/m 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 pg/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-
-------
mate that 0.47 ng/man/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/Fe/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).
-------
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.
ed.
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, O.E.H. and J.E. Boyd. 1967. Use of Daphnia magna for the microbio-
assay and pesticides. I. Development of standardized techniques for rearing
Daphnia 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.
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. Contam. 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-60'0/3-76-099.
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.
-------
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. Prot. Agency, Washington, O.C.
Radmoski, J.L. and 8. Davidow. 1953. The metabolite of heptachlor, its
estimation, storage, and toxicity. Jour. Pharmacol. 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. Prot. Agency.
Tashiro, S. and F. Matsumura. 1978. Metabolism of trans-nonachlor and re-
lated chlortane components in rat and man. Arch. Environ. Contam. Toxicol.
7: 113
U.S. EPA. 1977. Risk assessment of chlordane and heptachlor. Carcinogen
Assessment Group. U.S. Environ. Prot. Agency, Washington, D.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
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Yamaguchi, I., et al. 1979. Inhibition of synaptic atpases by heptachlor
epoxide in rat brain. Pest. Biochem. Physiol. 11: 285.
U-
-------
No. 110
Hexachlorobenzene
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
-/a.
-------
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.
-------
SPECIAL NOTATION
U.S. EPA1s Carcinogen Assessment Group (GAG) has evaluated
hexachlorobenzene and has found sufficient evidence to
indicate that this compound is carcinogenic.
-------
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 placental 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,
Lagodon rhoimboides, bioconcentrated this compound 23,000 times in 42 days
of exposure. The concentration of HCB in muscle of pinfish was reduced only
16 percent after 28 days of depuration, a rate similar to that for DDT in
fish.
-------
HEXACHLOROBENZENE
I. INTRODUCTION
This profile is based on the Ambient Water Quality Criteria Document
for Chlorinated Benzenes (U.S. EPA, 1979).
Hexachlorobenzene (HCB; CgClgj 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^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 supplies at
-------
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 yg 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 *he
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).
-------
D. 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).
B. 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
burden is not lowered (Villeneuve, 1975).
-------
C. Metabolism
Hexachlorobenzene is metabolized after i.p. administration in the
rat to pentachlorophenol, tetrachlorohydroquinone and pentachlorothiophenol
(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 hexachlorobenzene 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 i ty
Carcinogenic activity of hexachlorobenzene was assessed in hamsters
fed 4.8 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,
the only increase in tumors was in hepatomas (Cabral, et al. 1978). How-
-------
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).
8. 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). CD-I mice receiving 100 mg/kg/day hexachlorobenzene orally
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. 1978). 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 (HCB) 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 ug/1. In a 42-day
exposure, the pinfish, Laqodon rhgmboides, BCF was 23,000. The concen-
tration of HCB in pinfish muscle was reduced only 16 percent after 28 days
of depuration; this slow rate is similar to that for DDT 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.
-------
A. Human
The value of 0.6 pg/kg/day hexachlorobenzene was suggested by
FAD/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 ip-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.
8. Aquatic
Pertinent information concerning aquatic criteria could not be
located in the available literature.
-------
HEXACHLOROBENZENE
REFERENCES
Albro, P.W., and R. Thomas. 1974. intestinal absorption of
hexachlorobenzene and hexachlorocyciohexane isomers in rats.
Bull. Environ. Contain. 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. Contain. Toxicol. 16: 248.
Food and Agriculture Organization. 1974. 1973 evaluations
of some pesticide residues in food. FAO/AGP/1973/M/9/1; WHO
Pestic.. 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.r and W. Koransky. 1975. Studies on the toxicology
of hexachlorobenzene. I. Pharmacokinetics. 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.
-/ 3L 9?-
-------
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 Pish 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.
s
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 hexachlorobenzene in the rat. Toxicol.
Appl. Pharmacol. 31: 313.
-------
No. Ill
Hexachlorobutadiene
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.
-------
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 Ell 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 LC5Q 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.
-1300-
-------
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
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
1361
-------
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 mg/kg. Studies on HCBD contamination of food in several
European countries have measured levels as high as 42 u9/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
-------
although values as high as 460 u9/m 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. 1974).
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 f
high dose developed renal carcinomas. Metastasis to the lung
was observed in one case each for both male and female rats.
-/303-
-------
No carcinomas were observed in controls, however, a nephro-
blastoraa 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 glomerulonephritis 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,
-------
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 rag/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 Toxicity
Goldfish, (Carassius auratus), had an observed 96-
hour LCgQ 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 EC50 values of 210 and 130 u.g/1, respective-
ly (Leeuwangh, et al., 1975). No acute studies with marine
species have been conducted.
B. Chronic Toxicity
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 card iacum, .160 times after 7 days exposure {Laseter, et
al., 1976). Residue data on saltwater organisms are not
available.
-------
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 1CT5, 10~6, and 10~7 (U.S. EPA, 1979).
The corresponding criteria are 0.77 ug/1, 0.077 U9/1i 0.0077
ug/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).
-1307"
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HEXACHLOROBUTADIENE
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, D., 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
(HCBD). 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, D.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. Delo. 4: 130. HAPAB:72:820. (Translation)
-------
Poteryaeva, G.E. 1973. Toxicity of hexacblorobutadiene 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. Cno vol.): 73. CA:81:
73128E. (Translation)
Schwetz, B.A., et al. 1977. Results of a reproduction study in rats fed
diets containing nexachlorobutadiene. 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 - Hexacnlorobutadiene. 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).
•/30T-
-------
No. 112
Heachloro eye1ohe xane
/I
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.
-------
SPECIAL NOTATION
U.S. EPA*s Carcinogen Assessment Group (GAG) has evaluated
hexachlorocyclohexane and has found sufficient evidence to
indicate that this compound is carcinogenic.
-------
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.
-13/3 -
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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 (C^H^Cl^; 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).
/
-------
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 ug/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 isomers to be 1 to 3 ug/kg
body weight (Ouggan 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 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).
B. Distribution
After administration to experimental animals, lindane was detected
in the brain at higher concentrations than in other organs (Laug, 1948;
.'
Oavidow and Frawley, 1951; Koransky, et al. 1963; Huntingdon Res. Center,
-1316-
-------
1972). At least 75 percent of an intraperitonial dose of C-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 CChadwick, 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).
0. 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).
-1317-
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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 (Suselmair, 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).
-------
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).
D. 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
F~ and F females. In addition, some of the F^ and F2 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).
-131?-
-------
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/m3) 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_n values from one flow-
through and 24 static • bioassays for the gamma isomer of hexachloro-
cyclohexane ranged from 2 pg/1 for the- brown trout (Salmo trutta) (Macek and
McAllister, 1970) to 152 jug/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. Zebrafish
(Brachydanio rerio) showed a lindane LC value of 120 ng/1, but rainbow
trout (Salmo gairdneri) evidenced respiratory distress at 40 ng/1 (Slooff,
1979). Technical grade HCH was much less toxic than pure lindane; LC5Q
-------
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 8CH (15.5 percent gamma isomer) (Henderson, et al. 1959). Static
tests on freshwater invertebrates revealed a range of LC5Q values of from
A.5 pg/1 (96-hour test) (Sanders and Cope, 1968) for the stonefly
(Pteronarcys californica) to • 880 ^ug/1 (48-hour test) (Sanders and Cope,
1968) for the clado- ceran (Simocephalus serralatus) for lindane. Canton
and Slooff (1977) re- ported an LC5Q value for the pond snail (Lymnaea
stagnalis) of 1,200^/1 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 jjg/1 for the Atlantic
silversides (Henidia 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 LC _ values from 7.3 ^g/1 for the striped
bass (Morone saxatilis) (Korn and Earnest, 1974) to 240 ^g/1 for the long
nose killifish (Fundulus similis) (Butler, 1963). A single species, the
pinfish (Lagodon rhomboides), tested with technical grade hexachlorocyclo-
hexane, produced a 96-hour flow-through LC5Q 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 jug/1 for the grass shrimp (Palaemonetas
vulgaris) (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 ^ug/1 based on shell decom-
position (Butler, 1963).
-735.'
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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/1 were obtained for Chironomus tentans,
Gammarus fasciatus, and Daphnia maqna (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 jug/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 ^ig/1
(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 (Laqodojn rhomboides), the American oyster
-I32&.-
-------
(Crassostrea virginica), and offal tissue of the pinfish (Schimmel, et al.
1977). Sugiura, et al. (1979a) found alpha-, beta-, a'nd 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 jjg/1 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
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-
time cancer risk for humans of 10" is 21 ng/1, "based on the data of
Nagasaki, et al. (1972).
133.3-
-------
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 1Q~5 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 ug/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 0.5jjg/m3 (skin) for air (U.S. EPA, 1979).
B. Aquatic
For lindane, freshwater criteria have been drafted as 0.21 ug/1
with 24-hour average concentration not to exceed 2.9 pg/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.
-------
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 tox icologicaT
bioassays (model substance o^-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.
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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. Jous.
Pharmacol. Exp. Therap. 100: 59.
Goto, M. , et al. 1972. Ecological chemistry. Toxizitat
von a-HCH in mausen. Chemosphere 1: 153.
-------
Hanada, M. , et al. 1973. Induction of hepatoma in mice
by benzene hexachloride. Gann. 64: 511.
Hansen, P.O. 1979. Experiments on the accumulation of
lindane (gamma BHC) by the primary producers Chlorella spec.
and Chlorella pyrenoidosa. Arch. Environ. Contain. Toxicol.
8: 72~n
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. Progc. Pestic.
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Herbst, M., and G. Bodenstein. 1972. Toxicology of lin-
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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.
Ito, N., et al. 1973. Histologic and ultrastructural 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. Contain. Toxicol. 22: 522.
r'
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. Acute toxicity of twenty
insecticides to striped bass, Marone saxatilis. Calif.
Fish Game 60: 128.
-------
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. Bipchem. 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: 2346.
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-
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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. iMcAllister. 1970. Insecticide sus-
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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
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Nagasaki, H., et al. 1972. Carcinogenicity of benzene
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National Academy of Sciences - National Research Council.
1977. Safe Drinking Water Committee. Drinking Water and
Health. p. 939.
732.7-
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National Cancer Institute. 1977a. A bioassay for possible
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National Cancer Institute. 1977b. Bioassay of lindane
for possible carcinogenicity. NCI Carcinogenesis Technical
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feeding on testicular tissue on pure inbred Swiss mice.
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biphenyls and organochlorine insecticides in human embryos
and fetuses. Pediatrician 6: 45.
Palmer, A.K., et al. 1978. Effect of lindane on pregnancy
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Petrescu, S., et al. 1974. Studies on the effects of long-
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Poradovsky, R., et al. 1977. Transplacental permeation
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405.
Reuber, M.D. 1979. Carcinogenicity of lindane. Environ.
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prolonged exposure to BHC. Vrach. Delo. 10: 133.
Schimmel, S.E., et al. 1977. Toxicity and bioconcentration
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Environ. Contam. Toxicol. 6: 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 6: 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
gannna-Hexachlor o eye1ohexane
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 acc-uracy.
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Disclaimer Notice
Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
-/33JL-
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GAWA-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 mg/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-BHC, 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 cyclig
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-
Tently 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 pg/1 (U.S. EPA, 1975).
B. Food
The daily intake of lindane has been reported at 1 to 5 ;jg/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).
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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;
Oavidow and Frawley, 1951; Koransky, et al. 1963; Huntingdon Research Cen-
ter, 1971). At least 75 percent of an intraperitoneal dose of C-labe^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
-------
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
J9 -xX- C
Nagasaki, et al. (1972b) fed "^ , f , ~T, and £ 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 ppm^ -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 *f 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 <==<, $', y~ 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. The^p -isomer-treated animals had no tumors.
-------
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/x -, J~-, or & -HCH with the highly carcinogenic action of c^-
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 (U 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 o -HCH to Osborne-Mendel rats and B6C3F1 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 "2T"-HCH in mice is ques-
tionable (Natl. Cancer Inst., 1977).
B. Mutagenicity
Some alterations in mitotic 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). 2" -HCH was mutagenic in assays using Salmonella
typhimurium with metabolic activation, the host-mediated assay, and the
133?-
-------
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).
D. 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 postimplantation 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
-133?-
-------
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 (Litterst and Miller, 1975) and by treatment
with silymarin (Szpunar, et al. 1976), and various tranquilizers (Ulmann,
1972).
V. AQUATIC TOXICITY
A. Acute Toxicity
The range of adjusted LC5Q values for one flow-through and '124
static bioassays for lindane in freshwater fish ranged from 1 jug/1 for the
brown trout Salmo trutta (Macek, et al. 1970) to 83 jjg/1 for the goldfish
(Carassius auratus). and represents the results of tests on 16 freshwater
fish species (ILS. EPA, 1979). Zebrafish (Srachydanio rerio) showed an
LC5Q value of 120 jjg/1 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, stoneflies (Pteronarcys californi-
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
(Daonnia pulex. 0_._ magna and Simocephalus serralatus) were most resistant
with LC5Q values of 390 to 745 jjg/l. The midge (Chironomus tentans) was
intermediate in sensitivity with LC5Q values of 175 jjg/1 (U.S. EPA, 19*79).
-------
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 pg/1 (Eisler, 1970), while the striped mullet (Mugil
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 LC5Q of 7.3 ug/1 (Korn and Earnest, 1974); and the
longnose killifish (Fundulus similis) was most resistant with a reported
LC5n of 240 jug/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 ^ig/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 (Pimepnales 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, Acetafaularia mediterranea,
were 1,000 and 10,000 ^ig/1, respectively. Irreparable damage to Chlorella
spec, occurred at concentrations greater than 300 ;jg/l (Hansen, 1979).
D. 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
-------
(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 10~5 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.
B. Aquatic
The criterion has been drafted to protect freshwater organisms as a
0.21 jjg/1 24-hour average concentration not to exceed 2.9 pg/l. Data are
insufficient to draft criterion for the protection of marine life from gam-
ma-hexachlorocyclohexane (lindane).
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GAMMA-HEXACHLOROCYCLOHEXANE(LINDANE)
REFERENCES
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Barney, J.E. 1969. Chem. Eng. News, Vol. 42.
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Chadwick, R.W., and J.J. Freal. 1972. The indentification
of five unreported lindane metabolites recovered from rat
urine. Bull. Environ. Contain. Toxicol. 7: 137.
Chadwick, R.W., et al. 1975. Dehydrogenation, a previously
unreported pathway of lindane metabolism in mammals. Pestic.
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Copeland, M.F., and R.W. Chadwick. 1979. Bioisomerization
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Davidow, B., and J.P. Frawley. 1951. Proc. Soc. Exp. Biol,
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Didenko, G.G., et al. 1973. Investigation of the possible
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Duggan, R.E., and M.B. Duggan. 1973. Residues of pesti-
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Eisler, R. 1970. Acute toxicities of organochlorine and
organophosphorus insecticides to estuarine fishes. Bur.
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Engst, R., et al. 1977. Recent state of lindane metabolism.
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Fitzhugh, O.G., et al. 1950. Chronic toxicities of benzene
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Pharmacol. Exp. Therap. 100: 59.
Goto, M., et ai. 1972. Ecological chemistry. Toxizitat
von a-HCH in mausen. Chemosphere 1: 153.
Hanada, iM., et al. 1973. Induction of hepatoma in mice
by benzene hexachloride, Gann. 64: 511.
-------
Hansen, P.O. 1979 Experiments on the accumulation of lin-
dane ( -BHC) by the primary producers Chlocella spec.
and Chorella pyrenoidosa. Arch. Environ. Contam. Toxicol.
8: 72T!
Henderson, C-, et al. 1971. Organochlorine pesticide resi-
dues in fish-fall 1969: Natl. Pestic. Monitor. Progr. Pestic.
iMonitor. Jour. 5: A.
Herbst, M., and G. Bodenstein. 1972. Toxicology of lindane.
Page 23 in; E. Ulraann, (ed.) Lindane. Verlag K. Schillinger
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.
-------
Hansen, P.O. 1979 Experiments on the accumulation of lin-
dane ( -BHC) by the - primary producers Chlorella spec.
and Chorella pyrenoidosa. Arch. Environ. Contain. Toxicol.
8: 721.
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. Toxicology of lindane.
Page 23 in; E. Ulmann, (ed.) Lindane. Verlag K. Schillinger
Publ., Freiburg.
Heyroth, F.F. 1952. In: Leland, S.J., Chem. 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.
Itq, N^, ejt 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. 195=-. Absorption of some chlorinted
hydrocarbon insecticides from soils into various crops.
Jour. Agric. Food Chem. 7t 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 cession fish family representatives.
Trans. Am. Fish. Soc. 99: 20.
Macek, K.J., et al. U-76. Chronic toxicity of lindane
to selected aquatic invertebrates and fishes. EPA 600/3-
76-046. Q.3. 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^rkm. 20: 28.
Mather, S.P., and J.G. Sana. 1975. Microbial degradation
of lindane-C-14 in a flooded sand loam soil. Soil Sci,
120: 301.
Nagasaki, H. , et al. 1972. Carcinogenicity of benzene
hexachloride {BHC}. Top. Chem. Carcinog., Proc. Int. Symp.,
2nd. 343.
National Academy of Sciences - National Research Council^
1977. Safe Drinking Wa~er Committee. Drinking Water and
Health p. 939.
National Cancer Institute. 1977. A bioassay for possible
carcinogenicity of lindane.. Fed. Reg. Vol. 42. No. 218.
Nigam, S.K., et al. 1979. Effect of hexachlorocyclohexane
feeding on testicular tissue on pure inbred Swiss mice.
Bull. Environ. Contam. Tcsicol. 23: 431..
Nishimura, H., et al. 1977. Levels of polychlorinated
biophenyls and organochlrrine insecticides in human embryos
and fetuses. Pediatrician 6: 45.
Palmer, A.K., et al. 1S78. 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-
term administration of chlorinated organic pesticides (lin-
dane, DDT) on laboratory white rats. Rev. Med. - Chir.
78: 831.
Poradovsky, R. , et al. 1977. Transplacental permeation
of pesticides during ncrmal pregnancy. Cesk Gynekol. 42:
405.
-------
Lichtenstein, E.P. 1959. Absorption of some chlorinted
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.
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-
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.
Mather, S.P., and J.G. Saha. 1975. Microbial degradation
of lindane-C-14 in a flooded sand loam soil. Soil Sci.
120: 301.
Nagasaki, H., et al. 1972. Carcinogenicity of benzene
hexachloride (BHC). Top. Chem. Carcinog., Proc. Int. Symp.,
2nd. 343.
National Academy of Sciences - National Research Council.
1977. Safe Drinking Water Committee. Drinking Water and
Health p. 939.
National Cancer Institute. 1977. A bioassay for possible
carcinogenici,ty of lindane. Fed. Reg. Vol. 42. No. 218.
Nigam, S.K., et al. 1979. Effect of hexachlorocyclohexane
feeding on testicular tissue on pure inbred Swiss mice.
Bull. Environ.. Contain. Toxicol. 23: 431.
Nishimura, H. , et al. 1977. Levels of polychlorinated
biophenyls and organochlorine insecticides in human embryos
and fetuses. Pediatrician 6: 45.
Palmer, A.K., et al. 1978. Effect of lindane on pregnancy
in the rabbit and rat. Toxicology 9: 239.
Petrescu, 5., et al. 1974. Studies on the effects of long-
term administration of chlorinated organic pesticides {lin-
dane, DDT) on laboratory white rats. Rev. Med. - Chir.
78: 831.
Poradovsky, R. , et al. 1977. Transplacental permeation
of pesticides during normal pregnancy. Cesk Gynekol. 42:
405.
7.3/7-
-------
Reuber, M.D. 1979. Carcinogenicity 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
of BHC and lindane in selected estuarine animals. Arch.
Environ. Contain. Toxicol. 6: 355.
Sloof, W. 1979. Detection limits of a biological monitor-
ing system based on fish respiration. Bull. Environ. Contain.
Toxicol. 23: 517.
Sugiura, R. , et al. 1979. Accumulation of organochlorine
compounds in fishes. Difference of accumulation factors
by fishes. Chemosphere 6: 359.
Szpunar, K., et al. 1976. Effect of silymarin on hepatoxic
action of lindane. Herba. Pol. 22: 167.
Thorpe, £., and A.I. Walker. 1973. The toxicology of diel-
drin (HEOD). II. In mice with dieldrin, DDT, phenobarbitone,
beta-BCH, and gamma-BCH. Food Cosmet. Toxicol. 11: 433.
Tsoneva-Maneva, M.T., et al. 1971. Influence of Diazinon
and lindane on the mitotic activity and the karyotype of
human lymphocytes cultivated in vitro. Bibl. Haematol.
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
by soil microorganisms. Arch. Microbiol. 108: 259.
Ulmann, E. 1972. Lindane: Monograph of an insecticide.
Verlag K. Schillinger Publishers, Freiburg, West Germany.
U.S EPA. 1979. Hexachlorocyclohexane: Ambient Water Quality
Critera (Draft).
-------
No. 114
Hexachlorocyclopentadiene
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.
-13 &
-------
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 ug/1
to 104 jug/1. The chronic value for fish in an embryo-larval test is 2.6
ug/1.
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HEXACHLOROCYCLQPENTADIENE
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.
-J3S2-
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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.
B. 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
t
necrotic changes, and that the major route of elimination of HEX is through
the lungs. This information is based on morphological changes in rats
-------
administered HEX by gavage. 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 evefi 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 hexaneiisopropanol mixture.
D. 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.
-------
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).
-------
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 mg/kg/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.
D. 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 other abnormalities were
reported. Naishstein and Lisovskaya (1965) applied 0.5 to 0.6 ml of a solu-
tion of 20 ppm 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.
-------
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 LC^ values for the fathead minnow
(Pimgphales promelas) under static and flow-through conditions with larval
and adult fish range from 7.0 jjg/1 to 104 ug/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 pg/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.
-------
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).
B. Aquatic
For HEX, the draft criterion to protect freshwater aquatic life is
0.39 ug/1 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.
-------
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.
Sell, 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,
-------
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, D.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 l4C-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. Dev. 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, D.M. 1978. Letter to R. A. Ewing, Battelle Columbus Labora-
tories, dated August 9, 1978. Comments on document 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, D.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
'1361-
-------
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.
-------
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.
-S3 63
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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 Qthmer,
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).
-------
C. Teratogenicity
Teratogenic effects were not observed, in pregnant rats exposed to
hexachloroethane by inhalation or intubation (Weeks, et al. 1979),
D. 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
LC^Q value of the freshwater Cladoceran (Daphnia maqna) was reported as
8,070 jjg/l (U.S. EPA, 1978). For the marine fish, the sheepshead minnow
(Cyprinodon varieqatus), a 96-hour LC5Q value of 2,400 pg/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 940 jug/1 (U.S. EPA, 1978).
8. Chronic Toxicity
Pertinent data could not be located in the available literature*.
-------
C. Plant Effects
For the freshwater algae, Selenastrum capricornutum, the 96-hour
EC5Q effective concentrations based on chlorophyll and cell number were
87,000 and 93,200 pg/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 pg/1 being
reported.
D. Residues
A bioconcentration factor of 139 was detained for the freshwater
bluegill sunfish (U.S. £PA, 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 pg/1.
The eight-hour TWA exposure standard established by OSHA for hexa-
chloroethane is 1 ppm.
B. Aquatic Toxicity
The proposed criterion to protect freshwater aquatic life is 62
pg/1 as a 24-hour average and should not exceed 140 pg/1 at any time. The
»
drafted criterion for saltwater aquatic life is a 24-hour average concen-
tration of 7 pg/1 not to exceed 16 pg/1 at any time.
-------
HEXACHLOROETHANE
REFERENCES
Dickson, A.G,, and J.P. Riley. 1976. The distribution
of short-chain halogenated aliphatic hydrocarbons in some
marine organisms. Mar. Pollut. Bull. 79: 167.
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 Publishers, Inc. Sandusky, Ohio.
National Cancer Institute. 1978. Bioassay of hexachloro-
ethane for possible carcinogenicity. NatJ.. Inst. Health,
Natl. Cancer Inst. DHEW Publ. No. (NIH) 78-1318. Pub.
Health Serv. U.S. Dept. 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.
U.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: 187.
-------
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.
-------
I. INTRODUCTION
Hexachlorophene
-------
III. PHARMACOKINETICS
A. Absorption
Systemic toxicity following dermal application or ingestion of
hexachlorcphene 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 ;jg/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).
D. 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-percent and 50-percent so-
.-
lutions of hexachlorophene to mice failed to produce significant car-
cinogenic effects (Stenback, 1975); the levels of compound used caused bigh
toxicity. Ruo'ali 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).
-i373-
-------
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
v
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.
V
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).
-------
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.
-------
REFERENCES
American Medical Association. 1977. AMA Council on Drugs, Chicago.
Arnold, D., et al. 1975. Mutagenic evaluation of hexachlorophene.
Toxicol. Appl. Pharmacol. 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.
Desi, 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. Contam. 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 Cnem. 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.O. 1979. Importance of hepatic function on the plasma
disappearance and biliary excretion of hexachlorophene. Toxicol. Appl.
Pharmacol. 49: 113.
-/37V-
-------
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.
*.
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 D. 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.
-IS 77-
-------
No. 117
Hydrofluoric Acid
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.
-737?
-------
HYDROFLUORIC ACID
Summary
Hydrofluoric acid (HF) has produced mutagenic effects in plants and
Orosophila, 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 bloassay test demonstrated that a concentration of
50,000 ug/1 HF was lethal to bluegill sunfish in one hour.
-------
• HYDROFLUORIC ACID
I. INTRODUCTION
Hydrofluoric acid (CAS registry number 7664-39-3) (HF) is a colorless,.
clear, fuming corrosive liquid made by treating fluorspar (CaF2) with 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/HjO
Molecular Weight: 20.01 —
Melting Point: -83.550C —
Boiling Point: 19.Sloe —
Density: 0.987 1.15 - 1.18
Vapor Pressure: 1 atm ® 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.
B. 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 CCI_2F2 with NO and humid air (Saburo, et al. (1977). It is
present in the stratosphere (Zander, et al. 1977; Drayson, et al. 1977;
Fanner and Raper, 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. PHARMACQKINETICS
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 (MAS, 1971; NIOSH, 1975;
NIQSH, 1976)_ Fluoride reaches fetal circulation via the placenta and is
deposited in the fetal skeleton (MAS, 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 gsnsrs*
-------
health (MAS, 1971). It is recognized that urinary fluoride levels are di-
rectly related to the concetration of absorbed fluoride (NAS, 1971).
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, 1971).
IV. EFFECTS
A. Carcinogenic!ty
Pertinent data were not found in the available literature.
B. Mutagenicity
Mohamed (1968) 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 (1976).
Rats inhaling 0.1 mg HF/m3 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. 1973).
Weak mutagenic effects in the offspring of Drosophila exposed to
air bubbled through 2.5 percent HF have been reported (Mohamed, 1971).
C. Teratogenicity
Pertinent data were not found in the available literature.
D. Other Reproductive Effects
Reduced fertility in Drosophila and decreased egg hatch have been
reported following exposure to 2.9 ppm HF (Gerdes, et al. 1971).
E. Chronic Toxicity
Among the adverse physiologic effects of long-term exposure to HF
are skeletal fluorosis, dental mottling and pulmonary impairment (NAS, 1971;
NIOSH, 1975; NIOSH, 1976). Skeletal fluorosis is characterized by increased
JS
-------
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/nv5 devel-
oped pulmonary, kidney and hepatic pathology (Machle and Kitzmiller, 1935;
Machle, et al. 1934), while animals exposed to 24.5. mg HF/nv5 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).
V5V-
-------
V. AQUATIC TQXICITY
A. Acute Toxicity
McKee and Wolf (1963) reported that HF was toxic to fish
(unspecified at concentrations ranging from 40,000 to 60,000 jug/1. Bonner
and Morgan (1976) observed that 50,000 ^ig/1 HF was lethal to bluegill sun-
fish (Lepomis macrochirus) in one hour.
8. 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/m3 (3 ppm) 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/m5 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.
-------
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, Cookeville,
Tennessee. Prepared for the Office of Water Research and Technology.
Available from NTIS: PB261-253.
Burke, W.J., et al. 1973. Systemic fluoride poisoning resulting from a
fluoride skin bum. 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.
Chem. Market. Rep. August 21.
Collins, G.H., Jr., et al. 1951. Absorption and excretion of inhaled
fluorides. Arch. Ind. Hyg. Occup. Med. 4: 585.
Oinman, 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
Nostranu Reinhoid Co., New York.
-------
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-
drogen fluoride —II. The response following exposure to low concentra-
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. Cont. Assoc. 18: 395.
Mohamed, A. 1971. Induced recessive lethals in second chromosomes of
Drosophila melanogaster by hydrogen fluoride. In: Englung, H., Berry, W.,
eds. Proc. 2nd Internet. Clean Air Cong." New York": 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.
DHEW, 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.
DHEW National Institute for Occupational Safety and Health, March 1976.
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. Ln: 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.
-------
U.S. EPA. 1977a. Industrial process profiles for environmental use:
chapter 16. The fluorocarbon-hydrogen fluoride industry. U.S. Environ.
Prot. Agency. U.S. DHEW PB281-483,
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.
-------
Hydrogen Sulfide
Summary
Pertinent information could not be located on the
carcinogenicity, mutagenicity, or teratogeniclty of H2S•
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 (H2S; 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
fossil fuels (Standen, 1969).
-------
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 sys terns .
B. Food
H2S may be formed within the gastrointestinal tract
after the ingestion 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 H2S
in sewers and waste from industrial plants where animals
products are handled. Thus, there has been accidental poisoning
from H2S 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
the 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).
-------
IV. EFFECTS
A. Carcinogenic!ty
Pertinent information could not 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 ha's 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
1396-
-------
poisoning results in headache, Inflammatioa of the conjunctivae
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 teen reported to
die at a concentration of 1 ppm after long time exposure in
*
hard water= Verschueren (1974) reports a 96-hour LG50 value of
10 ppm for goldfish. Verschueren also reports on a large number
of fresh water fish with toxic effects resulting from exposure
13 9?-
-------
to H2S 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 Toxlcity, 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 H2S in wastewater treatment plants.
VI. EXISTING GUIDELINES AND STANDARDS
A. Human
The 8-hour, time-weighted average occupational
exposure limit for l^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
H2S 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 l^S (Verschueren, 1974).
-------
References
Baxter, C. F. and R. Van Reen. 1958a. 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. E., 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,
Niahi-Shimbashi, Mlnato-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 (l,2,3-cjl)pyrene
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.
-------
SPECIAL NOTATION
U.S. EPA"s Carcinogen Assessment Group (GAG) has evaluated
indeno(l,2,3-c,d}pyrene and has found sufficient evidence to
indicate that this compound is carcinogenic.
-------
INDENOf1 ,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 CLF.S.
EPA, 1979b).
Indenod ,2,3-cd]pyrene (IP; £22^12^ is one of tne 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.,
benz[b]fluoranthene), weak carcinogens (benz[a]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, 1979C).
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,
benzo[k]fluoranthene, benzo[j]fluoranthene, benzo[a]pyrene, benzo[g,h,i]-
perylene, and indeno[1,2,3-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 1.31* ng/m3 (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 m^ 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.
-------
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, 1979a)..
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, benz[b]-
fluoranthene, benzCh]fluoranthene, indeno[1,2,3-cd]pyrene, and benztalp^yrene)
not exceed 0.2 .ug/1. Occupational exposure criteria have been established
-------
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 (1979a) draft recommended criteria for PAH in water are
based upon the extrapolation of animal carcinogenicity data for benz[a]-
pyrene and dibenz[a,h]anthracene.
B. Aquatic
There are no standards or guidelines concerning allowable concen-
trations of IP in aquatic environments.
-------
INDENOC1,2,3-cd]PYRENE
REFERENCES
Basu, D.K., and J. Saxena. 1977. Analysis of raw and drinking water
samples for polynuclear aromatic hydrocarbons. EPA P.O. Mo. 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:
1050.
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.
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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.
-------
Isobutyl Alcohol
I. Introduction
Isobutyl alcohol (2-methyl-l-propanol, C,H1Q0; molecular weight
74.12) is a flammable, colorless, refractive liquid with an odor like of
amyl alcohol, but weaker. Isobutyl alcohol is used in the 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 the intestinal tract and
the lungs.
B. Distribution
No data were readily available.
C. Metabolism
Isobutyl alcohol is oxidized to isobutyraldehyde and isobutyric
acid in the rabbit, with further metabolism proceeding to acetone and carbon
dioxide. Some conjugation with glucuronic acid occurs in the rabbit and dog.
D. Elimination
Approximately 14% of isobutyl alcohol is excreted as urinary
conjugates in the rabbit.
IV, Effects
A. Garcinogenicity
Rats receiving isobutyl alcohol, either orally or subcutaneously,
one to two times a week for 495 to 643 days showed liver carcinomas and
-------
sarcomas, spleen sarcomas and myeloid leukemia (Gibel, e_t 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 racs developed Mallory's
alcoholic hyaline bodies in the liver, and were observed to have decreases
in fat, glycogen, and RNA in che liver.
E. Other Relevant 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 LDgQ of isobutyl alcohol for rates if 2.46 g/kg,
(Smith £t. 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
-------
VII. Information Sources
1. NCM Toxicology Data Bank,
2. Merch. Index, 9th ed.
3. NIOSH Registry of Toxic Effects of Chemical Substances, 1978.'
4. NCM Toxline.
5. Sax, I. "Dangerous Properties of Industrial Materials."
6. Proctor, N. and Hughes, J. " Chemical Hazards of the Workplace"
Lippincott Co., 1978.
7. Occupational Diseases. A Guide to Their Recognition, NIOSH
publication No. 77-181, 1977. • • •
8. Hunter, D. "The Diseases of Occupations" 5th ed., Hodder and
Stoughton, 1975.
-------
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.
-11K>-
-------
Ho. 121
Lead
Health and Environmental Effects
U.S. ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
APRIL 30, 1980
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LEAD
SUMMARY
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 LC5Q values
with the common fathead minnow, Pimephales p rgme_laj3, ranged
from 2,400-7,480 /jg/1 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 ug/1 for six species of freshwater fish. Lead
at 500 ug/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
-If 17-
-------
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.
-------
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 Health (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 106 jag/1 at pH 5.5 to 1 ^g/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 +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) .
-------
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 pg/1.
Lead does not move readily through stream beds
because i't '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 jug/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 8.5 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 pg/m , with excursions
of 10 pg/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/m3 (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
JW
-------
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 species 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-
-------
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. Hutagenicity
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'Riordan and Evans, 1974; 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 been published
(U.S. EPA, 1979). In experimental animals, on the other
hand, lead has repeatedly produced teratogenic effects (Cat-
-------
zione and Gray, 1941; Karnofsky ana Ridgway, 1952; iMcClain
and Becker, 1975; Carpenter and Ferm, 1977; Kimmel, et al.
1976). Positive results were shown by injection into the
yolk sac of chick embryos and by intravenous and intraperi-
toneal injection in rats and hamsters. Chronic administra-
tion of lead in the drinking water of pregnant rats at concen-
trations up to" 2sO pg/1 resulted in delayed fetal development
and fetal resorption without teratologic effects (Kimmel,
et al. 1976).
D. Other Reproductive Effects
Lead has caused miscarriages and stillbirths among
women working in the lead trades (Lane, 1949; Nogaki, 1953).
In addition, decreased sperm quality in lead-exposed human
males (Lancranjan, et al. 1975) and reduced fertility in
animals of both sexes (Stowe ana Goyer, 1971; Jacquet, et
al. 1975) have been reported.
E. Other Chronic Toxicity
There is considerable information in man concern-
ing the renal effects of lead in both adults and children
(Clarkson and Kench, 1956; Chisolm, 1968; Cramer, et al.
1974; Wedeen, et al. 1975). Two distinctive effects on
the kidney occur with lead absorption. One is reversiole
proximal tubular damage, which-is seen mainly with short-
term exposure. The other effect is reauced glomerular fil-
tration, which has generally been considered to be of a
slow, progressive nature. Human exposures to high concen-
trations of lead have also been associated with cerebrovas-
cular disease (Dingwall-Fordyce ana Lane, ±yb3), heart faiiut<-
-------
(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 TOKICITY
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 LC50 values, and that LC5Q values for
lead are greatly different in hard and soft water. Between
soft and hard water, the LC5Q 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 LCcg values for freshwater fish are
2,400 to 7,480 pg/1 for fathead minnows in soft water (Tarz-
well and Henderson, I960; Pickering and Henderson, 1966),
482,000 for fathead minnows in hard water (Pickering and
Henderson, 1966), 23,800 ^g/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 (Tubifjey sn.) obtained from tests conducted at pH
-------
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 LC5C, 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 pg/1 for lake trout (Sauter, et al. 1976) to 87
jug/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 Daphnj^a 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- '
sured the lead concentration which reduced C02 fixation
by 50 percent. These values range from 500 pg/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 ug/nr.
The U.S. EPA (1979) has derived a draft criterion
for lead of 50 ;ug/l 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.
-------
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. Comra. 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 in Proc. Int.
Symp. Environ. Health, Aspects of Lead. Amsterdam, 2-6
Oct., 1972. Comra. 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. Perm. 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 gaj.rdneri) in hard and soft
water. Water Res. 10: 1991
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 le^d. Revised criteria 1973,
National Institute for Occupational Safety and Health,
DHEW (NIOSH) Publication No. 78-158.
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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-
lism by use of stable isotope tracers. Environ. Health
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 Myt LI is 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-
cal technology. Interscience Publishers, New York,
Stowe, H.D., and R.A. Goyer. 1971. The reproductive ability
and progeny of FT lead-toxic rats. Fertil. Steril. 22:
755. l
Tarzwell, C.M. , and C. Henderson. I960, Toxicity of less'
common metals to fishes. Ind. Wastes 5: 12.
-------
U.S. EPA. 1979. Lead: Ambient Water Quality Criteria.
U.S. Environ. Prot. Agency, Washington, D.C.
Van Esch, G.J., et al. 1962. The induction of renal tumors
by feeding basic lead acetate to rats. Br. Jour. Cancer
16: 289.
Wedeen, R.P.r et al. 1975. Occupational lead nephropathy,
Am. Jour-. Med. 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 cetension 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-
hungen zu Den Entsprechenden Neubildung des Menschen. (Kid-
ney adenomas and carcinomas in rats caused by chronic lead
poisoning and their relationship to corresponding human
neoplasma). Virchow Arch. Pathol. Anat. 323: 694.
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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 acc-uracy.
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151
MALE1C 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 {C4H2O3; 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 that
-------
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 et 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).
jVThis 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).
-------
III. PHARMACOKINETICS
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 C02 and H20.
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-
f
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).
-------
V. AQUATIC EFFECTS
The 24 to 96-hr median threshold limit (TLm) for maleic
anhydride in mosquito fish is 230-240 rag/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 ppra in air (39CFR23540).
-------
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.
Br. 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. J7:645-659
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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