ฃEPA
United States
Environmental Protection
Agency
Office of Water
Regulations and Standards
Criteria and Standards Division
Washington DC 20460
EPA 440.'5-80-047
October 1980
Ambient
Water Quality
Criteria for
Endrin
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AMBIENT WATER QUALITY CRITERIA FOR
ENDRIN
Prepared By
U.S. ENVIRONMENTAL PROTECTION AGENCY
Office of Water Regulations and Standards
Criteria and Standards Division
Washington, D.C.
Office of Research and Development
Environmental Criteria and Assessment Office
Cincinnati, Ohio
Carcinogen Assessment Group
Washington, D.C.
Environmental Research Laboratories
Corvalis, Oregon
Duluth, Minnesota
Gulf Breeze, Florida
Narragansett, Rhode Island
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DISCLAIMER
This report has been reviewed by the Environmental Criteria and
Assessment Office, U.S. Environmental Protection Agency, and approved
for publication. Mention of trade names or commercial products does not
constitute endorsement or recommendation for use.
AVAILABILITY NOTICE
This document is available to the public through the National
Technical Information Service, (NTIS), Springfield, Virginia 22161.
11
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FOREWORD
Section 304 (a)(l) of the Clean Water Act of 1977 (P.L. 95-217),
requires the Administrator of the Environmental Protection Agency to
publish criteria for water quality accurately reflecting the latest
scientific knowledge on the kind and extent of all identifiable effects
on health and welfare which may be expected from the presence of
pollutants in any body of water, including ground water. Proposed water
quality criteria for the 65 toxic pollutants listed under section 307
(a)(l) of the Clean Water Act were developed and a notice of their
availability was published for public comment on March 15, 1979 (44 FR
15926), July 25, 1979 (44 FR 43660), and October 1, 1979 (44 FR 56628).
This document is a revision of those proposed criteria based upon a
consideration of comments received from other Federal Agencies, State
agencies, special interest groups, and individual scientists. The
criteria contained in this document replace any previously published EPA
criteria for the 65 pollutants. This criterion document is also
published in satisifaction of paragraph 11 of the Settlement Agreement
in Natural Resources Defense Counci 1, et. alI. vs. Train, 8 ERC 2120
(D.D.C. 1976), modified, 12 ERC 1833 (D.D.C. 1979).
The term "water quality criteria" is used in two sections of the
Clean Water Act, section 304 (a)(l) and section 303 (c)(2). The term has
a different program impact in each section. In section 304, the term
represents a non-regulatory, scientific assessment of ecological ef-
fects. The criteria presented in this publication are such scientific
assessments. Such water quality criteria associated with specific
stream uses when adopted as State water quality standards under section
303 become enforceable maximum acceptable levels of a pollutant in
ambient waters. The water quality criteria adopted in the State water
quality standards could have the same numerical limits as the criteria
developed under section 304. However, in many situations States may want
to adjust water quality criteria developed under section 304 to reflect
Tocal environmental conditions and human exposure patterns before
incorporation into water quality standards. It is not until their
adoption as part of the State water quality standards that the criteria
become regulatory.
Guidelines to assist the States in the modification of criteria
presented in this document, in the development of water quality
standards, and in other water-related programs of this Agency, are being
developed by EPA.
STEVEN SCHATZOW
Deputy Assistant Administrator
Office of Water Regulations and Standards
111
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ACKNOWLEDGEMENT
Aquatic Life Toxicology:
Donald I. Mount, ERL-Duluth
U.S. Environmental Protection Agency
David J. Hansen, ERL-Gulf Breeze
U.S. Environmental Protection Agency
Mammalian Toxicology and Human Health Effects:
Robert Kavlock, HERL
U.S. Environmental Protection Agency
Steven D. Lutkenhoff, ECAO-Cin
U.S. Environmental Protection Agency
Jeanne Manson
University of Cincinnati
Robert E. Menzer
University of Maryland
Parrel 1 R. Robinson
Purdue University
Sorrel 1 Schwartz
Georgetown University
Jerry F. Stara, ECAO-Cin
U.S. Environmental Protection Agency
Fred Oehme (author)
Kansas State University
Caryn Woodhouse (doc. mgr.) ECAO-Cin
U.S. Environmental Protection Agency
Bonnie Smith (doc. mgr.) ECAO-Cin
U.S. Environmental Protection Agency
Joseph Arcos
Tulane University
Robert M. Bruce, ECAO-RTP
U.S. Environmental Protection Agency
Edward Calabrese
University of Massachusetts
Jacqueline V. Carr
U.S. Environmental Protection Agency
William B. Deichmann
University of Miami
Patrick Durkin
Syracuse Research Corporation
Pamela Ford
Rocky Mountain Poison Center
Larry Fradkin, ECAO-Cin
U.S. Environmental Protection Agency
Earl Gray, HERL
U.S. Environmental Protection Agency
Technical Support Services Staff: D.J. Reisman, M.A. Garlough, B.L. Zwayer,
P.A. Daunt, K.S. Edwards, T.A. Scandura, A.T. Pressley, C.A. Cooper,
M.M. Denessen.
Clerical Staff: C.A. Haynes, S.J. Faehr, L.A. Wade, D. Jones, B.J. Bordicks,
B.J. Quesnell, C. Russom, R. Rubinstein.
IV
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TABLE OF CONTENTS
Page
Criteria Summary
Introduction A-l
Aquatic Life Toxicology B-l
Introduction B-l
Effects B-2
Acute Toxicity B-2
Chronic Toxicity B-5
Plant Effects B-8
Residues B-8
Miscellaneous B-ll
Summary B-ll
Criteria 8-12
References 8-36
Mammalian Toxicology and Human Health Effects C-l
Introduction C-l
Exposure C-4
Ingest ion from Water C-4
Ingestion from Food C-5
Inhalation C-ll
Dermal C-12
Pharmacokinetics C-12
Absorption C-12
Distribution C-13
Metabolism C-15
Excretion C-17
Effects C-18
Acute, Subacute and Chronic Toxicity C-18
Synergism and/or Antagonism C-29
Teratogenicity C-29
Mutagenicity C-32
Carcinogenicity C-33
Criterion Formulation C-34
Existing Standards and Guidelines C-34
Current Levels of Exposure C-35
Special Groups at Risk C-35
Basis and Derivation of Criteria C-36
References C-41
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CRITERIA DOCUMENT
ENDRIN
CRITERION
Aquatic Life
For endrin the criterion to protect freshwater aquatic life as
derived using the Guidelines is 0.0023 ug/1 as a 24-hour average,
and the concentration should not exceed 0.18 ug/1 at any time.
For endrin the criterion to protect saltwater aquatic life as
derived using the Guidelines is 0.0023 ug/1 as a 24-hour average,
and the concentration should not exceed 0.037 ug/1 at any time.
Human Health
The ambient water quality criterion for endrin is recommended
to be identical to the existing water standard which is 1.0 ug/1.
Analysis of the toxic effects data resulted in a calculated level
which is protective of human health against the ingestion of con-
taminated water and contaminated aquatic organisms. The calculated
value is comparable to the present standard. For this reason a
selective criterion based on exposure solely from consumption of
6.5 g of aquatic organisms was not derived.
VI
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INTRODUCTION
Endrin is the conunon name of one member of the cyclodiene
group of pesticides. It is a cyclic hydrocarbon having a chlorine-
substituted methanobridge structure. Chemically pure endrin is a
white crystalline solid, while the technical compound is a light
tan powder. The specific gravity of this compound is 1.7 at 20ฐC;
the vapor pressure is 2.7 x 10~ at 25ฐC; and the substance begins
to decompose at 200ฐC. Endrin was introduced into the United
States in 1951. The endrin sold in the United States is a technical
grade product, containing not less than 95 percent active ingre-
dient, available in a variety of diluted formulations (Brooks,
1974). Jarvinen and Tyo (1978) found the solubility of endrin to
be about 200 yg/1.
Known uses of endrin in the United States are as an avicide,
rodenticide, and insecticide, the latter being the most prevalent.
The largest single use of endrin domestically is for the control of
lepidopteron larvae attacking cotton crops in the southeastern and
Mississippi delta states. Its persistence in soil led to its dis-
continuation for control of tobacco worms. Thus, endrin enters the
environment primarily as a result of direct applications to soil
and crops. Waste material discharge from endrin manufacturing and
formulating plants and disposal of empty containers also contribute
significantly to observed residue levels. In the past several
years, endrin utilization has been increasingly restricted and pro-
duction has continued to decline. In 1978, endrin production was
approximately 400,000 Ibs (U.S. EPA, 1978).
A-l
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REFERENCES
Brooks, G.T. 1974. Chlorinated Insecticides. Vol. I. Technology
and Application. CRC Press, Cleveland, Ohio.
Jarvinen, A.W. and R.M. Tyo. 1978. Toxicity to fathead minnows of
endrin in food and water. Arch. Environ. Contam. Toxicol. 7: 409.
U.S. EPA. 1978. Endrin-position document 2/3. Special Pest.
Rev. Div., Off. Pest. Prog., Washington, D.C.
A-2
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Aquatic Life Toxicology*
INTRODUCTION
Endrin is one of a group of chlorinated hydrocarbon pesticides developed
after the broad scale use of DDT, and it was increasingly used during the
1950's. Perhaps because endrin has a high acute toxicity to aquatic organ-
isms, it was more frequently tested in aquatic toxicity tests than related
insecticides such as chlordane, heptachlor, and aldrin.
Because it is a broad spectrum pesticide, endrin was used to control
many pests including termites, mice, and army worms. In the latter 1960's
it was extensively used for cotton bollworm control. Its persistence in
soil, while good for termite control, led to its discontinuation for control
of tobacco worms. Early testing identified its high toxicity to mammals.
Endrin is very insoluble in water. Recently, Jarvinen and Tyo (1978)
used a saturator in their toxicity tests and found the solubility in fresh-
water to be about 200 wg/l. Nearly all of the early work with endrin and
aquatic animals used acetone or some other carrier solvent, and in those few
tests where concentrations were measured, the actual concentrations were
frequently considerably lower than the calculated ones. Some workers used a
wetting agent such as Tritonฎ X-100 in the acetone-endrin solution to im-
prove dispersion in the test water. Because concentrations were not mea-
sured, the toxicity data reported may not reflect the true toxicity.
Ferguson and co-workers at Mississippi State University have published
numerous articles on endrin resistance that developed in natural populations
*The reader is referred to the Guidelines for Deriving Water Quality Cri-
teria for the Protection of Aquatic Life and Its Uses in order to better un-
derstand the following discussion and recommendation. The following tables
contain the appropriate data that were found in the literature, and at the
bottom of each table are calculations for deriving various measures of tox-
icity as described in the Guidelines.
B-l
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of freshwater- aquatic organisms exposed to high water concentrations of en-
drin as a result of its use on cotton. Nearly all of their work used stat-
ic, unmeasured test procedures, and the data, at best, can be used only for
relative toxicity purposes. Clearly, they did demonstrate a marked in-
creased tolerance to endrin of a variety of species. None of the data on
resistant populations has been included here since the criterion is expected
to protect unacclimated populations as well. Ferguson and co-workers also
showed that where resistant populations of organisms were found, top preda-
tors were absent; this demonstrates that acquisition of resistance is costly
to species most important to humans.
The acute toxicity of endrin to saltwater organisms has been relatively
well studied, particularly in the 1960's, although data on bioaccumulation
of endrin and its chronic toxicity have been available only recently. Al-
though the criterion for endrin is based on its bioconcentration, acute and
chronic toxicity to invertebrate and fish species is only slightly above the
Final Chronic Value. The similarity of these values is significant because
only slight excursions above the Final Chronic Value may result in acute or
chronic toxicity.
EFFECTS
Acute Toxicity
Twenty-two standard LC5Q values have been reported for 15 freshwater
invertebrate species (Table 1). None of the 22 data were based on measured
concentrations, and only two were based on flow-through procedures. Most of
the invertebrate species tested were substantially more tolerant than
fishes, with few exceptions. Glass shrimp and stoneflies were comparable to
fishes in sensitivity. Daphnia magna was among the more tolerant species.
The generally higher tolerance of the insects and related groups was un-
expected since endrin was an effective insecticide.
8-2
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Table 1 also lists data on the acute toxicity of endrin to 13 freshwater
fish species. Only six of these data from three different papers were based
on measured concentrations and flow-through procedures. Of the seven data
points for fathead minnows, two were derived from static tests with unmea-
sured concentrations, and these were not used in calculating the species
mean acute value for that species. Few results of flow-through tests based
on measured concentrations are available, due in part to the limited use of
gas chromatography during the earlier work and the high toxicity which re-
quired very low detection limits which were not achievable until analytical
procedures were improved.
All of the species mean acute values for freshwater fish species are be-
tween 0.15 and 2.1 ug/1, suggesting a relatively narrow range of species
sensitivity for fishes. The Freshwater Final Acute Value for endrin, de-
rived from the species mean acute values listed in Table 3 using the proce-
dure described in the Guidelines, is 0.18 ug/1.
Acute toxicity tests with saltwater invertebrate species also demon-
strate that endrin is very toxic (Tables 1 and 6). The variability in
LC50 or EC50 values was greater than that for fishes, ranging from 0.037
ug/1 for pink shrimp to 790 ug/1 for the American oyster. The sensitivity
of arthropods to endrin was not much different from the sensitivity of
fishes. The penaeid shrimp was the most sensitive invertebrate family
tested, with LC values from 0.037 to 0.3 ug/l (Schimmel, et al. 1975;
Lowe, unpublished; Butler, 1963). LC5Q values for five other arthropod
species ranged from 0.3 ug/1 for Korean shrimp (Schoettger, 1970) to 25 ug/1
for blue crab (Butler, 1963). The sensitivity of different life stages of
grass shrimp is similar, differing by only a factor of 3.4 (Tyler-Schroeder,
1979). American oysters were less sensitive than arthropods, with EC50
B-3
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values ranging from 14.2 to 790 ug/1, based on decreased shell deposition or
abnormal development of larvae (Table 1).
The toxicity of endrin to oysters may vary depending on water tempera-
ture. Concentrations decreasing growth by 50 percent were: 400 ug/1 at
12'Ci 33 ug/1 at 24*C; and 14 ug/1 at 22ฐC (Table 1); these differences may
however, be reflective of different laboratories' procedures. The range in
LC50 values for saltwater invertebrate species was only slightly greater
than the range for 17 species of freshwater molluscs and arthropods; the
pink shrimp was the most sensitive invertebrate species tested.
Acute toxicity tests have been conducted with 17 species of saltwater
fishes, and sensitivity varies (Tables 1 and 6) from 0.048 ug/1 for chinook
salmon (Schoettger, 1970) to 3.1 ug/1 for northern puffer (Eisler, 1970b).
Only two (usually tolerant) species, the sheepshead minnow (Hansen, et al.
1977) and the sailfin molly (Schimmel, et al. 1975) have been tested for 96
hours in flow-through tests with measured endrin concentrations. Sheepshead
minnow fry, juveniles, and adults did not differ in their sensitivity to
acute exposure to endrin (Hansen, et al. 1977).
Data on LCgg values for saltwater invertebrate species from acute tox-
icity tests on endrin support the hypothesis that the acute toxicity of en-
drin is underestimated by static tests and by not measuring concentrations
of endrin in test water. Acute values based on nominal concentrations for
grass shrimp, and American oysters were higher than acute values for mea-
sured concentrations (Table 1). Additionally, LCrg values based on static
tests were greater than LC50 values for flow-through tests of the same
duration for sheepshead minnow, sailfin mollies, shiner perch, dwarf perch,
Korean shrimp, pink shrimp, and grass shrimp (Eisler, 1969; Schoettger,
1970; Earnest and Benville, 1972; Schimmel, et al. 1975; Tyler-Schroeder,
1979).
B-4
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The Saltwater Final Acute Value for endrin, derived from the species
mean acute values listed in Table 3 using the procedures described in the
Guidelines, is 0.037 ug/1.
Chronic Toxicity
Life-cycle chronic tests have been completed with fathead minnow and
flagfish, giving chronic values of 0.19 and 0.26 yg/1, respectively (Table
2). Mount (1962), working with the bluntnose minnow, a species closely re-
lated to the fathead minnow, found a no-observed-effect concentration be-
tween 0.1 and 0.4 ug/1 for a 291-day exposure (Table 6). Spawning did not
occur in this test, but the results are consistent with those for the fat-
head minnow.
Jarvinen and Tyo (1978) demonstrated that: (1) when food is contaminated
with endrin, the toxicity of endrin in the water is greater than when uncon-
taminated food is fed; (2) the contribution of endrin to the body burden by
food is only 10 to 15 percent of that contributed by water, and (3) residues
contributed by food were additive to those contributed by water. Unfortun-
ately, the existing data base is not sufficient to make a precise allowance
for exposure through both routes for various species.
One saltwater invertebrate species, grass shrimp, has been exposed to
endrin in a partial-life-cycle toxicity test (Table 2). Surivival of the
parental generation was reduced by exposure to 0.11 yg/1 (Tyler-Schroeder,
1979). Onset and duration of spawning were significantly delayed and
lengthened for female grass shrimp at all exposure concentrations (0.03 to
0.79 ug/1). The number of females depositing embryos was less than that of
controls, but embryo production and hatching success apparently were not af-
fected. Larval mortality increased, time to metamorphosis increased, and
growth of juvenile shrimp was decreased by endrin concentrations of 0.11
ug/1 and higher. A chronic value of 0.039 yg/1 endrin was obtained for
B-5
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grass shrimp (Table 2), even though all tested concentrations significantly
impaired some life-cycle function. A lower limit of 0.03 yg/1 was selected
because the only effect was a delay in onset of spawning of about one week;
a delay of one week probably would not affect natural populations. The up-
per limit of 0.05 ug/1 was set based on decrease in number of ovigerous fe-
males and delay in spawning of 3 to 4 weeks.
Sheepshead minnows (Schimmel, et al 1975; Hansen, et al. 1977), spot
(Lowe, 1966), and mummichog (Eisler, 1970a) have been exposed to endrin for
10 days or longer (Tables 2 and 6). Of these tests with saltwater fish spe-
cies, only the life-cycle exposure of sheepshead minnows (Hansen, et al.
1977) is suitable for obtaining a chronic value (Table 2). In this test,
embryos exposed to 0.31 and 0.72 ug/1 hatched early; all fry exposed to 0.72
ug/1, and about one-half of those exposed to 0.31 yg/1, died. Females died
during spawning, fewer embryos were fertile, and survival of exposed progeny
decreased in 0.31 ug/1. No significant effects were observed on survival,
growth, or reproduction at an exposure concentration of 0.12 ug/1. The
chronic limits, 0.12 to 0.31 ug/1, were not much less than the 96-hour
LC,JQ of 0.34 ug/1. indicating that there is little difference between en-
drin concentrations that produce acute effects and the highest that produce
no observed effect in chronic tests. Life-cycle tests with the freshwater
fish species, fathead minnow and flagfish, also show little difference be-
tween acute and chronic toxicity of endrin (Table 2).
An early-life-stage test with sheepshead minnows (Schimmel, et al. 1975)
was not used to obtain a chronic value because only LCcn values and nomi-
nal observed no-effect concentrations were reported (Table 6); however, re-
sults were similar to those reported in the life cycle test (Table 2). The
LCc;Q value based on measured concentrations for fry on the 33rd day of the
B-6
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experiment was 0.16 ug/1. Although mortality in fish exposed to a nominal
concentration of 0.31 ug/1 was not significant, they were visibly affected
by endrin.
Fifty-seven percent of the juvenile spot exposed to 0.075 ug/1 of endrin
died within the first 19 days of an eight-month test; those exposed to 0.05
ug/1 were apparently not affected (Lowe, 1966) (Table 6). Spot exposed to
0.05 ug/1 for eight months exhibited no signs of poisoning, and their survi-
val, length, and weight did not differ from those of control fish. The
nominal, no-observed-effect concentration of 0.05 ug/1 was 0.11 of the
LCcQ of 0.45 ug/U and this also tends to support the conclusion of a min-
imal difference between the acute and chronic toxicity of endrin to fishes.
The only other datum on >96-hour effects of endrin on a saltwater fish
species is a 10-day LC^Q of 0.33 ug/1 for the mummichog based on nominal
concentrations (Eisler, 1970a); this is little different from the 96-hour
LC5Q values of 0.6 and 1.5 ug/1 (Eisler, 19706).
The acute-chronic ratios for the fathead minnow and the flagfish are 2.2
and 3.3, respectively (Table 2). For saltwater species, the acute-chronic
ratios for the sheepshead minnow and grass shrimp are 1.9 and 18, respect-
ively. The species mean acute values and acute-chronic ratios are summar-
ized in Table 3.
Dividing the Freshwater Final Acute Value of 0.18 yg/l by the geometric
mean (4.0) of the four acute-chronic ratios (Table 2) gives the Freshwater
Final Chronic Value of 0.045 ug/1 (Table 3). Dividing the Saltwater Final
Acute Value of 0.037 ug/1 by the geometric mean of acute-chronic ratios
(4.0) gives the Saltwater Final Chronic Value of 0.0093 ug/l (Table 3).
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Plant Effects
Data on the toxicity of endrin to five freshwater species of algae are
listed in Table 4. Apparently, algae are not sensitive to endrin, and the
lowest effect level for plants is 475 ug/1 based on growth inhibition of
Anacystis nidularas.
Three published studies on five species of saltwater algae and a natural
phytoplankton community (Table 4) indicate that, like for freshwater spe-
cies, effects of endrin on these plant species are unlikely at concentra-
tions protective from acute effects on most invertebrate and fish species.
Menzel, et al. (1970) in tests with four phytoplankton species found effects
at concentrations greater than 1 ug/1. Productivity of natural phytoplank-
ton communities was reduced by 46 percent in 1,000 ug/1 (Butler, 1963).
Growth rate of Agmenellum quadruplicatum was reduced in as little as 0.2
ug/1 (Batterton, et al. 1971). Because none of the values is based on
measured concentrations, a Final Plant Value has not been established.
Residues
Steady-state bioconcentration factors (BCF) have been measured for seven
species of freshwater organisms (Table 5) including algae (140-122), mussels
(3,000), and fishes (1,640-15,000).
Endrin seems to enter the body rapidly as indicated by the short time
required for the tissues to reach equilibrium with the water concentration
(Jarvinen and Tyo, 1978). The short biological half-life, as observed by
Jackson (1976) (Table 6), demonstrates that endrin is different from related
pesticides such as DOT. Jarvinen and Tyo (1978) observed metabolites of
endrin in the tissues of their test fish suggesting an important rate of
degradation as well as elimination.
The bioconcentration of endrin from water into the tissues of saltwater
organisms has also been well studied (Tables 5 and 6). Steady-state biocon-
8-8
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centration factors are available from studies with American oysters (Mason
and Rowe, 1976), grass shrimp (Tyler-Schroeder, 1979), sheepshead minnows
(Hansen, et al. 1977; Schimmel, et al. 1975), and spot (Lowe, 1966). Addi-
tional endrin BCF data (Table 6) are available from 96-hour exposures of oy-
sters, grass shrimp, pink shrimp, sheepshead minnows, and sailfin mollies
(Wilson, 1966; Schimmel, et al. 1975).
Biconcentration factors (Table 5) for endrin in American oysters ex-
posed for seven days ranged from 1,670 to 2,780 (Mason and Rowe, 1976). En-
drin accumulated rapidly, reaching steady-state after about 48 hours of ex-
posure. Oysters placed in endrin-free water depurated endrin at a rate of
0.005 vg/g/hour, resulting in a biological half-life of 67 hours. Based on
this experiment, the oysters exposed to endrin for 10 days in a flow-through
test by Wilson (1966) were probably at steady-state and had a BCF of 1,000,
based on a nominal water concentration. Oysters exposed for only 96 hours
contained 1,200 times the concentration in the exposure water (Schimmel, et
al. 1975).
Bioconcentration factors for endrin from two experiments with grass
shrimp averaged 1,490 and 1,600 (Tyler-Schroeder, 1979). In the first ex-
periment, steady-state was reached after 2.5 days of a 21-day exposure.
Ninety percent of the endrin was depurated within 4.2 days. In the second
experiment, the average BCF of endrin was 1,600 in parental generation
shrimp from a partial-life-cycle exposure lasting five months. Average bio-
concentration factors after a 96-hour exposure were 830 for grass shrimp and
980 for pink shrimp (Schimmel, et al. 1975).
Bioconcentration data for two of three species of saltwater fishes dif-
fer little from those for invertebrate species. Bioconcentration factors
calculated from nominal water concentrations were 1,340 for spot exposed for
eight months and 1,560 for spot exposed five months (Lowe, 1966). The aver-
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age BCF for juvenile sheepshead minnows exposed for 28 days was 2,500; for
adults exposed for for 141 to 161 days the BCF was 6,400 (Hansen, et al.
1977), and for juvenile exposed for four days the BCF was 2,600 (Schimmel,
et al. 1975). Sailfin mollies exposed to endrin for four days had an aver-
age BCF of 2,400 (Schimmel, et al. 1975).
The geometric mean of normalized BCF values for endrin for freshwater
and saltwater aquatic life is 1,324 (Table 5). This value was obtained by
first dividing each BCF for which a percent lipid value is available by that
percent lipid value to obtain a normalized BCF, which is what the BCF would
be if the percent lipids were 1 percent. Normalized BCF values obtained
were: fathead minnow, 1,892; sheepshead minnow, 694 and 1,778; spot, 1,318.
The geometric mean of all freshwater and saltwater normalized BCF values was
then calculated.
Dividing the FDA action level of 0.3 mg/kg for edible fish and shellfish
by the geometric mean of normalized BCF values (1,324) and by a percent
lipid value of 15 for freshwater species (see Guidelines) gives a freshwater
residue value of 0.015 yg/1. Dividing the FDA level by the geometric mean
of normalized BCF values and by a percent lipid value of 16 for saltwater
species (see Guidelines), a saltwater residue value of 0.014 ug/l is calcu-
lated similarly. Dividing the FDA action level of 0.3 mg/kg by the highest
BCF for edible portion of an edible species, 2,780 for oyster (Mason and
Rowe, 1976), provides an additional residue value for saltwater species of
0.11 ug/1.
Dividing the FDA action level of 0.3 mg/kg for fish oil by the geometric
mean of normalized BCF values (1,324) and by a percent lipid value of 100
for fish oil gives a residue value for freshwater and saltwater of 0.0023
ug/1.
8-10
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Other available residue data for effect levels are not appropriate for
calculation of freshwater or saltwater residue values for wildlife protec-
tion. Therefore, the lowest residue value of 0.0023 ug/1 is taken as the
Freshwater Final Residue Value and the Saltwater Final Residue Value. The
Final Residue Value may be too high because, on the average, the concentra-
tion in 50 percent of species similar to those used to derive the value will
exceed the FDA action level.
Miscellaneous
Table 6, containing additional data for other effects not listed in the
first five tables, does not indicate any significant effect levels that
would alter the conclusions discussed previously.
Summary
Acute data are available for 28 freshwater species including a wide var-
iety of organisms normally performing a spectrum of community functions.
Only one of the 28 species has an acute value above 100 ug/1, the lowest
species mean acute value is 0.15 ug/1, and most values are clustered near
1.0 wg/1. The data are predominantly from static tests in which toxicant
concentrations were not measured and so probably underestimate true tox-
icity. The Freshwater Final Acute Value is 0.18 ug/1.
There are acute data for 21 species of saltwater organisms. None of the
values is above 14.2 vgf\ and four are below 0.1 ug/1. The Saltwater Final
Acute Value is 0.037 ug/1, one-fifth that of the freshwater.
Life cycle tests with two freshwater fish species gave chronic endpoints
near 0.2 yg/1 and acute-chronic ratios of 3.3 and 2.2. Chronic data for the
sheepshead minnow gave comparable estimates of 0.19 ug/1 and 1.9 for the
chronic value and acute-chronic ratio, respectively. The saltwater grass
shrimp was more sensitive (0.039 ug/1) and gave a much larger acute-chronic
ratio of 18. The geometric mean of these four estimates of the acute-
B-ll
-------
chronic ratio is 4.0. Using this value and the Freshwater and Saltwater
Final Acute Values, the Freshwater Final Chronic Value is calculated to be
0.045 ug/1 and the Saltwater Final Chronic Value is 0.0093 ug/1.
The residue data for freshwater and saltwater are similar and show rela-
tively low bioconcentration factors as compared to related insecticides such
as dieldrin. Further, the data agree that endrin uptake reaches steady-
state quickly and is depurated quickly. Using the FDA action level of 0.3
mg/kg for fish oil, the geometric mean of normalized bioconcentration fac-
tors (1,324), and a percent lipid value of 100 for fish oil, a Final Residue
Value of 0.0023 ug/1 is calculated for both freshwater and saltwater. The
Final Residue Value may be too high because, on the average, the concentra-
tion in 50 percent of species similar to those used to derive the value will
exceed the FDA action level.
The plant data clearly indicate that plants are much more resistant than
animals. Effect levels for plants are above 475 wg/1. Other data do not
reveal any more sensitive effects. Saltwater algae appear more sensitive
than freshwater, but all values are above 1 wg/1 except one. Therefore,
plant protection seems certain if animals are protected. Other data avail-
able do not suggest any lower effect levels.
CRITERIA
For endrin the criterion to protect freshwater aquatic life as derived
using the Guidelines is 0.0023 ug/1 as a 24-hour average, and the concentra-
tion should not exceed 0.18 ug/1 at any time.
For endrin the criterion to protect saltwater aquatic life as derived
using the Guidelines is 0.0023 ug/1 as a 24-hour average, and the concentra-
tion should not exceed 0.037 ug/1 at any time.
B-12
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-------
Table I. (Continued)
Species Mean
LC50/EC50 Acutซ Value
Sp*cles Method* (ug/l) (fg/O Reference
Shiner perch, FT, U 0.12
CymatcKjaster aggregate
Dwarf perch, S, U 0.6
Mlcrometrus minimus
Dwarf perch, FT, U 0. 13
Mlcrometrus minimus
aiuehead, S, U O.I
Thalassoma blfasclatum
Striped mullet, S, U 0.3
Mug II cephalus
Northern puffer, S, U 3.1
Sphaeroldes maculatus
0.31 Earnest & Benvllle,
J972
Earnest & Benvllle,
1972
0.28 Earnest & Benvllle,
1972
0.1 Elsler, 19706
0.3 Elsler, I970b
3.1 Elsler, I970b
* S = static, FT = flow-through, 0 = unmeasured, M = measured
*'Abnormal development of oyster larvae; decreased growth of oyster; or loss of equilibrium
of brown shrimp or blue crabs.
B-20
-------
Table 2. Chronic values for endrln
Species
Fathead minnow,
Plmephales promelas
Flagflsh,
Jordanella tIorIdae
Test"
Limits
(tifl/l)
Chronic Value
(ng/l)
FRESHWATER SPECIES
LC 0.14-0.25 0.19
LC 0.22-0.3 0.26
Reference
Jarvlnen & Tyo, 1978
Hermanutz, 1978
Sheepshead minnow, LC
Cyprlnodon varlegatus
Grass shrimp, LC
Palaemonetes puglo
SALTWATER SPECIES
0.12-0.31 0.19
0.03-0.05** 0.039
Han sen, et al. 1977
Tyler-Schroeder, 1979
* LC = life cycle or partial life cycle
"Onset of spawning was delayed about one week In shrimp exposed to 0.03 ug/l.
Because a delay of one week would probably not affect natural populations,
limits were set on decreases In number of ovlgerous females and delayed spawning
of 3-4 weeks In 0.05 ug/l of endrln.
Acute-Chronic Ratios
Spec 1 es
Fathead minnow,
Plmephales proms las
Flagflsh,
Jordanella florldae
Sheepshead minnow.
Acute
Value
(ug/l)
0.42***
0.85
0.36
Chronic
Value
(ug/l)
0.19
0.26
0.19
Ratio
2.2
3.3
1.9
Cyprlnodon varlegatus
B-21
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Table 3. Species ซean acute values and acute-chronic ratios for endrln
ink*
2B
27
26
25
24
23
22
21
20
19
18
17
16
15
Species
FRESHWATER
Cladoceran,
Daphnla magna
Mayfly,
Hexagon la bJllneata
Copepod (cyclopold),
(unidentified)
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S 1 mocepha 1 us serrulatus
Crayfish,
Orconectes nals
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Gamma r us tacustrls
Mayfly,
Ephemera 1 1 a grand 1 s
Scud,
Gammarus fasclatus
Goldfish,
Carasslus auratus
Damself ly,
Ischnura vert leal us
Sow bug,
Asellus brevlcaudus
Glass shrimp,
Palaemonetes kadlakensls
Guppy,
Poecl Ha retlculata
Chinook salmon,
Oncorhynchus tshawytscha
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insecticides on two species of stonefly naiads. Jour. Water Pollut. Control
Fed. 38: 1273.
Katz, M. 1961. Acute toxicity of some organic insecticides to 3 species of
salmonids and the threespine stickleback. Trans. Am. Fish. Soc. 90: 264.
Katz, M. and G.G. Chadwick. 1961. Toxicity of endrin to some Pacific
Northwest fishes. Trans. Am. Fish. Soc. 90: 394.
B-39
-------
Korn, S. and R. Earnest. 1974. Acute toxicity of 20 insecticides to
striped bass, Morone saxatilis. Calif. Fish and Game. 60: 128.
Lincer, J.L., et al. 1970. DDT and endrin fish toxicity under static ver-
sus dynamic bioassay conditions. Trans. Am. Fish. Soc. 99: 13.
Lowe, J.I. Results of toxicity tests with fishes and macroinvertebrates.
Data sheets available from U.S. Environmental Protection Agency, Environ-
mental Research Laboratory, Gulf Breeze, Florida 32561.
Lowe, J.I. 1966. Some effects of endrin on estuarine fishes. Proc. 19th
Annual Conf. S.E. Assoc. Game and Fish Comm. p. 271.
Ludke, J.L., et al. 1968. Some endrin relationships in resistant and sus-
ceptible populations of golden shiners, Notemigonus crysoleucas. Trans. Am.
Fish. Soc. 97: 260.
Macek, K.J., et al. 1969. Effects of temperature on the susceptibility of
bluegills and rainbow trout to selected pesticides. Bull. Environ. Contam.
Toxicol. 4: 174.
Mason, J.W. and D.R. Rowe. 1976. Accumulation and loss of dieldrin and en-
drin in the eastern oyster. Arch. Environ. Contam. Toxicol. 4: 349.
Menzel, D.W., et al. 1970. Marine phytoplankton vary in their response to
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-------
Mount, O.I. 1962. Chronic effects of endn'n on bluntnose minnows and gup-
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pi kill. Proc. 31st North Am. Wildl. Nat. Res. Conf. p. 176.
Naqvi, S.M. and D.E. Ferguson. 1968. Pesticide tolerances of selected
freshwater invertebrates. Jour. Miss. Acad. Sci. 14: 121.
Nebeker, A.F. and A.R. Gaufin. 1964. Bioassays to determine pesticide tox-
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salmonid species. Bull. Environ. Contam. Toxicol. 6: 144.
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species of cladocerans. Trans. Am. Fish. Soc. 95: 165.
B-41
-------
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B-42
-------
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8-43
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Mammalian Toxicology and Human Health Effects
INTRODUCTION
Wild and domestic mammals are exposed to endrin primarily
through ingestion of treated foliage, although dermal contact and
inhalation also occur. Endrin shows little tendency to accumulate
in tissues other than adipose tissue; levels of up to 23.7 pg/g
have been detected both in internal and external fat in a variety
of species following ingestion of endrin-contaminated feed. Endrin
was still detectable in the fat of these animals 42 days after the
exposure (Long, et al. 1961).
Metabolism of endrin has been studied extensively in rats.
Endrin is readily metabolized in the liver and excreted as hydro-
philic metabolites. However, certain toxic metabolites such as 12-
ketoendrin (also known as 9-ketoendrin) can be retained for longer
periods of time. Rats excrete endrin and its metabolites primarily
in the feces; in rabbits, excretion is primarily via the urine.
Endrin is highly toxic to all animals regardless of the route
of exposure (Treon, et al. 1955). The primary toxic effect of
acute exposure is on the central nervous system. When lethal con-
centrations are administered to experimental animals, convulsions
may occur as soon as 30 minutes after exposure, and may culminate
in death through respiratory failure in about 48 hours. The dose
lethal to 50 percent of the experimental animals ranges from 3
mg/kg for the monkey to 50 mg/kg for the goat.
Many cases of mammalian fatalities have been reported outside
the laboratory. For example, field application of endrin at rates
of
-------
various species of wild mice inhabiting the target area (Dana and
Shaw, 1958).
The chronic toxicity of endrin to mammals is greater than that
of other organochlorine pesticides. Sublethal effects in wild ani-
mals manifest primarily as behavioral and reproductive disorders,
i.e., improper maternal care, temporary loss of normal activity,
increased vulnerability to predators, reduced reproductive poten-
tial, increased post-natal mortality, and fetal death. Chronic
exposure to endrin may also be fatal. Doses of 0.49 to 0.81 mg/kg
in the diet was fatal to dogs in 5 to 6 months. Twelve mg/kg in the
diet for life decreased the survival time for mice. Deer mice suc-
cumbed to a diet which contained only 2 mg/kg endrin.
No malignancies attributable to endrin exposure have been
reported in the literature; however, endrin has been found to cause
chromosomal aberrations in rats following intratesticular injec-
tion. Teratogenesis, growth retardation, and increases in fetal
mortality have been observed in mice and hamsters following endrin
administration.
Human exposure to endrin occurs through the diet, from inhala-
tion, and through dermal contact. The average dietary intake in
the United States in 1973 was 0.033 yg/day (0.0005 yg/kg/day) for a
69.1 kg man. This is far below the maximum daily intake of 138.2
yg/day (2 yg/kg/day) established by the World Health Organization
(WHO). Respiratory and/or dermal exposure to endrin occur during
manufacture and distribution, but are more likely to result from
agricultural uses.
C-2
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Outbreaks of human poisoning have resulted from acfcidental
contamination of foods and have been traced to doses as low as 0.2
mg/kg body weight. Endrin toxicity seems to result primarily from
the effects of endrin and its metabolites on the central nervous
system. Symptoms usually observed in victims of endrin poisoning
were convulsions, vomiting, abdominal pain, nausea, dizziness, and
headache. Respiratory failure was the most common cause of death.
Significantly increased activity of the hepatic microsomal drug-
metabolizing enzymes has occurred in individuals employed in the
manufacture of endrin. No reports of irreversible adverse effects
of occupational exposure to endrin have been found in the available
literature.
Food contamination by endrin still occurs, but to a decreasing
extent. At present, levels are approximately 4,000 times lower
than those acceptable to the World Health Organization. Background
concentrations in the atmosphere, hydrosphere, and lithosphere, far
removed from agricultural areas where endrin is used and industri-
alized areas where endrin is manufactured, are generally below the
levels of detection.
Humans ingest endrin-treated agricultural produce as well as
meat from domesticated and wild animals and fish which feed on con-
taminated vegetation. Ingestion of 20 mg endrin per day by cows
resulted in levels of up to 0.25 ug/g of endrin in milk. Aquatic
invertebrates and fish bioconcentrate considerable quantities of
endrin from water and pass it on to predatory birds. This contami-
nated fowl (or the fish themselves) may, in turn, be ingested by
humans.
C-3
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In animals, chronic exposure to endrin may result in damage to
the liver, kidneys, heart, brain, lung, adrenal glands, and spleen.
Effects, secondary to central nervous system disorders, have also
been observed following chronic exposure of mammals to sublethal
doses of endrin. These include behavioral abnormalities, changes
in carbohydrate metabolism, and changes in the composition of the
blood. Although no reports of malignancies attributable to endrin
have been found, chromosomal abnormalities and teratogenesis have
been induced by endrin in several mammalian species.
EXPOSURE
Ingestion from Water
Occasionally, groundwater may contain more than 0.1 ug/1 of
endrin, but levels as high as 3 yg/1 have been correlated with pre-
cipitation and runoff following endrin applications (U.S. EPA,
1978). Drinking water from Franklin, Louisiana, an area of high
endrin usage, was found to contain a maximum of 23 ng/1 (Lauer, et
al. 1966).
In a study conducted between March 1964 and June 1967, more
than 500 grab samples of finished drinking water and corresponding
raw water were collected from 10 selected municipal water treatment
plants whose source was either the Mississippi or the Missouri
Rivers. Of the 458 finished water samples assayed, 156 (34 per-
cent), contained detectable concentrations of endrin. However, the
number of finished water samples containing concentrations of
endrin in excess of 0.1 ug/1 decreased from 23 (10 percent) to 0 in
a three year period from 1964 to 1967 (Schafer, et al. 1969).
C-4
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A recent study of endrin contamination of drinking water was
conducted by the U.S. EPA (1974). Endrin was detected in the
finished water from the Carrollton Water Plant in New Orleans,
Louisiana. The highest concentration measured from all samples was
4 ng/1.
Ingestion from Food
The general population has little exposure to endrin in the
diet. In a series of analyses of total diets determined from "mar-
ket basket" samples in five regions of the United States, the total
average intake from food ranged from approximately 0.009 ug/kg body
weight per day in 1965 to 0.0005 ug/kg body weight per day in 1970
(Table 1) (Duggan and Lipscomb, 1969; Duggan and Corneliussen,
1972). The six year average intake was 0.005 ug/kg body weight per
day. A market basket consisted of 117 food items grouped into 12
composites required for the 14-day diet for a 16- to 19-year-old
male. All foods were treated normally before analysis, i.e., meats
were cooked, etc. The average daily intake remained at trace
levels throughout the period 1965 to 1970; however, the frequency
of occurrence decreased somewhat (Table 1) . The breakdown of
dietary endrin intake levels by food class is given in Table 2.
Processing of some foods before human consumption signifi-
cantly changed endrin residues. Endrin increased in soybean oils
(0.28 ppm) relative to whole crop levels (0.07 ppm) following the
extraction process (Hill, 1970). Storage longer than 12 weeks
decreased endrin residues in Irish and sweet potatoes by 20 percent
(Solar, et al. 1971). Heat processing and freezing further lowered
potato residues 65 and 52 percent, respectively. Studies on
C-5
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TABLE 1
Average Incidence and Daily Intake of Endrin*
Year
1965
1966
1967
1968
1969
1970
Percent Positive
Composites
2.8
2.0
1.7
1.1
3.3
1.4
Daily Intake
(rag)
Ta
T
T
0.001
T
T
mg/kg
body wt/day
0.000009
0.000004
0.000004
0.00001
0.000004
0.0000005
Source: Duggan and Corneliussen, 1972
aT = Trace, < 0.001 mg
C-6
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turnips (Wheeler, et al. 1969) and carrots (Hermanson, et al. 197C)
identified 50 to 80 percent of the endrin in the peels.
Endrin disappearance from growing and harvested crop is so
variable that half-life data for endrin persistence on food plants
should be viewed with skepticism (Hill, 1970). The loss of endrin
from crops depends on the sum of many factors, including tempera-
ture, volatilization, metabolism, and dislodgement by wind and
rain. Since generalizations cannot be made that endrin on a given
crop will always "disappear" at the same rate, residue analyses on
harvested crops are the most effective means of determining poten-
tial human exposure.
A bioconcentration factor (BCF) relates the concentration of a
chemical in aquatic animals to the concentration in the water in
which they live. The steady-state BCFs for a lipid-soluble com-
pound in the tissues of various aquatic animals seem to be propor-
tional to the percent lipid in the tissue. Thus, the per capita
ingestion of a lipid-soluble chemical can be estimated from the per
capita consumption of fish and shellfish, the weighted average per-
cent lipids of consumed fish and shellfish, and a steady-state BCF
for the chemical.
Data from a recent survey on fish and shellfish consumption in
the United States were analyzed by SRI International (U.S. EPA,
1980) . These data were used to estimate that the per capita con-
sumption of freshwater and estuarine fish and shellfish in the
United States is 6.5 g/day (Stephan, 1980). In addition, these
data were used with data on the fat content of the edible portion of
the same species to estimate that the weighted average percent
C-8
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lipids for consumed freshwater and estuarine fish and shellfish is
3.0 percent.
Two laboratory studies, in which percent lipids and a steady-
state BCF were measured, have been conducted on endrin. The mean
of the BCF values, after normilization to 1 percent lipids, is
1,324 (see Table 5 in Section B). An adjustment factor of 3 can be
used to adjust the mean normalized BCF to the 3.0 percent lipids
that is the weighted average for consumed fish and shellfish.
Thus, the weighted average BCF for endrin and the edible portion of
all freshwater and estuarine aquatic organisms consumed by Ameri-
cans is calculated to be 3,970.
Because of the dynamic state of endrin in the biological tis-
sues of lower animals (Mount, et al. 1966), the bioaccumulation is
short-lived, and tissue burdens diminish rapidly once the environ-
mental source is removed. (Toxic endrin metabolites, such as 12-
ketoendrin, may persist for longer periods of time.) Commercial
catfish from Arkansas and Mississippi were reported to contain
average residues in the edible portions, ranging from 0.01 to 0.41
yg/g. Four percent of the samples exceeded the U.S. Food and Drug
Administration (FDA) action level for maximum permissible endrin
concentration of 0.3 yg/g in the edible portion of fish (Hawthorne,
et al. 1974; Crockett, et al. 1975).
Humans may also be exposed to endrin in cow milk and steer,
lamb, and hog meat. However, endrin is so rapidly metabolized and
excreted that edible tissue levels are usually at or below the
dietary concentrations of endrin. Residue levels in excess of 0.25
ug/g on a fat basis were detected in the milk of 40 Wisconsin dairy
C-9
-------
herds between 1964 and 1967 (Moubry, et al. 1968) . Endrin was pre-
sumably retained in the milk fat for up to four weeks. However, the
quantities of endrin ingested during that period were not control-
led. Williams and Mills (1964) studied the excretion of endrin in
cows' milk under controlled feeding conditions. Endrin concentra-
tions in the milk increased progressively during the first few days
of feeding until they plateaued at 13 to 15 days. When ingestion of
endrin ceased, residues in milk declined sharply and following 20
days on an endrin-free diet, detectable (>0.001 ug/g) levels were
present only in milk samples from cows fed the highest levels of
endrin (0.3 mg/kg). However, in this study animals were fed a mix-
ture of pesticides, thus, interactions may have occurred. Endrin
is apparently excreted in milk in higher concentrations when fed as
a residue on hay than when fed dissolved in soybean oil (Ely, et al.
1957). However, in general, a total daily endrin intake of > 20 mg
as a residue sprayed in forage is necessary for excretion of mea-
surable quantities of endrin in milk. In another study (Saha,
1969), the ratio of residue in milk to feed was 0.07.
Studies by Brooks (1969) demonstrated that steers, lambs, and
hogs receiving 0.1 mg/kg endrin in the diet for 12 weeks showed
little tendency to deposit endrin in body tissues. Continuous
feeding of up to 2 mg/kg resulted in a maximum body fat content of 1
Vig/g. Long, et al. (1961) reported high levels of storage (23.7
ug/g) in the adipose tissue of lambs. Higher levels were detected
in the internal fat surrounding the stomach and thoracic cavity
than in external fat deposits. After the lambs were transferred to
untreated pasture, endrin levels in fat decreased somewhat, but
C-10
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levels of approximately 6.4 to 13.8 ug/g were still present 42 days
after termination of exposure. Pigs receiving 510 mg endrin over
30 days had endrin fat levels of no more than 2 ug/g, and no endrin
was detected in any other tissue (Brooks, 1974).
Inhalation
Agricultural workers, home gardeners, and those involved in
the manufacture or distribution of endrin might become exposed
through the inhalation route. Respiratory exposure during periods
of orchard spraying may generally be expected to reach 0.01 mg/hour
(Wolfe, et al. 1963, 1967).
Wolfe, et al. (1963) reported that spraying of potatoes with a
1 percent solution of endrin dust produced levels of 0.41 mg/hour
for respiratory exposure. During the high pressure spraying of row
crops, the respiratory exposure rate was below the limits of detec-
tion of the analytical method employed (Jegier, 1964).
Another possible means of inhalation exposure to endrin is
from the residues on tobacco plants used for smoking materials.
Bowery, et al. (1959) found that tobacco retained an average of 0.2
yg of endrin per commercial cigarette. Forty percent of the resid-
ual endrin disappears during the curing process, but the remainder
persists throughout the cigarette manufacturing process. Endrin
residues in pipe tobacco increased approximately threefold from
1969 (0.05 ug/g) to 1971 (0.114 ug/g). Residues of endrin in
cigars remained at approximately 0.06 ug/g from 1969 to 1972.
Endrin residues in cigarettes decreased from 0.18 ug/g to 0.09 ug/g
from 1969 to 1971 (Bowery, et al. 1959; Domanski and Guthrie,
1974).
C-ll
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In a survey of 45 sites in 1971, the highest level of endrin in
ambient air was 25.6 ng/irr in Greely, Colorado (U.S. EPA, 1971).
In a separate survey of three sites in 1975, the highest reported
level of endrin in the ambient air was 0.5 ng/m3 (U.S. EPA, 1979).
Dermal
The most significant occupational dermal exposure to endrin
occurs during field applications. During dusting or spray-machine
operations, dermal exposure is almost always greater than respira-
tory exposure. Dermal exposure during orchard spraying is likely
to be as high as 3 mg/body/hour, for workers wearing standard pro-
tective clothing in which 3.15 ft of the body is exposed. Poten-
tially the greatest hazard associated with the use of endrin, how-
ever, occurs during measuring and pouring the emulsifiable concen-
trate solution (Wolfe, et al. 1963, 1967).
Wolfe, et al. (1963) studied exposure to endrin during several
field situations. These situations included: spraying orchard
cover crops for mouse control by various methods, dusting potatoes,
spraying row crops, and piloting an airplane during application.
The highest total exposure (dermal 18.7 mg/hr and respiratory 0.41
mg/hr) to endrin occurred during the dusting of potatoes with 1
percent endrin powder. In another study, a dermal exposure of 0.15
mg/hr was noted during the application of endrin to row crops
(Jegier, 1964).
PHARMACOKINETICS
Absorption
Endrin is known to be absorbed by the skin, the lungs, and the
gut (U.S. EPA, 1979) , however , the rates of the absorption have not been
adequately documented.
C-12
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Distribution
Humans do not tend to store endrin in significant quantities.
No residues were detected in plasma, adipose tissue, or urine of
workers occupationally exposed to endrin (Hayes and Curley, 1968).
Measurable levels of endrin have not been detected in human subcu-
taneous fat or blood, even in those areas where it is used exten-
sively, such as India or the lower Mississippi delta area (Brooks,
1974). Despite its high acute toxicity, endrin is a relatively
nonpersistent pesticide in humans. Endrin residues have only been
detected in the body tissues of humans immediately after an acute
exposure. However, little is known concerning the persistence and
toxicity of endrin metabolites.
As a result of acute human poisoning, high levels of endrin
have been observed in both blood and urine but not in cerebral
spinal fluid (Coble, et al. 1967). Endrin-poisoned humans have
been reported to have endrin levels as high as 400 yg/g in fat tis-
sue and 10 vig/g in other tissues (Coble, et al. 1967). However, the
400 ug/g value was obtained using a bioassay technique presently
regarded as unreliable (Curley, et al. 1970).
Much lower values of endrin were obtained from an autopsy of
victims poisoned by eating endrin-contaminated bread (endrin levels
ranged from 48 to 1,807 ppm) in Saudi-Arabia (Table 3). Blood and
urine samples taken from patients 29 to 31 days after the outbreak
were uniformly negative for endrin (Curley, et al. 1970) . Low
blood levels were detected in three humans who recovered after
accidental ingestion of endrin. In one case, the concentration of
endrin in the blood 30 minutes after convulsions occurred was
C-13
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TABLE 3
Endrin Concentrations Found in Victims
of Endrin Poisoning in Saudi Arabia*
Sample Endrin Concentrations (ug/g)
Blood 0.007-0.032
Urine 0.004-0.007
Vomitus 5.24
Tissues (autopsy) from:
Stomach 0.16
Liver 0.685
Kidney 0.116
*Curley, et al. 1970
C-14
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0.053 yg/g and 20 hours after convulsions it was recorded at 0.038
yg/g. This same patient excreted 0.02 yg/g endrin via the urine
during the following 24 hours (Coble, et al. 1967).
Richardson, et al. (1967) fed endrin to 9-raonth-old dogs for
128 consecutive days at a level of 0.1 mg/kg body weight per day.
Blood concentrations during the experiment ranged from 0.001 to
0.008 yg/g. At the termination of the experiment, concentrations
in the adipose tissue ranged from 0.3 to 0.8 yg/g; heart, pancreas,
and muscle were at the lower end of this range, while the concen-
tration in the hepatic tissue was 0.077 to 0.085 ug/g. The kidneys
and lungs had similar concentrations.
The amounts of endrin detected in the tissues of dogs that
were fed diets containing endrin in concentrations of 4 to 8 mg/kg
for approximately six months were as follows: 1 yg/g in the fat; 1
yg/g in the liver; and 0.5 yg/g in the kidneys (Treon, et al. 1955).
Metabolism
Endrin is metabolized and excreted more rapidly than other
chlorinated hydrocarbon insecticides (Jager, 1970). There is good
evidence that endrin is quickly metabolized in mammals (probably in
the liver) and excreted as a hydrophilic metabolite.
IH vitro studies appear to support the hepatic metabolism of
endrin. A metabolite behaving as a mono-hydroxy derivative was
produced when endrin was incubated at 30ฐC for several hours with
both rat liver and pig liver microsomes and NADPH (Brooks, 1969).
Formation of the mono-hydroxy derivative was suppressed by sesamex,
an inhibitor of microsomal oxidations.
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Information regarding the metabolic fate of endrin in vivo is
conflicting. Baldwin, et al. (1970) found that endrin is metabo-
lized in the rat to at least three metabolites. One is 9-ketoen-
drin, which is found in tissues and in urine. The other two metabo-
lites are excreted in the feces and have not been found in body tis-
sues. The acute oral LDcg of 9-ketoendrin in rats (62 mg/kg) is
higher than that of endrin (25 mg/kg), and the reaction appears to
be a detoxication step (Brooks, 1969). Oxidation without skeletal
rearrangement is the major metabolic route in mammals although
details remain to be worked out (Brooks, 1974).
Bedford, et al. (1975) studied oral LD values based on 10-
day mortalities for endrin and three of its mammalian metabolites
(anti-12-hydroxyendrin, syn-12-hydroxyendrin, and 12-ketoendrin)
in rats. All of the metabolites were more toxic than the parent
compound. Rapidity of intoxication, sex differences, and analysis
of the brain tissue indicated that 12-ketoendrin may be the acute
toxicant in each case. Thus, the oxidative metabolism of endrin
may be responsible for its acute toxicity.
Jager (1970) found, in feeding experiments with rats, that
females metabolize endrin more slowly than males. When carbon-14
labeled endrin was fed to male and female rats, the males excreted
60 percent of it in the feces within the first 24 hours and the
females only 39 percent. Less than 1 percent was excreted in the
urine. Of the total radioactivity excreted in the feces, 70 to 75
percent occurred in the form of hydrophilic metabolites; the
remainder was in unchanged endrin. Twenty-four hours after the
last dose, only hydrophilic metabolites were excreted.
C-16
-------
Sex differences in the rate of endrin metabolism in rats were
also found by Hutson, et al. (1975). Although the major metabolite
in both sexes was anti-12-hydroxyendrin, excreted via the bile as
the glucuronide, male rats produced the metabolite at a higher rate
than did females. A minor metabolite was trans-4,5-dihydroisodrin-
4,5-diol. 12-Ketoendrin was the major urinary metabolite in male
rats, whereas the major urinary metabolite in female rats was anti-
12-hydroxyendr in-o-sulf ate. These authors also found the formation
of 12-ketoendrin to be directly related to the acute toxicity of
endrin.
Excretion
At higher dosage levels in experimental animals, excretion of
endrin appears to be slower. Tissue content of endrin declines
fairly rapidly after a single dose or when -a-continuous feeding
experiment is terminated (Brooks, 1969).
The major metabolite in both male and female rats was anti-12-
hydroxyendrin, which was excreted via the bile as the glucuronide
(Hutson, et al. 1975); trans-4,5-dihydroisodrin-4,5-diol is a minor
biliary metabolite. 12-Ketoendrin was observed as the primary uri-
nary metabolite in the male rat; the major urinary metabolite in
female rats was anti-12-hydroxyendrin-o-sulfate. Syn-12-hydroxy-
endrin was not detected.
Cole, et al. (1968) also studied rates of excretion of carbon-
14 labeled endrin in whole rats, bile-fistulated rats, and isolated
perfused rat livers. Over 90 percent of the excreted radioactivity
was found in the feces of the intact animals and in the bile of the
fistulated animals. Fifty percent of the radioactive endrin was
C-17
-------
excreted within the first 24 hours, in the fistulated animals,- 50
percent of the endrin radioactivity was excreted in the bile in
approximately one hour in the perfused experiments (Cole, et al.
1968).
With the exception of endosulfan, endrin is the least persis-
tent of any of the chlorinated hydrocarbon pesticides in mammals.
It is rapidly metabolized and eliminated from the tissues of verte-
brates. Excretion occurs through the milk as well as through the
urine and the feces (Brooks, 1974). Endrin metabolites, one of
which is known to be several times more toxic than endrin itself,
may persist for longer periods of time.
EFFECTS
Acute, Subacute, and Chronic Toxicity
Endrin is classified as "very highly hazardous"; meaning that,
any contact with very small amounts of the substance may result in
severe systemic toxicity or death (Thompson, 1971). Endrin is the
most acutely toxic of the cyclodiene insecticides and yet, except
for endosulfan, is least persistent in mammals (Brooks, 1974).
Endrin toxicity can be elicited from any route of exposure. When
ingested in one dose by rats, endrin is about three times as toxic
as aldrin and about 15 times as toxic as DDT (Treon, et al. 1955).
Upon intravenous administration to mice, endrin was five times as
toxic as dieldrin (Walsh and Fink, 1972).
The onset of endrin toxicity symptoms is rapid. The return to
normal among those who survive is also rapid. The recovery from
endrin intoxication is faster than from other cyclodiene pesticides
(Brooks, 1974) .
C-18
-------
Symptoms of acute endrin poisoning in mammals clearly indicate
that endrin is a neurotoxicant. The first indication of acute
endrin poisoning is usually central nervous system excitation as
evidenced by hypersensitivity to external stimuli associated with
generalized tremors and followed by severe tonic-clonic convulsions
(Brooks, 1974). These convulsions may occur as early as 30 minutes
after acute endrin exposure (Brooks, 1974). Convulsions can culmi-
nate in death from respiratory failure (Brooks, 1974). In the
range of the acute oral I^Q (17 to 43 mg/kg) , death of rats may
result after 48 hours (Boyd and Stefec, 1969).
Other symptoms of acute endrin poisoning include bradycardia
(slowed heartbeat); increase in blood pressure, salivation, and
body temperature; leukocytosis (increase in number of white blood
cells); increased hemoconcentration; decreased blood pH; increased
cerebrospinal fluid pressure and cerebral venous pressure; in-
creased renal vascular resistance with decreased renal blood flow
and glomerular filtration rate; decrease in catecholamine concen-
tration of the adrenals; and increased levels of circulating epi-
nephrine and norepinephrine (Emerson, et al. 1964; Reins, et al.
1966). Histopathologic examinations of rat tissue at autopsy re-
veal signs of a stress reaction, degenerative changes in kidneys,
liver and brain capillaries, and venous congestion, and loss of
weight and dehydration of some organs (Boyd and Stefec, 1969) .
The symptoms in man include headache, dizziness, abdominal
disturbances, nausea, vomiting, mental confusion, muscle twitch-
ing, and epileptiform convulsions which may occur suddenly and
without prior warning (Brooks, 1974; Coble, 1967).
C-19
-------
Mammalian susceptibility to endrin toxicity varies greatly
with age, sex, and species as shown in Table 4. The LDcri values
range from 1.37 to 43 mg/kg. Apparently, mice and monkeys are most
sensitive, and guinea pigs are more resistant. Rabbits seem to be
somewhat more resistant than monkeys to a single dose of endrin.
The acute toxicity of endrin is, however, high for all these
species.
In rats and guinea pigs, females are more susceptible than
males. The greater susceptibility of female rats six months of age
than that of younger female rats is the reverse of the more normal
relationship between age and susceptibility found in males.
When endrin was maintained in contact, as a dry 100-mesh pow-
der, with either intact or abraded skin of female rabbits for 24
hours, the minimum lethal dosage was found to be greater than 60
and less than 94 mg/kg. Poisoned animals had convulsions, but
there was not evidence of gross or microscopic damage to the skin.
Degeneration of the cells in the central zones of the lobules of
the livers in the rabbits was observed (Treon, et al. 1955).
Graves and Bradley (1965) determined an LD of 5.6 mg/kg for
endrin injected into the peritoneal cavity of Swiss albino mice.
An intravenous LDcQ of 2.3 mg/kg was determined by Walsh and Fink
(1972) for adult male mice. Endrin injected into dogs intrave-
nously at a dosage of 3 mg/kg resulted in death in approximately 75
percent of the animals (Hinshaw, et al. 1966) .
Target organs found in acute experiments are not always the
same as those following repeated exposure over long periods of
C-20
-------
TABLE 4
Acute Oral Toxicity of Endrin to Mammals
Animal LD50
(age, sex) (mg/kg)
Mouse 1.37a
Rats (6 months, M) 43b
Rats (6 months, F) 7b
Rats (30 days, M) 30b
Rats (30 days, F) 17b
Rat 3a
Rabbits (F) 7-10b
Hamster 10a
Guinea pigs (F) 16b
Guinea pigs (M) 36b
Monkey 3b
fNIOSH, 1977
bTreon, et al. 1955
C-21
-------
time. The central nervous system is the target of acute endrin
poisoning. When an animal is repeatedly exposed to low doses (0.8
to 3.5 mg/kg/day) of endrin, it can often make compensatory adjust-
ments to cope with the initial nervous system injury until damage
to liver or other organs intervenes. However, Chernoff, et al.
(1979) found that the threshold level for convulsions in hamsters
was 10 mg endrin/kg body weight. This convulsive dose was approxi-
mately twice that required for the production of teratogenic
effects.
Revzin (1968) found that chronic administration of endrin can
lead to convulsions. He administered endrin to squirrel monkeys at
a minimum rate of 0.2 mg/kg/day, which caused a characteristic
change in the electroencephalogram (EEC) after seven days. With
continued daily dosing electrographic seizures developed. Endrin
administration was stopped after seizures, but after one month
EEC's and behavior were still abnormal.
The chronic toxicity of endrin is greater than that of other
organochlorine pesticides. In prolonged feeding experiments, rats
can consume diets containing approximately three times as much
aldrin and 12 times as much DDT as endrin without increase in rela-
tive weights of specific organs. On the basis of organ weights
dogs are at least 10 times as susceptible to the toxic effects of
endrin as to those of DDT (Treon, et al. 1955) . Species and sex
differences exist in susceptibility to chronic endrin toxicity.
Females are generally more susceptible than males. Rabbits and
dogs are more susceptible than rats (Treon, et al. 1955).
C-22
-------
Mammalian species appear to be sensitive to the toxic effects
of endrin at low levels in their diet. Significant mortality dur-
ing a 7-month period appeared in deer mice when fed 2 mg/kg endrin
in the diet (Morris, 1968). The deer mice exhibited symptoms of
hypertension, uncoordination, muscle tremors, and convulsions
which increased in intensity until death occurred. A 48-hour star-
vation period at the end of the feeding study increased mortality
of young mice and suggests possible translocation of endrin from
fatty tissues.
Endrin fed throughout the life to Osborne-Mendel rats at 12
mg/kg in the diet decreased viability. Mean survival time fell
from 19.7 months to 17.6 months for males and from 19.5 months to
18.2 months for females. The endrin-fed rats experienced moderate
increases in incidence of congestion and focal hemorrhages of the
lung; slight enlargement, congestion and mottling of the liver;
slight enlargement, discoloration or congestion of the kidneys
(Deichmann, et al. 1970).
The paper published by Treon, et al. (1955) is perhaps the
most extensive long-term toxicological study of endrin and will be
reviewed in detail. This paper includes acute, subacute (3 to 6
months), and chronic (2 year) feeding studies in rats (male and
female) ; a subacute oral study in rabbits (8 to 10 weeks) ; and a 19-
month oral study in dogs (male and female). Body weights, organ
weights, and histopathologic data are included.
The rabbit studies were limited to a single dose level. Four
of five female rabbits given 1 mg/kg/day of endrin in peanut oil
died in 8 to 10 weeks. The surviving animals sacrificed after
C-23
-------
50 doses over 10 weeks showed "diffuse degenerative and fatty vacu-
olization of the hepatic and renal cells" and degeneration of the
heart. Thus, to the rabbit a dose of 1 mg/kg/day of endrin is ex-
tremely toxic.
Initial subacute rat studies gave the following results. All
rats survived 50 doses (in peanut oil) over 10 weeks at the 1 mg/
kg/day level. At 2 mg/kg/day 1 of 2 young female rats and 1 of 3
adult female rats died during the 10 week study. All six of the
male rats (young and old) survived the 10 week dosing at 2 mg/kg/
day. Three of three male rats also survived a similar study at a
dose level of 5 mg/kg/day. Thus, the female rat is apparently more
sensitive to the effects of endrin than is the male. In addition,
this study indicates that the rat is more resistant to multiple
doses than is the rabbit.
In a 2-year rat feeding study (Treon, et al. 1955) , animals
were given 100, 50, 25, 5, 1, and 0 ppm of endrin in the diet.
Groups of 20 male and 20 female rats (Carworth strain) were fed at
each dosage level (total rats = 240). The mortality among these
groups of rats is shown in Table 5. Since these dosage levels are
given in ppm, it is necessary to calculate approximate daily intake
on a mg/kg basis. If one estimates that a 200 g rat eats 20 g
food/day then 100 ppm (100 ug/g) in the food translates to 10
mg/kg/day intake of endrin; 50 ppm to 5.0 mg/kg/day; and 25, 5, and
1 ppm to 2.5, 0.5, and 0.1 mg/kg/day, respectively. Endrin in the
diet of female rats at 100, 50, or 25 ppm caused significant mor-
tality at 80 weeks (Table 5) . The male rats were somewhat less
susceptible showing increased mortality only at the 100 and 50 ppm
C-24
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level. Dietary levels of 100 or 50 ppm resulted in the early deaths
of all but a few resistant rats. Body weight gains were not partic-
ularly altered by these dosages of endrin, nor was the rate of live
weight to body weight changed. In the male rats fed 25 ppm (2.5
mg/kg/day) or 5 ppm (0.5 mg/kg/day) the average liver weight to
body weight ratios were significantly different (p. 0.05-0.01) from
comparable controls. This was not true at the 1 ppm (0.1 mg/kg/
day) dietary level, nor was there any effect in female rats at the
0.5 or 0.1 mg/kg/day level. Hypersensitivity to external stimuli
and occasional convulsions were noted in rats at the 5 and 10
mg/kg/day level. Convulsions were not noted in the animals fed 2.5
mg/kg/day or less. Animals that died when fed at the three higher
dosage levels (10, 5, and 2.5 mg/kg/day) exhibited "diffuse degen-
eration of brain, liver, kidneys, and adrenal glands." Survivors
at the two highest dosage levels showed degenerative changes in the
liver only. A single statement notes that the incidence of neo-
plasia was not greater among experimental rats than among the con-
trols.
Treon, et al. (1955) also conducted an extensive dog study
summarized in Table 6 which is taken directly from the published
paper. This table provides the dosage in both ppm in the diet and
daily intake as mg/kg body weight. All dogs died when fed 0.5 to
4.0 mg/kg/day (10 to 50 ppm) in the diet and more than half of those
fed 0.20 to approximately 0.5 mg/kg/day (5 to 8 ppm) also died. All
dogs survived when their diets contained 4 ppm (0.15 to 0.21
mg/kg/day) or less for periods up to 18.7 months. All dogs fed
10 ppm (0.49 to 0.81 mg/kg/day) suffered extensive weight loss;
C-26
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those fed at 8 ppm (0.29 to 0.62 mg/kg/day) gained weight intially,
but eventually failed to continue growing. Those fed 4 ppm (0.15
to 0.21 mg/kg/day) did not grow normally, but those at 3 (0.12 to
0.25 mg/kg/day) or 1 ppm (0.045 to 0.12 mg/kg/day) grew as well as
control dogs. Affected dogs became emaciated, developed respira-
tory distress, and signs of irritation of the central nervous sys-
tem (hypersensitivity to stimulation, tremors, twitching, and
severe convulsions). Dogs fed at the 4, 3, or 1 ppm level exhib-
ited no such toxic manifestations. Dogs fatally poisoned were
found to have "diffuse degenerative lesions in the brain, heart,
liver, and kidneys, together with pulmonary hyperemia and edema."
Renal damage was severe and characterized by diffuse degeneration
and necrosis of the convoluted tubules. The liver exhibited dif-
fuse degeneration, fatty vacuolization, and, in some instances,
necrosis.
Dogs fed diets containing 8 ppm endrin (0.29 to 0.62 mg/kg/
day) for six months had enlargement of the liver, kidney, and
brain. At 3 ppm (0.12 to 0.25 mg/kg/day) the kidney and heart were
significantly enlarged at sacrifice (18.7 months). Dogs fed at the
1 ppm level (0.045 to 0.120 mg/kg/day) for 18.7 months were compar-
able to controls by all parameters of comparison.
In summary, this paper (Treon, et al. 1955) demonstrates that
dogs are apparently more susceptible to endrin than rats. Minimal
effects (organ enlargement) were seen at the 3 ppm (0.12 to 0.25
mg/kg/day) level in dogs after 18.7 months. At higher dosage
levels, effects were more severe with mortality beginning with the
5 ppm (0.20 to 0.27 mg/kg/day) group and no dogs surviving doses
C-28
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greater than 10 ppm (0.49 to 0.81 mg/kg/day) . The dog study
included a total of 25 dogs, both male and female, with dosages
ranging from 1 ppm (0.045 to 0.12 mg/kg/day) to 50 ppm (2.5 to 4.0
mg/kg/day) and demonstrated a no-effect level at the lowest dose of
1 ppm (0.045 to 0.12 mg/kg/day).
Although two monkeys were used in the Treon, et al. (1955)
study, no data is included in their report other than the minimum
lethal dosage of 1 to 3 mg/kg single oral dose for one male and one
female monkey (unspecified). Thus, on an acute basis, the monkey
appears more susceptible than the rodents.
Synergism and/or Antagonism
The acute oral toxicity (LE>5Q) of equitoxic doses of combina-
tions of 15 pesticides was examined by Keplinger and Deichmann
(1967). The results are presented in Table 7._.Endrin plus diazi-
non, endrin plus toxaphene, and endrin plus malathion showed addi-
tive effects; while endrin plus parathion, endrin plus DDT and,
particularly, endrin plus delnav showed lower than expected LDcgS,
suggestive of antagonistic effects. Joint administration of endrin
and its closely related compound aldrin showed a more than additive
effect, and endrin plus chlordane was found to exert a potentiating
effect.
No other information is available on synergistic and/or antag-
onistic effects of endrin.
Teratogenicity
Rats and mice were given 0.58 mg endr in/kg body weight four
times weekly for a month, and then after a week or more without
endrin treatment, the animals were allowed to become pregnant
C-29
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TABLE 7
Expected and Observed Oral LD^s of Endrin
plus other Pesticides in Mice*
Other
Pesticides
Chlordane
Aldrin
Dieldrin
Diazinon
Malathion
Toxaphene
Parathion
DDT
Delnav
Expected
LD50
(mg/kg)
473
63
63
93
703
63
12
213
87
Observed
LD50
(mg/kg)
211
34
50
93
820
77
18
400
195
Ratio
E/0
2.22
1.83
1.25
1.00
0.85
0.81
0.65
0.53
0.44
*Source: Keplinger and Deichmann, 1967
C-30
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(Nodu, et al. 1972) . A reduced fetal survival rate was found in
both species. Nine mouse fetuses with club foot were found in the
treated group of 177, while only one fetus with club foot was in the
control group of 303.
Endrin exerted embryocidal and teratogenic effects on pregnant
hamsters. Both soft and skeletal tissue malformations were pro-
duced. Single oral doses of endrin (5 mg/kg) administered to preg-
nant Syrian golden hamsters on day 7, 8, or 9 of gestation caused a
high incidence of fetal death, congenital abnormalities and growth
retardation. Thirty-two percent of the implantations resulted in
fetal mortalities. Teratogenic effects were observed in 28 percent
of the fetuses from hamsters treated on day eight. Open eye
occurred in 22 percent, webbed foot in 16 percent, cleft palate in
5 percent, cleft lip in 1 percent, and fused ribs in 8 percent
(Ottolenghi, et al. 1974).
Ottolenghi, et al. (1974) also found endrin to be teratogenic
in mice, but frequency and gravity of the defects produced were
less pronounced than in the hamsters when a single dose (2.5 mg/kg
in mice and 5 mg/kg in hamsters) of half the LD^g was administered.
Abnormalities in the mice included open eye and cleft palate. No
significant effects were found with respect to fetal survival or
fetal weight.
Golden hamsters, intubated with endrin (0.75 and 1.5 mg/kg) on
days 5 to 14 of gestation, had less reactive locomotor activity
than controls during gestation but not at weaning (Gray, et al.
1979). The offspring of these dams were tested in open field at 15,
z<3, ZT, 34, and 44 days of age. Fif teen-day-old pups at the
C-31
-------
1.5 mg/kg dose were approximately 90 percent more active than con-
trols but this difference disappeared by day 34. Prenatal endrin
exposure appeared to have behavioral effects in hamsters and their
offspring.
Chernoff, et al. (1979) found that a single dose of endrin
administered to pregnant hamsters on day eight, produced meningoen-
cephaloceles at doses above 1.5 mg/kg and fused ribs at doses above
5.0 mg/kg. Open eyes, cleft palates, and webbed feet were not
noted. It was suggested that a teratogenic level of endrin in
humans could be lower than the levels estimated to cause human con-
vulsions since the convulsive dose in hamsters was approximately
twice that required for the induction of terata.
Mutagenicity
Endrin, as well as aldrin and dieldrin,_can cause chromosome
damage (Grant, 1973) . Evidence of cellular degeneration has been
observed in germinal tissue of male albino rats treated with 0.25
mg endrin per testes administered intratesticularly (Dikshith and
Datta, 1972). The most conspicuous effects were hypertrophy, chro-
mosomal aberrations, including stickiness, bizarre configurations,
formation of chromosome fragments, and abnormal restitution of
chromosomes. Formation of single and double bridges with acentric
fragments was very common, disturbing the normal disjunction of
chromosomes and eventually affecting the chromosome complements of
the division products (Dikshith and Datta, 1973). Unequal distri-
bution of chromosomes at anaphase was also observed. Severe cell
C-32
-------
damage resulted in liquefication and transformation of the chroma-
tin mass into an amorphous lump (Dikshith and Datta, 1972). These
were the only instances reported of mutagenicity related to endrin.
However, chlorinated cyclopentadienes, such as endrin, may
undergo metabolic conversion forming acylating and, possibly, muta-
genic tetrachlorocyclopentadienone although no data exists to sup-
port this hypothesis. Using mouse liver microsomes for metabolic
activation and E. coli K12(343/113) to detect mutagenicity, tetra-
chlorocyclopentadiene and pentachlorocyclopentadiene were highly
mutagenic after metabolic activation, whereas hexachlorocylcopenta-
diene was not (Goggelman, et al. 1978).
Carcinogenicity
No malignancies attributed to endrin exposure have been
reported. In 2-year feeding studies in rats at dosage levels of
100, 50, 25, 5, I, and 0 ppm Treon, et al. (1955) reported that the
incidence of neoplasia was no greater among treated animals than
among controls. The high dosage level (100 ppm) approximates 10
mg/kg/day. Sndrin fed to weanling Osborne-Mendel rats for a life-
time at dietary levels of 2, 6, or 12 mg/kg was neither tumorigenic
nor carcinogenic (Deichmann, et al. 1970; Deichmann and MacDonald,
1971; Deichmann, 1972).
A recently completed National Cancer Institute (NCI) bioassay
for possible endrin carcinogenicity concluded that endrin was not
carcinogenic for Osborne-Mendel rats or for B6C3F1 mice (NCI,
1979) .
C-33
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CRITERION FORMULATION
Existing Guidelines and Standards
In 1965, maximum permissable levels were assigned to each of
the organochlorine compounds based on the "maximum acceptable con-
centrations" suggested on July 9, 1965, by the subcommittee on
Toxicology to the Public Health Service Advisory Committee on
Drinking Water Standards (Schafer, et al. 1969). This concentra-
tion for endrin was 0.001 ppm. In 1967, the "maximum reasonable
stream allowance" for endrin of 0.1 ppb (0.1 ug/1) was suggested by
Ettinger and Mount (1967) and was accepted as a guideline.
A maximum acceptable level of 0.002 mg/kg body weight/day was
established by a Joint Food and Agriculture Organization/World
Health Organization (FAO/WHO) Meeting on Pesticide Residues in Food
held in Rome, November, 1972 (FAO, 1973).
A threshold limit value of 100 ug/m was set for atmospheric
levels of endrin by the American Conference of Governmental Indus-
trial Hygienists (ACGIH) for 1971 (Yobs, et al. 1972). A threshold
limit value of 100 ug/m for an 8-hour time-weighted average occu-
pational exposure has also been established by the Occupational
Safety and Health Administration (OSHA) (29 CFR 1910.1000).
Toxic pollutant effluent standards (40 CFR 129.102) were pro-
mulgated by the U.S. EPA. These allowed an effluent concentration
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 work-
ing day's effluent. In addition, discharge is not to exceed 0.0006
kg per 1,000 kg of production.
C-34
-------
Current Levels of Exposure
While no recent data are available on levels of exposure of
humans to endrin it appears that the risk of exposure is decreasing
because of the decreased usage of the pesticide.
In a survey of over 500 drinking water samples, the number of
samples containing concentrations of endrin in excess of 0.1 ug/1,
which has been established as a maximum reasonable stream allow-
ance, decreased from 23 in the period 1964 to 1965 to 0 in the
period 1966 to 1967 (Schafer, et al. 1969). The most recent study-
found only 4 ng/1 in contaminated drinking water (U.S. EPA, 1974).
In a series of analyses of total diets, the average daily in-
take of endrin remained at trace levels (<: 0.001 mg) during the
period 1965 to 1970, but the frequency of occurrence decreased con-
siderably (Duggan and Lipscomb, 1969; Duggan and Corneliussen,
1972) .
Exposure of the general public to endrin in the air decreased
from a maximum level of 25.6 yg/m in 1971 at Greeley, Colorado, to
a maximum of 0.5 ug/m in 1975 in Jackson, Mississippi (U.S. EPA,
1979).
Special Groups at Risk
Agricultural workers, home gardeners, and those involved in
endrin manufacture and distribution are the most likely to be
exposed to endrin. They may be exposed through inhalation or der-
mal exposure. The most significant occupational exposure comes
during spraying of fields, and dermal exposure is almost always
greater than respiratory exposure. Probably the greatest hazard
associated with the use of endrin occurs when measuring and pouring
the emulsifiable concentrate material. Because endrin has been
C-35
-------
shown to cause teratogenic effects, pregnant women, particularly
those whose diets may contain large amounts of fish, must also be
considered a special group at risk. Evidence that endrin may cause
chromosomal damage in germinal tissue suggests that men and women
of child-bearing intent may also be a special risk group.
Endrin concentrations are highest in the atmosphere over agri-
cultural areas and probably reach their peak levels during the
pesticide use season. Of all urban communities, those surrounded
by farm lands run the highest-risk of atmospheric contamination.
Endrin adsorbed to particulates could not be detected in the air
over representative communities but, may be present at very low
concentrations in the vapor phase. Urban communities far removed
from agricultural areas are unlikely to experience significant con-
tamination. The homes of occupationally exposed workers have
higher levels of atmospheric contamination than do those of the
general public.
Basis and Derivation of Criterion
Carcinogenicity studies with endrin have all been negative.
The limited teratogenic and mutagenic studies on endrin suggest
that effects are induced with high endrin doses. However, an unus-
ual administration route was used in the positive mutagenic
studies. More toxicological data must be gathered about these
potential effects of endrin before a final conclusion can be
reached.
On the basis of long-term dietary studies in mammals, a rea-
listic drinking water criterion may be proposed. Maximum no-
C-36
-------
observed-effect and gradual dosage dietary levels of endrin re-
ported for experimental animals are shown in Table 8.
The data in Table 8 suggest that there is considerable species
difference in the response to endrin. A 1.0 mg/kg single dose pro-
duced 4/5 deaths in rabbits, yet this amount is reported as a no-
observed-effect level in the rat and mouse by other investigators.
Obviously, the results of various studies are sensitive only to the
extent to which the investigators pursue the study. In the Treon,
et al. (1955) study large numbers of animals were used, both male
and female, a range of dosages was fed and the animals followed by
observation, body weight, organ weights, and histopathologic exami-
nation of tissues at sacrifice.
The rat study by Treon, et al. (1955) suggests a no-observed-
effect level (NOEL) in a 2-year feeding study between 0.1 and 0.5
nig/kg/day. Dogs in an 18.7 month study were somewhat more sensi-
tive with the NOEL at approximately 0.1 mg/kg/day. Monkeys may be
more sensitive than the rat, but chronic studies in monkeys have
not been reported. However, using two monkeys Treon, et al. (1955)
found that single doses of 1 to 3 mg/kg were fatal.
Thus, long-term studies in both the rat and the dog suggest
that the NOEL is approximately 0.1 mg/kg/day. Extrapolation from
these two animal studies to man appears to be reasonable. Since
data on chronic human ingestion are not available, but valid long-
term feeding studies in more than one animal species have been
reported, an uncertainty factor of 100 is appropriate in the
absence of any indication of carcinogenicity for calculating a
water criterion. Human exposure to endrin was calculated on the
C-37
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basis of daily ingestion of 2 i of water and 6.5 g fish with a BCF
of 3,970 for endrin. Using a no-effect dosage level if 0.1
mg/kg/day the total acceptable daily intake (ADI) for a 70 kg per-
son is:
0.1 mg endr in/kg x 70 kg = 10Q (uncertainty" factor) - 70
The criterion for endrin is thus:
* - 2 1 + (ooOgkgx 3,970) - 2'51 "9/1 2'5
This approximates closely the 1 yg/1 maximum allowable concentra-
tion for endrin proposed by the Public Health Service for drinking
water. It is therefore, recommended that the endrin criterion be
established at 1 yg endrin/1 of ambient water (1 ppb) .
This calculation assumes that 100 percent of man's exposure is
assigned to the ambient water pathway. Although it is desirable to
establish a criterion based upon total exposure potential, the data
for other exposure conditions have not been factored into this
analysis.
In summary, based upon the use of toxicologic data for dogs
and rats, and an uncertainty factor of 100, the initial level for
endrin corresponding to daily intake of 70 yg/day, is 2.5 yg/1.
Since the existing 1 yg/1 allowable concentration in the drinking
water standards is reasonably close to 2.5 yg/1, it is recommended
that 1.0 yg/1 be used as the criterion with notation that there are
C-39
-------
special groups at risk.* Drinking water contributes 5 percent of
the assumed exposure while eating contaminated fish products
accounts for 95 percent.
*If endrin was present in waters from which edible fish were
located and if these fish concentrate endrin by a factor of 3,970,
this criterion may not be sufficient to protect a special high
risk group i.e., pregnant women who consume a single dose of en-
drin contaminated fish. Given the BCF, fish in water at the maxi-
mum recommended concentration of 1 ug/1, may contain 3.8 ug/g en-
drin. A 250 g portion of fish would contain approximately 1.0 mg
endrin (or 0.02 mg/kg for a 50 kg female). This dose provides a
margin of safety of only 75 over the NOEL of 1.5 mg/kg for terato-
genicity in the hamster (Chernoff, et al. 1979) . The adequacy of
this margin of safety is highly questionable, especially given the
likelihood of consumption of more than 250 g of fish at a given
time. The recommended water quality criterion of 1 yg/1 was based
on a chronic exposure study, teratologic outcomes are more likely
to occur with acute exposures at critical times in gestation.
C-40
-------
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C-44
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C-49
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o U. S GOVERNMENT PRINTING OFFICE : 1ซ80 7K).016/4386
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