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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 

EPA Ambient 

Water Quality 

Criteria for 

Endrin 
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 

i 
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. 

ii 
FOREWORD 

Section 304 (a)(1) 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 satisfaction of paragraph 11 of the Settlement Agreement 

in Natural Resources Defense Council, et. al. 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 effects. 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 

local 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 

iii 
ACKNOWLEDGEMENT 

Aquatic Life Toxicology: 

Donald I. Mount, ERL-Duluth David J. Hansen, ERL-Gulf Breeze 

U.S. Environmental Protection Agency U.S. Environmental Protection Agency 

Mammalian Toxicology and Human Health Effects: 

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 

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 

Farrell R. Robinson 

Purdue University 

Sorrell Schwartz 

Georgetown University 

Jerry F. Stara, ECAO-Cin 

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 
TABLE OF CONTENTS 

Page 

Criteria Summary 

Introduction A-1 

Aquatic Life Toxicology 

Introduction 

Effects 

Acute Toxicity 

Chronic Toxicity 

Plant Effects 

Residues 

Miscellaneous 

Summary 

Criteria 

References 

Mammalian Toxicology and Human Health Effects 

Introduction 

Exposure 

Ingestion from Water 

Ingestion from Food 

Inhalation 

Dermal 

Pharmacokinetics 

Absorption 

Distribution 

Metabolism 

Excretion 

Effects 

Acute, Subacute and Chronic Toxicity 

Synergism and/or Antagonism 

Teratogenicity 

Mutagenicity 

Carcinogenicity 

Criterion Formulation 

Existing Standards and Guidelines 

Current Levels of Exposure 

Special Groups at Risk 

Basis and Derivation of Criteria 

References 

B-1 

B-1 

B-2 

B-2 

B-5 

B-8 

B-8 

B-11 

B-11 

B-12 

B-36 

C-1 

C-1 C-4 

C-4 

C-5 

C-11 

C-12 

C-12 

C-12 

C-13 

C-15 

C-17 

C-18 

C-18 

C-29 

C-29 

C-32 

C-33 

C-34 

C-34 

C-35 

C-35 

C-36 

C-41 

V 
CRITERIA DOCUMENT 

ENDRIN 

CRITERION 

Aquatic Life 

For endrin the criterion to protect freshwater aquatic life as 

derived using the Guidelines is 0.0023 μg/l as a 24-hour average, 

and the concentration should not exceed 0.18 μg/l at any time. 

For endrin the criterion to protect saltwater aquatic life as 

derived using the Guidelines is 0.0023 μg/l as a 24-hour average, 

and the concentration should not exceed 0.037 μg/l 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 μg/l. 

Analysis of the toxic effects data resulted in a calculated level 

which is protective of human health against the ingestion of contaminated 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 
INTRODUCTION 

Endrin is the common name of one member of the cyclodiene 

group of pesticides. It is a cyclic hydrocarbon having a chlorinesubstituted 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 20°C; 

the vapor pressure is 2.7 x 10 -7 at 25°C; and the substance begins 

to decompose at 200°C. 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 ingredient, available in a variety of diluted formulations (Brooks, 

1974). Jarvinen and Tyo (1978) found the solubility of endrin to 

be about 200 μg/l. 

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 discontinuation 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 production has continued to decline. In 1978, endrin production was 

approximately 400,000 lbs (U.S. EPA, 1978). 

A-1 
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 
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 organisms, 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 freshwater to be about 200 μg/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 Triton® X-100 in the acetone-endrin solution to improve dispersion in the test water. Because concentrations were not measured, 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 Criteria for the Protection of Aquatic Life and Its Uses in order to better understand 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 toxicity as described in the Guidelines. 

B-1 
of freshwater aquatic organisms exposed to high water concentrations of endrin as a result of its use on cotton. Nearly all of their work used static, unmeasured test procedures, and the data, at best, can be used only for 

relative toxicity purposes. Clearly, they did demonstrate a marked increased 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 predators 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. Although 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 LC50 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 unexpected since endrin was an effective insecticide. 

B-2 
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 unmeasured 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 required 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 between 0.15 and 2.1 μg/l, suggesting a relatively narrow range of species 

sensitivity for fishes. The Freshwater Final Acute Value for endrin, derived from the species mean acute values listed in Table 3 using the procedure described in the Guidelines, is 0.18 μg/l. 

Acute toxicity tests with saltwater invertebrate species also demonstrate 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 

μg/l for pink shrimp to 790 μg/l 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 LC50 values from 0.037 to 0.3 μg/l (Schimmel, et al. 1975; 

Lowe, unpublished; Butler, 1963). LC50 values for five other arthropod 

species ranged from 0.3 μg/l for Korean shrimp (Schoettger, 1970) to 25 μg/l 

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 
values ranging from 14.2 to 790 μg/l, based on decreased shell deposition or 

abnormal development of larvae (Table 1). 

The toxicity of endrin to oysters may vary depending on water temperature. Concentrations decreasing growth by 50 percent were: 400 μg/l at 

12°C; 33 μg/l at 24°C; and 14 μg/l at 22°C (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 mollusks 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 μg/l for chinook 

salmon (Schoettger, 1970) to 3.1 μg/l 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 LC50 values for saltwater invertebrate species from acute toxicity tests on endrin support the hypothesis that the acute toxicity of endrin 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 measured concentrations (Table 1). Additionally, LC50 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 
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 pg/l. 

Chronic Toxicity 

Life-cycle chronic tests have been completed with fathead minnow and 

flagfish, giving chronic values of 0.19 and 0.26 pg/l, respectively (Table 

2). Mount (1962), working with the bluntnose minnow, a species closely related to the fathead minnow, found a no-observed-effect concentration between 0.1 and 0.4 ug/l for a 291day exposure (Table 6). Spawning did not 

occur in this test, but the results are consistent with those for the fathead 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 uncontaminated 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. Unfortunately, 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). Sur 

parental generation was reduced by exposure to 0.11 Pg/l (Ty 

1979). Onset and duration of spawning were significantly 

lengthened for female grass shrimp at all exposure concentrat i 

viva1 of the 

erSchroeder, 

delayed and 

ons (0.03 to 

0.79 lJg/l). The number of females depositing embryos was less than that of 

controls, but embryo production and hatching success apparently were not affected. Larval mortality increased, time to metamorphosis increased, and 

growth of juvenile shrimp was decreased by endrin concentrations of 0.11 

ug/l and higher. A chronic value of 0.039 pg/l endrin was obtained for 

I3 -5 
grass shrimp (Table 2), even though all tested concentrations significantly 

impaired some life-cycle function. A lower limit of 0.03 pg/l 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 upper limit of 0.05 "g/l was set based on decrease in number of ovigerous females and delay in spawning of 3 to 4 weeks. 

Sheepshead minnows (Schimrnel, et al 1975; Hansen, et al. 1977), spot 

(Lowe, 1966), and mumnichog (Eisler, 1970a) have been exposed to endrin for 

10 days or longer (Tables 2 and 6). Of these tests with saltwater fish species, 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/l hatched early; all fry exposed to 0.72 

u!3/1, and about one-half of those exposed to 0.31 ug/l, died. Females died 

during spawning, fewer embryos were fertile, and survival of exposed progeny 

decreased in 0.31 ug/l. No significant effects were observed on survival, 

growth, or reproduction at an exposure concentration of 0.12 ug/l. The 

chronic limits, 0.12 to 0.31 ug/l, were not much less than the 96-hour 

LC50 of 0.34 ug/l, indicating that there is little difference between endrin 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 between 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 LCso values and nominal observed no-effect concentrations were reported (Table 6); however, results were similar to those reported in the life cycle test (Table 2). The 

LCso value based on measured concentrations for fry on the 33rd day of the 

B-6 
experiment was 0.16 ug/l. Although mortality in fish exposed to a nominal 

concentration of 0.31 pg/l was not significant, they were visibly affected 

by endrin. 

Fifty-seven percent of the juvenile spot exposed to 0.075 pg/l of endrin 

died within the first 19 days of an eightmonth test; those exposed to 0.05 

ug/l were apparently not affected (Lowe, 1966) (Table 6). Spot exposed to 

0.05 ,,g/l for eight months exhibited no signs of poisoning, and their survival, length, and weight did not differ from those of control fish. The 

nominal, no-observed-effect concentration of 0.05 vg/l was 0.11 of the 

Lc50 of 0.45 ug/l, and this also tends to support the conclusion of a minimal 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 lOday LC50 of 0.33 ug/l for the mumnichog based on nominal 

concentrations (Eisler, 1970a); this is little different from the 96-hour 

LCso values of 0.6 and 1.5 pg/I (Eisler, 1970b). 

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, respectively. The species mean acute values and acute-chronic ratios are summarized in Table 3. 

Dividing the Freshwater Final Acute Value of 0.18 *g/l by the geometric 

mean (4.0) of the four acute-chronic ratios (Table 2) gives the Freshwater 

Final Chronic Value of 0.045 pg/l (Table 3). Dividing the Saltwater Final 

Acute Value of 0.037 ug/l by the geometric mean of acute-chronic ratios 

(4.0) gives the Saltwater Final Chronic Value of 0.0093 pg/l (Table 3). 

B-7 
Data on the toxicity of endrin to five freshwater spec ies of a 

listed in Table 4. Apparently, algae are not sensitive to endrin, 

lgae are 

and the 

Plant Effects 

lowest effect level for plants is 475 pg/l 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 species, effects of endrin on these plant species are unlikely at concentrations 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/l. Productivity of natural phytoplankton communities was reduced by 46 percent in 1,000 ug/l (Butler, 1963). 

Growth rate of Agmenellum quadruplicatum was reduced in as little as 0.2 

Ug/l (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 DDT. 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- 
centration factors are available from studies with American oysters (Mason 

and Rowe, 1976), grass shrimp (Tyler-Schroeder, 1979), sheepshead minnows 

(Hansen, et al. 19775 Schimnel, et al. 1975), and spot (Lowe, 1966). Additional endrin BCF data (Table 6) are available from 96-hour exposures of oysters, grass shrimp, pink shrimp, sheepshead minnows, and sailfin mollies 

(Wilson, 1966; Schimmel, et al. 1975). 

Bioconcentration factors (Table 5) for endrin in American oysters exposed for seven days ranged from 1,670 to 2,780 (Mason and Rowe, 1976). Endrin accumulated rapidly, reaching steady-state after about 48 hours of exposure. Oysters placed in endrin-free water depurated endrin at a rate of 

0.005 pg/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 (Schimnel, 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 experiment, steady-state was reached after 2.5 days of a 21day exposure. 

Ninety percent of the endrin was depurated within 4.2 days. In the second 

experiment, the averaqe BCF of endrin was 1,600 in parental generation 

shrimp from a partial-life-cycle exposure lasting five months. Average bioconcentration factors after a 96-hour exposure were 830 for grass shrimp and 

980 for pink shrimp (Schimnel, et al. 1975). 

Bioconcentration data for two of three species of saltwater fishes differ 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 averB-9 
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 (Schimnel, 

et al. 1975). Sailfin mollies exposed to endrin for four days had an average 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 ug/l. 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 pg/l is calculated 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 llg/l. 

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 

u!j/l l 

B-10 
Other available residue data for effect levels are not appropriate for 

calculation of freshwater or saltwater residue values for wildlife protection. Therefore, the lowest residue value of 0.0023 ug/l 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 concentration 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. 

Summarv 

Acute data are available for 28 freshwater species including a wide variety of organisms normally performing a spectrum of community functions. 

Only one of the 28 species has an acute value above 100 ug/l, the lowest 

species mean acute value is 0.15 vg/l, and most values are clustered near 

1.0 vg/l. The data are predominantly from static tests in which toxicant 

concentrations were not measured and so probably underestimate true toxicity. The Freshwater Final Acute Value is 0.18 ug/l. 

There are acute data for 21 species of saltwater organisms. None of the 

values is above 14.2 pg/l and four are below 0.1 pg/l. The Saltwater Final 

Acute Value is 0.037 pg/l, one-fifth that of the freshwater. 

Life cycle tests with two freshwater fish species gave chronic endpoints 

near 0.2 ug/l and acute-chronic ratios of 3.3 and 2.2. Chronic data for the 

sheepshead minnow gave comparable estimates of 0.19 pg/l and 1.9 for the 

chronic value and acute-chronic ratio, respectively. The saltwater grass 

shrimp was more sensitive (0.039 ug/l) and gave a much larger acute-chronic 

ratio of 18. The geometric mean of these four estimates of the acuteB-11 
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 ,,g/l and the Saltwater Final Chronic Value is 0.0093 ug/l. 

The residue data for freshwater and saltwater are similar and show relatively low bioconcentration factors as compared to related insecticides such 

as dieldrin. Further, the data agree that endrin uptake reaches steadystate quickly and is depurated quickly. Using the FDA action level of 0.3 

mg/kg for fish oil, the geometric mean of normalized bioconcentration factors (1,324), and a percent lipid value of 100 for fish oil, a Final Residue 

Value of 0.0023 vg/l is calculated for both freshwater and saltwater. The 

Final Residue Value may be too high because, on the average, the concentration 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 ug/l. Other data do not 

reveal any more sensitive effects. Saltwater algae appear more sensitive 

than freshwater, but all values are above 1 vg/l except one. Therefore, 

plant protection seems certain if animals are protected. Other data available 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/l as a 24Slour average, and the concentration should not exceed 0.18 pg/l at any time. 

For endrin the criterion to protect saltwater aquatic life as derived 

using the Guidelines is 0.0023 ug/l as a 24-hour average, and the concentration should not exceed 0.037 ug/l at any time. 

B-12 
m * 

0 

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d 


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Table 1. Khnt hued) 

Spacles 

Shiner perch, 

Cymatogaster sregata 

Dwarf perch, 

Hlcranetrus mlnlmus 

Owarf perch, 

Hlcfctnetrus mlnimus 

a I behead, 

Thalassoma blfasclatum 

Strlped mu1 let, 

Mugll cepha I us 

Northern puffer, 

Sphaeroldes maculatus 

ueihod* 

f-f, u 

s, u 

n, u 

s, u 

s, u 

s, u 

Lc50fic50 

@I/l 1 

0.12 

0.6 

0. I3 

0. I 

0.3 

3. I 

speccles mean 

Acute Value 

(lea/t) 

0.31 

0.28 

0. I 

0.3 

3.1 

Reference 

Earnest 6 t)envl I le, 

1972 

Earnest & Benvll le, 

1972 

Earnest 6 Ebnvllle, 

1972 

Els ler, l970b 

Elsler, 1970b 

Elsler, l970b 

l S = static, FT = flow-through, U = unmeasured, H = measured 

**Abnormal development of oyster larvae; decreased gouth of oyster; or loss of equi librlum 

of brown shrlap or blue crabs. 

B-20 
Table 2. Chronic values for endrln 

Specl es 

Fathead minnow, 

Pimephales promelas 

Flagfish, 

Jordanella florldae 

Sheepshead mlnnou, 

Cyprinodon varleqatus 

Grass shrlrap, 

Palaemonetes puglo 

Test* 

Llmlts 

(rg/l) 

Chronic Value 

(pg/l) Ret arence 

FRESHWATER SPECIES 

LC 0.14-0.25 

LC 0.22-0.3 0.26 

0.19 

SALTWATER SPECIES 

LC 0.12-0.31 0.19 

LC 0.03-O. 05” 0.039 

Jarvlnen & Tyo, 1978 

Hermanutz, 1978 

Hansen, et al. I977 

Tyler-Schroeder, 1979 

u LC = life cycle or parflal life cycle 

**Onset of spawning was delayed about one week In shrimp exposed to 0.03 rig/l.. 

Because a delay of one reek would probably not af feet natural populations, 

llmlts were set on decreases In number of ovigerous females and delayed spawnlng 

of 3-4 weeks in 0.05 ug/l of endr In. 

Acute-Chronic Ratios 

Spec I es 

Fathead mlnnow, 

Plmephales promelas 

Acute Chron I c 

Value Value 

01g/ I) tug/ I) Ratlo 

0.42’** 0.19 2.2 

Flagflsh, 0.85 0.26 3.3 

Jordanella florldae 

-- 

Sheepshead minnow, 0.36 0. I9 1.9 

Cyprlnodon varfegatus 

B-21 
L * 
Table 3. Species mean acute values and acute-chronic ratlos for endrln 

Rank* 

28 

27 

26 

25 

24 

23 

22 

21 

20 

19 

I0 

17 

16 

Species 

Species Mean Spades Maan 

Acute Value Acute-Chron I c 

(&I 1 Ratio 

FRESHWATER SPECIES 

Cladoceran, 

Daphnla magna 

&Yf IY, 

Hexagenla bl I lneata 

352 

64 

Copepod fcyclopold), 

(unidenilfied) 

60 

Cladoceran, 

Simocephalus serrulatus 

34 

Crayf Ish, 

Orconectes na I s 

32 

Scud, 

Gamarus lacustrls 

5.9 

kYf IY, 

Ephemerel la grandls 

4.7 

Scud, 

Gamnarus fasclatus 

3. I 

Goldflsh, 

Carasslus auratus 

2. I 

Damself ly, 

lschnura verticalus 

1.8 

Sowbug, 

Asellus brevlcaudus 

1.5 

Glass shrimp, 

Falaemonetes kadlakensls 

1.3 

GUPPY # 

Poecl lia retlculata 

1.2 

ch lnook sa Item, 

Oncorhynchus tshawytscha 

I. I 

B-23 
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REFERENCES 

Argyle, R.L., et al. 1973. Endrin uptake and release by fingerling channel 

(Ictalurus punctatus). Jour. Fish. Res. Board Can. 30: 1743. 

Batterton, J.C., et al. 1971. Growth response of bluegreen algae to aldrin, dieldrin, endrin, and their metabolites. Bull. Environ. Contam. Toxicol. 6: 589. 

Bennett, H.J. and J.W. Day, Jr. 1970. Absorption of endrin by the bluegill 

sunfish, Lepomis macrochirus. Pestic. Monitor. Jour. 3: 201. 

Brungs, W.A. and G.W. Bailey. 1966. Influence of suspended solids on the 

acute toxicity of endrin to fathead minnows. Proc. Zlst Purdue Ind. Waste 

Conf., Part 1. 50: 4. 

Butler, P.A. 1963. Commercial fisheries investigations, pesticide-wildlife 

studies: A review of Fish and Wildlife Service investigations during 1961 

and 1962. U.S. Dep. Inter. Fish Wildl. Circ. 167: 11. 

Cutkomp, L.K., et al. 1971. ATPase activity in fish tissue homogenates and 

inhibitory effects of DOT and related compounds. Chem. Biol. Int. 3: 439. 

Davis, H.C. and H. Hidu. 1969. Effects of pesticides on embryonic development of clams and oysters and on survival and growth of the larvae. U.S. 

Dep. Inter. Fish Wildl. Fish. Bull. 67: 393. 

B-36 
Davis, P.W., et al. 1972. Organochlorine insecticide, herbicide, and polychlorinated biphenyl (PCB) inhibition of NaK4TPase in rainbow trout. Bull. 

Environ. Contam. Toxicol. 8: 69. 

Earnest, R.D. and P.E. Benville, Jr. 1972. Acute toxicities of four organochlorine insecticides to 2 species of surf perch. Calif. Fish and Game. 

58: 127. 

Eisler, R. 1969. Acute toxicities of insecticides to marine decapod crustaceans. Crustaceana. 16: 302. 

Eisler, R. 1970a. Factors affecting pesticide-induced toxicity in an estuarine fish. U.S. Dep. Inter. Bur. Sport Fish. Wildl. Tech. Paper 45. 

Eisler, R. 1970b. Acute toxicities of organochlorine and organophosphorous 

insecticides to estuarine fishes. U.S. Dep. Inter. Bur. Sport Fish. 

Wildl. Tech. Paper 46. 

Eisler, R. and P.H. Edmunds. 1966. Effects of endrin on blood and tissue 

chemistry of a marine fish. Trans. Am. Fish. Sot. 95: 153. 

Fabacher, D.L. 1976. Toxicity of endrin and an endrinmethyl parathion 

formulation to largemouth bass fingerlings. Bull. Environ. Contam. Toxicol. 16: 376. 

Ferguson, D.E. and C-R. Bingham. 1966. Endrin resistance in the yellow 

bullhead, Ictalurus natalis. Trans. Am. Fish. Sot. 95: 325. 

B.37 
Ferguson, D.E., et al. 1965. Tolerance of five chlorinated hydrocarbon insecticides in 2 species of fish from a transect of the lower Mississippi 

River. Jour. Miss. Acad. Sci. 11: 239. 

Ferguson, D.E., et al. 1966. Dynamics of endrin uptake and release by resistant and susceptible strains of mosquitofish. Trans. Am. Fish. Sot. 

95: 335. 

Gaufin, A.R., et al. 1965. The toxicity of ten organic insecticides to 

various aquatic invertebrates. Water Sewage Works. 12: 276. 

Grant, B.F. and P.M. Mehrle. 1970. Chronic endrin poisoning in goldfish. 

Carassius auratus. Jour. Fish. Res. Board Can. 27: 2225. 

Grant, B.F. and P.M. Mehrle. 1973. Endrin toxicosis in rainbow trout (Salmo qairdneri). Jour. Fish. Res. Board Can. 30: 31, 

Hansen, D.J. 1969. Avoidance of pesticides by untrained sheepshead minnows. Trans. Am. Fish. Sot. 98: 426. 

Hansen, D.J. 1980. Memorandum to C. Stephan. U.S. EPA. July. 

Hansen, D.J., et al. 1973. Avoidance of pesticides by grass shrimp (Palaemonetes pugio). Bull. Environ. Contam. Toxicol. 9: 129. 

Hansen, D.J., et al. 1977. Endrin: Effects on the entire life-cycle of 

saltwater fish, Cyprinodon variegatus. Jour. Toxicol. Environ. Health. 

3: 721. 

B-38 
Henderson, 

bon insect 

C ., et al. 1959. Relative toxicity of ten 

Hermanutz, 

chlorinated hydrocaricides to four species of fish. Trans. Am. Fish. Sot. 88: 23. 

R. 1978. Endrin and malathion toxicity to flagfish (Jordanella 

floridae). Arch. Environ. Contam. Toxicol. 7: 159. 

Iyatomi, K.T., et al. 1958. Toxicity of endrin to fish. Prog. Fish. 

Cult. 20: 155. 

Jackson, G.A. 1976. Biologic half-life of endrin in channel catfish tissues. Bull. Environ. Contam. Toxicol. 16: 505'. 

Jarvinen, A.W. 1980. Memorandum to R.C. Russo. June 13. 

Jarvinen, A.W. and R.M. Tyo. 1978. Toxicity to fathead minnows of endrin 

in food and water. Arch. Environ. Contam. Toxicol. 7: 409. 

Jensen, L.D. and A.R. Gaufin. 1966. Acute and long-term effects of organic 

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. Sot. 90: 264. 

Katz, M. and G.G. Chadwick. 1961. Toxicity of endrin to some Pacific 

Northwest fishes. Trans. Am. Fish. Sot. 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 versus dynamic bioassay conditions. Trans. Am. Fish. Sot. 99: 13. 

Lowe, J.I. Results of toxicity tests with fishes and macroinvertebrates. 

Data sheets available from U.S. Environmental Protection Agency, Environmental 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 Comrn. p. 271. 

Ludke, J.L., et al. 1968. Some endrin relationships in resistant and susceptible populations of golden shiners, Notemigonus crysoleucas. Trans. Am. 

Fish. Sot. 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 endrin in the eastern oyster. Arch. Environ. Contam. Toxicol. 4: 349. 

Menzel, D.W., et al. 1970. Marine phytoplankton vary in their response to 

chlorinated hydrocarbons. Science. 167: 1724. 
Mount, D.I. 1962. Chronic effects of endrin on bluntnose minnows and guppies. U.S. Dep. Inter. Fish Wildl. Serv. Res. Rep. 58. p. 38. 

Mount, D.I. and G.J. Putnicki. 1966. Summary report of the 1963 Mississippi 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 toxicity to the amphipod crustacean, Gamrnarus lacustris. Utah Acad. Proc. 

41: 64. 

Post, G. and T.R. Schroeder. 1971. Toxicity of four insecticides to four 

salmonid species. Bull. Environ. Contam. Toxicol. 6: 144. 

Sanders, H.O. 1969. Toxicity of pesticides to the crustacean Gammarus 

lacustris. U.S. Dep. Inter. Bur. Sport Fish. Wildl. Tech. Paper 25. 

Sanders, H.O. 1972. Toxicity of some insecticides to four species of malacostracan crustaceans. U.S. Dep. Inter. Bur. Sport Fish. and Wildl. Tech. 

Paper 66. 

Sanders, H.O. and O.B. Cope. 1966. Toxicities of several pesticides to two 

species of cladocerans. Trans. Am. Fish. Soc. 95: 165. 

B-41 
Sanders, H.O. and 0.6. Cope. 1968. The relative toxicities of several 

pesticides to naiads of three species of stoneflies. Limnol. Oceanogr. 

13: 112. 

Schimmel, S.C., et al, 1975. Endrin: Effects on several estuarine organisms. Proc. 28th Annu. Conf. S.E. Assoc. Game and Fish Comm., 1974. p. 187. 

Schoettger, R.A. 1970. Fish-pesticide research laboratory, progress in 

sport fishery research. U.S. Dep. Int. Bur. Sport Fish Wildl. Res. Publ. 

106. 

Solon, J.M., et al. 1969. The effect of sublethal concentration of LAS on 

the acute toxicity of various insecticides to the fathead minnow (Pimephales 

promelas Rafinesque). Water Res. 3: 797. 

Tyler-Schroeder, D.B. 1979. Use of Grass Shrimp, Palaemonetes pugio, in a 

Life-Cycle Toxicity Test. In: L.L. Marking and R.A. Kimerle (eds.), Proceedings of a Symposium on Aquatic Toxicology and Hazard Evaluation. Am. 

Sot. Testing and Materials (ASTM), October 31 - November 1, 1977. p. 159. 

U.S. Food and Drug Administration. 1977. Administrative Guideline 7426.04, 

Attachment F. 

U.S. Food and Drug Administration. 1978. Administrative Guideline 7420.08, 

Attachment E. 

B-42 
Vance, B.D. and W. Drummond. 1969. Biological concentration of pesticides 

by algae. Jour. Am. Water Works Assoc. 61: 360. 

Wilson, A.J. 1966. Chem. Assays. Annu. Rep. Bur. Corrmercial Fisheries 

Biol. Lab., Gulf Breeze, Fla. U.S. Bur. Commercial Fish. Circ. 247: 6. 

Bd3 
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 values μg/l 

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 hydrophilic 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 concentrations 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 0.55 to 2.75 kg/ha resulted in the death of 33 to 100 percent of 

C-1 
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 animals manifest primarily as behavioral and reproductive disorders, 

i.e., improper maternal care, temporary loss of normal activity, 

increased vulnerability to predators, reduced reproductive potential, 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 succumbed 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 injection. 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 inhalation, and through dermal contact. The average dietary intake in 

the United States in 1973 was 0.033 μg/day (0.0005 μg/kg/day) for a 

69.1 kg man. This is far below the maximum daily intake of 138.2 

μg/day (2 μg/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 
Outbreaks of human poisoning have resulted from accidental 

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 drugmetabolizing 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 industrialized 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 contaminated vegetation. Ingestion of 20 mg endrin per day by cows 

resulted in levels of up to 0.25 μg/g of endrin in milk. Aquatic 

invertebrates and fish bioconcentrate considerable quantities of 

endrin from water and pass it on to predatory birds. This contaminated fowl (or the fish themselves) may, in turn, be ingested by 

humans. 

C-3 
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. 

Ingestion from Water 

EXPOSURE 

Occasionally, groundwater may contain more than 0.1 μg/l of 

endrin, but levels as high as 3 μg/l have been correlated with precipitation 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 3 ng/l (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 percent), contained detectable concentrations of endrin. However, the 

number of finished water samples containing concentrations of 

endrin in excess of 0.1 μg/l decreased from 23 (10 percent) to 0 in 

a three year period from 1964 to 1967 (Schafer, et al. 1969). 

C-4 
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/l. 

Ingestion from Food 

The general population has little exposure to endrin in the 

diet. In a series of analyses of total diets determined from "market basket" samples in five regions of the United States, the total 

average intake from food ranged from approximately 0.009 μq/kg body 

weight per day in 1965 to 0.0005 μg/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 μg/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 significantly 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) l Storage longer than 12 weeks 

decreased endrin residues in Irish and sweet potatoes by 20 percent 

(Solar, et al. 1971) l Heat processing and freezing further lowered 

potato residues 65 and 52 percent, respectively. Studies on 

C-5 
TABLE 1 

Average Incidence and Daily Intake of Endrin* 

Percent Positive Daily Intake w/kg 

Year Composites (mg) body wt/day 

1965 2.8 Ta 0.000009 

1966 2.0 T 0.000004 

1967 1.7 T 0.000004 

1968 1.1 0.001 0*00001 

1969 3.3 T 0.000004 

1970 1.4 T 0.0000005 

*Source: Duggan and Corneliussen, 1972 

aT = Trace, < 0.001 mq 

C-6 
I 

-- - 40 ’ I 0 . . I cu 3 

ll!za,Il - - i i - I c 

I-- 

:, 
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 temperature, 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 potential human exposure. 

A bioconcentration factor (ECF) relates the concentration of a 

chemical in aquatic animals to the concentration in the water in 

which they live. The steady-state l3CFs for a lipid-soluble compound in the tissues of various aquatic animals seem to be proportional 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 percent lipids of consumed fish and shellfish, and a steady-state RCF 

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 consumption 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 
lipids for consumed freshwater and estuarine fish and shellfish is 

3.0 percent. 

Two laboratory studies, in which percent lipids and a steadystate 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 Americans is calculated to be 3,970. 

Because of the dynamic state of endrin in the biological tissues of lower animals (Mount, et al. 1966), the bioaccumulation is 

short-lived, and tissue burdens diminish rapidly once the environmental 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 

w/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 pg/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 presumably retained in the milk fat for up to four weeks. However, the 

quantities of endrin ingested during that period were not controlled. Williams and Mills (1964) studied the excretion of endrin in 

cows ' milk under controlled feeding conditions. Endrin concentrations 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 (>O.OOl pg/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 mixture 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 measurable 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/kq resulted in a maximum body fat content of 1 

Isq/q l Long, et al. (1961) reported hiqh levels of storage (23.7 

pg/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 
levels of approximately 6.4 to 13.8 pg/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 pg/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 detection 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 

pg of endrin per commercial cigarette. Forty percent of the residual endrin disappears during the curing process, but the remainder 

persists throughout the cigarette manufacturing process. Endr in 

residues in pipe tobacco increased approximately threefold from 

1969 (0.05 pg/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 pg/g 

from 1969 to 1971 (Bowery, et al. 1959; Domanski and Guthrie, 

1974). 

c-11 
In a survey of 45 sites in 1971, the highest level of endrin in 

ambient air was 25.6 ng/m3 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 respiratory exposure. Dermal exposure during orchard spraying is likely 

to be as high as 3 mg/body/hour, for workers wearing standard protective clothing in which 3.15 ft2 of the body is exposed. Potentially the greatest hazard associated with the use of endrin, however, occurs during measuring and pouring the emulsifiable concentrate 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 ratesof the absorption have notbeen 

adequately documented. 

c-12 
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). 

LYeasurable levels of endrin have not been detected in human subcutaneous fat or blood, even in those areas where it is used extensively, 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 pg/g in fat tissue and 10 pg/g in other tissues (Coble, et al. 1967). However, the 

400 pg/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 
TABLE 3 

Endrin Concentrations Found in Victims 

of Endrin Poisoning in Saudi Arabia* 

Sample Endrinconcentrations (pg/g) 

Blood 

Urine 

vomitus 

Tissues (autopsy) from: 

stomach 

Liver 

Kidney 

0.007-0.032 

0.004-0.007 

5.24 

0.16 

0.685 

0.116 

*Curley, et al. 1970 

c-14 
0.053 ug/g and 20 hours after convulsions it was recorded at 0.038 

lJg/g ' This same patient excreted 0.02 pg/g endrin via the urine 

during the following 24 hours (Coble, et al. 1967). 

Richardson, et al. (1967) fed endrin to g-month-old 

128 consecutive days at a level of 0.1 mg/kg body weight 

Blood concentrations during the experiment ranged from 

dogs for 

per day. 

0.001 to 

0.008 pg/g. At the termination of the experiment, concentrations 

in the adipose tissue ranged from 0.3 to 0.8 vg/g; heart, pancreas, 

and muscle were at the lower end of this range, while the concentration 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 ug/g in the fat; 1 

Dg/g in the liver; and 0.5 pg/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. 

In 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 30°C 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. 

c-15 
Information regarding the metabolic fate of endrin in vivo is -- 

conflicting. Baldwin, et al. (1970) found that endrin is metabolized in the rat to at least three metabolites. One is 9-ketoendrin, which is found in tissues and in urine. The other two metabolites are excreted in the feces and have not been found in body tissues. The acute oral LD50 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 LD50 values based on loday 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-dihydroisodrin4,5-diol. 12-Ketoendrin was the major urinary metabolite in male 

rats, whereas the major urinary metabolite in female rats was anti12-hydroxyendrin-o-sulfate. 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-z-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-dial is a minor 

biliary metabolite. 12-Ketoendrin was observed as the primary urinary metabolite in the male rat; the major urinary metabolite in 

female rats was anti-12-hydroxyendrin-o-sulfate. Syn-12-hydroxyendrin was not detected. 

Cole, et al. (1968) also studied rates of excretion of carbon14 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 persistent of any of the chlorinated hydrocarbon pesticides in mammals. 

It is rapidly metabolized and eliminated from the tissues of vertebrates. 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 culminate in death from respiratory failure (Brooks, 1974). In the 

range of the acute oral LDsO (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; increased renal vascular resistance with decreased renal blood flow 

and glomerular filtration rate; decrease in catecholamine concentration of the adrenals; and increased levels of circulating epinephrine and norepinephrine (Emerson, et al. 1964; Reins, et al. 

1966). Histopathologic examinations of rat tissue at autopsy reveal 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 twitching, and epileptiform convulsions which may occur suddenly and 

without prior warning (Brooks, 1974; Coble, 1967). 

c-19 
*Nammalian susceptibility to endrin toxicity varies greatly 

with age, sex, and species as shown in Table 4. The LD50 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 loo-mesh powder, 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 LDSO of 5.6 mg/kg for 

endrin injected into the peritoneal cavity of Swiss albino mice. 

An intravenous LDsO of 2.3 mg/kg was determined by Walsh and Fink 

(1972) for adult male mice. Endrin injected into dogs intravenously 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 

(age, sex) 

Mouse 1.37a 

Rats (6 months, M) 43b 

Rats (6 months, F) 7b 

Rats (30 days, M) 3ob 

Rats (30 days, F) 17b 

Rat 3a 

Rabbits (F) 7-lob 

Hamster 1oa 

Guinea pigs (F) 16b 

Guinea pigs (M) 36b 

Monkey 3b 

aNIOSH, 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 endtin, it can often make compensatory adjustments to cope with the initial nervous system injury until damage 

to liver or other organs intervenes. HOWeVeK, Chernoff, et al. 

(1979) found that the threshold level for convulsions in hamsters 

was 10 mg endrin/kg body weight. This convulsive dose was approximately twice that KeqUiKed fOK 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 (EEG) after seven days. With 

continued daily dosing electrographic seizures developed. Endrin 

administration was stopped after seizures, but after one month 

EEG'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 relative 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 (Peon, 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 during a 7-month period appeared in deer mice when fed 2 mg/kg endrin 

in the diet (MOKKiS, 1968). The deer mice exhibited symptoms of 

hypertension, uncoordination, muscle tremors, and convulsions 

which increased in intensity until death occurred. A 48-hour starvation 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 OK congestion of the kidneys 

(Deichmann, et al. 1970). 

The paper published by Treon, et al. (1955) is perhaps the 

most extensive long-term tOXiCOlOgiCa 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 vacuolization 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 extremely 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 (TKeOn, et al. 1955), animals 

were given 100, 50, 25, 5, 1, and 0 ppm of endrin in the diet. 

GKOU~S 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 mortality at 80 weeks (Table 5). The male rats were somewhat less 

susceptible showing increased mortality only at the 100 and 50 ppm 

C-24 

level. Dietary levels of 100 OK 50 ppm resulted in the early deaths 

of all but a few resistant rats. Body weight gains were not particularly 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) OK 5 ppm (0.5 mg/W/dW 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 OK 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 OK less. Animals that died when fed at the three higher 

dosage levels (10, 5, and 2.5 mg/kg/day) exhibited "diffuse degeneration of brain, liver, kidneys, and adrenal glands." SUKViVOKS 

at the two highest dosage levels showed degenerative changes in the 

liver only. A single statement notes that the incidence of neoplasia was not greater among experimental rats than among the contKOlS. 

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 respiratory distress, and signs of irritation of the central nervous system (hypersensitivity to stimulation, tremors, twitching, and 

severe convulsions). Dogs fed at the 4, 3, OK 1 ppm level exhibited 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 diffuse degeneration, fatty vacuolization, and, in some instances, 

necrosis. 

Dogs fed diets containing 8 ppm endrin (0.29 to 0.62 mg/kg/ 

day) fOK 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 comparable to controls by all parameters of comparison. 

In summary, this paper (TKeOn, 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 
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 (LD 50) of equitoxic doses of combinations of 15 pesticides was examined by Keplinger and Deichmann 

(1967). The results are presented in Table 7,.Endrin plus diasinon, endrin plus toxaphene, and endrin plus malathion showed additive effects: while endrin plus parathion, endrin plus DDT and, 

particularly, endrin plus delnav showed lower than expected LDsOs, 

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 antagonistic effects of endrin. 

Teratogenicitv 

Rats and mice were given 0.58 mg endrin/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 
TABLE 7 

Expected and Observed Oral LDSOs of Endrin 

plus other Pesticides in Mice* 

Other 

Pesticides 

Expected Observed 

LD50 LD50 

(w/kg 1 (mg/kg ) 

Ratio 

E/O 

Chlordane 

Aldrin 

Dieldrin 

Diazinon 

Malathion 

Toxaphene 

Parathion 

DDT 

Delnav 

473 211 2.22 

63 34 1.83 

63 50 1.25 

93 93 1.00 

703 820 0.85 

63 77 0.81 

12 18 0.65 

213 400 0.53 

87 195 0.44 

*source: Keplinger and Deichmann, 1967 

c-30 
(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 produced. Single oral doses of endrin (5 mg/kg) administered to pregnant Syrian golden hamsters on day 7, 8, OK 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 

OCCUKKed 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 LDSO was administered. 

Abnormalities in the mice included open eye and cleft palate. NO 

significant effects were found with respect to fetal survival OK 

fetal weight. 

Golden hamsters, intubated with endrin (0.75 and 1.5 mg/kg) on 

days 5 to 14 of gestation, had less reactive 1OCOmOtOK 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, 

ZG, z7, 34, and 44 days of age. Fifteen-day-old pups at the 

c-31 
1.5 mg/kg dose were approximately 90 percent more active than controls 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 meningoencephaloceles at doses above i-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 convulsions 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, chromosomal 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 distribution of chromosomes at anaphase was also observed. Severe cell 

C-32 
damage resulted in liquefication and transformation of the chromatin mass into an amorphous lump (Dikshith and Datta, 1972). These 

were the oniy instances reported of mutagenicity related to endrin. 

However, chlorinated cyclopentadienes, such as endrin, may 

undergo metabolic conversion forming acylating and, possibly, mutagenic tetrachlorocyclopentadienone although no data exists to support this hypothesis. Using mouse liver microsomes for metabolic 

activation and E. coli K12(343/113) to detect mutagenicity, tetrachlorocyclopentadiene and pentachlorocyclopentadiene were highly 

mutagenic after metabolic activation, whereas hexachlorocylcopentadiene was not (Goggelman, et al. 1978). 

Carcinogenicitv 

No malignancies attributed to endrin exposure have been 

reported. In 2-year feeding studies in rats at dosage levels of 

100, 50, 25, 5, 1, and 0 ppm Treont 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. Endrin fed to weanling Osborne-Mendel rats for a lifetime 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 B6C3Fl mice (NCI, 

1979). 

c-33 
CRITERION FORMULATION 

Existing Guidelines and Standards 

In 1965, maximum permissable levels were assigned to each of 

the organochlorine compounds based on the "maximum acceptable concentrations" 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 concentration for endrin was 0.001 ppm. In 1967, the "maximum reasonable 

stream allowance" for endrin of 0.1 ppb (0.1 ug/l) 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 pg/m' was set for atmospheric 

levels of endrin by the American Conference of Governmental Industrial Hygienists (ACGIH) for 1971 (Yobs, et al. 1972). A threshold 

limit value of 100 ug/m3 for an 8-hour time-weighted average occupational 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 promulgated by the U.S. EPA. These allowed an effluent concentration 

of 1.5 pg/l per average working day calculated over a period of one 

month, not to exceed 7.5 bg/l in any sample representing one working 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 pg/l, 

which has been established as a maximum reasonable stream allowance, 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 rig/l in contaminated drinking water (U.S. EPA, 1974). 

In a series of analyses of total diets, the average daily intake of endrin remained at trace levels (<O.OOl mg) during the 

period 1965 to 1970, but the frequency of occurrence decreased considerably (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 ug/m3 in 1971 at Greeley, Colorado, to 

a maximum of 0.5 pg/m3 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 derma1 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 agricultural 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 contamination. 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 unusual 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 realistic drinking water criterion may be proposed. Max imum noC-36 
observed-effect and gradual dosage dietary levels of endrin reported 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 produced 4/5 deaths in rabbits, yet this amount is reported as a noobserved-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 examination of tissues at sacrifice. 

The rat study by Treon, et al. (1955) suggests a no-observedeffect level (NOEL) in a 2-year feeding study between 0.1 and 0.5 

w/kg/day. Dogs in an 18.7 month study were somewhat more sensitive 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 longterm 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 

basis of daily ingestion of 2 1 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 person is: 

0.1 mg endrin/kg x 70 kg = 1oO 7 mg/day 

(uncertainty factor) = 7o ug/day 

The criterion for endrin is thus: 

70 pg/day 

' = 2 1 + (0.0063 kg x 3,910) = *05’ Pg/l -2.5 pg/l 

This approximates closely the 1 pg/l maximum allowable concentration for endrin proposed by the Public Health Service for drinking 

water. It is therefore, recommended that the endrin criterion be 

established at 1 pg endrin/l 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 pg/day, is 2.5 pg/l. 

Since the existing 1 vg/l allowable concentration in the drinking 

water standards is reasonably close to 2.5 pg/l, it is recommended 

that 1.0 pg/l 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 endrin contaminated fish. Given the BCF, fish in water at the maximum recommended concentration of 1 ug/l, may contain 3.8 pg/g endrin. A 250 g portion of fish would contain approximately 1.0 mg 

endrin (or 0.02 mq/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 teratogenicity 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 pg/l 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-41 
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c-44 
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C-46 
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C-47 
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C-48 
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c-49 
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e “. S GOVERNMENT PIUNTMC OFFICE : i980 IM-016/1386 

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