# Click here for DISCLAIMER Document starts on next pageUnited States Environmental Protection Agency Office ofWater Regulations and Standards Criteria and Standards Division Washington DC20460 EPA 440/5-80-047 October 1980 

# EPA Ambient

Water Quality Criteria for EndrinAMBIENT WATER QUALITYCRITERIAFOR ENDRIN

PreparedBy U.S.ENVIRONMENTAL

PROTECTION

AGENCY Office ofWaterRegulations andStandards Criteria andStandardsDivision Washington,D.C. Office of ResearchandDevelopment Environmental Criteria andAssessmentOffice

Cincinnati, Ohio CarcinogenAssessmentGroup Washington,D.C. Environmental ResearchLaboratories

Corvalis, Oregon Duluth, Minnesota GulfBreeze, Florida Narragansett, RhodeIsland

iDISCLAIMER

This report hasbeenreviewedbythe Environmental Criteria and AssessmentOffice, U.S.Environmental Protection Agency,andapproved for publication. Mentionoftradenamesorcommercialproductsdoesnot constitute endorsementorrecommendationfor use.

AVAILABILITYNOTICE This documentis available to the public through the National Technical Information Service, (NTIS), Springfield, Virginia 22161.

iiFOREWORD

Section 304(a)(1) ofthe CleanWaterActof 1977(P.L. 95-217), requires theAdministrator of the Environmental Protection Agencyto publish criteria for water quality accurately reflecting the latest scientific knowledgeonthekindandextent ofall identifiable effects onhealth andwelfare which maybeexpected from the presence of pollutants inanybodyofwater, including groundwater. Proposedwater quality criteria for the65toxic pollutants listed undersection 307 (a)(l) of the CleanWaterActweredevelopedandanotice of their availability waspublished for public commentonMarch15,1979(44FR 15926),July25,1979(44FR43660),andOctober1,1979(44FR56628). Thisdocumentis arevision of those proposedcriteria basedupona consideration of commentsreceived fromother FederalAgencies, State agencies, special interest groups, andindividual scientists. The criteria contained inthis documentreplaceanypreviously published EPA criteria for the 65pollutants. This criterion document is also published insatisfaction ofparagraph11oftheSettlement Agreement in Natural ResourcesDefenseCouncil, et. al. vs. Train, 8ERC2120 (D.D.C.1976),modified, 12ERC1833(D.D.C. 1979). Theterm "water quality criteria" is usedintwosections ofthe CleanWaterAct,section 304(a)(l) andsection 303(c)(2). Thetermhas adifferent programimpactineachsection. Insection 304,theterm represents anon-regulatory, scientific assessmentof ecological effects. Thecriteria presented inthis publication aresuchscientific assessments. Suchwater quality criteria associated with specific streamuseswhenadoptedasStatewaterquality standardsundersection 303becomeenforceable maximumacceptable levels of apollutant in ambientwaters. Thewaterquality criteria adoptedintheStatewater quality standards couldhavethesamenumerical limits asthecriteria developedundersection 304. However,inmanysituations Statesmaywant toadjust waterquality criteria developedundersection 304toreflect local environmental conditions andhumanexposure patterns before incorporation into water quality standards. It is not until their adoption aspart oftheStatewaterquality standardsthat thecriteria becomeregulatory. Guidelines to assist the States in the modification of criteria presented in this document, in the development of water quality standards, andinotherwater-related programsofthis Agency,arebeing developedbyEPA.
STEVEN SCHATZOW DeputyAssistant Administrator Office ofWaterRegulations andStandards

iiiACKNOWLEDGEMENT
Aquatic Life Toxicology: Donald I. Mount, ERL-Duluth DavidJ. Hansen,ERL-Gulf Breeze U.S. Environmental Protection Agency U.S. Environmental Protection Agency MammalianToxicology andHumanHealth Effects: FredOehme(author) KansasState University CarynWoodhouse(doc. mgr.) ECAO-Cin U.S. Environmental Protection Agency Bonnie Smith (doc. mgr.) ECAO-Cin U.S. Environmental Protection Agency JosephArcos Tulane University Robert M.Bruce, ECAO-RTP U.S. Environmental Protection Agency EdwardCalabrese University of Massachusetts Jacqueline V.Carr U.S. Environmental Protection Agency William B.Deichmann University of Miami Patrick Durkin Syracuse Research Corporation PamelaFord RockyMountain 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 JeanneManson 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.ivTABLEOFCONTENTS Page

Criteria Summary Introduction A-1 Aquatic Life Toxicology Introduction Effects AcuteToxicity ChronicToxicity Plant Effects Residues Miscellaneous Summary Criteria References MammalianToxicology andHumanHealth Effects Introduction Exposure Ingestion fromWater Ingestion fromFood Inhalation Dermal Pharmacokinetics Absorption Distribution Metabolism Excretion Effects Acute, SubacuteandChronicToxicity Synergismand/orAntagonism Teratogenicity Mutagenicity Carcinogenicity Criterion Formulation Existing StandardsandGuidelines Current Levelsof Exposure Special GroupsatRisk BasisandDerivation ofCriteria 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

VCRITERIA 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 a24-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 a24-hour average, and the concentration should not exceed 0.037 μg/l at any time. HumanHealth 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 acalculated level which is protective of human health against the ingestion of contaminated water andcontaminated aquatic organisms. Thecalculated value is comparable to the present standard. For this reason a selective criterion based on exposure solely from consumption of 6.5 gof aquatic organisms wasnot derived.

viINTRODUCTION

Endrin is the common name of one member of the cyclodiene group of pesticides. It is acyclic hydrocarbon having achlorinesubstituted 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. Theendrin sold in the United States is atechnical grade product, containing not less than 95 percent active ingredient, available in a variety of diluted formulations (Brooks, 1974). Jarvinen andTyo (1978) found the solubility of endrin to beabout 200 μg/l. Known uses of endrin in the United States are as an avicide, rodenticide, and insecticide, the latter being the most prevalent. Thelargest single useof 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 anddisposal of empty containers also contribute significantly to observed residue levels. In the past several years, endrin utilization has been increasingly restricted andproduction has continued to decline. In 1978, endrin production was approximately 400,000 lbs (U.S. EPA, 1978).

A-1REFERENCES

Brooks, G.T. 1974. Chlorinated Insecticides. Vol. I. Technology andApplication. CRCPress, Cleveland, Ohio.
Jarvinen, A.W. andR.M. Tyo. 1978. Toxicity to fathead minnows of endrin in food andwater. 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-2Aquatic Life Toxicology* INTRODUCTION

Endrin isoneofagroupofchlorinated hydrocarbonpesticides developed after the broadscale useof DDT,andit wasincreasingly usedduring the 1950's. Perhapsbecauseendrin hasahigh acutetoxicity toaquatic organisms, it wasmorefrequently tested in aquatic toxicity tests than related insecticides suchaschlordane, heptachlor, andaldrin. Becauseit is abroadspectrum pesticide, endrin wasusedto control manypests including termites, mice, andarmyworms. In the latter 1960's it wasextensively usedfor cotton bollworm control. Its persistence in soil, while goodfor termite control, ledto its discontinuation for control oftobaccoworms. Earlytesting identified its hightoxicity tomammals. Endrin is very insoluble in water. Recently, Jarvinen andTyo(1978) usedasaturator intheir toxicity tests andfoundthe solubility infreshwater to beabout200μg/l. Nearly all of the early workwith endrin and aquatic animals usedacetoneorsomeother carrier solvent, andinthosefew tests where concentrations weremeasured, the actual concentrations were frequently considerably lowerthanthecalculated ones. Someworkersuseda wetting agentsuchasTriton® X-100in the acetone-endrin solution to improvedispersion in thetest water. Becauseconcentrations werenot measured, thetoxicity datareported maynotreflect thetrue toxicity. Fergusonandco-workers at Mississippi State University havepublished numerousarticles onendrin resistance that developedinnatural populations

*Thereader is referred to the Guidelines for Deriving WaterQuality Criteria for theProtection ofAquatic Life andIts Usesinorder tobetter understand the following discussion andrecommendation. Thefollowing tables contain the appropriate datathat werefound in the literature, andat the bottomof eachtable arecalculations for deriving various measuresoftoxicity asdescribed intheGuidelines.

B-1of freshwater aquatic organisms exposed to high water concentrations of endrin as aresult of its use on cotton. Nearly all of their work used static, unmeasured test procedures, and the data, at best, can beused 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 andfish 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 mayresult in acute or chronic toxicity. EFFECTS Acute Toxicity Twenty-two standard LC50 values have been reported for 15 freshwater invertebrate species (Table 1). Noneof the 22 data were based on measured concentrations, and only two were based onflow-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 magnawas amongthe more tolerant species. The generally higher tolerance of the insects and related groups was unexpected since endrin wasaneffective insecticide.

B-2Table1also lists dataontheacutetoxicity ofendrin to 13freshwater

fish species. Onlysix ofthesedatafromthree different paperswerebased onmeasuredconcentrations andflow-through procedures. Ofthe sevendata points for fathead minnows,twowerederived fromstatic tests with unmeasured concentrations, andthese werenot usedin calculating the species meanacutevaluefor that species. Fewresults of flow-through tests based onmeasuredconcentrations areavailable, duein part tothe limited useof gaschromatographyduring the earlier workandthe high toxicity which required very lowdetection limits whichwerenotachievable until analytical procedureswereimproved. All ofthespeciesmeanacutevaluesfor freshwater fish speciesarebetween0.15 and2.1 μg/l, suggesting arelatively narrow rangeof species sensitivity for fishes. TheFreshwater Final AcuteValuefor endrin, derived fromthespeciesmeanacutevalues listed inTable3usingthe proceduredescribed intheGuidelines, is0.18μg/l. Acute toxicity tests with saltwater invertebrate species also demonstrate that endrin is very toxic (Tables 1and6). Thevariability in LC50or EC50values wasgreater than that for fishes, ranging from 0.037 μg/l for pink shrimpto 790μg/l for theAmericanoyster. Thesensitivity of arthropods to endrin wasnot muchdifferent from the sensitivity of fishes. Thepenaeid shrimp wasthe most sensitive invertebrate family tested, with LC50values from 0.037 to 0.3 μg/l (Schimmel, et al. 1975; Lowe, unpublished; Butler, 1963). LC50values for five other arthropod speciesrangedfrom0.3μg/l for Koreanshrimp (Schoettger, 1970)to25μg/l for blue crab (Butler, 1963). Thesensitivity of different life stagesof grassshrimpis similar, differing byonly afactor of3.4 (Tyler-Schroeder, 1979). American oysters were less sensitive than arthropods, with EC50

B-3values ranging from14.2to790μg/l, basedondecreasedshell deposition or

abnormaldevelopmentoflarvae (Table 1). Thetoxicity of endrin to oysters mayvary dependingonwater temperature. Concentrations decreasing growth by50percent were: 400μg/l at 12°C;33μg/l at24°C;and14μg/l at22°C(Table 1); thesedifferences may however,bereflective ofdifferent laboratories' procedures. Therangein LC50values for saltwater invertebrate species wasonly slightly greater than the rangefor 17species of freshwater mollusks andarthropods; the pinkshrimpwasthemostsensitive invertebrate speciestested. Acutetoxicity tests havebeenconductedwith 17species of saltwater fishes, andsensitivity varies (Tables 1and6)from0.048μg/l for chinook salmon(Schoettger, 1970)to 3.1 μg/l for northern puffer (Eisler, 1970b). Onlytwo (usually tolerant) species, the sheepsheadminnow(Hansen,et al. 1977)andthesailfin molly (Schimmel,et al. 1975)havebeentested for 96 hoursinflow-through tests with measuredendrin concentrations. Sheepshead minnowfry, juveniles, andadults did not differ in their sensitivity to acuteexposuretoendrin (Hansen,etal. 1977). DataonLC50values for saltwater invertebrate species fromacutetoxicity tests onendrin support the hypothesis that the acutetoxicity of endrin is underestimated bystatic tests andbynotmeasuringconcentrations of endrin intest water. Acutevalues basedonnominal concentrations for grass shrimp, andAmericanoysters werehigher thanacutevalues for measuredconcentrations (Table 1). Additionally, LC50values basedonstatic tests were greater than LC50values for flow-through tests of the same duration for sheepsheadminnow,sailfin mollies, shiner perch, dwarf perch, Koreanshrimp, pink shrimp, andgrass shrimp (Eisler, 1969; Schoettger, 1970; Earnest andBenville, 1972; Schimmel, et al. 1975;Tyler-Schroeder, 1979).

B-4The 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, aspecies 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 10to 15percent of that contributed bywater, 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 iviva1 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 -5grass shrimp (Table 2), even though all tested concentrations significantly

impaired somelife-cycle function. Alower limit of 0.03 pg/l was selected because the only effect was adelay in onset of spawning of about oneweek;
adelay of oneweek probably would not affect natural populations. The upper limit of 0.05 "g/l was set based ondecrease in number of ovigerous females anddelay in spawning of 3to 4weeks. Sheepshead minnows (Schimrnel, et al 1975; Hansen, et al. 1977), spot (Lowe, 1966), and mumnichog (Eisler, 1970a) have been exposed to endrin for 10days or longer (Tables 2and 6). Ofthese tests with saltwater fish species, only the life-cycle exposure of sheepshead minnows (Hansen, et al. 1977) is suitable for obtaining achronic 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. Nosignificant 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 andchronic toxicity of endrin (Table 2). Anearly-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-6experiment was 0.16 ug/l. Although mortality in fish exposed to anominal

concentration of 0.31 pg/l was not significant, they were visibly affected byendrin. Fifty-seven percent of the juvenile spot exposed to 0.075 pg/l of endrin died within the first 19days 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 aminimal 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 LCsovalues 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-7Data on the toxicity of endrin to five freshwater species 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 onfive species of saltwater algae and anatural 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 onmost invertebrate andfish species. Menzel, et al. (1970) in tests with four phytoplankton species found effects at concentrations greater than 1ug/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, aFinal 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), andfishes (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 aswell aselimination. The bioconcentration of endrin from water into the tissues of saltwater organisms has also been well studied (Tables 5and 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 BCFdata (Table 6) are available from 96-hour exposures of oysters, grass shrimp, pink shrimp, sheepshead minnows, andsailfin 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 abiological half-life of 67 hours. Based on this experiment, the oysters exposed to endrin for 10days in aflow-through test byWilson (1966) were probably at steady-state and had a BCFof 1,000, based on anominal 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 apartial-life-cycle exposure lasting five months. Average bioconcentration factors after a96-hour exposure were 830 for grass shrimp and 980for 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-9age BCFfor juvenile sheepshead minnows exposed for 28days was 2,500; for

adults exposed for for 141 to 161 days the BCFwas 6,400 (Hansen, et al. 1977), and for juvenile exposed for four days the BCFwas 2,600 (Schimnel, et al. 1975). Sailfin mollies exposed to endrin for four days had an average BCFof 2,400 (Schimmel, et al. 1975). The geometric mean of normalized BCFvalues for endrin for freshwater and saltwater aquatic life is 1,324 (Table 5). This value was obtained by first dividing each BCFfor which apercent lipid value is available by that percent lipid value to obtain anormalized BCF,which is what the BCFwould be if the percent lipids were 1 percent. Normalized BCFvalues obtained
were: fathead minnow, 1,892; sheepshead minnow, 694 and 1,778; spot, 1,318. The geometric meanof all freshwater and saltwater normalized BCFvalues was then calculated. Dividing the FDAaction level of 0.3 mg/kg for edible fish andshellfish by the geometric mean of normalized BCFvalues (1,324) and by a percent lipid value of 15for freshwater species (see Guidelines) gives afreshwater residue value of 0.015 ug/l. Dividing the FDAlevel by the geometric mean of normalized BCFvalues and by a percent lipid value of 16 for saltwater species (see Guidelines), asaltwater residue value of 0.014 pg/l is calculated similarly. Dividing the FDAaction level of 0.3 mg/kg by the highest BCFfor 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 FDAaction level of 0.3 mg/kg for fish oil bythe geometric
meanof normalized BCFvalues (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-10Other 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 maybe too high because, on the average, the concentration in 50percent of species similar to those used to derive the value will exceed the FDAaction 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 28freshwater species including awide 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. Noneof 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 wasmore sensitive (0.039 ug/l) and gave amuch larger acute-chronic ratio of 18. The geometric mean of these four estimates of the acuteB-11chronic 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 andthe 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 FDAaction level of 0.3 mg/kg for fish oil, the geometric mean of normalized bioconcentration factors (1,324), and apercent lipid value of 100for fish oil, aFinal Residue Value of 0.0023 vg/l is calculated for both freshwater and saltwater. The Final Residue Value maybe too high because, onthe average, the concentration in 50percent of species similar to those used to derive the value will exceed the FDAaction level. The plant data clearly indicate that plants are muchmore 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 donot 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 a24Slour average, andthe 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 a24-hour average, and the concentration should not exceed 0.037 ug/l at any time.

B-12m * 0 A In d0 -I .- I

R CD'

a W-T I I I



### In ui

a VJ-Table 1. Khnt hued)
Spacles Shiner perch, Cymatogaster sregata Dwarf perch, Hlcranetrus mlnlmus Owarf perch, Hlcfctnetrus mlnimus

aIbehead, Thalassoma blfasclatum Strlped mu1let, Mugll cepha Ius 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 6t)envl Ile, 1972 Earnest &Benvll le, 1972 Earnest 6Ebnvllle, 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-20Table 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) Retarence FRESHWATERSPECIES LC 0.14-0.25
LC 0.22-0.3 0.26 0.19

SALTWATERSPECIES 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

uLC = 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 affeet 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
SpecIes Fathead mlnnow, Plmephales promelas Acute Chron Ic 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-21L *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 Ic (&I 1 Ratio FRESHWATERSPECIES Cladoceran, Daphnla magna &Yf IY, Hexagenla blIlneata 352
64 Copepod fcyclopold), (unidenilfied) 60 Cladoceran, Simocephalus serrulatus 34 Crayf Ish, Orconectes naIs 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
chlnook saItem, Oncorhynchus tshawytscha I.I



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2* ZZ -- --P 3;;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 bythe 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 DOTand 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-36Davis, 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 amarine 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.37Ferguson, 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 SewageWorks. 12: 276.
Grant, B.F. andP.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 (Salmoqairdneri). 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. Memorandumto 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-38Henderson,

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. Memorandumto R.C. Russo. June 13.

Jarvinen, A.W. and R.M. Tyo. 1978. Toxicity to fathead minnows of endrin in food andwater. 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. WaterPollut. Control Fed. 38:1273.

Katz, M. 1961. Acute toxicity of someorganic insecticides to 3species of salmonids andthe 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-39Korn, 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. DDTand 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. Someeffects of endrin onestuarine fishes. Proc. 19th Annual Conf. S.E. Assoc. Gameand Fish Comrn. p. 271.

Ludke, J.L., et al. 1968. Someendrin 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 andendrin 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 onbluntnose minnowsandguppies. U.S.Dep.Inter. FishWildl. Serv.Res.Rep.58. p.38.
Mount,D.I. andG.J. Putnicki. 1966. Summaryreport ofthe 1963Mississippikill. Proc.31stNorthAm.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. andA.R.Gaufin. 1964. Bioassaystodetermine pesticide toxicity to the amphipodcrustacean, Gamrnaruslacustris. UtahAcad. Proc. 41:64.
Post, G.andT.R. Schroeder. 1971. Toxicity of four insecticides tofour 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.SportFish. Wildl. Tech.Paper25.
Sanders,H.O. 1972. Toxicity ofsomeinsecticides tofour species ofmalacostracan crustaceans. U.S.Dep.Inter. Bur. Sport Fish. andWildl. Tech. Paper66.
Sanders,H.O.andO.B.Cope. 1966. Toxicities ofseveral pesticides totwo speciesofcladocerans. Trans.Am.Fish. Soc. 95:165.

B-41Sanders, 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. Gameand 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 LASon the acute toxicity of various insecticides to the fathead minnow (Pimephales promelas Rafinesque). Water Res. 3: 797.
Tyler-Schroeder, D.B. 1979. Useof 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-42Vance, B.D. and W. Drummond. 1969. Biological concentration of pesticides

byalgae. Jour. Am.Water WorksAssoc. 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.

Bd3Mammalian Toxicology and HumanHealth 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 valuesμg/l have been detected both in internal and external fat in avariety of species following ingestion of endrin-contaminated feed. Endrin wasstill detectable in the fat of these animals 42days 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 as12- 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 onthe central nervous system. Whenlethal concentrations are administered to experimental animals, convulsions mayoccur as soon as 30minutes after exposure, and mayculminate 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 50mg/kg for the goat. Manycases 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 33to 100percent of

C-1various species of wild mice inhabiting the target area (Dana and

Shaw, 1958). Thechronic 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 mayalso befatal. Doses of 0.49 to 0.81 mg/kg in the diet wasfatal to dogs in 5to 6months. Twelve mg/kg in the diet for life decreased the survival time for mice. Deer mice succumbed to adiet which contained only 2mg/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. Humanexposure to endrin occurs through the diet, from inhalation, and through dermal contact. The average dietary intake in the United States in 1973was0.033 μg/day (0.0005 μg/kg/day) for a 69.1 kg man. This is far below the maximumdaily 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-2Outbreaks 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 seemsto 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 wasthe most commoncause of death. Significantly increased activity of the hepatic microsomal drugmetabolizing enzymes has occurred in individuals employed in the manufacture of endrin. Noreports of irreversible adverse effects of occupational exposure to endrin have been found in the available literature. Foodcontamination byendrin still occurs, but to adecreasing 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 andwild animals andfish which feed oncontaminated vegetation. Ingestion of 20 mgendrin 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 andpass it onto predatory birds. This contaminated fowl (or the fish themselves) may, in turn, be ingested by humans.

C-3In animals, chronic exposure to endrin mayresult in damageto

the liver, kidneys, heart, brain, lung, adrenal glands, andspleen. 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 ashigh as3μg/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, wasfound to contain amaximumof 3 ng/l (Lauer, et al. 1966). In a study conducted between March 1964 and June 1967, more than 500grab samples of finished drinking water andcorresponding rawwater were collected from 10selected 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 0in athree year period from 1964 to 1967 (Schafer, et al. 1969).

C-4Arecent 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 NewOrleans, Louisiana. Thehighest concentration measured from all samples was 4ng/l. Ingestion from Food The general population has little exposure to endrin in the diet. In aseries 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). Thesix year average intake was0.005 μg/kg body weight per day. Amarket 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 andsweet potatoes by20percent (Solar, et al. 1971)l Heat processing and freezing further lowered potato residues 65 and 52 percent, respectively. Studies on

C-5TABLE 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 ---

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

I-- :,turnips (Wheeler, et al. 1969) andcarrots (Hermanson, et al. 197C)

identified 50to 80percent 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 beviewed with skepticism (Hill, 1970). The loss of endrin from crops depends on the sumof many factors, including temperature, volatilization, metabolism, and dislodgement by wind and rain. Since generalizations cannot bemadethat endrin onagiven crop will always "disappear" at the samerate, residue analyses on harvested crops are the most effective meansof determining potential humanexposure. Abioconcentration factor (ECF) relates the concentration of a chemical in aquatic animals to the concentration in the water in which they live. The steady-state l3CFsfor a lipid-soluble compound in the tissues of various aquatic animals seemto beproportional to the percent lipid in the tissue. Thus, the per capita ingestion of alipid-soluble chemical can beestimated from the per capita consumption of fish andshellfish, the weighted average percent lipids of consumed fish and shellfish, and asteady-state RCF for the chemical. Data from arecent survey onfish 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 usedwith data onthe fat content of the edible portion of the same species to estimate that the weighted average percent

C-8lipids for consumed freshwater and estuarine fish and shellfish is

3.0 percent. Twolaboratory studies, in which percent lipids and asteadystate BCFwere measured, have been conducted on endrin. The mean of the BCFvalues, after normilization to 1 percent lipids, is 1,324 (see Table 5in Section B). Anadjustment factor of 3can be used to adjust the mean normalized BCFto the 3.0 percent lipids that is the weighted average for consumed fish and shellfish. Thus, the weighted average BCFfor endrin andthe edible portion of all freshwater and estuarine aquatic organisms consumed by Americans is calculated to be3,970. 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). Humansmay 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 onafat basis were detected in the milk of 40Wisconsin dairy

c-9herds between 1964 and 1967 (Moubry, et al. 1968). Endrin waspresumably retained in the milk fat for upto four weeks. However, the quantities of endrin ingested during that period were not controlled. Williams andMills (1964) studied the excretion of endrin in cows' milk under controlled feeding conditions. Endrin concentrations in the milk increased progressively during the first fewdays of feeding until they plateaued at 13to 15days. Wheningestion of endrin ceased, residues in milk declined sharply and following 20 days onanendrin-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 amixture of pesticides, thus, interactions mayhave occurred, Endrin is apparently excreted in milk in higher concentrations whenfed as aresidue onhaythan whenfed dissolved in soybean oil (Ely, et al. 1957). However, in general, atotal 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 was0.07. 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 upto 2mg/kq resulted in amaximumbody 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-10levels of approximately 6.4 to 13.8 pg/g were still present 42days

after termination of exposure. Pigs receiving 510 mgendrin over 30days hadendrin fat levels of nomore than 2pg/g, and noendrin wasdetected 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 maygenerally beexpected to reach 0.01 mg/hour (Wolfe, et al. 1963, 1967). Wolfe, et al. (1963) reported that spraying of potatoes with a 1percent 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 wasbelow 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 anaverage of 0.2 pgof endrin per commercial cigarette. Forty percent of the residual endrin disappears during the curing process, but the remainder persists throughout the cigarette manufacturing process. Endrin residues in pipe tobacco increased approximately threefold from 1969 (0.05 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-11In asurvey of 45sites in 1971, the highest level of endrin in

ambient air was 25.6 ng/m3 in Greely, Colorado (U.S. EPA, 1971). In aseparate survey of three sites in 1975, the highest reported level of endrin in the ambient air was0.5 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 beas high as3mg/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 mousecontrol byvarious 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, adermal exposure of 0.15 mg/hr was noted during the application of endrin to row crops (Jegier, 1964). PHARMACOKINETICS Absorption Endrin is known to beabsorbed bythe skin, the lungs, andthe gut(U.S.EPA,1979),however, the ratesof theabsorption havenotbeen adequately documented. c-12Distribution

Humansdo 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 humansubcutaneous 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 ashigh as400pg/g in fat tissueand 10pg/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 byeating endrin-contaminated bread (endrin levels ranged from 48to 1,807 ppm) in Saudi-Arabia (Table 3). Blood and urine samples taken from patients 29to 31days 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-13TABLE3 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-140.053 ug/g and 20hours after convulsions it wasrecorded at 0.038

lJg/g' This samepatient excreted 0.02 pg/g endrin via the urine during the following 24hours (Coble, et al. 1967). Richardson, et al. (1967) fed endrin to g-month-old 128 consecutive days at alevel 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 was0.077 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 4to 8mg/kg for approximately six months were asfollows: 1ug/g in the fat; 1 Dg/g in the liver; and0.5 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 asahydrophilic 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 wassuppressed bysesamex, aninhibitor of microsomal oxidations.

c-15Information 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. Oneis 9-ketoendrin, which is found in tissues and in urine. Theother two metabolites are excreted in the feces andhave not been found in body tissues. The acute oral LD50of 9-ketoendrin in rats (62 mg/kg) is higher than that of endrin (25 mg/kg), and the reaction appears to beadetoxication step (Brooks, 1969). Oxidation without skeletal rearrangement is the major metabolic route in mammals although details remain to beworked out (Brooks, 1974). Bedford, et al. (1975) studied oral LD50values 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 maybe the acute toxicant in each case. Thus, the oxidative metabolism of endrin maybe responsible for its acute toxicity. Jager (1970) found, in feeding experiments with rats, that females metabolize endrin more slowly than males. Whencarbon-14 labeled endrin wasfed to male and female rats, the males excreted 60 percent of it in the feces within the first 24 hours and the females only 39percent. Less than 1percent was excreted in the urine. Of the total radioactivity excreted in the feces, 70to 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-16Sexdifferences in the rate of endrin metabolism in rats were

also found byHutson, 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 ahigher rate than did females. Aminor metabolite wastrans-4,5-dihydroisodrin4,5-diol. 12-Ketoendrin was the major urinary metabolite in male rats, whereas the major urinary metabolite in female rats wasanti12-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). Themajor metabolite in both male and female rats wasanti-12- hydroxyendrin, which was excreted via the bile as the glucuronide (Hutson, et al. 1975); trans-4,5-dihydroisodrin-4,5-dial is aminor biliary metabolite. 12-Ketoendrin wasobserved 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 carbon14labeled endrin in whole rats, bile-fistulated rats, and isolated perfused rat livers. Over 90percent of the excreted radioactivity wasfound in the feces of the intact animals and in the bile of the fistulated animals. Fifty percent of the radioactive endrin was

c-17excreted within the first 24hours. 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 andeliminated from the tissues of vertebrates. Excretion occurs through the milk aswell 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, maypersist for longer periods of time. EFFECTS Acute, Subacute, andChronic Toxicity Endrin is classified as"very highly hazardous"; meaning that, any contact with very small amounts of the substance mayresult 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 beelicited from any route of exposure. When ingested in one dose by rats, endrin is about three times as toxic asaldrin and about 15times astoxic as DDT(Treon, et al. 1955). Upon intravenous administration to mice, endrin was five times as toxic asdieldrin (Walsh and Fink, 1972). Theonset 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-18Symptoms of acute endrin poisoning in mammalsclearly 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 andfollowed by severe tonic-clonic convulsions (Brooks, 1974). These convulsions mayoccur asearly as30minutes 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 48hours (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 astress reaction, degenerative changes in kidneys, liver and brain capillaries, and venous congestion, and loss of weight anddehydration of someorgans (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 LD50values range from 1.37 to 43mg/kg. Apparently, mice andmonkeys are most sensitive, andguinea pigs are more resistant. Rabbits seemto be somewhat more resistant than monkeys to asingle 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. Thegreater 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. Whenendrin wasmaintained in contact, asadry 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 wasnot 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 wasobserved (Treon, et al. 1955). Graves andBradley (1965) determined anLDSOof 5.6 mg/kg for endrin injected into the peritoneal cavity of Swiss albino mice. An intravenous LDsOof 2.3 mg/kg wasdetermined by Walsh and Fink (1972) for adult male mice. Endrin injected into dogs intravenously at adosage of 3mg/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-20TABLE4 Acute Oral Toxicity of Endrin to Mammals
Animal (age, sex) Mouse 1.37a Rats (6months, M) 43b Rats (6months, F) 7b Rats (30days, M) 3ob Rats (30days, 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-21time. The central nervous system is the target of acute endrin

poisoning. Whenananimal is repeatedly exposed to low doses (0.8 to 3.5 mg/kg/day) of endtin, it canoften makecompensatory 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 was10mgendrin/kg body weight. This convulsive dose wasapproximately twice that KeqUiKed fOK the production of teratogenic effects. Revzin (1968) found that chronic administration of endrin can lead to convulsions. Headministered 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 12times asmuchDDTasendrin without increase in relative weights of specific organs. Onthe basis of organ weights dogs are at least 10times 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-22Mammalian species appear to besensitive to the toxic effects

of endrin at low levels in their diet. Significant mortality during a7-month period appeared in deer mice when fed 2mg/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. A48-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 andwill be reviewed in detail. This paper includes acute, subacute (3 to 6 months), and chronic (2 year) feeding studies in rats (male and female) ;asubacute oral study in rabbits (8to 10weeks): anda19- month oral study in dogs (male and female). Body weights, organ weights, and histopathologic data are included. The rabbit studies were limited to asingle dose level. Four of five female rabbits given 1mg/kg/day of endrin in peanut oil died in 8 to 10 weeks. The surviving animals sacrificed after

C-2350doses over 10weeks showed "diffuse degenerative andfatty vacuolization of the hepatic and renal cells" and degeneration of the heart. Thus, to the rabbit adose of 1mg/kg/day of endrin is extremely toxic. Initial subacute rat studies gave the following results. All rats survived 50doses (in peanut oil) over 10weeks at the 1mg/ kg/day level. At 2mg/kg/day 1of 2young female rats and 1of 3 adult female rats died during the 10week study. All six of the male rats (young and old) survived the 10week dosing at 2mg/kg/ day. Three of three male rats also survived asimilar study at a dose level of 5mg/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 ppmof endrin in the diet.
GKOU~S of 20male and 20female 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 200g rat eats 20g food/day then 100 ppm (100 ug/g) in the food translates to 10 mg/kg/day intake of endrin; 50ppmto 5.0 mg/kg/day; and25, 5, and 1ppmto 2.5, 0.5, and 0.1 mg/kg/day, respectively. Endrin in the diet of female rats at 100, 50, or 25ppmcaused significant mortality at 80 weeks (Table 5). The male rats were somewhat less susceptible showing increased mortality only at the 100and 50ppm

C-24level. Dietary levels of 100 OK 50ppmresulted in the early deaths

of all but afew resistant rats. Bodyweight gains were not particularly altered bythese dosages of endrin, nor wasthe rate of live weight to body weight changed. In the male rats fed 25 ppm (2.5 mg/kg/day) OK 5ppm (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 1ppm (0.1 mg/kg/ day) dietary level, nor wasthere any effect in female rats at the 0.5 OK0.1 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 OKless. Animals that died whenfed 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. Asingle statement notes that the incidence of neoplasia wasnot greater amongexperimental rats than amongthe contKOlS. Treon, et al. (1955) also conducted an extensive dog study summarized in Table 6which is taken directly from the published paper. This table provides the dosage in both ppmin the diet and daily intake as mg/kg body weight. All dogs died when fed 0.5 to 4.0 mg/kg/day (10to 50ppm) in the diet and more than half of those fed 0.20 to approximately 0.5 mg/kg/day (5to 8ppm)also died. All dogs survived when their diets contained 4ppm (0.15 to 0.21 mg/kg/day) or less for periods up to 18.7 months. All dogs fed 10ppm (0.49 to 0.81 mg/kg/day) suffered extensive weight loss;

C-26Gl

. 

### cr

. z rl l-l U 5 2 $ a .d Er kl I:

tc-4 d I 0 N d

ln l-l
c

0 0 4I =:

. cv

0 In r. U 5 m Lc 2 -4 m

d dd CL I I cv z:

. . . 0 00

23 N cvthose fed at 8ppm (0.29 to 0.62 mg/kg/day) gained weight intially,

but eventually failed to continue growing. Those fed 4ppm (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 1ppm (0.045 to 0.12 mg/kg/day) grew aswell 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 1ppmlevel 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 8ppmendrin (0.29 to 0.62 mg/kg/ day) fOK six months had enlargement of the liver, kidney, and
brain. At 3ppm(0.12 to 0.25 mg/kg/day) the kidney andheart were significantly enlarged at sacrifice (18.7 months). Dogsfed at the 1ppmlevel (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 3ppm (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 5ppm (0.20 to 0.27 mg/kg/day) group and no dogs surviving doses

C-28greater 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 1ppm (0.045 to 0.12 mg/kg/day) to 50ppm (2.5 to 4.0 mg/kg/day) anddemonstrated ano-effect level at the lowest dose of 1ppm (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 1to 3mg/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 (LD50) 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 DDTand, particularly, endrin plus delnav showed lower than expected LDsOs, suggestive of antagonistic effects. Joint administration of endrin andits closely related compound aldrin showed amore than additive effect, and endrin plus chlordane wasfound to exert apotentiating effect.
No other information is available onsynergistic and/or antagonistic effects of endrin. Teratogenicitv Rats and mice were given 0.58 mgendrin/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-29TABLE7 Expected andObserved Oral LDSOsof 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 onefetus with club foot wasin the control group of 303. Endrin exerted embryocidal andteratogenic effects onpregnant hamsters. Both soft and skeletal tissue malformations were produced. Single oral doses of endrin (5mg/kg) administered to pregnant Syrian golden hamsters onday 7, 8, OK9of gestation caused a high incidence of fetal death, congenital abnormalities andgrowth retardation. Thirty-two percent of the implantations resulted in fetal mortalities. Teratogenic effects were observed in 28percent of the fetuses from hamsters treated on day eight. Open eye OCCUKKedin 22percent, webbed foot in 16percent, 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 beteratogenic in mice, but frequency and gravity of the defects produced were less pronounced than in the hamsters whenasingle dose (2.5 mg/kg in mice and 5mg/kg in hamsters) of half the LDSOwasadministered. 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 5to 14 of gestation, had less reactive 1OCOmOtOKactivity than controls during gestation but not at weaning (Gray, et al. 1979). Theoffspring of these damswere tested in open field at 15, ZG, z7, 34, and 44 days of age. Fifteen-day-old pups at the

c-311.5 mg/kg dose were approximately 90percent 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 onday 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 humanconvulsions 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 mgendrin 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-32damage resulted in liquefication and transformation of the chromatin mass into anamorphous 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 nodata 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 0ppmTreont 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 alifetime at dietary levels of 2, 6, or 12mg/kg wasneither tumorigenic nor carcinogenic (Deichmann, et al. 1970; Deichmann andMacDonald, 1971; Deichmann, 1972). Arecently 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-33CRITERION 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) wassuggested by Ettinger andMount (1967) andwasaccepted asaguideline. Amaximum 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 onPesticide Residues in Food held in Rome, November, 1972 (FAO, 1973). Athreshold 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). Athreshold limit value of 100 ug/m3 for an8-hour time-weighted average occupational exposure has also been established by the Occupational Safety andHealth Administration (OSHA) (29CFR1910.1000). Toxic pollutant effluent standards (40 CFR129.102) were promulgated by the U.S. EPA. These allowed aneffluent concentration of 1.5 pg/l per average working day calculated over aperiod 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 kgper 1,000 kgof production.

c-34Current Levels of Exposure

while no recent data are available on levels of exposure of humansto endrin it appears that the risk of exposure is decreasing because of the decreased usage of the pesticide. In asurvey of over 500drinking 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). Themost recent study found only 4rig/l in contaminated drinking water (U.S. EPA, 1974). In aseries 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 amaximum level of 25.6 ug/m3 in 1971 at Greeley, Colorado, to amaximum 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 maybeexposed 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 useof endrin occurs whenmeasuring andpouring the emulsifiable concentrate material. Because endrin has been c-35shown to cause teratogenic effects, pregnant women, particularly

those whose diets maycontain large amounts of fish, must also be considered aspecial group at risk. Evidence that endrin maycause chromosomal damage in germinal tissue suggests that menand women of child-bearing intent mayalso 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, anunusual 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. Onthe basis of long-term dietary studies in mammals, a realistic drinking water criterion may be proposed. Maximum noC-36observed-effect and gradual dosage dietary levels of endrin reported for experimental animals are shown in Table 8. Thedata in Table 8suggest that there is considerable species difference in the response to endrin. A1.0 mg/kg single dose produced 4/5 deaths in rabbits, yet this amount is reported as anoobserved-effect level in the rat andmouse byother 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, arange of dosages wasfed and the animals followed by observation, body weight, organ weights, andhistopathologic examination of tissues at sacrifice. The rat study byTreon, et al. (1955) suggests ano-observedeffect level (NOEL) in a2-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 NOELat approximately 0.1 mg/kg/day. Monkeys maybe 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 1to 3mg/kg were fatal. Thus, long-term studies in both the rat and the dog suggest that the NOELis approximately 0.1 mg/kg/day. Extrapolation from these two animal studies to manappears 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. Humanexposure to endrin was calculated on the

c-37basis of daily ingestion of 21of water and 6.5 gfish with aBCF

of 3,970 for endrin. Using a no-effect dosage level if 0.1 mg/kg/day the total acceptable daily intake (ADI) for a70 kg person is: 0.1 mgendrin/kg x70kg =1oO 7mg/day (uncertainty factor) =7oug/day The criterion for endrin is thus: 70pg/day ' =2 1+ (0.0063 kgx3,910) =*05’ Pg/l -2.5 pg/l This approximates closely the 1pg/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 1pgendrin/l of ambient water (1ppb). This calculation assumes that 100percent of man's exposure is assigned to the ambient water pathway. Although it is desirable to establish acriterion 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 1vg/l allowable concentration in the drinking water standards is reasonably close to 2.5 pg/l, it is recommended that 1.0 pg/l beused asthe criterion with notation that there are

c-39special groups at risk.* Drinking water contributes 5percent of

the assumed exposure while eating contaminated fish products accounts for 95percent.

*If endrin was present in waters from which edible fish were located and if these fish concentrate endrin byafactor of 3,970, this criterion may not be sufficient to protect a special high risk group i.e., pregnant womenwhoconsume asingle dose of endrin contaminated fish. Given the BCF, fish in water at the maximumrecommended concentration of 1ug/l, maycontain 3.8 pg/g endrin. A250gportion of fish would contain approximately 1.0 mg endrin (or 0.02 mq/kg for a50kg female). This dose provides a margin of safety of only 75over the NOELof 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 250gof fish at agiven time. The recommended water quality criterion of 1pg/l wasbased onachronic exposure study, teratologic outcomes are more likely to occur with acute exposures at critical times in gestation.

c-40REFERENCES

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