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Ambient
Water Quality
Criteria for
Aldrin/Dieldrin
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AMBIENT WATER QUALITY CRITERIA FOR
ALDRIN/OIEIDRIN
Prepared By
U.S. ENVIRONMENTAL PROTECTION AGENCY
Office of Mater Regulations and Standards
Criteria and Standards Oivision
Washington, O.C.
Office of Research and Development
Environmental Criteria and Assessment Office
Cincinnati, Ohio
Carcinogen Assessment Group
Washington, D.C.
Environmental Research Laboratories
Corvalis, Oregon
Duluth, Minnesota
Gulf-Breeze, Florida
Narragansett, Rhode Island
i
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DISCLAIMER
This report has been reviewed by the Environmental Criteria and
Assessment Office, U.S. Environmental Protection Agency, and approved
for publication. Mention of trade names or conmercial products does not
constitute endorsement or. recommendation for use.
AVAILABILITY NOTICE
This document is available to the public through the National
Technical Information Service, (NTIS), Springfield, Virginia 22161.
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FOREWORD
Section 304 (a)(1) of the Clean Water Act of 1977 (P.L. 95-217),
requires the Administrator of the Environmental Protection Agency to
publish criteria for water quality accurately reflecting the latest
scientific knowledge on the kind and extent of all identifiable effects
on health and welfare which nay be expected from the presence of
pollutants 1n any body of water, Including ground water. Proposed water
quality criteria for the 65 toxic pollutants listed under section 307
(a)(1) of the Clean Water Act were developed and a notice of their
availability was published for public comment on March 15, 1979 (44 FR
15926), July 25, 1979 (44 FR 43660), and October 1, 1979 (44 FR 56628).
This document 1s a revision of those proposed criteria based upon a
consideration of comments received from other Federal Agencies, State
agencies, special interest groups, and individual scientists. The
criteria contained in this document replace any previously published EPA
criteria for the 65 pollutants. This criterion document is also
published in satisfaction of paragraph 11 of the Settlement Agreement
in Natural Resources Defense Council, et. al. vs. Train. 8 IRC 2120
(O.O.C. 1976), modified, 12 EM 1533 (O.D.C. 1979). "~
The term "water quality criteria" is used in two sections of the
Clean Water Act, section 304 (a)(1) and section 303 (c)(2). The term has
a different program impact in each section. In section 304, the term
represents a non-regulatory, scientific assessment of ecological ef-
fects. The criteria presented in this publication are such scientific
assessments. Such water quality criteria associated with specific
; stream uses when adopted as State water quality standards under section
303 become enforceable maximum acceptable levels of a pollutant in
ambient waters. The water quality criteria adopted in the State water
quality standards could have the same numerical limits as the criteria
developed under section 304. However, in many situations States may want
to adjust water quality criteria developed under section 304 to reflect
local environmental conditions and human exposure patterns before
incorporation into water quality standards. It is not until their
adoption as part of the State water quality standards that the criteria
become regulatory.
Guidelines to assist the States in the modification of criteria
presented in this document, in the development of water quality
standards, and in other water-related programs of this Agency, are being
developed by EPA.
STEVEN SCHATZOW
Deputy Assistant Administrator
Office of Water Regulations and Standards
iii
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ACKNOWLEDGEMENTS
Aquatic life Toxicology:
Charles i. Stephen, ERL-Duluth
U.S. Environmental Protection Agency
David J. Hansen, IRL-Gulf Breeze
U.S. Environmental Protection Agency
Mammalian Toxicology and Human Health Effects:
Thomas Connor (author)
University of Texas Medical Branch
Michael L. Dourson (doc. mgr.)
ECAO-Cin
U.S. Environmental Protection Agency
Karen Blackburn (doc. mgr.), ECAO-Cin
U.S. Environmental Protection Agency
Gordon Chesters
University of Wisconsin
David J. Hansen, ERL-Gulf Breeze
U.S. Environmental Protection Agency
Steven D. Lutkenhoff, ECAO-Cin
U.S. Environmental Protection Agency
David J. McKee, ECAO-RTP
U.S. Environmental Protection Agency
Shane S. Que Hee
University of Cincinnati
Jerry F. Stara, ECAO-Cin
U.S. Environmental Protection Agency
James Barnett
University of Texas Medical Branch
William B. Buck
University of Illinois
Gary Chapman, ERL-Duluth
U.S. Environmental Protection Agency
Patrick Durkin
Syracuse Research Corporation
Alfred Garvin
University of Cincinnati
Fumio Matsamura
University of Michigan
W. Bruce Peirano, HERL
U.S. Environmental Protection Agency
Herbert Schumacher
National Center for Toxicological Res
Roy E. Albert*
Carcinogen Assessment Group
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.
Paynes, S.J. Faehr, L.A. Wade, D. Jones, B.J. Bordicks,
B.J. Quesnell, C. Russom, R. Rubinstein.
*CAG Participating Members:
Elizabeth L. Anderson, Larry Anderson, Dolph Arnicar, Steven Bayard,
David L. Bayliss, Chao W. Chen, John R. Fowle III, Bernard Haberman,
Charalingayya Hiremath, Chang S. Lao, Robert HcGaughy, Jeffrey Rosen-
blatt, Dharm V. Singh, and Todd W. Thorslund.
1v
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TABLE OF CONTENTS
Page
Criteria Suamary
Introduction A-l
Aquatic Life Toxicology 8-1
Introduction 8-1
Effects . 8-1
Acute Toxicology 8-1
Chronic Toxicology 8-4
Plant Effects B-6
Residues 8-6
Miscellaneous 8-9
Summary 8-10
Criteria B-12
References 8-43
Manrnalian Toxicology and Human Health Effects C-l
Introduction C-l
Exposure C-2
Ingestion from Water C-3
Ingestion from Food C-5
Inhalation C-10
Dermal C-ll
Pharmacokinetics C-ll
Absorption C-ll
Distribution C-12
Metabolism C-19
Excretion C-25
Effects C-32
Acute, Subacute, and Chronic Toxicity C-32
Synergism and/or Antagonism C-35
Teratogenicity C-36
Mutagenicity C-38
Carcinogenicity C-44
Criterion Formulation C-61
Existing Guidelines and Standards C-61
Current Levels of Exposure C-62
Special Groups at Risk C-62
Basis and Derivation of Criteria C-63
References C-66
Appendix C-81
v
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CRITERIA DOCUMENT
ALDRIN-DIELDRIN
CRITERIA * *
Aquatic Life
Dieldrin
For.dieldrin the criterion to protect freshwater aquatic life
as derived using the Guidelines is 0.0019 fig/1 as a 24-hour aver-
$.0
age, and the concentration should not exceed £s5 jug/1 at any time.
For dieldrin the criterion to protect saltwater aquatic life
as derived using the Guidelines is 0.0019 jug/1 as a 24-hour aver-
age, and the concentration-should not exceed 0.71 jig/1 at any time.
Aldrin
For freshwater aquatic life the concentration of aldrin should
1-G
not exceed pg/1 at an? time. No data are available concerning
the chronic toxicity of aldrin to sensitive freshwater aquatic
life.
For saltwater aquatic life the concentration of aldrin should
not exceed 1.3 /jg/1 at any time. No data are available concerning
the chronic toxicity of aldrin to sensitive saltwater aquatic life.
Human Health
For the maximum protection of human health from the potential
carcinogenic effects due to exposure of aldrin through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentration should be zero based on the non-threshold
assumption for this chemical. However, zero level may not be
attainable at the present time. Therefore, the levels which may
-------
result in incremental increase ot cancer risk over the lifetime are
estimated at 10~S, 10"6 and 10"7. The corresponding recommended
criteria are 0.74 ng/1, 0.074 ng/1, and 0.0074 ng/1, respectively.
If the above estimates are made for consumption of aquatic organ-
isms only, excluding consumption of water, the levels are 0.79
ng/1, 0.079 ng/1, and 0.0079 ng/1, respectively.
For the maximum protection of human health from the potential
carcinogenic effects due to exposure of dieldrin through ingestion
of contaminated water and contaminated aquatic organisms, the
ambient water concentration should be zero based on the non-
threshold assumption for this chemical. However, zero level may
not be attainable at the present time. Therefore, the levels which
may result in incremental increase of cancer risk over the lifetime
— § —6
are estimated at 10 , 10 and 10 . The corresponding recom-
mended criteria are 0.71 ng/1, 0.071 ng/1, and 0.0071 ng/1, respec-
tively. If the above estimates are made for consumption of aquatic
organisms only, excluding consumption of water, the levels are 0.76
ng/1, 0.076 ng/1, 0.076 ng/1, and 0.0076 ng/1, respectively.
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INTRODUCTION
Aidrin and dieldrin have bean two of tha most widely used
domastic pesticides. They ara chlorinatad hydrocarbon compounds.
Although aldrin is usad in graatar quantity thin dieldrin, aldrin
quickly transforms into dialdrin in tha environment. Hence, there
is concern with both compounds. The primary use of the chemicals
¦i in the past was for control of corn pests, although they were also
* used by.the citrus industry. Uses are restricted to those where
there is no effluent discharge.
Aldrin use in the United States peaked at 19 million pounds in
1966 but dropped to about 10.5 million pounds in 1970. During that
same period dieldrin use decreased from 1 million pounds to about
670,000 pounds. The decreased use has been attributed primarily to
* increased insect resistance to the two chemicals and to development
* and availability of substitute materials.
Aldrin and dieldrin have been the subject of litigation bear-
ing upon the contention that these substances cause severe aquatic
environmental change and are potential carcinogens. Zn 1970, the
U.S. Department of Agriculture cancelled all registrations of these
pesticides based upon a concern to limit dispersal in or on aquatic
areas. In 1972, under the authority of the Fungicide, Insecticide,
Rodenticide Act as amended by the Federal Pesticide Control Act of
"*¦ 1972, USCS Section 135, et. sec., an EPA order lifted cancellation
*'> of all registered aldrin and dieldrin for use in deep ground inser-
tions for termite control, nursery clipping of roots and tops of
non-food plants, and mothproofing of woolen textiles and carpets
where Jbhere is no effluent discharge. In 1974, cancellation
A-1
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proceedings.disclosed the sever* hazard to human health and suspen-
sion of registration of aldrin and dieldrin use was ordered; pro-
duction was restricted for all pesticide products containing aldrin
or dieldrin. Bowever, formulated products containing aldrin and
dieldrin are imported from Europe each year solely for subsurface
soil injection for termite control. Therefore, limits that protect
all receiving vater uses must be placed on aldrin and dieldrin.
The litigation has produced the evidentiary basis for the Adminis-
trator's conclusions that aldrin/dieldrin are carcinogenic in mice
and rats, approved the Agency's extrapolation to himtans of data
derived from tests on animals, and affirmed the conclusions that
aldrin and dieldrin pose a substantial risk of cancer to humans,
which constitutes an "imminent hazard" to man.
Aldrin and dieldrin are white crystalline substances with
aldrin melting at 104°C and dieldrin melting between 176 to 17?°C.
Both are soluble in organic solvents with dieldrin the least solu-
ble of the two. The chemical name for aldrin is 1, 2, 3, 4, 10, 10-
hexachloro-1, 4, 4a, 5, 8, 8a-hexahydro-l, 4: S, 8-exo-dimethano-
naphathalene. The chemical name for dieldrin is 1, 2, 3, 4, 10, 10-
hexachloro-6, 7-epoxy-l, 4, 4a, 5, 6, 7, 8, 8a-octahydro-endo, exo-
1, 4: 5, 8-dimetharionaphthalene.
Aldrin is metabolically converted to dieldrin. This epoxida-
tion has been shown to occur in several species including mammals
and poultry, houseflies, locusts, soil microorganisms, a large
number of Lepidoptera species, freshwater fish (Gakstatter, 1968),
and a number of freshwater invertebrates including protozoa, co-
elenterates, worms, arthropods, molluscs, and lobsters. The aldrin
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molecule is biologically altered in the environment to a more sta-
ble and at least equally toxic form, dieldrin. Oieldrin is known-
to be metabolically degraded as shown by Matsumura and-Boush (1967)
, i
and Patilr et al. (1972)j however, its persistence in the environ-
ment is due to its extremely low volatility (i.e., a vapor pressure
of 1.78 x 1
-------
Woolson, 1967). When applied to sandy soil at a rata of 100 ppa,
residues could b« found 15 years latar. Hatsunrara and Boush (1967)
found that of 577 bactarial isolates collected from araas heavily
contaminated with dieldrin, 10 isolatas would altar dialdrin to two
to nina unidentified metabolites. The microbes were members of
Pseudomonas, Bacillus# and Trichoderma genera. Subsequent micro-
biological studies by Wedemeyer (1968) revealed that Aerobacter
aerogenes also will alter dialdrin similarly to £«7- trans-dihydro-
xydihydroaldrin. Chacko, et al. (1966) tested this capability of
17 species of fungi and-actinomycetes. Though most degraded penta-
chloronitrobenzene (PCNE) or DOT or both, none degraded dieldrin.
Patil, et al. (1972), studied the metabolic transformations of
aldrin/dieldrin by marine algae, surface film, sediments, and wa-
ter. They found that the insecticide was not degraded or metabo-
lized in sea water or polluted waters. Some sarine algal popula-
tions were shown to degrade aldrin to dieldrin.
Alterations of dieldrin by bacterial systems result in the
formation of at least one acidic product (Matsumura and Boush,
1967). Once in the fatty tissue of organisms, dieldrin remains
stable, according to Sanborn and 7u (1973). However, dieldrin can
be mobilized from fatty tissue as demonstrated by Brockway (1973);
for example, when fish are placed in an environsent without diel-
drin, there is an elimination from the tissue (Brockway, 1973).
The elimination rate depends upon the diet with fasted fish elim-
inating dieldrin more rapidly than fed fish because of the utiliza-
tion of fat stores (Grzenda, et al. 1972).
A-4
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The dieldrin eliminated from the tissues reenters the water
and thus becomes available for bioconcentration by other organisms.
The movement of dieldrin among organisms, water, and sediment is
dynamic# with equilibrium attained when the chemical concentration
is constant.
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REFERENCES
Brockvaj, D.C. 1973. Th* uptake# storage and release of dieldrin
r
and some effects of its release in the fish, C'iehlosoaa bimaculatum
(Linnaeus). Diss. Abstr. Int. 33: 4323B.
Chacko, C.I. # et al. 1966. Chlorinated hydrocarbon pesticides:
Degradation by microbes. Science. 154: 893.
Edwards, C.A. 1966. Insecticide residues in soils. Residue Rev.
13: 83.
Gakstatter, J.H. 1968. Rates of accumulation of 14C-dieldrin resi-
dues in tissues of goldfish exposed to a single sublethal dose of
14C-aldrin. Jour. Pish. Res. Board Can. 25: 1797.
Grzenda, A.R., et al. 1972. The elimination and turnover of 14C-
dieldrin by different goldfish tissues. Trans. Am. Fish. Soc.
101s 686.
International Agency for Research on Cancer. 1974. Dieldrin. IARC
monographs on the evaluation of carcinogenic risk of chemicals to
man: Some organochlorine pesticides. 5: 125.
Matsumura, F., and G.M. Boush. 1967. Dieldrin: Degradation by soil
microorganisms. Science. 156: 959.
A-6
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Aquatic Life Toxicology*
INTRODUCTION
Aldrin and dieldrin are members of a group of synthetic cyclic hydrocar-
bons called cyclodienes. The group includes other insecticides such as
chlordane, heptachlor, endosulfan, and endrin. Until recently, aldrin and
dieldrin were the most widely used domestic pesticides, with aldrin being
applied in much greater- quantities tfan dieldrin. Aldrin was applied to
soils and foliage using soil injection or aerial techniques. Since leaching
by water was minimal, soil erosion and sediment transport were the two major
routes for aldrin to enter aquatic environments. However, these pesticides
are often considered together because aldrin is rapidly converted to diel-
drijj by metabolism by animals and plants or by photodecomposition. This
conversion is accomplished through the addition of an epoxide group to the
aldrin molecule.
Since aldrin is rapidly ¦converted to dieldrin and since adeauate data
are not available for the species required by the methodology, no criterion
has been developed for aldrin. The following discussion is based on diel-
drin data only except where specifically noted otherwise.
EFFECTS
Acute Toxicity
Results of 14 freshwater acute toxicity tests on dieldrin and inverte-
brate species are presented in Table 1. All of these tests were conducted
under static conditions, and concentrations were not measured. The results
*The reader is referred to the Guidelines for Deriving Water Quality Crite-
ria for the Protection of Aauatic Life and Its Uses in order to understand
this section better. The attached tables contain pertinent available data,
and at the bottoms of the appropriate tables are calculations deriving vari-
ous measures of toxicity as described in the Guidelines. ,
8-1
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ranged from a 96-hour LCgg value of 5.0 yg/1 for the isopod Asellus brevi-
caudus to 740 yg/1 for a crayfish (Sanders, 1972)*. This range of about 150
times demonstrates definite differences 1n species sensitivity to this
compound.
Results of 12 acute toxicity tests with freshwater invertebrate species
and aldrin are also presented in Table 1. Each test was conducted so that
data could be compared with data obtained from comparable tests with diel—
drin. Al'drin 96-hour LCgg values range from 8 yg/1 for an isopod (San-
ders, 1972) to 38,500 yg/1 for the scud, Gammarus lacustris (Gaufin, et al.
1965). Generally, the LCjq values for aldrin are higher than those for
dieldrin, except for cladoceran species which are more sensitive to aldrin.
Sixty-five acute toxicity tests on dieldrin and freshwater fish species
are reported in Table 1. The tests were conducted with eight species of
fishes including both coldwater and warmwater fishes. All of the tests were
static, and none included measured concentrations.
The most sensitive fish species tested was the rainbow trout with
96-hour LC50 values between 1.1 and 9.9 yg/1. The other salmonids, coho
and Chinook salmon; had 96-hour LCgQ values of 10.8 and 6.1 yg/1, respec-
tively. The most resistant fish species was the goldfish with a 96-hour
LCgQ value of 41 yg/1. In the middle of the range, between the salmonids
and the goldfish, were fathead minnows (range 16 to 36 yg/1) and the blue-
gill (range 8 to 32 yg/1). Special attention should be given to the data on
the guppy in the report by Chadwick and Kiigemagi (1968) concerning the de-
velopment of a toxicant delivery system. To determine the efficiency of the
system, toxicity tests with the guppy were conducted over an extended time
period, and the data are included in Table 1. Thirty-eight of the six-
ty-five test results are from this study; they range from 2.3 to 10 yg/1.
B-2
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Twenty tests were conducted on aldrin with 12 freshwater fish species.
The range of the 96-hour LCgQ values (2.2 to 45.9 ug/1) is similar to the
range (1.1 to 41 ug/1) obtained for dieldrin. Comparison of results from
tests on both aldrin and dieldrin with the same fish species by the same
author shows that the toxicities of these two chemicals to a given fish spe-
cies are generally very similar (Henderson, et al. 1959; Katz, 1961; Macek,
et.al. 1969).
Acute toxicity tests with aldrin and dieldrin have established that
these compounds are toxic to freshwater aquatic life at low concentrations.
Based on species mean acute values summarized in Table 3, the Freshwater
Final Acute Value for dieldrin, derived from the species mean acute values
4
using the procedure described in the Guidelines, is 2.5 vg/1. Similarly, the
Freshwater Final Acute Value for aldrin is 3.0 ug/1.
Saltwater invertebrate species are acutely sensitive to both aldrin and
dieldrin, but there are greater differences in reported LC50 values for
these species than for .saltwater fish species (Table 1). Saltwater inverte-
brate acute values ranged from 0.37 to 33.0 ug/1 for aldrin and from 0.28 to
50 ug/1 for dieldrin (Tables 1 and 6). The most sensitive species to aldrin
in a 96-hour test (Table 1) was Korean shrimp with IC50 values of 0.74 and
3.0 ug/1 (Schoettger, 1970). The commercially important pink shrimp was the
most sensitive species to dieldrin in a 96-hour test (Table 1) with an
LCgg value of 0.7 ug/1 (Parrish, et al. 1973). Other invertebrate species
were less sensitive to dieldrin, and their acute LC50 values ranged from
3.7 to 50 vg/1 (Table 1)..
All species of saltwater fishes tested were sensitive to acute exposures
to aldrin or dieldrin (Table 1). In aldrin exposures, the 96-hour ICgQ
values for 11 fish species ranged from 2.03 ug/1 for dwarf perch (Earnest
B-3
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and Benville, 1972) to 100 yg/1 for striped mullet (Eisler, 1970b). The
acute LCgQ values for 13 fish species exposed«to dieldrin ranged from 0.9
yg/1 for American eel to 34.0 yg/1 for northern puffer (Eisler, 1970b).
Generally, the LCgQ values for aldrin are slightly higher than those for
dieldrin in tests where the same species were tested.
Based on species mean acute values summarized in Table 3, the Saltwater
Final Acute Value for dieldrin is 0.71 vg/1 as calculated according to the
procedure described in the Guidelines; that for aldrin 1s 1.3 yg/1.
Chronic Toxicity
Only one chronic study with a freshwater invertebrate species was
found. Adema (1978) exposed the cladoceran, Daphnia magna, to dieldrin in a
life-cycle test; a chronic value of 57 yg/1 was obtained from his results.
This value was not used in determining final chronic values because no acute
toxicity information for £. magna was available in the literature, and the
"acute-chronic ratio required by the Guidelines could not be calculated.
Two chronic toxicity tests with freshwater fish species have been con-
ducted with dieldrin. One was an early-life-stage exposure using steel head
(rainbow) trout (Chadwick and Shumway, 1969). A chronic value of 0.22 yg/1
was calculated from their data. This was the most sensitive freshwater spe-
cies according to the acute studies (Table 3). Because Chadwick and Shumway
did not provide an" acute value for this species, the species mean acute val-
ue of 2.5 yg/1 is divided by the chronic value of 0.22 yg/1 to give an
acute-chronic ratio of 11 for this species (Table 2). The other chronic ex-
posure was a three-generation study using the guppy (Roelofs, 1971). A
B-4
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chronic value of 0.45 ug/1 was obtained. The geometric mean of 38 96-hour
LCm values (Table 1) using the same source of test water is 4.1 ug/1
5u
(Chadwick and Kiigemagi, 1969); the acute-chronic ratio is 9.1 (Table 2).
No chronic studies were found for any freshwater invertebrate species,
other than Daphnia magna previously discussed. Based on measured concentra-
tions however, Jensen and Gaufin (1966) determined a 30-day LCgg value of
2 ug/1 for the stonefly, Pteronarcys californica, (Table 6) in flowing
water, their typical habitat. This compares to an acute value of 39 ug/1
(Jensen and Gaufin, 1966) from a static test in which concentrations were
not measured. A lower 30-day LC^q value of 0.2 ug/1 was also obtained for
another stonefly, Acroneuria pacifica (Table 6). These data indicate that
some chronic values for larval insects may be lower than those determined
for fishes, which might be expected because the primary use of dieldrin was
as an insecticide.
The only chronic data found for saltwater species was a 28-day life
cycle study on the mysid shrimp with dieldrin (Table 2). In that study
(U.S. EPA, 1980) the chronic limits were 0.49 and 1.1 ug/1 based on cumula-
tive mortality. Effects on reproduction (fecundity) were not observed in
any of the test concentrations. The geometric mean of these two values,
0.73 ug/1, becomes the chronic value for mysid shrimp. Dividing this value
into the acute value for this species of 4.5 ug/1 gives an acute-chronic
ratio of 6.2 (Table'2).
The Final Acute-Chronic Ratio for dieldrin of 8.5 is the geometric mean
of the three acute-chronic ratios (Tables 2 and 3). The Freshwater Final
Acute Value for dieldrin of 2.5 ug/1 divided by the Final Acute-Chronic
Ratio of >8.5 results in the Freshwater Final Chronic Value for dieldrin of
0.29 ug/1. The Saltwater Final Acute Value for dieldrin of 0,71 ug/1
B-5
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divided by the Final Acute-Chronic Ratio of 8.5 results in the Saltwater
Final Chronic Value for dieldrin of 0.084 pg/1. •
Plant Effects
Four toxicity tests have been conducted on dieldrin with three fresh-
water plant species (Table 4). The alga, Scenedeswus quadricaudata. was the
most sensitive species tested with a 22 percent reduction in biomass after
exposure to 100 yg/1 (Stadnyk and Campbell, 1971). The other species, a
diatom and the water-meal, were affected only at concentrations of 128 times
and 100 times higher than that affecting the alga. Because fish and inver-
tebrate species were affected at concentrations over 100 times lower than
that affecting the alga, the plants should be protected by the ani-
mal-derived criteria.
Information on the sensitivity of saltwater aauatic plants, including
algae and phytoplankton {Table 4), indicates, as was true for freshwater
species, that they are much less sensitive than are saltwater fish and in-
vertebrate species. Productivity and growth rates were reduced at concen-
trations of approximately 950 to 1,000 tig/1 in three 4- to 36-hour static
tests using one algal species and mixed population communities (Batterton,
et al. 1971; Butler, 1963).
Residues
Table 5 contains the results of 11 freshwater residue studies with diel-
drin. No comparable aldrin data were found. The 11 studies include plant,
invertebrate, and fish species. The range of the bioconcentration factors
(BCF) is from 128 for an alga (Reinert, 1972) to 68,286 for whole body of
yearling lake trout (Reinert, et al. 1974). All of the authors (except
Reinert, et al. 1974) indicate that a steady-state condition was reached in
their studies.
B-6
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The analysts of the freshwater residue data can be divided into two
broad groups, the plant-Invertebrate and the fish data. The lower plant-in-
vertebrate BCF values range from 128 to 5,558. The two values representing
the algal and diatom community accumulations are perhaps the most ecologi-
cally applicable data 1n this group. The studies were conducted in open
channels under field conditions, whereas the other algal study was a short-
exposure laboratory test. The two 8CF values for Invertebrate species show
a comparatively low b1©accumulation potential.
The BCF values for freshwater fish species range from 2,385 to 68,286.
Although all but one of the authors reported that steady-state had been
reached in each of their exposures, there seems to be a'relationship between
length of exposure and total residue accumulation. For example, guppies ex-
posed for 32 days had a BCF of 12,708, whereas exposure for 160 to 230 days
resulted in a BCF of 28,408. The same relationship may explain the high BCF
for the lake trout. The b1 ©concentration of dieldrin by this species may
become greater since the fish may not have reached steady-state when the
study was terminated. The channel catfish BCF is the lowest, of the values
for fish species (Shannon, 1977a,b). This is probably a result of the ex-
perimenter analyzing dorsal muscle rather than whole fish (with its higher
lipid content) as was done by the others.
Bioconcentration factors for dieldrin and saltwater species (Tables 5
and 6) range from 400 to 8,000 for fish or shellfish (Lane and Livingston,
1970; Epifanio, 1973; Parrish, et al. 1973; Parrish, 1974; Mason and Rowe,
1976). Bioconcentration factors for oysters were higher for longer exposure
periods because dieldrin concentrations in tissues reached steady-state con-
ditions after extended periods (several weeks) of exposure (Parrish, 1974;
Mason and Rowe, 1976). Therefore, long exposures are necessary to attain
B-7
-------
steady-state t>1©concentration factors. After 34 weeks of exposure to diel-
dr1nt sailfin mollies exhibited BCF values of 2,867 to 4,867 in muscle;
values for liver, brain, gill. Intestine, and blood ranged from 10,500 to
50,000 (Lane and Livingston, 1970). Spot exposed to .dieldrin for 35 days
depurated the chemical to non-detectable body-burdens within 13 days of
holding in d1eldrin-free saltwater (Parrish, it al. 1973). Concentrations
In edible tissues were about 15 percent less than concentrations in whole
spot; however, concentrations in liver were 2 to 13 times that in spot
muscle.
Dividing a BCF value by the percent lipid value for the same species
provides a BCF value based on one percent lipid content: this resultant BCF
value is referred to as the normalized BCF. The two BCF values for which
percent lipid data are available (1,160 for freshwater mussel and 2,300 for
spot) (Table 5) were normalized by dividing the BCF values by their corre-
sponding percent lipid values. The geometric mean of the normalized BCF
values was then calculated to be 1,557. The action level established by the
U.S. Food and Drug Administration (FDA) for dieldrin in fish and shellfish
1s 0.3 mg/kg. Dividing the FOA action level of 0.3 mg/kg by the geometric
mean of normalized BCF values (1,557) and by a percent lipid value of 15 for
freshwater species (see Guidelines) gives a freshwater residue value of
0.013 vg/1. Similarly, dividing the FOA action level of 0.3 mg/kg by the
geometric mean of normalized BCF values (1,557) and by a percent lipid value
of 16 for saltwater species (see Guidelines) gives a saltwater residue value
of 0.012 ug/1. The highest BCF value for the edible portion of a consumed
freshwater species is 2,993 for channel catfish (Shannon, 1977a). Dividing
this value into the FDA action level of 0.3 mg/kg gives a freshwater residue
value of 0.10 ug/1. The highest BCF value for the edible portion of a con-
B-8-
-------
sumed saltwater species is the value of 8,000 for Eastern oyster (P&rrish,
1974). Dividing this Into the FDA action level of 0.3 mg/kg gives a salt-
water residue value of 0.038 wg/1.
the U.S. FDA has established an action level of 0.3 mg/kg for dleldrln
1n fish oil. Dividing this value by the geometric mean of normalized BCF
values (1,557) and by a percent lipid value of 100 for fish oil gives a res-
idue value of 0.0019 «g/l for both freshwater and saltwater.
The lowest residue value of those calculated 1s 0.0019 wg/1, and this
value 1s then the Freshwater Final Residue Value and Saltwater Final Residue
Value (Table 5) The Final Residue Value may be too high because, on the
average, the concentration in 50 percent of species similar to those used to
derive the value will exceed the FDA action level.
Miscellaneous
The freshwater data presented in Table 6 do not conflict with data used
to calculate the Freshwater Final Acute and Chronic Value. However, a spe-
cial sensitivity of aquatic insects to dieldrin is reflected in the values
obtained in 30-day exposures of Pteronarcys dorsata and Acroneuria padfl-
ea. With these insects the IC50 values were 2 and 0.2 wg/1, respective-
ly. A 24-hour exposure of the midge, Chironomus tentans, resulted in an
LC5q of 0.9 wg/1. These three values are below the Final Acute Value of
2.5 wg/1, which indicates that some aquatic insects may not be protected by
this value.
For saltwater species, two pink shrimp studies by Lowe (undated)(Table
6) give acute values for aldrin (0.37 wg/1) and dieldrin (0.28 wg/1) that
are lower than any in Table 1. Parrish, et al. (1973) produced an LC50
value of 0.7 wg/1 dieldrin for pink shrimp based on measured values in a
flow-through test; this test should take precedence over that of Lowe, in
B-9
-------
which the test concentrations were not measured. If one can assume that the
relationship between the dieldrin ICgQ values for Korean shrimp (flow-
through test) and pink shrimp (I.e., 6.9 to 0.7 ug/1) would; hold for the
same two species exposed to aldrin, then one would expect the aldrin LCS0
for pink shrimp to be 1/10 that (3.0 wg/1} of Korean shrimp, or 0.3 yg/1.
In fact, a 24-hour EC5Q of 0.37 wg/l has been reported for pink shrimp
(Lowe, undated)(Table 6). Because this test does not meet the criteria in
the Guidelines for an acceptable acute test (the duration was 24 hours), it
was not placed 1n Table 1. However, pink shrimp are commercially valuable
as well as ecologically important, and the Saltwater Final Acute Value may
be too high to protect this important species.
Sumnary
Acute values are available for 19 freshwater fish and invertebrate spe-
cies. The data are all from static exposures in which aldrin and dieldrin
concentrations were calculated but not measured. The species list repre-
sents all of the major functional and taxonomic classifications. The most
resistant fish species is the goldfish at 41 wg/1, and the most sensitive is
the rainbow trout at 2.5 wg/1. A similar comparison for the invertebrate
species shows a range from 5 wg/1 for the isopod, Asellus brevicaudus. to 39
wg/1 for the stonefly, Pteronarcys californica. The Freshwater Final Acute
Value for dieldrin is 2.5 wg/1; that for aldrin is 3.0 ug/1.
The three freshwater chronic values for dieldrin are 0.22, 0.45, and 57
wg/1 for the rainbow trout, guppy, and Daphnia magna, respectively. The
acute-chronic ratios for rainbow trout and guppy are 11 and 9.1,
respectively.
The freshwater residue data for dieldrin show a wide range of bioconcen-
tration factors. The highest factor was for yearling lake trout which may
B-10
-------
not have reached steady-state at a bloconcentratlon factor of 68,286. This
factor may underestimate the b1oconcentration potential of older, larger
lake trout and thus 1s a conservative estimate for this species. The Fresh-
4
water Final Residue Value for dieldrln of 0.0019 yg/1 was calculated using
the FOA action level of 0.3 mg/kg for fish oil, a percent lipid value of 100
for fish oil, and the geometric mean of normalized bloconcentratlon factors
(1,557). The Final Residue Value may be too high because, on the average,
the concentration 1n 50 percent of species similar to those used to derive
the value will exceed the FOA action.
The freshwater plant data clearly Indicate that plants are more resis-
tant than animals. The lowest plant value of 100 yg/1 for 10 days would
certainly destroy most animal life 1n the water.
The acute toxicities of aldrin and dieldrin to saltwater organisms and
the persistence and bioaccumulation potential for dieldrin have been studied
using saltwater plants and animals. Bioaccumulation by saltwater organisms
and/or subsequent transfer to other animals 1n saltwater food-webs have been
documented 1n field studies and laboratory experiments. Results from
>96-hour tests Indicate that dieldrin is chronically toxic, to saltwater
fishes and crabs, although the exact mechanism of toxicity is not known.
The Saltwater Final Acute Value for dieldrin is 0.71 yg/1; that for aldrin
is 1.3 yg/1.
No chronic study on any saltwater fish species has been reported. One
saltwater test on dieldrin using the mysid shrimp, Mysidopsis bahia, pro-
duced a chronic value of 0.73 yg/1, and the acute-chronic ratio for the spe-
cies is 6.2. The Saltwater Final Chronic Value for dieldrin is 0.084 yg/1.
Dieldrin bioconcentration factors for saltwater species range from 400
to 8,000. The Saltwater Final Residue Value of 0.0019 yg/1 was calculated
B—11
-------
using the FDA action level of 0.3 mg/kg for fish oil, a percent lipid value
of 100 for fish oil, and the geometric mean of normalized bloconcentratlon
factors (1,557). The Final Residue Value may be too high because, on the
average, the concentration 1n 50 percent of species similar to those used to
derive the criteria will exceed the FDA action level.
CRITERIA
01eldr1n
For dleldrln, the criterion to protect freshwater aquatic life as
derived using the Guidelines 1s 0.0019 jig/1 as a 24-hour average, and the
concentration should not exceed 2.5 yg/1 at any time.
For dieldrin the criterion to protect saltwater aquatic life as de-
rived using the Guidelines is 0.0019 wg/1 as a 24-hour average, and the con-
centration should not exceed 0.71 wg/1 at any time.
Aldrln
For freshwater aquatic life, the concentration of aldrin should not
exceed 3.0 «g/l at any time. No data are available concerning the chronic
toxicity of aldrin to sensitive freshwater aquatic life.
For saltwater aquatic life, the concentration of aldrin should not
exceed 1.3 wg/1 at any time. No data are available concerning the chronic
toxicity of aldrin to sensitive saltwater aquatic life.
8-12
-------
Tab la 1. Acuta values for aldrtn/dlaldrln
Speclas
Method'
Chaalcal
LC50/ESC30
CM/I)
MmM
Acuta Value
Iwi/ll
Rafaranca
FRESHWATER SPECIES
Cladocaran,
Oaphnla carlnata
S, U
Technical grade
dlaidrln
130
130
Santharaa, at al.
1976
Cladocaran,
Oaphnla putax
S, U
Dlaidrln
m
290
Sandar* A Copa, 1966
Cladocaran,
Slaocaphalus sarrulatus
S, U
Olaidrln
240
-
Sandars i Copa, 1966
Ciadocaran,
Slaocaphalus sarrulatus
S, U
Olaldrln
190
213
Sanders 4 Copa, J3S6
Isopod,
Asallus bravlcaudus
S, U
Olaidrln
5
5
Sandars, 1972
Scud,
Ga«*arus fasctatus
S, U
Olaldrln
640
-
Sandars, 1972
Scud,
6n—arus fasclatus
S, U
Olaidrln
600
620
Sandars, 1972
Scud,
Gamarus laaistrls
S, U
Olaldrln
700
-
Gaufln, at al. 1965
Scud,
6awurus lacustrls
S, U
Olaldrln
460
567
Sandars, 1969
Glass strlap,
Palaaaonatas kadlakensis
S, U
Olaldrln
20
20
Sandars, 1972
Crayfish,
Orconactas nals
S, U
Olaldrln
740
740
Sandars, 1972
Mayfiy,
Ephaaaralla prandls
s, u
Olaldrln
8
8
Gaufln, at al. 1969.
Stonefly,
Acroneurla pad flea
s, u
100* dlaldrln
24
24
Jansan i Gaufln,
1964
-------
Tafcta 1. (Contln—d)
Spiclw
Stonaf ly,
Pt»ronycv» eallfornlca
Rainbow trout,
Sal wo galrdnr I
Rainbow trout, ¦
Sal wo oalrdnrl
Rainbow trout,
Sal wo galrdnarl
Rainbow trout,
Sal— aalrdnar I
Coho salaon,
Oncorhynchus klsutch
Chinook salami,
Oncorhynchus tshawytscha
Goldfish,
Carasslus auratut
Fathead alnrtow,
Pt—phala* pro—las
Fathaad alnnow,
PI—phalas pro—las
Fathaad Minnow,
PI—phalas pro—las
Fathaad Minnow,
PI—p ha I as pro—las
Fathaad alnnow,
PI—phalas pro—las
Fathaad alnnow,
Plaaphalas pro—las
Hatkodf Chwlcal
S, U 100* dlaldrln
S, U 90* dlaldrln
S, U 85)1 dlaldrln
S, U 69| dlaldrln
S, U SS| dlaldrln
S, U 90| dlaldrln
S, U 90* dlaldrln
S, U 90| dlaldrln
S, U 90| dlaldrln
S, 0 90| dlaldrln
S, U 651 dlaldrln
S, U 85* dlaldrln
S, U 85* dlaldrln
S, U 851 dlaldrln
LC50/BC50 Aort. V«l—
lua/il
-------
Table I* (Co*tMi
lESEla.
Fathead ¦Inncw,
Plwpliatw wawlat
Guppy,
Poecl I la reticulata
Guppy,
Poecl Ha rat leu lata
Guppy.
Poecl Ha reticulata
GuPW.
PoeclI la reticulata
Guppy,
PoeclI la ratIcutata
Guppy,
PoeclI la reticulata
Guppy,
Poecllla reticulata
Gwpy#
PoeclI la reticulata
Guppy,
PoeclI la reticulata
Guppy,
PoeclI la reticulata
Guppy,
PoeclI la reticulata
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Chew leaI
SSJt dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
technical grade
dleldrln
Technical (fade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
IC90/EC9O
23
3.9
4.7
3.9
5.1
3.9
3.7
3.2
3.9
4.2
4.3
4.3
4.1
3.5
Spaclw
Acuta Value
22
Reference
Tarxvel I I Henderson,
1957
ChadNlck 4 Kllyagl,
1968
ChadNlck i Kllgaaagl,
1968
ChadNlck i KllgMegl,
1968
ChadNlck 4 Kllgaaagl,
1968
Chadwlck 4 Kl la»inl,
1968
Chadvick 4 KlloMmgl,
1968
ChadNlck 4 Kllgaaagl,
1968
ChadNlck 1 Kllgaaagl,
1968
ChadNlck 4 Kl Igeatagl,
1968
ChadNlck 4 Kllgaaagl,
1968
Chadnlck 4 Kligaaagl
1968
ChadNlck 4 Kllgaaagl,
1968
ChadNlck 4 Kllgaaagl,
1968
-------
Tab!* 1. (Ccetlnued)
Specie*
Guppy,
Poecllla
ratleu lata
G«ppr»
Poecl1 la
raticulata
Guppy,
PoeclI la
reticulata
Guppy,
Poecl Ha
reticulata
Guppy,
PoeclIla
reticulata
Guppy,
PoeclI la
reticulata
Guppy,
PoeclI la
reticulata
Guppy,
PoeclI la
reticulata
Guppy,
PoeclI la
reticulata
Guppy,
Poecllla
reticulata
Guppy,
Poecllla
reticulata
Guppy,
PoeclI la
reticulata
Guppy,
PoeclI la reticulata
Guppy,
PoeclIla
reticulata
Method*
S, U
S, U
S, U
S, U
5, U
S, U
S, U
S, U
S, U
S, U
S, tl
S, U
s, u
s, u
Chealcel
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Technical grade
dleldrln
Specie* Me*a
LC50/EC50 Acute Vale*
(tm/l) Cttq/I) Refareece
4.7 - Chadulck ft Kllgaaagl,
1968
3.2 - Chadwlck i Kllga—jl,
1968
2.9 - Chadwlck ft Klfgaaagl,
1968
2.6 - Chadwlck ft Kllganagl,
1968
2.9 - Chad»lck A Kllgeeegl,
1968.
2.4 - Ctiafolck ft Kllgeaegl,
1968
2.6 - Chadvlck ft KllgaMfll,
1968
2.3 - ChadHlck ft KllgaMgl,
1968
2.7 - Chadniek ft Kllgaoagt,
1968
2.3 - Chadelck ft Kllgaaaal,
1968
2.7 - Chadnlck ft Kllgeaegl,-
1968
2.7 - Chadulck ft Kllgaaegl,
1968
4.8 - Chadwlck ft Kllgaaegl,
1968
6.1 - Chadulck ft Kllgaaegl,
1968
-------
TabU 1, (Continued)
Species
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Gupoy.
Poecllla reticulata
Green sunflsh,
top qui Is cyanel lus
Method* Chenlcal
S, U Technical grade
dleldrln
S, U 99*% dleldrln
S, U 99+* dleldrln
S, 0 99*! dleldrln
S, U 99*| dleldrln
S, U 99*% dleldrln
S, U 99+* dleldrln
S, U 99*! dleldrln
S, U 99*! dleldrln
S, U 99t| dleldrln
S, U 99*! dleldrln
S, U 90! dleldrln
S, U Dleldrln
S, U 85! dleldrln
Species Keen
LC50/EC50 Acute Value
Inn/1) (uo/l) Reference
3.2 - Chadwlck & Kl IgeMogI,
1968
6.6 - Chadwlck A Kllgeiaagl,
1968
9.6 - Chadwlck,!. Kllgamagl,
1968
6.1 - Chadwlck A Kllgetwgl,
1968
7.5 - Chadwlck ft Kl IgeMgl,
1968
10 - Chadwlck ft Kl IgeMgl,
1968
6.6 - Chadwlck ft Kllgeaagl,
1968
6.6 - Chadwlck ft Kl Igeautgl,
1968
6.9 - Chadwlck & Kl IgeMgl,
1968
4.7 - Chadwlck A Kllgeaagl,
1968
7.5 - Chadwlck A Kl IgeMgl
1968
75 - Henderson, et al.
1959
21 4.5 Cairns ft Loos,
1966
6 - Tarzwel I ft Henderson
1957
-------
Trtlt I. (Coatliwad)
60
I
H
CD
Saaclas Hatfiod* Chwlcil
Grmm unflsh, S, U 85| dlaldrln
tapo»ls cyanallus
Graan *u*f Ith, S, II 851 dlaldrln'
lapotls cvaflij las
Bluaglll, S, U 90f dlaldrln
Lwli —crochlrus
Bluaglll, S, U 85* dlaldrln
tapo»ls ¦acrochirus
BluaglII, S, U 65* dleldrln
L dpi It ¦acrochlrus
Bluaglll, S, U 85* dlaldrln
Lapowls ma crochlrus
Bluaglll, S, U 85| dlaldrln
Lappa I * »acrocttlrus
Bluaglll, S, U 85* dlaldrln
Ldo*Is wacrochlrus
Bluaglll. S, U 85)1 dlaldrln
Lapcwls —crochlrus
Bluaglll, S, U 85* dlaldrln
tapcals wacrochlrus
Cladooaran, S, U Aldrln
Daphnla pulax
Cladocaran, S, If Aldrln
Slwocaphalm serrulatus
Cladocaran, S, U Aldrln
Slwocaphalus sarrulatus
Isopod, S, U Aldrln
Asallus bravlcaudus
SmmsImi 1
UC50/B30 Acuta Valw
(m/D liig/il Rafaraaoa
11 - Tarz«al I 4 Handarson,
195?
8 8.1 Terzwal I 4 Handarson,
1957
9 - Handarson, at at.
1999
17 - Macak, at al. 1969
14 - Macak, at al. 1969
8.8 - Macak, at al. 1969
32 - Tariwal I 4 Handarson,
1951
18 - Tarzuall t Handarson,
1957
8 - Tarzwal I t Handarson,
1957
22 15 Tarmal I 4 Handarson,
1957
28 28 Sandars 4 Copa, 1966
23 - Sandars 4 Copa, 1966
32 27 Sandars 4 Copa, 1966
8 8 Sandars, 1972
-------
Table I. (Continued)
Species
Scud,
Gamisrus fasclatus
Scud,
Gammarus
fasclatus
Scud,
Garmnarus
lacustrls
Scud,
Gammarus
lacustrls
Glass shrlap,
Palaemonates kadlakensls
Mayfly,
Ephemeralla grandIs
Stonefly»
Acroneurla paclflca
Stonefly,
Pteronarcys callfornlea
American eel,
Angullla rostrata
Rainbow trout,
Salmo galrdnerl
Rainbow trout,
Salmo galrdnarl
Rainbow trout,
Salmo galrdnarl
Rainbow trout,
Salmo galrdnerl
Coho salmon,
Oncorhynchus klsutch
Method* Chiiil ca I
S, U Aldrln
S, U Aldrln
S, U Aldrln
S, U Aldrln
S, U Aldrln
S, U Aldrln
S, U Aldrln
S, U 93f aldrln
S, U Aldrln
S, U 88.4J aldrln
S, U 95* aldrln
S, U 95* aldrln
S, U 95* aldrln
S, U 88.4* aldrln
Species Mean
LC50/EC50 Acute Value
(pg/l) Cuo/I) Reference
4,300 - Sanders, 1972
5,600 4,900 Sanders, 1972
38,500 - Gaufln, et al. 1965
9,800 19,000 Sanders, 1969
50 50 Sanders, 1972
9 9 Gaufln, et al. 1965
143 143 Jensen & Gaufln, 1964
180 180 Jensen & Gaufln, 1964
16 16 Rehwoldt, et al. 1977
17.7 - Katz, 1961
3.2 - Macek, ®t al. 1969
3.3 - Macek, et al. 1969
2.2 4.5 Macek, et al. 1969
45.9 45.9 Katz, 1961
-------
Table I. (Continued)
Spaclee
Chinook salmon,
Oncorhvnchus tshawytscha
Goldfish,
Carasslus auratus
Corp,
Cyprlnus carpIo
Fathaad alnnow,
Plwephales pro—las
Fathaad Minnow,
Plwephales prone las
8andad kllllflsh,
Fundulus dlaphanus
Guppy,
Poecllla ratleulata
Guppy,
Poacllla ratleuIata
Nhlta parch,
Roccus awarlcanus
Strlpad bass,
Horone saxatllls
Blueglll,
Lapowls eacrochlrus
Blueglll,
Lepowls MBcrochlrus
Blueglll,
Lepowls ¦acrochlrus
Blueglll,
Lepowls wacrochlrus
Method*
S, U
S, U
S, U
S, U
S, U
S, U
s, u
s, u
s, u
s, u
s, u
s, u
s, u
s, u
Ch—leal
68.4? aldrln
68.4? aldrln
Aldrln
68.4? aldrln
68.4? aldrln
Aldrln
88.4? aldrln
Aldrln
Aldrln
Aldrln
88.4? aldrln
95? aldrln
95? aldrln
95? aldrln
Specie* Meee
LC50/BC5O Acute Valwe
(Hfl/I) (w/H Reference
6.1 6.1 Katz, 1961
32 32 Henderson, at al.
1959
4 4 Rehwoldt, at al. 1977
37 - Henderson, et al*
1959
32 34 Henderson, et al.
1959
21 21 Rehwoldt, et al. 1977
37 - Henderson, et al.
1959
20 27 Rehwoldt, et al. 1977
42 42 Rehwoldt, et al. 1977
10 10 Rehwoldt, et al. 1977
15 - Henderson, et al.
1959
7.7 - Hacek, et al. 1969
5.8 - Hacek, et al. 1969
4.6 7.4 Macek, et al. 1969
-------
Tabla 1. (Coatlauad)
Spaclas
Eastern oyster,
Crassoatraa vlrqlnlca
Eastarn oystar,
Crassostraa vlrqlnlca
Mysld shrlap,
Hysldopsls bah I a
Mysld shrlap,
Mysidopsls bah.Is
Sand shrlap,
Cranqon sopt—splnoaa
Haralt crab,
Paqurus lonqlcarpus
w —
to Grass shrlup,
H Palaa«onatas vulaarls
Grass shrlap,
Pataawonatas twalo
Koraan shrimp.
Pa I a won wacrodactylus
Koraan shrlap,
Palaawon wacrodactylus
Pink shrlap,
Panaaiis duorarun
Aaorlcan aal.
Annul I la rostrata
Chinook salaon,
Oneorhynehus tshawytscha
Atlantic siIvarslda,
Hen Id to wenldla
Hathod*
FT, U
FT, M
S, U
FT, H
S, U
S, U
S, U
FT, N
5, U
FT, U
FT, M
S, (I
FT, U
S, U
Spaclaa Naaa
ummm Ac*ta Yalta
Chawlcal (wg/l) Chq/O
SALTWATER SPECIES
Dial dr In 34*«
Ololdrln 31.2«« 31.2
Dlaldrln 3>T
Dialdrln 4.5 4.3
Dlaldrln 7.0 7.0
Dlaldrln 18.0 18.0
Dlaldrln 50.0 90.0
Dlaldrln 8.6 8.6
Dlaldrln 16.9
Dlaldrln 6.9 10.8
Dlaldrln 0.7 0.7
Dlaldrln 0.9 0.9
Dlaldrln 1.5 1.5
Olaldrln 5.0 5.0
Butlar, 1963
Parrlsh, at at. 1973
U.S. EPA, 1980
U.S. 0*A, 1980
Eislar, 1969
Elslar, 1969
Eislar, 1969
Parrlsh, at al. 1973
Schoattgar, 1970
Schoattgar, 1970
Parrlsh, at al. 1973
Eislar, 1970b
Schoattgar, 1970
Eislar, 1970b
-------
Table I* (Continued)
Species
Sheepshead minnow,
Cyprlnodon varleqatus
MuiMlchog,
Fundulus haterocI Itus
Mwmlcbog,
Fundulus heteroclItus
Striped kllllflsh,
fundulus ma Jails
Threesplne stickleback,
Gastwostaus aculeatus
Threasplna stickleback,
Oastarosteus aculeatus
Striped bass,
Horone saxatllls
Shinor perch,
Cvmatogaster aggregate
Shiner perch,
Cywatoqaster aggregate
Dwarf perch,
Mlcrometrus nilnlmus
Dwarf perch,
Hlcrometrus minimus
Bluehaad,
Thalassoma blfasclatum
Striped Millet,
Mugll cephalus
Northern puffer,
Sphaeroldes maculatus
LC50/EC30
10.0
5.0
16.0
5.0
15.3
13.1
19.7
3.7
1.5
5.0
2.44
6.0
23.0
34.0
Species Mean
Acute Value
Ciig/ll
10.0
8.9
5.0
14.2
19.7
2.3
3.5
6.0
23.0
34.0
Reference
Parrlsh, et al. 1973
Elsler, 1970a
Elsler, 1970b
Elsler, 1970b
Katz, 1961
Katz, 1961
Korn 4 Earnest, 1974
Earnest & Benvllle,
1972
Earnest & Benvllle,
1972
Earnest & Benvllle,
1972
Earnest I Benvllle,
1972
Elsler, 1970b
Elsler, 1970b
Elsler, 1970b
-------
Table I. (Continued)
Species Method* ' Ch—leal
Eastern oyster, FT, U Aldrln
Crassostrea vlrglnlca
Sand shrlap, S, U Aldrln
Crangdon septewsplnosa
Hernlt crab, S, (I Aldrln
Pagtirus looglcarpus
Grass shrlep, S, U Aldrln
Palaeoonates vulgaris
Korean air lap* S, II Aldrln
Pa I a won eacrodactylus
Korean shrlep, FT, (I Aldrln
Palacwon wacrodactylus
Aaerlcan eel, S, U Aldrln
Angullla rostrate
Mum*Ichog, S, U Aldrln
FundyIms heteroclltus
Muaalchog, S, U Aldrln
Fundulus heteroclltus
Striped kllllflsh, S, U Aldrln
Fundulus mjalls
Atlantic siIvarslde, S, U Aldrln
Menldla wanldla
Threes pine stickleback, S, II Aldrln
Gasterosteus aculaatut
Threes pine stickleback, S, II Aldrln
Gasterosteus aculeatus
Striped toss, FT, U Aldrln
Morons saxatllls
UC50/BC50
25.0**
8.0
35.0
9,0
Q*
3.0
9.0
8.0
4.0
17.0
13.0
39.8
27.4
7.2
Species Mean
Acute Value
25.0
8.0
33.0
9.0
1.5
5.0
5.6
17.0
13.0
33.0
7.2
Reference
Butler, 1963
Elsler, 1969
Elsler, 1969
Elsler, 1969
Schosfttgsr, 1970
Schoettger, 1970
Elsler, 1970b
Elsler, 1970b
El tier, 1910a
Elsler, 1970b
Elsler, 1970b
Katz, 1961
Katz, 1961
Korn t Earnest, 1974
-------
Tabl* 1. (CoatlMMd)
»
i
K>
t-
Spaclaa Hathod' Ctadcil
Shlnar parch, S, U Aldrln
Cyiaatogaster aggregate
Shiner parch, FT, U Aldrln
Cvwatoqastar aqqraqata
Dwarf parch, S, U Aldrln
Nlcrowatrua minims
Dwarf parch, FT, U Aldrln
Hlcroaietrus ilnlwus
Bluehaad, S, U Aldrln
Thalassona bifasclatua
Strlpad «il1st, S, U Aldrln
Muqll cephalus
Northern puflar, S, I) Aldrln
Sphaoroldes waculatus
* S ¦ static, FT » flow-through, U ¦ unmeasured, M
ssgCS0 based on shell deposition
LC50/EC50
7.4
2.26
18.0
2.03
12.0
100.0
36.0
Specie* Hmhi
Acuta Velee
W)
4.1
6.0
12.0
100.0
36.0
Reference
Earnest & Benvllle,
1972
Earnest i BenvlI la,
1972
Earnest A BanvlIle,
1972
Earnast ft BanvlI la
1972
Elslar, 1970b
Elslar, 1970b
Elslar, 1970b
measured
-------
Table 2. Chronic values for ileidrl*
Species
Cladoceran,
Paphnla enqna
Rainbow trout,
Sal wo galrdnerl
Guppy,
PoeclI la reticulata
Units Chronic Value
Tost* (pg/l) lug/1 ) Reference
FRESHWATER SPECIES
IC 32-100
ELS 0.12-0.39
IC 0.2-1.0
57 Aden. 1978
0.22 Chad wick A StUMway,
1969
0.45 Roelofs, 1971
Mysld shrlep,
Mysldopsls bahIa
LC
SALTWATER SPECIES
i
0.49-1.1
0.73
U.S. EPA, I960
• LC « life cycle or partial life cycle, ELS ¦ early life stage
Acute-Chronic Ratios
Species
Rainbow trout,
Sal wo galrdnarI
Guppy,
Foecllla reticulata
Mysld stirlnp,
Mysldopsls bahIa
Acuta
Value
Cng/»
2.5
4.1
4.5
Chronic
Value
(wg/l) Ratio-
0.22
0.45
0.73
II
9.1
6.2
-------
Tabla 3. Spaclaa mm scut* value* and scuta-chronic ratios for aldrln/dlaldrl*
Rank* Special
Specie* H*m
Acuta Value
(Hi/I)
FRESHWATER SPECIES
Dleldrln
19 Crayfish, 740
Oreonactas nals
18 Scud, 620
Gawmarus fnsclatus
17 Scud, 567
Gawnartis lacustrls
16 Cladoceran, 250
Paphnla pulax
15 Cladocaran, 213
Sliaocephatiis sarrulatua
14 Cladocaran, 130
Paphnla carlnata
13 Goldfish, 41
Carass Ius auratus
12 Stomfly, 39
Ptaronarcys call tornlea
It Stonafly, 24
Acroneurla paclflea
10 Fathead Minnow, 22
Plwephalas prowalas
9 Glass shrlap, 20
Palaewonetas kadlakansls
a Blueglll, 15
Lepowls aacrochlrus
7 Coho salmon, 10.8
Oncorhynchus klsutch
Specie* Mean
Aoita-Chroalc
Ratio
-------
T«bU 3.
Rank*
6
5
4
3
2
I
21
20
19
18
17
16
15
IContlMwd)
SpaclM Nmm SpcclM Nmh
Aorta Valu* Aattfr-Clironlc
Sp«c1— WP Ratio
Gram sunfIsh, 8.1
Lapowli cvanaliui
Mayfly, 8
EphamaraHa or and Is
Chinook salnon, 6.1
Oncorhynchua tshawytscha
l&opod, 5.0
Attllw brevlcaudus
Guppy, 4.5 9.1
PoecHIa rat leu lata
Rainbow trout, 2.5 11
Sain galrdnarl
Aldrln
Scud, 19,000
6a—aru» lacustrls
Scud, 4,900
Ca—arus fosclatui
Stomally, 180
PtTowarcvt cal Ifornlca
StonaHy, 143
Acronaurla pad flea
Glass slrlMp, 50
Palaawowatns kadtakensls
Coho salami, 45.9
Onoorhynchus klsutch
Wilt* perch, 42
Roccus americanus
-------
T«t>l« 3. (CmtlMfWl)
Spaclaa N«m Spaclaa Nmm
text* Vain* Aoita-Chroalc
Rank" Stfcl— <»g/l) Ratio
14 Fathead alnrtow, 34
Plwaphalas prowalas
13 Goldfish, 32
Carasslus auratu*
12 Cladocaraa, 28
Daphnla pulax
II ' Guppy, 27
Poacllla reticulata
10 Cladoceran,, 27
Slwocephalus wrulatus
9 Bandad kllllflsh, 21
Fundulus dlaphanus
8 Aaterlcan esl, 16
w Angullla rostrata
NJ
oo 7 Striped bass, 10
Norone saxatllis
6 Mayfly, 9
Ephemeral la grand Is
5 Isopod, 8
Asa11us bravlcaudus
4 Bluaglll, 7.4
tap owls nacrochlrus
3 Chinook saleon, 6,1
Oncorhynchus tshawytscha
2 Rainbow trout, 4.5
Sal no qalrdnerl
I Carp, 4
Cyprlnus carplo
-------
Table 3. (Continued)
a
NJ
VO
Rank*
21
20
19
18
17
16
15
14
13
12
II
10
9
Suaelas
Species Dim
Acuta Valua
Wl)
Species Nmii
Acute-Chronic
Ratio
SALTWATER SPECIES
Ololdrln
Grass shrimp,
Palaemooetes vulgaris
Northern puffer,
Sphaeroldes waculatus
Eastern oyster,
Crassostrea vlrglnlca
Striped mullet.
Hug11 cephalus
Striped bass,
Horone saxatllls
HeraIt crab,
Pagurus longIcarpus
Threesplne stickleback,
Gastarostaus aculatus
Korean shrimp,
Palaemon ¦aerodactvlus
Stteepshead minnow,
Cyprlnodon varleqatus
Mtumlchog,
fundulus heteroclItus
Grass shrimp,
Palaeroonetes puglo
Sand shrimp,
Crangon septeosplnosa
Bluehead,
Thalassoma bl fasclatun
50.0
34.0
31.2
23.0
19.7
18.0
14.2
10.8
10.0
8.9
8.6
7.0
6.0
-------
Table 3.
Rank"
8
1
6
5
4
3
2
I
16
15
14
13
12
(CofttlMWd)
Species
Striped kllllflsh,
Fundulus ¦a.lalls
Atlantic silver-side,
Henldla —nldla
Mysld (hrlep,
Hysldopsl* bahIa
Dwarf parch,
Hlcrowetrus wlnlins
Shlnar parch,
Cyamatogaster aqqreqata
Chinook salaon,
Oncorhynchus tshawytscha
American sal,
AngulI la rostrata
Pink slrlap,
Penaaus duoraruai
Specie* Naen
taita Value
5.0
5.0
4.5
3.5
2.3
1.5
0.9
0.7
Specie* Haan
Acvte-Ctroelc
Ratio
6.2
Aldrln
Strlpad nil let, 100.0
Wunll cephalus
Northern puffer, 36.0
Sphaeroldes —ctilatus
Hermit crab, 33.0
Pagurus longIcarpus
Three-splnad stickleback, 33.0
Gastarosteus aculeatut
Eastern oyster, 25.0
Crassostrea vlrglnlca
-------
Tab!* 3. (Continued)
i Mlcrometrus minimus
Specie* Newt Species Mean
Acuta Value Acute-Chronic
Rank* . SwcIm
-------
Table 3. (Continued)
Final Acute-Chronic Ratio tor dleldrln ¦ 8.3
Freshwater Final Acuta Valiw for dleldrln » 2.5 pg/l
Frashwatsr Final Chronic Value tor dleldrln » 2.5 tig/1 »• 8.5 ¦ 0.29 jig/1
Saltwater Final Acuta Value for dleldrln " 0.71 |tg/l
Saltwater Final Chronic Value for dleldrln ¦ 0.71 tig/I » 8.5 - 0.064 pg/t
-------
Species
Alga,
Scenedesnus quadrI caudate
Dlato«,
Navlcula SMlnulun
Water-weal,
Wolffia papullfera
Water-eeaI,
Wolffia papullfera
Alga,
AqwenalI urn guadruplleaf tin
Phytoplankton cownunlty
Table 4* Plant values for •Idrln/dleldrln
eh—leal Effect
FRESHWATER SPECIES
Dleldrln 22$ reduction
In blouss In
10 days
Dleldrln 50$ reduction
in growth in
5 days
Dleldrln Reduced popula-
tion growth In
12 «toys
Aldrln Reduced popula-
tion growth In
12 days
Result
(no/1)
100
12,000
10,000
10,000
Reference
Stadiyk i Campbel I,
1971
Cairns, 1968
Worth ley I Schott,
1971
Worth ley 4 Schott,
1971
SALTWATER SPECIES
Dleldrln Reduced growth
ratio
Aldrln 84.6-84.81
decrease In
productivity
after 4 Irs
950 Batterton, et al.
1971
1,000 Butler, 1963
-------
Table 5. Residue* for dfeldrln
Species
Alga,
Scwwlwui obi Iguus
Comunlty dominated by
the alga,
Trlbofi—w» Minus
CoMunlty of alga and
dlatoa* Including
Stlo*oclonliM suDsecundua,
SvnSrla ulna, Epllh-la
sorax. docconals placentula
var. euqlypta, and
NltischTa sp.
Cladoceran,
Daphnla Magna
Freshwater Mussel,
Laepsllls slllguotdea
Steelhead trout
(newly hatched alevln),
Salwo qalrdneri
lake trout (yearling),
SalvetInus nawaycush
Channel catfish,
I eta I ur us punetatus
Channel catfish,
Ictalurus punctatus
Guppy,
Poecllla reticulata
Guppy,
Poecllla reticulata
Tissue
whole body
Dhole an I Ml
MhoIa body
dorsal Muscle
dorsal Muscle
whole anlaaf
whole an Im|
Lipid Blocoeceetratlon
(it) Factor
FRESHWATER SPECIES
128
5,558
3,188
1,395
1,160
3,225
68,286
2,385
2,993
12,708
28,408
Duration
(days)
2,5
4-6 «ks
Reference
Relnert, 1972
Rose & Mclntlre, 1910
4-6 wks Rose & Mclntlre, 1970
3 Relnert, 1972
7-12 Bedford I Zablk, 1973
35 ChadMlck t Shwway,
1969
152 Relnert, et at. 1974
70 Shannon, 1977b "
28 Shannon, 1977a
32 Relnert, 1972
160-230 Roelofs, 1971
-------
Table 5. (ContlnuMi)
Spec!*
Eattorn oyster,
Crassostrea vlrglnlca
Crab,
Leptodlus (lorldanus
Sal If In aolly,
PoeclI la latlplnna
Spot,
Lelostomus xanthurus
Tliww
edible tissue
Mhole body
edible tissue
whole body
Lipid Blocoooentratlon
ill Factor
SALTWATER SPECIES
8,000
400*
4,86?
I.I** 2,300
Duration
(days) Refereaca
392
16
238
35
Parrlsh, 1974
Eplfanlo, 1973
Lane & Livingston,
1970
Parrlsh. at al. 1973
* Converted from dry to wet weight basis
""Data tor J lipid fro* Hansen, 1980
Maxima Permissible Tissue Concentration
Action Level or Effect
Fish and shellfish
Fish oil
Altered ammonia
detoxifying mechanls*
of rainbow trout,
Salwo galrdnerl
Altered phenylalanine
mechanism of rainbow
trout, Saloo galrdnerl
ConcentratIon
(«g/hg)
0.3
0.3
0.36 of
diet
0.36 of
diet
Reference
U.S. FDA Guideline
7420.08, 1978
U.S. FDA Guldal Ine
7426.04, 1977
Mehrle I Btooaflaid,'
1974
Mehrle & OeClue, 1972
-------
Table 3. (Continued)
Goaoetr Ic mean of normalIzad BCF values Ism text) • 1,557
Marketability lor human consumption: FDA action level for fish and shellfish • 0,3 mg/kg
Percent lipid values for freshwater species Isee Guidelines) ¦ 15
. Percent lipid value for saltwater species (see Guidelines) ¦ 16
Freshwater: 0.3 » 0.000013 mn/ka ¦ 0.013 tig/1
1,557 x 15
Saltwater: 0.3 » 0.000012 mg/ka ¦ 0.012 w«j/l
1,557 x 16
Using highest BCF for edible portion of a consumed species
Freshwater: Channel catfish • 2,993 (Shannon, 1977a)
0.3 - 0.00010 mg/kg - 0.10 ug/l
Saltwater: Eastern oyster - 8,000 (Parrlsh, 1974)
0.3 - 0.000038 mg/kg - 0.038 ug/l
8, Mo
FDA action level for fish oil * 0.3 mg/kg
Percent lipid value tor fish oil * 100
Freshwater and Saltwater: 0.3 * 0.0000019 mg/kg * 0.0019 Ufl/I
1,557 x m
Freshwater Final Residue Value * 0.0019 i»g/l
Saltwater Final Residue Value = 0.0019 wg/l
-------
Sb«cIw
Amoeba,
Acanthamoeba casta 11anH
Tublfields (Mixture),
Tublfax and Llwoodrllus
Ostracod,
Cypretta kawatal
Ostracod,
Cypretta kawatal
Aquatic Insects
Stonefly,
Pteronarcy* call fornice
Stonefly,
Acronaurla paclflea
Hldga,
Chlronowua tentans
Rainbow trout,
Sal wo galrdnerl
Rainbow trout,
Sal wo galrdnerl
Rainbow trout,
Salwo galrdnerl
Rainbow trout,
Salmo galrdnerl
Table 6.
Cheilcal
Oleldrln
Oleldrln
Oleldrln
Oleldrln
Dleldrln
Oleldrln
Oleldrln
Oleldrln
Oleldrln
Oleldrln
Oleldrln
Oleldrln
Other data for aldrlri/Jleldrln
Duration Effect
FRESHWATER SPECIES
6 days
96 hrs
24 hrs
72 hrs
6 DOS
30 days
No effect on
survival
LC50
IC50
LC50
Bloconcenfrot Ion
In naturally
exposed anlaals
LC50
Result
Cwfl/U Refer eoce
10,000 Prescott, at al. 1977
6,700 Hhltten ft Goodnight,
1966
185 Hansen ft Kawatskl,
1976
12.3 Hansen ft Kawatskl,
1976
4,620 Bulk ley, at al. 1974
Jensen ft Gauf In. 1966
30 days
24 hrs
17-23
days
140 days
140 days
168 days
LC50
LC50
Lethal Muscle
tissue concentra-
tion 7.7 ag/kg
0.2 Jensen ft Gauf In, 1966
0.9 Karnak ft Collins,
1974
2.3 Hoiden, 1966
Altered conoen- I agAg/"k Mehrle, at al. 1971
tratlons of II
aalno acids
Increased lipid 0.2 agAg/ Macek, at al. 1970
wk
Equltlbrluw bio- 0.2 mg/kg/ Macek, at al. 1970
accuaulatlon of ok
1.05 agAg
-------
Tabl# 6, (ContliuMd)
Spaclas
Corp,
Cyprlnos carpIo
Channel catfish,
Ictalurus punctatus
Ch—leal
Olaldrln
Olaldrln
Black bullhead,
Ictalurus wolas
Mosqultof Ish,
Gawbutla afflnls
Graan sun fish.
Lopowls cyanallus
Graan sunflsh,
Leo owls cyanallus
Walleye,
Stlzostadlon vltraua
Oteldrln
Olaldrln
Dlaldrln
Olaldrln
Dlaldrln
Toad (tadpoles),
Bwto woodhouil
Frog (tadpolas),
Psaudacrls trlserlata
Amoeba,
Acanthaiaoeba castalllanll
Cladocaran,
Daphnla aagna
Mayfly,
Haxaowla billneata
Stonafly,
Ptaronarcys calI tornlea
Olaldrln
Olaldrln
Aldrln
Aldrln
Aldrln
Aldrln
Effact
I00> mortality
of aotryos
Raducad growth
LC50
Rasult
SmOL
5,000
4 |ig/g of
dl«t
fdry wt.)
Rafarewca
Ma lone & Blaytocfc,
1970
Argyla, 1975
2.5 Farguson, at al. 1965
LC50
Concantratlon In 5.65 itg/g
blood at death
Concantratlon In 10.31 |ig/g
brain at death
Behavioral Nar-
rations of yolk
sac fry
LGS0
UC50
12.2
150
too
No affact on 10,000
survival
Bi ©concent rat I or. 14,100
Blocoocantratlon 6,300
LC50 2.5
Cullay I Farguson,
1989
Hogan I Roalofs, 1971
Hogan I Roalofs, 1971
Hair, 1972
Sandars, 1970
Sandart, 1970
Praacott, at al. 1977
Johnson, at al. 1971
Johnson, at al. 1971
Jansan i Gaufln, 1966
-------
Table 6. (Continued)
Swclw
Stonefly,
Aaroneurla paclflea
Midge,
Chlronowus sp.
Corp,
Cyprtnus carplo
Black bullhead,
Ictalurus aw Ias
MosquitoHsh,
Gambusla afflnls
MosqultofIsh,
Gambusla afflnls
Blueglll,
Lepowls wacrochlrus
Toad (tadpoles),
Bufo MoodhouslI
Alga,
Skatetonania costatuw
Alga,
Tatraselels chull
Alga,
Isochrysls oalbana
Alga,
Ollsthodlscus luteus
Alga,
Cvclotella nana •
Ch—leal
Aldrln
Aldrln
Aldrln
Aldrln
Aldrln
Aldrln
Aldrln
Aldrln
Oleldrln
Oleldrln
Oleldrln
Oleldrln
Dleldrln
Duration
50 days
3 days
96 hrs
48 Irs
24 hrs
Effect
Result
(uo/l) Reference
LC50 22 Jensen & GaufIn, 1966
BI ©concentration 4,600 Johnson, et al> 1971
Significant 180 McBrlde & Richards,
increase of sodlua 1975
In perfused gl 11
LC50
LC50
LC50
96 hrs
50* Inhibition
dose of Ma -K
ATPase
L.C50
12.9 Ferguson, et al. 1965
36 Culley & Ferguson,
1969
270 Krleger & Lee, 1973
30 mM Yap, et al. 1973
150 Sanders, 1970
SALTWATER SPECIES
2 hrs Bloooncentratlon
factor - 1,588*
2 hrs BloconcentratIon
factor " 859*
2 hrs Bloconoantrat Ion
factor » 824*
2 hrs Bloconoantrat Ion
factor » 490*
2 hrs . Bloconoantrat Ion
factor ¦ 481*
Rice A Slkka, 1973
Rice 4 Slkka, 1973
Rice 4 Slkka, 1973
Rice t Slkka, 1973
Rice i Slkka, 1973
-------
Table 6. (Continued)
Specie*
Alga,
Aaphldlnliwi carterl
Clew,
Ranqla cwinta
Eastsrn oyster,
Crassostrea vlrglnlca
Eastern oyster,
Crassostrea vlrglnlca
Eastern oyster,
Crassostrea vlrglnlca
Eastern oyster,
Crassostrea vlrglnlca
Pink strtap,
Penaeus duoraruw
Brotm shrlaip,
Penaeus aztecus
Brown shrl*p,
Crangon crangon
Shore crab.
Card bus waenus
Fiddler crab,
Uca pugllator
Ch—lcal
Oleldrln
Oleldrln
Dleldrln
Oleldrln
Oleldrln
Oleldrln
Oleldrln
Dleldrln
Oleldrln
Oleldrln
Dleldrln
Crab larvae,
Leptodlus t lorIdanus
Dleldrln
Crab larvae,
Leptodlus florIdanus
Crab larvae,
Leptodlus florIdanus
Dleldrln
Oleldrln
Duration
2 Irs
72 hrs
7 days
7 days
I day
1 day
2 days
2 days
2 days
2 (toys
15 days
18 days
6 days
16 days
Effect
Bloconcentratlon
factor » 98"
Bloconcentratlon
factor « 1,600 "
Bloconcentratlon
factor ¦ 2,070
Bloconcentratlon
factor » 2,880
EC50
EC50
EC50
EC50
LC50
LC50
Result
W*
Reference
Rice & Slkka, 1973
Petrocelll, et al.
1973
Mason 4 Rowe, 1976
Mason 4 Rons, 1976
13.0 Lowe, undated
240.0 Lowe, undated
0.28 Lowe, undated
3.2 Lowe, undated
>10, <33 Portaann 4 Hilton,
1971
>I0# <33 Portaann & Mllson,
1971
Dleldrln In food O.t Mfl/9 Klein 4 Linear, 1974
affected running
behavior
Bloaccuaulated
after consimlna
food with 213
ug/kg
Approximate LC50
Bloconcentratlon
factor ¦ 7,052
217 ug/9 Epl fan lo, 1973
I Epl fan lo( 1971
Eplfanlo, 1973
-------
Table 6. (Continued)
Species
Blue crab,
Cal1Inactas sap I dm
Blue crab (juvenile),
CalItnoctes sapldus
S heaps ha ad minnow,
CyprInodon varlegatus
Shoepshaad minnow,
CyprInodon varlegatus
Sal if In molly,
Poecllla latlplnna
Sal I tin *olly,
Poecllla latlplnna
Spot,
La I ostonus xanthurus
White Mil let.
Hug 11 curawa
Striped millet.
Mug11 cephalus
Striped Millet,
Mug11 cephalus
Striped Millet,
Muqll cephalus
NInter Hounder,
PseudopIeuronectes awarlcanus
Chewlcat
Oleldrln
Dleldrln
Dleldrln
Oleldrln
Dleldrln
Oleldrln
Dleldrln
Dleldrln
Dleldrln
Dleldrln
Dleldrln
Oleldrln
Pink shrimp,
Panaeus duoraruw
Aldrln
Duration
10 days
2 days
2 days
2 days
2 days
34 wks
1 day
2 days
2 days
2 days
2 days
I day
Effect
Bloaccuawilated 4
to 7 tlaes the
dally dose In food
EC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
1.21 mg/kg In
eaibryos caused
86? reduction In
fertilization
EC50
Result
(uo/l) Reference
PetrocelII, et al„
1975
23.0 Lowe, undated
5.82 Made,' 1969
24.0 Lowe, undated
10.8 Made, 1969
>1.5, <3.0 Lane & Livingston.
1970
3.2 Lowe, undated
7.1 Butler, 1963
3.2 Lows, undated
3.2 Lowe, undated
0.66 Lowe, undated
Smith 1 Cole, 1973
0.37 Lowe,, undated
-------
Table 6* (Continued)
Specie*
Blue crab (Juvenile),
CalHnectes sapldus
Spot,
l»Iosteons xanthurus
White mullet,
Huqll eurewa
Striped aullet,
Hugll cephalus
Ch—teal Duration
Aldrln 2 days
Aldrln 2 days
Aldrln 2 days
Aldrln 2 days
« Corraction factor (0.1) for dry weight analysis
Effect
EC 50
LC50
IC50
LC50
Result
(ua/l) Reference
23 Lone, undated
3.2 Lowe, undated
2.8 Butler, 1963
2.0 Lowe, undated
-------
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3«»4S
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%
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B-46
-------
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B-47
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B-48
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B-49
-------
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*
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B-50
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B-51
-------
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•«
«
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B-52
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Mammalian Toxicology and Human Health Effects
INTRODUCTION
During the past decade, considerable information has been gen-
erated concerning the toxicity and potential carcinogenicity of the
two organochlorine pesticides aldrin and dieldrin. These two pes-
ticides are usually considered together since aldrin is readily
i .
expoxidized to dieldrin in the environment. Both are acutely toxic
to most forms of life including arthropods, mollusks, inverte-
brates, amiphibians, reptiles, fish, birds, and mammals. Dieldrin
is extremely persistent in the environment. By means of bioaccum-
mulation it is concentrated manyfold as it moves up the food chain.
Aldrin and dieldrin are manmade compounds belonging to the
group of cyclodiene insecticides. They are a sub-group of the
chlorinated cyclic hydrocarbon insecticides which include DDT, BHC,
etc. They were manufactured in the United States by Shell Chemical
Company until the U.S. EPA prohibited their manufacture in 1974 (39
FR 37246) under the Federal Insecticide, Fungicide and Rodenticide
Act. They are currently manufactured by Shell Chemical Company in
Holland. Prior to 1974, both insecticides were available in the
United States in various formulations for broad-spectrum insect
control. They were used for control of soil pests and grass-
hoppers, protection of vegetables and fruits, and control of dis-
ease vectors including locusts and termites (Int. Agency Res. Canr
cer, 1974a,b). In 1974, the U.S. EPA restricted the use of al-
drin/dieldrin to termite control by direct soil injection and non-
food seed and plant treatment.
C-l
-------
Early work by Treon and Cleveland (1955) suggested that aldrin
and dieldrin may have tumor-inducing potential, especially in the
liver. Since that time, several conflicting reports of the hepato-
carcinogenicity in mice, rats, and dogs have appeared in litera-
ture. Studies have been carried out mainly by the U.S. Food and
Drug Administration, the National Cancer Institute (NCI), and by
the manufacturer, Shell Chemical Company. There has been much
debate over the type and significance of hepatic damage caused by
aldrin and dieldrin. In order to ascertain the human risks asso-
ciated with aldrin and dieldrin, evaluations of the toxic effects
of these pesticides have been carried out on workers in the Shell
Chemical Company. The evaluations include epidemiological studies
in addition to the more routine toxicity studies. However, it is
felt that the number of workers with high exposures was too small
and the time interval too short to determine whether or not aldrin
and dieldrin represent a cancer threat to humans.
The objective of this report is to examine published studies
so as to utilize the most relevant data to develop a criterion for
human risk assessment.
EXPOSURE
Exposure to aldrin and dieldrin is from contaminated waters,
food products, and air. Because of its persistence, dieldrin has
become widespread in the aquatic environment. It is also spread
great distances by wind. Since aldrin and dieldrin are used
throughout much of the world beyond the United States, it must be
assumed that imported food stuffs, such as meat products, contain
residues of these pesticides.
C-2
-------
Use of aldrin and dieldrin peaked at 19.3 million lbs. in
1966, and 3.6 million in 1956, respectively (39 FR 37251). The
¦e
subsequent decline in dieldrin use vas due, in part, to increased
resistance of boll weevils to chlorinated insecticides (Table 1).
The use of dieldrin vas preferred to aldrin because it required
less application due to its persistence.
Ingestion from Water
Aldrin and dieldrin have been applied to vast areas of agri-
cultural land and aquatic areas in the United States and in most
parts of the world. These pesticides have therefore found their
way into most fresh and marine waters. Unlike DDT, aldrin and
dieldrin are somewhat more soluble in water (27 and 186 mg/1, re-
spectively) (Park and Bruce, 1968). Gunther, et al. (1968) report-
ed dieldrin to be slightly more soluble at 250 mg/1.
In early studies (Weaver, et al. 1965), dieldrin was found in
all major river basins (mean concentration 7.5 ng/1) in the United
States and it was found more often than any other pesticide. It was
also found in the Mississippi delta (U.S. Dep. Agric. , 1966) at
10.0 ng/1 while aldrin was found as high as 30 ng/1. Marigold and
Schulze (1969) reported aldrin and dieldrin at 40 and 70 ng/1,
respectively, in streams in the western United States. Leichten-
berg, et al. (1970) found levels of dieldrin and aldrin as high as
114 and 407 ng/1, respectively, in surface waters in the United
States.
More recently, dieldrin has been reported to be present in
many fresh waters in the United States with mean concentrations
C-3
-------
TABLE 1
Domestic Sales of Aidsin and DieMrin
From 1950 Through July 1, 1974
Year
Aidrin
(1,000 lbs)
Dieldrin
(1,000 lbs)
1950
1,456
0
1951
3,238
185
1952
814
750
1953
1,234
1,135
1954
2,993
1,777
1955
4,372
2,585
1956
6,495
8,635
1957
2,431
2,673
1958
4,971
3,074
1959
5,566
3,008
1960
8,109
2,650
1961
9,926
2,764
1962
10,886
2,990
1963
12,152
2,685
1964
12,693
2,052
1965
14,278
1,814
1966
19,327
1,908
1967
18,092
1,478
1968
13,690
1,332
1969
9,902
1,206
1970
8,909
749
1971
11,615
705
1972
11,868
740
1973
(to July 1)
8,721
432
1973
estimated (to Dec. 31)
(10,000)
(576)
1973
9,900
------
1974
(to July 1)
9,700
Source: 39 PR, 1974
C-4
-------
ranging from 5 to 395 ng/1 in surface water and from 1 to 7 ug/1 in
drinking water (Epstein, 1976).
In 1975 a survey in the United States of aldrin, dieldrin,
DDT, and DDT metabolite levels in raw and drinking water was car-
ried out (U.S. EPA, 1976). Dieldrin was found in 117 of 715 samples
analyzed (Table 2). The six samples in the highest range were all
taken from the same location, three from raw waters and three from
finished waters. Three of these six samples also contained aldrin
in concentration of 15 to 18 ng/1.
Harris, et al. (1977) summarized the distribution of various
chemicals in drinking water in several cities in the United States.
Dieldrin was found in concentrations of 1 ng/1 in Seattle, Washing-
ton, and Cincinnati, Ohio; 2 ng/1 in Miami, Florida, and Ottumwa,
Iowa; and as high as 50 ng/1 in New Orleans, Louisiana.
It has been estimated (MacKay and Wolkoff, 1973) that unlike
many chlorinated hydrocarbons that evaporate rapidly from shallov
waters, dieldrin has by far the longest half-life of these com-
pounds in water 1 meter in depth. They calculated that the half-
life for aldrin and dieldrin would be 10.1 days and 723 days, re-
spectively, compared to 3.5 days for DDT and 289 days for lindane.
This long half-life in water combined with the potential for bio-
concentration by aquatic organisms such as microorganisms, phyto-
plankton, mollusks, and fish further enhances the hazard of these
two pesticides (Wurster, 1971).
Ingestion from Food
Although aldrin is readily converted to dieldrin, dieldrin
itself is stable and persistent in the environment. Because it is
C-5
-------
TABLE 2
*
Dieldrin Concentrations in Raw and Drinking Hater
0
1
0»
No. of Samples 598 94 13 4 6
ng/1 4 4-10 11-20 21-29 56-110
* Source: U.S. EPA,. 1976
-------
lipophilic,-dieldrin accumulates in the food chain (Wurster, 1971).
The persistence of aldrin and dieldrin in different soils varies
with the type of soil and with movement to other areas by water,
wind, etc. (Matsumura and Boush, 1967). Dieldrin has been shown to
be one of the most persistent of all the organochlorine pesticides
(Nash and Woolson, 1967).
It has been estimated that 99.5 percent of all human beings in
the United States have dieldrin residues in their tissues (U.S.
EPA, 1971). Although there are other origins of contamination,
these residue levels are mainly due to contamination of foods of
animal origin (Wurster, 1971). The levels of aldrin/dieldrin in
several types of food have been summarized by Edwards (1973),
Matsumura (1974), and Manske and Johnson (1975). The overall con-
centration of dieldrin in the diet in the United States has been
calculated to be approximately 43 ng/g of food consumed (Epstein,
1976). Table 3 lists the estimated daily dietary intake for aldrin
and dieldrin of a late teen-aged male (National Academy of Sciences
(NAS), 1975).
A bioconcentration factor (BCF) relates the concentration of a
chemical in aquatic animals to the concentration in the water in
which they live. The steady-state BCFs for a lipid-soluble com-
pound in the tissues of various aquatic animals seem to be propor-
tional to the percent lipid in the tissue. Thus, the per capita
ingestion of a lipid-soluble chemical can be estimated from the per
capita consumption of fish and shellfish, the weighted average per-
cent lipids of consumed fish and shellfish, and a steady-state BCF
for the chemical.
-------
TABLE 3
Estimated Daily Dietary Intake (mg) of a Young Hale*
1965
1966
1967
1968
1969
1970
Aldrin
0.001
0.002
0.001
trace
trace
trace
Dieldrin
0.005
0.007
0.001
0.004
0.005
0.005
•Sources MAS, 1975
-------
Data from a recent survey on £ish and shellfish consumption in
the United States were analyzed by SRI International (U.S. EPA,
1980). These data were used to estimate that; the per capita con-
sumption of freshwater and estuarine fish and shellfish in the
United States is 6.5 g/day (Stephan, 1980). In addition, these
. data were used with data on the fat content of the edible portion of
the same species to estimate that the weighted average percent
lipids for consumed freshwater and estuarine fish and shellfish is
3.0 percent.
Two laboratory studies, in which percent lipids and a steady-
state BCF were measured, have been conducted on dieldrin. The mean
i of the BCF values, after normalization to one percent lipids, is
1,557 (see Table 5 in Section B). An adjustment factor of 3 can be
used to adjust the mean normalized BCF to the 3.0 percent lipids
that is the weighted average for consumed fish and shellfish.
Thus, the weighted average bioconcentration factor for dieldrin and
the edible portion of all freshwater and estuarine aquatic organ-
isms consumed by Americans is calculated to be 4,670.
No useful measured bioconcentration factor can be obtained for
aldrin because it is rapidly converted to dieldrin by aquatic or-
ganisms. In addition, because aldrin is converted to dieldrin in
soil, aquatic organisms are rarely exposed to aldrin.
However, the equation "Log BCF » (0.85.Log P) - 0.70" can be
used (Veith, et al. 1979) to estimate the BCF for aquatic organisms
that contain about 7.6 percent lipids (Veith, 1980) from the
octano]<-water partition coefficient (P). Based on a measured
log P value of 3.01 (Hansch and Leo, 1979), the steady-state
C-9
-------
bioconcentration factor for aldrin is estimated to be 72. -An
adjustment factor of 3.0/7.6 «0.395 can be used to adjust the es-
timated BCF from the 7.6 percent lipids or Which the equation is
based to the 3.0 percent lipids that is the weighted average for
consumed fish .and shellfish. Thus, the weighted average bioconcen-
tration factor for aldrin and the edible portion of all freshwater
and estuarine aquatic organisms consumed by Americans is calculated
to be 72 x 0.395 ¦ 28.
Inhalation
Aldrin and dieldrin' enter the air through various mechanisms
such as spraying, wind action, water evaporation, and adhesion to
particulates. Stanley, et al. (1971) reported levels of aldrin and
dieldrin in air samples in nine cities in the United States. One
sample of the air in Iowa City, Iowa had detectable levels of al-
drin (8.0 ng/m ), and 50 samples taken in Orlando, Florida had de-
3
tectable amounts of dieldrin, the largest being 29.7 ng/m . Vari-
ous other studies of the air carried out during the 1960's were
summarized by Edwards (1973).
In a study conducted by the O.S. EPA from 1970 to 1972
(Epstein, 1976), dieldrin was found in more than 85 percent of the
air samples tested. The mean levels ranged from 1 to 2.8 ng/m^.
From these levels, the average daily intake of dieldrin by respira-
tion was calculated to be 0.035 to 0.098 yg;
Although aldrin/dieldrin are no longer used in the United
States, there is still the possibility of air borne contamination
from other parts of the world. Edwards (1973) showed that dieldrin
has been transported long distances in the air. Exposure due to
C-10
-------
inhalation o£ aldrin and dieldrin from the application of these
pesticides was, of course, much greater before the festriction of
their use. Pesticide applicators and individuals living, near agri-
cultural areas were exposed to aldrin/dieldrin through inhalation.
In a recent report, Domanski, et al. (1977) reported no in-
crease in dieldrin concentration in adipose tissue of cigarette
smokers as compared to nonsmokers although tobacco has high resi-
dues of pesticides and is stored many years before use.
Dermal
Dermal exposure to aldrin or dieldrin is limited to those
involved in manufacturing or application of these pesticides.
Wolfe, et al. (1972) reported that exposure to workers, both manu-
facturers and applicators, was mainly through dermal absorption
rather than from inhalation. Due to the ban on manufacturing of
the pesticides in the United States, the possibilities of dermal
exposure have been greatly reduced.
PHARMACOKINETICS
Absorption
Heath and Vandekar (1964), using ^^Cl-dieldrin (4 percent in
arachis oil) showed that absorption by the upper part of the gas-
troinestinal tract begins almost immediately after oral administra-
tion in rats and that the absorption varies with the solvent used.
Barnes and Heath (1964) demonstrated that the LDgQ varies with the
dieldrin-to-solvent ratio. Heath and Vandekar (1964) also demon-
strated that absorption is by the portal vein and not the thoracic
lymph duct. Initially, dieldrin is widespread but within a few
hours it is redistributed in favor of the fat. They also stated
C-ll
-------
that following .oral treatment at 25 mg/kg, "®°Cl-dieldrin could be
recovered from the stomach, small intestine, large intestine, and
feces after 1 hour. :
Distribution
It is well known that dieldrin has a low solubility in water
and a high solubility in fat. At 1 and 2 hours after treatment.
Heath and Vandekar (1964) detected the highest concentration of
36Cl-dieldrin in fat tissue. They also reported high concentra-
tions in the liver and kidney with moderate concentrations in the
brain at these times.
Deichmann, et al.(1968} studied the retention of dieldrin in
blood, liver, and fat. Female Osborne-Mendel rats were fed a diet
containing 50 mg/kg dieldrin (87 percent purity). The rats were
killed on various days of feeding up to 183 days. The concentra-
tion of dieldrin in the blood and liver increased for nine days and
then leveled off until the end of the six-month period. The con-
centration of dieldrin in the fat took approximately 16 days to
reach a level that was maintained throughout the experiment. The
fat had the highest concentrations of dieldrin followed by the
liver. The mean concentration in the fat was 474 times that in the
blood, while the concentration in the liver was approximately 29
times the blood concentration.
Walker, et al. (1969) studied the distribution of dieldrin in
rats and dogs over a two-year period. Dieldrin (99 percent purity)
was incorporated into the diet of CFE male and female rats at
0.1, 1.0, and 10 mg/kg and was fed to dogs in gelatin capsules at
concentrations equivalent to 0.1 and 1.0 mg/kg of their daily
C-12
-------
dietary intake. The authors measured the dieldrin residues in
whole blood, fat, liver, and brain and found significantly in-
*
creased concentrations in all tissues compared'to those in the con-
trols (Table 4).
The concentrations in the tissues increased with an increase
in the dietary concentrations, and the concentrations in the female
rats were considerably higher than those in the males. The diel-
drin concentrations reached a plateau by the end of the 6th month
and remained fairly constant for the remaining 18 months..
In dogs, the blood concentrations increased in both treatment
groups during the first 12 weeks. With the higher dose (1.0
mg/kg/diet) the concentration leveled off between 18 and 30 weeks
of treatment. However, with the lower dose (0.1 mg/kg/diet) the
plateau was reached between 12 and 18 weeks. In the group receiv-
ing 1.0 mg/kg/diet the dieldrin concentration in the blood in-
creased significantly during the final 6 weeks of exposure. The
dieldrin concentrations in the liver and brain were also dose-
related but, as opposed to the results from the rats, showed no
i
significant sex differences. As in other studies, the concentra-
tion in the fat was much greater than that in the liver, which in
turn, was greater than in the brain.
Additional studies on the distribution of dieldrin were car-
ried out by Robinson, et al. (1969). In this study Carworth rats
were fed dieldrin (99+ percent purity) at 10 mg/kg of their diet
for 8 weeks. At the end of this time, they were returned to a diel-
drin-free diet and killed randomly in pairs up to 12 weeks after
withdrawal of the dieldrin diet. The fatty tissue clearly had the
C-13
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TABLE 4
Mean Geometric Dieldrin Concentration (ug/g)
in Rat Tissues after 104 weeks*
Dietary
Level (mg/kg)
Blood
Pat
Liver
Brain
Males
0
0.0009
0.0598
0.0059
0.0020
0.1
0.0021
0.02594
0.0159
0.0069
1.0
0.0312
1.493
0.01552
0.1040
10.0
0.1472
19.72
1.476
0.4319
Females
0
0.0015
0.3112
0.0112
0.0077
0.1
0.0065
0.8974
0.0348
0.0224
1.0
0.0861
13.90
0.4295
0.2891
10.0
0.3954
57.81
2.965
1.130
•Sources Walker, et al. 1969
-------
highest concentration of dieldrin followed by the liver, brain, and
blood. Concentrations of dieldrin in fat returned to control lev-
els after 12 weeks and the decline in dieldrin concentrations was
'approximately exponential in nature.
Matthewsr et al. (1971) investigated the distribution of diel-
drin and some of its metabolites in several organs and tissues of
both male and female Charles River rats. Three animals of each sex
were fasted for eight hours and then given 3 g of food containing
14
10 mg/kg C-dieldrin (96 percent purity). The animals were killed
after nine days, and'dieldrin and metabolic product concentrations
were determined. In general, the amount of radioactivity per gram
was higher for the female rats. The kidneys and stomachs of the
males contained more radioactivity than those of the females. Lev-
els in the lungs and intestines showed similar differences. The
other organs and tissues of the females had three to four times the
radioactivity of the males. In the females, storage was predomi-
nantly as dieldrin, but in males other metabolites, identified as
keto dieldrin, and trans-dihydro-aldrin, and a polar metabolite
were detected in various tissues.
Hayes (1974) determined the concentration of dieldrin in the
fat, liver, kidney-, brain, muscle, and plasma following a single
oral dose in rats. Male Sprague-Dawley rats were given 10 mg/kg
dieldrin (86 percent purity) by stomach tube. The animals were
killed at various intervals up to 240 hours and the dieldrin con-
centration in the tissues was determined. The concentrations in
the brain at 4 and 16 hours were 1.5 and 1.0 ug/g, respectively.
Hayes assigned a value of one to the concentrations in the brain
C-15
-------
and calculated the ratio of the concentrations in other tissues to
the concentrations in the brain at 4 and 16 hours (Table S). The
concentrations in the tissues remained relatively constant for 24
hours and began to decline at 48 hours. Mo further samples were
taken until 240 hours when all the dieldrin concentrations were
below 0.2 ug/g except the concentration in the fat which was
5 pg/g.
In a study done in 1963 on 30 individuals from three different
statesr the concentrations of chlorinated hydrocarbon pesticides in
body fat were determined (Dale and Quinby, 1963). Twenty-eight
individuals were from the general population while one had previous
DDT exposure and one had aldrin exposure. The mean (—SE) for the
general population was 0.15 — 0.02 ug/g dieldrin, while that for
the aldrin-exposed individual was 0.36 ug/g dieldrin (see discus-
sion on aldrin metabolism to dieldrin in the Metabolism section of
this report).
In a study of aldrin and dieldrin concentrations in 71 workers
involved in pesticide manufacturing, Hayes and Curley (1968) mea-
sured the plasma; fat, and urine concentrations by gas-liquid
chromatography. Their findings were in accordance with the earlier
animal studies. The fat contained the highest concentration of the
pesticides followed by the urine and plasma. The mean concentra-
tions of dieldrin in the fat, urine, and plasma of the pesticide
workers were 5.67 + 1.11, 0.0242 + 0.0063, and 0.0185 + 0.0019
ug/g* respectively. These were significantly different from those
reported for the general population. The authors reported a high
correlation between total hours or intensity of exposure and
C-16
-------
TABLE 5
Concentrations of Dieldrln in Tissues of Rats*
(Single Oral Dose)
Hr. Brain Muscle Liver Kidney Plasma Fat
4 1.00 + 0 0.62 ± 0.05 2.30 + 0.11 1.55 + 0.22 0.20 + 0.02 7.20 + 1.18
16 1.00 + 0 0.55 + 0.06 3.17 ± 0.25 2.02 + 0.56 1.35 + 1.11 17.96 + 3.23
•Sources Hayes, 1974
-------
concentration of dieldrin. However, no correlation cof-uld be found
between dieldrin concentrations and amount of sick leave.
Another study (Hunter, .et al. 1969) involving adult males
ingesting 10, SO, or 211 jig dieldrin per day*for 18 or 24 months,
again found a relationship between the dose and the length of expo-
sure and concentration of dieldrin in the fat and blood. In gen-
eral, the concentration of dieldrin in the samples increased during
the first 18 months and either leveled off or rose slightly during
the remaining time. The control and 10 ug groups, both of which
were given 211 pg/day for the final 6 months, demonstrated a rise
in concentrations similar to the rise demonstrated by those who
were given 211 *ig/day initially. The authors stated that there was
no effect on the general health of the individuals receiving the
dieldrin for the two-year test.
In the previously-mentioned studies, blood concentrations of
aldrin or dieldrin were determined using whole blood (Deichmann, et
al. 1968; Robinson, et al. 1969? Hunter, et al. 1969? Walker, et
al. 1969), or plasma (Hayes and Curley, 1968). Mick, et al. (1971)
measured the aldrin and dieldrin concentrations in erythrocytes,
plasma, and the alpha- and beta-1ipoprotein fractions of the blood
of six aldrin workers after the workers had formulated 2 million
pounds of aldrin over a five-week period. The six workers were
exposed to aldrin by both inhalation and dermal contact. The blood
samples were collected at the conclusion of the five-week exposure
and blood plasma concentrations as high as 312 ng/1 were measured.
No immediate health problems were reported during this time. In
all cases, dieldrin concentrations were higher than the aldrin
C-18
-------
concentrations due to the epoxidation of aldrin to dieldrin. The
dieldrin residue in the plasma averaged approximately four times
higher than that in the erythrocytes. As the dieldrin residue in
« •
the blood increased, the amount in the plasma became proportionally
higher. In addition, the beta-lipoprotein fraction usually con-
tained more dieldrin than the alpha fraction.
The work of Mick, et al. (1971) was confirmed in part by
Skalsky and Guthrie (1978). Using labelled pesticides of 98 per-
cent purity incubated with various fractions of human blood in
vitro Skalsky and Guthrie (1978) were able to demonstrate that
dieldrin and DDT bind to albumin and beta-lipoprotein.
Metabolism
Aldrin and its epoxidation product, dieldrin, are both cyclo-
pentadiene insecticides. Since epoxidation of aldrin to dieldrin
was first reported by Radomski and Davidow (1953), there have been
many reports in the literature of the ability of various organisms
(i.e., soil microorganisms, plants, fish, and animals, including
man) to epoxidize this type of double bond. Winteringham and
Barnes (1955) first reported this reaction with aldrin in mice.
Wong and Terriere (1965) were able to demonstrate the in vitro con-
version of aldrin to its epoxide, dieldrin, using microsomes^* from
male and female rats. The reaction was NADPH-dependent and the
enzymes were heat-labile. Winteringham and Barnes (1955) also
showed that males converted aldrin to dieldrin at a higher rate.
No other metabolic products were detected, although the authors
*In this document microsomes refers to the cell-free homogenized
liver (including soluble enzymes and microsomes) and not to puri-
fied microsomes.
C-19
-------
noted that polar products could have been overlooked by the methods
used. Nakatsugawa, et al. (1965) confirmed the work of Hong and
Terrier* (1965) using microsomes from male rats and rabbits. They
also demonstrated a requirement for NADPH and Stated that dieldrin
was not further metabolized by the microsomes. They reported that
lung homogenate was only one-tenth as active as liver in epoxidase
\
activity and that no activity was detected in the kidney, spleen,
pancreas, heart, or brain.
Korte (1963) identified one of the metabolic products of al-
drin as aldrin diol in studies with rabbits. Heath and Vandekar
(1964) reported the existence of a somewhat polar metabolite which
is excreted in the feces. They stated that the feces are the main
route of excretion and that little dieldrin is excreted unchanged.
They were able to detect other polar metabolites in both urine and
feces.
1 i
Ludwig, et al. (1964) administered C-aldrin to male rats at
4.3 ug/day for three months. The compounds excreted into the urine
consisted of aldrin, dieldrin, and unidentified hydrophilic meta-
bolic products. These unidentified products made up 75 percent of
the dose excreted in the feces and 95 percent excreted in the
urine. Two different products were found in the feces and two in
the urine. Two of these four products appeared to be identical by
paper and thin-layer chromatography.
Korte and Arent (1965) isolated six urinary metabolites from
•I M
rabbits treated orally with C-dieldrin for 21 weeks. The major
metabolite (86 percent) was one of the two enantiomorphic isomers
of 6,7-trans-dihydroxy-dihydro-aldrin.
e-20
-------
Richardson, et al. (1968) were able to identify two metabo-
lites in urine and £eces from male CF rats fed a diet containing 100
mg/kg dieldrin for seven months. Metabolites.-were isolated from
the urine and feces collected during the last month. They deter-
mined that the urinary metabolite had a keto group on the number 12
carbon and the epoxide was unchanged. The fecal metabolite was a
mono-hydroxyderivative of dieldrin at either the 4a or 4 position.
A similar study was carried out (Matthews and Matsumura, 1969) in
which male rats were fed a diet of 20 mg/kg purified dieldrin for
one month, with the dosage increased to 100 mg/kg for 18 days while
the urine and feces were collected. Two metabolites were isolated
from the feces and two from the urine. The major fecal metabolite
was similar to the mono-hydroxy-derivative isolated by Richardson,
et al. (1968) in the feces. The major urinary metabolite was iden-
tical to the ketone compound identified by Richardson, et al.
(1968) in the urine. The minor urinary and fecal metabolites were
identical and similar to the 6,7-trans-dihydroxy-dihydro-aldrin
described by Korte and Arent (1965).
Matthews and Matsumura (1969) also conducted _iji vitro experi-
14
ments using C-dieldrin incubated with rat liver microsomes and
various co-factors. Thin-rlayer chromatography of the water-soluble
components isolated six metabolites in addition to the unchanged
dieldrin. Analysis of the water-soluble metabolites revealed a
glucuronide conjugate which accounted for approximately 45 percent
of the radioactivity. Comparison of the Rj, values for the iri vivo
and in vitro studies showed that the minor urinary/fecal metabolite
(i.e., the 6,7-trans-dihydroxy-dihydro-aldrin) was produced in
C-21
-------
vitro and that the metabolite freed from the glucuronic acid was
also present in the in vitro system in the unconjugated form.
The products identified by Richardson, et al. (1968) and
• ~
Matthews and Matsumura (1969) represent an oxidized form of diel-
drin in the urine and feces and a hydroxylated dechlorination meta-
bolite which had lost the intact dieldrin ring system.
Hedde, et al. (1970) were able to isolate six metabolic pro-
ducts in the urine of sheep dosed with C-dieldrin. Three cas-
trated sheep were given unlabeled dieldrin orally at 2 mg/kg/day
14
, for five days before dosing with C-dieldrin. Four other sheep
were fed a single oral dose of labeled dieldrin at 20 mg/kg/day.
Urine and feces were collected up to six days after treatment with
the labeled dieldrin. Although other determinations were made,
only the urine was analyzed quantitatively. After hexane extrac-
tion at pH 1 followed by other clean-up procedures, the four
hexane-soluble metabolites were separated on Sephadex LH-20 gel.
The LH-20 was again used to separate the two water soluble metabol-
ites after they were purified by several procedures, including
paper chromatography. The authors postulated that these water-
soluble metabolites were a glucuronic acid conjugate of the trans-
diol and an unidentified conjugate of glucuronic acid and, possi-
bly, glycine.
Feil, et al. (1970) were able to identify two of the hexane-
soluble metabolites found by Hedde, et al. (1970) in sheep urine.
One was the 6,7-trans-dihydroxy-dihydro-aldrin described by
Richardson, et al (1968) and the other was the 9-mono-hydroxy-
derivative. Further work on the metabolism of dieldrin (Matthews,
C-22
-------
et al. (1971) is discussed in the Distribution .section of this re-
port where details of treatment are given. Matthews, et al. docu-
mented the production of several metabolite^, of dieldrin including
*
the 6,7«*t r ans-dihy droxy-dihy dro-aidrin and a second unidentified
polar metabolite excreted in the feces. The mono-hydroxylated com-
pound represented the greatest percentage of the radioactivity
extracted from the feces of both male and female rats. In. male
rats, the chloroform extract of the urine consisted of the keto-
metabolite described by Klein, et al. (1968). Also, initially,
trans-dihydroxy-dihydro-aldrin was found in the urine of the male
rats along with unchanged dieldrin. Host of the radioactivity
extracted from the urine of the female rats was in the form of the
trans-dihydroxy-dihydro-aldrin, and initially contained up to 20
percent dieldrin.
The metabolism and excretion of dieldrin *$peaxs to be more
rapid in male than in female rats. Investigators attribute this to
the males1 ability to produce the more polar metabolites, especial-
ly the keto-product which is excreted into the urine.
A recent paper has appeared on the comparative metabolism of
dieldrin in rodents. Baldwin, et al. (1972) treated a male CFE rat
with 3 mg/kg of 14C-labelled dieldrin and two male CF1 mice with 10
mg/kg. The urine and feces were collected for tlie following seven
.or eight days. The authors reported that the CFE rat excreted the
pentachloroketone derivative in the urine but that the CF1 mice did
not. Conversely, the mice produced an unidentified urinary metabo-
lite which the rat did not. The 6,7-trans-dihydroxy-dihydro-aldrin
C-23
-------
was found in the feces of the mice and the rat, and a dicarboxylic
derivative was found in the urine of all three animals.
A review of the literature on the metabolism of dieldrin and
i
endrin in rodents has been compiled by Bedford and Hutson (1976).
They summarized the four known metabolic products of dieldrin as
the 6,7-traris-dihydroxy-dihydro-aldrin (trans-diol) and the tri-
cyclic dicarboxylic acid (both of which are products of the trans-
formation of the epoxy group), the syn-12-hydroxy-dieldrin (a mono-
hydro-derivative), and the pentachloroketone.
In comparing dieldrin metabolism in acute or short-term stud-
ies versus chronic, low-dose exposure, it must be mentioned that
organochlorine compounds, including dieldrin, have been shown to
induce the mixed function oxidases (MFO) found in the liver (Kohli,
et al. 1977). It is therefore possible, in the long-term animal
studies, that investigators have been observing the results of high
levels of these enzymes and that the percentages and amounts of
certain metabolites may be misleading. Baldwin, et al. (1972) in a
limited study, were able to show some inducibility in the CFE male
rat but not in the CF1 male mouse. They induced the enzymes by
prefeeding the animals for 21 days with low doses (i.e., 10 or 25
mg/kg in diet) of dieldrin. If the results of the Kohli, et al.
study are to be accepted, then one may assume that since man is sub-
ject to chronic, low-dose exposure to many MFO inducers (including
various organochlorine pesticides), this exposure may affect stud-
ies of dieldrin metabolism.
C-24
-------
Excretion *
A: mentioned in the Distribution and Metabolism sections of
this report, aldrin and/or dieldrin are excreted mainly in the
£eces and to some extent in the urine in the form bf several metabo-
lites that are more polar than the parent compounds. Usually, a
plateau is reached in most tissues when the dose is held relatively
constant. However, if the dosage increases, the body concentra-
tions will increase and vice versa.
The early work of Ludwig, et al. (1964) demonstrated that male
14
Wistar rats administered daily low doses of C-labeled aldrin (4.3
ug for 12 weeks) excreted approximately nine times as much of the
radioactivity in the feces as in the urine. After about two weeks
of treatment, the rats were excreting 80 percent of the daily dose
of aldrin and this increased to 100 percent after eight weeks.
Twenty-four hours after the final dose (12 weeks), the animals had
excreted 88 percent of the total radioactivity fed. This increased
to 98 percent after six weeks and greater than 99 percent after 12
weeks. It appears that after eight weeks of feeding aldrin, a
saturation level was attained which did not increase with continued
feeding at the same concentration. The concentrations in the body
decreased rapidly once the feeding was terminated.
14
In a study with rabbits administered C-dieldrin orally over
a 21-week period (total dose 56 to 58 mg/kg), Korte and Arent
(1965) reported somewhat conflicting results. . At the end of the
feeding (22nd week) 42 percent of the total radioactivity had been
excreted with two to three times as much in the urine as in the
C-25
-------
feces. The level in the feces was negligible after 24 weeks while
the amount in the urine was up to 43 percent at 52 weeks.
It must be kept in mind that aldrin is metabolized to dieldrin
*
which is then converted to more polar metabolites for excretion.
It is possible that the increased amount of radioactivity noted by
Korte and Arent (1965) in the feces after treatment with aldrin
could be due to the less polar aldrin or dieldrin as compared to the
more polar metabolites excreted in the urine or to a basic differ-
ence in metabolism of dieldrin in the rabbit.
The work of Robinson, et al. (1969) on the metabolism of diel-
drin has been summarized in the Metabolism section of this report.
These investigators also studied the loss of dieldrin (99+ percent
purity) from the liver, blood, brain, and adipose tissue of male
CFE rats fed 10 mg/kg in their diet for eight weeks. Figure 1 il-
lustrates the loss of dieldrin from these tissues. During the per-
iod of observation, approximately 99 percent of the dieldrin was
excreted at various rates from the tissues. However, it must be
noted that the analysis was performed by gas-liquid chromatography
and that later investigators (Matthews, et al. 1971) have found
liver can contain approximately 30 percent of products other than
dieldrin, a fact which may have been overlooked by Robinson, et al.
(1969). The fat and brain contained greater than 99 percent of the
dieldrin and the'excretion times correspond to those for the rat
observed by Korte and Arent (1965) in their work six years earlier.
It can be seen from Figure 1 that three of the four slopes for
dieldrin loss were not linear and that with the blood and liver,
loss was rapid at first and then slowed down. Estimates for the
C-26
-------
*
Blood
01
ex 10*S4 2 «xp (-0 *535!)
*298«*p C-0- 05 291)
60
SO
40
Uv«r
10
e»fc» 71«>p t-0-5415
~o*233•«p (-o-oeat)
>
01
s
I
4.
a
©
23
5 001
e
e
p
S
o
s
o
0001
20
60
40
•§, 2001-|
1 soo
S
** i-o
| MUlfll.
C» 1 3 • 5« *p (
Adipost tiuut
0-067t)
lal
e
•.
u
0!
0-01
x
Brain
C«0*035-M0-230-0* 035)
* •xp(-Q• 23tJ
g 010
30 60 90 o 0 10 20
Ooys sines H100 ftiding etostd
FIGURE 1
The Loss of Dieldrin (HOED) from the Liver,
Blood, Brain, and Adipose Tissue of Male Rats
Source: Robinson, et al. 1969
C-27
-------
half-life of dieldrin in the liver and blood were 1.3 flays for the
period of rapid elimination and 10.2 days for the slower period.
The estimated half-life for dieldrin was 10.3 days in the adipose
tissue and 3.0 days in the brain.
In the study of 14C-dieldrin metabolism in sheep (Hedde, et
al. 1970) mentioned in the Metabolism section of this report, the
excretion of dieldrin or its metabolites was higher in the feces
than in the urine. This ratio varied considerable due partially to
the different doses used. The authors noted that in two very fat
sheep the ratio of labelled dieldrin in feces to urine was greater
than 10 to 1 but in two thin sheep receiving the same dose, it was
slightly greater than 1 to 1. The amount of radioactivity that was
-I a
exhaled as C02 was only 0.25 percent of the total dose. This
indicates that virtually none of the dieldrin is broken down to
C02. With the sheep, less than 50 percent of the total radioactiv-
ity was recovered after the five or six days of collection.
Several investigators have shown that removal of dieldrin from
the diet results in rapid loss of dieldrin or metabolites from the
body, especially the adipose tissue. Barron and Walton (1971) fur-
ther studied the loss of dieldrin from the body of the rat and also
looked at the role of dieldrin in the diet with respect to loss from
the adipose tissue. For this study, male Osborne-Mendel rats were
fed a diet containing 25 mg/kg dieldrin (99+ percent purity) for 8
weeks. They were then placed on a normal diet and given four daily,
oral doses of 14C-dieldrin equivalent to 25 mg/kg in their diet.
After these four days, one-half of the animals were then returned
to the dieldrin diet (25 mg/kg) while the rest remained on the
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normal diet. Groups of five animals were sacrificed on the four
days when they received the labeled-dieldrin and on days 7, 9, 11,
16, and 23 after the conclusion of the eight-ireek feeding. The,
concentration of dieldrin found in the adipose tissue from the rats
receiving the dieldrin diet was approximately SO ug/g and remained
at this level throughout the 23 days following the feeding period.
The concentrations in the rats on the normal diet decreased to 4
ug/g at day 23. The authors reported that the half-life of diel-
drin in the adipose tissue was about 4.S days, which is somewhat
lower than the 10.3 days calculated by Robinson, et al. (1969) with
rats fed only 10 mg/kg dieldrin.
14
Cole, et al. (1970) measured the appearance of C-dieldrin
and 14C-endrin in the urine and feces of male Holtzman rats for
seven days after a single intravenous dose of 0.25 mg/kg of either
chemical. They reported that greater than 90 percent of the radio-
activity occured in the feces. Approximately 80 percent of the
total dose of labeled dieldrin was excreted in the feces after the
seven days, compared with approximately 100 percent for the endrin.
Cole, et al. (1970) conducted a similar experiment during a four-
day period using bile-fistula rats. They also reported that these
rats produced a time course of excretion similar to those observed
in the first experiment; greater than 90 percent of the excreted
radioactivity was found in the bile.
In a comparison of the excretion of dieldrin in the CF1 mouse
and CPE rat, Baldwin, et al. (1972) found that after seven or eight
days the amount of labelled dieldrin excreted was similar in both
species. Also, the feces contained approximately two times as much
radioactivity as the urine, and 50 to 70 percent of the total
C-29
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activity was excreted during the collection period. As mentioned
in the Metabolism section of this report, the proportion of metabo-
lites varied between the mouse and the rat.
Although there has been extensive work done on the metabolism
and excretion of dieldrin in animals, there is understandably less
known about the fate of dieldrin in humans. Early work by Cueto and
Hayes (1962) demonstrated that dieldrin and some of its metabolites
could be detected in the urine of occupationally exposed workers.
A later report by Cueto and Biros (1967) compared the levels of
dieldrin and other chlorinated insecticides in the urine of 5 men
and 5 women in the general population to that of 14 men with differ-
ent degrees of occupational exposure. The concentrations of diel-
drin found in the urine of men and women in the general population
were 0.8 — 0.2 mg/1, and 1.3 — 0.1 mg/1, respectively. The concen-
trations found in male workers with low, medium, and high degrees
of exposure were 5.3 mg/1 (5), 13.8 mg/1 (4), and 51.4 mg/1 (5),
respectively (numbers in parentheses represent the number of indi-
viduals per sample). The degrees of exposure were only expressed
as relative and no data on the exposures were given.
Hayes and Curley (1968) measured the plasma, fat, and urine
concentration of various chlorinated pesticides in workers with
occupational exposure to these chemicals. In 14 urine samples, al-
drin was present at less than 0.2 mg/1 and dieldrin was present at
1.3 to €6.0 mg/1. This is compared to the mean for dieldrin in the
general population of 0.8—0.2 mg/1 determined in the same labor-
atory by Cueto and Biros (1967).
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A Study by Hunter, et al. (1969) concluded that dieldrin had a
relatively. long half-life in humans. This compares with a half-
life of less than ten days reported in animal studies. In the
Hunter, et al. study, 12 human volunteers ingested various doses of
dieldrin for up to 24 months. The blood and adipose concentrations
were determined over this time and the blood levels were followed
for eight additional months after termination of the treatment.
The authors reported that during this period concentrations of
dieldrin in the blood of three of the volunteers did not change
significantly. (These concentrations were not given). In the
other nine subjects, the half-life of dieldrin in the blood ranged
from 141 to 592 days with a mean of 369 days. These estimates were
made on a limited number of samples.
Jager (1970) reported that DeJonge, in an unpublished report,
studied the half-life of dieldrin in the blood of 15 aldrin/diel-
drin workers who were transferred to other areas. Prior to trans-
fer, these workers had had high exposures to the pesticides and
concentrations of aldrin/dieldrin in their blood had reached equi-
librium. Measurements of the dieldrin blood concentrations were
taken every six months for three years following the transfer. The
mean half-life was 0.73 years (approx. 266 days). This is somewhat
in agreement with the estimates of Hunter, et al. (1969) of 369
days based on limited data.
It has been reported by these and other authors (Robinson, et
al. 1969; Walker, et al. 1969) that there is a direct relationship
between the concentration of dieldrin in the blood' and that in
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adipose and other tissues. It seems likely that the half-life in
the blood may reflect the overall half-life in other tissues.
EPPECTS
Acute, Subacute, and Chronic Toxicity
The acute toxicity of aldrin and dieldrin has been extensively
summarized by Hodge, et al. (1967) and Jager (1970). In many
cases, aldrin and dieldrin are considered similar due to the rapid
conversion of aldrin to dieldrin (see Metabolism section). Diel-
drin, in turn, is metabolized to a variety of more polar products.
In some cases, the toxicity of the metabolites has been compared to
the parent compound but this information is rather sparse (Soto and
Deichmann, 1967).
After ingestion, aldrin and dieldrin are rapidly absorbed from
the gastro-intestinal tract. Following absorption, the pesticides
are transported from the liver to different sites in the body.
They have been found at various levels in the brain, blood (includ-
ing erythrocytes), liver, and especially the adipose tissue (Mick,
et al. 1971; Walker, et al. 1969). In addition, dieldrin has been
shown to cross the placenta to the fetus (Hathaway, et al. 1967).
Hunter, et al. (1969) demonstrated that a relationship between in-
take and storage exists and that a plateau is maintained in the
tissues unless the dose changes considerably.
It was shown early that the pesticide-to-solvent ratio affects
the (Barnes and Qeath, 1964) and that some variation is caused
by the solvent employed (Heath and Vandekar, 1964). There is a
pronounced variation in toxicity related to route of administra-
tion. Toxicity is highest by the intravenous route, followed by
C-'32
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oral, then dermal. This is aost likely due to the high blood and
central nervous system concentrations produced from intravenous
injection. Oral and dermal toxicity is lower due to lower blood
concentrations brought about by resorption and storage in adipose
tissue. For most species the acute oral toxic dose is between 20
and 70 mg/kg. This includes the rat, mouse, dog, monkey, sheep,
and man (Hodge, et al. 1967).
With both aldrin and dieldrin, toxicity in anlnals appears to
be related to the central nervous system. According to Hodge:
...a characteristic pattern has been described of stimu-
lation, hyperexcitability, hyperactivity, incoordina-
tion, and exaggerated body movement, ultimately leading
to convulsion, depression, and death.
There apparently is a direct correlation between blood concen-
trations and clinical signs of intoxication. Keane, et al. (1969)
reported that in dogs fed daily doses of dieldrin, the first signs
of muscle spasms occurred at 0.38 to 0.50 ug/ml blood and convul-
sions at 0.74 to 0.84 ug/ml.
The symptoms of intoxication in man are similar to those found
in mice, rats, and dogs. Jager (1970) described the symptoms re-
sulting from oral or dermal exposure that occur from 20 minutes to
24 hours as:
...headache, dizziness, nausea, general malaise, vomit-
ing, followed later by muscle twitching, myoclonic jerks
and even convulsions. Death may result from anoxemia.
Changes in the electroencephalogram (EEG) usually result after
insecticide intoxication and generally return to normal after dis-
continuance of exposure (Hoogendam, et al. 1962). The transitory
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change In the SEG has been challenged by several investigators (see
Burchfiel, et al. (1976) for recent summary). Work carried out in
Rhesus monkeys (Burchfielr *t al. 1976) using technical grade diel-
• •
drin (4 mg/kg, i.v. one time or 1 mg/kg i.m. administered once a
.i4
week for 10 weeks) demonstrated that dieldrin can alter the EEG for
up to 1 year.
The acute lethal dose of aldrin in man was reported by Jager
(1970) and Hayes (1971) based on the summary of Bodge, et al.
(1967) to be 5 g or 70 mg/kgr respectively. However, Hodge# et al.
only speculated on possible human toxic effects from a 1-year feed-
ing study in monkeys. It is known that persons have recovered from
acute oral doses of 26 mg/kg aldrin and.44 mg/kg dieldrin so that
the acute lethal human dose might be somewhat higher (Hayes, 1971).
The subacute or chronic toxicity of low doses of aldrin and
dieldrin to mice, rats, dogs and, to some extent, monkeys, has been
reported in many of the carcinogenicity studies included herein.
The resulting effects include shortened life span, increased liver-
to-body weight ratio, various changes in liver histology, and in-
duction of hepatic enzymes. Another effect that has been observed
is teratogenicity (Ottolenghi, et al. 1974).
Some information . is available concerning the subacute or
chronic exposure of humans to aldrin and dieldrin. Based on infor-
mation gained from monitoring workers at the Shell Chemical Com-
pany, Jager (1970) reported that 33.2 iig/kg/day can be tolerated by
workers for up to IS years. Above this level some individuals may
show signs of intoxication, although others can tolerate two times
this level. In another study involving 12 volunteers who ingested
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dieldrin Cor up to two years, 3.1 ;ug/kg/day was tolerated and pro-
duced no increase in plasma alkaline phosphatase activity (Hunter,
et al. 1969).
Synergism and/or Antagonism
Since aldrin and dieldrin are metabolized by way of the mixed
function oxidases (MFO), it must be assumed that any inducer or
inhibitor of these enzymes will affect the metabolism of aldrin or
dieldrin. Dieldrin and other organochlorine pesticides have been
reported to induce the MFO (Kohli, et al. 1977). Baldwin, et al.
(1972) reported that prefeeding low doses of dieldrin to rats
altered the metabolic products produced after acute dosing. Sev-
eral reports have appeared on the combined effect of aldrin or
dieldrin on the storage of DDT in tissues (Street, 1964; Street and
Blau, 1966; Deichmann, et al. 1969).
In the Deichmann, et al. (1969) study when aldrin was given
along with DDT or after a plateau had been reached in the blood and
fat by chronic DDT feeding. The retention of DDT by the blood and
fat increased considerably in the animals given both chemicals as
compared to the animals only given DDT. The authors suggest that
this increase in tissue DDT concentrations is due to a reduced rate
of excretion of DDT.
Walker, et al. (1972) fed groups of mice 50 or 100 mg/kg/diet
DDT or a mixture of 5 mg/kg/diet dieldrin and 50 mg/kg/diet DDT for
112 weeks. The highest incidence of tumors was in the dieldrin/DDT
group, although it is difficult to determine whether the effect
between dieldrin and DDT was additive or synergistic.
Clark and Krieger (1976) studied the metabolism and tis-
14
sue accumulation of C-labeled aldrin (99.3 percent purity) in
C-35
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combination with an inhibitor of oxidative biotransformation (i.e.,
SKF 525-A). They reported that pretreatment of male^ Swiss-Webster
mice with either 50 or 100 mg/kg SKF 525-A significantly increased
the accumulation of radioactivity in the blood, brain, kidney, and
liver. The SKF 525-A blocked the epoxidation of aldrin to diel-
drin. However, the authors did not feel that differences in meta-
bolite formation or excretion alone could account for the increased
accumulation in the tissues.
Teratogenicity
In 1967, Hathaway, et al. established that 14C-dieldrin could
cross the placenta in rabbits. Eliason and Posner (1971a,b) demon-
strated that ^^C-dieldrin crossed the placenta in the rat and that
the concentration in the maternal plasma increased as gestation
progressed. Deichmann (1972) reported that 25 mg/kg/diet aldrin
and dieldrin fed to mice for six generations markedly affected such
parameters as fertility, gestation, viability, lactation, and sur-
vival of the young, while mice fed lower doses showed fewer or no
effects.
In a study by Ottolenghi, et al. (1974) pregnant golden ham-
sters and pregnant CD-I mice were given single oral doses of puri-
fied aldrin, dieldrin, or endrin at one-half the (hamsters 50,
30, 5 mg/kg, and mice 25, 15, 2.5 mg/kg, respectively). The ham-
sters were treated orally on day seven, eight, or nine of gestation
and the mice on day nine. All three pesticides caused a signifi-
cant increase in fetal death in hamsters treated on days seven and
eight. Only dieldrin gave significant results on day nine. Ham-
sters treated on day eight also had the highest number of anomalies
C-36
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(i.e., open eye, webbed foot, cleft palate, and others). These
increased anomalies were noted for all three pesticides. The three
pesticides also reduced the fetal weight in the^hamsters treated on
the three different days. No significant difference was observed,
in the weight or survival of fetuses of treated and control mice;
-however, a teratogenic effect was observed in mice for all three
*.pesticides. It was less pronounced in the mice than in the hci§||p£
'\sters. The author reasoned that the reduced teratogenic effect in
mice may be due to the lower doses used in the mice.
Two later studies on the teratogenicity of dieldrin have
reached different conclusions. The studies of Chernoff, et al.
•(1975) and Dix, et al. (1977) both concluded that dieldrin was not
^teratogenic. Chernoff, et al. tested dieldrin (87 percent purity)
; and the photo-product, photodieldrin (95 percent purity) in CD-I
mice and CD rats orally at doses lower than those used by Otto-
lenghi, et al. (1974). The actual doses of dieldrin based on 87
-.percent purity were 1.3, 2.6, and 5.2 mg/kg/day over a ten-day per-
iod (i.e., days 7 to 16 of gestation). The compounds were dis-
solved in peanut oil. The control animals also received peanut
oil. The highest doses of dieldrin produced 41 percent mortality
in rats. In mice the highest doses induced significant increases
in liver-to-body weight ratios, reduced the weight gain, and pro-
duced some fetal toxicity. Photodieldrin at 0.6 mg/kg/day for 10
days also induced a significant increase in the liver-to-body
weight ratio in rats but caused no fetal toxicity. However, no
teratogenic effects were observed in the mice or rats at any of the
doses employed.
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Dix, «t al. (1977") "exwained the umToX'two solvents (corn oil
and dine thy lsulf oxide (DMSO)) with various doses of-dieldr-in -in CF1
mice. The corn oil groups received 1.5 or 4.0«'mg/kg/day of 99 per-
cent pure dieldrin orally with suitable controls of corn oil or no
treatment. The OMSO groups received 0.25, 0.5, or 1.0 mg/kg/day
with similar controls. Both solvent groups were treated on days 6
through 14 of gestation. In the corn oil group;" young (7-week)
virgin animals were ased and -the pregnancy rate was very low. With
the few animals that survived to term, the only significant effect
was delayed ossification in fetuses ct the mice administered the 4
mg level. The DMSO experiments were conducted with older animals
(ten weeks) of proven fertility. Fetuses of these animals demon-
strated a significant increase in incidences of delayed ossifica-
tion and extra ribs. However, the OMSO controls also had a high
incidence of these tvo anomalies. The authors attributed this to
the toxic effect of this solvent. DMSO also produced a reduction
in maternal and fetal body weights whereas the corn oil did not. No
differences were observed in the mean litter size, number of re-
sorptions, or fetal death with either solvent.
Mutagenicity
Relatively little work has been done on the mutagenicity of
aldrin or dieldrih. Of "the limited data available, most are con-
cerned with the mutagenicity of dieldrin. This may be sufficient,
since aldrin is readily converted to dieldrin in both .in vivo and
in vitro systems. Fahrig (1973) summarized the microbial studies
carried out up to 1973 on aldrin, dieldrin, and other organochlor-
ine pesticides including DDT and the metabolites of DDT. Aldrin
C-38
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and di«ldrin gave negative results with gene conversion in Sa-
ceharomvces cerevisae, back-nutation in Serratia marceacena. for-
ward mutation (Gal R8) in lachericia coli an<* forward mutation to
streptomycin resistance in £• coli. It is important to note that
DDT and several of its metabolites also gave negative results in
thaae microbial tasts and that no mention of any type of activation
system , mammalian liver enzymes) was made -in this summary.
Bidwell, et al. (1975) reported in an abstract that dieldrin
was not found to be mutagenic in five atrains of Salmonella tvoh-
imurium with or without the addition of a liver activation system,
although the authors did not give dose levels. They also stated
that dieldrin was negative in the host-mediated assay, blood and
urine analysis, micronucleus test, metaphase analysis, dominant
lethal test, and heritable tranalocation test. The doses used were
0.08, 0.8, and 8.0 mg/kg in corn oil with corn oil used as the con-
trol and triethylene melamine (0.5 mg/kg five times) serving as the
positive mutagenic control. The pesticide was given orally on a
subacute basis.
Dean, et al. (1975) evaluated dieldrin (99+ percent purity) in
two dominant lethal assays in CF1 mice, for chromosomal damage in
male and female Chinese hamsters and in the host-mediated assay
with Saccharomyces'oeravisiae in CF1 male mice.
Two dominant lethal assays were carried out, one with a single
oral dose of 12.5 or 25 mg/kg and the other with a single oral dose
of 12.5, 25 or 50 mg/kg. The treatment groups consisted of 8 males
and the DMSO solvent control groups of 16 males. In both experi-
ments, each male was caged with three females for 7 days. This was
C-39
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repeated for 8 weeks in the first study and for 5 weeks in the sec-
ond. Also, in the first experiment 8 mice received cyclophospha-
mide at 100 mg/kg orally as a positive control. ; In all cases, diel-
drin was dissolved in DMSO and the control animals were given DMSO.
• - • • • <• -- V -s, - • • . f* .
The''female's were killed and examined 13 days after the mid-week of
being caged with the males. All of the dieldrin-treated males
demonstrated signs of intoxication. One of the cyclophosphamide-
treated males died 7 days after treatment. Neither dieldrin nor
cyclophosphamide produced significant changes in the pregnancy rate
•of the female mice. However, when overall means of the total fetal
implants per pregnancy were examined, the 12.5 mg/kg and the cyclo-
phosphamide-treated groups were significantly lower than the con-
trols (P 0.05 and P 0.001, respectively) in the first experiment.
Conversely, the overall means for the 25 mg/kg group in the second
experiment was significantly higher than the control group
(P 0.05).
In the cytogenetic-studies using Chinese hamsters, four males
and four females were administered either DMSO, or dieldrin dis-
solved in DMSO at 30 or 60 mg/kg orally. Two animals of each sex
were killed at 8 arid 24 hours after treatment and slides were pre-
pared from the femurs. . One hundred cells were analyzed from the
bone marrow of each animal. While there is some problem determin-
ing the actual number of animals employed and the number of cells
examined , there appears to be no significant differences in gaps
or polyploidy between treated or control hamsters. It should be
2
The authors state in the results that 4,800 cells were analyzed
from 48 animals. However, from the methods section it appears that
only 24 animals were used in this study.
C-40
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noted that it appears that only two males and two £emales «ere exam-
ined at each time/dose point and this is a very small sample size
when trying to determine small increases as the authors are doing.
«
Another part of this study looked at mitotic gene conversion
in Saccharomyces cerevisiae D4 at the adej and trpg loci in a host-
mediated assay using male, CF1 mice. The experiments were divided
into three single dose and three multiple dose (5) protocols. In
the single dose treatments, mice received either DMSO, or dieldrin
dissolved in DMSO, orally at 25 or 50 mg/kg orally. The multiple
treatments consisted of DMSO or 5 or 10 mg/kg dieldrin orally for 5
days. There were two. mice per treatment group and the yeast were
injected i.p. either immediately after the single treatment or the
final multiple treatment. Ethyl methane-sulfonate (EMS) was given
orally at 400 mg/kg as a single dose. A small proportion of the
animals receiving dieldrin at 10 mg/kg for 5 days did not survive
but this is not reflected in the results given. The table summa-
rizing the results of the host-mediated assay states that two mice
per group were used but the number for the three experiments is
obviously less than six if all the mice did not survive. Of those
that did survive, only the EMS treatment groups had significant
increases in adenine and tryptophan convertants.
Three reports on the mutagenicity of aldrin or dieldrin have
recently been published. The first examined the mutagenicity of
dieldrin and several other pesticides with four strains of S.
typhimurium (i.e., TA1535, TA1536, TA1537, and TA1538) with the
addition of a rat liver activating system (Marshall, et al. 1976).
The second, an in-depth study of nearly 200 pesticides, utilized
several microbial indicators and, in some cases, the addition of an
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activating system (Shirasu, et al. 1977). The third study dealt
primarily with strains of S,. typhimurium (TA1535, TA100, and TA98)
plus a mouse liver activating system (Majumdar, et al. 1977).
, Zn ..the .Marshall, et al. (1976). study, dieldrin was tested at
only one concentration, 1,000 pg per plate, With and without the
addition of phenobarbital induced rat liver homogenate. In all
four strains tested, no increase in mutagenicity was observed at
this concentration.
Shirasu, et al. (1977) assayed aldrin with metabolic activa-
tion using E. coli B/r WP2 try-hcr+ and WP try-hcr" and S. typhim-
urium strains TA1535, TA1537, TA98, and TA100. Dieldrin was as-
sayed without metabolic activation using the E. coli WP2 her4*, WP2
her" and S5. typhimurium TA1535, TA1536, TA1537, and TA1538.
According to the authors, both aldrin and dieldrin were considered
nonmutagenic in these tests.
Wade, et al. (1979) have evaluated dieldrin using S. typhi-
murium strain TA100 and TA98 both with and without a rat liver
activating system. Their assay was in the form of a spot test at
50 and 1,000 ug per plate. At these two levels, dieldrin failed to
produce any mutagenic response.
Majumdar, et al. (1977), on the other hand, have reported that
dieldrin was somewhat mutagenic for S. typhimurium strains TA1535,
TA100, and TA98 without metabolic activation and that it was
strongly mutagenic for all three strains when liver enzymes from
Aroclor-1254^-induced mice were added to the mixtures.
3Aroclor-1254 is a mixture of PCBs, which induce the MFO in liver
(Ames, et al. 1975).
C-42
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in summarizing the limited microbial mutagenicity studies on
aldrin and dieldrin, it must be mentioned that the only refer-
ence to any mutagenicity in the Majumdar studifs contains several
notable inconsistencies. The inconsistencies are: (1) the cul-
tures used were grown for 24 hours rather than the recommended 16
hours; (2) the plates were incubated for 72 hours rather than the
conventional 48 hours; and (3) the control values for TA1535 and
TA98 were not consistent with those recommended by Ames, et al.
(1975).
Zt is not possible to say that these inconsistencies could
account for the positive mutagenic findings but they should be
taken into consideration in view of the fact that several other
similar, although not identical, studies reported no mutagenic
findings with dieldrin. Zt should be kept in mind that mice appar-
ently metabolize dieldrin differently than do rats (see the Metabo-
lism section of this report). Zt is possible that the use of the
mouse liver enzymes by Majumdar, et al. (1977) may be producing a
mutagenic metabolite not seen in other studies.
Studies on the mutagenic effects of dieldrin in organisms
other than microorganisms were also somewhat varied. Scholes
(1955) reported that dieldrin had no effect on onion root mitosis.
However, Markaryan (1966) observed an increase in the cytogenic
effects of dieldrin in mouse bone marrow nuclei and Bunch and Low
(1973) reported chromosomal aberrations in semi-domestic mallard
ducks.
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Recentlyf Majumdar, et al. (1976) studied (1) the e£fect of
dieldrin on chromosomes in mouse bone marrow Ln vivo and in cul-
tured human WI-38 lung cells, and (2) the cytopathic. effect of
dieldrin on the cultured human WI-38 cells. They reported a de-
crease in the mitotic index in both the Ui vivo mouse bone marrow
and i£ vitro human lung cells with the increasing concentration of
dieldrin used. In each test, an increase in chromosome aberrations
was observed with the lowest doses employed (1 mg/kg in mouse bone
marrow and 1 ug/ml in human cell cultures). The authors also re-
ported a dose- and time-dependent cytotoxic effect on the WI-38
human lung cells.
In addition, Ahmed, et al. (1977a) measured unscheduled DNA
synthesis (UDS) in SV-40 transformed VA-4 human fibroblasts in
vitro with and without an. uninduced rat liver activating system
using aldrin, dieldrin, DDT, and other pesticides. Both aldrin and
dieldrin produced a significant increase in UDS either with or
without the activating system at all the doses used.
Another study by this group (Ahmed, et al. 1977b) demonstrated
that dieldrin induced ouabain resistance in Chinese hamsters V79
cells when tested at a concentration of 0.01 M. With a cell sur-
vival of 77.8 percent, they obtained a mutation frequency of 16.4
mutants per 106 survivors as compared to 1.8 per 10® for the con-
trols.
Carcinogenicity
During the 1960's and the early part of the 1970's, numerous
studies on the carcinogenicity of aldrin and dieldrin appeared in
literature. These reports include studies on mice, rats, dogs, and
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monkeys. Of these species, mice appear to be the most susceptible
to aldrin/dieldrin. Various strains of both sexes have been exam-
ined at different dose levels. The effects range from benign liver
tumors to hepatocarcinomas with, transplantation confirmation
to pulmonary metastases. The data on carcinogenicity, have been
evaluated and discussed extensively, mainly by Epstein (1975a,b,
1976).
Six major studies using various strains of mice have been car-
ried out mainly by long-term feeding at low doses (i.e., 0.1 to 20
mg/kg in the diet). The earliest of these studies was conducted by
the U.S. Food and Drug Administration (FDA) (Davis and Fitzhugh,
1962). Using C^HeB/Fe (C^H) mice, both males and females were fed
either aldrin or dieldrin at 10 mg/kg in the diet for two years.
Both aldrin and dieldrin shortened the average life span by two
months. The experimental and control group death rate was high,
possibly due to overcrowding. Significantly more hepatomas were
observed in the treated groups than in the controls for both sexes.
In addition, the number of mice with tumors may have been under-
estimated due to the high mortality which left fewer animals for
evaluation.
. In an FDA follow-up study, Davis (1965) examined 100 males and
females of the C^H mice treated with aldrin or dieldrin at the same
concentrations as the first study. Again, survival was reduced
compared to the control group and there was an increase in benign
hyperplasia and benign hepatomas. A re-evaluation of the histolog-
ical material of both of these studies was carried out by Rueber in
1973 (Epstein, 1975a,b, 1976 ) . He concluded that the hepatomas
C-45
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were malignant and that both aldrin and dieldrin were hepatocar-
cinogenic for male and female C^H mice.
In a 1964 abstract. Song and Harville reported some indication
of hepatocarcinogenicity in CjH and CBA mice with aldrin (15 mg/kg)
and dieldrin (15 mg/kg) although minimal data are given. Epstein
(1975a,b, 1976) reviewed an unpublished study of MacDonald, et al.
on technical grade dieldrin in Swiss-Webster mice. The authors
concluded that dieldrin was noncarcinogenic but that there was some
questions as to the type of lesions.
Walker, et al. .(1972) conducted a multi-part study of dieldrin
in CPl mice of both sexes. In this study, the dieldrin used was 99+
percent pure and 4-amino-2,3-dimethylazobenzene (ADAB) was used as
the positive control. In the first part of the study, diets were
prepared containing 0, 0.1, 1.0, and 10 mg/kg dieldrin although
0.01 mg/kg dieldrin was found in the control (0 mg/kg) diet along
with low concentrations of other pesticides. The treatment groups
were made up of 600, 250, 250, and 400 mice, respectively, and con-
tained equal numbers of males and females. The ADAB group, which
contained 50 mice ' equally divided as to sex, received 600
rag/kg/diet for six months. Initially, the animals were housed five
to a cage, but after the sixth week they were placed in individual
cages. The positive controls were maintained separately from the
other groups. After nine months, the mice receiving 10 mg/kg ih
the diet dieldrin demonstrated palpable intra-abdominal masses, and
by the fifteenth month, half the males and females in the group had
died or had been killed when the masses became large. This period
of 15 months is short compared to the 20 to 24 months that elapsed
C-46
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before one-half of the control group had died. The life spans of
members' of the 0.1 mg/kg and 1.0 mg/kg groups were similar to those
of the controls. All the ADAB mice were dead by the 15th month.
An increased number of liver tumors was observed at all the
concentrations of dieldrin including 0.1 mg/kg, with the highest
increase occurring in the 10 mg/kg group. The tumors were classi-
fied by the authors as type (a) "...solid cords of closely packed
parenchymal cells with a morphology and staining affinity little
different from the rest of the parenchyma," or (b) "...areas of
cells proliferating in confluent sheets and often with foci of
necrosis. These lesions were distinguished from the previous types
of growth by the presence of areas of papilliform and adenoid for-
mations of liver cells with wide and irregular vascular channels
within the growth." This classification appears somewhat arbi-
trary. Nonetheless, the presence of tumors was dose-related and
effects were detected at the lowest dieldrin level tested (0.1
mg/kg). In addition to the increase in hepatic tumors there was an
increase in the incidence of tumors at other sites.
In the second part of the Walker, et al. (1972) study, groups
of 30 male and 30 female CFl mice received ethylene oxide-steri-
lized diets containing 1.25, 2.5, 5, 10, or 20 mg/kg dieldrin for
128 weeks. . The control group consisted of 78 males and 78 females
and the conditions and observations were similar to those in the
first experiment. In this part of the study, the mice that re-
ceived 20 mg/kg dieldrin in the diet had a high mortality rate.
About 25 percent of the males and 50 percent of the females showed
signs of intoxication and died during the first 3 months. Liver
C-47
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masses were detected at 36 weeks, and all the nice either died or
were killed at 12 months. Masses were not detected until 40 weeks
in the 10 mg/kg mice, 75 weeks in the 5 mg/kg m^ce, and 100 weeks in
the 2.5 mg/kg mice. In the 10 and 20 mg/kg groups, few animals were
available for examination due to the acute toxicity or their being
used in another study. The 5 mg/kg group had a higher incidence of
tumors than the 2.5 mg/kg group.
The third part of the study was carried out under similar con-
ditions. Groups of 60 mice received gamma-irradiated diets con-
taining 0 or 10 mg/kg/diet dieldrin for 120 weeks. Also, groups of
48 mice received gamma-irradiated diets and litter for 110 weeks or
unsterilized diets and litter for 104 weeks. The authors stated
that liver enlargement occurrence and mortality were similar to
those of the previous study.
The next section of the Walker, et al. (1972) study concerned
the combined effect of dieldrin and DDT treatment on CF1 mice.
Initially, the mice were fed diets containing 200 mg/kg DDT or
10:200 mg/kg dieldrinsDDT. This resulted in high mortality. The
diets were subsequently reduced to 50 and 100 mg/kg DDT and 5:50
mg/kg dieldrin:DDT. There were 47 males and 47 females in the con-
trol group and 32 males and 32 females in each of the treatment
groups. In mice on the 5:50 mg/kg diet and 100 mg/kg DDT diet,
liver enlargements were detected after 65 weeks of exposure. Both
of these doses were toxic to males but only the 5:50 mg/kg dose was
toxic to females. At 50 mg/kg DDT, masses were detected by the 96th
week but the mortality was similar to that of the controls. In this
experiment, the highest incidence of liver tumors was in the
C-48
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dieldrin:DDT group. However, because only one combination was
tested, it is difficult to determine whether the effect was syner-
gistic or additive. In a re-evaluation of the: experiment, Reuber
(see Epstein, 1975a,b, 1976), believes that Walker, et al. (1972)
over-estimated the incidence of liver tumors in the control and DOT
groups, thus minimizing the effect of the combined dieldrin/DDT.
In the last section of the Walker, et al. (1972) study, groups
of 58 mice were fed dieldrin at 10 mg/kg for 2, 4, 8, 16, 32, and 64
weeks and sacrificed after 2 years. The control group consisted of
156 mice. All groups were equally divided between males and fe-
males. In the mice receiving dieldrin for 64 weeks, liver enlarge-
ments were detected after 60 weeks in six males and two females.
These enlargements remained after the termination of the feeding.
No other enlargements were detected and the mortality of all the
groups was similar throughout the 2 years. It is important to note
that type b tumors were detected after only 4 or 8 weeks of treat-
ment and that the liver enlargements did not appear after the feed-
ing was terminated.
A similar study of dieldrin and other chemicals in CFl mice
was carried out by the same group (Thorpe and Walker, 1973). The
treatment groups were comprised of 30 males and 30 females and the
controls of 45 mice of each sex. Dieldrin was tested at one concen-
tration (10 mg/kg/diet) only, and the animals were not sacrificed
when abdominal masses were large as in the previous studies. The
study was terminated after 100 weeks of feeding. The authors re-
ported that there were no signs of intoxication in the dieldrin
groups; however, mortality increased after 22 months of exposure.
C- 49
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Also liver enlargements were detected in both sexes by the 50th
week. In this study, the cumulative tumor incidence and the number
o£ dead mice were given at 17, 21, 25, and 26 months. Dieldrin at
4
10 mg/kg produced a high incidence of liver tumors. All the males
and one-half the £emales that had died by 17 months had liver tu-
mors. By the end o£ the study, 100 percent o£ the males and 87
percent of the £emales had liver tumors.
In a recent evaluation of both aldrin and dieldrin by the Na-
tional Cancer Institute, aldrin and dieldrin were £ound to produce
hepatic carcinomas in male mice. Female mice responded to low
doses o£ dieldrin, but showed no effects from aldrin. No carcino-
mas were observed in either male or female rats of two different
species (43 FR 2450) when the subjects were exposed to both aldrin
and dieldrin. In the study on mice, groups of 50 male and 50 female
B6C3F1 mice were fed either aldrin (technical grade) or dieldrin
(technical grade) at various doses. The females received aldrin at
3 and 6 mg/kg/diet and the males received aldrin at 4 and 8 mg/kg.
Both sexes were given dieldrin at 2.5 and 5 mg/g. Aldrin controls
consisted of 20 untreated males and 10 females and dieldrin con-
trols had 20 animals per group. In addition, pooled controls con-
sisted of 92 males and 78 females. The animals were fed the pesti-
cide diets for 80 weeks and then observed for 10 to 13 weeks. All
survivors were killed at 90 to 93 weeks.
In the male mice administered aldrin, there was a significant,
dose-related increase in the incidence of hepatic carcinomas. The
values were: matched controls 3/20 (15 percent); pooled controls
17/92 (19 percent); 4 mg/kg 16/49 (33 percent); and 8 mg/kg 25/45
C-50
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(56 percent.). The mean body weights of the aldrin- and dieldrin-
fed mice were similar in the control and treated groups. There was
a dose-related mortality in £emale mice at the high dose o£ aldrin.
With the male mice £ed dieldrin, a significant increase in hepatic-
carcinomas was observed in the 5 rag/kg group. The incidences were
12/50 (24 percent) for the 2.5 mg/kg group and 16/45 (36 percent)
for the 5 mg/kg group.
There have also been six carcinogenicity studies of aldrin
and/or dieldrin done in various strains of rats. In an early paper
by Treon and Cleveland (1955) aldrin and dieldrin were fed to male
and female Carworth rats at 2.5, 12.5, and 25 mg/kg. The authors
reported a significant increase in mortality and an increase in
liver-to-body weight ratios at all concentrations tested. No data
on tumor incidences were given, although some liver lesions were
detected. Later Cleveland (1966) summarized the work on aldrin and
dieldrin conducted at the Kettering Laboratory. Although little
data and details were given, Cleveland stated that aldrin and diel-
drin were not tumorigenic in their rat studies.
A study was carried out by the U.S. Food and Drug Administra-
tion on aldrin and dieldrin in rats and dogs (Fitzhugh, et al.
1964) to determine the toxicity of these pesticides. Groups of 12
male and 12 female Osborne-Mendel rats were fed diets containing
either aldrin (99+ percent purity) or dieldrin (100 percent purity)
at 0, 0.5, 2, 10, 50, 100, or 150 mg/kg for two years. The animals
were housed individually and the survivors were killed after two
years. None of the dose levels of aldrin or dieldrin affected the
growth of the rats but both chemicals at 50 mg/kg or greater
C-51
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reduced the survival. A significant increase in liver-to-body
weight ratios was observed in both males and females for several
doses of both chemicals. The authors reported no increase in liver
«
tumors; however, there was a high incidence of multiple site tumors
at lower concentrations of both aldrin and dieldrin.
Deichmann, et al. (1967) carried out a study in which 5 mg/kg
aldrin (technical grade) was fed to male and female Osborne-Mendel
rats, either individually or in combination with 200 mg/kg aramite,
200 mg/kg DDT, and 1,000 mg/kg methoxychlor. There were 30 males
and 30 females in each treatment group and .they were housed in
pairs. No increase in mortality over the controls was observed in
any of the treated groups. Aldrin alone had no significant effect
on liver-to-body weight ratio, but an increase in the ratio was
noted in the groups treated with the pesticide mixtures. The
authors state that one-half (13 females and 2 males) of the aldrin-
treated rats had one tumor; however, only the tumors in survivors
were listed.
Walker, et al. (1969) fed dieldrin (99+ percent purity) to
Carworth rats at concentrations of 0, 0.1, 1.0, and 10 mg/kg in the
diet for two years. There were 25 males and 25 females in each
treatment group and 45 rats of each sex in the control group. The
animals were housed individually and dying animals were killed and
examined. The authors reported that some irritability, tremo.rs,
and convulsions occurred after two to three months but that the
animals remained in good health for the two years. None of the
dieldrin doses had any effect on body weight. Mortality was the
same for the control and treated groups; however, all the groups
C- 52
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had an overall, high rate of mortality. This resulted in only a few
. animals being available for examination at the conclusion of the
feeding. At 1 and 10 mg/kg there were increases in liver-to-body
weight ratios. Only one male rat and four female rats at the 10
mg/kg level demonstrated any liver cell changes. However, at the
0.1 and 1.0 mg/kg levels there were high but not significant in-
* creases in total tumors even though few animals were examined his-
tologically.
In another study with the Osborne-Hendel rat, Deichmann, et
. al. (1970) examined aldrin, dieldrin, and endrin in a lifetime
exposure. Aldrin (technical, 95 percent) and dieldrin (technical4,
100 percent active ingredients) were fed in the diet to groups of
50 males and SO females. The concentrations during the first two
weeks were 10, 15, and 25 mg/kg aldrin and 10, 15, and 25 mg/kg
dieldrin. After this time all the dose concentrations were doubled
' for the remainder of the treatment time. The control groups con-
tained 100 rats of each sex. Any animals that appeared ill were
sacrificed. Both aldrin and dieldrin produced some dose-related
toxicity, tremors, and clonic convulsions, especially in females.
However, these doses had no effect on mean gain in body weight
although some animals had marked loss of weight. The mean survival
rate was somewhat lower in the aldrin and dieldrin rats; again,
predominantly in females receiving the high concentrations. There
were significant increases in liver-to-body weight ratios in
males fed aldrin at 30 and 50 mg/kg and dieldrin at 30 mg/kg and a
4This is somewhat contradictory since "technical" dieldrin is
actually 85 percent pure.
C-53
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significant decrease in liver-to-body weight ratios in females fed
aldrin at 20 mg/kg. A moderate increase in hepatic centrilobular
cloudy swelling and necrosis was observed ia both male and female
rats fed aldrin and dieldrin as compared to the controls. However,
there was no increase in the number of liver tumors or other site
tumors. In fact, a decrease in total tumors was observed in both
the males and females fed aldrin and dieldrin. The authors stated
that this was possibly due to increased microsomal enzyme activity.
It should be noted that limited re-evaluation of this data was car-
ried out by Reuber who disagreed with the findings of Deichmann, et
al. (1970). However, he re-evaluated only one group (dieldrin, 30
mg/kg) and there has been no independent re-evaluation of the
material.
A two-year study by the National Cancer Institute (1976)(43 FR
2450) studied the effects of technical grade aldrin and dieldrin on
Osborne-Mendel and Fisher 344 rats. The first part of the study
used groups of 50 Osborne-Mendel rats of each sex for aldrin (30 or
60 mg/kg) and dieldrin (29 or 65 mg/kg). Aldrin was fed to the
males for 74 weeks. The rats were then observed for an additional
37 to 38 weeks. All survivors were killed at 111 to 113 weeks. The
same doses of aldrin were administered to the female rats for 80
weeks, followed by 32 to 33 weeks of observation. All survivors
were killed at- 111 to 113 weeks. The dieldrin rats were treated for
59 weeks at 65 mg/kg followed by 51 to 52 weeks of observation, or
80 weeks at 29 mg/kg followed by 30 to 31 weeks of observation. All
survivors were killed at 110 to 111 weeks. For both pesticides,
the controls consisted of 10 untreated rats of each sex plus pooled
C- 54
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controls consisting the matched control groups combines with SB
untreated males and 60 untreated females from similar bioassays of
other chemicals.
During the first year of the rat studies, the mean body
weights for the aidrin- and dieldrin-fed rats did not differ from
those of the controls. However, during the second year-, the body
.. weights of the treated rats were lower than those of the untreated.
For both aldrin and dieldrin, no significant increase in hepatic
carcinomas was observed in either sex. There was a significant
increase in adrenal cortical adenoma in the low-dose aldrin- and
dieldrin-treated female rats.
In the second part of the study on rats, 24 male and 24 female
Fisher 344 rats were fed purified dieldrin at 2, 10, or 50 mg/kg of
diet for 104 to 105 weeks. Matched controls consisted of 24 rats of
each sex. All survivors were killed at 104 to 105 weeks. The body
weights of the treated and control rats were similar and survival
was not greatly affected. The high-dose males and females demon-
strated signs of intoxification at 76 and 80 weeks, respectively.
A variety of neoplasms occurred in both the control and treated
rats? however, there were no significant dose-related increases in
the neoplasms.
To date, there has been only one carcinogenicity study report-
ed on either aldrin or dieldrin in hamsters. Cabral, et al. (1979)
carried out lifetime feeding studies in Syrian golden hamsters with
dieldrin (99 percent purity). Groups of nearly equal size (i.e.,
32-41 per group) of male and female hamsters were fed a diet con-
taining 0, 20, 60 or 80 mg/kg for up to 120 weeks at which time the
C- 55
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remaining survivors were killed. While there was no decrease in
survival at 50 weeks, the numbers of females remaining at 70 weeks
was one-half or less than the males. At 90 weeks the survival rate
was about 10 percent for all groups except the males of the 180
mg/kg level which had 32 percent survivors. Both males and females
at the low and high doses demonstrated a marked retardation of
growth. The authors state that there was no significant difference
between the percentage of control animals with tumors and the
treated animals with tumors. However, in the treated groups, more
animals had more than one tumor than in the control groups. Al-
though there was an increase in the number of animals with adrenal
tumors, especially males, again this was not statistically signifi-
cant. In the animals receiving the high dose of dieldrin, there
was one male and one female which had hepatomas. It was also noted
by these authors that there was a dose-related increase in the
incidence of hepatic cell hypertrophy in the dieldrin-treated ham-
sters.
There has been minimal work on the carcinogenicity of aldrin
or dieldrin in dogs. A limited, short-term study was conducted by
Treon and Cleveland (1955). Aldrin and dieldrin were fed to two
male and two female beagles at 1 and 3 mg/kg/diet. The dogs were
killed between 15 and 16 months. Although the growth rates of the
treated dogs were similar to those of the controls, liver weights
were increased at 1 mg/kg. These doses were toxic to the dogs and
mortality was high. The study provides few data on the necropsy
and the treatment was too short to adequately evaluate carcinogen-
icity.
C-56
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In another study using dogs, Fitzhugh, et al- (1964) treated
26 animals with aldrin or dieldrin at dosages of 0.2 to 1.0
mg/kg/day, 6 days a week# up to 25 months. Aldoses of 0.5 mg/kg
and greater, toxic effects including weight loss, convulsions, and
death were observed. At 1 mg/kg/day or higher no animals survived
over 49 days, and at 2.5 and 10 mg/kg/day all dogs died within 10
weeks. However, dogs fed 0.2 mg/kg/day of aldrin or dieldrin
showed no ill-effects during the 2 years of the study. In the dogs
fed aldrin at 1.0 mg/kg/day or dieldrin at 0.5 mg/kg/day, fatty
degeneration was observed in the liver and kidneys. This study
also was too short-termed to determine tumorigenic properties of
aldrin and dieldrin. The number of animals surviving at the end of
the study was inadequate to make any type of evaluation.
A third short-termed study on dieldrin in dogs was carried out
by Walker, et al. (1969). Dieldrin (99+ percent purity) was admin-
istered to groups of five male and five female dogs in gelatine
capsules at 0.005 and 0.05 mg/kg/day. After two years, the health
and body weight of the treated dogs, as compared to the controls,
was normal. A variety of physiological tests confirmed the general
good health of the dogs. In dogs administered the higher concen-
tration of dieldrin, liver-to-body weight ratios were increased
significantly over the controls. The report stated that no lesions
were seen in the tissues but provided no data on this.
There has been one report on the effects of dieldrin on Rhesus
monkeys. The unpublished work of Zavon and Stemmer (1975), from
the Kettering Laboratory, reports on a study in which -six con-
trol monkeys (five male, one female) and groups of five monkeys
C-57
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received 0, 0.01, 0.1, 0.5, 1.0 or 1.75 mg/kg dieldrin in their
diet £o'r 5.5 to 6 years. The group at 1.75 mg/kg received 5.0 mg/kg
for £our months, then 2.5 mg/kg for approximately 2.5 months, and
then 1.75 mg/kg for the remainder of the exposure. Additionally,
one monkey in this group had its dieldrin intake progressively in-
creased to the 5 mg/kg concentration. The authors state that this
animal and three others died during the study. These animals had
received 5.0, 1.0 or 0.1 mg/kg dieldrin in their diets. The re-
maining animals survived until they were killed.
Fat biopsies were taken on selected animals at various inter-
vals. Dieldrin blood levels and other parameters were determined
throughout the study.
The authors concluded that there were no significant hepatic
changes other than alterations in cytochrome P-450 levels. They
also stated that there was no indication of dieldrin-associated
malignancies although admittedly this was not considered a cancer
study. It was also the opinion of the authors that the premature
deaths were not related to the ingestion of dieldrin.
Versteeg and Jager (1973) summarized health studies carried
out on pesticide workers in the Shell plant in Holland. These
workers had occupational exposure to aldrin/dieldrin over periods
of up to 12.3 years with a mean of 6.6 years. The average time that
had elapsed from the end of exposure was 7.4 years (maximum, 16
years). The average age of the group was 47.4 years. The report
states that 233 long-term workers were involved in this study and
that no permanent adverse effects (including cancer) on the work-
ers' health were observed.
C-58
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Recently, Van Raalte (1977) published a follow-up on presum-
ably the same group of workers reported on by Versteeg and Jager
(1973). This study listed the various physiological parameters
which were examined in workers with more than four years of expo-
sure. These workers were examined in terms of two categories, one
- having workers with more than four years exposure and more than IS
years of observation (a total of 166 men) and the other with more
than ten years exposure and more than 15 years of observation (69
men). While this is the same number of subjects in both studies
done at the same plant, Versteeg and Jager listed aldrin/dieldrin
and other pesticide exposures while Van Raalte only mentioned diel-
' drin. This study appears to be a continuation of the previously
- reported work with an additional number of years of exposure and
observation. The author states that again there were no persisting
medical problems in the workers and no increase in cancer. Van
Raalte also goes on to point out that several other of the human
carcinogens have been detected in limited populations after rela-
tively short times. He suggests that the lack of early adverse
health signs and the lack of an increase in* cancer at this time
strengthens the assumption that dieldrin is not a human carcinogen.
While it is most likely correct to assume that these workers
are probably the most highly exposed group available for study, the
total number is again rather small and the observation times are
still less than 20 years.
Epstein (1975a) states that the epidemiological aspects of the
Study carried out by Shell have been reviewed by several experts
who have criticized the study as inadequate due to the number of
C-59
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#
workers at risk and the short duration of exposure and/or time
after exposure."
C-60
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CRITERION PORMOLATION
existing Guidelines and Standards
Prior to 1974, aldrin and dieldrin were^approved for use on 46
agricultural crops and for treatment of soil around fruits, grains,
nuts, and vegetables (Int. Agency Res. Cancer, 1974a,b). In 1974
•the registration'of aldrin and dieldrin vas suspended on the basis
of adverse health affects in rodents (39 PR 37251). As a result,
production is restricted for all pesticide products containing al-
drin or dieldrin. Aldrin and dieldrin can no longer be used for
spraying and dusting, or for mothproofing in which the residues are
discharged into waterways. All uses in structures occupied by hu-
mans or livestock, uses upon turf, and any use involving applica-
tion to any aquatic environment are also restricted. Aldrin and
dieldrin can be used for termite treatment which involves direct
application to the soil and therefore little movement of the pesti-
cides. They may also be used for treatment of some nonfood seeds
and plant dipping during transplantation.
The current exposure level for both aldrin and dieldrin set by
s
the Occupational Safety and Sealth Administration is an air time-
3
weighted average (TVA) of 250 jig/nr for skin absorption (37 FR
22139). In 1969, the.U.S. Public Health Service Advisory Committee
recommended that the drinking water standards for both aldrin and
dieldrin be 17 jig/1 (Mrak, 1969). Also, the U.N. Food and Agricul-
ture Organization/World Health Organization's acceptable daily
intake for aldrin and dieldrin is 0.0001 rag/kg/day (Mrak, 1969).
C-61
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Current Levels of Exposure
Th. people of the United States are exposed to aldrin and
dieldrln in air, water, and Cood. As mentioned earlier, aldrin or
dieldrin has been found in store than 85 percent of the air samples
tested by the U.S. EPA (Epstein, 1976). The levels were as high as
3
2.8 ng/m resulting in an intake of up to 0.098 jig/day. Dieldrin
can travel great distances in the air, especially when absorbed to
particulate natter. Thus people can potentially be exposed to
pesticide treatments from other countries.
*
Waters recently sampled in the United States contained aldrin
or dieldrin in amounts up to 0.05 jig/1 (Harris, et al. 1977). The
standard diet in the United States has been calculated to contain
approximately 43 ng/g of dieldrin. According to Epstein (1976)
tolerances for dieldrin in cattle-meat fat, milk fat, meat, and
meat by-products have been petitioned for at levels of 0.3, 0.2,
and 0.1 ppm, respectively.
Special Groups at Risk
¦¦aaMHiMWMHaMBiknMawaaHiaMaa*
Children, especially infants, have a high dairy product diet
that has been shown to contain dieldrin (Manske and Johnson, 1975).
It has also been demonstrated that human milk contains dieldrin
residues and that some infants may be exposed to high concentra-
tions of dieldrin from that source alone (Savage, 1976).
In early studies, Curley and Kimbrough (1969) and Zavon, et
al. (1969) reported that dieldrin and several other chlorinated
hydrocarbon pesticides were present in the tissues of stillborn
infants. Curley, et al. (1969) also reported that dieldrin and
other pesticides could be found in the blood of newborn infants.
C-62
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Ho votk has been carried out on neonatal animals with either
aldrin or dieldrin; however, due to the sensitivity of neonatal
animals to other carcinogens, this should toe an area of treat con-
cern.
Basis and Derivation of Criteria
The aldrin and dieldrin carcinogenicity data of Walker, et al.
(1972) and the National Cancer Institute (1976} (43 FR 24S0) were
analyzed using a linearized multistage model as discussed in the
Human Health Methodology Appendices to the October 1980 Federal
Register notice which announced the availability of this document.
It should be noted that the Walker, et al. study used 99 percent
pure dieldrin while the NCI study used technical grade dieldrin.
tinder the Consent Decree in NRDC v. Train, criteria are to
state "recommended maximum permissible concentrations (including
where appropriate, zero) consistent with the protection of aquatic
organisms, human health, and recreational activities." Both aldrin
and dieldrin are suspected of being human carcinogens. Because
there is no recognized safe concentration for a human carcinogen,
the recommended concentration of aldrin/dieldrin in water for maxi-
mum protection of human health is zero.
Because attaining a zero concentration level may be infeasible
in some cases and in order to assist the Agency and states in the
possible future development of water quality regulations, the con-
centrations of aldrin and dieldrin corresponding to several incre-
mental lifetime cancer risk levels have been estimated. A cancer
risk level provides an estimate of the additional incidence of can-
cer that may be expected in an exposed population. A risk of 10~5
C-63
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for example/ indicates a probability of one additional case of can-
cer for every 100,000 people exposed, a risk of 10~6 indicates one
additional case of cancer for every million people exposed, and so
forth.
In the Federal Register notice of availability of draft.ambi-
ent water quality criteria, EPA stated that it is considering set-
ting criteria at an interim target risk level of 10~^, 10~® or 10~7
as shown in the table below.
Exposure Assumptions Risk Levels and Corresponding Criteria
0 10~7 10~6 10"5
ng7l ng7l ng7l ng7l
(
2 liters of drinking water
and consumption of 6.5 grams
of fish and shellfish (2)
Aldrin 0 0.0074 0.074 0.74
Dieldrin 0 0.0071 0.071 0.71
Consumption of fish
and shellfish only.
Aldrin 0 0.0079 0.079 0.79
Dieldrin 0 0.0076 0.076 0.76
(1) Calculated by applying a linearized multistage model as dis-
cussed above. Appropriate bioassay data used in the calcula-
tion of the model are presented in Appendix I. Since the
extrapolation model is linear at low doses, the additional
lifetime risk is directly proportional to the water concentra-
tion. Therefore, water concentrations corresponding to other
risk levels can be derived by multiplying or dividing one of
the risk levels and corresponding water concentrations shown
in the table by factors such as 10, 100, 1,000, and so forth.
C-64
-------
(2) Ninety-four percent of aldrin exposure results from the con-
sumption of aquatic organisms which exhibit an average biocon-
centration potential of 28-fold, but for purposes of criterion
development are assumed to bioconcentrate aldrin at 4670,
because aldrin is converted to, and stored as dieldrin in
these organisms (see Appendix I). The remaining 6 percent of
aldrin exposure results from drinking water.
Ninety-four percent of dieldrin exposure results from the con-
sumption of aquatic organisms which exhibit an average biocon-
centration potential of 4670-fold. The remaining 6 percent of
dieldrin exposure results from drinking water.
Concentration levels were derived assuming a lifetime exposure
*to various amounts of aldrin/dieldrin, (1) occurring from the con-
sumption of both drinking water and aquatic life grown in water
containing the corresponding aldrin/dieldrin concentrations and,
(2) occurring solely from the consumption of aquatic life grown in
the waters containing the.corresponding aldrin/dieldrin concentra-
tions.
Although total exposure information for aldrin and dieldrin is
discussed and an estimate of the contributions from other sources
of exposure can be made, this data will not be factored into the
ambient water quality criteria formulation because of the tenuous
estimates. The criteria presented, therefore, assume an incremen-
tal risk from ambient water exposure only.
C-65
-------
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C-73
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•
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•
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APPENDIX 1
*
Summary and Conclusions Regarding the
Carcinogenicity of Aldrin and Dieldrin*
Aldrin has induced liver tumors in males and females of three
strains of mice according to reports of four separate chronic feed-
ing studies. It has failed to induce a statistically significant
carcinogenic response in rats at any site according to reports of
five studies in two different strains. In two bacterial assays
with and without activation (£5. typhimurium and E. coli) it was
found to be nonmutagenic, but it did produce unscheduled DNA syn-
thesis in human fibroblasts with and without activation. The in-
duction of hepatocellular carcinoma in both male and female mice
from the administration of aldrin leads to the conclusion that it
is likely to be a human carcinogen.
Dieldrin, which is readily formed from aldrin in the environ-
ment and by metabolism of aldrin in rats, mice, fish, and many
other species, has produced liver tumors in four strains of mice
according to six reports of chronic feeding studies and possible
liver tumors in an unpublished study with a fifth strain. In rats
it has failed to induce a statistically significant excess of tum-
ors at any site in six chronic feeding studies in three strains. It
was found to be mutagenic in S. typhimurium after metabolic acti-
vation with mouse liver enzymes, but it was not mutagenic in
two other studies of the same bacterial strain with a rat liver
*
This summary has been prepared and approved by the Carcinogens
Assessment Group, U.S. EPA, on July 25, 1979.
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enzyme activation mixture. The induction of hepatocellular carci-
nomas in mice leads to the conclusion that dieldrin i.s likely to be
a human carcinogen.
Both aldrin and dieldrin have been found to be nonmutagenic in
several test systems as follows: a) gene conversion in S. •cere-
visie; b) back mutations in j3. marcescens, and c) foward mutations
at two loci in £. coli. Several other organochlorine pesticides
which produce mouse liver tumors are also nonmutagenic in the same
systems.
The induction of liver tumors in mice of both sexes by aldrin
and dieldrin is sufficient evidence that they are likely to be hu-
man carcinogens.
The water quality criterion for aldrin is based on the hepato-
cellular carcinoma incidence in male B6C3F1 mice in the NCI chronic
test, and on this same response in groups of female CF-1 mice in the
Walker, et al. (1972) experiment, because aldrin is converted to
and stored as dieldrin in fish. It is concluded that the water con-
centration of aldrin should be less than 0.74 ng/1 in order to keep
the lifetime cancer risk below 1Q~5. For dieldrin the criterion is
based on the response in groups of female CF-1 mice in the Walker,
et al. (1972) experiment. The corresponding concentration for
dieldrin is 0.71 ng/1.
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Summary of Pertinent Data for Aldrin
The water quality criterion for aldrin is derived from the
hepatocellular carcinoma response of the B6C3P1 male mice given
aldrin in the MCI bioassay test. The slope of the one-hit dose-
response curve for aldrin is calculated from the following para-
meters:
Dose Incidence
(mq/kq/day) (no. responding/no. tested)
0.0 3/30
0.52 16/49
1.04 25/46
le ¦ 80 weeks w • 0.035 kg
Le ¦ 90 weeks
L m 90 weeks
With these parameters the carcinogenic potency factor for hu-
mans, q^ , for aldrin is 11.45 (mg/kg/day) .
The conversion of aldrin to dieldrin in fish results in the
accumulation of dieldrin residues in fish exposed to aldrin. This
makes it necessary to consider the risk resulting from intake of
dieldrin stored in fish due to the presence of aldrin in water.
Thus, the criterion for aldrin also depends upon the carcinogenic
potency factor for humans, q^*, for dieldrin, which is 30.37
(mg/kg/day)"1.
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The equation describing the risk due to aldrin in water is
derived Srom the general relationship:
P » BgD and D •» 1/70 kg, thus
P « Bh 1/70 kg and
PC70 kg) - Bgi
where
P « individual lifetime risk (set at 10~5 for criterion
calculation)
I « average daily human intake of the substance in
question
BH » estimated carcinogenic potency factor for humans
70 kg » average weight of humans
Since aldrin in water leads to the accumulation of dieldrin
residues in fish, the equation describing the risk due to aldrin
is;
Pa (70 kg) » Bga Ca (2.0 l/day)+ BHa Cft Rad (0.0065 kg/day) +
Bgd Ca Rad (0.0065 kg/day)
where
P * risk due to aldrin (set at 10"5 for criterion calcu-
lation)
Bs « 11.45 (mg/kg/day)"1, the aldrin carcinogenic potency
factor for humans
BHd » 30.37 (mg/kg/day)"1, the dieldrin carcinogenic po-
tency factor for humans
C « criterion concentration for aldrin (to be calculated)
cl
r * 28 1/kg, the fish bioconcentration of aldrin from
aldrin
R d " 4642 1/kg, the fish bioconcentration of dieldrin
from aldrin
2.0 1/day - average daily intake of water for humans
0.0065 kg/day ¦ average daily intake of fish for humans
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The term containing Rad represents intake of dieldrin resulting
from the presence of aldrin in the water, and is thus multiplied by
the dieldrin dose-response slope. Rad is estimated by assuming
that in the absence of conversion to dieldrin, aldrin would biocon-
centrate 4670 times (as dieldrin does), and that since aldrin only
accumulates 28 times, the remainder of the expected aldrin residues
are being stored as dieldrin (i.e., 4670 - 28 ¦ 4642).
The result is that the water concentration of aldrin should be
less than 0.74 ng/1 in order to keep the individual lifetime risk
-5
below 10 .
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Summary of Pertinent Data £or Dieldrin
•The water quality criterion for dieldrin is based on the hepa-
tocellular carcinoma response of the female CF-1 mice given various
*
concentrations of dieldrin continuously in the diet in the experi-
ment of Walker, et al. (1972). The parameters of the dose-response
Incidence
(no. responding/no. tested)
39/297
24/90
32/87
136/148
w ¦ 0.030 kg
R - 4670 1/kg
With these parameters the carcinogenic potency factor for hu-
nans, q^, is 30.37 (mg/kg/dayl"1. The result Is that the water
concentration should be less than 0.71 ng/1 in order to keep the
individual lifetime risk below 10~5.
model are:
Dose .
(ag/kg/day)
0.0013
0.013
0.128
1.28
le » 924 days
Le * 924 days
L » 924 days
Doses are concentrations determined to be in the diet. The first
dose group, the control, was found to have a level of contamination
in the diet equivalent to 0.0013 mg/kg/day.
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