297 916:
ALDRIN /DIELDRIN
Ambient Water Quality Criteria
Criteria and Standards Division
Office of Water Planning and Standards
U.S. Environmental Protection Agency
Washington, D.C.
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CRITERION DOCUMENT
ALDRIN-DIELDRIN
CRITERIA
Aquatic Life
For aldrin/dieldrin the criterion to protect freshwater
aquatic life as derived using the Guidelines is 0.0019 ug/1 as a
24-hour average and the concentration should not exceed 1.2 ug/1
at any time.
For aldrin/dieldrin the criterion to protect saltwater
.aquatic life as derived using procedures other than the Guidelines
is 0.0069 ug/1 as a 24-hour average and the concentration should
not exceed 0.16 ug/1 at any time.
Human Health
For the maximum protection of human health from the potential
carcinogenic effects of exposure to aldrin through ingestion of
water and contaminated aquatic organisms, the ambient water con-
centration is zero. Concentrations of aldrin estimated to result
in additional lifetime cancer risks ranging from no additional
risk to an additional risk of 1 in 100,000 are presented in the
Critrion Formulation section of this document. The Agency is con-
sidering setting criteria at an interim target risk level in the
range of 10~5, 10~6, or 10~7 with corresponding criteria of 4.6 x
10~2 ng/1, 4.6 x 10~3 ng/1, and 4.6 x 10~4 ng/1, respectively.
For the maximum protection of human health from the potential
carcinogenic effects of exposure to dieldrin through ingestion of
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water; and .contaminated aquatic organisms, the ambient water con-
centration :is zero. Concent rat ions of dieldrin estimated to re-
sult :in ^additional lifetime cancer risks ranging from no^addi-
tional risk to an additional risk of 1 in 100,000 are presented in
the Criterion Formulation section of this document. The Agency is
considering setting criteria at an interim target risk level in
the range of 10~5, 10~6, or 10~7 with corresponding criteria of
4.4 x lO-2 ng/1, 4.4 x 10~3 ng/1, and 4.4 x 10~4 ng^Jf respec-
tively.
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Introduction
Aldrin and dieldrin have been two of the most widely
used domestic pesticides. They are chlorinated hydrocarbon
compounds. Although aldrin is used in greater quantity
than dieldrin, aldrin quickly transforms into dieldrin in
the environment. Hence, there is concern with both compounds.
The primary use of the chemicals 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 resis-
tance to the two chemicals and to development and availability
of substitute materials.
Aldrin and dieldrin have been the subject of litigation
bearing upon the contention that these substances cause
severe aquatic environmental change and are potential carcino-
gens. In 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
insertions for termite control, nursery clipping of roots
and tops of non-food plants, and mothproofing of woolen
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textiles and carpets where there is no effluent discharge*
In 1974, cancellation proceedings disclosed the severe hazard
to human health and suspension of registration of aldrin
and' dJie.ldrin use was ordered; production was restricted
fLor. all pesticide products containing aldrin or dieldziit.,
, formulated products containing aldrin and di.el.dtr.ijt
imported, from Europe e.ac.h. year solely for subsurfcace:
.s,Q-iJL lo.ject.ion.- for termite- control.,. Therefore., limits tfrat
.^ aJL.1 neceiving: wa-tejc. uses, must he* placed on addr.in.
el^lr.in:. The li ti.gat'i:on has produced, the. evidentriarcy
for th.e Administrator-' s conclusions that- aldrin/dieldr.in
are c^rc.feno,genic in mice and. rats., approved the Agency's
e-xt-r:a:pAl-a,tion. to humans of. data, derived from tests on an-imal.s,
a.jn;d- .a.f;fi,rmed, the conclusjons that aldrin. and di.eldrin pose-
a SMteafc-a/nivlal. r:lsk of canc.er to humans, which consrtridt.ut.es.
an: "'ijciainent; ha.zacd" to. man..
n- and. dieldrin are white crystralline substances
, ald:mn. melting, at 104°c and die.-ldr.in melting between
to, l;7'7"QGu. Hoth are; soluble in organic solvents with:
thej le-ast soluble of. the- two. The chemical name:
fojrr al.drin i.s 1, 2, 3, 4,, 10", 10-hexachloro-l, 4, 4a, 5,- 8;, 8.a-
hex^ah;y.dEQr-l.,. 4: 5, 8-ex.o-dime.thanonaphathalene. The chemical
name; for dieldrin is 1., 2, 3, 4, 10, lO-hexachloro-6, 7—
epToxvy'-l, 4.-, 4a, 5, 6, 7, 8, Sia-octahydro-endo, exo-1, 4:
5. ,, ai-d;imeit hanonaph t h a 1 ene- .
Aldrin, i.s-, metabolic ally converted to. dieldrin. This,
Qpo.xida.tion has been shown to occur, in several species includ-
ing; mamm-als: and poultry, house^flies-, locusts, soil microorgan-
., a larg^e number of. Lepidoptera species, freshwater.
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fish (Gakstatter, 1968), and a number of freshwater inverte-
brates including protozoa, coelenterates, worms, arthropods,
molluscs, and lobsters. The aldrin molecule is biologically
altered in the environment to a more stable and at least
equally toxic form, dieldrin. Dieldrin is known to be meta-
bolically degraded as shown by Matsumura and Boush (1967)
and Patil, et al. (1972); however, its persistence in the
environment is due to its extremely low volatility (i.e.,
a vapor pressure of 1.78 x 10-7 mm mercury at 20°C) and
low solubility in water (186 ;ug/l at 25 to 29°C) (Int. Agency
Res. Cancer, 1974). In addition, dieldrin is extremely
apolar, resulting in a high affinity for fat which accounts
for its retention in animal fats, plant waxes, and other
such organic matter in the environment. The fat solubility
of dieldrin results in the progressive accumulation in the
food chain which may result; in a concentration in an organism
which would exceed the lethal limit for a consumer.
Many organisms not in direct contact with contaminated
water and sediment accumulate aldrin/dieldrin from the food
supply. This biological concentration results in tissue
concentrations many times those found in the surrounding
environment (Sanborn and Yu, 1973). Concentrations increase
in the food chain reaching the carnivores at the top including
man.
Dieldrin is probably the most stable insecticide among
the cyclodienes (i.e., isodrin-endrin; heptaclor-heptachlor
epoxide) . The time required for 95 percent of the dieldrin
to disappear from soil has been estimated to vary from 5
to 25 years depending upon the microbial flora of the soil
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(Edwards, 1966). Dieldrin applied at 100 ppm has been shown
to persist in soil for more than six years (Westlake and
San Antonio, 1960) , while at 25 ppm in a different soil
type, a 50 percent loss was found at seven years (Nash and
Woolson, 1967). When applied to sandy soil at a rate of
100 ppm, residues could be found 15 years later. Matsumura
and Boush (1967) found that of 577 bacterial isolates collect-
ed from areas heavily contaminated with dieldrin, 10 isolates
would alter dieldrin to two to nine unidentified metabolites.
The microbes were members of Pseudomonas, Bacillus, and
Trichoderma genera. Subsequent microbiological studies by
Wedemeyer (1968) revealed that Aerobacter aerogenes also
will alter dieldrin similarly to 6,7- trans-dihydroxydihydro-
aldrin. Chacko, et al. (1966) tested this capability of
17 species of fungi and actinomycetes. Though most degraded
pentachloronitrobenzene (PCNE) or DDT or both, none degraded
dieldrin.
Patil, et al. 1972, studied the metabolic transformations
of aldrin/dieldrin by marine algae, surface film, sediments,
and water. They found that the insecticide was *not degraded
or metabolized in sea water or polluted waters. Some marine
algal populations 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 Yu (1973).
However, dieldrin can be mobilized from fatty tissue as
demonstrated by Brockway (1973); for example, when fish
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are placed in an environment without dieldrin, there is
an elimination from the tissue (Brockway, 1973). The elimina-
tion rate depends upon the diet with fasted fish eliminating
dieldrin more rapidly than fed fish because of the utilization
of fat stores (Grzenda, et al. 1972).
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
Brockway, D.C. 1973. The uptake, storage and release of
dieldrin and some effects of its release in the fish, Cichlo-
soma 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
residues in tissues of goldfish exposed to a ssingle 'sublethal
dose of 14C-aldrin. Jour. Fish. 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. 101: 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.
Nash, R.G., and.E.A. Woolson. 1967. Persistence of chlorinated
hydrocarbon insecticides in soils. Science 157: 924.
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Patil, K.C., et al. 1972. Metabolic transformation of DDT,
dieldrin, aldrin, and endrin by marine microorganisms. Environ.
Sci. Technol. 6: 631.
Sanborn, J.R., and C.C. Yu. 1973. The fate of dieldrin in
a model ecosystem. Bull. Environ. Contain. Toxicol. 10: 340.
Wedemyer, G. 1968. Partial hydrolysis of dieldrin by Aerebacter
aerogenes. Appl. Microbiol. 16: 661.
Westlake, W.E., and J.P. San Antonio. 1960. Insecticide
residues in plants, animals and soils. Page 105 in. The nature
and fate of chemicals applied to soils, plants, and animals.
U.S. Dep. Agric. 20: 9.
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AQUATIC LIFE TOXICOLOGY*
FRESHWATER ORGANISMS
Introduction
Aldrin and dieldrin are members of a group of synthetic
cyclic hydrocarbons 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 than dieldrin. However, these pesticides are often
considered together since aldrin is rapidly converted in animal or
plant tissue and soil to dieldrin. This conversion is accom-
plished through the addition of an epoxide group to the aldrin
molecule.
Since aldrin is rapidly converted to dieldrin and there are
no adequate data in all the criterion areas, no criterion has been
developed for aldrin. The following discussion is based on diel-
drin data only except where specifically noted.
*The reader is referred to the Guidelines for Deriving Water Qual-
ity Criteria for the Protection of Aquatic Life [43 FR 21506 (May
18, 1978) and 43 FR 29028 (July 5, 1978)] in order to better
understand the following discussion and recommendation. The fol-
lowing tables contain the appropriate data that were found in the
literature, and at the bottom of each table are the calculations
for deriving various measures of toxicity as described in the
Guidelines.
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Acute Toxicity
Sixty-seven acute toxicity tests using dieldrin are reported
in Table 1. The tests were conducted with ten species of fish
ranging from coldwater fish such as the rainbow trout, coho and
chinook salmon to warmwater fish such as the goldfish and carp.
All of the tests were static and none included measured concentra-
tions. The adjustment of a 48-hour LC50 to a 96-hour value was
necessary only for the. exposure of the mosquitofish.
Dieldrin is acutely toxic at low concentrations. Only 10 of
the 67 adjusted LC50 values are greater than 10 ug/1 and a major-
ity of the values below 10 ug/1 are in the range of 0.6 to 5.5
ug/1. There are, however, species differences. The most sensi-
tive fish tested was the rainbow trout with 96-hour LC50 values
between 0.6 ug/1 and 5.4 ug/1. The other salmonids (coho and
chinook salmon) had 96-hour LC50 values of 3.3 and 5.9 ug/l» re-
spectively. The most resistant fishes were the carp and the gold-
(
fish with 96-hour LC50 values of 33 and 22 ug/lr respectively. In
the middle of the range, between the salmonids and the carp, were
fathead minnows (range 9 to 20 ug/D and the bluegill (range 4.8
! I
to 17 ug/D. Special attention should be given to the data on the
guppy in the report by Chadwick and Kiigemagi (1968) concerning
i
the development of a toxicant delivery system. To determine the
efficiency of the system, guppy toxicity tests were conducted over
an extended time period and the data are included in Table 1.
Thirty-eight of the 67 test results are from this study and range
from 1.3 to 5.5 ug/1.
Thirteen fish species were tested and 23 tests were com-
pleted using aldrin. The range of the adjusted values (1.2 to
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25.1 ug/1) is similar to the values obtained for dieldrin. One
test, not included in the range, had a 96-hour LC50 value of 97
ug/lo This test used the mosquitofish which is well-known for
pesticide-resistant wild populations. When the geometric means
from Table 1 are divided by the sensitivity factor 3.9, the
resulting Final Fish Acute Values are 1.6 and 2.4 ug/1 for diel-
drin and aldrin, respectively. Only 11 of the 67 dieldrin tests
are lower than this concentration (1.6 ug/1) and of these 11, 8
are with the guppy (Chadwick and Kiigemagi, 1968) and are balanced
by 30 values above 1.6 ug/1. The other three LC50 values are for
the rainbow trout. These results suggest that the adjustment fac-
tors from the Guidelines are appropriate.
Nineteen acute toxicity test results for dieldrin and inver-
tebrate species are presented in Table 2. All of these tests were
conducted under static water conditions and the concentra- tions
i
were not measured. The adjusted concentrations range from a 96-
hour LC50 value of 0.4 ug/1 for the stoneflies Pteronarcella badia
and Pteronarcys californica (Sanders and Cope, 1968) to 627 ug/1
for the crayfish (Sanders, 1972). This wide range in concentra-
tion of over 1,500 times demonstrates definite differences in in-
terspecific sensitivity to this compound.
Intraspecific variation is apparent for the stonefly and
ostracod data. This variation may have resulted from differences
in experimental procedures used in stonefly testing and Guideline
applications to the ostracod data. Sanders and Cope (1968) deter-
mined a 96-hour LC50 value of 0.5 ug/1 dieldrin at 15°C for the
stonefly Pteronarcys californica. They did not aerate the test
water. Jenson and Gaufin (1964) used aeration and a slightly
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higher test temperature of 15.5°C and determined a much larger
(about 78 times) 96-hour LC50 value of 39 ug/1 for this species.
Since this insect inhabits well oxygenated flowing water the non-
aerated static test may have potentiated toxic effects. Hansen
and Kawatski (1976) report a 24-hour LC50 value of 185 ug/1 and a
i
72-hour LC50 value of 12.3 ug/1 with the ostracod Cypretta
kawatai. These tests were conducted under similar conditions but
were of different duration.
The geometric mean of the dieldrin data, 26 ug/lf was divided
by the sensitivity factor of 21 from the Guidelines to obtain a
concentration of 1.2 ug/1. This concentration is higher than 3 of
the 1:9 adjusted concentrations for the tested invertebrate species;
this result appears to support the procedures in the Guidelines
for the sensitivity factor.
i
Results of 13 acute toxicity tests with aldrin are also pre-
sented in Table 2. Each test was conducted so that data could be
compared with data obtained from similar tests with dieldrin.
Adjusted aldrin 96-hour LC50 values range from 1.1 ug/1 for the
•stonefly (Sanders and Cope, 1968) to 32,609 ug/1 for the scud
i '
(Ga.ufin, et al, 1965). The cladocerans were relatively more sen-
1966). In all other cases the invertebrates were relatively more
sensitive to d.ieldrin. For aldrin, the estimated concentration at
or below the 96-hour LC50 value for 95 percent of all invertebrate
species is 3,8 ug/1•
Acute toxicity tests with aldrin and dieldrin have estab-
lished that these compounds are toxic to aquatic life at low con-
centrations. The data indicate that dieldrin is slightly more
toxic than aldrin for both fish and invertebrates. The Final Fish
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Acute Value is 1.6 ug/1 and the Final Invertebrate Acute Value is
,1.2 ug/1. Because the Invertebrate Acute Value is the lowest, the
Final Acute Value is 1.2 ug/1.
Chronic Toxicity
Two chronic toxicity tests have been conducted with dieldrin.
.One was an embryo-larval exposure using steelhead (rainbow) trout
(Chadwick and Shumway, 1969). This species was the most sensitive
rspecies according to the acute studies (Table 1). The other
chronic exposure was a three-generation study using the guppy
(Roe-lpfs, 1971). Fortunately, the 96-hour LC50 concentration is
.well-established for this fish (Table 1) and is about 2.9 ug/1.
The geometric mean (0.21 ug/1) of the two chronic concentrations
divided by the sensitivity factor (6.7) results in a 95 percent
protection concentration or Final Fish Chronic Value of 0.031
ug/1 (Table 3). Since the two tested species include the most
sensitive and a moderately sensitive species, the calculated con-
centration should confer adequate protection for the non-tested
fish species.
No chronic studies were found for these important animals.
i
Because of the lack of chronic data, it is necessary to reexamine
the invertebrate test results. All of the acute invertebrate
values are greater than the fish geometric mean chronic concentra-
tion of 0.21 ug/1. However, three stonefly species have adjusted
acute values (0.4 to 0.5 ug/1) which are close to the fish geomet"
ric mean value. These data were obtained under static water con-
ditions without aeration and the dieldrin concentrations were not
measured. More meaningful data for assessing the risk of chronic
exposure of dieldrin to stoneflies was obtained by Jensen and
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Gaufin (1966)o They determined a 30-day LC50 value of 2 ug/1/
based on measured concentrations,, for one of the three species, £.
californica (Table 6), in flowing water to which stoneflies are
adapted. A lower 30-day LC50 value of 0.2 ug/1 was also obtained
for another stonefly Acroneuria pacifica. These data indicate
that the insect chronic value might be less than that calculated
for fisho A lower value might be expected because the primary use
of dieldrin was as an insecticide.
After applying the sensitivity factor the Final Fish Chronic
Value is 0.031 ug/1. The extent of protection for the inverte-
brates is unknown but it can be estimated from the acute toxicity
test that many would be safe if exposed at the concentration of
0.031 ug/1.
Plant Effects *
Four dieldrin toxicity tests using three plant species were
found (Table 4). The alga, Scenedesmus quadricaudata, was the
most sensitive species tested with a 22 percent reduction in
biomass after exposure to 100 ug/1 of dieldrin (Stadnyk and
Campbell, 1971). The other species, diatom and water meal, were
affected only at concentrations 100 times higher than the alga.
Since fish and invertebrate species were affected at concentra-
tions 100 times lower than the alga, the plants should be pro-
tected by the animal-derived data.
Residues
Table 5 contains the results of 10 residue studies with diel-
drin. No comparable aldrin data were found. The 10 studies in-
clude plant, invertebrate and fish species. The range of the bio-
concentration factors (BCF) are from 128 for an alga (Reinert,
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1972) to 68,286 for lake trout (Reinert, et al. 1974). All of the
authors (except Reinert, et al. 1974) indicate that an equilibrium
had occurred in their specific study. An examination of the data
in the reports supports the conclusion of the individual authors.
The analysis of the residue data can be divided into two
broad groups, the plant-invertebrate and the fish data. The
plant-invertebrate BCF values range from 128 to 5,558. The two
values representing the algal and diatom community accumulations
are perhaps the most ecologically applicable data in this group.
The studies were conducted in open channels under field conditions
whereas the other algal study was a short-exposure laboratory
test. The invertebrate BCF values show a comparatively low bio-
accumulation potential for the two species.
The fish BCF values range from 2,385 to 68,286. Although all
but one of the authors report that equilibrium had occurred in
each of their exposures, there seems to be a relationship between
length of exposure and total residue accumulation. For example,
guppies exposed for 32 days had a BCF of 12,708 while exposure for
160 to 230 days resulted in a BCF of 28,408. The same relation-
ship may explain the high BCF for the lake trout. The bioconcen-
tration of dieldrin by this species may become greater since the
fish had not reached an equilibrium when 'the study was terminated.
The channel catfish BCF is the lowest of the fish values (Shannon,
1977a,b). This is probably a result of the experimenter analyzing
dorsal muscle rather than whole fish as was done by the others.
The residue limit established by the Food and Drug Adminis-
tration (FDA) for dieldrin in domestic animal feed is 0.03 mg/kg,
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and was used to calculate the Residue Limited Toxicant Concentra-
tion (RLTC). : The FDA domestic animal feed concentration of 0.03
mg/kg .divided by the average fish bioconcentration factor of
.15,482 gives a RLTC of 0.0000019 mgAg or 0.0019 iig/1.
.The .lowest of .the Final Fish Chronic Value (0.0.31 ug/D -,
:Final Invertebrate Chronic Value (none), Final Plant Value (100
,ug-/D ...and 'the RLTC (0.0019 ug/1) is -used to determine the Final
Chronic Value,. Tor dieldrin the Final-Chronic Value is '0.0019
.ug/1.
' Ma:,scellan.e:ous
Data•;presented ..in Table 6 do not conflict with the selection
o.'f 0.001;9 ug/1 -:as .the .Final 'Chronic Value.
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CRITERION FORMULATION
Freshwater-Aquatic Life
Summary of Available Data
The concentrations below have been rounded to two significant
figures.
Final Fish Acute Value =1.6 ug/1
Final Invertebrate Acute Value = 1.2 ug/1
Final Acute Value = 1.2 ug/1
Final Fish Chronic Value = 0.031 v.g/1
Final Invertebrate Chronic Value = not available
Final Plant Value = 100 ug/1
Residue Limited Toxicant Concentration = 0.0019 ug/1.
Final Chronic Value = 0.0019 ug/1
0.44 x Final Acute Value = 0.53 ug/1
The maximum concentration of dieldrin is the Final Acute
Value of 1.2 ug/1 which is based on the more acutely sensitive in-
vertebrate organisms. Since 0.44 times the Final Acute Value (0.44
x 1.2 ug/1 = 0.53 ug/D is not lower than the Final Chronic Value
(0.0019 ug/Df the latter is the recommended 24-hour average con-
centration. No important adverse effects on freshwater aquatic
organisms have been reported to be caused by concentrations lower
than the 24-hour average concentration.
CRITERION: For dieldrin the criterion to protect freshwater
aquatic life as derived using the Guidelines is 0.0019 ug/1 as a
hour 24-average and the concentration should not exceed 1.2 ug/1
at any time.
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Table 1. Freshwater fish acu.te values for aldrin/dieldrin
Adjusted
Bioasaay __Test
'Organism HfeifiS^Jl Cone.**
Rainbow trout, S U
Salmo gairdneri
Rainbow trout,
Salmo Rairdneri
Rainbow trout ,
Salmo gairdneri
Rainbow trout:,
Salmo gairdneri
Coho salmon,
Oncorhynchus kisutch
to
l!j Chinook salmon,
o Oncorhynchus tshawytscha
Goldfish,
Carassius auratus
Carp,
Cyprlnus carpio
Fathead minnow,
Pimephales promelas
Fathead minnow,
I'imephalea promelas
Fathead minnow,
Pimephales promelaa
Fathead minnow,
Pimephales promelas
Fathead minnow,
Pimephales promelas
Fathead minnow,
Pimophales promelas
S
S
S
S
S
S
S
S
S
S
S
S
S
U
U
U
u
u
u
u
u
u
u
u
u
u
Chemicai
Description - •
Dieldrin
90% dieldrin
85% dieldrin
85% dieldrin
85% dieldrin
90% dieldrin
90% dieldrin
90% dieldrin
15% dieldrin
90% dieldrin
90% dieldrin
85% dieldrin
85% dieldrin
85% dieldrin
85% dieldrin
Time
ms>
*
96
96
96
96
96
96
96
96
96
96
96
96
96
96
LCbU
• (^^v
9.9
2.4
1.1
10.8
6.1
41
60
18
18
36
,24 „ .
16
25
LC!>0
(uq/il
5.4
1.3
0.6
0.8
5.9
3.3
22
33
10
10
20
13
9
14
• Reference
Katz, 1961
Macek, et al.
1969
Macek, et al.
1969
Macek. et al.
1969
Katz, 1961
Katz, 1961
Henderson,
et al. 1959
Rao, et al.
1975
Henderson.
et al. 1959
Henderson .
et al. 1959
Tarzwell
Henderson, 1957
Tarzwell &
Henderson, 1957
Tarzwell &
Henderson, 1957
Tarzwell &
Henderson, 1957
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Table 1. (Continued)
Organism
Bioaseay Test Cnemical Tame LC5U
Adjusted
XCiO
jug/11 Keterence
Fathead minnow,
Pimfcphales promelas
Hosquitofish,
Gambusia af finis
Guppy,
Poecilia
Cuppy,
Poecilia
Cuppy,
Poecilia
Guppy ,
Poecilia
CD
jjj Guppy,
i_i Poecilia
Guppy ,
Poecilia
Guppy ,
Poecilia
Guppy ,
Poecilia
Guppy ,
Poecilia
Guppy,
Poecilia
Guppy ,
Poecil la
Guppy,
Poecilia
Guppy,
Poecilia
reticulata
reticulata
reticulata
-rctlTrul-ata
-reticulata
reticulata
reticulata
reciculata
reticulata
reticulata
reticulata
reticulata
reticulata
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
u
u
"
u
u
u
u
u
u
u
u
u
u
u
u
85% dieldrin
957. dieldrin
"Technical
dieldrin
Technical
-->die-ldrin
"Technical
dieldrin
Technical
-d-fcel-drin
Technical
dieldrln
Technical
dieldrin
Technical
dieldrin
• Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
grade
grade
grade
•grade
grade
grade
grade
grade
grade
grade
grade
grade
grade
96
48
-96
96
96
96
96
96
96
96
96
96
96
96
96
23
8
3
4
3
5
3
3
3
3
4
4
4
4
3
*
n
.9
.1
.9
.7
.2
.9
.2
.3
.3
.1
.5
13
3
*
^
2
1
2
2
1
2
2
2
2
2
1
•*
6
.1
.8
.1
.0
.7
.1
.3
.3
.3
.2
.9
Tarzwell &
-Henderson. 1957
Culley &
Ferguson, 1969
Chadwick &
Kiigeraagi. 1968
Chadwick &
Xilgemagl. 1968
Chadwick &
Kligemagl. 1963
Chadwick A
XHgenagl, 1958
Chadwick &
Kiigemagij 19.68
Chadwick &
Kii£emagi. 1968
Chadwick &
Kiigemagi, 1968
Chadwick &
Kiigemagi, 1968
Chadwick &
Kiigeraagi, 1968
Chadwick 6r
Kiigemagi. 1968
Chadwick &
Kiigemagi, 1968
Chadwick &
. Kiigemagi, 1968
Chadwick &
Kiigemagi, 1968
-------
Table 1. (Continued)
Organism
Bioassay Test Chemical
_ ConCj.** Description jhrs)
LC5U
(uy/il
Adjusted
LCbO
hetexence
Guppy,
Poecilia
reticulata
DO
I
!-•
tO
Guppy.
Poecilia
reticulaca
Guppy
Poecilia
reticulata
Guppy,
Poecilia
reticulata
Guppy,
Poecilia
reticulata
Guppy.
Poecilia
Guppy,
Poecilia
reticulata
reticulata
Guppy,
Poecilia
reticulata
Guppy.
Poecilia
reticulata
Guppy,
Poecilia
reticulata
Guppy,
Poecilia
reticulnta
Guppy,
Poecilia
Guppy,
Poecilia
Guppy.
Poecilia
Guppy.
Poecilia
reticulata
reticulata
reticulata
reticulata
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
U
U
U
U
U
U
U
U
U
U
U
U
U
I)
U
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
Technical
dieldrin
grade 96
grade 96
grade 96
grade 96
grade 96
grade 96
grade 96
grade 96
grade 96
grade 96
grade, 96
grade 96
grade 96
grade 96
grade 96
4.7
3.2
2.9
2.6
2.9
2.4
2.6
2.3
2.7
2.3
2.7
2.7
4.8
6.1
3.2
2.6 Chadwick &
Kiigemagi, 1968
1. 7 Chadwick &
Kiigemagi. 1968
1.6 Chadwick &
Kiigemagi, 1968
1.4 Chadwick &
Kiigemagi. 1968
1.6 Chadwick &
Kiigemagi, 1968
1.3 Chadwick &
Kiigemagi. 1968
1.4 Chadwick &
Kiigemagi, 1968
1.3 Chadwick &
Kiigemagi, 1968
1.5 Chadwick &
Kiigemagi. 1968
1.3 Chadwick &
Kiigemagi. 1968
1.5 Chadwick &
Kiigemagi, 1968
1.5 Chadwick &
Kiigemagi. 1968
2.6 Chadwick &
Kiigemagi. 1968
3.3 Chadwick &
Kiigemagi. 1968
1.7 Chadwick &
Kiigemagi. 1968
-------
Table 1. (Continued)
LC50
Adjusted
LCbO
(uq/i) heterence
CD
1
(jJ
Guppy ,
Poecilia reciculaca
Guppy.
Poecilia reciculaca
Guppy,
Poecilia reciculaca
Guppy .
Poecilia reciculatd
Guppy.
Poecilia reciculaca
Guppy.
Poecilia reciculaca
Guppy.
Poecilia reciculata
Guppy.
Poecilia reciculaca
Guppy ,
Poecilia reciculaca
Guppy.
Poucilla reticulaca
Guppy.
Poecilia reciculoCa
Guppy .
Poecilia reciculata
Green uunfish,
l.epomis ey_anellus
Green bunt'isli.
l.epomis eyanellus
Green sun fish,
S
S
S
s
s
s
s
s
s
s
s
s
s
,
s
s
rf-"^T -TT
u
u
u
u
u
u
u
u
u
u
u
u
u
u
u
99+% dieldrin
99+% dieldrin
99+7. dieldrin
99+7. dieldrin
99+7. dieldrin
99+% dieldrin
99+7. dieldrin
99+% dieldrin
99+7. dieldrin
99+% dieldrin
90% dieldrin
Dieldrin
85% dieldrin
85% dieldrin
85% dieldrin
• • i , i'f
96
96
96
96
96
96
96
96
96
96
96
96
96
96
,
96
6.6
5.6
6.1
7.5
10
6.6
6.6
6.9
A. 7
7.5
25
21
6
11
8
3.6
3.1
3.3
4.1
5.5
3.6
3.6
3.8
2.6
4.1
14
11
3
6
A
Chadwick &
Kiigemagi, 1968
Chadwick &
Kiigemagi, 1968
Chadwick &
Kiigemagi, 1968
Chadwick &
Kiigemagi, 1968
Chadwick &
Kiigemagi. 1968
Chadwick &
Kiigemagi, 1968
Chadwick &
Kiigemagi, 1968
Chadwick &
Kiigemagi, 1968
Chadwick &
Kiigemagi. 1968
Chadwick &
Kiigemagi, 1968
Henderson.
et al. 1959
Cairns & Loos,
1966
Tarzwell &
Henderson, 1957
Tarzwell &
Henderson, 1957
Tarzwell &
I.cnomi s c
Henderson, 1957
-------
Table 1. (Continued)
CD
I
Oraarjisro
Bluegill.
Lcpomis macrochirua
Bluegill.
Lepomls macrochlrua
Bluegill,
Lcpomis maci'ochiru3
Bluegill.
Lepomls macrochlrus
Bluegill.
Lepomls macrochirus
Bluegill,
Lepomiu macrochirus
Bluegill.
Lepornia macrochirus
Bluegill.
Lepomis macrochirus
American eel,
AnjjuLHa rostrata
Rainbow crouc,
Salino gairdnerl
Rainbow Crouc,
Sal mo j'.aircl fieri
Rainbow trout,
Saliiio galrdneri
Rainbow trout.
SaInio ga irdneri
Colio salmon,
Oncorliynclms klsucch
Bioataay Test Chemical Time
(nra)
Adjusted.
LCbc LCiu
(ug/it ^ heterence
S
s
s
s
s
s
s
s
s
s
s
s •
s
s
U
U
U
U
U
U
IJ
U
U
U
U
U
U
11
90% dleldrln
85% dieldrin
85% dieldrin
85% dleldrln
85% dieldrin
85% dieldrlp
85% dieldrin
85% dieldrin
Aldrin
Aldrin
88.4% aldrin
95% aldrin
95% aldrin
95% aldrin
88. *4% aldrin
.••i • "f
96
96
96
96
96
96
96
96
96
96
96
96
96
96
9
17
14
8.8
32
18
8
22
16
17.7
3.2
3.3
2.2
45.9
5
9
8
4.8
17
10
4
12
9
9.7
1.7
1.8
1.2
25.1
Henderson,
ec al. 1959
Macek,
ec al. 1969
Macek,
ec al. 1969
Macek,
ec al. 1969
Tarzwell &
Henderson, 1957
Tarzwell &
Henderson, 1957
Tarzwell &
Henderson, 1957
Tarzwell &
Henderson, 1957
RehwoldC,
ec al. 1977
Kacz. 1961
Macek.
ec al. 1969
Macek.
ec al. 1969
Macek,
ec al. 1969
Kacz. 1961
-------
Table 1. (Continued)
03
I
Organism
Chinook salmon,
Oncorhynchua tshawytscha
Goldfish.
Carassius auratua
Carp.
Cyprinus carpio
Carp.
Cyprinus carpio
Fathead minnow,
Pimephales promelas
Fathead minnow,
Pimephales promelas
Banded killifish,
Fundulus diaphanus
Mosquitofish,
Gambusia affinis
Mosquitofish,
Gambusia affinis
Guppy,
Poecilia reticulata
Guppy,
Poecilia reticulata
Uhite perch,
Koccjuis americanus
Striped bass,
Morone saxatilis
lilucgill.
Lcpomls macrochirus
BluegiJl.
l.upoiiiis macrochirus
assay
hod*_
S
S
S
S
S
S
S
S
S
S
S
S
S
S
S
Test
Cone .**
U
U
U
U
U
U
U
U
U
U
U
U
U
U
U
Chemical
Description
88.4% aldrin
88.4% aldrin
30% aidrin
Aldrin
88.4% aldrin
88.4% aldrin
Aldrin
95% aldrin
Aldrin
88.4% aldrin
Aldrin
Aldrin
Aldrin
88.4% aldrin
95% aldrin
Time
(nrs)
96
96
^
96
96
96
96
96
48
24
96
96
96
96
96
96
LCbo
(Uj/11
6.1
32
3.7
4
37
32
21
36
270
37
20
42
10
15
7.7
Adjusted
LC'jO
(uq/l)
3.3
17
2
2.2
20
17
11
16
97
20
11
23
7
8
4.2
Keterence
Katz, 1961
Henderson.
et al. 1959
Rao, et al.
1975
Rehwoldt,
et al. 1977
Henderson,
et al. 1959
Henderson,
et al. 1959
Rehwoldt ,
et al. 1977
Culley &
Ferguson, 1969
Krieger &
Lee. 1973
Henderson ,
et al. 1959
Rehwoldt,
et al. 1977
Rehwoldt,
et al. 1977
Rehwoldt,
et al. 1977
Henderson,
et al. 1959
Macek,
et al. 1969
-------
Table 1. (Continued)
00
I
M
a\
test cnemicai Time
Method* couc.** Description tnta)
Adjusted
'LCb'i, LC!>0
lug/I) tuq/il heterfence
Bluegill, S 11 95% aldrin 96
Lepomis macrochirus
Bluegill. S U 95% aldrin 96
Lepomis macrochirus
5.8 3.2 Macek-.
et al. 1969
A. 6 2.5 Macek,
et al. 1969
* S - static
** I) = unmeasured
5.9
Geometric mean of adjusted valuesi Dieldrin » 5.9 pg/1 LL± «• 1.6
•
Aldrin - 9.4 »,g/l - 2.4 Mg/l
-------
Table 2. Freshwater intertebrate acute values for aldrtn/dieldrln
Bioassay Test Chemical Time
Method* Cone .** Description Hire)
LCt>0
Adjusted
LCbO
(ug/l> Heterence
03.
-J
Cladoceran, S U
Daphnia carinata
Cladoceran, S U
Daphnia pulex
Cladoceran, S U
Simocephalus serrulatus
Cladoceran, S U
Simoccphalus serrulatus
Ostracod, S U
Cypretta kawatai
Ostracod, S U
Cypretta kawatai
Isopod, S U
Asellus breicaudus
Scud. S U
Gammarus fasciatus
Scud. S U
Gammarus fasciatus
Scud, S U
Ganimarus lacustris
Scud, S U
Gammarus lacustris
Glass shrimp, S U
Palaemonetcs kadiakensis
Crayfish, " S U
Orconecies nais
Mayfly, S U
Ephemeralla grandis
Dieldrin
Technical grade 48
dieldrin
Dieldrin
Dieldrin
Dieldrin
48
48
43
99+7. dieldrin 24
99+% dieldrin 72
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Dieldrin
96
96
96
96
96
96
96
96
130 110 Santharara,
et al. 1976
250 212 Sanders &
Cope, 1966
240 203 Sanders &
Cope. 1966
190 161 Sanders 6.
Cope, 1966
185 41 Hansen &
Kawatski, 1976
12.3 6.3 Hansen &
Kawatski, 1976
5 4 Sanders, 1972
640 542 Sanders, 1972
600 508 Sanders. 1972
700 593 Gaufin, et al.
1965
460 390 Sanders, 1969
20 17 Sanders, 1972
740 627 Sanders. i972
8 7 Gaufin, et al.
1965
-------
Table 2. (Continued)
Adjusted
Uioabsay.
Organism ' Mgtnod*'
Stonefly, S
Acroneuria paclfica
Sconefly.
Claasaenta sabulosa
Stonefly,
Pteronarcella badia
Stonefly,
Pteronarcys callfornlca
Stonefly.
Pteronarcys callfornica
~"~~L- ' " . e - -_ .
CD
!
h* Cladoceran,
* Daphnia pulcx
Cladoceran,
Slmocephalus aerrulatua
Cladoceran,
Simocephalua serrulatus
Isopod.
Asellus bretcaudus
Scud .
Gammarus fasciatus
Scud.
Ganuuarus fasciatus
Scud,
Gammarus lacustris
Scud,
Ganuuarus lacuatris
Class shrimp.
S
S
S
S
S
S
S
S
S
S
S —
S
S
_ Test Cliemicai
Cone t** bebcri ption-
U
U
U
u
u
u
u
u
u
I)
u
- u -
u
u
100% dieldrln
Oieldrin
Dleldrin
100% dieldrln
Technical grade
dieldrin '
Aldrin
Aldrin
Aldrin
Aldrin
Aldrin
Aldrin
Aldrin
Aldrin
Aldrin
Aldrin
Time
(lira)
96
96
96
96
96
u
48
48
96
96
96
96
96
96
LCbu
24
0.58
0.5
39
0.5
28
23
32
8
4,300
5.600
38 . 500
9.800
50
^q/il
20
P. 5
0.4
33
0.4
24
19
27
7
3.642
4,743
32,609
8.301
42
inference
Jensen & .
Gaufln. 1964
Sanders &
Cope, 1968
Sanders &
Cope , 1968
Jensen &
Gaufin, '1964
Sanders &
Cope, 1968
Sanders &
Cope, 1966
Sanders & .
Cope. 1966
Sanders &
Cope, 1966
Sanders, 1972
Sanders. 1972
Sanders, 1972
Gaufln, et al
1965
Sanders, 1969
Sanders, 1972
Palaeiuonctett kadiakensia
-------
Table 2. (Continued)
Adjusted
Organism
Mayfly.
Ephemerella grandis
Stonefly,
Acroneuria pacifica
Stonefly.
Pteronarcys californica
Sconefly.
Pteronarcys californica
* S = static
*"* U = unmeasured
CO
1 Geometric mean of ad
Bioassay
HetJiod*
S
S
S
S
Ttst Chemical
ConCj.** Description
U Aldrin
U Aldrin
U 93% aldrin
U Technical grade
aldrin
lusted values:
Dleldrln = 26 wg/l
Time LCbO
(t>r§) (ug/il
96 9
"96 143
96 180
96 1.3
^ - 1.2 pg/1
LCbO
(uq/i) heterence
8 Gaufin, et al.
1965
121 Jensen &
Gaufin. 1964
152 Jensen &
Gaufin, 1964
1.1 Sanders &
Cope, 19ba
Aldrin = 80 Mg/l
80
3.8 Mg/l
-------
00
to
o
Table 3. Freshwater fish chronic values for aldrln^dieldrin
CJuonic
Limits value
Orqaniarii ' jest*
Steelhead trout, E-L 0.12-0.39 Q.ll** Chadwlck & Shniftway> 1969
Salmo palrdnert
Guppy, LC 0.2-1.0 0.4** Roelbfs, 1971
Poecilla reciculata
*E-L - embryo-larval, LC - life cycle or partial life cycle
**All chronic data are for dleldrin .
Geometric mean of chronic values = 0.21 cg/1 0.H - 0.031 (.g/1
Lowest chronic value = 0.11 vg/1
-------
Table A. Freshwater plant effects for aldrin/dieldrin
CO
1
M
Organism
Alga.
Scenedesmus
quadrlcauHata
Diatom.
Navicula seminulum
Water meal,
Uolffia papulifera
Mater meal ,
Uolffia papulifera
Lowest plant value for
Concentration
Effect (uq/it
22% reduction
In blomass in
10 days
50% reduction
In growth In
5 days
Reduced popula-
tion growth in
12 days
Reduced popula-
tion growth in
12 days
dieldrin - 100 Mg/l
100
(dieldrin)
12,800
(dieldrin)
10.000
(dieldrin)
10.000
(aldrin)
Reference
Stadnyk £> Campbell. 1971
Cairns, 1968
Worthley & Schott, 1971
Worthley & Schott. 1971
Lowest plant value for aldrin = 10,000 pg/1
-------
fattier S.- Fice'sl^te"*' i-Gsiditi-a fclf
Onianism
Time
12$
SceneJfesmus bbltqju'us
Community domiriaued by the alga;
tribonema minus
Community of alga arid diatoms
including Sfcieeoctorilum
subsecuridum, SyneJria ulna,
i.-iSlf
4-6 wks Rtis£ & tfc'ifttir'ej i970f
4-6 wks Rose & Mclhtire, 1970
Epitnemifr sorex, Coccohe tg.
placeritula vaf •' englyptari
and Ni tzachi-a S£.
Cladoceran, 1,395 3 Rftihert, 1972
Paphnia magna
Freshwater mussel, 1,030 7-21 Bedford & Zabitc. I973f
Lampsi lis siliguoidea
Ci> .
1 Steelhead trout (newly hatched 3,225 35 Ghadwick & Shumway, 1969
^ alevin)f,
Sal mo gairdneri
Lake trout (year ling) ,-
Salvelinus riamaycttsh
Channel catfish',
ictatluru's punctatus
Channel catfish,
Ictalurus punctatus
Guppy
Poecilia rcticulata
Cuppy ,
Poecilia reticulata
Organism
Man
68,286** 152 Reiner t. ec al. 1974
2,385*** 70 Shatiiioh, 19776
2,993*** 28 Shannon. 1977a
9,862 32 Reinert., 1972
28.787 160-230 Rbelofs. 1971
Maximum Permissible Tissue Concentration
Concentration
Action Level or Effect (ing/kg) Reference
Fish and shellfish - 0.3 FDA Admin. Guideline
smoked,- frozen or canned
7420.08
-------
Table 5, (Continued)
03
I
M
LO
Organism
Man
Domestic animals
Rainbow trout,
Sal mo gairdneri
Rainbow trout,
Salmo gairdneri
Act ion" Level or Effect
Fish and shellfish -
raw edible portion
Animal feed
Altered amonia
detoxifying mechanism
- Altered phenylalanine
metabolism
Concentration
(mg/kg)
0.3
0.03
0.36 of
diet
0.36 of
diet
Reference
FDA Admin. Guideline
7420.09
FDA Admin. Guideline
7426.04
Mehrle & Bloomfield.
1974
Mehrle & DeClue.
1972 '
* All bioconcentration factor data are for dieldrin
** May not be at equilibrium
***Data are for dorsal muscle.
Geometric mean bioconcentration factor for all species =3,238
Geometric mean whole fish bioconcentration factor - 15,482
Lowest residue concentration = 0.03 nig/kg
;
-,482
0.0000019 mg/kg or 0.0019 ng/1
-------
Table 6.
freshwater data for aldrinj'dieldrin
Organism '
Test
Result
Amoeba,
Acanthamoeba
casFeTlanti
Tubtficids (mixture),
Tubifex and
timrioOrilua
Aquatic insects
Oieldrln
6 days No effect on survival 10.000 Prescott. et al. 1977
96 hrs LC50
6 mos Bioconcentratlon in
naturally exposed
animals
Stonefly. 30 days
Pteronarcys californica
CD
K>
it^.
Stonefly.
Acroneuria pacifica
Midge .
Chironomus tentans
Rainbow trout,
Salino gairdneri
30 days
24 hrs
17-23
days
Rainbow trout,
Salmo gairdneri
LC50
LC50
I.C50
Lethal muscle tissue
bioconcentration
1AO days Altered concentrations
of 11 amino acids
6,700 bitten & Goodnight, 1966
4.620 Buikley. et al. 1974
2 Jensen & Gaufin. 1966',
0.2 Jensen & Gaufin. 1966
0.9 Karnak 6. Collins. 1974
3,348 Holden. 1966
1 mg/kg/ Mehrle, et al. 1971
wk
Rainbow trout,
Salino gairdneri
Rainbow trout,
Sano
Carp. ........
Cypriiuis carpio
Channel catfish.
Ic-talurus ptinctatus
IIlack bnl]head.
lcr,-ilt
140 days Increased lipid
control
0.2 mg/kg/ Macek. et al. 1970
wk
168 days Equilibrium bioaccumu- 0.2 mg/kg/ Macek, et al. 1970
lation of 1.05 ppm v;k
96 hrs 1007. mortality of
embryos
210 days Reduced growth
36 hrs I.C50
5,000 Halone & Blaylock, 1&70-
4 pg/g of Argyle, 1975
diet
(dry wt.)
2.5 Kcrguson. c-t al. 1965
-------
Table 6. (Continued)
Organism
Green sunfish,
Leggings cyanellus
Green sunfish,
Teet
Kesuit
Ill hrs Concentration In blood 5.65 wg/g Hogan & Roelofs, 1971
at death
111 hrs Concentration in brain 10.31 ug/g Hogan & Koelofs. 1971
at death
Walleye, embryonic Behavioral aberrations
Stizostedion vitreum atage of of yolk sac fry
develop.
Toad (tadpoles),
Bufo woodhoust
Frog (tadpoles),
Pseudacris triseriata
96 hrs I.C50
96 hrs I.C50
12.2 Hair. 1972
150 Sanders, 1970
100 Sanders, 1970
CO
Amoeba,
Aienthamoeba
casteTIani
Cladoceran,
Daphnia magna
Mayfly,
ll£xageni_a bilineata
Stonefly,
Pteronarcys californica
Stonefly ,
Acroneuria
Aldrin
6 days No effect on survival 10,000 Prescott, et al. 1977
Midge,
Chiconomus sp.
Carp,
Cyp_r_hois_ carpio
Black bullhead,
Ict.'ilurus mel as
blaegill,
macrochirus
3 days
3 days
30 days
30 days
3 days
Bioconcencration
Bioconcencration
LC50
LC50
Bioconcentration
14.100 Johnson, et al. 1971
1 6.300 Johnson, et al. 1971
2.5 Jensen i< Gaufin, 1966
22 Jensen 6> Gaufin, 1966
4,600 Johnson, et al. 1971
Significant increase of 180 McBride & Richards, 1971
sodium in profused gill
36 hrs LC50
12.5 Ferguson, et al. 1965
Aldrin 50% inhibition 30
dose or Na+-K+ ATPase
Yap. et al. 1975
-------
'fable 6.
Test flesult
Organism Duration
Lowesic aldrin value = 2.5
Toad (tadip,ole§}, 9$hrs( LC5Q \$Q ganders. 19.7.0.
t>itfo wobdhoua'ii " '" ••':-;-- •-••>
LovjesiU dieldrln value -0.2 yg/
6v
-------
SALTWATER ORGANISMS
Introduction
Aldrin and dieldrin are chlorinated cyclodiene compounds that
have in the past, been two of the most widely used insecticides.
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. Aldrin and dieldrin are
often considered together, because aldrin is rapidly converted to
dieldrin by metabolism by plants and animals or by photo- decom-
position. Therefore, although aldrin and dieldrin are considered
separately for purpose of comparison, dieldrin is of the greater
concern in the aquatic environment.
The acute toxicities of aldrin and dieldrin and the persis-
tence and bioaccumulation potential for dieldrin have been studied
•
using estuarine plants and animals. Bioaccumulation by estuarine
organisms and/or subsequent transfer to other animals in estuarine
food-webs have been documented in field-studies and laboratory
experiments. Long-term test results indicate that dieldrin is
chronically toxic to estuarine fishes and crabs, although the
exact mechanism of toxicity is not known.
Acute Toxicity
All species of saltwater fish tested were sensitive to acute
exposures to aldrin (13 species) or dieldrin (16 species) .(Table 7)
In flow-through exposures, the unadjusted 48- or 96-hour LC50
values for six fishes ranged from 2.0 to 7.2 ug aldrin/1 (Butler,
1963; Earnest and Benville, 1972; Korn and Earnest; 1974; and
Lowe, data sheets). The unadjusted acute LC50 values for eight
B-27
-------
fishes exposed to dieldrin differed and ranged from 0.66 to 24.0
ug/1 in flow-through tests (Butler, 1963; Earnest and Benville,
1972; Korn and Earnest, 1974; Lowe, data sheets; Parrish, et al.
1973; Schoettger, 1970; and Wade, 1969). Generally, LC50 values
for ^aldrin are slightly higher than those for dieldrin in tests
where the same species were tested, but for practical purposes,
the acute toxicities for these two chemicals can be considered the
same •..
i *
Es.tuarine invertebrate species are acutely sensitive to both
1 i
aldrin and. dieldrin, but there is greater differences in reported
LC50 values for these species than for fishes (Table 8).' Unad-
justed invertebrate LC50 or EC50 values ranged from 0.37 to 33.0
ug aldrin/1 and 0.28 to 240.0 ug dieldrin/1. The most sensitive
species tested was the commercially important pink shrimp; the
24-h:our LC50 value for aldrin was 0.37 ug/1, while the 48-hour
LC50 value for dieldrin was 0.28 ug/1 (unmeasured), and the 96-
ho.ur LC50 value was 0.7 ug/1 (measured) in flowing water exposures
(•Lowe, data sheets; Parrish., et al. 1973). Other crustaceans were
less sensitive and their acute LC50 values ranged from 3.0 to
.240.0 ug/1 (Butler, 1963; Lowe, data sheets; Parrish, et al. 1973;
Schoettger, 1970).
Acute toxicity test conditions can affect the results of
tests with fishes and invertebrates. For example, LC50 values
based on static exposures of aldrin or dieldrin with three fish
and two invertebrate species are higher than LC50 values based on
flow-through exposures where comparable data are available
(Earnest and Benville., 1.072; Eisler, 1969, 1970b; Lowe, data
sheets; and Parrish, et al. 1973). In addition, LC50 values for
B-28
-------
dieldrin based on unmeasured concentrations were higher than those
based on measured concentrations in tests with sheepshead minnows
and two shrimp species (Eisler, 1969; Parrish, et al. 1973).
Therefore, if relative sensitivities of species are to be under-
stood, knowledge of test procedures is necessary.
Chronic Toxicity
No entire life-cycle or embryo-larval tests have been re-
ported for aldrin or dieldrin. However, results (Table 11) of
long-term exposures of invertebrate species and a fish species to
dieldrin in food (Klein and Lincer, 1974) or water (Epifanio,
1971; Lane and Livingston, 1970) indicate a need for such data.
The LC50 value of dieldrin to the sailfin molly after 34 weeks of
exposure was approximately one-fourth that after 48-hours (Tables
7 and 11). Fiddler crabs (Uca pugilator) fed 100 ng dieldrin/g
for 15 days (Table 11) demonstrated unusual running behavior
(Klein and Lincer, 1974).
Plant Effects
Information on the sensitivity of aquatic plants, including
algae and rooted vascular plants, indicates that they are much
less sensitive than are fish and invertebrate species. Product-
ivity and growth rates were reduced at concentrations of approxi-
mately 950 to 1,000 ug/1 in three 4- to 36-hour static tests using
one alga and mixed-population communities (Batterton, et al. 19'71;
Butler, 1963).
Residues
Bioconcentration factors (BCF) for dieldrin (Tables 10 and
11) range from 400 to 8,000 for fish or shellfish (Epifanio, 1973;
Lane and Livingston, 1970; Mason and Rowe, 1976; Parrish, 1974;
B-29
-------
and Parrish, et al. 1973). Bioconcentration factors for oysters
were higher for long exposure periods because dieldrin concentra-
tions in tissues reached steady-state after extended periods
(several weeks) of exposure (Mason and Rowe, 1976; Parrish, 1974;
Parrish, et al. 1973). Therefore, long exposures are necessary to
attain steady-state bioconcentration factors. After 34 weeks of
exposure to dieldrin, sailfin mollies exhibited BCF's of 3,867 to
4,867 in muscle; BCF's 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 dieldrin-
free saltwater (Parrish, et al. 1973). Concentrations in edible
tissues- were slightly less (about 15 percent) than concentrations
in whole spot; however, concentrations in liver were two to 13
times that in spot muscle.
Data Interpretation and Use of Guidelines
Acute toxicity of aldrin and dieldrin will be underestimated
by static tests and by toxicity tests in which the concentration
of aldrin or dieldrin is not measured by chemical analysis. After
applying adjustment factors for test conditions, the variability
in sensitivity of fishes to aldrin and dieldrin was reduced so
they differed by less than a factor of 50 for all species. When
the geometric mean of the LC50 value for aldrin is divided by the
Guideline's species sensitivity factor of 3.7, a value of 1.4 ug/1
results. The Guidelines adjustment factors for test conditions
and species sensitivity seem reasonable because none of the geo-
metric mean adjusted LC50 values for any species is lower than the
B-30
-------
Final Fish Acute Value of 1.4 ug aldrin/1 although some are close.
The Guidelines are designed to obtain a Final Acute Value that
provides an estimate of an LC50 value that is less than that of 95
percent of all fish species. When the geometric mean of the LC50
value for dieldrin is divided by the Guidelines species sensi-
tivity factor of 3.7, a Final Fish Acute Value of 0.85 ug/1 re-
sults,, This value is lower than the geometric mean adjusted LC50
values for 13 of 16 species tested. Therefore, since the Guide-
lines are designed to provide a Final Fish Acute Value which is
lower than or equal to the LC50 value of 95 percent of the
species, the test conditions and sensitivity adjustment factors
appear appropriate.
Invertebrate acute values must also be adjusted for test con-
ditions and species sensitivities. When the Final Invertebrate
Acute Value is obtained from the geometric mean of the adjusted
LC50 values divided by the species sensitivity factor of 49, a
Final Invertebrate Acute Value of 0.084 ug aldrin/1 results. The
adjustment factors seem reasonable because the geometric mean ad-
justed LC50 values for six of the seven tested species are greater
than the Final Invertebrate Acute Value; the adjusted LC50 value
of 0.074 ug/1 for pink shrimp is only slightly lower.
The geometric mean of the adjusted LC5C values for dieldrin,
when divided by the species sensitivity factor of 49, gives a
Final Invertebrate Acute Value of 0.16 ug dieldrin/1. The adjust-
ment factors seem reasonable because the geometric mean adjusted
LC50 values for seven of the eight tested species are greater than
B-31
-------
the Final Invertebrate Acute Value.; the adjusted LC50 value for
one test with pink shrimp was less than the Final Invertebrate
Acute Value.
Dieldriri was bioconcentrated in edible portions of fish and
shellfish by; 2,000 to 8,000 times the concentration in water. The
acceptable residue level, 0.03 ug/g for animal feed, divided; by
the geometric mean- bioconcentration factor of 4,367 for whole
fish,, gives* a' Residue Limited Toxicant Concentration (RLTC) of
0.,006:9 ug/'l«
B-32
-------
CRITERION FORMULATION
Saltwater-Aquatic Life
Summary of Available Data
The concentrations below have been rounded to two significant
figures.
Final Fish Acute Value = 0.85 ug/1
Final Invertebrate Acute Value = 0.16 ug/1
Final Acute Value = 0.16 ug/1
Final Fish Chronic Value = not available
Final Invertebrate Chronic Value = not available
Final Plant Value = 950 ug/1
Residue Limited Toxicant Concentration = 0.0069 ug/1
Final Chronic Value = 0.0069 ug/1
0.44 x Final Acute Value = 0.070 ug/1
No saltwater criterion can be derived for dieldrin using the
Guidelines because no Final Chronic Value for either fish or
invertebrate species or a good substitute for either value is
available.
However, results obtained with dieldrin and freshwater
organisms indicate how a criterion may be estimated. For
freshwater organisms the Final Fish Chronic Value divided by the
Final Fish Acute Value is 0.031/1.6 = 0.019. When this value is
multiplied times the saltwater Final Fish Acute Value, an
estimated Final Fish Chronic Value of 0.85 x 0.019 = 0.016 ug/1 is
obtained. Therefore, the Final Chronic Value of 0.0069 ug/lf
based on the RLTC, should not cause adverse chronic effects on
fish or invertebrate species.
B-33
-------
To. estimate: a criterion for dieldrin, the maximum concen-
tration is. the Final, Acute Value of: 0.16- ug/1 and the 24-hour.
average concentration- is the Final Chronic Value of- 0.0069 ug/1.
No important': adverse effects; on saltwater aquatic organisms have
been, re.pQ,r,teid to- be; caused; by concentrations" lower* than the
2.4.--hour, ave.rag.ei- concentration-:.
GHfcTEKEON:; Eor; dieldrin the/ criterion to. protect saltwater
aquatsiar Ifijffe as§ decdived. us-ing- procedures: other than the Guidelines
iissi Oiiv03016^ ug^lt as'j as 2;4-^hour. averager and- the concentration- should-
nts.tr reaccee'd,; Oi..16ft ug/1; at any time:.
-------
Table 7. Marine fish acute values for aldrin/dieldrin
03
1
U)
cn
B
Organism H
American eel,
Anguilla rostrata
Mummlchog,
Fundulus heteroclitus
Mummichog,
Fundulus heteroclitus
Striped killifish.
Fundulus majalis
Atlantic silverside,
Menidia menidia
Threespine stickleback,
Gasterosteus aculeatus
Threespine stickleback,
Gasterosteus aculeatus
Striped bass,
Morone saxatilis
Spot,
Lciostomus xanthurus
Shiner perch,
Cyma togas ter aggregata
Shiner perch,
C^iiiato^aster aggregata
Dwarf perch,
Micrometrus minimus
Dwarf perch,
MLcrometrus minimus
111 uehedil ,
Thalassoma bifasciatum
Uhite mullet,
Mu^il cureina
ioaeeay
etfiod*
S
S
S
S
S
S
S
FT
Ft
S
FT
S
FT
S
FT
Teat
Cone***
U
U
U
U
U
U
U
U
'U
U
U
U
U
U
U
Time
Ihra)
Aldrin
96
96
96
96
96
96
96
96
48
96
96
96
96
96
48
LCSO
lug/l|
5.0
8.0
4.0
17.0
13.0
39.8
27.4
7.2
3.2
7.4
2.26
18.0
2.03
12.0
2.8
Adjusted
LC50
2.73
4.37
2.19
9.29
7.11
21.76
14.98
5.54
2.0
4.05
1.74
9.84
1.56
6.56
1.75
heier fence
Eisler, 1970b
Eisler, 1970b
Eisler. 1970a
Eisler, 1970b
Eisler, 1970b
Katz, 1961
Katz. 1961
Korn & Earnest. 1974
Lowe, undated
Earnest & Benville,
1972
Earnest & Benville,
1972
Earnest & Benville,
1972
Earnest & lienville,
1972
Eisler, 1970b
-------
7, (Continued)
vTest.
Mttjiod*_ Cone,** Jii£S)
'Adjusted
.cfc
03
1
co
CT\
Striped mullet,
Mujjil cephalua
Striped jnulJlet,
Mugil cepjialus
Northern puffer-5-'
Sphaeroidus maculatus
American eel,' :
Anpuilla rostrata
Chi'iro'ok aalnion,'' ' :
Oncorhyncluis tshawytacha
Sheepsh&ad minnow. •
Cyp_rinodou varie^ntua
Sheejpuheud minnow, •"
Cygrinodon vuriegatus
SheepsheiH minnow.
Cygrinodon varie^-atua
Munmiichog,
Kiindulus heteroclltus
Munuiiichog , • ' •
Kniululus heterocll tua
Mimimi cliog ,
Kundtilti.s heterocll tua
Striped killifish.
l-'undnlus majalis
Sail fin molly,
I'oucilia Ijlipinna
A 1 1 antic si Iverside,
S
,
FT
S
S
PC
S
FT
FT
S
S
S
S
S
S
U
U
U
U
U
U
U
M
U
U
U
U
U
U
96
48
_,.:,
96
Dieldrin
96
96
48
48
96
96
96
96
96
48
96
100.0
; ;-,
2.0
.•: ••'
36.0
0.9
1.47
5 . 82***
24.0
10.0
5.0
16.0
5.0
4.0
10.81***
5.0
54.67
1.25
19.68
0.49
1.13
2.58
15.0
10.0
2.73
8.75
2.73
2.19
4.79
2.73
Eisler. I970b
f .-;.... ••••• :-:i.- .i
Lowe, undated
' • ^ - • •-* .'
Eisler. 1970b
Eisler. i970b
Schoettger, 1970
Wade. 1969
Lowe, undated
Parrish. et al. 1973
Eisler. 1970a
Eisler. 1970b
Eisler. 1970b
Eisler. 1970b
Wade. 1969
Eisler. 1970b
tloaiilla
-------
7. (Continued)
Bioassay
Oiqjinism Method*
Threespine stickleback,
Gasterosteus aculeatus
Threespine stickleback,
Gasterosteus aculeatus
Striped bass.
Morone saxatilis
Spot,
Leiostomus xanthurus
Shiner perch,
Cymatogaster aggregata
Shiner perch.
^ C^r.ia togas ter aggregata
-j Dwarf perch,
Micrometrus minimus
Dwarf perch,
Micrometrus minimus
Bluehead ,
Tlialassoma bifasciatum
White mullet,
Mugil curema
Striped mullet,
Mui>il cephalus
Striped mullet,
Mugil cephalus
Striped mullet,
Mugi 1 cepliai us
Striped mullet.
S
S
FT
FT
S
FT
S
FT
S
FT
FT
FT
FT
S
Teat
Conc^**
U
U
U
U
U
U
U
U
U
U
U
U
U
U
Time
(lirst
96
96
96
24
96
, 96
96
96
96
48
48
48
48
96
LC50
(iiq/j \
15.3
13.1
19.7
*
3.2
3.7
1.5
5.0
2.44
6.0
7.1
3.2
3.2
0.66
23.0
Adjusted
LCbO
Juq/1)
8.36
7.16
15.17
1.63
1.53
1.16
2.73
1.88
3.28
4.43
2.0
2.0
0.41
12.57
Kfeierei.ce
Katz. 1961
Katz, 1961
Korn & Earnest, 1974
Lowe, undated
Earnest & Benville,
1972
Earnest & Benville,
1972
Earnest & Benville,
1972
Earnest & Benville,
1972
Eisler, 1970b
Butler, 1963
Lowe, undated
Lowe, undated
Lowe, undated
Eisler, 1970b
-------
Organism
T. (Continued)
Bipassay Teat
Method* ;
Northern puffer,
Sphaeroldfcs maculatus
U
LC50
Adjusted
1 •• •--
34,0 18.59 Elsier. 19,70b
•:. I:.'-' ' * 1 - - ' A ••" •' '-
CO
I
u>
op
* S = static; FT f flow-through
** M = measured; U ° unmeasured
***Geotnecric mean of 18 means
Geometric mean of adjusted values:
aldrin = 5.2 jig/1
dieldrin =3.2 pg/1
1.4
-------
Table 8. Marine invertebrate acute values :for aldrin/dleldrln
t
Orcjanisjji £
Eastern oyster,
Crassostrea virginica
Sand shrimp,
Crangdon septemspinosa
Hermit crab,
Pagurus longicarpus
Crass shrimp,
Palaemonetes vulgaris
Korean shrimp,
Palaemon macrodactylua
CD
^ Korean shrimp,
(£> Palaemon macrodactylus
Pink shrimp,
Penacus duorarum
Blue crab (juvenile) ,
Callinecte£ sapidus
Eastern oyster.
Crassostrea virginica
Eastern oyster,
Crassostrea virginica
Eastern oyster,
Crassostcea virginica
Eastern oyster.
Crassostrea virginica
Sand shrimp,
Ik' rin if crab,
licassay
FT
S
S
S
S
FT
FT
FT
FT
FT
FT
Ft
S
S
Test
Cone.*
U
U
U
U
U
U
U
U
•
U
U
U
M
U
U
Time
Aldrin
96
96
96
96
96
96
24
48
Dieldrin
96
24
24
96
96
96
LC50
25.0***
8.0
33.0
9.0
0.74
3.0
0.37
23.0
34 . 0***
15.0***
240 . 0***
31.2***
7.0
18.0
Adjusted
LCbO
19.25
6.78
27.95
7.62
0.63
2.3
0 . 074
7.62
26.2
3.0
48.0
31.2
5.9
15.2
Kctfctence
Butler. 1963
Eisler. 1969
Eisler. 1969
Eisler. 1969
Schoettger. 1970
Schoettger, 1970
Lowe, undated
Lowe, undated
Butler, 1963
Lowe, undated
Lowe, undated
Parrluli. ct al. 1973
Eisler, 1969
Eisler, 1969
Pa^urus lon^lcjirpus
-------
Table 8. (Continued)
Biodusay Teat
Method*
Time
Adjusted
LC50 IX. 50
ii__ hfeterenre
(irass shrimp.
i'alaeiiioiietes vulj-.irla
(iraaa shrimp,
I'ijiaeiiionetes pujjlo
Korean shrimp,
I'alaeiiion macrodacljlus
Korean shrimp.
I'.-j 1 ciiMiiuu iiiacrodactyl us
1'Jnk shrimp.
Penaeus duo fa rum
I'lnk shriinp.
to Fenaeus duoraruin
1
** Brown shrimp,
Fencieus a^tecus
lilue crab (juvenile).
Calllnccles :>a(>idus
s u
FT M
S U
FT U
FT M
FT U
FT U
FT U
96 50 ;0
96 8.64
96 16.9
96 6.9
96 0.7
48 0.28
48 3 . 2
48 240.0
42.4
8.64
14.3
5.3
0.7
0.093
1.06
79.5
Elsler. 1969
Parrlsh, et al.
Sehoettger. 1970
Schoettger, 1970
Parribli , et al .
Lowe, undated
Lowe, undated
Lowe, undated
1973
1973
* S - static; FT - flow-through
** M = inc.-
-------
00
I
9. Marine plant effects for aldrin/dieldrin
Organ lain
Alga.
Agmenellum
quadrupl Lea turn
Phytoplankcon
community
Ettect
Reduced
growth
rate
84.6-84.8%
decrease In
productivity
after 4 hrs
Concentration
(uq/ij Reference
950*: Batterton, et al. 1971
1,000**' Butler, 1963
* Dieldrin
** Aldrin
Lowest plant value =, 950 pg/1
-------
J
Table 10, Marine residues for aldrin/dieldrin
Organism
Eastern oyster,
Crassostrea virgintca
Crab,
Leptodius floridanus
Sailfin molly,
Pqecilia latipinna
Spot,
I.elostomus xanthurus
Organism
Man
to
1
M Man
Domestic animals
Bioconcentration Factoi
TIME
(days)
netetence
8,000** 392 Parrish, 1974
400*** 16 Epifanlo, 1973
4,367 238 Lane & Livingston,
1970
2,300** 35 Parrish, et al. 1973
Maximum Permissible Tissue Concentration
Action Level or Effect
Fish and shellfish -
smoked, frozen or canned
Fish and shellfish -
raw edible portion
Animal feed
Concentration
-- • (mg/kg)
0.3
0.3
0.03
Reference
FDA Admin. Guideline
7420.08
FDA Admin. Guideline
7420.09
FDA Admin. Guideline
7426.04
All data are for dieldrin
**
Edible tissue
*** Converted from dry to wet weight basis
Ccomctric mean whole fisli bioconcentration factor
4,367
Geometric mean bioconcentration factor for all species (does not include edible portion of
fish) = 2,409
Lowest residue concentration = 0.03 mg/kg
0.0000069 mg/kg or 0.0069 ug/1
-------
Table 11. Other marine data for aldrln/dieldrin*
CD
.U
U>
Organ!am
Teat
Duration ptfect
Alga,
Skeletonema costatum
Alga,
Tetraaelrais chuii
Alga.
Isochrysis galbana
Alga,
Oltsthodlacus luteua
Alga,
Cyclotella nana
Alga.
Amphldlnium carter!
Clam,
Rang! a cuneata
Eastern oyster,
Crassostrea virginica
Eastern oyster,
Crassostrea virginica
Brown shrimp.
Crangou crangon
Shore crab,
Carcinus macnus
Fiddler crab.
lie a pugilator
Crab larvae.
Leptodius floridanus
2 hrs
2 hrs
2 hrs
2 hrs
2 hrs
2 hrs
72 hrs
7 days
7 days
48 hrs
48 hrs
IS days
18 days
1,588**
859**
824**
490**
481**
98**
Bioconcentration
factor - 1.600
Bioconcentrat ion
factor <* 2,070
Bioconcentration
factor - 2,880
LC50
LC50
Dieldrin in food
affected running
Bioaccumulated al
consuming food wi
Result
Crab larvae, 6 days
Leptodivis flortdanus
213 ng/kg
LC50, approximately
- 1
peierencfc
Rice & Sikka, 1973
Rice & Sikka, 1973
.Rice & Sikka. 1973
Rice & Sikka. 1973
Rice & Sikka, 1973
Rice & Sikka. 1973
Petrocelli. et al. 1973
Mason & Rowe. 1976
Mason & Rowe, 1976
>10, <33 Portroann & Wilson. 1971
>10, <33 Portmann & Wilson. 1971
0.1 tig/g Klein & Liueer. 1974
>r
217 ng/g Epifanio, 1973
Epifanio, 1971
-------
Oranic*
Table 11. (Continuet»)
Test i**Ult
Blue cfife. 10 days Bioaccuaulated 4 to « Pe trope 111, ft al. lf||
Calltn«(CUa aapidu» 1 tiuwi (h* daily
—-•--"—•---- ----- - dose in fQod
S.llfln molly, 34 wk. LC50 ?|.| O.O Lag* % Llvin^ton. 1970
Poacilta latiplnna r
Winter flounder. -r 1.21 pg/kg in egga - Smith 6> Cole.
Raeudopleuronectea caused &&t reduction
|pftri-gi>nug"^-"--"'-t'7 in"f?rtil'i?§ti«ii"J
to contrqla
D
I
* All data are for dieldrln
** Correccion faccor> (0.1) for dry weight analys^a,
-------
Cairns, J., Jr. 1968. The effects of dieldrin on diatoms.
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3-4'
-------
Hansen, C.R., Jr., and J.A. Kawatski. 1976. Application
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B-48
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Karnak, R.E., and W.J. Collins. 1974. The susceptibility
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B-49
-------
Lowe, J.I. Results of toxicity tests with fishes and macro-
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14 14
elimination of dietary C-DDT and C-Dieldrin in rainbow
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I
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B-50
-------
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B-51
-------
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4 •
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B-52
-------
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B-53
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9
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j
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ALDRIN AND DIELDRIN
Mammalian Toxicology and Human Health Effects
Introduction
During the past decade, considerable information has
been generated concerning the toxicity and potential carcino-
genicity of the two organochlorine pesticides aldrin and
dieldrin. These two pesticides are usually considered to-
gether since aldrin is readily expoxidized to dieldrin in
the environment. Both are acutely toxic to most forms of
life including arthropods, mollusks, invertebrates, amiphi-
bians, reptiles, fish, birds, and mammals. Dieldrin is
extremely persistent in the environment. By means of bio-
accummulation 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 grasshoppers, protection of vegetables and fruits, and
control of disese vectors including locusts and termites
C-l
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(Int. Agency Res. Cancer, 1974a,b). In 1974, the U.S.EPA
restricted the use of aldrin/dieldrin to termite control
by direct soil injection and non-food seed and plant treatment,
Early work by Treon and Cleveland in 1955 suggested
that aldrin and dieldrin may have tumor-inducing potential,
especially in the liver. Since that time, several conflicting
reports of the hepatocarc.inogenicity in mice, rats, and
dogs have appeared in literature. Studies have been carried
out mainly by the U.S. Food and Drug Administration, the
National Cancer Institute, and by the manufacturer, Shell
i
Chemical Company. There has been much debate over the type
and significance of hepatic damage caused by aldrin and
dieldrin. IIY order to ascertain the human risks associated
with aldrin and dieldrin, evaluations of the toxic effects
of. these pesticides have been carried out on workers in
i
the Shell Chemical Company. The evaluations include epide-
miological 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.
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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 beyoimd
the United States, it must be assumed that imported food
stuffs, such as meat products, contain residues of these
pesticides.
Use of aldrin and dieldrin peaked at 19.3 million Ibs.
in 1966, and 3.6 million in 1956, respectively (39 FR 37251).
The subsequent decline in dieldrin use was due, in part,
to increased resistance of boll weevils to chlorinated insect-
icides (Table 1). The use of dieldrin was preferred to
aldrin because it required less application due to its persis-
tance.
C-3
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TABLE 1
Year
Aldr in
(1,000 Ibs)
Dieldrin
(1,000 Ibs)
1950
1951
1952
1953
1954
1955
1956
1957
1958
1.959
1960
1961
1962
1963
1964
1965
1-966
1967
1968
1969
1.970
L971
1972
1973 (to July 1) .
1973 estimated (to Dec. 31)
1973
1974 (to July 1)
1,456
3,288
814
1,234
2,993
4,372
6,495
2,431
4,971
5,566
8,109
9,26
10,886
12,152
12,693
14,278
19,327
18,092
13,690
9,902
8,909
11,615
11,868
8,721
(10,000)
9,900
9,700
0
185
750
1,135
1,777
2,585
8,635
2,673
3,074
3,008
2,650
2,764
2,990
2,685
2,052
1,814
1,908
1,478
1,332
1,206
749
705
740
432
(576)
'Domestic sales of aldrin and dieldrin from 1950 through July 1,
1974 (3.9 FR, 19T4);...
C-4
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Ingestion from Water
Aldrin and dieldrin have been applied to vast areas
of agricultural land and aquatic areas in the United States
and in most parts of the world. These pesticides have there-
fore found their way into most fresh and marine waters.
Unlike DDT, aldrin and deildrin are somewhat more soluble
in water (27 and 186 mg/1, respectively) (Park and Bruce,
1968). Gunther, et al. (1968) reported 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. Leichtenberg,
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 concentra-
tions ranging from 5 to 395 ng/1 in surface water and from
1 to 7 jjg/1 in drinking water (Epstein, 1976) .
In 1975 a survey in the United States of aldrin, diel-
drin, DDT, and DDT metabolite levels in raw and drinking
water was carried out (U.S.EPA, 1976). Dieldrin was found
in 117 of 715 samples analyzed (Table 2). The six samples
C-5
-------
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.
TABLE 2
Dieldrin Concentrations in Raw and Drinking Water
(U.S. EPA, 1976)
No. of Samples 598 94 13 4 6
ng/1 4 4-10 11-20 21-29 56-110
.-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/l in Seattle, Washington, and Cincinnati, Ohio;
2 ,ng/l 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
firom shallow waters, dieldrin has by far the longest half-
lii'fe of these compounds in water 1 meter in depth. They
calculated that the half-life for aldrin and dieldrin would
be 10.1 days and 723 days, respectively, compared to 3.5
days -for DDT and 289 days for lindane. This long half-life
in water combined with the potential for bioconcentration
by aquatic organisms such as micororganisms, phytoplankton,
mollusks, and fish further enhances the hazard of these
two pesticides (Wurster, 1971).
C-6
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Ingestion from Food
Although aldrin is readily converted to dieldrin, diel-
drin itself is stable and persistent in the environment.
Because it is lipophilic, dieldrin accumulates in the food
chain (Wurster,. 1971). The persistence of aldrin and diel-
drin 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 concentration of
dieldrin in the diet in the United States has been calcu-
lated 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 16-to 19-year-old male (Natl.
Acad. Sci., 1975).
A bioconcentration factor (BCF) relates the concentration
of a chemical in water to the concentration in aquatic organisms,
but BCF's are not available for the edible portions of all
four major groups of aquatic organisms consumed in the United
States. Since data indicate that the BCF for lipid-soluble
C-7
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TABLE 3
Daily Dietary intake (ing)
1965 1966 1967 1968 1969 1970
Aldrin 0.001 0.002 0.001 trace trace trace
V Dieldrin 0.005 0.007 0.001 0.004 0.005 0.005
00
-------
compounds is proportional to percent lipids* BCF's can be
adjusted to edible portions using data on percent lipids
and the amounts of various species consumed by Americans.
A recent survey on fish and shellfish consumption in the
United States (Cordle, et al. 1978) found that the per capita
consumption is 18.7 g/day. From the data on the 19 major
species identified in the survey and data on the fat content
of the edible portion of these species (Sidwell, et al.
1974), the relative consumption of the four major groups
and the weighted average percent lipids for each group can
be calculated:
Consumption ' Weighted Average
Group (Percent) Percent Lipids
Freshwater fishes 12 4.8
Saltwater fishes 61 2.3
Saltwater molluscs 9 1.2
Saltwater decapods 18 1.2
Using the percentages for consumption and lipids for each
of these groups, the weighted average percent lipids is
2.3 for consumed fish and shellfish.
Measured steady-state bioconcentration factors were
obtained for dieldrin using five species:
C-9
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Organisms
Eastern oyster,
Crassostrea virginica
Percent Adjusted
BCF Lipids BCF Reference
8,000
1.5 12,266 Parrish, 1974
Spot,
Leiostomus xanthurus
Lake trout (yearling) ,
Salvelinus namaycush
Channel catfish,
Ictalurus punctatus
Channel catfish,
Ictalurus punctatus
2,300
68,286
2,385
2,993
3.1
14.9
3.2
3.2
1,706
10,540
1,714
2,151
Parrish
1973
Reinert
et al.
Shannon
Shannon
, et al
1974
, 1977b
, 1977a
Each of these measured BCF's was adjusted from the percent
lipids of the test species to the 2.3 percent lipids that is
the weighted average for consumed fish and shellfish. The
geometric mean was obtained for each species, and then for
all species. Thus, the mean bioconcentration factor for
dieldrin and the edible portion of all aquatic-organisms
consumed by Americans is calculated to be 4,500.
No useful measured bioconcentration factor can be obtained
for aldrin because it is rapidly converted to dieldrin by
aquatic organisms. In addition, because aldrin is converted
to dielrin in soil, aquatic organisms are rarely exposed
to aldrin.
However, the equation "Log BCF =0.76 Log P - 0.23"
can be used (Veith, et al. Manuscript) to estimate the BCF
for aquatic organisms that contain about eight percent lipids
from the octanol-water partition coefficient (P). Based
on an octanol-water partition coefficient of 1,000, a steady-
state bioconcentration factor for aldrin would be estimated
to be 110. An adjustment factor of 2.3/8.0 = 0.2875 can
C-10
-------
be used to adjust the estimated BCF from the 8.0 percent
lipids on which the equation is based to the 2.3 percent
lipids that is the weighted average for consumed fish and
shellfish. Thus, the weighted average bioconcentration
factor for aldrin and the edible portion of all aquatic
organisms consumed by Americans would be calculated to be
110 x 0.2875 = 32.
Inhalation
Aldrin and dieldrin enter the air through various mech-
anisms 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,
» ' 3
Iowa had detectable levels of aldrin (8.0 ng/m ), and 50
samples taken in Orlando, Florida had detectable amounts
of dieldrin, the largest being 29.7 hg/m . Various other
studies of the air carried out during the 1960's were sum-
marized by Edwards (1973).
In a study conducted by the U.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/m3. From these levels, the average daily intake
of dieldrin by respiration was calculated to be 0.035 to
0.098 pg.
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
C-ll
-------
in the air. Exposure due to inhalation of aldrin and diel-
drin from the application of these pesticides was, of course,
much greater before the restriction of their use. Pesticide
applicators and indiviudals living near agricultural areas
were exposed to aldrin/dieldrin through inhalation.
In a recent report, Domanski, et al. (1977) reported
no increase in dieldrin concentration in adipose tissue
of cigarette smokers as compared to non-smokers although
tobacco has high residues 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
pestic.ides. Wolfe, et al. (1972) reported that exposure
to workers, both manufacturers and applicators, was mainly
through dermal absorption rather than from inhalation.
Due to the ban on manufaturing of the pesticides in the
United States, the possibilities of dermal exposure have
been greatly reduce.
PHARMACOKINETICS
Absorption
Heath and Vandekar (1964) , using Cl-dieldrin (4 percent
in arachis oil) showed that absorption by the upper part
of the gastroinestinal tract begins almost immediately after
oral administration in rats and that the absorption varies
with the solvent used. Barnes and Heath (1964) demonstrated
that the LD50 varies with the dieldrin-to-solvent ratio.
Heath and Vandekar (1964) also demonstrated that absorption
is by the poartal vein and not the thoracic lymph duct.
' C-12
-------
Initially, dieldrin is widespread but within a few hours
it is redistributed in favor of the fat. They also stated
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
h
in water and a high solubility in fat. At 1 and 2 hours
after treatment, Heath and Vandekar (1964) detected the
highest concentration of Cl-dieldrin in fat tissue. They
also reported high concentrations in the liver and kidney -
with moderate concentrations in the brain at these times.
Deichmann, et al.(1968) studied the retention of diel-
drin 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 concentration of dieldrin in the blood
and liver increased for nine days and then leveled off until
the end of the six-month period. The concentration of diel-
drin 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 diel-
drin in rats and dogs over a two-year period. Dieldrin
(99 percent purity) was incorporated into the diet of CFE
C-13
-------
male and female rats at 0.1, 1.0, and 10 rag/kg and was fed
to dogs in gelatin capsules at concentrations equivalent
to 0.1 and 1.0 mg/kg of their daily dietary intake. The
authors measured the dieldrin residues in whole blood, fat,
liver, and brain and found signifrcantly increased concentra-
tions in all tissues compared to those in the controls (Table 4)
TABLE 4
fe-.H Mean Geometric Dieldrin Concentration (mg/kg) in Rats
104 weeks
Dietary
Level (mg/kg) Blood
Males
Females
0
0.1
1.0
10.0
0
0.1
1.0
10.0
0.0009
0.0021
0.0312
0.1472
0.0015 '
0.0065
0.0861
0.3954
Fat
0.0598
0.02594
1.493
19.72
0.3112
0.8974
13.90
57.81
Liver
0.0059
0.0159
0.01552
1.476
0.0112
0.0348
0.4295
2.965
Brain
0.0020
0.0069
0.1040
0.4319
0.0077
0.0224
0.2891
1.130
(Walker, et al. 1969)
The concentrations in the tissues increased with an
increase in the dietary concentrations, and the concentra-
tions in the female rats were considerably higher than those
in the males. The dieldrin 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 of 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
C-14
-------
18 weeks. In the group receiving 1.0 mg/kg/diet the dieldrin
concentration in the blood increased 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 significant sex
differences. As in other studies, the concentration 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 carried out by Robinson, et al. (1969). In this study
Carworth rats were fed dieldrin (99+ percent purity) at
10 mg/kg in their diet for 8 weeks. At the end of this
time, they were returned to a dieldrin-free diet and killed
randomly in pairs up to 12 weeks after withdrawal of the
dieldrin diet. The fatty tissue clearly had the highest
concentration of dieldrin followed by the liver, brain,
and blood. Concentrations of dieldrin in fat returned to
control levels after 12 weeks and the decline in dieldrin
concentrations was approximately exponential in nature.
Matthews, et al. (1971) investigated the distribution
of dieldrin 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
14
then given 3 g of food containing 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
C-15
-------
of the males contained more radioactivity than those of
the females. Levels 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 predominantly 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 concentration in the tissues was determined.
The concentrations in the brain at 4 and 16 hours were 1.5
and 1.0 jug/g, respectively. Hayes assigned a value of one
to the concentrations in the brain and calculated the ratio
of the concentrations in other tissues to the concentrations
in the brain at 4 and 16 hours (Table 5).
TABLE 5
Hr. Brain
Muscle
Liver
Kidney
Plasma
Fat
4
16
1.00+0
1.00+0
0.62+0.05
0.55+0.06
2.30+0.11
3.17+0.25
1.55+0.22
2.02+0.56
0.20+0.02
1.35+1.11
7.20+1.18
17.96+3.23
C-16
-------
The concentrations in the tissues remained relatively con-
stant for 24 hours and began to decline at 48 hours. No
further samples were taken until 240 hours when all the
dieldrin concentrations were below 0.2 pg/g except the con-
centration in the fat which was 5 pg/g.
In a study done in 1963 on 30 individuals from three
different states, the concentrations of chlorinated hydro-
carbon 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 the aldrin exposure was
0.36 jug/g dieldrin (see discussion 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) measured the plasma, fat, and urine concent-
rations 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 concentrations of diel-
drin in the fat, urine, and plasma of the pesticide workers
were 5.67+1.11, 0.242+0.0063, and 0.0185+0.0019 mg/g respect-
ively. These were significantly different from those re-
ported for the general population. The authors reported
a high correlation between total hours or intensity of .exposure
and concentration of dieldrin. However, no correlation
could be found between dieldrin concentrations and amount
of sick leave.
I C-17 .
-------
Another study (Hunter, et al. 1969) involving adult
males ingesting 10, 50, or 211 jig dieldrin per day for 18
or 24 months again found a relationship between the dose
and the length of exposure and concentration of dieldrin
in the fat and blood. In general, 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 ;ug/day for the final 6 months, demonstrated a
rise in concentrations similar to the rise demonstrated
by those who were given 211 jug/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 above-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-lipoprotein 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
C-18
-------
concentrations were higher than the aldrin 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 contained 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 percent purity incubated with various fractions of
human blood £n vitro Skalsky and Guthrie were able to demon-
strate that dieldrin and DDT bind to albumin and beta-lipopro-
tein.
Metabolism
Aldrin and its epoxidation product, dieldrin, are both
cyclopentadiene insecticides. Since epoxidation of aldrin
to dieldrin was first reported by Radomski and Davidow in
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 iin vitro conversion
of aldrin to its epoxide, dieldrin, using microsomes* from
*In this document microsomes refers to the cell-free homo-
genized liver (including soluble enzymes and microsomes)
and not to purified microsomes.
C-19
-------
male and female rats. The reaction was NADPH-dependent
and the enzymes were heat-labile., Winteringham and Barnes
also showed that males converted alrdin to dieldrin at a
higher rate. No other metabolic products were detected,
although the authors noted that polar products could have
been overlooked by the methods used. Nakatsugawa, et al.
(1965) confirmed the work of Wong and Terriere using micro-
somes 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 aldrin 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 excretio'n and that
little dieldrin is excreted unchanged. They were able to
detect other polar metabolites in both urine and feces.
14
Ludwig, et al. (1964) fed C-aldrin to male rats at
4.3 jug/day for three months. The compounds excreted into
the urine consisted of aldrin, dieldrin, and unidentified
hydrophilic metabolic 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
g-20
-------
four products appeared to be identical by paper and thin-
layer chromatography.
Korte and Arent (1965) isolated six urinary metabolites
14
from rabbits treated orally with C-dieldrin for 21 weeks.
The major metabolite (86 percent) was one of the two enantio-
morphic isomers of 6,7 trans-dihydroxy-dihydro-alrdin.
Richardson, et al. (1968) were able to identify two
metabolites in urine and feces 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 determined that the urinary metabolite
had a keto group on the number 12 carbon and the epoxide
was unchanged. The fecal metabolite was a mono-hydroxyder-
ivative 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 metabo-
lites 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 identical to the
ketone compound identified by Richardson, et al. in the
urine. The minor urinary and fecal metabolites were ident-
ical and similar to the 6,7 trans-dihydroxy-dihydro-aldrin
described by Korte and Arent (1965).
C-21
-------
Matthews and Matsumura (1969) also conducted in vitro
14
experiments using C-dieldrin incubated with rat liver
microsomes and various co-factors. Thin-layer chromatography
of the water-soluble components producted six metabolites
in addition to the unchanged dieldrin. Analysis of the
water-soluble metabolites revealed a glucoronide conjugate
which accounted for approximately 45 percent of the radio-
activity. Comparsion of the Revalues for the in vivo and
iji vitro studies showed that the minor urinary/fecal metabo-
lite (i.e., the 6, 7 trans-dihydroxy-dihydro-aldrin) was
produced in vitro and that the metabolite freed from the
glucoronic acid was also present in the in vitro system
in the unconjugated form.
The products identified by Richardson, et al. (1968)
0
and Matthews and Matsumura (1969) represent an oxidized
form of dieldrin in the urine and an oxidated, dechlorinated
metabolite in the feces which had lost the intact dieldrin
ring system.
Hedde, et al. (1970) were able to isolate six metabolic
14
products in the urine of sheep dosed with C-dieldrin.
Three castrated sheep were given unlabelled dieldrin orally
14
at 2 mg/kg for five days before dosing with C-dieldrin.
Four other sheep were fed a single oral dose of labelled
dieldrin at 20 mg/kg. Urine and feces were collected up
to six days after treatment with the labelled dieldrin.
Although other determinations were made, only the urine
was analyzed quantitatively. After hexane extraction of
pH 1 followed by other clean-up procedures, the four hexane-
C-22
-------
soluble metabolites were separated on Sephadex LH-20 gel.
The LH-20 was again used to separate the two water soluble
metabolites after they were purified by several procedures,
including paper chromatography. The authors postulated
that these water-soluble metabolites were a glucoronic acid
conjugate of the transdiol and an unidentified conjugate
of glucoronic acid and, possibly, glycine.
Feil, et al (1970) were able to identify two to 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-,momo-hydroxy-derivative. Further work on the
metabolism of dieldrin (Matthews, et al. (1971) is discussed
in the Distribution section of this report where details
of treatment are given. Matthews, et al. documented the
•
production of several metabolites of dieldrin including
the 6,7-trans-dihydroxy-dihydro-aldrin and a second unident-
ified polar metabolite excreted in the feces. The mono-
hydroxy-lated compound 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. Most 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.
C-23
-------
The metabolism and excretion of dieldrin appears to
be more rapaid in male than in female rats. Investigators
attribute this to the males' ability to produce the more
polar metabolites, especially the keto-product which is
excreted into the urine.
A recent paper has appeared on the comparative meta-
bolism of dieldrin in rodents. Baldwin, et al. (1972) treated
14
a male CFE rat with 3 mg/kg of C-labelled dieldrin and
two male CF1 mice with 10 mg/kg. The urine and feces were
collected for the following seven or eight days. The authors
reported that the CFE rat excreted the pentachloroketone
derivative in the urine but that the CFl mice did not.
Conversely, the mice produced an unidentified urinary metabolite
which'the rat did not. The 6,7-trans-dihydorxy-dihydro-
aldrin 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 diel-
drin and endrin in rodents has been compiled by Bedford
and Hutson (1976) . They summarized the four known metabolic
products of dieldrin as they 6,7-trans-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 of short-
term studies versus chronic, low-dose exposure, it must
be mentioned that organochlorine compounds, including diel-
C-24
-------
drin, have been shown to induce the mixed function oxidases
(MFO) found in the liver (Kohli, et al. 1977). It is there-
fore possibel, in the long-term animal studies, that investi-
gators 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. (1971) in
a limited study, were able to show some inducability in
the CFE male rat but not in the CFl 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 subject to chronic, low-
dose exposure to many MFO inducers (including various organo-
chlorine pesticides), this exposure may affect studies of
dieldrin metabolism.
Excretion
As mentioned in the Distribution and Metabolism sections
of this report, aldrin and/or dieldrin are excreted mainly
in the feces and to some extent in the urine in the form
of several metabolites 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 concentrations will increase
and vice versa.
The early work of Ludwig, et al. (1964) demonstrated
that male Wistar rats administered daily with low doses
of C-labelled aldrin (4.3/jg for 12 weeks) excreted appro:
mately nine times as much of the radioactivity in the feces
C-25
-------
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 percenc 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-deildrin 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 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 metabo-
lites for excretion. It is possible that the increased
amount of radioactivity noted by Korte and Arent in the
feces after treatment with aldrin could be due to the less
polar aldrin or deldrin as compared to the more polar metabo-
lites excreted in the urine or to a basic in metabolism
of deildrin in the rabbit.
C-26
-------
The work of Robinson, et al. (1969) on the metabolism
of dieldrin 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 illustrates the
loss of dieldrin from these tissues. During the period
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. The fat and
brain contained greater than 99 percent dieldrin and the
excretion times correspond to those for the rat observed
by Korte and Arent, 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 half-life of dieldrin in the liver
and blood were 1.3 days 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 C-dieldrin metabolism in sheep (Hedde,
et al. 1970) mentioned in the Metabolism section of this
report, the excretion of dieldrin or its metabolites was
C-27
-------
01
0-01
o
o
Ul
"S 0001
e
0-0001
t
V.
Blood
cXI04«542«xp(-0-535l)
.+ 298 exp (-005291)
-
-X
2O 40
60
_J
80
°"
a
o
X
^
§
§
a
0001
\
e«07lexp(-O54t)
-f 0-233 exp(-O068t)
20 tO 60
G
200
10-0
-o
o-,
o
-------
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
14
that was exhaled as CO^ was only 0.25 percent of the total
dose. This indicates that virtually none of the dieldrin
is broken down to CO^. With the sheep, less than 50 percent
of the total radioactivity was recovered after the five
or six days of collection.
Several investigators have shown that removal of diel-
drin from the diet results in rapid loss of dieldrin or
metabolites from the body, especially the adipose tissue.
Barren and Walton (1971) further studied the loss of diel-
drin from the body of the rat and also looked at the role
of dieldrin in the diet with respect to loss from the adi-
pose 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
14
given four daily, oral doses of C-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 normal diet.Groups
of five animals were sacrificed on the four days when they
received the labelled-dieldrin and on days 7, 9, 11, 16,
and 23 after the conclusion of the eight-week feeding.
The concentration of dieldrin found in the adipose tissue
from the rats receiving the dieldrin diet was approximately
C-29
-------
50 ;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 >ig/g at day 23.
The authors reported that the half-life of dieldrin in the
adipose tissue was about 4.5 days/ which is somewhat lower
than the 10.3 days calculated by Robinson, et al. (1969)
with rats fed only 10 mg/kg dieldrin.
Cole, et al. (1970) measured the appearance of C-dieldrin
14
and C-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 radioactivity occured in the feces. Approx-
imately 80 percent of the total dose of labelled 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 patterns of excretion similar to
those observed in the first experiment.
In a comparison of the excretion of dieldrin in the
CF1 mouse and CFE rat, Baldwin, et al. (1972) found that
after seven or eight days the amount of labelled dieldrin
excreted was similar to both species. Also, the feces contained
approximately two times as much radioactivity as the urine,
and 50 to 70 percent of the total activity was excreted
during the collection period. As mentioned in the Meta-
bolism section of this report, the proportion of metabo-
lites varied between the mouse and the rat.
C-30
-------
Although there has been extensive work done on the meta-
bolism and excretion of dieldrin in animals, there is under-
standably less known about the fate of dieldrin in humans.
Early work by Cueto and Hayes (1962) demonstrated that diel-
drin 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 different degrees of occupational exposure. The concen-
trations of dieldrin 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 concentrations 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 individuals
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, aldrin was present at less than 0.2
mg/1 and dieldrin was present at 1.3 to 66.0 mg/1. This
is compared to the mean for dieldrin in the general popu-
lation of 0.8 — 0.2 mg/1 determined in the same laboratory
by Cueto and Biros (1967).
A study by Hunter, et al. (1969) concluded that dieldrin
had a relatively long half-life in humans. This compares
C-31
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with a half-life of less than ten days reported in animal
(t
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 additi-
onal 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 sign-
ificantly. (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/dieldrin workers who were transferred to other
areas. Prior to transfer, these workers had had high ex-
posures to the pesticides and concentrations of aldrin/diel-
drin in their blood had reached equilibrium. 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 (Robin-
son, et al. 1969; Walker, et al. 1969) that there is a
direct relationship between the concentration of dieldrin
in the blood and that in adipose and other tissues. It
seems likely that the half-life in the blood may reflect
the overall half-life in other tissues.
C-32
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EFFECTS •-.-..•;•'.:•/•-• ^; ^.. .
Acute, Subacute, and Chronic Toxicity
The acute toxicity of aldrin and dieldrin has been ex-
tensively summarized by Hodge, etal. (1967) and Jager (1970)
In many cases, aldrin and dieldrin are considered similar
due to the rapid conversion of aldrin to dieldrin (see Meta-
bolism section) . Dieldrin, in turn, is metabolized to a
variety of more polar products. In"some cases, the toxi-
city 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 absorb-
ed 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 (including 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 LD50 (Barnes and Heath, 1964) and that some
variation is caused by the solvent employed (Heath and Vand-
ekar, 1964). There is a pronounced variation in toxicity
related to route of administration. Toxicity is highest
by the intravenous route, followed by oral, then dermal.
C-33
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This is most likely due to the high blood and central ner-
vous system concentrations produced from intravenous injec-
tion. 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 animals 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
concentrations 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 jug/ml blood and convulsions at 0.74 to 0.84 jug/ml.
The symptoms of intoxication in man are similar to
those found in mice, rats, and dogs. Jager (1970) described
the symptoms resulting from oral or dermal exposure that
occur from 20 minutes to 24 hours as:
"...headache, dizziness, nausea, general malaise, vomiting,
followed later by muscle.twitching, myoclonic jerks
and even convulsions. Death may result from anoxemia."
Changes in the electroencephalogram (EEC) usually result
after insecticide intoxication and generally return to normal
after recovery (Hootsman, 1962). The transitory change
C-34
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in the EEC has been challanged by several investigators
(see Burchfiel, et al. (1976) for recent summary). Work
carried out in Rhesus monkeys (Burchfiel, et al. 1976) using
technical grade dieldrin (4 mg/kg, i.v. one time or 1 mg/kg
i.m. administeredonce a week for 10 weeks) demonstrated
0
that dieldrin can alter the EEC 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
Hodge, et al. (1967) to be 5 g or 70 mg/kg respectively.
However, Hodge, et al. only speculated on possible human
toxic effects from a 1-year feeding 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 must be somewhat higher (Hayes, 1971).
The sub'acute 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 induction 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 information gained from monitoring workers at the Shell
Chemical Company, Jager (1970) reported that 33.2 /ig/kg/day
can be tolerated by workers for up . to 15 years. Above this
level some individuals may show signs of intoxication, al-
C-35
-------
though others can tolerate two times this level. In another
study involving 12 volunteers who ingested dieldrin for
up to two years, 3.1 ;ug/kg/day was tolerated and produced
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 organ-
ochlorine 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 meta-
bolic products produced after acute dosing. Several reports
have appeared on the combined effect of aldrin or dieldrin
on the storage of DDT in tissues (Street, 1964; Street and
Blau, 1966; and 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. The
authors suggest that this increase in tissue dieldrin concen-
trations 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.
036
-------
Clark and Krieger (1976) studied the metabolism and
tissue accumulation of C-labelled aldrin (99.3 percent
purity) in combination with an inhibitor of oxidative bio-
transformation (i.e., SKF 525-A). They reported that pre-
treatment 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 dieldrin.
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
14
In 1967, Hathaway, et al. established that C-dieldrin
could cross the placenta in rabbits. Eliason and Posner
14
(1971a,b) demonstrated that C-dleldrin 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 survival of the young,
while mice fed lower doses showed fewer or no effects.
In a study by Ottolenghi, et al. (1974) pregnant golden
hamsters and pregnant CD-I mice were given single oral doses
of purified aldrin, dieldrin, or endrin at one-half the
LD50 (hamsters 50, 30, 5 mg/kg, and mice 25, 15, 2.5 mg/kg,
respectively). The hamsters were treated orally on day
seven, eight, or nine of gestation and the mice on day nine.
All three pesticides caused a significant increase in fetal
C-37
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death in hamsters treated on days seven and eight. Only
dieldrin gave significant results on day nine. Hamsters
treated on day eight also had the highest number of anomalies
(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 hamsters. 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 diel-
drin (87 percent purity) and the photo-product, photodiel-
driri (95 percent, purity) in CD-I mice and CD rats orally
at doses lower than those used by Ottolehghi, 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 period (i.e.,
days 7 to 16 of gestation). The compounds were dissolved
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 signi-
ficant increases in liver-to-body.weight ratios, reduced
the weight gain, and produced some fetal toxicity. Photodiel-
C-38
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drin 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.
Dix, et al. (1977) examined the use of two solvents
(corn oil and dimethylsulfoxide (DMSO)) with various doses
of dieldrin in CF1 mice. The corn oil groups received 1.5
or 4.0 mg/kg/day of 99 percent pure dieldrin orally with
suitable controls of corn oil or no treatment, The DMSO
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 used and the pregnancy rate was very
low. With the few animals that survived to term, the only
significant effect was delayed ossification in the mice
administered the 4 mg dose. The DMSO experiments were con-
ducted with older animals (ten weeks) of proven fertility.
These animals demonstrated a significant increase in inci-
dences of delayed ossification and extra ribs. However,
the DMSO controls also had a high incidence of these two
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 resoprtions, or fetal death with either solvent.
039
-------
Mutagenicity
Relatively little work has been done on the mutageni-
city of aldrin or dieldrin. Of the limited data available,
most are concerned with the mutagenicity of dieldrin. This
may be sufficient, since aldrin is readily converted to
dieldrin in both in vivo and j.£ vitro systems. Fahrig (1973)
summarized the microbial studies carried out up to 1973
on aldrin, dieldrin, and other organochlorine pesticides
including DDT and the metabolites of DDT. Aldrin and diel-
drin gave negative results with gene conversion in Sacchar-
omyces cerevisae, back-mutation in Serratia marcescens,
forward mutation (Gal Rs) in Eschericia coli and forward
mutation to streptomycin resistance in E^ coli. It is impor-
tant to note that DDT and several of its metabolites also
gave negative results in these microbial tests and that
no mention of any type of activation system (i.e., 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 strains of
Salmonella typhimurium 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, micro-
nucleus test, metaphase analysis, dominant lethal test,
and heritable translocation test. The doses used were 0.08,
0.8, and 8.0 mg/kg in corn oil with corn oil used as the
control and triethylene melamine (0.5 mg/kg five times)
serving as the positive mutagenic control. The pesticide
was given orally on a subacute basis.
i •
C-40
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Three reports on the mutagenicity of aldrin or dieldrin
have recently been published. The first examined the mutagen-
icity 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 activating
system (Shirasu, et al. 1977). The third study dealt pri-
marily with strains of S^ typhimurium (TA1535, TAlOO, and
TA98) plus a mouse liver activating system (Majumdar, et
al. 1977) .
In the Marshall, et al. (1976) study, dieldrin was
tested at only one concentration, 1000 ug 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
activation using £_._ coli B/r WP2 try-hcr+ and WP try-hcr~
and S^ typhimurium strains TA1535, TA1537, TA98, and TAlOO.
Dieldrin was assayed without metabolic activation using
the E^ coli WP2 hcr + , WP2 her" and S^ typhimurium TA1535,
»
TA1536, TA1537, and TA1538. According to the authors, both
aldrin and dieldrin were considered non-mutagenic in these
tests.
Majumdar, et al. (1977), on the other hand, have reported
that dieldrin was somewhat mutagenic for S_._ typhimurium
strains TA1535, TAlOO, and TA98 without metabolic activation
C-41
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and that it was strongly mutagenic for all three strains
when liver enzymes from Aroclor-1254*-induced mice were
added to the mixtures.
In summarizing the limited microbial mutagenicity studies
on aldrin and dieldrin, it must be mentioned that the only
reference to any mutagenicity in the Majumdar studies con-
tains several notable inconsistencies. The inconsistencies
are: (1) the cultures 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 consis-
tent with those recommended by Ames, et al. (1975).
It 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 re-
ported no mutagenic findings with dieldrin. It should be
kept in mind that mice apparently metabolize dieldrin differ-
ently than do rats (see the Metabolism section of this report).
It 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 organ-
isms 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
*Aroclor-1254 is a mixture of PCB's, which induce the MFO
in liver (Ames, et al. 1975).
C-42
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in the cytogenic effects of dieldrin in mouse bone marrow
nuclei and Bunch and Low (1973) reported chromosomal aberra-
tions in semi-domestic mallard ducks.
.gecently, Majumdar, et al. (1976) studied (1) the effect
of dieldrin on chromosomes in mouse bone marrow iin vivo
and in'cultured human WT-38 lung cells, and (2) the cyto-
pathic effect of dieldrin on the cultured human WI-38 cells.
They reported a decrease in the mitotic index in both the
iS v^-vo mouse bone marrow and in 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 jug/ml in human cell cultures). The authors also re-
ported a dose- and time-dependent cyt:otoxic effect on the
WI-38 human 'lung cells.
In addition, Ahmed, et al. (1977) measured unscheduled
• DNA synthesis (UDS) in SV-40 transformed VA-r4 human fibro-
blasts 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 £he doses used.
Carcinogenicity —
During the 1960's and the early part of the 1970's,
numerous studies on the carcinogenicity of aldrin and diel-
drin appeared in literature. These reports include studies
on mice, rats, dogs, and monkeys. Of these species, mice
appear to be the most susceptible to aldrin/dieldrin. Various
C-43
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strains of both sexes have been examined at different dose
levels. The effects range from benign liver tumors to hepat-
ocarcinogenicity with transplantation confirmation to pul-
monary 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 carried 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 Administ-
ration (FDA) (Davis and Fitzhugh, 1962) . Using C-jHeB/Fe
(C-aH) 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 underestimated due to the high mortality which
left fewer animals for evaluation.
In an FDA followup study, Davis (1965) examined 100
males and females of the C-jH 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 histological material
of both of these studies was carried out by Rueber in 1973
(Epstein, 1975a,b, 1976) . He concluded that the hepatomas
C-44
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were malignant and that both aldrin and dieldrih were hepato-
carcinogenic for male and female, CoH mice.
In a 1964 abstract/ Song and Harville reported some
indication of hepatocarcinogenicity in CjH ancj 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 quest-
ions as to the type of lesions.
Walker, et al. (1972) conducted a multi-part study
of dieldrin in CF1 mice of both sexes. In this study, the
dieldrin used was 99+ percent pure and 4-aminq-2,3-dimethy-
lazobenzene (ADAB) was used as the positive control. In
ttte first part of the study, diets were prepared containing
0, 0.1, 1.0, and 10 mg/kg dieldrin although 0.01 mg/kg diel-
drin 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 contained equal numbers of males and females. The ADAB
group, which contained 50 mice equally divided as to sex,
received 600 mg/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 pqsitive controls
were maintained separately from the other groups. After
nine months, the mice receiving 10 mg/kg in the diet dieldrin
demonstrated palpable intrar-abdominal masses, and by the
fifteenth month, half the males and females in the group
C-45
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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 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 classified 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 formations
of liver cells with wide and irregular vascular channels
within the growth." This classification appears somewhat
arbitrary. 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-sterilized 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 observa-
G-46
-------
tions were similar to those in the first experiment. In
this part of the study, the mice that received 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 masses were detected at 36 weeks, and all the mice
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 mice, 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 toxieity 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 simi-
lar conditions. Groups of 60 mice received gamma-irradiated
i
diets containing 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 CFl mice. Initially, the mice were fed diets containing
200 mg/kg DDT or 10/200 mg/kg dieldrin/DDT. 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 control group and 32
C-47
-------
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 dieldrin/DDT group.
However, because only one combination was tested, it is
difficult to determine whether the effect was synergistic
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 DDT 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 females. In the mice receiving
dieldrin for 64 weeks, liver enlargements were detected
after 60 weeks in six males and two females. These enlarge-
ments 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 import-
ant to note that type b tumors were detected after only
4 or 8 weeks of treatment and that the liver enlargements
did not appear after the feeding was terminated.
C-48
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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 concentration (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
reported that there were no signs of intoxication in the
dieldrin groups; however, mortality increased after 22 months
of exposure. Also liver enlargements were detected in both
sexes by the 50th week. In this study, the cumulative tumor
incidence and the number of dead mice were given at 17,
21, 25, and 26 months. Dieldrin at 10 mg/kg produced a
high incidence of liver tumors. All the males and one-half
the females that had died by 17 months had liver tumors.
By the end of the study, 100 percent of the males and 87
percent of the females had liver tumors.
In a recent evaluation of both aldrin and dieldrin
by the National Cancer Institute, aldrin and dieldrin were
found to produce hepatic carcinomas in male mice. Female
mice responded to low doses of dieldrin, but showed no effects
from aldrin. No carcinomas 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 BgC-jF-^
mice were fed either aldrin (technical grade) or dieldrin
(technical grade) at various doses. The females received
049
-------
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 controls had 20 animals per
group. In addition, pooled controls consisted of 92 males
and 78 females. The animals were fed the pesticide 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 signif-
icant, 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 (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 female mice at the high dose of aldrin. With
the male mice fed dieldrin, a significant increase in hepatic-
carcinomas was observed in the 5 mg/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 in-
cidences were given, although some liver lesions were detected,
C-50
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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 dieldrin were not tumorigenic in their rat studies.
A study was carried out by the U.S. Food and Drug Admin-
istration 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 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 1000 mg/kg methoxy-
chlor. There were 30 males and 30 females in each treatment
group and they were housed in pairs. No increase in mor-
tality over the controls was observed in any of the treated
groups. Aldrin alone had no significant effect on liver-
C-51
-------
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
rng/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, tremors, 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 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
increases in total tumors even though few animals were examined
histologically.
In another study with the Osborne-Mendel rat, Deichmann,
et al. (1970) examined aldrin, dieldrin, and endrin in a
lifetime exposure. Aldrin (technical, 95 percent) and dieldrin
C-52
-------
(technical*, 100 percent active ingredients) were fed in
the diet to groups of 50 males and 50 females. The concentra-
tions 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 contained 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 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 in 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 carried out
*This is somewhat contradictory since "technical" diel-
drin is actually 85 percent pure.
C-53
-------
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 weefcs, 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 controls
consisting the matched control groups combined with 58 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 aldrin-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
C-54
-------
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 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 demonstrated signs of intox-
ification 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.
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 carcinogenicity.
C-55
-------
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. At
doses 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 and dieldrin showed
no ill-effects during the 2 years of the study. In the
dogs fed aldrin at 1.0 mg/kg/day and 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 tumori-
genic 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 administered 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 concentration
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 work of Zavon in 1970, which appears
C-56 • >.
-------
to be unpublished, has been summarized by Epstein (1975b).
Epstein reports that six control monkeys (five male, one
female) and groups of five monkeys each received 0, 0.1,
0.5, 1.0, and 1.75 mg/kg dieldrin in their diet for 5.5
to 6 years. The group at 1.75 mg/kg received 5.0 mg/kg
for 4 months, then 2.5 mg/kg for approximatley 2.5 months,
and then 1.75 mg/kg for the remainder of the exposure.
Epstein further states that four of the monkeys died during
the study, two of which had received 5 mg/kg. The remaining
animals survived until they were killed. No data on his-
tology are given although it is reported that no differences
were observed between control and treated monkeys.
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 workers' health
were observed.
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 workers at risk and the short duration
of exposure and/or time after exposure.
C-57
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CRITERION FORMULATION
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 was suspended on the basis of adverse health affects
in rodents (39 FR 37251). As a result, production is restricted
for all pesticide products containing aldrin or dieldrin.
Aldrin and dieldrin can no longer be used for spraying and
dusting, or for mothproofing in which the residues are dis-
charged into waterways. All uses in structures occupied
by humans or livestock, uses upon turf, and any use involving
f
application 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 pesticides. They may also be used for treat-
ment of some non-food seeds and plant dipping during trans-
plantation.
The current exposure level for both aldrin and dieldrin
set by the Occupational Safety and Health Administration
is an air time-weighted average (TWA) of 250 ;ug/m for skin
absorption (37 FR 22139). In 1969, the U.S. Public Health
Service Advisory Committee recommended that the drinking
water standards for both aldrin and dieldrin be 17 /ag/1
(Mrak, 1969). Also, the U.N. Food and Agriculture Organization/World
Health Organization's acceptable daily intake for aldrin
and dieldrin is 0.0001 mg/kg/day (Mrak, 1969).
C-58
-------
Current Levels of Exposure
The people of the United States are exposed to aldrin
and dieldrin in air, water, and food. As mentioned earlier,
aldrin or dieldrin has been found in more than 85 percent
of the air samples tested by the U.S. EPA (Epstein, 1976).
The levels were as high as 2.8 ng/m resulting in an intake
of up to 0.098 jug/day. Dieldrin can travel great distances
in the air, especially when absorbed to particulate matter.
Thus people can potentially be exposed to pesticide treat-
ments 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
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 concentrations of dieldrin from that
source alone (Savage, 1976).
In early studies, Curley, et al. (1969) and Zarvon,
et al. (1969) reported that dieldrin and several other chlori-
nated hydrocarbon pesticides were present in the tissues
of stillborn infants. Curley, et al. also reported that
C-59
-------
dieldrin and other pesticides could be found in the blood
of newborn infants.
No work 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 be
an area of great concern.
Basis and Derivation of Criterion
The aldrin and dieldrin carcinogenicity data of Walker,
et al. (1972) and the National Cancer Institute (1976) were
analyzed using a linear dose-response model to calculate
that concentration of dieldrin in water which is estimated
to result in an excess lifetime risk of 10" in man (see
Appendix I). It should be noted that Walker, et al. study
used 99 percent pure dieldrin while the NCI study used techni-
cal grade dieldrin.
Under the .Consent Decree in NRDC vs. Train, criteria
are to state "recommended maximum permissible concentrations
(including where appropriate, zero) consistent with the
protection of aquatic organisms, human health, and recreation-
al 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 maximum 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 concentrations of aldrin and dieldrin corre-
C-60
-------
spending to several incremental lifetime cancer risk levels
have been estimated. A cancer risk level provides an estimate
of the additional incidence of cancer that may be expected
in an exposed population. A risk of 10~ for example, indi-
cates a probability of one additional case of cancer for
every 100,000 people exposed, a risk of 10 indicates one
additional case of cancer for every million people exposed,
and so forth.
In the Federal Register notice of availability of draft
ambient water quality criteria, EPA stated that it is con-
sidering setting criteria at an interim target risk level
of 10" , 10 or 10 as shown in the table below.
Exposure Assumptions Risk Levels and Corresponding Criteria^ '
0_ 1£~7 1£~6 1£~5
2 liters of drinking water
and consumption of 18.7
grams of fish and shellfish (2)
Aldrin 0 4.6 x 10"4 4.6 x 10~3 4.6 x 10~2
Dieldrin 0
Consumption of fish
and shellfish only.
Aldrin 0 4.6 x 10~4 4.6 x 10~3 4.6 x 10"2
ng/1 ng/1 ng/1
Dieldrin 0 4.5 x 10~4 4.5 x 10~3 4.5 x 10~2
ng/1 ng/1 ng/1
(1) Calculated by applying a modified "one hit" extrapolation
model described in the FR 15926, 1979. Appropriate bioassay
data used in the calculation of the model are presented
in Appendix I. Since the extrapolation model is linear
C-61
ng/1
4.4 x 10~4
ng/1
ng/1
4.4 x 10~3
ng/1
ng/1
4.4 x 10~2
ng/1
-------
to low doses, the additional lifetime risk is directly propor-
tional to the water concentration. Therefore, water concen-
trations corresponding to other risk levels can be derived
by multiplying or dividing one of the risk levels and corres-
ponding water concentrations shown in the table by factors
such as 10, 100, 1,000, and so forth.
(2) 99.9 percent of aldrin exposure results from the consump-
tion of aquatic organisms which exhibit an average bioconcen-
tration potential of 4500 fold. The remaining 0.1 percent
of aldrin exposure results from drinking water.
Ninety-eight percent of dieldrin exposure results from
the consumption of aquatic organisms which exhibit an average
bioconcentration potential of 4500 fold. The remaining
2 percent of dieldrin exposure results from drinking water.
0
Concentration levels were derived assuming a lifetime
0
exposure to various amounts of aldrin/dieldrin, (1) occurring
from the consumption 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 concentrations.
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 incremental risk from ambient
water exposure only.
C-62
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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 feeding studies. It has failed to induce a statisti-
cally 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 (S. Typhi-
murium and E. Coli) it was found to be non-mutagenic, but
it did produce unscheduled DNA synthesis in human fibroblasts
with and without activation. The induction 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
environment 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 tumors at any site in
six chronic feeding studies in three strains. It was found
to be mutagenic in S. typhimurium after metabolic activation
with mouse liver enzymes, but it was not mutagenic in two
*This summary has been prepared and approved by the Carcinogens
Assessment Group, U.S. EPA, on July 25, 1979.
-------
other studies of the same bacterial strain with a rat liver
enzyme activation mixture. The induction of hepatocellular
carcinomas in mice leads to the conclusion that dieldrin
is likely to be a human carcinogen.
Both aldrin and dieldrin have been found to be non-
mutagenic in several test systems as follows: a) gene conver-
sion in SA cerevisie; b) back mutations in S. marcescens
and c) foward mutations at two loci in £_._ coli. Several
other organochlorine pesticides which produce mouse liver
tumors are also non-mutagenic 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 human carcinogens.
The water quality criterion for aldrin is based on
the hepatocellular carcinoma incidence in male B6C3F1 mice
of the low dose group in the NCI chronic test, and on the
response in the 0.1 ppm group of female CF-1 mice in the
Walker, et al. (1972) experiment (because aldrin in converted
to and stored as dieldrin in fish). It is concluded that
the water concentration of aldrin should be less than 4.6
_2
x 10 ng/1 in order to keep the lifetime cancer risk below
— 5
10 . For dieldrin the criterion is based on the response
in the 0.1 ppm group of female CF-1 mice in the Walker,
et al. (1972) experiment. The corresponding concentration
_2
for dieldrin is 4.4 x 10 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 B6C3F1 male
mice given the low dose of aldrin in the NCI bioassay test,
and on the response in the 0.1 ppm group of female CF-1
mice in the Walker, et al. (1972) experiment. In the NCI
study, a time-weighted average dose of 4 ppm was given in
the feed for 80 weeks and the animals were observed for
an additional 10 weeks before terminal sacrifice. The inci-
dence of hepatocellular carcinoma was 3/20 and 16/49 in
the control and treated groups, respectively. The slope
of the one-hit dose-response curve for aldrin is calculated
from the following parameters:
n =16 Le = 90 weeks
Nfc =49 le = 80 weeks
n = 3 d = 4 ppm x 0.13 = 0.52 mg/kg/day
N = 20 L = 90 weeks
c
w = 0.035 kg
With these parameters the slope of the one-hit dose-
response curve for aldrin is 6.349 (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 one-hit dose-response curve
for dieldrin, which has a slope of 183.6 (mg/kg/day)~ as
calculated previously from the talker, et al. (1972) study.
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The equation describing the risk due to aldrin in water
is derived from the general relationship
P = BUD and D = 1/70 kg, thus
ri
P = Bu 1/70 kg and
rl
where
P{70 kg) = BHI
P = individual lifetime risk (set at 10~ for criterion
calculation)
I = average daily human intake of the substance
in question
Bu = estimated slope of the human one-hit dose-response curve
n
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) = BRa CQ (2.0 l/day)+ BRa Cfl Rad (0.0187 kg/day) +
BHd Ca Rad (°-0187 k9/day)
where
P = risk due to aldrin (set at 10 for criterion
a calculation)
Bu = 6.349 (mg/kg/day)~ , the aldrin dose-response slope
tia
BHd = 183.6 (mg/kg/day)~ , the dieldrin dose-response slope
C = criterion concentration for aldrin (to be calculated)
a
R = 32 I/kg, the fish bioconcentration of aldrin
from aldrin
R , = 4468 I/kg, the fish bioconcentration of dieldrin
a from aldrin
2.0 I/day = average daily intake of water for humans
0.0187 kg/day = average daily intake of fish for humans
C-tfO
-------
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. R , is estimated by
aci
assuming that in the absence of conversion to dieldrin,
aldrin would bioconcentrate 4500 times (as dieldrin does),
and that since aldrin only accumulates 32 times, the remainder
of the expected aldrin residues are being stored as dieldrin.
The result is that the water concentration of aldrin
-2
should be less than 4.6 x 10 ng/1 in order to keep the
individual lifetime risk below 10 .
C--dl
-------
Summary of Pertinent Data for Dieldrin
The water quality criterion for dieldrin is based on
the hepatocellular carcinoma response of the female CF-1
mice given 0.1 ppm of dieldrin continuously in the diet
in the experiment of Walker, et al. (1972). In that group
the incidence of type a and type b liver tumors in the 0.1
'ppm group of females was 24 out of 90 animals, whereas in
the controls it was 39 out of 297 animals. Assuming a fish
bioconcentration factor of 4500, the parameters of the dose-
response model are:
nfc = 24 d = 0.1 ppm x 0.13 = 0.013 mg/kg/day
Nfc = 90 L = 132 weeks
n = 39 w = 0.025 kg
c
NC = 297 R = 4500
Le = 132 weeks F = 0.0187 kg/day
le = 132 weeks
With these parameters the slope of the one-hit dose-
response curve for dieldrin is 183.6 (mg/kg/day)
The result is that the water concentration should be
_o
less than 4.4 x 10 ng/1 in order to keep the individual
lifetime risk below 10~ .
<3PO 860 727
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