EPA
United States
Environmental Protection
Agency
Office of Water
Regulations and Standards
Criteria and Standards Division
Washington DC 20460
EPA 440/5-80-038
October 1980
Ambient
Water Quality
Criteria for
DDT
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AMBIENT WATER QUALITY CRITERIA FOR
DDT
Prepared By
U.S. ENVIRONMENTAL PROTECTION AGENCY
Office of Water Regulations and Standards
Criteria and Standards Division
Washington, D.C.
Office of Research and Development
Environmental Criteria and Assessment Office
Cincinnati, Ohio
Carcinogen Assessment Group
Washington, D.C.
Environmental Research Laboratories
Corvalis, Oregon
Duluth, Minnesota
Gulf Breeze, Florida
Narragansett, Rhode Island
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DISCLAIMER
This report has been reviewed by the Environmental Criteria and
Assessment Office, U.S. Environmental Protection Agency, and approved
for publication. Mention of trade names or commercial products does not
constitute endorsement or recommendation for use.
AVAILABILITY NOTICE
This document is available to the public through the National
Technical Information Service, (NTIS), Springfield, Virginia 22161.
11
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FOREWORD
Section 304 (a)(l) of the Clean Water Act of 1977 (P.l. 95-217),
requires the Administrator of the Environmental Protection Agency to
publish criteria for water quality accurately reflecting the latest
scientific knowledge on the kind and extent of all Identifiable effects
on health and welfare which may be expected from the presence of
pollutants 1n any body of water, Including ground water. Proposed water
quality criteria for the 65 toxic pollutants listed under section 307
(a)(l) of the Clean Water Act were developed and a notice of their
availability was published for public comment on March 15, 1979 (44 FR
15926), July 25, 1979 (44 FR 43660), and October 1, 1979 (44 FR 56628).
This document is a revision of those proposed criteria based upon a
consideration of comments received from other Federal Agencies, State
agencies, special Interest groups, and Individual scientists. The
criteria contained In this document replace any previously published EPA
criteria for the 65 pollutants. This criterion document is also
published in satisifaction of paragraph 11 of the Settlement Agreement
in Natural Resources Defense Council, et. al. vs. Train. 8 ERC 2120
(O.D.C. 1976), modified, 12 ERC 1833 (D.D.C. 1979).
The term "water quality criteria" is used in two sections of the
Clean Water Act, section 304 (a)(l) and section 303 (c)(2). The term has
a different program impact in each section. In section 304, the term
represents a non-regulatory, scientific assessment of ecological ef-
fects. The criteria presented in this publication are such scientific
assessments. Such water quality criteria associated with specific
stream uses when adopted as State water quality standards under section
303 become enforceable maximum acceptable levels of a pollutant in
ambient waters. The water quality criteria adopted in the State water
quality standards could have the same numerical limits as the criteria
developed under section 304. However, in many situations States may want
to adjust water quality criteria developed under section 304 to reflect
local environmental conditions and human exposure patterns before
incorporation Into water quality standards. It is not until their
adoption as part of the State water quality standards that the criteria
become regulatory.
Guidelines to assist the States in the modification of criteria
presented in this document, in the development of water quality
standards, and in other water-related programs of this Agency, are being
developed by EPA.
STEVEN SCHATZOW
Deputy Assistant Administrator
Office of Water Regulations and Standards
111
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ACKNOWLEDGEMENTS
Aquatic Life Toxicology:
William A. Brungs, ERL-Narragansett
U.S. Environmental Protection Agency
David J. Hansen, ERL-Gulf Breeze
U.S. Environmental Protection Agency
Mammalian Toxicology and Human Health Effects:
Maria Rabello (author)
University of Texas Medical Branch
James Barnett (author)
University of Texas Medical Branch
Michael L. Dourson (doc. mgr.)
ECAO-Cin
U.S. Environmental Protection Agency
Jerry F. Stara (doc. mgr.) ECAO-Cin
U.S. Environmental Protection Agency
Alfred Garvin
University of Cincinnati
Steven D. Lutkenhoff, ECAO-C1n
U.S. Environmental Protection Agency
Donald P. Morgan
University of Iowa
W. Bruce Pierano, HERL-Cin
U.S. Environmental Protection Agency
Donna Sivulka, ECAO-Cin
U.S. Environmental Protection Agency
Joseph Arcos
Tulane Medical Center
Tom Conner
University of Texas Medical Branch
Patrick Ourkin
Syracuse Research Corporation
William Oykstra, OTS
U.S. Environmental Protection Agency
Wayland Hayes
Vanderbllt University
Fumio Matsamura
Michigan State University
Gary Osweiler
University of Missouri
Shane Que Hee
University of Cincinnati
Roy E. Albert, CAG*
U.S. Environmental Protection Agency
Technical Support Services Staff: D.J. Reisman, M.A. Garlough, B.L. Zwayer,
P.A. Daunt, K.S. Edwards, T.A. Scandura, A.T. Pressley, C.A. Cooper,
M.M. Denessen.
Clerical Staff: C.A. Haynes, S.J. Faehr, L.A. Wade, 0. Jones, B.J. Bordlcks,
B.J. Quesnell, P, Gray, R. Rubinstein.
*CAG Participating Members: Elizabeth L. Anderson, Larry Anderson, Ralph Arnicar,
Steven Bayard, David L. Bayliss, Chao W. Chen, John R. Fowle III, Bernard Haberman,
Charalingayya Hiremath, Chang S. Lao, Robert McGaughy, Jeffrey Rosenblatt,
Dharm V. Singh, and Todd W. Thorslund.
1v
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TABLE OF CONTENTS
Criteria Summary
Introduction A-l
Aquatic Life Toxicology B-l
Introduction B-l
Effects B-2
Acute Toxicity B-2
Chronic Toxicity 8-5
Plant Effects B-6
Residues B-6
Miscellaneous B-ll
Summary B-ll
Criteria B-ll
References B-53
Mammalian Toxicology and Human Health Effects C-l
Exposure C-l
Ingestion from Water C-l
Ingestion from Food C-3
Inhalation C-9
Dermal C-12
Summary C-12
Phannacokinetics C-13
Absorption C-13
Distribution C-14
Metabolism C-21
Excretion C-29
Effects C-31
Acute, Subacute and Chronic Toxicity C-31
Synergism and/or Antagonism C-33
Teratogenicity C-35
Mutagenicity C-38
Carcinogenicity C-45
Criterion Formulation C-64
Existing Guidelines and Standards C-64
Current Levels of Exposure C-66
Special Groups at Risk C-68
Basis and Derivation of Criterion C-69
References C-74
Appendix C-93
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CRITERIA DOCUMENT
DDT AND METABOLITES
CRITERIA
Aquatic Life
DDT
For DDT and its metabolites the criterion to protect fresh-
water aquatic life as derived using the Guidelines is 0.0010 ug/1
as a 24-hour average and the concentration should not exceed 1.1
ug/1 at any time.
For DDT and its metabolites the criterion to protect saltwater
aquatic life as derived using the Guidelines is 0.0010 ug/1 as a 24
hour average and the concentration should not exceed 0.13 ug/1 at
any time.
TDE
The available data for TDE indicate that acute toxicity to
freshwater aquatic life occurs at concentrations as low as 0.6 ug/1
and would occur at lower concentrations among species that are more
sensitive than those tested. No data are available concerning the
chronic toxicity of TDE to sensitive freshwater aquatic life.
The available data for TDE indicate that acute toxicity to
saltwater aquatic life occurs at concentrations as low as 3.6 ug/1
and would occur at lower concentrations among species that are more
sensitive than those tested. No data are available concerning the
chronic toxicity of TDE to sensitive saltwater aquatic life.
DDE
The available data for DDE indicate that acute toxicity to
freshwater aquatic life occurs at concentrations as low as 1,050
M9/1 and would occur at lower concentrations among species that are
vi
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more sensitive than those tested. No data are available concerning
the chronic toxicity of DDE to sensitive freshwater aquatic life.
The available data for DDE indicate that acute toxicity to
saltwater aquatic life occurs at concentrations as low as 14 ug/1
and would occur at lower concentrations among species that are more
sensitive than those tested. No data are available concerning the
chronic toxicity of DDE to sensitive saltwater aquatic life.
Human Health
For the maximum protection of human health from the potential
carcinogenic effects due to exposure of DDT through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentration should be zero based on the non-threshold
assumption for this chemical. However, zero level may not be
attainable at the present time. Therefore, the levels which may
result in incremental increase of cancer risk over the lifetime are
estimated at 10 , 10 and 10~ . The corresponding recommended
criteria are 0.24 ng/1, 0.024 ng/1, and 0.0024 ng/1, respectively.
If the above estimates are made for consumption of aquatic organ-
isms only, excluding consumption of water, the levels are 0.24
ng/1, 0.024 ng/1, and 0.0024 ng/1, respectively.
vii
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INTRODUCTION
DDT, first synthesized in Germany in 1874, has been used ex-
tensively world-wide for public health and agricultural programs.
Its efficacy as a broad spectrum insecticide and its low cost make
it the insecticide for those measures for most of the world.
Following an extensive review of health and environmental haz-
ards of DDT, U.S. EPA decided to ban its further use. This decision
was based on several well evidenced properties such as: (1) DOT and
its metabolites are toxicants with long-term persistence in soil
and water, (2) it is widely dispersed by erosion, runoff and vola-
tilization, (3) the low-water solubility and high lipophilicity of
DDT result in concentrated accumulation of DDT in the fat of wild-
life and humans which may be hazardous. Agricultural use of DDT
was cancelled by the U.S. EPA in December, 1972. Prior to this, DDT
had been widely used in the U.S. with a peak usage in 1959 of 80
million pounds. This amount decreased steadily to less than 12
million pounds by 1972. Since the 1972 ban, the use of DDT in the
U.S. has been effectively discontinued.
Table 1 gives abbreviations and their meanings as used in the
text of this document. The physical properties of DDT isomers are
listed as well.
A-l
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TABLE 1
DDT and Its Metabolites
DDT refers to technical DDT, which is usually composed of:
77.1% p,p'-DDT
14.9% o,p'-DDT
0.3% p,p'-DDD R
0.1% 0,p'-DDD
4.0% p,p'-DDE
0.1% 0,p'-DDE
3.5% unidentified compounds
DDT l,l'-(2,2,2-trichloroethyli-
dene)-bis/4-chlorobenzene/
DDE l,l'-(2,2-dichloroethenyli-
dene)-bis/4-chlorobenzene/
ODD 1,1' - (2,2-d ichloroe thylidene)
bis/4-chlorobenzene/
DDMU l,l'-(2-chloroethenylidene)-
bis/4-chlorobenzene/
DDMS l,l'-(2-chloroethylidene)-
bis/4-chlorobenzene/
DDNU l,l-bis(4-chlorophenyl)
ethylene
DDOH 2,2-bis(4-chlorophenyl)
ethanol
DDA 2,2-bis(4-chlorophenyl)-
acetic acid
R
•Cl
•Cl
•Cl
•Cl
•Cl
•Cl
•Cl
•Cl
R'
-H
None
-H
None
-H
None
-H
-H
R"
-cci3
-cci2
-CHC12
-CHC1
-CH2C1
-CH2
-CH2OH
-C-OH
II
0
A-2
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Physical Properties
The general physical properties of
below:
Molecular weight
(Windholz, M. (ed.), 1976)
Melting point
(Gunther and Gunther, 1971)
Boiling Point
(Gunther and Gunther, 1971)
Vapor pressure
(Martin, 1972)
(Spencer, 1975)
(Metcalf, 1972) at 20°C
Solubility in water at 25°C
(Weil et al., 1974)
(Biggar and Riggs, 1974)*
Metcalf, 1972)
(Bowman et al., 1960)
Log octanol/water
partition coefficient
(O'Brien, 1974)
(Kengaga and Goring, 1978)
(Wolfe et al., 1977)
(Kapoor et al., 1973)
the DDT isomers are given
354.5
108.5-109.0°C (pp1)
74-74.5°C (opf)
185°C (pp')
1.9
,-7
x 10 : torr (pp1) at 25"C
7.3 x 10"^ torr (pp») at 30°C
5.5 x 10", torr (op1) at 30°C
1.5 x 10"' torr (pp') at 20°C
5.5 ppb (pp()
26 ppb (op1)
25 ppb (ppf)
85 ppb (op'}
~2 ppb
<1.2 ppb (pp1)
6.19 (pp1, calc.)
5.98
4.89
3.98 (pp1, measured)
*Particle size <5.0 um.
A-3
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REFERENCES
Bowman, M.C., et al. 1960. Solubility of carbon-14 DDT in water.
Jour. Agric. Food Chem. 8: 406.
Gunthr, F.A. and J.D. Gunther. 1971. Residues of pesticides and
other foreign chemicals in foods and feeds. Res. Rev. 36: 69.
Kapoor, I.P., et al. 1973. Structure activity correlations of
biodegradability of DDT analogs. Jour. Agric. Food Chem. 21: 310.
Kenaga, E.E. and C.A. Goring. 1978. Relationship between water
solubility, soil-sorption, octanol/water partitioning, and biocon-
centration of chemicals in biota. Am. Soc. Test. Mater. Third
Aquatic Tox. Symp., New Orleans, Louisiana.
Martin, H., (ed. ) 1972. Pesticide Manual. 3rd ed. Br. Crop Prot.
Counc., Worcester, England.
Metcalf, R.L. 1972. DDT substitutes. Grit. Rev. Environ. Con-
trol. 3: 25.
O'Brien, R.D. 1974. Nonenzymic Effects of Pesticides on Mem-
branes. In: R. Hague and V.H. Freed, (eds.) Environmental Dynamics
of Pesticides. Plenum Press, New York. p. 331.
A-4
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Spencer, W.F. 1975. Movement of DDT and its derivatives into the
atmosphere. Res. Rev. 59: 91.
Weil, L., et al. 1974. Solubility in water of insecticide chlo-
rinated hydrocarbons and polychlorinated biphenyls in view of water
pollution. Z. Wasser Abwasser Forsch. 7(6): 169.
Windholz, M. (ed.) 1976. The Merck Index, 9th ed. Merck and Co.,
Inc. Rahway, New Jersey.
Wolfe, N.L., et al. 1977. Methoxychlor and DDT degradation in
water: Rates and products. Environ. Sci. Technol. 7(110): 1077.
A-5
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Aquatic Life Toxicology*
INTRODUCTION
DDT, a chlorinated hydrocarbon insecticide, was at one time the most
widely used chemical for the control of insect pests. It was applied for
more than 30 years to a variety of environments, including the aquatic en-
vironment, in many forms such as powders, emulsions, and encapsulations.
DDT has probably been subjected to more investigations than any of the other
chlorinated hydrocarbon pesticides such as aldrin, dieldrin, endrin, chlor-
dane, and toxaphene.
DDT is a persistent, lipid-soluble pesticide. Long-lived pesticides
provide control of target organisms over extended periods of time and reduce
the need for reapplication, but may also affect non-target flora and fauna
for long periods of time. Because of its persistent nature, coupled with
hydrophobic properties and solubility in lipids, DDT and its metabolites are
concentrated by aquatic organisms at all trophic levels from water, enter
the food web, and are bioaccumulated by organisms at higher trophic levels.
DDT has several metabolites; the two most frequently found in nature are
TDE (ODD or Rhothane) and DDE. TOE was manufactured as an insecticide and
used for a number of years. Most of the available aquatic toxicity data are
for DDT. However, because of their widespread occurrence and particularly
their toxicities to consumer species, TDE and DDE are included in this cri-
terion document.
DDT is intermediate in toxicity to fishes in comparison to other chlori-
nated hydrocarbon pesticides. It is less toxic than aldrin, dieldrin, endrin,
*The reader is referred to the Guidelines for Deriving Water Quality Cri-
teria for the Protection of Aquatic Life and Its uses in order to understand
this section better. The attached tables contain pertinent available data,
and at the bottoms of the appropriate tables are calculations deriving vari-
ous measures of toxicity as described in the Guidelines.
B-l
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and toxaphene, but more toxic than chlordane, lindane, and methoxychlor
(Henderson, et al. 1959; Katz, 1961).
Most acute toxicity data for DOT are from static tests; few
flow-through studies have been conducted. Relatively few data are available
that describe the chronic effects of DDT on aquatic animals. Chronic test
data are available for only one species of freshwater fish, and no life-cy-
cle toxlcity test has been conducted on a freshwater invertebrate species
nor on any saltwater species. Few data are available on effects of DDT on
plants.
Many references on bioconcentration data are available. However, a num-
ber of these were not usable, either because it appeared that a steady-state
condition was not reached in laboratory experiments or, in the case of field
monitoring, adequate documentation of the concentration of DDT in the water
was not available.
Derivation of a DOT criterion must consider not only acute and chronic
toxicity to aquatic organisms, but also its propensity for bioaccumulation,
its breakdown Into long-lived metabolites, and the toxicity of DDT and its
metabolites to organisms at higher trophic levels, such as birds of prey, as
a result of food chain bioaccumulation.
Data discussed in the following sections are for DDT unless otherwise
specified.
EFFECTS
Acute Toxlcity
Acute toxlcity data are available for 18 freshwater invertebrate species
for a total of 46 data points (Table 1). Invertebrate species for the most
part are more sensitive than fish species, but the range of invertebrate
species LC5Q values (10,000 times) is greater than that (300 times) for
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fishes. The least sensitive invertebrate species is the stonefly, Pteronar-
cys californica, for which a 96-hour LCcQ value of 1,800 gg/1 was
determined (Gaufin, et al. 1965). This LCcQ is almost 16 times greater
than the geometric mean of the other three LC5Q values for the same spe-
cies (Table 1); however, no valid reason to discount this value could be
found. The most sensitive aquatic invertebrate species is a crayfish, Or-
conectes nais, with an LC5Q of 0.18 vg/1 for 1-week-old organisms (San-
ders, 1972). However, 10-week-old crayfish of the same species had an
LC5Q of 30 yg/1.
Only two of the acute values for freshwater invertebrate species (Table
1) were derived from flow-through tests, and none were from a test with mea-
sured toxicant concentrations. The result of one flow-through test in Table
1 is one-fourth of the static test result for the same species of scud, Gam-
marus fasciatus (Sanders, 1972), whereas in another comparison the result of
a static test is lower than the result from a flow-through test with the
glass shrimp, Palaemonetes kadiakensis (Sanders, 1972). This difference may
be due to a difference between species or to experimental variability. TDE
is more toxic than DDT to three invertebrate species (a glass shrimp, PaTae-
monetes kadiakensis, and two species of scud, Gammarus fasciatus and Gam-
marus lacustris), but less toxic than DDT to the cladocerans, Daphnia pulex
and Simocephalus serrulatus, and the sowbug, Asellus brevicaudus (Table 1).
Data are available for 24 freshwater fish species for a total of 107
values (Table 1). Two of the LC5Q values are from flow-through tests, and
the rest are from static tests. The flow-through LC5Q value (unmeasured
concentrations) for rainbow trout fry (Tooby, et al. 1975) is equal to or
less than 85 percent of the 13 static values for the same species. The only
flow-through test with a measured toxicant concentration (Jarvinen, et al.
B-3
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1977) is for the fathead minnow, and the LCrr, value is greater than 87
percent of all static LC5Q values for the same species. Since the water
solubility of DDT is not high, it would be expected that static tests would
underestimate toxicity as indicated by the rainbow trout data. The fathead
minnow data, however, are in contrast to this, perhaps because of large var-
iability for this species with DOT. Lincer, et al. (1970) demonstrated that
the fathead minnow was more sensitive to DOT in the static than in the flow-
through test (48-hour static = 7.4 ug/1, 48-hour flow-through - >40 ug/1),
and Macek and Sanders (1970) determined that among five fish species tested,
variation in susceptibility to DDT was greatest in the fathead minnow. In-
terspecific variability, shown by the LCgg values in Table 1, indicates
that the fathead minnow is more variable than 87 percent of the 24 species
for which there are data available. Only three species are more variable:
goldfish, guppy, and brook trout, with the goldfish being the most variable.
The yellow perch is the fish species most sensitive to DOT (96-hour
LCrQ of 0.6 ug/1) (Marking, 1966), whereas the least sensitive species is
the goldfish (96-hour LC5Q of 180 ug/1) (Marking, 1966). Therefore, the
range of species sensitivity for the tested fishes is 300 times.
The Freshwater Final Acute Value for DOT, derived from the species mean
acute values listed in Table 3 using the procedure described in the
Guidelines, is 1.1 ug/1. Acute data for TOE and DDE are insufficient to
determine a Freshwater Final Acute Value for these compounds.
Acute toxicity tests on six saltwater invertebrate species (Table 1)
produced acute LC50 values from 0.14 to 9.0 ug/U the lowest value is the
96-hour LCgQ for the brown shrimp (Penaeus aztecus). Data are available
for a mollusc and four different families of arthropods. Table 6 reports
24- or 48-hour EC™ values for five species giving EC™ values ranging
from 0.6 to 10 ug/1.
B-4
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Acute 96-hour toxicity tests with 11 species of saltwater fishes (repre-
senting nine fish families) gave LCgg values ranging from 0.26 to 89
ug/1. The northern puffer was by far the least sensitive; most other IC™
values for fish species range between 0.5 and 7 vg/1. Table 6 reports
48-hour LCjQ values for six species with LC50 values from 0.32 to 3.2
The Saltwater Final Acute Value for DOT, derived from the species mean
acute values listed in Table 3 using the procedure described in the Guide-
lines, is 0.13 pg/1.
In tests on TOE, 96-hour l_C50 values are reported for three saltwater
species. The acute values range from 1.6 to 25 wg/1. Results of 48-hour
tests on pink shrimp (Penaeus duorarum) and the longnose killifish (Fundulus
slmilis) provide LCgo values of 2.4 and 42 yg/1, respectively (Table 6).
Test data on TOE are insufficient to provide a Saltwater Final Acute Value
according to the Guidelines.
In the only available 96-hour test on DDE, the ECgQ (based on shell
deposition) for the oyster, Oassostrea virginica, was 14 yg/l. In tests
lasting 48 hours (Table 6), two species were exposed to DOE. The 48-hour
LCg0 value for the brown shrimp was 28 wg/1; that for spot was 20 yg/1.
Test data on DDE are insufficient to provide a Saltwater Final Acute Value
according to the Guidelines.
Chronic Toxicity
Chronic toxicity data for DDT for are available for only one freshwater
fish species, the fathead minnow (Jarvinen, et al. 1977). The chronic value
for this study is 0.74 yg/1 (Table 2). The comparable 96-hour LC50 value
(48 ug/1) from the same study is 65 times higher than the chronic toxicity
value.
8-5
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No chronic toxicity data were found for any freshwater invertebrate spe-
cies nor for any saltwater animal species.
Because the available data do not meet the minimum data base requirement
set forth in the Guidelines, no Final Chronic Values can be determined for
DOT, TOE, or ODE.
Plant Effects
Four species of freshwater algae (Table 4) have a wide range of sensi-
tivity (2,700 times), with most plant values being above the Final Acute
Value for aquatic animals. The lowest effect value for plants is 0.3 ug/1,
determined from the growth and morphology data for Chlorella sp. (Sodergren,
1968).
Information on the sensitivity of saltwater aquatic plant species, in-
cluding algae and rooted vascular plants, is limited (Table 4) but indicates
that they are much less sensitive to DOT than are fish or invertebrate spe-
cies. DOT at a concentration of 10 ug/1 has been found to reduce photosyn-
thesis in saltwater diatoms, green algae, and dinoflagellates (Wurster,
1968).
Residues
Twenty-four field-generated data points for 22 freshwater fish and in-
vertebrate species are available, whereas 18 laboratory-generated data
points for 16 fish and invertebrate species are available (Table 5). Fresh-
water fish species bioconcentration in the field was much greater than in
laboratory tests, which may be due to a difference in the physical form of
the toxicant between field and laboratory studies, the many additional tro-
phic levels involved in field exposures, or a difference in lipid content of
the tissues.
B-6
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Bioconcentration factors are available for three saltwater invertebrate
and nine fish species (Table 5). Odum, et al. (1969) fed fiddler crabs a
diet of natural detritus containing DDT residues of 10 mg/kg. After five
days, crabs fed DDT-contaminated detritus exhibited extremely poor coordina-
tion. They concluded that although no crabs died, such behavior would
"almost certainly affect survival under natural conditions." After 11 days
on the diet, concentrations of DDT and metabolites increase threefold in
their tissues to 0.885 mg/kg. Odum, et al. (1969) speculated that the re-
sults of this study may help to explain the disappearance of this species
from a Long Island marsh sprayed with DDT for more than 15 years.
Bioconcentration factors from laboratory tests with DDT and saltwater
organisms ranged from 1,200 to 76,300 for fish and shellfish (Lowe, et al.
1970; Nimmo, et al. 1970). Eastern oysters provided BCF values from 42,400
in a 252-day exposure to 76,300 in a 168-day exposure (Lowe, et al. 1970).
For saltwater organisms bioconcentration factors for DDT determined from an-
imals captured from their natural environments were comparable with those
from laboratory studies (Table 5). BCF values in these studies ranged from
4,750 times for Cancer magister to 46,500 times for the dwarf perch (Earnest
and Benville, 1971).
Data for DDE in Table 5 pertaining to maximum permissible tissue concen-
trations indicate that long-term dietary dosage at 2.8 to 3 mg/kg DDE (wet
weight) can have adverse effects on reproduction of mallards (Heath, et al.
1969; Haseltine, et al. 1974), black ducks (Longcore, et al. 1971; Longcore
and Stendell, 1977), and screech owls (McLane and Hall, 1972). DDE has been
found to constitute 50 to 90 percent of the DDT analogs present in fish
(Jarvinen, et al. 1977).
B-7
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Avian species that feed on saltwater animals containing DOT and metabo-
lites (particularly ODE) have exhibited reductions in their reproductive
capacity. For example, a colony of Bermuda petrels, a species which feeds
primarily on cephalopods in the North Atlantic, suffered a significant de-
cline in their population from 1958 to 1967 (Wurster, 1968). Analysis of
unhatched eggs and dead chicks revealed an average concentration of 6.4
mg/kg DDT and metabolites (62 percent DOE). No data are available on the
concentrations of DDT and metabolites in the cephalopods consumed by the
petrels.
Studies have been conducted to evaluate the effects of DOT and metabo-
lites in eggs of the brown pelican and the subsequent decrease in reproduc-
tive success. Blus, et al. (1974) reported that their reproductive success
was normal only when concentrations of DDT (including metabolites) and the
insecticide, dieldrin, were less than 2.5 mg/kg and 0.54 mg/kg, respective-
ly. The DDE concentration causing shell thinning was estimated to be 0.5
mg/kg or less in eggs of brown pelicans (Blus, et al. 1972). Much higher
concentrations in the eggs than concentrations that were fed for several
months have been found for other species. Ten times higher concentrations
were observed in black duck eggs {Longcore, et al. 1971; Longcore and Sten-
dell, 1977) and almost eight times higher in sparrow hawk eggs (Lincer,
1975).
Anderson, et al. (1975) studied the breeding success of the brown peli-
can in relation to residues of DOT and metabolites in their eggs and in
their major food source, the northern anchovy. Their analyses of data col-
lected from 1969 to 1974 included the following observations: (1) residues
of DDT and metabolites (the major compound was ODE) in northern anchovies
dropped steadily from a mean of 3.4 mg/kg (wet weight) in 1969 to 0.15 mg/kg
8-8
-------
1n 1974; (2) during that same period, DOT and metabolites in intact eggs
averaged 907 mg/kg (lipid weight) in 1969 to 97 mg/kg in 1974, and higher
residues were associated with crushed eggs; (3) productivity of pelicans
increased from a total of four young fledged in 1969 to 1,115 fledged in
1974, with a concurrent increase in eggshell thickness. Anderson, et al.
(1975) stated that even the lowest concentration of DOT and metabolites in
northern anchovies (0.15 mg/kg) and the subsequent 97 mg/kg concentration in
pelican eggs were unacceptably high, because the pelican eggshell thickness
was below normal and productivity was too low for population stability.
Dividing a BCF value by the percent lipid value for the same species
provides a BCF value adjusted to 1 percent lipid content; this resultant BCF
value is referred to as the normalized BCF. The geometric mean of normal-
ized bioconcentration factors for DDT for freshwater and saltwater aquatic
life is 17,870 (Table 5).
It is noteworthy that only one of the fish species listed in Table 5 is
a species that belongs to the order Clupeiformes. Clupeids are a major food
source for brown pelicans and are high in lipid content. Due to the lipo-
philic nature of DDT and its metaboltes, it is likely that these fishes
would contain higher concentrations of the insecticide than would fishes of
lower lipid content; indeed, the BCF for alewife, a clupeid, (1,296,666) is
nearly two orders of magnitude higher than the geometric mean BCF. There-
fore, the mean normalized bioconcentration factor of 17,870 may underesti-
mate the bioconcentration factors likely to occur in clupeid species.
Because no safe concentration for DDT and metabolites in food of pelicans is
known and because the mean bioconcentration factor may be too low, the resi-
due values based on these data may be underprotective.
B-9
-------
Dividing the FDA action level of 5.0 mg/kg for fish by the geometric
mean of normalized BCF values (17,870) and by a percent lipid value of 15
for freshwater species (see Guidelines) gives a freshwater residue value
based on marketability for human consumption of 0.019 yg/1 (Table 5). Di-
viding the FDA action level (5.0 mg/kg) by the geometric mean of normalized
BCF values (17,870) and by a percent lipid value of 16 for saltwater species
(see Guidelines) gives a saltwater residue value of 0.017 ug/1. Also based
on marketability for human consumption, using the FDA action level and the
highest BCF for edible portion of a consumed fish species (458,259 for lake
trout for freshwater), a residue value of 0.011 vg/1 is obtained for fresh-
water (Table 5). No appropriate BCF value for edible portion of a consumed
fish species is available for saltwater.
A residue value for wildlife protection of 0.0010 yg/1 is obtained for
both freshwater and saltwater using the lowest maximum permissible tissue
concentration of 0.15 mg/kg based on reduced productivity of the brown peli-
can (Anderson, et al. 1975). Average lipid content of pelican diets is un-
available. Clupeids usually constitute the major prey of pelicans, and the
percent lipid value of the clupeid, northern anchovy, is 8 (Reintjes,
1980). The northern anchovy is in some areas a major food source of the
brown pelican. Therefore, the percent lipid value of 8 was used for the
calculation of the Final Residue Value. The value of 0.15 mg/kg divided by
the geometric mean of normalized BCF values (17,870) and by a percent lipid
value of 8 gives a residue value of 0.0010 ug/1 (Table 5).
Selection of the lowest freshwater and saltwater residue values from the
above calculations gives a Freshwater Final Residue Value of 0.0010 ug/1 and
a Saltwater Final Residue Value of 0.0010 ug/1. The Final Residue Values
may be too high because they are based on a concentration which reduced the
productivity of the brown pelican.
8-10
-------
Miscellaneous
Table 6 contains additional data concerning the effect of DOT on 23 spe-
cies of freshwater and 13 species of saltwater aauatic life. The values
range from LCcQ values for time oeriods that are either less or greater
than specified in the Guidelines to physiological and behavioral effects.
The lowest value in Table 6 is a hyperactive locomotor response observed by
Ellgaard, et al. (1977) for the bluegill exposed at 0.008 ug/1. This value
is slightly higher than the Freshwater Final Residue Value.
Results of acute toxicity tests shown in Table 6 indicate that the pin-
fish (Lagodon rhomboides) was the species most sensitive to DOT (48-hour
LCcQ = 0.32 ug/1,)(Lowe, undated). This is the lowest value of all acute
values with fishes (Tables 1 and 6); however, this value is not below the
Saltwater Final Acute Value of 0.13 ug/1. The LC5Q values for other spe-
cies in acute tests lasting less than 96 hours lie between 0.4 and 5.5 ug/1
for DDT, and between 2.4 and 42 ug/1 for TOE and DDE. No other data from
Table 6 suggest any more sensitive effects or greater bioconcentration than
that found in the previous tables.
Summary
Acute toxicity data for DDT are available for 18 freshwater invertebrate
species; a wide range in species sensitivity was found, with acute values
ranging from 0.18 to 1,800 ug/1. Acute toxicity tests on 24 freshwater fish
species also showed a wide range of species sensitivity, with LC5Q values
ranging from 0.6 ug/1 for yellow perch to 180 ug/1 for goldfish. Few data
are available concerning effects on freshwater plants, and those that are
available indicate a wide range of concentrations at which effects occur.
A Freshwater Final Acute Value of 1.1 ug/1 was obtained for DDT based on
data for 42 species. A single chronic value of 0.74 ug/1 DOT was obtained
for the fathead minnow. Based on a maximum permissible tissue concentration
R-11
-------
of 0.15 mg/kg for wildlife protection, the geometric mean of normalized bio-
concentration factors (17,870), and a percent lipid value of 8, the Fresh-
water Final Residue Value for DDT is 0.0010 ug/1.
Acute toxicity data for DOT and six saltwater invertebrate species indi-
cate that the brown shrimp, with a 96-hour LC,-0 of 0.14 ug/1 is the most
sensitive species of those tested. Acute tests on 11 saltwater fish species
gave LC5Q values ranging from 0.26 to 89 ug/1. No chronic data are avail-
able for any saltwater species. From limited data, saltwater plants appear
to be much less sensitive than fish or invertebrate species to DOT.
A Saltwater Final Acute Value of 0.13 ug/1 was obtained for DDT based on
data for 17 species. No DOT Saltwater Final Chronic Value can be calculated
because insufficient data are available. Based on a maximum permissible
tissue concentration of 0.15 mg/kg, the geometric mean of normalized biocon-
centration factors (17,870), and a percent lipid value of 8, the Saltwater
Final Residue Value for DOT is 0.0010 ug/l.
It should be pointed out that the Final Residue Values may be too high
because. Average lipid content of pelican diets is unavailable. Clupeids
usually constitute the major prey of pelicans, and the percent lipid value
of the clupeid, northern anchovy, is 8 (Reintjes, 1980). The northern
anchovy is in some areas a major food source of the brown pelican.
Therefore, the percent lipid value of 8 was used for the calculation of the
Final Residue Value.
CRITERIA
DDT
For DDT "and its metabolites the criterion to protect freshwater aauatic
life as derived using the Guidelines is 0.0010 ug/1 as a 24-hour average,
and the concentration should not exceed 1.1 ug/1 at any time.
8-12
-------
For DOT and its metabolites the criterion to protect saltwater aquatic
life as derived using the Guidelines is 0.0010 wg/1 as a 24-Jiour average,
and the concentration should not exceed 0.13 ug/1 at any time.
IDE
The available data for TDE indicate that acute toxicity to freshwater
aquatic life occurs at concentrations as low as 0.6 vg/1 and would occur at
lower concentrations among species that are more sensitive than those
tested. No data are available concerning the chronic toxicity of TDE to
sensitive freshwater aquatic life.
The available data for TDE indicate that acute toxicity to saltwater
aquatic life occurs at concentrations as low as 3.6 jig/1 and would occur at
lower concentrations among species that are more senstive than those
tested. No data are available concerning the chronic toxicity of TDE to
sensitive saltwater aquatic life.
DDE
The available data for DDE indicate that acute toxicity to freshwater
aquatic life occurs at concentrations as low as 1,050 ug/1 and would occur
at lower concentrations among species that are more sensitive than those
tested. No data are available concerning the chronic toxicity of DDE to
sensitive freshwater aquatic life.
The available data for DDE indicate that acute toxicity to saltwater
aquatic life occurs at concentrations as low as 14 ug/1 and would occur at
lower concentrations among species that are more sensitive than those
tested. No data are available concerning the chronic toxicity of DDE to
sensitive saltwater aquatic life.
B-13
-------
Table 1. Acute values for DOT and metabolites
Species
Method"
LC50/EC50
Species Mean
Acute Value
(ug/l) Reference
FRESHWATER SPECIES
Cladoceran,
Daphnla maqna
Cladoceran,
Paphnla magna
Cladoceran.
Daphnla pulex
Cladoceran,
S I mocepha 1 us serrulatus
Cladoceran,
Slmocephalus serrulatus
Sow bug,
Asellus brevlcaudus
Scud,
Gamnarus fasclatus
Scud,
Gommarus fasclatus
Scud,
Ganmarus fasclatus
Scud,
Gommarus 1 acustr 1 s
Scud,
Gamnarus 1 acustr Is
Seed shrimp,
Cyprldopsls vldua
Glass shrimp,
Palaemonetes kadlakensls
S.
s.
s.
s.
s,
s,
s,
FT.
s,
s.
s,
s,
s.
u
u
u
u
u
u
u
u
u
u
u
u
u
DOT
4
1.48
0.56
2.5
2.8
4
3.2
0.8
1.8
9
1
54
4.2
Macek &
Sanders
, 1970
2.4 Prlester, J965
0.36 Sanders
Sanders
2.6 Sanders
4.0 Sanders,
Sanders ,
Sanders,
1.7 Sanders,
Gaufln,
3.0 Sanders,
54 Macek &
Macek &
& Cope,
& Cope,
& Cope,
1972
1972
1972
1972
et al .
1969
Sanders
Sanders
1966
1966
1966
1965
, 1970
, 1970
B-14
-------
Table 1. {Continued)
Species
Glass shrimp,
Palaemonetes kadlakensls
Glass shrimp,
PalaeMonetes kadlakensls
Crayfish,
Orconectes nals
Crayfish (1-day-old),
Orconectes nals
Crayfish (1-wk-old),
Orconectes nals
Crayfish (2-wk-old),
Orconectes nals
Crayfish (3-wk-old),
Orconectes nals
Crayfish (5-wk-old),
Orconectes nals
Crayfish (8-wk-old),
Orconectes nals
Crayfish (10-wk-old),
Orconectes nals
Crayfish,
Procambarus acutus
Mayfly,
Ephemerel la grand Is
Stonefly,
Acroneurla paclflca
Stonefly,
Acroneurla paclflca
Method*
s. u
FT. U
S, U
S. U
S, U
S, U
S, U
s, u
S, U
S. U
S, U
s, u
s, u
s, u
LC50/EC50
(to/I)
2.3
3.5
100
0.30
0.16
0.20
0.24
0.90
28
30
3
25
410
320
Species
Acute Value
3.2
1.9
3
25
362
Reference
Sandors, 1972
Sanders, 1972
Sanders, 1972
Sanders, 1972
Sanders, 1972
Sanders, 1972
Sanders, 1972
Sanders. 1972
Sanders, 1972
Sanders, 1972
Albaugn, 1972
Gaufln, et al. 1965
Gaufln, et al. 1961
Gaufln, et al. 1965
B-15
-------
Table I. (Continued)
Stonefly,
Claassenla sabulosa
Stonefly,
Pteronarcalla bad Ia
Stonefly,
Pteronarcys ca11f ornIca
Stonefly,
Pteronarcys callfornlca
StonefIy,
Pteronarcys callfornlca
Caddlsfly,
Arctopsyche grandIs
Caddlsfly,
Hydropsyche ca11fornIca
Planarlan,
Polycells fellna
Cono salmon,
Oncorhynchus klsutch
Cono salmon,
Oncorhynchus klsutch
Cono salMon,
Oncorhynchus klsutch
Cono salmon,
Oncorhynchus klsutch
Cono salmon,
Oncorhynchus klsutch
Chinook salmon,
Oncorhynchus tshawytscha
Method*
S, U
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
Species Mean
LC50/EC50 Acute Value
(ua/D (uq/l)
3.5 3.5
1.9 1.9
1,600
7
560 192
175 175
48 48
1,230 1,230
44
4
11.3
18.5
13 14
11.5 12
Reference
Sanders & Cope, 1968
Sanders & Cope, 1968
Gaufln, et al. 1965
Sanders & Cope, 1968
Gaufln, et al. 1961
Gaufln, et al. 1965
Gaufln, et al. 1965
Kouyoumjfan & Uglov,
1974
Katz. 1961
Macek & McAllister,
1970
Post & Schroeder,
1971
Post & Schroeder,
1971
Schaumburg, et al.
1967
Katz, 1961
B-16
-------
Table 1. (Continued)
Spec las
Cutthroat trout.
Sal mo dark)
Cutthroat trout.
Sal mo clarkl
Rainbow trout.
Sal mo galrdnerl
Rainbow trout.
Sal mo galrdnerl
Rainbow trout.
So (no (joirdnerl
Rainbow trout,
Sal mo galrdnerl
Rainbow trout.
Sol mo galrdnerl
Rainbow trout.
Sal no galrdnerl
Rainbow trout,
Sal mo galrdnerl
Rainbow trout.
Sal mo galrdnerl
Rainbow trout.
Sal mo galrdnerl
Rainbow trout.
Sal mo galrdnerl
Rainbow trout,
Sal mo galrdnerl
Rainbow trout,
Sal mo galrdnerl
Species Mean
LC50/EC50 Acute Value
Method* (ug/l) (ug/l) Reference
S, U 0.85 - Post A Schroeder,
1971
S, U 1.37 I.I Post 4 Schroeder,
1971
S, U 42 - Katz, 1961
S. U 7 - Macek & McAllister,
1970
S, U 7.2 - Macek & Sanders, 1970
S, U U - Marking, 1966
S, U 4.6 - Marking, 1966
S, U 7.2 - Marking, 1966
5. U 15 - Marking, 1966
S, U 17 - Marking, 1966
S, U 13 - Marking, 1966
S, U 12 - Marking, 1966
S, U 2.4 - Marking, 1966
S, U 1.7 - Post & Schroeder,
1971
B-17
-------
Table 1. (Continued)
Species
Rainbow -trout (fry),
Sal mo qalrdnerl
Brown trout (f Inger 11 ng),
Sal mo trutta
Brown trout,
Salmo trutta
Brown trout,
Salmo trutta
Brook trout,
SnivelInus fontlnalls
Brook trout,
SalveIInus fontlnalls
Brook trout,
Salve! Inus fonttnaljs
Brook trout,
Salvellnus fontlnalls
Brook trout,
Salvellnus fontlnalls
Brook trout,
SalvolInus fontlnalls
Lake trout,
SaI ye 11nus namaycush
Lake trout,
Sal veilnus namaycush
Northern pike,
Esox luclus
Goldfish,
Carassius auratus
Method*
FT, U
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
LC50/EC50
CUO/M
2.4
17.5
2
10.9
7.2
17
20
1.8
7.4
11.9
9.1
9.5
1.7
21
Spectes MMH
Acute Value
(UQ/I)
7.B
-
-
7.3
-
-
-
-
-
8.5
-
9.3
1.7
_
Reference
Tooby, et al. 1975
King, 1962
Macek & McAl lister,
1970
Marking, 1966
Marking, 1966
Marking, 1966
Marking, 1966
Marking, 1966
Post A Schroeder,
1971
Post & Schroecter,
197t
Marking, 1966
Marking, 1966
Marking, 1966
Macek & McAl lister,
1970
B-18
-------
Table 1. (Continued)
Species
Goldfish,
Car ass 1 us auratus
Goldfish,
Car ass 1 us auratus
Goldfish,
Car ass 1 us auratus
Goldfish,
Car ass 1 us auratus
Goldfish,
Car ass 1 us auratus
Goldfish,
Carasslus auratus
Goldfish,
Carasslus auratus
Goldfish.
Car ass I us auratus
Northern redbel ly dace,
Chrosomus eos
Carp,
Cyprlmis carplo
Cerp,
Cyprlnus carplo
Carp,
Cyprlnus carplo
Carp,
Cyprlnus carplo
Carp,
Cyprlnus carplo
Method*
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
(J
u
u
LC50/EC50
(iig/l)
76
27
32
180
40
35
21
36
68
10
9.2
4.0
11.3
12
Species Mean
Acute Value
(uq/l) Reference
Marking, 1966
Marking, 1966
Marking, 1966
Marking, 1966
Marking, 1966
Marking, 1966
Marking, 1966
40 Henderson, et al.
1959
68 Marking, 1966
Macek & McAllister,
1970
Marking,
Marking,
Marking,
Marking,
1966
1966
1966
1966
B-19
-------
T»ble 1. (Continued)
Species
Carp,
Cyprlnus carplo
Carp,
Cyprlnus carplo
Fathead minnow,
Plmephales promelas
Fathead minnow,
Plmephales protnelas
Fathead minnow,
Plmephales promelas
Fathead minnow,
Plmephales promelos
Fathead minnow,
Plmephates promelos
Fathead minnow,
Plmephales promelas
Fathead minnow,
Plmephales promelas
Fathead minnow,
Plmephales promelas
Black bullhead,
Ictalurus melas
Slack bullhead,
Ictalurus melas
Black bullhead,
tctalurus melas
Slack bullhead,
Ictalurus melas
Method*
s.
s.
FT,
s.
s.
s.
s,
5,
s.
s.
s.
s,
s,
s,
u
u
M
U
U
U
U
U
U
u
u
u
u
u
LC50/EC50
(uo/D
6.9
6
48
19
19.9
58
42
45
26
26
5
42
23.5
17
Specie* Mewi
Acute Value
(Ml/I) Reference
Marking, 1966
8.0 Marking, 1966
Jarvlnen, et al . 1977
Macek & McAllister,
1970
Macek A Sanders, 1970
Pr tester, 1965
Henderson, et al.
1959
Henderson, et al .
1959
Henderson, et al.
1959
48 Henderson, et al.
1959
Macek & McAllister,
1970
Marking. 1966
Marking, 1966
Marking, 1966
B-20
-------
TabU I. (Conttnuad)
Spaclas
Black bullhaad,
Ictalurus Ml as
Channal catfish,
Ictalurus punctatus
Channal catfish,
Ictalurus punctatus
ChaiMMl catfish,
Ictalurus punctatus
ChaniMl catfish,
Ictalurus punctatus
Guppy,
PoacMla ratlculata
Guppy.
Poacllla ratlculata
Brook stick! •back,
Cul««a Inoomtwis
Graan suftflsh,
LapoMis cyanvllus
Oraan sonflsh,
LafMMls cyanallus
Graan cut fish,
Lapoails cyanallus
Graan sunflsh,
Lapoal* cyanallus
Graan munflsh,
Lapoails cyanallus
Graan sunflsh,
Lapoails cyanallus
Mathod*
s, u
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
LC90/EC90
Cwo/n
20
16
17.4
17.5
17.5
19.5
56
67
2.8
3
3.9
6.7
6.4
4.4
Spaclac Mas*
Aoita Valita
(wa/|) ftafaranca
18 Marking. 1966
Hacak & McAlllstar,
1970
Mac* 4 Sandars, 1970
Marking, 1966
17 Marking, 1966
King. 1962
33 Handarson, at at.
1959
67 Marking, 1966
Marking, 1966
Marking, 1966
Marking, 1966
Marking, 1966
Marking. 1966
Marking, 1966
B-21
-------
TabU 1. CContlMMd)
LCM/EC90
Sp«ci«s Method* (no/11
Green sunfish,
LepoMls cyanel lus
Green sunflsh
L«pcssis cysfisiius
L ape* Is glbbosus
Lappals glbbosus
PiMpklnseed,
Lepo»ls glbbosus
PuMpk 1 nso«d ,
Leponls glbbosus
PtMpklnseed,
Lepo»ls glbbosus
Blueglll,
Lepomls macrochlrus
Bluegl 1 1,
Lepomls macrochlrus
Bluegl It,
Lepomls macrochlrus
Bluegl 1 1.
Lepomls macrochlrus
Blueglll,
Lepomls macrochlrus
Blueglll,
Lepomls macrochlrus
Blueglll,
Lepomls macrochlrus
5,
s.
s.
s,
s.
s,
s.
s,
s,
s,
s.
s.
s,
s.
U 3.6
U 5
U J.I
U 6.7
U 2.6
U 3.6
U 1.6
U 8
U 9.5
U 4.3
U 3.6
U 1.7
U J.2
U 3
Sp«clM MM*
Acut* Value
(wo/I) ftof«r«n
Marking
4.3 Marking
Marking
Marking
Marking
Marking
3.9 Marking
Macek &
1970
Macek &
Marking
Marking
Mark 1 ng
Mark 1 ng
Mark 1 ng
c*
, 1966
. 1966
, 1966
. 1966
. 1966
, 1966
, 1966
McAllister,
Sanders, 1970
, 1966
, 1966
, 1966
, 1966
. 1966
B-22
-------
Table I. (Continued)
Species
Blueglll,
Lepcntis wsacrochlrus
Bluerjtll.
Lepomls macrochlrus
Blueglll,
Lepomls macrochlrus
Blueglll,
Lepomls macrochlrus^
Blueglll,
Lepomls macrochlrus
Blueglll,
Lepomls macrochlrus
Longaar sunflsh,
Lepomls mega lot Is
Longaar sunflsh,
Lepomls mega lot Is
Redear sunflsh.
Lepomls ulcrolophus
Largemouth bass.
Ml crop tar us salMoldes
Largemouth bass,
Ml crop tar us salMoldes
Largeoouth bass.
Mlcropterut ta Isoldes
Yal low parch,
Parca flavascans
Yel low parch.
Perca flavascans
LC50/EC50
Method* (lig/l)
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
4.6
7
9.4
7
2.8
21
4.9
12.5
5
2
t.8
0.8
9
0.8
Species Mean
Acute Valas
(ug/D Reference
Marking, 1966
Marking, 19&6
Marking, 1966
Marking, 1966
Marking, 1966
4.9 Henderson, at at.
1959
Marking, 1966
7.6 Marking, 1966
5.0 Macek & McAllister,
1970
Macek 8. McAllister,
1970
MoceV, £ Sanders, 1970
1.4 Marking, 1966
Macek & McAllister,
1970
Marking, 1966
B-23
-------
Table I. (Continued)
Species
Yel low perch,
Perca flavescens
Yellow perch,
Perca flavescens
Freshwater druei,
Aplodlnotus grunnlens
Cladoceran,
Daphnla pulex
Cladoceran,
Sle-jcephalus serrulatus
Clacoderan,
Sleocephalus serrulatus
Sowbug,
Asa II us brevlcaudus
Scud,
GeMMrus fasclatus
Scud,
Gamer us fasclatus
Scud,
GeM-erus lecustrls
Glass shrl«p,
Palaenonetes kadlakensls
Stonefly.
Pteronarcys callforntca
Planar Ian,
Method'
^•••••••••••K
S, U
S, U
s, u
s. u
s, u
s, u
s, u
s, u
s, u
s, u
s, u
s, u
s. u
Sptcftt MBMI
LC9Q/EC90 Acute Velue
fHO/1) (MB/D
0.6
1.9 1.6
10 10
TP€
3.2 3.2
4.5
5.2 4.8
10 10
0.6
0.86 0.72
0.64 0.64
0.68 0.68
380 380
740 740
Reference
Marking, 1966
Marking, 1966
Marking, 1966
Sanders J, Cope,
Sanders i Cope,
Sanders i Cope,
Sanders, 1972
1966
1966
1966
Sanders, 1972
Senders, 1972
Sanders, 1969
Sanders, 1972
Sanders & Cope, 1968
KouyouMJfan i Ugfow,
Polycells feline
1974
B-24
-------
Table 1. (Continued)
Species
Planar Ian,
Polycells fellna
Eastern oyster,
Crassostrea virgin lea
Eastern oyster,
Crassostrea virgin lea
Brown shrimp,
Penaeus aztecus
Grass shrimp,
Palaemonetes vulgar Is
Sand shrimp,
Crangon septemsp 1 nosa
Korean shrimp,
Palaemon macrodacty 1 us
Korean shrimp,
Palaemon macrodacty 1 us
Hermit crab,
Pegurus long! car pus
American eel,
Angullla rostrata
Chinook salmon,
Oncorhynchus tshawytscha
Mummlchog,
Fundu 1 us hater oc 1 1 tus
Method"
S, U
FT, U
FT, U
FT, M
S. U
S, U
S, U
FT. U
S, U
S, U
FT, U
S, U
Species Mean
LC50/EC50 Acute Value
(UO/I) (ug/l)
DDE
1,050 1
SALTWATER SPECIES
DOT
7.0
9.0
0.14
2.0
0.6
0.86
0.17
6.0
4.0
0.68
3.0
,050
-
7.9
0.14
2.0
0.6
0.38
6.0
4.0
0.68
Reference
Kouyoumjlan & Uglow,
1974
Lowe, undated
Lowe, undated
Schlmmel & Patrick,
1975
Elsler, 1969
El si or, 1969
Schoettger, 1970
Schoettger, 1970
Elsler, 1969
Elsler, 1970b
Schoettger, 1970
Elsler, 1970a
B-25
-------
Table !. (Continued)
MtMMlchog,
Fundulus heteroclltus
Strips) Mlllflsh,
Fundulus Mjalls
Atlantic sllverslde,
MMldla Mnldla
Strlp*l bass.
Moron* saxatllls
Shlnar parch,
Cyaatogaatar aggragata
Shiner parch,
CynatooMter aggragata
Dwarf parch,
Mlcroaatrus alnlaus
Dwarf parch,
Mlcronetrus alnlwus
Bluahaad,
ThalasscM blfasclatum
Striped Mullat,
Mug 1 1 cepha 1 us
Striped MJ Hat,
Mug II capha 1 us
Northern puffer,
Sphaeroldes oaculatus
Eastern oyster,
Crassostrea vlrnlnlca
Korean shrimp,
Palaemon macrodacty 1 us
Method"
s. u
s. u
s, u
FT, U
S, U
FT» U
S. U
FT. 0
S, U
S. U
s, u
S. U
FT, U
S, U
Spaclaa Mean
LCM/EC90 Acute Value
(1*0/1) (WO/I)
5.0 3.9
1.0 1.0
0.4 0.4
0.53 0.53
7.6
0.45 1.8
4.6
0.26 1.1
7.0 7.0
0.9
3.0 1.6
89 89
TDE
25 25
8.3
Reference
Elsler, 1970b
Elsler, 1970b
Elsler, 1970b
Korn & Earnest, 1974
Earnest & Benvllle.
1971
Earnest & Benvllle,
1971
Earnest 4 Benvll le,
1971
Earnest & Benvll le,
1971
Elsler, 1970b
Elsler, 19706
Elsler. 1970b
Elsler, I970b
Lowe, undated
Scho«ttg«r, 1970
B-26
-------
Table t. (Continued)
Specie*
Korean shrlnp.
Palaoenn iMcrodactylus
Striped bass.
Morone saxatllls
Eastern oyster,
Crassostree vlrgtnlca
LC50/EC30
Method" (wo/1)
FT. U 1.6
FT, U 2.5
DOE
FT, U 14
Species Mawt
Acute Value
3.6
2.5
14
Reference
Schoettger, 1970
Korn & Earnest, 1974
Lowe, undated
*S » static; FT * t low-through; M » measured; U * unmeasurod
B-27
-------
Table 2. Chronic values for DOT (Jarvlnen, et al. 1977)
Limits Chronic Value
Test* (ug/l) (ug/l)
FRESHWATER SPECIES
Fathead minnow, 1C 0.37-1.48 0.74
Plmephales promelas
• LC » life cycle or partial life cycle
Acute-Chronic Ratio
Acute Chronic
Value Value
Species (M9/I) (ug/l) Ratio
Fathead minnow 48 0.74 65
Plmephales promelas
B-28
-------
Tabu 3. Spaclas MM acuta valuas a*d •cuta-chroftlc ratios for'DOT and a»tabolltas
Mk*
42
41
40
39
38
37
36
35
34
33
32
31
30
FRESHWATER
DDT
Planar Ian,
Polycalls fallna
Stonafly,
Acronaurla pact flea
Stonafly.
PtaroMrcys callfornlca
Caddlsfly,
Arctopsycha grand Is
Ncrtnarn radbafly daca,
ChrosoMus aos
Brook stick (aback,
Culaaa Inconstans
Saad shrlap,
Cyprldoasls vldua
Caddlslty,
Hydroasycha California
Fathaad minnow,
Plaaohalas pronalas
Goldfish,
Carasslus auratus
Poacl'lla ratlculata
Mayfly,
EphaiaaraHa grandls
Black bullhaad.
d^VCI W ^^alBM
Acuta Valua
(IM/I>
SPECIES
1,230
362
192
175
69
67
54
46
48
40
33
25
18
Saaclas Maaa
Acuta-Chroalc
Ratio
-
-
-
-
-
65
-
-
Ictalurus «alas
B-29
-------
Table 3. (Continued)
Rank*
29
28
27
26
25
24
23
22
21
20
19
18
17
16
Specie* MMN
Acute Value
Specie* (wa/l)
Channel catfish,
Ictalurus punctutus
Coho salmon,
Oncer hyncnus Msutch
Chinook salmon,
Oncorhynchus tshawytscha
Fr«sh«at«r druM,
Aplodlnotus grunnlans
Lak« trout,
Salvallnua naanycush
Brook trout,
Salv«Hnua footlnaHs
Carp,
Cyprlnus carplo
Rainbow trout.
Sal no gairdncrl
Long«ar sunflsh,
Lapouls •agalotts
Brown trout.
Sal MO trutta
Rwtoar sunflsh,
L«pools olcrotophus
Blueglll,
LapoMls awcrochirus
Green sunflsh,
Lepoails cyon^Hus
Sowbug,
17
14
12
10
9.3
8.9
8.0
7.8
7.8
7.3
5.0
4.9
4.3
4.0
Specie* Mean
Acute-Chronic
Ratio
-
-
-
-
-
-
-
-
As*11 us brevlcaudus
B-30
-------
Table 3. (Continued)
Rank*
15
14
13
12
11
10
9
8
7
6
5
4
3
2
Pumpk 1 nseed ,
lepoails qlbbosus
Stonetly,
Claassenla sabulosa
Glass shrimp,
PalaeMonetes kadiakensls
Crayfish,
Procamfaarus acutus
Scud,
GaMMrus lacustrls
Cladoceran,
SlMocephalus serru lotus
Cladoceran,
Daphnla «aqna
Crayfish,
Orconectes nals
Stonefly,
Pteronarcel 1 a badla
Northern pike,
Esox luclus
Scud,
GamMrus fasclatus
Yatlow perch,
Parca flavascens
Largemouth bass,
Mlcropterus sa Into Ides
Cutthroat trout,
Sal»o clarki
Species Mean
Acute Value
3.9
3.5
3.2
3.0
3.0
2.6
2.4
1.9
1.9
1.7
1.7
1.6
1.4
1.1
Species Mean
Acute-Chronic
Ratio
-
-
-
-
-
-
B-31
-------
T«bl« 5. (ContltMMMf)
DOT
17 NorttMrn puffer, 89
Rank*
1
8
7
6
5
4
3
2
1
1
Acut* ValiM
Cladocaran, 0.36
Daphnla j>ulax
TDE
Planarlan, 740
PolycalU fallna
STCXM* »y» 3ou
Ptwonarcyc calif arnica
SoMbug, «0
A»«l III* tr«vfCMldu»
Cladocaran, 4.6
SlK>c«ph»luc sarrulatws
Cl«doe«ran( 3.2
Oapanla £ul*x
Scud, 0.72
Glass ihrlMp, 0.68
Scud. 0.64
GaiMwrus lacmtrls
ODE
Planarfan. 1,090
Polyc«ll> fallfi*
SALTWATER SPECIES
Acuta-Cliroftic
Ratio
-
B-32
-------
T«h(* 3, (Continued)
ftanfc*
16
15
U
15
12
It
10
9
8
7
6
5
4
3
Soecles Mean Soecles Mean
Acute Value Acute-Chronic
Soeeles (wo/I) Ratio
Eastern oyster,
Crassoatrea vlrglnlca
Bluehead,
Thalassoma blfasclatiM
Hemtt crab,
Pag.tr us 8 ongi carpus
American eei,
Anguil la rostrata
ftjMlchog,
Fundulus heteroclltui
Grass shrl«p,
Palaavonetes puqlo
Shiner perch,
Cyaatoqastar aggregate
Striped mil let,
Mug II cephalus
Dwarf perch,
MIcroMetrus wlnlpus
Striped MlllfUh,
Fundulus wijal Is
Chinook salmon,
Oncorhynchus tshawytscha
Sand shrlap,
Crangort septemsplnosa
Striped bass,
Morone saxatl I Is
Atlantic sllverslde.
7.9
7.0
6.0
4.0
3.9
2.0
1.8
1.6
1.1
t.O
0.68
0.6
0.53
0.4
-
-
-
-
-
Menldla men Id la
B-33
-------
TabU 3. (Continued)
Species NMA Species Mean
Acute Value Acute-Orowlc
Rank* Species Cua/t) tUtlo
2 Korean shrlap. 0.38
Palaa»on aacrodactylus
I Brown sir l*p, 0.14
Panaeus attacu*
TOE
3 Eastarn aystar, 29
Crasaoatraa vlrqlnlca
2 Koraan shrlap. 3.6
Palaaaon Kacrodactylut
1 Strlpad bast, 2.5
Morooa «axatiI>»
OK
1 Eastarn oystar, 14
Orasaoatraa virginlea
•Ranked tram least sensitive to Most sensitive based on special «aan
acute value.
Freshwater Final Acute Value for DOT - 1.1 wg/1
Saltwater Final Acute Value for DOT • 0.13 ng/1
B-34
-------
TabU 4. Plant valtiM for DOT
Species
Alga,
Anacystls nldulans
Alga,
Chloral la sp.
Alga,
ScenedesMis quadrtcauda
Alga,
Selenastrum capr 1 cor nu turn
Diatom,
Skeletonema costatum
Coccol Ithophore,
Coccollthus huxleyl
Green alga,
Pyramlwonas so.
Nerltlc dlnoflagellate,
Perldlnlu* trocholdeuM
Effect
FRESHWATER SPECIES
Growth
Growth and
morphology
Growtti
Photosynthesis
SALTWATER SPECIES
Reduced photo-
synthesis (1-day)
Reduced photo-
synthesis (1-day)
Reduced photo-
synthesis (1-day)
Reduced photo-
synthesis (1-day)
Result
(UQ/I)
800
0.3
100
3.6
10
10
10
to
Reference
Batterton, et al,
1972
Sodergren, 1968
Stadnyk, et al.
1971
Lee, et al. 1976
Wurster, 1968
Wurster, 1968
Wurster, 1968
Wurster, 1968
B-35
-------
Tabl* 5. Residues for DOT ami metabolites
Species
Coontall.
Ceratophyl tum demersum
Cladophora,
Cladophora sp.
Duckweed,
Lemon minor
Water milfoil,
Hyr 1 qphy 1 1 urn sp.
Curly leaf pond weed,
Potamogeton crlpus
Narrow- leaf pondweed,
Potamogeton fol losus
Sago pondweed,
Potamogeton pectlnatus
Soft stem bulrush,
Sclrpus val Idus
Bur reed,
Spar qnnl urn eurycarpum
Bladder wort,
Utrl cut aria vulgar Is
Mussel,
Anodonta grand Is
Clams (five species
compos 1 te) ,
Latiysllls slllquoldea
Lamps 1 1 Ts veotr 1 cosa
Lasmlqonn costata
Fusconala flava
Llpld Bloconcsntratlon
Tissue (|) Factor
FRESHWATER SPECIES
DOT
1 ,9»
21,580
1,210
1,870
14,280
781
6,360
495
623
2,200
Whole body 1.0 2,400
Whole body 1.0 12,500
Duration
<
30
30
30
30
30
30
30
30
30
30
21
56
Reference
Cberhardt, et at.
1971
Eberhardt, et at.
1971
Eberhardt, et al.
1971
Eberhardt, et al.
1971
Eberhardt, et al.
1971
Eberhardt, et al.
1971
Eberhardt, et al.
1971
Eberhardt, et al.
1971
Eberhardt, et al.
1971
Eberhardt, et al.
1971
Bedford 4 Zablk, 1973
Jarvlnen, et al . 1977
Llqomla recfa
B-36
-------
Table 5. (Continued)
Species
C 1 adoceran ,
Daphnla magna
Freshwater prawn,
Pataemonetes paludosus
Crayfish,
Orconectes punctata
Crayfish,
Procambarus allenl
Mayfly (nymph).
Ephemera dan lea
Dragonfly (nymph),
Tetragon eur la sp.
Bloodworm,
Teodlpes sp.
Red leech,
Erpobdella punctata
Mewlfe.
Alosa pseudoharengus
Lake herring,
Corecjnus artedl
Lake uhlteflsh,
Coregonus clupeaformls
Bloater,
Coregonus hoyl
Klyl,
Coregonus klyl
Cisco,
Coregonus sp.
Tissue
Whole body
Whole body
Whole body
Whole body
Whole body
-
-
Whole body
Whole body
Whole body
Whole body
Whole body
Muscle
Lip Id Bloconcantratlon
(%} Factor
9,923"
7,000
5,060
1,947
4,075
2,700
4,750
7,520
10.0 1,296.666
5.3 2,236,666
7.6 260,000
20.0 2,870,000
4,426,666
6.4 368,777
Duration
(days)
14
Field
30
Field
5
20
30
30
Field
Field
Field
Field
Field
Field
Reference
Prlester, 1965
Kollplnskl, et al.
1971
Eberhardt, et al.
1971
Kollplnskl, et al.
1971
Sodergren & Swenson,
1973
Wllkes & Weiss, 1971
Eberhardt, et al .
1971
Eberhardt, et al .
J971
Relnert, 1970
Relnert, 1970
Relnert, 1970
Re Inert, 1970
Re Inert, 1970
Miles & Harris, 1973
B-37
-------
Tabl* 9. (Continued)
Species
Coho salmon,
Oncorhynchus klsutch
Rainbow trout.
Sol mo gairdnerl
Rainbow trout,
Sal no gairdnerl
Brown trout.
Sal mo trutta
Lake trout,
Safveffnus namaycush
Lake trout,
Salvellmis namaycush
Lake trout,
Salvellnus namaycush
American smelt,
Osmerus mordax
Carp,
Cyprinus carp I o
Common shlnor (composite),
Notropls cornutus
Northern redbelly dace,
Chrosomus eos
Fathead minnow,
Plmephales promelas
Wh I te sucker ,
Catostomus commersonl
White sucker,
Catostomus commersonl
Trout-perch,
Per cops Is omlscomaycus
Tissue
Whole body
Muscle
Whole body
Muscle
Muscle
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Muscle
Whole body
Whole body
Llpltt Bloconcentratlon
(?) Factor
1, 563, 571
1.0 11,607
6.6 38,642
1.8 45,357
4.6 458,259
11.0 1,168,333
47,428
3.9 770,000
6.2 640,000
363,000
3. 1 99,000
2.8 1 JO, 000
2.8 96,666
313,333
Duration
(days)
Field
Field
84
Field
Field
Field
152
Field
Field
40
266
Field
Field
Field
Reference
Lake Michigan Inter-
state Pestle. Conm.,
1972
Mites & Harris, 1973
Reiner t. et al. 1974
Miles & Harris, 1973
Miles & Harris, 1973
Re Inert, 1970
Re Inert & Stone, 1974
Relnert, 1970
Re inert, 1970
Hanellnk, et al. 1971
Jarvlnen, et al . 1977
MHes 4 Harris, 1973
Relnert, 1970
Relnert, 1970
B-38
-------
TabU 5. (ContimMd)
Species
Flagflsh,
Jordanel la f lorldae
Mosqu 1 tof 1 sh,
Ganbusla nfflnls
Rock bass,
Ajnbloplltes rupestrls
Green sun fish,
Lepomls cy one II us
Green sun fish (composite),
I apoai|s cy ana II us
f*ufflpklns~eed,
Lepomls glbbosus
BluegHI,
Lepomls mocrochlrus
Largemouth bass
(young of year),
Mlcropterus sal mo Ides
Yellow perch,
Perca ftavescens
Slimy sculp In,
Cottus cognatus
Zoop lank ton (mixed),
Paphnla sp.
Keratel la sp.
Rainbow trout,
Sal mo qalrdnerl
Bluegll 1,
Lepomls macrochlrus
Tl»su«
Whole body
Whole body
Muscle
Whole body
Whole body
Muscle
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Lip lit Bioconcentratlon
{%) Factor
14,526
21,411
4.0 17,500
17,500
59,210
4.0 16,071
317,000
7.9 1,073,333
763,333
DOE
63,500
181,000
110,000
Duration
(days)
Field
Field
Field
15
80
Field
40
Field
Field
21
108
60
Reference
Kolipinskl, et al.
1971
Kolipinskl, et al.
1971
Miles & Harris, 1973
Sanborn, et al . 1975
Hamellnk, et al. 1971
Miles & Harris, 1973
Hamellnk, et al. 1971
Relnert, 1970
Relnert, 1970
Hamellnk & Way brant,
1976
Hamellnk & Waybrant,
1976
Hamellnk & Waybrant,
1976
B-39
-------
Table 3. (Continued)
Species
Eastern oyster,
Crassostrea vlrglnfca
Eastern oyster,
Crassostrea virgin lea
Pink shrimp,
Penaeus duorartM
Market crab.
Cancer maglster
Market crab,
Cancer maglster
Atlantic croaker,
Mlcropogon undulatus
Shiner perch,
Cyma togas tar aggregata
Shiner perch,
Cymatogaster aggregate
Dwarf parch,
MIcroMetrus •Inlmus
Dwarf perch,
MIcroMtrus Minimus
White perch,
Phanerodon fureatus
White perch,
Phanerodon fureatus
Pile perch.
Raccoon II us vacca
Pile perch,
Raccochllus vacca
Tl«»ue
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Lip Id Bloconcentretlon
(JO Factor
SALTWATER SPECIES**
DDT
42,400
76,300
1,200
1.3 14,250
1.3 4,750
16,000
3.4 43,250
3.4 34,750
6.4 46,500
6.4 37,000
2.8 22,250
2.8 29,250
4.4 26,750
4.4 32,500
Duration
(day*)
252
168
56
Field
Field
21-35
Field
Field
Field
Field
Field
Field
Field
Field
Reference
Low, et al. 1970
Lowe, et al . 1970
Nlmro, et al. 1970
Earnest & Benvl 1 le,
J97I
Earnest & Benvl 1 le,
1971
Hans en & Wilson, 1970
Earnest & Benvl 1 le,
1971
Earnest & Benvl 1 le,
1971
Earnest & Benvll le,
1971
Earnest & Benvl 1 le,
1971
Earnest & Benvll le,
1971
Earnest & Benvll le,
1971
Earnest & Benvl 1 le,
1971
Earnest & Benvl 1 le,
1971
B-40
-------
Table 5. (Continued)
Species
Staghcrn sculpln,
Leptocotlus armatus
Staghorn sculpln,
Leptocotlus armatus
Speckled sanddab,
Clthar Ichthys stlgmaeus
Speckled sanddab,
Clthar Ichthys stlgreaeus
English sole,
Parophrys vetulus
English sole,
Parophrys vetulus
Starry flounder,
PI at Ichthys stellatus
Starry flounder,
PI at Ichthys stellatus
•Value converted from dry
"Data Include metabolites
Action
Tissue
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Whole body
Up Id Bloconcentratlon Duration
(*) Factor (days)
1.9
1.9
2.7
2.7
2.0
2.0
2.5
2.5
17,000
22,250
15,250
12,250
20,000
13,000
24,750
23,750
Field
Field
Field
Field
Field
Field
Field
Field
Refer em
Earnest
1971
Earnest
1971
Earnest
1971
Earnest
t971
Earnest
1971
Earnest
1971
Earnest
1971
Earnest
1971
:e
& Benvllle,
& Benvllle,
& Benvllle,
A Benvllta,
& Benvll le,
& Benvll le,
4 Benvllle,
& Benvllle,
weight to wet weight basis.
when given.
Maximum Permissible Tissue Concentration
Level or Effect
Concentration
(mo/kg) Reference
Fish 5.0
Reduced productivity. Brown 0.15
pelican, Pelecanus
occIdental Is
U.S. FDA Guideline
7420.08, 1978
Anderson, et al. 1975
B-41
-------
Table 9. COnrtlMMd)
Maxl»u» Permissible Tissue Concentration
Action level or Effect
Eggshell thinning. Brown
pelican, Pelecenu*
occldentaTTr
Inhibition of Na*-K* AlPese,
Rainbow trout, Salao
galrdnerl
R«duc«d duckling survival,
Black duck. Anam
Roducad survival, -Sparrow
hawk, Fa I co «parv«rlu«
Eggshell thinning. Screech
owl, Otus a»lo
Eggshell thinning. Mallard.
Anas platyrnynchos
Eggshell thinning. Mallard,
Anas platyrhynchos
Eggshell thinning. Black
duck. Anas rubrlpes
Eggshell thinning. Sparrow
hawk. Fa I co sparverlus
Reduced sac fry survival,
Cutthroat trout. Sal mo
ctorkl
Reduced fry survival,
Brown trout, Sal mo
trutta
Concentration
(«a/kg)
0.9
2.75
2.0
2.8
2.8
3.0*
3.0*
3.0*
3.0
3
3.4
Reference
Blue, et •!. 1972, 1974
Ce^bell. et al. 1974
Longcore & Stendell, 1977
Porter & Hleneyer, 1972
McLane & Hall, 1972
Haseltlne, et al. 1974
Heath, et al. 1969
Longcore, et al. 1971
Llncer, 1975
AM (son, et al. 1963
Burdlck, et al. 1972
B-42
-------
Table 5. (Continued)
Maximum Permissible Tissue Concentration
Action Level or Effect
Reduced fry survival.
Lake trout, Salvellnus
namaycush
Reduced survival, Coho
salmon (f Ingerl Ing),
Oncorhynchus klsutch
Reduced survival, Chinook
salmon (finger! Ing),
Oncer hynchus tshawytscha
Reduced light Intensity
discrimination. Rainbow
trout. Sal mo galrdnerl
Reduced phenoxyethanol
anesthetic Induction and
recovery tines. Rainbow
trout, Salmo galrdnerl
Concentration
(mg/kq) Reference
6 Burdlck, et al. 1972
6.25 Buhler, et al. 1969
6.25 Buhler, et al. 1969
9 McNfcholl 4 Mackay, 1975
11.36 Klaverkamp, et al . 1976
* Value converted fro* dry weight to wet weight basis
Geometric mman of normalized BCF values (see text) - 17,870
tor human consuwptlon: FOA action level tor fish* 9.0 ugAg
Percent llpld value for freshwater species (see Guidelines) » 15
Percent llpld value for saltwater species (see Guidelines) - 16
Freshwater: ___ 9.0 - 0.000019 «g/kg « 0.019 ug/l
17,870 x 15
Saltwater:
5.0
x 16
0.000017 ntg/Xg - 0.017 ug/l
B-43
-------
Table 3. (Continued)
Using highest 8CF for edible portion of a consumed species
Freshwater: Lake trout « 458,259 (Miles and Harris, 1973)
5.0 = 0.000011 «g/kg - 0.011 ug/l
458,259
Wildlife Protection: Lowest maximum permissible tissue concentration « 0.15 mg/kg (Anderson, et at. 1975)
Percent llpld value for northern anchovy - 8 (Relntjes, 1980)
Freshwater and Saltwater: 0.15 - 0.0000010 wg/kg - 0.0010 ug/l
17,870 x 8
Freshwater Final Residue Value * 0.0010 ug/l
Saltwater Final Residue Value = 0.0010 ug/l
8-44
-------
Tab la 6. Other data for DOT and •wtabolltos
5p»cl«»
Alga,
Chlorelta pyreooldosa
Duration EfUct
FRESHWATER SPECIES
DOT
Result
(ug/D
7 days No growth effect 100,000
Stonefly (naiad),
Acroneurla pad flea
Stonefly (naiad).
Pteronarcys cnl 1 forn I ca
Damsel fly,
Ischnora vertIcalls
Reference
Christie, 1969
Cladoceran,
Oaphnla maqrm
Ciadoceran,
Daphnla maqna
Cladoceran,
Daphnla maqna
Sow bog,
Asa II us brtvlcaudus
26 hrs
i4 days
14 days
46 hrs
LC50
LC50
50* Inhibition
of total young
produced
LC50
Scud, 48 hrs LC50
Gamnarus fasclatus
Scud, 120 hrs LC50
Gamnarus fasclatus
Glass shrimp, 36 hrs LC50
Patnemonates kadlak«nsls
Glass shrimp, 120 hrs LC50
Pa Iaemonetes Kadlakensls
30 days LC50
30 days LC50
48 hrs LC50
Planartan. 24 hrs
PolyCOI Is falIna
AsexuaI fIssI on
Inhibition
5.5 Crosby, at al. 1966
0.67 Maki & Johnson, 1975
0.50 Makl 4 Johnson, 1975
4.7 Macek & Sanders, 1970
3.6 Hacek & Sanders, 1970
0.6 Sanders. 1972
4.5 Ferguson, et al.
I965b
1.3 Sanders, 1972
72 Jensen & Gaufln, 1964
265 Jensen I Gaufln, 1964
22.5 Macek & Sanders, 1970
250 Kouyoumjlan & Uglo»,
1974
B-45
-------
TobU 6. (Continual)
RMUlt
Specie*
Mlcronetazoan,
Lepldoderwel la squanmata
Coho salmon,
Oncorhynchus klsutch
Coho salmon (juvenile),
Oncorhynchus klsutch
Coho salmon,
Oncorhynchus klsutch
Cutthroat trout,
Salmo ctarkl
Rainbow trout,
Salmo golrdner)
Rainbow trout,
Salmo qalrdnerl
Rainbow trout,
Salmo galrdnerl
Atlantic salmon
(gastrulae),
Salmo salar
Atlantic salmon,
Salmo salar
Atlantic salmon,
Salmo salar
Atlantic salmon.
Sal MO salar
Brown trout (al«vln),
Salmo trutta
Duration
96 hrs
7 days
125 days
24 hrs
5 hrs
30 days
24 hrs
24 hrs
24 hrs
48 hrs
Effect
Reproductive
lethality (25* OOT7
Reduced fry
survival
Increased cough
frequency
Estimated median
survival time -
106 days
Reduced sac fry
survival
Uncontrol led
reflex reaction
Cough response
threshold
Reduced sac fry
survival
Retarded behav-
ioral development
and Impaired
balance of alevlns
Altered temperature
selection
Altered temperature
selection tor 1 mo
Altered temperature
selection
LC50
3,000
1.09 Big/kg
In eggs
1.27 mg/kg
In food
>0.4 mg/kg
In «ggs
100
52-140
>0.4 mg/kg
In eggs
50
5
50
10
2.5
Reference
Huwion, 1974
Johnson & Pecor, 1969
Schaunfcurg, 1967
Buhler 4 Shanks, 1972
Cuerrler, et al. 1967
Peters & Weber , 1977
Lunn, et al . 1976
Cuerrler, et al. 1967
OKI & Sounders, 1974
Ogllvle 4 Anderson,
1965
Ogllvle A MM lor,
J976
Peterson, 1973
Alabaster, 1969
B-46
-------
Table 6. (Continued)
Species
Duration
Effect
Result
Reference
Brook trout,
Salvellnus fontlnalls
Brook trout,
Salvellnus fontlnalls
Brook trout,
Salvellnus fontlnalls
Brook -h-out,
Salvellnus fontlnalls
Brook trout,
Salvellnus fontlnalls
Brook trout,
Salvellnus fontlnalls
Brook trout,
Salvellnus fontlnalls
Brook trout,
Sa 1 ve 1 1 nus font 1 na 1 1 s
Lake trout (fry),
Salvellnus namaycush
Goldfish,
Carasslus auratus
Goldfish^
Carasslus auratus
Goldfish,
Carasslus auratus
Golden shiner,
Notemlgonus crysoleycas
24 hrs LC50
24 hrs Lateral line nerve
hyper sensitivity
24 hrs Visual conditioned
avoidance Inhibi-
tion
Reduced sac fry
survival
24 hrs Altered tempera-
ture selection
156 days Slight reduction
In sac fry
survival
24 hrs Altered tempera-
ture selection
24 hrs Altered tempera-
ture selection
Reduced survival
2.5 hrs Loss of balance
and decreased
spontaneous elec-
trical activity of
the cerebellum
4 days Exp 1 oratory behav-
lor Inhibition
7 days Schooling
Inhibition
24 hrs Schooling
Inhibition
54
100
20
X3.4 mg/kg
In eggs
20
2 mg/kg
In food
10
100
2.9 mg/kg
In fry
1,000
10
1
15
Mil ler i Ogl Ivle,
1975
Anderson, 1968
Anderson & Peterson,
1969
Cuerrler, at al . 1967
Gardner, 1973
Macek, 1968
Miller & Ogllvle,
1975
Peterson, 1973
Bur dick, et al. 1964
Aubln & Johansen,
1969
Oavy & Kleerekoper,
1973
Vtels & Wels, 1974
Bailey, 1973
B-47
-------
Table 6. (Continued)
pecles
Duration
Effect
Re*ult
(ya/D Reference
Golden shiner,
Notemlgonus crysoleucas
Fathead minnow,
PlMphaies prow*) as
Fathead winnow,
Pfuwphajes prowelos
Fathead minnow,
Plmepholes pronelas
Channel catfish
(flngerllng),
Ictaturus punctatus
Black bullhead,
Ictalurus melas
Mosqultof Ish,
Gae)busta afflnts
Mosqultof Ish,
Gambusla afflnls
Mosqultof Ish,
GaMbusIa afflnls
Green sunflsh,
Lepouls cyanellus
Blueglll,
Lepomls Macrochlrus
Blueglt 1,
Lepomls macrochlrus
Toad,
Bufo woodhousel fowlerl
36 nrs
48 hrs
48 hrs
266 days
96 hrs
36 hrs
46 hrs
36 hrs
40 Mln
36 hrs
36 hrs
16 days
36 hrs
LC50
LC50 (static)
LC50 (flow-
through)
Mg2+ ATPase
Inhibition
LC30 <50f DOT
dust)
LC50
LC50
LC50
Succlnlc dehydro-
genase activity
Inhibition
LC50
LC50
Hyperactive loco-
Motcr response
LC50
29.9
7.4
>40
0.5
>2,000
16.4
43
21.3
9 x IO"9
molar
23.5
28.7
0.008
560.0OO
Ferguson, et a). 1964
Linear, et al. 1970
Llncer, et al . 1970
Oesalah, et al. 1975
Clemens A Sneed, 1959
Ferguson, et al .
1965a
DzluK & Plapp, 1973
Ferguson, et al.
1965a
Moffett I
Yar trough. 1972
Ferguson, et al . 1964
Ferguson, et al . 1964
EUgnord, et al. 1977
Ferguson & Gilbert,
1967
B-48
-------
Table 6. (Continued)
Specie*
Toad (tadpole,
4-5-wK-old).
Bufo woodhousel fowler I
Toad (tadpole,
6-wk-old),
Bufo woodhousel fowler!
Toad (tadpole,
7-wk-old),
Bufo woodhouset fowler I
Prog,
ACT Is crop I tans
Frog (tadpole),
Pseudacrls trlserlata
Frog (tadpole),
Rana cIamitans
Frog,
Rana temper aria
Turtle,
Chrysemys pieta
Channel catfish
(finger I Ing),
Ictalurus punctatus
Toad (tadpole,
4-5-wk-old),
Bufo woodhousel fowler I
Frog (tadpole),
Pseudacrls trlserlata
Duration Effect
96 hrs LC50
96 hrs LC50
% hrs LC50
36 hrs
96 hrs
6 days
20 days
30 m\n
96 hrs
LC50
LC50
Result
(Ma/1) Reference
1,000 Sanders, 1970
100 Sanders, 1970
30 Sanders, 1970
620,000 Ferguson A Gilbert,
196?
600 Sanders, 1970
Increased pituitary 100 Peaslee, 1970
meIanocyte-stI mu I a-
tlng hormone levels
LC30
7.6 mg/kg Harrl, et al. 1979
In food
ATPase Inhibition 0.53 uH Phillips A Wells,
1974
TDE
LC50 (50$ TDE) <2,600 Clemens & Sneed, 1959
96 hrs LC50
96 hrs LC50
140 Sanders, 1970
400 Sanders, 1970
B-49
-------
Takla 6.
Easter* oyster,
Craasostrea vlrfllnlca
a»la and feaale
Penaeus duoraru*
Pink shrlMp,
Penaeus duoraruai
Pink shrlap,
Panaaus duoraru*
Pink shrlep.
Panaeus duoraru*
White sir Imp,
PanMus satlfarus
Ufact
SPCCICS
lit
(•fl/l)
DOT
12
12
12
BlooMCMtratlen
factor * 20.000*
BlocoMcantratloii
factor • 14,000*
BlocoHcantratlea
factor - 10.000*
•39*
apara)
Eastarn aystar,
CraMostraa vlrqlnlca
Eastarn oyster,
Qrassoatraa virgin lea
Pink slrlap,
12 days
t2 days
7 days
592 day*
30 days
B locoacantrat Ion
factor - 29,000*
B locoacantrat Ion
factor <• 9,000*
Af fact ad shall
deposition
B loooncantrat Ion
factor - 37,000**
Affactad cation
concentrations In
tMpatopancraas
tlssua
28 days LCI00
2 days EC5O
13 days Bloconcantratlon
factor - 1,500*
I day EC50
Butlar. I9tt
Butlar. 19*6
Butlar. l«66
Butlar. 19*6
Butlar. 1966
10.0 Butlar, 1966
Parrlsh, 1974
NlnMO ft Blackaan,
1972
0.12 NlMBD, at al. 1970
0.6 Lowa, undatad
, et al. 1970
0.7 Lowa, undated
B-50
-------
Table 6. (Continued)
Result
Species
Brown shrimp,
Penaeus aztecus
Grass shrimp,
Palaenonetes juigio
Blue crab,
Calllnectes sapJdus
Blue crab,
Ca 1 1 1 nectes sap 1 dus
Sheepshead minnow,
Cyprlnodon varlegntus
Sheepshead ml nnow,
Cyprlnodon varleqatus
Mummlchog,
Fundu 1 us heter oc ) 1 tus
Longnose kill If Ish,
Fundu 1 us slml Ms
Mosqultof Ish,
Gambusla afflnls
Plnf Ish,
Laqodon rhomboldes
Plnf Ish,
Lagodon rhomboldes
Plnf Ish,
Lagodon rhomboldes
Spot,
Lelostomus xanthurus
Striped mul let',
Hug II cephalus
Duration
2 days
2 days
2 days
36 wks
2 days
2 days
10 days
2 days
1 day
2 days
14 days
14 days
2 days
2 days
Effect
EC50
EC50
EC50
Mortality
LC50
LC50
LC50
LC50
Affected salinity
preference
LC50
B 1 oconcentr a 1 1 on
factor « 40,000*
8 1 oconcentr at 1 on
factor - M,000»
LC50
LC50
1.0
o.a
10
0.5
3.2
2.0
2.7
5.5
5.0-20
0.32
1.8
0.4
Reference
Lowe, undated
Lowe, undated
Lowe, undated
Lowe, 1965
Lowe, undated
Lowe, undated
Eisler, 1970a
Lowe, undated
Hansen, 1972
Lowe, undated
Hansen & Wilson, 1970
Hansen & Wilson, 1970
Lowe, undated
Lowe, undated
3-51
-------
Table 6. (Continued)
Striped mul let,
Mug 11 cephalus
Striped mullet,
Mug 11 cephalus
Duration Effect
2 days LC50
2 days
LC50
TDE
* DOT and metabolites
"•Results based on unmeasured water concentrations.
Result
(ug/l) Reference
0.55 Lowe, undated
0.4 Lowe, undated
Pink shrimp,
Penaeus duorarum
Longnose Mil (fish,
Fundulus slml 1 Is
Brown shrimp,
Penaeus aztecus
Spot.
Lelostomus xanthurus
2 days
2 days
2 days
2 days
EC50
LC50
DDE
EC50
LC50
2.4
42
28
20
Lowe,
Lowe,
Lowe,
Lowe,
undated
undated
undated
undated
B-52
-------
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Macek, K.J. and W.A. McAllister. 1970. Insecticide susceptibility of some
common fish family representatives. Trans. Am. Fish. Soc. 99: 89.
Macek, K.J. and H.O. Sanders. 1970. Biological variation in the suscepti-
bility of fish and aauatic invertebrates to DDT. Trans. Am. Fish. Soc.
99: 89.
Maki, A.W. and H.E. Johnson. 1975. Effect of PCB (Aroclor^ 1254) and
p,p'-OOT on production and survival of Daphnia magna Strauss. Bull. En-
viron. Contam. Toxicol. 13: 412.
Marking, L.L. 1966. Evaluation of p,p'-DDT as a Reference Toxicant in Bio-
assays. In; Investigations in Fish Control. U.S. Fish Wild!. Serv. Resour.
Publ. U.S. Oept. Inter. 14: 10.
McLane, M.A.R. and L.C. Hall. 1972. DDE thins screech owl eggshells.
Bull. Environ. Contam. Toxicol. 8: 65.
B-61
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McNIcholl, P.G. and W.C. Mackay. 1975. Effect of DOT on discriminating
ability of rainbow trout (Salmo gairdneri). Jour Fish. Res. Board Can.
32: 785.
Miles, J.R.W. and C.R. Harris. 1973, Organochlorine insecticide residues
in streams draining agricultural, urban-agricultural, and resort areas of
Ontario, Canada - 1971. Pestic. Monitor. Jour. 6: 363.
Miller, O.L. and O.M. Ogilvie. 1975. Temperature selection in brook trout
(Salvelinus fontinalis) following exposure to DOT, PCB, or pnenol. Bull.
Environ. Contam. Toxicol. 14: 545.
Moffett, G.B. and J.O. Yarbrough. 1972. The effects of DOT, toxaphene, and
dieldrin on succinic dehydrogenase activity in insecticide-resistant and
susceptible Gambusia affinis. Jour. Agric. Food Chem. 20: 558.
Nimmo, D.R. and R.R. Blackman. 1972. Effects of DOT on cations in the
hepatopancreas of penaeid shrimp. Trans. Am. Fish. Soc. 101: 547.
Nimmo, D.R., et al. 1970. Localization of DDT in the body organs of pink
and white shrimp. Bull. Environ. Contam. Toxicol. 5: 333.
Odum, W.E., et al. 1969. DOT residues absorbed from organic detritus by
fiddler crabs. Science. 164: 576.
Ogilvie, D.M. and J.M. Anderson. 1965. Effect of DOT on temperature selec-
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22: 503.
B-62
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Ogilvie, O.M. and D.L. Miller. 1976. Duration of a DDT-induced shift in
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Parrish, P.R. 1974. Aroclor^/ 1254, DDT and ODD, and dieldrin: Accumu-
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B-63
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B-64
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costracan crustaceans. Bur. Sport Fish. Wild!. Tech. Pap. 66: 19.
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B-65
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14
Stadnyk, L., et al. 1971. Pesticide effect on growth and C assimilia-
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Attachment C, October 5.
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B-66
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Mammalian Toxicology and Human Health Effects
BXPOSURE
Ingestion from Water
The solubility of DDT in water is approximately 1.2 ppb, al-
though the presence of salts, colloid and particulate material may
increase this solubility. An examination of Table 1 shows no in-
stance of natural water approaching the solubility limit (Bevenue,
1976). Lichtenberg, et al. (1970) noted that residues in surface
water peaked in 1966 and decreased in 1967 and 1968, and this trend
should be continuing. Since the primary source of DDT residues in
surface waters is runoff from drainage areas, the variations seen
in samplings range from nondetectable to 1 ppb. Variations result
in variable seasonal runoff patterns, sedimentation rates, amount
of pesticides on land areas, and distance from points of applica-
tion.
By utilizing the guidelines for deriving water quality criter-
ia for the protection of aquatic life (43 FR 29028), maximum con-
centrations of DDT in fresh water were calculated. To protect
freshwater aquatic organisms and consumers of these organisms, a
24-hour average concentration of DDT of 0.00023 ug/1 and a maximum
concentration of 0.41 ug/1 were proposed as standards. The chronic
levels proposed are near the limits of detection and subject to
significant analytical error (Gunther, 1969). The low chronic
level proposed may be a reflection of the large bioaccumulation
factor used in this model.
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TABLE 1
DDT and Metabolites in Waters of Different Areas*
Water Sources
Time
Period
PPt
Range
Galveston Bay (Gulf
of Mexico)
Selected Western
Streams (USA)
Selected Western
Streams (USA)
Surface Waters of
United States
Region:
Northeast
Middle Atlantic
Southeast
Ohio Basin
Great Lakes
Missouri Basin
South Central
Southwest
Northwest
Iowa Rivers (USA)
Arkansas Bay, Texas
(USA)
Big Creek, Ontario,
Canada
Seawater, California
Current System
Hawaii:
potable waters
marine waters
Rivers, Southern
California Bight area
1964
1965-1966
1966-1968
1967-1968
1968-1970
1969
1970
1970
1971
1970-1971
1971-1972
**N.D.-1,000
N.D.-120
N.D.-120
N.D.-30
N.D.-30
N.D.-60
N.D.-5
N.D.-270
N.D.-840
N.D.-110
N.D.-30
N.D.-20
N.D.-23
N.D.-100
3-67
2-6
ca 1
1-82
120-880
*Source: Bevenue, 1976
**N.D. = Non detectable
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The National Academy of Sciences Safe Drinking Water Committee
estimates the carcinogenic risk to man to be an excess death rate
of 63 persons per year at a 10 pg/1 exposure. These calculations
were for direct exposure from water intake and do not account for
bioconcentration effects. In 1976, the U.S. EPA recommended that
water levels not exceed 0.001 ug/1 on the basis of bioaccumulation
in food and adverse effects in birds.
According to Lichtenberg, et al. (1970), fresh water entering
treatment plants contained DDT residues in amounts of 0.01 to 0.002
of the permissible levels for public water supplies as described in
the Water Quality Criteria (Fed. Water Pollut. Control Adm. , 1968)
of 50 ng/1.
Assuming an average daily intake of 2 liters of water per
individual, Huang (1972) concluded that the maximum daily ingestion
would be 0.002 mg DDT, which is based on the highest recorded lev-
els in water. This would amount to approximately 5 percent of the
total daily dietary intake. Most of the evidence indicates that
DDT residues in drinking water are 1 to 3 orders of magnitude less;
therefore, it has been concluded that recorded DDT residues in
water probably make only a minor contribution to DDT ingestion by
human populations but may contribute to bioconcentration in aquatic
species and higher organisms in the food chain (Woodwell, et al.
1967). Recent monitoring studies of DDT in water are summarized in
Table 2.
Ingestion from Food
The accumulation of DDT in different species of widely differ-
ent phyla has made it the classical compound for study of biological
C-3
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TABLE 2
DDT Residues in Ocean Water
Location and Date
— 9
DDT, 10 g/liter Reference
Southern Calif., 1974
Irish Sea, 1974
Firth of Clyde, 1974
North Sea, 1974
English Channel, 1974
Mississippi Delta
Gulf Coast
Open Gulf of Mexico
Southern Calif. Bight
near Los Angeles, 1973
Bight western boundary,
1973
Bight western boundary,
1975
Near Los Angeles, 1975
San Francisco Bay, 1975
Mediterranean, 1974
Pacific offshore waters
of Mexico, 1975
0.30 - 1.80
<0.01 - 0.24
0.02 - 0.05
<0.01 - 0.04
<0.01 - 0.03
1.70 (mean)
0.35 (mean)
0.25 (mean)
40 - 60
0.44 - 1.40
<0.10 - 0.50
<0.30 - 8.00
Scura and McClure
(1975)
Dawson and Riley
(1977)
Dawson and Riley
(1977)
Dawson and Riley
(1977)
Dawson and Riley
(1977)
Giam, et al. (1978)
Giam, et al. (1978)
Giam, et al. (1978)
Risebrough, et al.
(1976)
Risebrough, et al.
(1976)
Risebrough, et al.
(1976)
Risebrough, et al,
(1976)
0.11 (mean of 26) Risebrough, et al.
(1976)
0.25 - 1.3
0.003 - <0.1
Risebrough, et al,
(1976)
Risebrough, et al.
(1976)
C-4
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magnification of pesticides. An abundance of literature attests to
the widespread movement of persistent residues along food chains in
natural environments coupled with the biological concentration of
the residue at each trophic level. Magnification of DDT occurs by
two routes: (1) direct absorption from contaminated water by
aquatic organisms and (2) transfer of residues through sequential
predator feeding.
Nontarget species, such as predatory birds, have been severely
affected through reproductive loss due to eggshell thinning. Al-
though in no way comprehensive, the following selected papers
illustrate the relative magnitude of bioconcentration of DDT.
14
Johnson, et al. (1971) introduced C-labeled DDT into fresh water;
within 3 days from initial exposure, the magnification factor in
two groups of invertebrates (Cladocera anci Diptera) ranged over
100,000 times; in two others (Amphipoda and Ephemeroptera), excess-
es of 20,000 occurred; and in Decapoda anC Odonata, magnification
was up to 3,000 times. Cope (1971) calculated the accumulation of
DDT in comparison to water for several species as follows: 70,000
times Eor oysters, 1,000,000 times for coho salmon, and 1,200 to
317,000 times in other fish. As a final example of bioconcentra-
tion, woodwell, et al. (1967) measured DDT residues in a Long
Island marsh area and observed the following pptn on a whole body
wet weight basis: for plankton, 0.04; water plants, 0.08; snail,
0.26; shrimp, 0.16; minnow, 0.94; bill fish, 2.07; heron, 3.5;
cormorant, 26.4; gull, up to 75.5. However, the decline of DDT in
the environment is reflected in the decrease in the residues of
various avian species (Johnson, 1974; Klass and Belisle, 1977;
C-5
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Anderson, et al. 1975; Spitzer, et al. 1978; Barber and Warlen,
1979).
The primary route of human exposure to DDT is from ingestion
of small amounts in the diet. These residues are transferred from
agricultural soils, of which 5 percent of the total area has been
heavily treated and has an estimated average content of 2 ppm
(Edwards, 1966). Since the half-life of DDT is approximately 3 to
10 years (Menzie, 1972) and sandy soils can retain 39 percent at 17
years (Nash and Woolson, 1967), the presence of DDT residues in
foodstuffs derived from contaminated soils will continue for some
time.
Monitoring programs by the Food and Drug Administration (FDA)
have been conducted in 80 markets nationwide in the period from
1965 to 1970, and the results are shown in Table 3 (Bevenue, 1976).
Meats, fish, poultry, and dairy products are the primary sources of
DDT residues.
As seen from these data, there have been co'ntinual decreases
in the overall levels of residues in all classes from 1965 to 1970.
Between 1970 and 1973, a significant drop in residues of DDT and
ODD occurred, constituting decreases of 86 and 89 percent, respec-
tively. DDE decreased only 27 percent. These decreases are re-
flected in the changing amounts of estimated dietary intake:
1965 - 0.062 mg/man/day, 1970 - 0.024 mg/man/day, and 1973 - 0.008
mg/man/day (U.S. EPA, 1975). This trend continued through 1977 as
reported by Johnson and Manske (1977). Compared to 49 percent of
the samples presently containing organochlorine residues, 54 per-
cent were observed in 1971. DDE in meat, fish and poultry has
:-6
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TABLE 3
DDT and Metabolite Residues in Food and Feed*
Product and
time period
DDT
Residue (ppm)
ODD DDE
*Source: Bevenue, 1976
Total
Dairy products (
1965-1966
1967-1968
1968-1969
1969-1970
Meat, fish, and
1965-1966
1967-1968
1968-1969
1969-1970
Grains
1965-1966
1967-1968
1968-1969
1969-1970
Leafy vegetables
1965-1966
1967-1968
1968-1969
1969-1970
fat basis
0.040
0.030
0.023
0.017
poultry (
0.299
0.103
0.101
0.072
0.008
0.004
0.005
0.004
0.012
0.015
0.010
0.007
Garden fruits (tomatoes,
1965-1966
1967-1968
1968-1969
1969-1970
Fruits
1965-1966
1967-1968
1968-1969
1969-1970
Oils (salad oil,
1965-1966
1967-1968
1968-1969
1969-1970
0.027
0.029
0.028
0.019
0.009
0.009
0.009
0.021
, 8-13% fat)
0.015
0.019
0.012
0.005
fat basis, 17-23%
0.139
0.062
0.043
0.049
0.002
0.001
0.001
0.001
0.016
0.007
0.001
0.001
cucumbers, squash,
0.017
0.015
0.012
0.016
0.003
0.001
0.004
0.001
margarine, peanut butter,
0.009
0.009
0.003
0.006
0.016
0.028
0.003
0.003
0.075
0.063
0.048
0.043
fat)
0.254
0.116
0.100
0.114
0.001
0.002
0.001
0.001
0.005
0.004
0.007
0.002
etc. )
0.005
0.002
0.002
0.002
0.002
0.002
0.001
0.001
etc. )
0.005
0.018
0.003
0.002
0.130
0.112
0.083
0.065
0.602
0.281
0.244
0.235
0.011
0.007
0.007
0.006
0.033
0.026
0.018
0.010
0.049
0.046
0.042
0.037
0.014
0.012
0.014
0.023
0.0.30
0.055
0.009
0.010
C-7
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declined from 0.114 to 0.033 ppm, and in dairy products from 0.043
to 0.017 ppm, while DOT remained constant in meat residues at 0.017
ppm. The decreases in pesticide residues in various food classes
indicate that the ban on DDT has indeed lowered the exposure of
humans via the diet. This decrease is paralleled by a lowering of
the total DDT equivalent in human tissues for the U.S. population
average from approximately 8 ppm to 5 ppm residue in fat from 1971
through 1974.
The acceptable daily intake of DDT established by WHO/FAO is
0.005 rag/kg/day. Duggan and Corneliussen (1972) reported the six-
year average from 1965 through 1970 in the U.S. diet of DDT and its
metabolites to be almost 10-fold less at 0.0007 mg/kg/day.
A bioconcentration factor (BCF) relates the concentration of a
chemical in aquatic animals to the concentration in the water in
which they live. The steady-state BCFs for a lipid-soluble com-
pound in the tissues of various aquatic animals seem to be propor-
tional to the percent lipid in the tissue. Thus the per capita
ingestion of a lipid-soluble chemical can be estimated from the per
capita consumption of fish and shellfish, the weighted average per-
cent lipids of consumed fish and shellfish, and a steady-state BCF
for the chemical.
Data from a recent survey on fish and shellfish consumption in
the United States were analyzed by SRI International (U.S. EPA,
1980). These data were used to estimate that the per capita con-
sumption of freshwater and estuarine fish and shellfish in the
United States is 6.5 g/day (Stephan, 1980). In addition, these
data were used with data on the fat content of the edible portion of
C-8
-------
the same species to estimate that the weighted average percent
lipids for consumed freshwater and estuarine fish and shellfish is
3.0 percent.
Numerous laboratory and field studies, in which percent lipids
and a steady-state BCF were measured, have been conducted on DDT.
The mean of the BCF values, after normalization to one percent
lipids, is 17,870 (see Table 5 in Aquatic Life Toxicology, Section
B). An adjustment factor of 3 can be used to adjust the mean nor-
malized BCF to the 3.0 percent lipids that is the weighted average
for consumed fish and shellfish. Thus, the weighted average bio-
concentration factor for DDT and the edible portion of all fresh-
water and estuarine aquatic organisms consumed by Americans is cal-
culated to be 53,600.
Inhalation
Levels of DDT found in the air are far below levels that add
significantly to total human intake. Stanley, et al. (1971) sam-
pled air in nine localities in both urban and agricultural areas in
the U.S. p,p'-DDT was found in all localities to range from 1 ng/m
of air to 2,520 ng/m . Generally, levels were highest in southern
agricultural areas and lower in urban areas. These samples were
taken during time of high usage of DDT. Most likely, air concen-
trations are much lower today. Kraybill (1969) estimated the con-
centration of DDT in the air to be 0.2 ng/m which is in the lower
range of Stanley's reported values. Several recent studies which
have monitored levels of DDT in the atmosphere are summarized in
Table 4.
C-9
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Location and Date
TABLE 4
Atmospheric DDT Residues
Number
of samples
DDT, 10"9 g/m3
Reference
Continental data
Mississippi Delta
1972
1973
1974
Kingston, FL, 1973-75
Sapelo Island, GA, 1975
Organ Pipe Natl. Pk.,
AZ, 1975
Hays, KS, 1974
Columbia, SC, 1976-79
156
(3-yr total)
99.5 mean of
16.0 monthly
11.9 average
levels
0.05 - 0.8
6
6
3
18
0.02
0.20
0.01
0.01
- 0.07
- 0.7
- 0.09
- 0.18
Arthur, et al.
(1976)
Bidleman, et al.
(1976)
Bidleman, et al.
(1976)
Bidleman, et al.
(1976)
Bidleman, et al.
(1976)
Bidleman and
Christensen (1980)
C-10
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TABLE 4 {cont.)
Location and Date
Number
of samples
DDT, 10~9 g/m3
Reference
2.
Marine data
Bermuda and North Atlantic
1973
Bermuda and North Atlantic
1974
Grand Banks, 1973
Chesapeake Bay, 1973
North Atlantic, 1976
Gulf of Mexico, 1977
English Channel, 1974
Barbados, W.I., 1977
Arabian Sea and Gulf of
Oman, 1977
11
25
5
10
6
13
0=009 - 0=053
- 0.062
<0.001
0.014 - 0.048
0.002 - 0.014
0.030 - 0.22
0.010 - 0.020
0.0024 (mean)
0.043 (mean)
Bidleman and Olney
(1974)
Harvey and
Steinhauer (1974)
Bidleman, et al.
(1976)
Harvey and
Steinhauer (1974)
Bidleman, et al.
(1976)
Giam, et al. (1978)
Giam, et al. (1980)
Dawson and Riley
(1977)
Bidleman (1979)
Bidleman (1979)
C-ll
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In a study on plant workers, Wolfe and Armstrong (1971) esti-
mated respiratory exposure from the contamination of filter pads
placed within respirators. The highest exposures reported were
33.8 mg/man/hour for the bagging operation, with a mean 14.11
mg/man/hour. The authors concluded that workers in formulating
plants not wearing respirators have significant intake of DDT via
inhalation. VJolfe, et al. (1967) used a similar method to deter-
mine inhalation exposure and found for airplane flaggers in dusting
operations 0.1 to 0.2 mg/man/hour levels.
Although inhalation may not be a significant source of expo-
sure to DDT in terms of the proportion of the daily dose, atmos-
pheric transport of DDT is apparently a significant route of envi-
ronmental transport (Woouwell, et al. 1971).
Dermal
Absorption of DDT through skin is minimal. Several factors
can influence the rate of absorption, such as the condition of the
skin or external factors such as temperature. Technical DDT was
less toxic dermally to white rats than a large percentage of other
pesticides tested by Gaines (1969). In Wolfe and Armstrong's study
(1971), most of the exposure was dermal with the exposure ranging
from 5 to 993 mg/man/hour. These high exposures did not correlate
with significant increases above the general population. This led
them to conclude that there was a minimal absorption of DDT in
exposed skin areas.
Summary
Hayes (1966) estimated the intake of DDT to be in the follow-
ing proportions: food - 0.04 mg/man/day, water - 4.6 x 10
C-12
-------
mg/man/day, and air - 9 x 10~ mg/man/day. Wessel (1972) calcu-
lated the daily dietary intake of DDT and analogues to be 0.027 ppm
DDT, 0.018 ppm DDE, and 0.012 ppm ODD. Kraybill (1969) estimated
DDT dietary intake to be approximately 85 percent of the total
exposure of 30 mg/year. Aerosols, dust and cosmetic exposure were
estimated as 5 mg/year, with air and water intakes of 0.03 and 0.01
mg/year, respectively.
From these estimates, it is concluded that the maximum total
intake of DDT and analogues does not exceed 0.1 mg/man/day and is
probably today considerably less, due to restriction in its use.
Since dermal, inhalation, and water intake account for less than 10
percent of the total dosage, and in most recent estimates, dietary
intakes are 0.008 mg/man/day, the actual total dose per day is
estimated to be approximately 0.01 mg/man/day or 3.65 mg/year.
PHARMACOKINETICS
Absorption
DDT and DDE are absorbed from the gastrointestinal tract with
high efficiency characteristic of dietary fat. Maximum lipid solu-
bilities reach 100,000 ppm. In as much as DDT and metabolites in-
gested are contained primarily in fat-bearing foodstuffs such as
dairy products, meat, and poultry, the absorption of dietary DDT
approaches the 95 percent absorptive values for these dietary fats.
Over 65 percent of labeled DDT and metabolites were found in the
9-day bile collections of treated rats (Jensen, et al. 1957).
Determinations of absorption and assimilation of ingested DDT
in humans have been studied by following the serum and adipose
lipid concentrations of the compound after chronic ingestion
C-13
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(Morgan and Roan, 1971). Highest concentrations were found in
serum three hours after ingestion of DDT. These concentrations
remained above pre-dose level for at least 14 hours but returned to
base level within 24 hours. Serum levels reflect a relatively slow
uptake and assimilation consistent with physiological dependence on
intestinal fat absorption. With a dosage of 20 mg intestinal ab-
sorption proceeded faster than transport from the vascular compart-
ment into tissue storage. Absorption of this entire dose was com-
pleted within 24 hours. One subject ingested a total of 2.82 g
technical DDT? approximately 85 percent was stored in body tissue
or excreted in the urine. The authors concluded that several fac-
tors collectively cause storage values of DDT to underestimate
absorptive efficiency.
Distr ibution
DDT and its metabolites have been found in virtually all body
tissues, approximately in proportion to respective tissue content
of extractable tissue lipid, except in the brain. Adipose/blood
ratios of DDT have been variously estimated from 140 to 1,000; more
recent estimations indicate that the ratio is approximately 280
(fat:plasma) (Morgan and Roan, 1977). This ratio represents a
dynamic equilibrium between DDT in plasma lipoprotein and in tri-
glycerides stored in fat cells.
Long-term admininstration of DDT to mice and its storage in
various tissues have been reported by Tomatis, et al. (1971).
Apart from o,p'-DDT, there is direct relationship between the con-
centration of each metabolite in each organ and the dose to which
the animal was exposed. The highest concentration of DDT and
C-14
-------
metabolites was found in fat tissue, followed by reproductive
organs, liver and kidneys together, and lastly, the brain. The
most prevalent stored compound was unaltered p,p'-DDT. Storage
levels of o,p'-DDT were proportionally higher in animals receiving
the control diet or exposed to the lowest DDT dose. In the repro-
ductive organs and fat, females had considerably greater levels of
all three compounds than males, with no storage differences in the
kidneys, brain, and liver.
In Rhesus monkeys, Durham, et al. (1963) noted that dosage
levels from 0.25 to 10 mg/kg/day technical DDT in the diet produced
a maximum storage in fat by six months, which was not increased by
DDT feeding for an additional period of seven years. Of interest
is the fact that no DDE was detected in the fat of these monkeys.
However, high levels of DDE storage were found in monkeys fed DDE,
indicating an inability to convert DDT to DDE.
Human adipose storage decreases in the order DDE ^p,p'-DDT^
DDD. Serum and adipose concentrations of DDE rise slowly to DDT
ingestion with the peak some months following termination of dos-
ing. In contrast, levels of DDT, DDD, and o,p'-DDT decline more
rapidly. Fitted exponential curves in man suggest that 25 percent
of stored material should be lost within a year after the last
administration. Elimination of very low levels from storage of DDT
proceeds much more slowly than disposition of the large stores of
DDT accumulated by occupationally exposed or dosed volunteers.
Thus, when DDT in fat amounts to 100 ppm, the chemical is lost at a
rate of 4.1 mg/day or 0.24 percent of the total store. When, after
two years, the load has decreased to 40 ppm, the loss rate falls to
C-15
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0.2 rag/day or 0.10 percent of store? projected to 5 ppm, storage
loss is 0.03 rag/day DDT or only 0.04 percent of body stores {Morgan
and Roan, 1971).
Hayes, et al. (1971} have shown that subjects ingesting high
doses up to 35 rag/kg/day DDT reach a storage plateau sometime be-
tween 18 to 22 months (Figure 1). Volunteers had mean adipose con-
centrations of 281 rag/kg with a high of 619 over a 21-month period.
DDE reached levels as high as 71 mg/kg with a mean of 25.8 mg/kg in
21 months, but the values increased during recovery to a peak of
563 mg/kg approximately two years after dosing, and fell only
slightly to 50.8 after a 3-year recovery. Over a 5-year recovery
period, the concentration of DDE in fat as a percentage of all DDT
derived material rose from 26 to 47 percent.
The preceding data are consistent with the known fact that DDE
is very slowly eliminated from the body and has the higher affinity
for storage. The average North American adult, with 17 kg of body
fat, contains approximately 25 mg of DDT and 75 mg of DDE. Storage
loss data predict that, if dietary intake were eliminated, most of
the DDT would be lost within one or two decades, but DDE would re-
quire an entire lifespan.
It has been suggested by a number of investigators that DDT
levels reflect recent exposure to DDT, while DDE levels correlate
well with long-term exposure and storage capacity of the human body
(Morgan and Roan, 1971; Edmundson, et al. 1969b). In occupation-
ally exposed workers, Laws, et al. (1967) determined the concentra-
tions in fat of DDE expressed as DDT to be 25 to 63 percent of total
C-16
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^000
500
I
O 35 my. pp'-OOT/man/day
• 3.5 mg. pp*-OOT/m»n/day
500
100
200 300
Tim* of treatment, days
FIGURE 1
400
600
Increase of the Concentration of p,p'-DDT in the Body Fat
of Men with Continuing Intake of p,p'-DDT*
Source: Hayes, et al. 1971
C-17
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DDT-related material. This is in contrast with 72 to 92 percent
found in the general population.
Tissue storages of DDE in the general population originate
almost entirely from dietary DDE rather than DDT conversion (Morgan
and Roan, 1971).
A comparison of DDT and DDE storage in the U.S population is
shown in Table 5 (U.S. EPA, 1975). Mean levels of DDT in human
adipose tissue show a downward trend from 7.95 ppm in 1971 to 5.89
ppm in 1973. Overall DDE levels on the other hand, do not show a
similai trend; long-term storage is reflected in the slightly in-
creased percentage of the total DDT found as DDE.
A simple linear model has been developed by Durham, et al.
(1965b) to describe the relationship between the concentration of
DDT in the body fat of man and the daily dcse of this compound. The
equation is: log C1 = 0.7 log I +1.3, where C^ is the fat storage
of DDT in ppm and I is the DDT intake in mg/man/day. This equation
is in good agreement with storage found by other investigators and
is represented in graphical form in Figure 2.
At high levels of exposure, human volunteers have demonstrated
a steady state of storage or plateau which is exponentially ap-
proached within 18 months. This plateau level is proportional to
the dose administered (Figure 1).
Harris and Highland (1977), in summarizing recent studies by
the U.S. EPA, reported mean DDT levels in human milk have been
measured to be 529 M9/kg fat (99 percent of 1,400 women) with a
maximum level at 34,369 ,ug/kg fat. In the same report, 100 percent
C-18
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TABLE 5
National Summary of Total DDT Equivalent Residues in Human Adipose Tissue*
(Total U.S. Population Basis)
Year
FY
FY
FY
FY
1970
1971
1972
1973
Sample Size
1
1
1
1
,412
,616
,916
,092
Frequency
99.
99.
99.
100.
3%
75%
95%
00%
Geometric Mean
7.
7.
6.
5.
87
95
88
89
ppm
ppm
ppm
ppm
Percent DDT
found as DDE
77.
79.
80.
81.
15%
71%
33%
19%
Total DDT equivalent = (o,p'-DDT + p,p'-DDT) + 1.114 (o,p'-DDD + p,p'-DDD + p,p'-DDE
+ o,p'-DDE)
*Source: U.S. EPA, 1975
C-19
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soo.o
i
o
Q
"3100.0
e
10.0
1.0
0.01
O Mean.
T standard error of mean
OJ 1.0
Daily dosv of DOT, mo/day
10.0
3 .0
FIGURE 2
Relationship Between the Concentration of DDT in the Body Fat
of Man and the Daily Dose of that Compound
Source: Hayes, et al. 1971; Durham, et al. 1965b
C-20
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of the women sampled had DDE residues in their milk. The mean and
maximum levels were 3,521 and 214,167 pg/kg fat, respectively.
Metabolism
The metabolism of DDT has been well established in several
mammalian species. Generally, two separate reductive pathways pro-
duce the primary endpoint metabolites, DDE and DDA. As seen in
Figure 3, a generalized outline of the metabolism of DDT, the pre-
dominant conversion is of DDT to ODD via dechlorination. This is
the first product in a series which results in metabolites which
are later excreted. The other primary pathway proceeds via reduc-
tive dehydrochlorination which results in the formation of DDE, the
major storage product in animals and humans.
Peterson and Robison (1964) showed convincingly that DDD was
the intermediate metabolite leading to DDA. Adult male rats were
treated acutely by gavage with 100 mg/kg purified DDT and sacri-
ficed 4 to 60 hours later. Liver samples yielded primarily DDT and
DDD, in a ratio of 14:1. Rats fed a diet of 1,500 mg/kg purified
DDT were sampled at 6 days; the livers yielded DDT, DDD, and DDE in
the ratios of approximately 3:5:1. Additional rats given 1,000
mg/kg DDD in identical manner of the DDT treatment showed DDD and
DDMU in a ratio of 1:13. Liver and kidney samples of DDE-treated
rats yielded only unchanged DDE, and the urine from a 2-week diet
of a 1,000 ppm DDE showed no detectable DDA. Furthermore, rats
treated acutely with DDMU were able to biologically convert this
compound to DDMS. Similarly, DDMS administration produced DDNU in
ratios of 2:5 in the kidney and 3:1 in the liver.
C-21
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•OOMS*
•OOWT
H—C—H
'DDNf
•DDOH*
Probable Inter-
mediate aldehyde
CHO
Probahtv
mrdiacc aldehyde
OOA
FIGURE 3
Metabolic Products of p,p'-DDT in the Rat
Source: Peterson and Robison, 1964
C-22
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The final conversion step of DDNU to DDA by hydroxylation
occurs more slowly. Short-term 6-hour exposure to DDNU produced
minimal amounts of DDOH. However, analyses of liver and kidney
tissue from rats fed 500 ppm DDNU diet contained equal quantities
of DDNU and DDOH, and the urine collected provided identification
of DDA. Each degradation product from DDT to DDNU when fed to rats,
was able to eventually exhibit DDOH and DDA in the urine. The alde-
hyde shown in Figure 3 was postulated by the authors as a briefly
existing intermediate between DDOH and DDA in mice.
14
Recent studies with pregnant rats using radiolabeled C-p,p'-
DDT give evidence of the sites in which a metabolite conversion
occurs. Thin layer chromatography of various tissues following
14
treatment with 0.9 mg C-DDT was utilized to determine the rela-
tive percentages of the metabolites produced. In the liver, from
12 to 24 hours the ratio of DDT, DDD, and DDE was unchanged at
approximately 3:3:1, a ratio similar to that found by Peterson and
Robison (1964) of 3:5:1 in male rats. Liver activity for DDT con-
version is much higher in the adults in comparison to neonates.
The results for the metabolites recovered from different tissues
and fetuses 8 to 10 hours post exposure are shown in Tables 6 and 7
(Pang, et al. 1977). DDE was the major metabolite in all tissues.
DDD was a minor metabolite, with the exception of spleen, in which
DDD and DDE were equal. DDA was detected in high levels in the
lung, intestine, kidney, and blood; in lower levels in the spleen,
placenta, and fetus; and was undetected in muscle tissue, the
heart, pancreas, and brain. These observations suggest that
enzymatic activity for the dehydrochlorination and reductive
C-23
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TADLE 6
14
Concentration of C-DDT and Its Hetabolitea in the Tisaues of Infant Rats after Consuming Milk fro* Daw that
received an Oral Dose of 0.9 mg C-DDT and in the tissues of the Dans (U9 DDT and Equivalent per g Dry Tissue)*
Elapsed
time
(days)*
1(4)
2(4)
3(4)
414)
7(4)
11(2)
14(3)
21(2)
28(2)
14
28(2)
StOMCh
content
(•ilk)D
15.26+3.27
4.56+0.46
3.49+1.09
3.73+0.35
1.62+0.18
1.73
0.75
1.33
0.48
-
-
Stoaach
4.82
2. 81+. 13
2. 31+. 06
2. 18+. IS
1.S9+.28
1.42
1.03
0.99
0.70
-
-
Dlood
0.72+. 34
1.59+.52
1.70+.15
1.59+.34
1.23+.29
1.21
0.44
0.40
0
0
0
Liver
13.93+2.50
14.23+6.94
12.13+4.26
8.29+1.73
5.96+1.78
5.93
4.64
2.39
1.11
0.68
0.50
Kidney
2.50+0.53
4.35+0.84
4.73*0.71
3.52+1.23
3.08+0.72
3.13
2.22
0.76
0.20
0.07
0.12
Intestine
Infant Rats
13.64+3.37
13.48+6.36
11.09+3.72
8.42+2.36
8.07+3.19
5.98
5.42
1.95
0.95
Oasis
0.20
0.78
Lung
3.30+0.97
5.05+1.60
6.23+1.78
5.25+0.75
4.26+0.79
4.25
1.97
1.13
0.52
1.12
0.40
Heart
1.21+.22
1.57+.56
1.91+.64
2. 00+. 85
1.13+. 26
1.81
1.04
0.60
0.18
0.28
0
Drain
1. 23+.3S
2. 24+. 94
1.83+.94
1.83+.72
1.11+.62
1.30
0.62
0.28
0.14
0.06
0.10
Caicass
6.34«2.
9. li + J .
10.98(1.
9.64» .
7-09* .
6. 16
4.39
2.42
1.87
-
-
22
12
63
57
ri7
aNu*ber of neonates used
Values are Means + standard deviation
•Source: Fang, et at. 1977
C-24
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14
TABLE 7
C-DDT and its Labeled Metabolites in Different Tissues of Pregnant Rats
8 or 10 hours after Receiving an Oral Dose of C-DDT*
Tissue
Blood
Brain
Fetus
Heart
Intestine
Kidney
Lung
Muscle
Pancreas
Placenta
Spleen
Radioactivity
recovered
(%)
83
100
86
100
93
88
100
99
100
100
83
0.02-0.04
DDA
26
0
8
0
39
24
41
0
0
4
11
0.36-0.43
ODD
10
18
20
10
18
5
6
0
5
9
32
Rp Values
0.46-0.52
DDT
31
36
25
67
11
24
8
9
15
5
14
0.56-0.61
DDE
33
46
35
20
31
34
32
72
59
49
36
75
1
0
12
3
1
13
14
19
21
27
0
*Source: Fang, et al. 1977
C-25
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dechlorination reactions transforming DDT to DDD and DDE are pre-
sent in all tissues, whereas the enzymes involved in the hydrogena-
tion and hydroxylation steps changing DDD to DDA are absent in the
brain, heart, pancreas, and muscle tissue of the rat.
The metabolism of o,p'-DDT in rats shows no striking differ-
ences to that of p,p'-DDT. Feil, et al. (1973) were able to detect
13 different metabolites in the rat excreta by nuclear magnetic
resonance spectra. Besides o,p'-DDD and o,p'-DDA, a number of
additional ring-hydroxylated DDA forms were present. Serine and
glycine conjugates and o, p'-dichloro-benz-hydrol were identified
in the rat urine. These results indicate that o,p'-DDT is exten-
sively metabolized.
Radiolabeled o,p'-DDD given orally in a 100 mg dose to rats
yielded, in both feces and urine, o,p'-DDA, aromatic 3,4-monohy-
droxy- and 3,4-dihydroxy-substitutec o,p'-DDA. Comparison of uri-
nary excretion of o,p'-DDD metabolites of rats and humans are fun-
damentally similar. Hydroxylation occurs primarily at the 3 and 4
positions. Humans show a higher percentage of total dose excreted
in the urine than rats, 10 to 50 percent versus 3 to 7 percent.
Serine and glycine conjugates are excreted in the urine of man and
rat (Reif and Sinsheimer, 1975).
The metabolism of DDT in the mouse follows essentially the
same pathways as the rat (Gingell and Wallcave, 1976). No species
differences in overall rates of metabolism of DDT, as measured by
urinary excretion of C were observed. Further studies inves-
tigating chronic exposure up to four months, have demonstrated fun-
damental differences in the metabolic and physiological handling of
C-26
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DDT among other rodent species. Both Swiss and CF, mice produce
small but significant amounts of DDE in urine, whereas none was
found in hamster urine (Gingell, 1976). With long-term feeding,
the mouse increasingly eliminates DDE, and at the termination of
the experiment, nearly as much DDE as DDA was found (Gingell and
Wallcave, 1976). The authors suggest that DDE may be the proximate
hepatotumorigenic metabolite in mice, in as much as hamsters are not
susceptible to DDT tumorigenesis and do not form DDE. Addition-
ally, hamsters are resistant to toxic effects of DDT up to 2,100
mg/kg (Agthe, et al. 1970).
Two major studies by Hayes, et al. (1971) and Morgan and Roan
(1977) are the basis for what is known of the metabolism of DDT in
man and are here described. Hayes, et al. (1971) performed two
studies, exposing volunteers from a U.S. penitentiary to technical
or recrystallized p,p'-DDT at rates from 3.5 mg to 35 mg/man/day.
In the first study, 10 subjects were studied: three for one year at
3.5 mg/man/day and seven for one year at 35 mg/man/day. In the
second study, 24 men could be followed for a period of over four
years. They consisted of four groups: Gl - a control, whose diet
was estimated as having 0.18 mg/man/day DDT; G2 - receiving 3.5 mg
technical DDT (85 percent p,p'-DDT); G3 - receiving 35 mg technical
DDT (85 percent p,p'-DDT); and G4 - receiving 35 mg recrystallized
p,p'-DDT.
Roan, et al. (1971) and Morgan and Roan (1977) measured the
concentrations of p,p'-DDE, p,p'-DDD and p,p'-DDA in blood, fat,
and urine in response to oral dosing with these compounds. Four
volunteers ingested technical DDT doses ranging from 5 to 20 mg/day
C-27
-------
for up to six months. The total dose ingested ranged from 0.06 g to
2.82 g. Two volunteers ingested a total dose of 0.45 g p,p'-DDE in
a 3-month period. A single volunteer was used for each dosing of
DDD and DDA for total dosages of 0.41 g and 0.105 g, respectively.
From these studies, Morgan and Roan (1977) concluded that the
conversion of DDT to DDE occurs with considerable latency. The
magnitude of conversion at these levels was estimated to be less
than 20 percent conversion in the course of three years. An upper
trend in DDE fat storage over this time course may be due to release
of stored DDT and further conversion to DDE, but no more than one-
fifth of the absorbed DDT ultimately undergoes this conversion.
The o,p'-isomer was not found to be present in fat and blood of the
subjects. DDE-dosed subjects did not exhibit any significant ex-
cretion of p,p'-DDA in excess of predose values. Dose-dependent
increases in DDD blood levels with DDT dosing indicated the exis-
tence of this metabolic pathway. Urinary DDA excretion and serum
DDD concentrations showed increases with DDT dosage and declined
after dosing ended. Conversely, DDE exhibited an upward trend for
months after dosing. These facts further support the mutually
exclusive role of DDD, rather than DDE, in the formation of the
urinary metabolite, DDA. Taken together, these results strongly
confirm that the metabolism of DDT in man is identical to the path-
ways reported by Peterson and Robison (1964) for the mouse. Meta-
bolic conversion of DDT by dechlorination to DDA proceeds more
rapidly and accounts for approximately one-fifth of the DDT load,
which is excreted in the urine. DDE, or the storage metabolite, is
C-28
-------
produced from DDT more slowly, via dehydrochlorination, and overall
conversion will be approximately 20 percent in three years.
Excretion
Studies were conducted by Wallcave, et al. (1974) on the
excretion of DDT metabolites in hamsters and mice. Of the ingested
dose of between 22 to 29 mg per animal over a 4-month period, 12 to
14 percent was recovered in the urine as DDA or DDE. Steadily in-
creasing amounts of DDE excretion were observed in mice with long-
term feeding, whereas the hamster had no DDE present. Approximate-
ly 9 percent of ingested DDT was found in fecal excretion as ODD or
DDT in mice, as compared to 3 percent in hamsters. These species
seem to have less biliary excretion than the rat, in which 65 per-
cent of a DDT dose can be found in the bile collections and large
amounts of DDT conjugate are found in the feces (Jensen, et al.
1957).
The excretion of DDT was investigated in human volunteer stud-
ies of Hayes, et al. (1971) and Roan, et al. (1971), previously
described. Excretion of DDA in the urine increased rapidly in the
first few days following a gradual increase in the subjects dosed
with 35 mg/man/day to a steady level of approximately 13 to 16 per-
cent of the daily dose. DDA excretion fell rapidly following ces-
sation of dosing. Since storage levels did not increase after
reaching steady state, these volunteers were apparently able to
excrete the entire dose of 35 mg/day. This is probably due primar-
ily to excretion of DDT from the gut, inasmuch as only 5.7 mg/day
of all DDT isomers were found in urine. Gut organisms have a
C-29
-------
demonstrated capacity for degradation of DDT to ODD and DDA and may
be important in fecal excretion.
Occupationally exposed workers have been shown to have signif-
icantly increased levels of DDA excretion in the urine. Ortelee
(1958) classified individuals as heavy, moderate and slight expo-
sure groups in formulating plants and found a good correlation
between exposure and DDA in the urine. Laws, et al. (1967) were not
able to find DDA in urine samples from all persons of the general
population due to insensitivity of analytical methods at the time.
In workers, increased levels of DDA excretion were found, but para-
doxically, DDE was found in only slightly higher concentrations in
exposed workers versus the general population with no correlation
with increasing work exposure. Estimations of total intake of DDT
based on DDA in urine are in good agreement with estimations of
intake based on the calculations of DDT in fat by Durham, et al.
(1965a).
Morgan and Roan (1977) have calculated from excretion measure-
ments the following rank order of loss rates from storage (from
fastest to slowest): DDA, ODD, o,p'-DDT, p,p'-DDT, and p,p'-DDE.
Differences in excretability from one end of the scale to the other
are very great, water solubility being a possible important varia-
ble. Interspecies differences also exist in the capacity for un-
loading stored DDT. Man, as compared to the rat, dog or monkey,
exhibits a considerably slower rate of loss, which may be related
to differences in renal handling of the pesticide. If dietary in-
take were completely eliminated, most of the DDT would be lost in
C-30
-------
10 to 20 years but DDE would require almost an entire lifespan for
removal.
EFFECTS
Acute, Subacute, and Chronic Toxicity
Acute toxic effects show central nervous system symptoms, such
as hyperexcitability/ generalized trembling, convulsions, and
paralysis within 5 to 10 minutes following intraveneous (i.v.)
administration and a latent period of several hours for oral dosing
in experimental animals. LD50 values for rats typically range from
100 to 400 mg/kg orally and 40 to 60 mg/kg i.v. (Negherbon, 1959;
Hayes, 1963). Dermal exposure in rats was toxic at 3,000 mg/kg.
DDE has an oral LD in rats of 380 mg/kg in males and 1,240 mg/kg
in females; DDA, 740 mg/kg in males and 600 mg/kg in females
(Hayes, et al. 1965). The oral LD5Q of DDT is 60 to 75 mg/kg in
dogs, 250-400 mg/kg in rabbits, and 200 mg/kg in mice (Pimentel,
1971).
Studies on acute toxicities in animals indicate that the cor-
relation between pathological symptomatic effect and pesticide
level is highest in the brain. Dale, et al. (1963) observed trem-
ors in male rats four hours after administration of DDT, when the
brain concentration reached 287 ppm on a lipid basis.
Acute poisoning in man is a rare event, and no well-described
case of fatal uncomplicated DDT poisoning has been reported. Gen-
eral symptoms are similar to those found in animals and include
dizziness, confusion, and, most characteristically, tremors. In
severe poisoning, convulsions and parasthesia of extremities may
intervene.
C-31
-------
Single ingestion of 10 rag/kg produces illness in some, but not
all, subjects. Smaller doses generally produce no illness. Con-
vulsions and nausea frequently occur in dosages greater than 16
mg/kg. Dosages as high as 285 mg/kg have been taken without fatal
result, but such large dosages are usually followed promptly by
vomiting, so the amount retained is variable (Hayes, 1963).
Although a number of pathological changes have been noted in
experimental animals, the most consistent finding in lifetime feed-
ing studies has been an increase in the size of liver, kidneys and
spleen, extensive degenerative changes in the liver and an in-
creased mortality rate. In rats, Laug, et al. (1950) observed
hepatic alteration with feedings in diet at 5 ppm DDT. At dose lev-
els of 600 and 800 ppm, significant decreases in weight gain and
increased mortality were observed in rats (Fitzhugh and Nelson,
1947). The observation that increased mortality results from doses
above 100 ppm DDT in the diet is well established in mice (Walker,
et al. 1972).
In contrast to the rodent models, Rhesus monkeys fed diets
with up to 200 ppm DDT showed no liver histopathology, no decrease
in weight gain or food consumption, or no clinical signs of ill-
ness. Several monkeys fed 5,000 ppm in the diet had some weight
loss prior to early death due to DDT poisoning (Durham, et al.
1963). In one animal, liver pathology consistent with DDT poison-
ing in other animals was found.
No clinical or laboratory evidence of injury to man by repeat-
ed exposure to DDT has been reported. Volunteers ingesting up to
35 mg/day for 21 months had no alterations in neurological signs,
C-32
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hematocrit, hemoglobin, and white blood cell counts. No changes in
cardiovascular status or liver function tests were noted (Hayes, et
al. 1971).
Studies of exposed workers by Laws, et al. (1967), Wolfe and
Armstrong (1971), and Almeida, et al. (1975) have demonstrated no
ill-effects from long-term high levels of exposure, as judged by
physical examination and chest X-ray.
Furthermore, the dermal toxicity of DDT in humans is prac-
tically nil. A few cases of allergic reaction have been observed,
which may be due to the extreme sensitivity of the individual.
Synergism and/or Antagonism
One of the primary concerns about pesticide residues is the
possibility that they may act synergistically with other chemicals
over a long period to produce cancer. The accumulation and summa-
tion of carcinogenic exposure from various sources may present a
health problem of great significance.
DDT, a strong inducer in the mixed function oxidase system,
potentially could enhance the biological effects of other chemicals
by activation or diminish their activities through detoxification
mechanisms. Weisburger and Weisburger (1968) were able to enhance
the incidence of hepatomas in rats caused by N-fluorenacetamide
(2-AAF) by co-administration of DDT. They had previously shown
that 2-AAF is metabolized by a mixed function oxidase system (MFO)
to the hydroxy intermediate which is carcinogenic. By stimulating
liver metabolism with 10 mg/day DDT which, by itself, causes no
hepatomas, the percentage of animals bearing tumors from a dose of
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1 mg/day 2-AAF for up to 52 weeks rose from 67 to 90 percent in
males and from 7 to 33 percent in females.
Conney (1967) observed decreases in phenobarbital-induced
sleeping times proportional to the dose of DDT given to rats two
days earlier. Doses of 1 and 2 mg/kg of body fat caused a 25 and 50
percent reduction in sleeping time, respectively. This response is
due to the greater capacity of the MFO system to detoxify pheno-
barbital to a more readily excretable form. Similar effects have
been seen for Librium, methyprylon, and meprobamate in rats (Datta
and Nelson-, 1968).
Enhancement of metabolic activity has been demonstrated in
workers occupationally exposed to several insecticides, DDT includ-
ed (Kolmodin, et al. 1969). In these workers, the half-life of
antipyrine was significantly decreased in comparison to controls.
Deichmann, et al. (1967) evaluated the synergistic effects of
aramite (200 ppm), DDT (200 ppm), methoxychlor (1,000 ppm), thio-
urea (50 ppm), and aldrin (5 ppm) given singly or in combination to
rats. These dosages were approximately 50 percent of the levels
reported to induce liver tumors. Rats fed combinations of aramite,
DDT, methoxychlor, and thiourea, with a total tumorigenic dose of
200 percent had a 17 percent tumor incidence. Similarly, a combi-
nation of aramite, DDT, methoxychlor, and aldrin had a 10 percent
tumor incidence. Single chemical feedings had the following inci-
dences of tumors: aramite - 23 percent, DDT - 17 percent, methoxy-
chlor - 18 percent, thiourea - 28 percent, and aldrin - 25 percent.
Control rats had 23 percent tumors. Since both total tumors
and liver tumors were essentially the same in control versus
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experimental groups, those authors concluded that the compounds did
not act in an additive manner and further suggested that the mix-
tures might have an antagonistic effect in the reduction of tumors
below control.
Walker, et al. (1972) produced liver tumors in mice with
either 100 ppm DDT or 5 ppm dieldrin. Two types of histology were
scored: simple nodular growth of parenchymal cells (A), and papil-
liform adenoid growth of tumor cells (B). Combination of the two
chemicals showed an overall increase in tumor numbers in males
only, 53 to 88, when compared to 100 ppm DDT alone. What is most
striking, however, is that for both males and females, there was a
significant shift in proportion to the more tumorigenic type D
phenotype with the combined feeding.
The induction of the hepatic enzymes occurs in animal models
and possibly in occupationally exposed workers, as shown by in-
creased drug metabolism. However, the tumorigenicity data present
inconsistent findings with respect to activation or detoxification,
depending on the agent used. This is not an uncommon paradox when
dealing with metabolic induction. The effects on human health as a
result of low level exposure and synergistic/antagonistic interac-
tions with other chemicals are unknown.
Teratogenicity
Minimal teratogenic effects have been reported following high
acute dosages. Hart, et al. (1971) showed that DDT has an effect on
prematurity and causes an increase in the number of fetal resorp-
tions in rabbits given 50 mg/kg on days 7, 8 and 9 of gestation. In
the experimental group, 25 percent of the implantations were
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reabsorption in utero in comparison to 2 percent in the controls.
The weight of the viable fetuses were significantly lower in the
treated animals. The dose used in the experiment corresponds to
one-sixth to one-tenth of the acute LD50 for the species.
Low level exposure to DDT exerts an adverse effect on repro-
duction of several avian species. While data for mammalian species
are meager, published reports to date indicate that dietary intake
has little or no effect on the reproductive success of laboratory
animals. Dietary DDT at 7 ppm was fed to BALB/C and CFW strains of
Swiss mice for 30 days prior and 90 days post-breeding. In the
BALB/C strains, there was a slight reduction in overall fertility,
but fecundity (litter size) was greater than control values. With
the CFW strain, no differences in fertility or fecundity were noted
(Ware and Good, 1967).
Ottoboni (1969) studied the effect of DDT at levels of 0, 20,
and 200 ppm on fertility, fecundity, neonatal morbidity, and mor-
tality through two successive generations in Sprague-Dawley rats.
Neither alteration in sex ratios nor any evidence of teratogenic
effect was found among live or stillborn young. Litter size,
weights at birth and weaning showed no differences between treated
and control. Poor survival of the newborn pups to weaning age in
the 200 ppm group was observed. This finding was compromised by
large losses in the control, yet the 20 ppm diet group was un-
affected. Viability of young was high for all three groups in the
F, generation breedings. Of the other indices studied, fecundity,
fertility and mortality, none was significantly affected. The only
significant finding was an increase in ring tail, a constriction of
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the tail followed by amputation, in the offspring of mothers whose
diets contained 200 ppm DDT.
Krause, et al. (1975) noted a damaging effect on spermatogene-
sis in rats which was somewhat persistent for 90 days, and fertil-
ity was markedly reduced. This followed acute 500 mg/kg dose on
days 4 and 5 of life or 200 mg/kg from day 4 to day 23. In this
experiment, the administered dosages are close to the LD for the
species; therefore, these results cannot be considered conclusive,
since acute toxicity will alter other physiological parameters that
could affect fertility.
Both p,p'-DDT and o,p'-DDT have been shown to possess estro-
genic activity in rodents and birds (Welch, et al. 1969; Bittman,
et al. 1968)., Increases in uterine wet weight, and uptake of
labeled glucose into various precursors which are in competition
with estrodiol 17B for uterine binding sites have been demon-
strated.
The importance of the estrogenic activity of low level DDT
exposure is difficult to estimate. Since fertility in mammals is
dependent upon complex hormonal interactions, chemical interfer-
ence may represent a hazard. As an example, Ottoboni (1969) sug-
gested that 20 ppm of DDT in the diet had an adverse effect on the
subfertile females in their reproductive prime and observed a
greater fertility or protective effect in aging female rats as com-
pared to controls. In a later study by Wrenn, et al. (1970) long-
term feeding of o,p'-DDT to rats did not interfere with normal
reproduction nor were estrogen-sensitive physiological parameters
significantly affected.
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Mutagenicity
DDT has not shown mutagenic activity in any of the bacterial
test systems thus far studied. McCann, et al. (1975) found no in-
creased frequency of reversions in Salmonella typhimurium strains
TA-1535, 1537, 98, or 100 with 4 jjg/plate DDT. In addition, DDE was
nonmutagenic in this system; neither DDT nor DDE were positive with
S-9 microsomal activation. Marshall, et al. (1976) confirmed these
studies with doses up to 2,500 ug/plate DDT and 1,000 pg/plate DDE.
No inhibition of growth was seen in the E. coli Pol-A strains with
500 ug of DDT and the metabolites ODD and DDE (Fluck, et al. 1976).
DDT was also negative in the rec-assay with Bacillus subtilis
(Shirasu, et al. 1976).
Fahrig (1974) reviewed the activity of DDT and-its metabolites
DDE, ODD, DDOH, and DDA in several other bacterial systems. All
metabolites were negative, as judged by resistance to 5-methyltryp-
tophane and streptomycin in liquid holding tests. Back mutation to
prototrophy was negative in two strains of Escher ichia marcescens
and was negative to galactose prototrophy in E. coli.
The only positive result found in any of the bacterial test
systems was reported by Buselmaier, et al. (1972) upon the adminis-
tration of ODD to mice and assaying for back mutation of Salmonella
typhimur ium and E. marcescens following incubation in the perito-
neum in the host-mediated assay. However, DDT, DDE, and DDA were
found negative by this method.
In summary, with the exception of the metabolite, DDD, in the
host-mediated assay, no genetic activity has been detected in the
prokaryotic test systems.
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Tests on eukaryotic yeast cells have been uniformly negative.
Fahrig (1974) investigated the effect of DDT and various metabo-
lites on mitotic gene conversion in Saccharomyces cerevisiae/ which
detects single strand breaks of the DNA. Host-mediated studies
with DDT, ODD, and DDE of cells incubated in the testis, liver, and
lung of rats were also negative. Clark (1974) found no significant
increases in mutagenicity of conidia of Neurospora crassa incubated
ijn vitro and ^_n vivo with the host-mediated assay.
Vogel (1972) measured X-linked recessive lethal mutations in
Drosophila melanogaster and found activity for DDT and DDA, with
negative results for DDE, ODD, and DDOH.
Clark (1974) examined the relationship between spermatogene-
sis stages in D. melanogaster and the effect of DDT on dominant
lethality and chromosome abnormalities. Sequential breedings of
the treated males with virgin females at three day intervals indi-
cated that DDT causes an increase in dominant lethality in early
spermatid and spermatocyte stages. This increased lethal effect
was correlated with an increase in nondisjunction.
In mammalian systems, the mutagenic activity of DDT and its
metabolites is relatively weak. This is evidenced by the fact
that, depending upon the dose and route of administration, and the
species sensitivity of the test organism, reported studies are
negative or marginally positive.
High doses of technical DDT administered orally to mice at 150
ing/kg/day for two days (acute) or 100 mg/kg DDT twice weekly for 10
weeks (chronic) showed significant increase in the number of dead
implants per female. Acute treatment showed maximum sensitivity in
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induction of dominant lethals in week 5 and chronic treatment in
week 2, with continued increases above control through week 6.
Chronic, but not acute, dosing caused significant reductions in
sperm viability and a reduction of cell numbers in all stages of
spermatogenesis (Clark, 1974).
Oral feeding of two strains of mice at lower levels {1.05
mg/kg/day) showed little effects in reproductive response. Both
CFW and BALB/C strains of Swiss mice fed DDT showed lesser parent
mortality than control. Neither fertility, as measured by pairs
producing young, or fecundity, as measured by litter size, was
statistically different from the contol. Number of litters per
pair was not diminished (Ware and Good, 1967).
Two additional studies have been reported with negative re-
sults for dominant lethality in mice (Epstein and Shafner, 1968?
Buselmaier, et al. 1972). Intraperitoneally (i.p.) treated male
rats in doses up to 80 mg/kg for five days showed no effect in domi-
nant lethality or fertility (Palmer, et al. L973). Five-day oral
doses of 25, 50, or 100 mg/kg given to males bred sequentially for
six weeks, showed a statistically significant effect in implanta-
tion loss only in week three at 100 mg/kg level.
Oral feeding of technical DDT at 20 and 200 ppm/body weight in
the diet of Sprague-Dawley rats for two generations produced no
apparent effect on fertility, fecundity, neonatal morbidity, or
mortality through two generations (Ottoboni, 1969). By contrast,
juvenile male rats of the Wistar Han strain, fed 500 mg/kg on days 4
and 5 after birth (acute) and 200 mg/kg pure DDT daily from day 4 to
23 (chronic) showed damaging effects on spermatogenesis: testicular
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weight, tubular diameter, wall thickness, and number of sperma-
togonia (Krause, et al. 1975).
There are relatively few papers reporting the effect of DDT
and metabolites on mammalian chromosomes. Johnson and Jalal (1973)
studied the effect of DDT on the bone marrow of i.p. injected
BALB/C mice exposed to one single administration of 100, 150, 200,
300, and 400 ppm/body weight. Doses of 150 ppm and greater caused a
significant increase in the number of cells with fragments; sticky
cells were significantly increased at all concentrations. Smaller
doses were tested by Larsen and Jalal (1974) in brown and BALB/C
mice: 25, 50, 100, and 250 ppm did not significantly affect the
number of gaps, stickiness or mitotic indices, but deletions and
gaps plus deletions were significantly higher or approached the
significant levels at 50 ppm and higher concentrations.
Rats treated by i.p. or by gavage with doses ranging from 20
up to 100 ppm/body weight did not show a dose-response relationship
or an increase in percent of chromosomal aberrations over the con-
trols (Legator, et al. 1973).
DDE, but not DDT, caused an increase in chromosome aberrations
in a Chinese hamster cell line (V79) at 30 and 35 ug/ral (Kelly-
Garvert and Legator, 1973).
Palmer, et al. (1972) found a significant increase in cells
with structural aberrations when an established cell line of the
kangaroo rat, Potorus tridactylis apicalis was exposed to 10 pg/ml
p,p'- and o,p'-DDT, p,p'- and o,p'-DDD, and p,p'-DDE. The p,p'-DDA
was the least toxic among DDT metabolites, since only a concentra-
tion of 200 pg/ml caused a cytopathic effect, whereas DDT, ODD, and
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DDE - p,p' and o,p' - were toxic at 20 and 50 ug/ml. Mitotic inhi-
bition was intense in cultures treated with o,p'- and p,p'-DDT (40
percent and 35 percent more, respectively, than in the control).
Cultures exposed to p,p'- and o,p'-DDD and DDE had indices of 20 to
25 percent below the control; almost no inhibition was observed
with p,p'-DDA. The rate of chromosomal aberrations depended upon
the isomer used: p,p'-DDT, DDD, and DDE caused a twofold increase
as compared to the o,p' isomers. At 10 pg/ml p,p'-DDT, DDD, and DDE
caused chromosome damage to 22.4, 15.5 and 13.7 percent of the
cells, respectively. Approximately 12 percent of the abnormal
cells produced by p,p'-DDT and p,p'-DDE had rearrangements. Only
10 percent of the cells treated with p,p'-DDD had rearrangements.
The o,p' isomers did not produce exchanges.
Mahr and Miltenburger (1976) confirmed the fact that DDA is
the least effective of DDT metabolites in producing cytogenetic
damage and inhibiting proliferation in the Chinese hamster cell
line B14F28. The proliferation rate after a four-hour treatment
was inhibited most strongly by DDD (at 75, 45, and 22 ppm) , fol-
lowed by DDT (81 ppm) and DDE (88 ppm); 100 ppm DDA did not produce
any effect. The continuous presence of DDT (8 ppm) in the medium
for three months did not result in an altered proliferation rate in
cultures. Chromosome damage (i.e., breakage and gap formation) was
observed with 41 and 81 ppm DDT, 45 and 75 ppm DDD, and 44 and 88
ppm DDE. Here again DDA was the least effective in producing
chromosomal damage; at the highest concentration chromosomal gaps,
but not breaks, were increased. No chromosomal structural anom-
alies were found in the experiment.
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Hart, et al. (1972) found no increase in chromosomal aberra-
tions in human or rabbit lymphocyte cultures exposed to 1, 5, 10,
30, 50, and 100 ug/ml DDT based on the analysis of 25 metaphases per
level in the human lymphocyte cultures. Liver cells from rabbit
fetuses whose mothers had been treated with DDT during pregnancy
showed no difference as to chromosome damage when compared to non-
treated controls.
Lessa, et al. (1976) exposed human lymphocytes i_n vitro to
very low concentrations of technical DDT ranging from 0.06 to 0.20
ug/ml and from 1 to 15 ug/ml. The lowest concentrations (0.06 to
0.20 ug/ml) are similar to those found in the plasma of individuals
of the general population in Brazil. No correlation was found be-
tween DDT dose and cells with chromosomal aberrations. At 0.20,
4.05, and 8.72 ;ag/ml the proportion of cells with structural aber-
rations was significantly greater than in controls. It is inter-
esting to note, though, that higher concentrations of approximately
12 and 15 ppm produced no such effect. Such effects may be caused
by precipitation of DDT in the culture medium or may reflect a
difference in the amount of binding of DDT and metabolites to the
lipid moiety in the serum, or even differences in cell permeabil-
ity.
Yoder, et al. (1973) reported an increase in chromatid lesions
in blood cultures from a group of 42 men occupationally exposed to
several pesticides, DDT included, during the spraying season, as
compared with cultures made six months before when the same indivi-
duals had not been in contact with the pesticides for 30 days.
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Rabello, et al. (1975) compared the frequency of cells witu
chromosomal aberrations in workers from three DDT plants, directly
and indirectly exposed to DDT. There was no significant difference
between these two groups. The total DDT and DDE levels in the
plasma were determined. In the 25 workers in direct contact with
DDT, the levels ranged from 0.16 ug/ml to 3.25 /ug/ml (mean 1.03
ug/ml _+ 0.79) total DDT and 0.03 to 1.77 ,ug/ml (mean 0.48 4- 0.52)
p,p'-DDE. In these 25 individuals not in direct contact with the
compound, values ranged from 0.03 to 1.46 ug/ml (mean 0.38 ug/ml +
0.15) total DDT and 0.01 to 0.41 ug/ml (mean 0.15 + 0.02) p,p'-DDE.
In one of the plants, though, not being in direct contact with DDT
did not prevent the workers from having DDT plasma levels as high
as those in workers who actually manipulated the substance. A sec-
ond comparison was then made between the groups with high and low
DDT plasma concentrations, which showed an increase in cells with
chromatid aberrations in the highly exposed group.
When another group of eight plant workers with total DDT
plasma levels ranging from 0.09 to 0.54 ug/ml (mean 0.24 ug/ml +
0.15) and DDE levels ranging from 0.02 to 0.09 ug/ml (mean 0.041 +
0.02) was compared to 10 individuals of the general population with
no detectable o,p'- or p,p'-DDT and DDE levels ranging from 0.02 to
0.04 ug/ml (mean 0.029 ug/ral + 0.01), no significant difference was
found in the cytogenetic analysis. A positive correlation was
found between DDT levels and length of exposure of all individuals,
but there was no correlation between DDT levels in the plasma and
frequency of cells having any type of chromosomal aberrations
(numerical or structural).
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No effect on unscheduled DNA synthesis was seen in SV40 trans-
formed human cells with concentrations up to 1,000 jiM DDT either
with or without S-9 microsomal metabolic activation (Ahmed, et al.
1977).
In summary, the evidence in prokaryotic and fungal systems
indicates that DDT and its metabolites do not produce point muta-
tions. Although the evidence is somewhat contradictory in the dom-
inant lethal studies, in vivo and in vitro cytogenetic studies seem
to indicate that DDT is a clastogenic (chromosome breaking)
substance.
Carcinogenicity
Fitzhugh and Nelson (1947) were the first to investigate the
carcinogenic potential by chronic feeding of DDT in rodents.
Osborne-Mendel weanling rats were fed diets containing 0, 10, 20,
40, and 80 mg/kg/day technical DDT for a period of two years. Path-
ologic examination revealed that the chief lesion was a moderate
degree of liver damage, which consisted of hypertrophy of centro-
lobular hepatic cells, hyalinization of the cytoplasm and focal
necrosis. Although no information as to dosage or sex of the
tumor-bearing animals was given, the authors concluded that defin-
ite but minimal hepatic tumor formation was evident. This conclu-
sion was based on comparison to many hundreds of similar aged rats
which spontaneously showed distinct hepatic tumors at a frequency
of one percent. By contrast, of the 75 rats surviving to 18 months,
15 exhibited either large adenomas or nodular ademonatous hyper-
plasia with similar microscopic morphologies, differing chiefly in
size. Chronic feeding produced degenerative changes in the liver
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at all doses. Acute admininstration of 1,000 mg/kg in the diet for
12 weeks produced the characteristic pathology which persisted for
2 weeks and reverted to a normal appearance when examined at 4, 6,
8, and 10 weeks post exposure.
Laug, et al. (1950) followed this study by administering lower
doses of technical DDT in the diet for periods of 15 to 27 weeks to
weanling rats. No hepatic cell alterations were noted in the con-
trols and 1 ppm levels, with minimal effects at 5 ppm. At doses of
10 and 50 ppm, definite hepatic hypertrophy was observed, but gross
alterations such as necrosis were not present. Ortega, et al.
(1956) confirmed that liver alterations can be observed in rats
with DDT levels as low as 5 ppm. However, this pathology was re-
versed to normal once the administration of the compound was
stopped.
The next major report on the carcinogenicity of DDT was the
work of Tarjan and Kemeny (1969) with BALB/C mice. Six generations
of mice were fed either the control diet, contaminated with 0,2 to
0.4 mg/kg DDT, or the test diet of 2.8 to 3.0 mg/kg p,p'-DDT. The
control group was comprised of 406 mice and the test group had 683
mice with a daily intake of 0.4 to 0,7 mg/kg. A striking increase
in the incidence of leukemias was seen for the diet supplemented
with pure DDT beginning at the F3 generation. Myeloid, lymphoid,
and aleukemias were found in 85 treated animals (12.4 percent) but
only the latter two types were found in 10 controls (2.5 percent).
In the F. and Fg generations, myeloid leukemias accounted for one-
third of the total malignancies. The authors further noted that in
BALB/C mice spontaneous leukemia is unknown. The induction of
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tumors in the experimental group was significant in the F- genera-
tion and increased almost logarithmically in successive generations
from F3. A total of 196 animals (28.7 percent) versus 13 (3.2 per-
cent) were found to have tumors in the exposed and control series,
respectively. The predominant tumor type was pulmonary carcinoma
(116/196 animals), and the authors claim that prior observation of
their colony shows incidence of malignant pulmonary tumors to be
below 0.1 percent. A variety of tumors was observed widely dis-
persed throughout the body and included malignant vascular tumors
(22/196) and reticulosarcomas (27/196) of the liver, kidney,
spleen, ovary, and other organs. The authors noted that these pos-
itive findings were somewhat complicated by the fact that fetal
exposure via placental passage and newborn intake through breast
milk may heighten adverse effects.
In a survey of 120 selected pesticides and industrial chemi-
cals to determine their potential carcinogenic!ty, five pesticides,
p,p'-DDT included, were among the 11 compounds that showed signifi-
cant increases in tumor incidence (Innes, et al. 1969). Two hybrid
strains of mice were bred by crossing C-57BL/6 with either C3H/Amf
or AKR strains; F, generations were designated strains X and Y,
respectively. From day 7 to 28, the animals were treated by gav-
age, at the maximum tolerated dose of 46.4 mg/kg in a 0.5 percent
gelatin suspension. From 4 weeks to 18 months, the chemical was
mixed directly in the diet to approximate this dose; the concen-
tration of DDT was calculated to be 21 mg/kg/day. The frequency of
mice with hepatomas in both strains as compared to controls is
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given in Table 8. Pulmonary tumors and lymphomas occurred in lower
frequencies but are not presented in the table.
The pattern of tumor type among several experimental compounds
was similar to the positive carcinogenic control compounds with the
major evidence for tumorigenicity arising from the increased inci-
dence of hepatomas. These increases were significant at the 0.01
level for the sum of both sexes and both strains, the sum of males
of both strains, and for the males of each separate strain of the
hybrids. Although incidence of lung and lymphatic tumors showed
fewer increases than hepatoma, the incidence of lymphomas was sig-
nificantly above negative controls for p,p'-DDT. The pulmonary
tumors consisted primarily of adenomas.
In 1967, the International Agency for Research on Cancer
(IARC) initiated a large investigation on the potential carcinogen-
icity of DDT in rodents. Studies were conducted in three different
strains of mice in Lyon, France, by Tomatis, et al. (1972) (CF^; in
Moscow (USSR) by Shabad, et al. (1973) (strain A); and by Terracini,
et al. (1973) in Milan (Italy) with BALB/C. In addition, a study was
performed on white rats in Leningrad (USSR) (Turusou, et al. 1973).
Although the rat study was negative, the long-term administration
of DDT to mice induced a significant increase in the frequency of
liver tumors, which constituted the strongest evidence to date for
the possible tumor igenicity of DDT. Tomatis, et al. (1972) and
Turusov, et al. (1973) fed six consecutive generations of CF^ mice
technical DDT in the diet, at doses of 0.3, 1.5, 7.5, and 37.5
mg/kg/day over the lifespan. CF^^ mice are characterized by a
rather high incidence of spontaneous tumors mainly of the lung,
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TABLE 8
Frequency of Animals with Hepatomas in Two Hybrid
Strains of Mice Exposed to 21.0 mg/kg/day p,p'-DDT
and to a Control Diet Without DDT*
Strain
C57 BL/6 x C3H/AmF
C57 BL/6 x AKR
Group
Exposed
Control
Exposed
Control
Total Number
of Animals
M F
18 18
79 87
18
90 82
Number of
Animals with
Hepatomas
M F
11 5
22 8
7 0
5 1
*Source: Innes, et al. 1969
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haematopoietic system, bone, and, in males, hepatomas. The per-
centage of animals bearing tumors of all types in DDT treated males
(89 to 94 percent) was somewhat higher than in the male controls
(78 percent). The DDT treated females had similar incidence (85 to
90 percent) to that of the female controls (89 percent). Only
liver tumor incidence was clearly affected by DDT treatment. DDT
treated male mice showed increases in liver hepatoma at all treat-
ment levels, with the peak at 37.5 mg/kg/day (301/350) and similar
incidence of 179/354, 181/362, and 214/383 (50 percent to 56 per-
cent) for the three lower doses. Control males by contrast had 30
percent liver tumor frequency (97/328). In the females, no effect
was seen at 0.3 and 3.0 mg/kg/day, but at the higher dose levels,
tumor rates were significantly increased at 7.5 mg/kg/day (43/328)
and 37.5 mg/kg/day (192/293). Liver tumors appeared earlier in the
F, through Fr generations than in the parental at higher dosages,
but tumor incidence did not show consistent increases with consecu-
tive generations as previously reported in BALB/C mice (Tarjan and
Kemeny, 1969).
Comparable lifetime studies were performed by Shabad, et al.
(1973) in A-strain mice. Technical DDT was given via gavage in
daily dosages of 1.5 and 7.5 mg/kg/day for the parent lifetime and
10 mg/kg/day for consecutive generations, F^ through F^. Dosing
with DDT in 0.1 ml sunflower oil began at 6 to 8 weeks of age for
each generation. Strain A, which is susceptible to spontaneous
lung adenomas, had an overall incidence of 7 percent in the control
group. The parental generation, which received the highest dose,
showed 37 percent incidence of lung adenomas. The frequencies of
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lung tumor formation in parents and generations up to F,- treated at
1.5 mg/kg/day were 19, 15, 24, 46, 43, and 13 percent, respective-
ly. Animals dying prior to six months in all of the control, paren-
tal and F, treated groups showed no tumors, whereas earlier appear-
ance of tumors in treated F~ to FS was seen in animals dying prior to
six months. No other tumors, including liver tumors, were detected.
A third multigeneration study on mice was performed by Terra-
cini, et al. (1973). Three dose levels of technical DDT in the diet
corresponding to 0.3, 3.0, and 37.5 mg/kg/day of DDT was admin-
istered to two separate colonies of BALB/C mice, beginning at 4 to
5 weeks of age, for their lifespan. The liver was the only target
organ to show significant increases in the proportion of animals
bearing tumors. Both males and females showed higher percentages
of tumors at the 37.5 mg/kg/day level, with no excess tumorigeni-
city at 0.3 and 3.0 mg/kg/day. Liver tumors were present in 28/63
of the female parents and 43/58 of the first generation females, at
the high dose only. Both colonies of mice showed identical results
at this dosage. Incidence of malignant lymphomas was approximately
50 percent in the control, 0.3 or 3.0 mg/kg/day treated mice. At
highest dosages, this incidence fell to 14 percent in one colony
and 36 percent in the other. The incidence of lung adenomas was not
affected by DDT treatment.
In order to determine if the liver tumors of mice would pro-
gress or regress after cessation of dosing, Tomatis, et al. (1974)
treated CF^ mice with dietary DDT of 37.5 mg/kg/day for 15 or 30
weeks. Autopsies were performed at 65, 95, and 120 weeks from the
beginning of the experiment. The data indicated that a limited
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period of exposure to 37.5 mg/kg/day results in an increased and
early appearance of hepatomas, similar to that caused by lifespan
exposure. The shorter the period of exposure, the lower the inci-
dence of liver tumors. In males treated for 15 weeks and killed at
65, 95, and 120 weeks, the incidence of hepatomas was 13/60, 25/60,
and 25/60, respectively. In males treated for 30 weeks the cor-
responding values were 38/60, 41/60, and 37/60, whereas the values
for the controls in the same periods were 12/70, 24/83, and 33/98.
In females, the incidence of hepatomas increased from the 65th to
the 120th week. Those treated for 15 weeks showed 3/60, 11/60, and
5/60 after 65, 90, and 120 weeks, respectively; the corresponding
values for the 30-week treated mice were: 4/54, 11/65, and 11/54;
control values were: 0/69, 0/72, and 1/90.
The size and multiplicity of the hepatomas were also corre-
lated with the duration of exposure and time of autopsy. In this
study, as in the mouse studies previously cited, the histology of
the hepatomas rarely shows signs of metastases and local invasive-
ness.
Further confirmation of the tumor igenicity of DDT to mouse
livers was reported by Walker, et al. (1972) and by Thorpe and
Walker (1973) in CF, strains. Incidences of tumors increased from
13 percent in controls to 37 percent at 7.5 mg/kg/day and 53 per-
cent at 15 mg/kg/day with slightly higher increases in females
(control, 17 percent; 15 mg/kg/day, 76 percent). In the second
study over 26 months, Thorpe and Walker (1973) reported that the
control values for both males and females were approximately 23
percent and rose to 77 percent for males and 87 percent for females
C-52
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when fed 15 mg/kg/day in the diet. In contrast to the considerable
shortening of lifespan seen in all previous mouse studies, minimal
reduction was observed in this study.
Lifespan studies of the effect of chronic exposure to the
metabolites DDE and DDD at 37.5 mg/kg/day in the diet and a mixture
of 18.75 mg/kg/day each have been reported (Tomatis, et al. 1974).
DDE showed marked effects in female CF, mice on liver tumors in-
creasing from 1 percent (1/90) to 98 percent (54/55) in control
versus treated; male incidence rose from 34 (33/98) to 74 percent
(39/53). DDD showed slight increases in males only, but lung ade-
nomas were markedly increased in both sexes. Control values for
lung adenomas were 54 and 41 percent for males and females, respec-
tively. Treatment with DDD plus DDE or DDE only showed a decrease to
approximately 15 percent of female mice with lung tumors. DDE
reduced incidence in males to 36 percent, but continued treatment
had no further effect. The combination of DDD and DDE increased
hepatoma incidence in both sexes to approximately 75 percent.
Since the most significant evidence implicating DDT as a pos-
sible carcinogen to date has been the formation of hepatic tumors
in the mouse, some criticism of the use of this model with high dos-
ages has been expressed (Deichmann, 1972). The use of animals with
high spontaneous rate of tumor formation confers an added sensitiv-
ity if increases are found following exposure. The use of animal
models with none or low spontaneous tumor incidences may be more
indicative of actual risk.
Breslow, et al. (1974) reviewed the multigeneration studies by
the IARC group to determine associations between tumor types
C-53
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following DDT exposure. A negative correlation was seen between
lymphomas and lung, mammary, and ovarian tumors, possibly due to
competing risk mortality of the diseases. Despite some spurious
results caused by grouping of animals, or age specific tumor preva-
lence, significant associations remained. Positive association
between lymphoma and bone tumor formation could be a reflection of
viral factors. Viruses isolated from some tumors of CF, mice have
produced tumors in neonate mice. Hepatoma formation was less af-
fected by lymphoma mortality. Histological examination of liver
tumors in the CF, mice showed that this hepatoblastoma is similar
in morphological resemblance to human hepatoblastoma. These tumors
were found in association with the ordinary type of hepatoma and
isolated primarily from older animals. The hepatoblastoma proved
to be more highly malignant than the hepatoma, with metastases
occurring in 10 to 20 percent versus 1 to 2 percent for hepatomas.
A progression from hyperplasia to neoplasia can occur spontaneously
with age in mice. The phenomena of induction of hyperplasia could
be attributable to age and spontaneous tumor formation or associ-
ated with early induction by DDT activity.
One other positive report on the possible carcinogenicity of
DDT in other species should be noted. Halver, et al. (1962) have
observed an increase in evidence of hepatomas in rainbow trout
being raised for lake stock. Following determinations of toxicity
in rodents, dose fractions or multiples of one-sixteenth, one-
fourth, 1, 4, and 16 times were fed in a synthetic diet of caseine
gelatin, minerals, etc. High doses of DDT, 2-AAF, carbon tet-
rachloride, and other substances exhibited toxic effects.
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Histopathologically confirmed hepatomas appeared in the inter-
mediate levels of DDT, DBS, and DMN. In a parallel study of fatty
extracts from commercial ratios fed to fish, fish developed tumors
also histologically resembling mammalian hepatorna.
In contrast to the positive results found in the rat, mouse,
and fish studies previously cited, a number of other studies have
shown no significant increase in tumor formation following DDT
exposure. Lifetime feeding studies with Syrian golden hamsters at
75 and 150 mg/kg/day DDT were conducted by Agthe, et al. (1970). No
increases in tumor incidences were observed, although there was a
slight decrease in survival in both males and females.
A number of negative studies have been reported for various
rat strains. Cameron and Cheng (1951) gave daily doses of 0.36,
3.6, and 36 mg/kg in oil for up to 63 weeks. Of the characteristic
lesions described by Fitzhugh and Nelson (1947) and Laug, et al.
(1950), only two female rats showed the centrolobular necrosis, and
no significant differences in the extent of the other pathological
changes could be made between treated and untreated groups.
Two long-term feeding studies utilizing Osborne-Mendel rats
have shown no significant tumorigenic response to three dosage lev-
els of DDT. In the first (Radomski, et al. 1965), DDT was fed at
7.5 and 12 mg/kg/day in the diet for two years. At 7.5 mg/kg/day, a
slight, but not significant increase in hepatic tumor was noted; at
12 mg/kg/day no liver tumors were noted, and no differences were
found between control and treated rats in tumors of other sites.
In addition, DDT was fed in a mixture with 12 mg/kg/day each of
C-55
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aramite, methoxychlor, and thiourea for two years, and no additive
or synergistic effect for tumor formation was found.
In a similar fashion, Deichmann, et al. (1967) repeated these
studies with a higher dosage of DDT - approximately 10 mg/kg/day
for 27 months (200 ppra in the diet). Despite the fact that the
treated animals displayed increased liver weights and the charac-
teristic liver pathology, actual tumor incidence in DDT-fed rats
was less than in the control. The majority of tumors were mammary
tumors in both control and treated animals. Liver tumors were
found only in rats fed DDT, aramite, or a mixture of these plus
methoxychlor and thiourea. Mixtures of these tumorigens also had
no significant effect in tumor incidence.
In order to determine the effect of diet and DDT on the devel-
opment of leukemia, Kimbrough, et al. (1964) fed rats purified high
fat, purified normal fat, and normal diets with and without DDT,
for varying time periods. Of the seven animals developing leuke-
mia, four were on the high fat diet, two were on purified high fat
and 35 mg/day pp'-DDT, and one was on normal fat diet and DDT. No
animals fed DDT and normal ratios developed leukemias. The authors
concluded that leukemic development in Sherman rats was a conse-
quence of diet and unrelated to DDT treatment.
Weisburger and Weisburger (1968) fed weanling Fisher rats 10
mg DDT/day (30-100 mg/kg/day) by gavage and found no liver tumors
nor evident hepatotoxicity. In combination with 0.1 mg/day 2-AAF,
hepatoma incidence increased from 67 to 90 percent in males and 7
to 33 percent in females compared to treatment with 2-AAF alone.
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Rossi, et al. (1977) were able to induce noninvasive nodular
liver tumors in Wistar rats by administering in their diet approxi-
mately 35 mg/kg/day of either technical DDT or sodium phenobarbi-
tal. None of the tumors were metastatic, and extrahepatic tumors
were slightly higher in controls than in treated animals. For DDT,
liver tumor incidences of 45 percent (24 of 53 animals) were ob-
served in treated rats while controls exhibited no liver tumors.
Interestingly, sodium phenobarbital at the same dosage level showed
a similar hisopathologic liver change in 44 percent (22/50) of the
rats. A compilation of long-term tumorigenicity studies in rats is
given in Table 9.
In a recently published report of the National Cancer Insti-
tute (NCI, 1978), bioassays of DDT, DDD, and DDE were conducted in
male and female Osborne-Mendel rats and B6C3F, mice by long-term
feeding. Approximately 50 animals of each sex were treated and 20
animals of each sex served as controls. The dosing period consist-
ed of 78 weeks in which there were dosage changes during the course
of the study, and dosing was reported as time-weighted averages.
High and low dietary concentrations of DDT were? 32.1 and 16.05
mg/kg/day for male rats, 21.0 and 10.5 for females; for DDD, males
were fed 164.7 and 82.4 mg/kg/day and females 85.0 and 42.5
^/kg/day. For DDE, males were fed 41.95 and 21.85 mg/kg/day and
females 23.1 and 12.1 mg/kg/day. Increased mortality was seen in
both sexes of rats dosed with DDE. No evidence of carcinogenicity
was found for DDT or DDE in either sex at the given doses. DDD had
no carcinogenic effects in the females, but in the males receiving
a low dose, a significant increase in the follicular cell adenomas
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TABLE 9
Long Term Tumorigenicity Studies in Rats
Dose Range
mg/ kg/day
5-40
0.36-36.0
0.12-1.2
1-2
7.5-12
10
30-100
35
10-32
Route of stratn
administration b"«ln
In diet
In oil by
gavage
In diet
In diet
In diet
In diet
In diet
In diet
In diet
Osborne-
Mendel
Os borne-
He ndel
Carvorth
Sherman
Os borne-
Mendel
Osborne-
Hendel
Fischer
Hictar
Osborne-
Mendel
Duration
2 years
63 weeks
2 years
Variable
2 years
2.25 yrs
1 year
2.9 yrs
78 weeks
Results
Increase in liver tumors at
unspecified close.
No effect.
No effect.
No increase in leukemia
incidence.
12 mg/kg/day. No effect.
Slight increase liver tumor
Incidence at 7.5 mg/kg/day.
No effect.
No effect.
Liver tumors in 45%
of animals.
DOT and DDE -
No significant tumor incidences
DDD - Increased thyroid tumors.
Reference
Fitzhugh and Nelson
(1947)
Cameron and Cheng
(1951)
Treon and Cleveland
(1955)
Kimbrougli, et al. (1964)
Radomski, et al. (1965)
Deichmann, et al. (1967)
Welsburger and
Weisburger (1968)
Rossi, et al. (1977)
NCI (1978)
C-58
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and carcinomas of the thyroid was observed. Because of high varia-
tion of thyroid lesions in control male rats, these findings are
considered only suggestive of a chemical related effect. Among
dosed rats no significant increases in other neoplasms were seen as
compared to controls. Administration of DDE did not result in sig-
nificant incidences of liver tumors, but the compound was hepato-
toxic, inducing centrolobular necrosis and fatty metamorphosis.
Time-weighted average high and low dietary concentrations of
DDT for the mice were; 6.6 and 3.3 mg/kg/day for male mice, and
26.25 and 13.05 mg/kg/day for female mice; high and low average
doses of DDD were 123.3 and 61.65 mg/kg/day for male and female
mice; and average high and low doses of DDE were 39.15 mg/kg/day
and 22.2 mg/kg/day for male and female mice. Significant positive
associations between increased doses and greater mortality in fe-
male mice dosed with DDT and DDE were observed. Poor survival was
seen in control and dosed male mice in the bioassays of DDT and DDE.
The only neoplasms occurring in statistically significant increased
incidence were hepatocellular carcinomas among groups receiving
DDE. The incidences of these tumors in control low-dosed and high-
dosed males were 0/19, 7/41 (17 percent), and 17/47 (36 percent),
respectively. Corresponding figures for females were 0/19, 19/47
(40 percent), and 34/48 (71 percent).
The National Cancer Insitiute (NCI) study presented no evi-
dence for the carcinogenicity of DDT in rats and mice, of DDD in
female rats or mice of either sex, or of p,p'-DDE in rats although
hepatotoxicity was evident. A possible carcinogenic effect of DDD
in inducing follicular cell tumors of the thyroid of male rats was
C-59
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suggested. DDE was carcinogenic in B6C3F, mice, causing hepato-
cellular carcinomas in both sexes (NCI, 1978).
Durham, et al. (1963) found no liver pathology in Rhesus mon-
keys fed 100 mg/kg/day or less for up to 1% years. Monkeys dosed at
2,500 mg/kg/day had cytoplasmic inclusions and necrosis in the
liver and brain pathology. These animals died in less than six
months from DDT poisoning.
There is evidence that DDT is an inhibitor of tumor takes in
transplant. Mice exposed to 5.5 mg/kg/day in the diet were sub-
jected to experimental transplantation of an ependymona. Compared
to controls, treated animals were less susceptible to tumor trans-
plantation and had increased longevity upon implantation (Laws,
1971).
In summary, the evidence for carcinogenicity of DDT in labora-
tory animals has been demonstrated only for the mouse in the pro-
duction of liver tumors. In several other species, such as the
rat, monkey, and hamster, no tumorigenic effect for DDT has been
shown at doses less than 50 mg/kg. At doses higher than that level,
evidence is equivocal for the rat (Fitzhugh and Nelson, 1947;
Radom ski, et al. 1965; Deichmann, et al. 1967; NCI, 1978).
The epidemiological studies in man cannot be considered con-
clusive in view of the small number of individuals studied.
Ortelee (1958) reported on a group of 40 men with extensive and
prolonged occupational exposure to DDT in manufacturing or formu-
lating plants. An exposure rate was given to each individual based
on observation on the job. The highest exposure rate was estimated
to be absorbed doses of approximately 42 mg/man/day. With the
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exception of minor skin irritations, physical, neurological, and
laboratory findings were within normal ranges, and no correlation
between DDT exposure and frequency and distribution of the few
abnormalities were seen. Laws, et al. (1967) found no evidence of
adverse health effects in 35 men with 11 to 19 years of high occupa-
tional exposure (3.6 to 18 mg/man/day). No case of cancer was
found.
Almeida, et al. (1975) have conducted a surveillance of work-
ers exposed to DDT for six or more years as spray men in a malaria
eradication campaign in Brazil. Although significant increases in
DDT and DDE residues in the blood serum levels were observed, phys-
ical examination showed no significant increases in adverse health
effects for the exposed versus control groups.
Edmundson, et al. (1969a) studied 154 individuals with occupa-
tional exposure to DDT and observed significant differences asso-
ciated with race and type of occupation. Nonwhite formulators and
agricultural sprayers showed greatly elevated serum concentra-
tions, but during the 2-year time of study no clinical effects re-
lated to DDT exposure were observed.
Hayes, et al. (1971) administered doses up to 35 mg/man/day to
volunteers for 21.5 months. Liver function studies of SCOT, plasma
cholinesterase, and BSP retention exhibited no significant change
from normal for these volunteers. A number of other health para-
meters were studied and no definite chemical or laboratory evidence
of injury by DDT was found at the prevailing levels of intake. This
led the authors to conclude that DDT had a considerable degree of
safety for the general population.
-------
Several authors have examined the storage of DDT in persons
with various diseases. Maier-Bode (1960) found no differences in
storage of DDT or DDE in 21 persons who died of cancer and 39 others
who died of other diseases.
The difficulty in making these kinds of associations is illus-
trated by the results of Radomski, et al. (1968). Pesticide con-
centrations in fat and liver were determined at autopsy for 271
patients previously exhibiting various pathology of liver, brain,
and other tissues. Another group that previously had infectious
diseases was examined. High significant elevations of DDT and DDE
were found in carcinomas of varying tissues. Fat concentrations of
DDE, DDT, ODD, and dieldrin were consistently elevated in cases of
hypertension. These observations were clouded by the great indivi-
dual variability of pesticide levels regardless of the disease
category.
Two further studies (Hoffman, et al. 1967; Casarett, et al.
1968) have been conducted on the levels of DDT in tissues of pa-
tients with cancer and other chronic diseases. One showed higher
DDT residues in cancer patients (Casarett, et al. 1968). No con-
clusions can be made from these studies as to a possible causal
relationship.
Sanchez-Medal, et al. (1963) noted 20 cases of aplastic anemia
over an 8-year period in a Mexico City Hospital. In 16 out of 20
cases, the patients had repeated contact with pesticides during the
prior six months. Insecticides implicated were DDT alone or DDT in
association with lindane, dieldrin, or DDVP. One 13-year-old boy
had been exposed repeatedly to DDT alone for two years and exposure
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was intensified to every other day in the prior four months. He was
accidentally exposed to 10 percent DDT spray in the hospital and
died 30 hours later due to a worsening blood discrasia. The Ameri-
can Medical Association Registry on Blood Discrasia reported 44
cases of aplastic anemia associated with pesticide exposure through
1963. Of these cases, 19 were related to DDT, and in three, DDT was
the sole agent (Erslev, 1964).
At the present time, no evidence of neoplasia has been found
in the studies performed in occupationally exposed or dosed volun-
teer subjects. Medical histories have been essentially normal.
However, these studies do not constitute an adequate basis to make
conclusions regarding human carcinogenicity because of small sample
size and short duration in terms of average human life span.
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CRITERION FORMULATION
Existing Guidelines and Standards
In 1958, the U.S. Department of Agriculture (USDA) began to
phase out the use of DDT in insect control programs. Spraying was
reduced from 4.9 million acres in 1957 to just over 100,000 acres
in 1967, and DDT was used as a persistent pesticide thereafter only
in the absence of an effective alternative. In 1964, the Secretary
of Interior issued a directive that use of chlorinated hydrocarbons
should be avoided in interior lands. This was extended in 1970,
when 16 pesticides, including DDT, were completely bann*3 for use
on Department of Interior lands. By 1969, DDT registration and
usage was curtailed by the USDA in various areas of the cooperative
Federal State pest control program. In November 1969, the USDA
announced its intention to discontinue all uses of DDT nonessential
to human health and for which there were safe and effective substi-
tutes. In 1970, the USDA cancelled Federal registrations of DDT
products for use on 50 food crops, domestic animals, finished wood
and lumber products, and use around commercial, institutional, and
industrial establishments.
Major responsibility for Federal regulation of pesticides
under the Federal Insecticide, Fungicide, and Rodenticide Act
(1947} was transferred to the U.S. EPA. In January, 1971, U.S. EPA
issued notices of intent to cancel all remaining Federal registra-
tions of products containing DDT. A hearing on the cancellation of
Federal registration of products containing DDT was held beginning
in August, 1971 and concluding in March, 1972. The principal par-
ties to the hearing were 31 DDT formulating companies, the USDA,
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the Environmental Defense Fund, and the U.S. EPA. This hearing and
other evidence from four Government reports including the December
1969 Mrak Commission Report were instrumental in the final cancel-
lation of all remaining crop usages of DDT in the U.S., effective
December 31, 1972. During the same period {'October 1972), a Fed-
eral Environmental Pesticide Control Act (FEPCA) was enacted which
provided EPA with more effective pesticide regulation mechanisms.
The cancellation order was appealed by the pesticides industry in
several U.S. courts. On December 13, 1973, the U.S. Court of
Appeals for the District of Columbia ruled there was substantial
evidence in the record to support the U.S. EPA ban on DDT. In April
1973, the U.S. EPA, in accordance with authority granted by FEPCA,
required that all products containing DDT be registered with the
Agency by June 10, 1973. Since that time, the U.S. EPA has granted
requests to the states of Washington and Idaho and to the Forest
Service to use DDT on the basis of economic emergency and no effec-
tive alternative to DDT being available.
Authority to regulate hazards arising from the manufacturing
and formulation of pesticides and other chemicals resides with the
Occupational Safety and Health Administration (OSHA). Under the
terms of the Occupational Safety and Health Act of 1970, the
National Institute for Occupational Safety and Health (NIOSH) has
been responsible for setting guidelines and criteria for occupa-
tional exposure. The OSHA exposure limit for DDT on skin has been
2
set a 1.0 mg/m . Further, DDT has been classified as a suspected
occupational carcinogen that should be cautiously handled in the
workplace.
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The decision to ban DDT was extensively reviewed relative to
scientific and economic aspects in 1975 (U.S. EPA, 1975). No new
evidence was found contradicting the original finding of the Admin-
istrator in 1972 (Table 10).
Current Levels of Exposure
Most of the reported DDT concentrations in air are associated
with high usage of DDT prior to 1972. Stanley, et al. (1971) ana-
lyzed air samples from nine localities. DDT levels ranged from 0.1
ng/m to 20 ng/m . Air samples collected in July 1970 over the
Atlantic Ocean had 0.00007 ng/m (Prospero and Seba, 1972). The
actual levels of DDT in the ambient air at the present time are dif-
ficult to estimate but are probably at the lowest ranges of Stan-
ley's estimates. The ambient air levels of DDT might be below
levels that might add significantly to the total human intake
(Spencer, 1975).
Kenaga (1972) gave the following relative values for residues
for DDT and its metabolites found in various types of waters: rain
water, 0.2 ug/1; fresh water, 0.02 P9/1? and s®a water, 0.001 jug/1.
Assuming average daily intake of water to be 2 liters in any given
year, the maximal DDT intake from water would be 0.015 mg. This
figure is approximately twice the estimated daily dietary intake of
DDT for a 19-year-old male (U.S. EPA, 1975). Therefore, it is con-
cluded that DDT intake from potable water does not contribute sig-
nificantly to the overall exposure.
Duggan and Corneliussen (1972) calculated the average daily
intake of total DDT residues in 1965 as 0.0009 mg/kg and decreasing to
0.0004 mg/kg in 1970. Market basket studies have shown significant
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TABLE 10
History of DDT Standard or Recommendation
Year
1971
1976
1977
Aqency
WHO
U.S. EPA
Natl. Acad. Sci. ,
Standard
0.005 mg/kg
body weight
0.001 ug/1
-
Remarks
Maximum Acceptable Daily
i n food
Intake
Quality Criteria for Water
In light of carcinogenic
risk
1978
1978
Natl. Res. Counc.
Occup. Safety
Health Admin.
(NIOSH, 1978)
U.S. EPA
(40 FR 17116)
1 mg/m'
0.41 ug/1
0.00023 ug/1
projection, suggested strict
criteria for DDT and DDE in
drinking water
Skin exposure
Final acute and chronic values
for water quality criteria for
protection of aquatic life
(fresh water)
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declines between 1970 and 1973 of DDT and DDD residues of 86 and 89
percent, respectively. DDE decreased by 25 percent over this per-
iod of time. Dairy, meat, fish, and poultry constitute 95 percent
of the total ingested DDT sources with dairy products contributing
30 percent of this amount. Average human fat storage for the time
period of 1970 to 1973 has decreased from approximately 8 ppm to 6
ppm in the U.S. population. Based on these declines and the most
current intake figures as of 1973, it is estimated that current
levels of dietary intake are approximately 0.0001 mg/kg/day, with
DDE comprising over 80 percent of this amount. Assuming the aver-
age male weighs 70 kg, the average daily intake would be 0.007
mg/day or 2.56 mg/year.
Human exposure to DDT is primarily by ingestion of contami-
nated food. Air and water intake is negligible and amounts to
probably less than 0.01 mg/year. Therefore, by estimation, total
intake of DDT per year for the average U.S. resident will be less
than 3 mg/year.
Special Groups at Risk
The entire population of the U.S. has some low level exposure
to dietary contaminants. Minimal exposure from air and water
sources, however, may be more important in previously heavily
sprayed agricultural areas, where large amounts of residues may
still be present.
In 1975, estimated DDT production was 30 to 49 million pounds
(NIOSH, 1978). Groups at special risk are workmen in manufacturing
and formulating plants, applicators, handlers, and sprayers. Dur-
ing such times when exceptions are granted by the U.S. EPA for crop
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usage or during use for public health measures, those involved in
handling or applying DDT may have considerable exposure.
Estimating the number of individuals at high risk due to occu-
pational exposure is difficult. It is estimated that 8,700 workers
are involved in formulating or manufacturing all pesticides. Since
DDT constitutes much less than 10 percent of the total, the maximal
number of exposed workers would be approximately 500. Since usage
of DDT is severely limited, persons exposed by application would
probably be fewer.
Basis and Derivation of Criteria
Since no epidemiological evidence for the carcinogenicity of
DDT in man has been reported, the results of animal carcinogenicity
studies conducted by feeding DDT or its metabolites over the life
span of the animal are regarded as the most pertinent data. Al-
though a number of studies have been reported for various species,
the major evidence for the tumorigenicity of DDT is its ability to
induce liver tumors in mice.
Under the Consent Decree in NRDC v. Train, criteria are to
state "recommended maximum permissible concentrations (including
where appropriate, zero) consistent with the protection of aquatic
organisms, human health, and recreational activities." DDT is sus-
pected of being a human carcinogen. Because there is no recognized
safe concentration for a human carcinogen, the recommended concen-
tration of DDT 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
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possible future development of water quality regulations, the con-
centrations of DDT corresponding to several incremental lifetime
cancer risk levels have been estimated. A cancer risk level pro-
vides an estimate of the additional incidence of cancer that may be
expected in an exposed population. A risk of 10 for example,
indicates 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 ambi-
ent water quality criteria, EPA stated that it is considering set-
ting criteria at an interim target risk level of 10 , 10 or 10~
as shown in the following table.
Exposure Assumptions Risk Levels and Corresponding Criteria^
(per day) 1Q-7 1Q-6 1Q-5
2 liters of drinking 0.0024 ng/1 0.024 ng/1 0.24 ng/1
water and consumption
of 6.5 grams of fish
and shellfish (2)
Consumption of fish 0.0024 ng/1 0.024 ng/1 0.24 ng/1
and shellfish only.
(1) Calculated by applying a linearized multistage model as dis-
cussed in the Human Health Methodology Appendices to the October
1980 Federal Register notice which announced the availability of
this document to the animal bioassay data presented in
Appendix I. Since the extrapolation model is linear at low
doses, the additional lifetime risk is directly proportional
to the water concentration. Therefore, water concentrations
corresponding to other risk levels can be derived by multiply-
ing or dividing one of the risk levels and corresponding water
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concentrations shown in the table by factors such as 10, 100,
1,000, and so forth.
(2) Greater than 99 percent of the DDT exposure results from the
consumption of aquatic organisms which exhibit an average bio-
concentration potential of 53,600-fold. The remaining less
than one percent of DDT exposure results from drinking water.
Concentration levels were derived assuming a lifetime exposure
to various amounts of DDT (1) occurring from the consumption of
both drinking water and aquatic life grown in water containing the
corresponding DDT concentrations, and (2) occurring solely from the
consumption of aquatic life grown in the waters containing the cor-
responding DDT concentrations. Although total exposure information
for DDT is discussed and an estimate of the contributions from
other sources of exposure can be made, this data will not be fac-
tored 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.
The case of DDT and its possible role as a human carcinogen is
complicated by several factors. Despite widespread use and expo-
sure over 30 years, no positive associations with human cancer have
been found to date, although the number of individuals studied is
not statistically large. It is a chemical with high efficacy and
has been extremely effective all over the world for public health
measures. However, its slow biodegradability and propensity to
accumulate in nontarget species have made it particularly hazardous
for many fish and bird species. For mammals, however, it has a low
acute toxicity as compared to other alternate pesticides.
C-71
-------
DDT has not been shown to produce point mutations or terato-
genic effects in a wide battery of tests. Some evidence for its
clastogenic properties/ however, make it suspect. The primary evi-
dence for the carcinogenicity of DDT and metabolites to date has
been the induction of liver tumors in mice. Studies in other spe-
cies have shown negative or inconsistent effects. The evidence for
the carcinogenicity of DDT would be much more convincing if tumor-
igenicity in other species or at other sites could be conclusively
demonstrated. This is in light of the fact that DDT has been proba-
bly the most extensively studied compound in modern science.
An alternative level based on toxicity data was calculated for
comparison as suggested in public comments. The Effects Section
of this document discusses several adequate chronic bioassays on
which to base this derivation. The Laug, et al. (1950) study was
chosen because: (1) male rats appear to be the most sensitive ani-
mals to DDT exposure; and (2) the study was of sufficient legnth to
observe toxic effects (approximately 27 weeks); and three, several
doses were administered in the diet over the range of the dose-
response curve. The highest no-observable-adverse-effeet level
(NOAEL) in this study was 1 ppm. An ADI can be determined for man
from this dose by the following calculations:
1 mg/kg of diet^l pp«) x 0.05 = Q<143 mg/kg/df
where 1 mg/kg is the highest NOAEL, 0.05 is the fraction of body
weight that a rat is assumed to eat of diet per day, and 0.350 kg is
the assumed weight of a rat.
ADI - 0.143 mg/kg/d x 70 kg = 1>Q
10
C-72
-------
where 70 kg is the average body weight of man and 10 represents an
uncertainty factor used because of the available data on human
exposure and other adequate chronic animal bioassays, as per Na-
tional Academy of Sciences guidelines (NAS/ 1977).
The ambient water quality concentration for DDT corresponding
to this ADI is:
Concentration = - - * 2.85
(2 1/d + 0.0065 kg/d x 53,600 I/kg)
Current levels of exposure would seem to pose extremely small
risk to persons in the U.S. In addition, DDT and DDE are preferen-
tially stored in fatty compartments that are not actively dividing,
suggesting less carcinogenic risk. However, the use of DDT has
been restricted in several countries because of its impact on the
environment and its tumor igenic effect in mice. This seems to be
reasonable based on numerous reports.
Therefore, the Agency recommends that the criterion for DDT to
be derived from the carcinogenic response in mice in the Tarjan and
Kemeny (1969) study. The criterion associated with a human life-
time cancer risk of 10 is 0.24 ng/1.
C-73
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APPENDIX I
Summary and Conclusions Regarding the
Carcinogenicity of DDT*
DDT is a synthetic, chlorinated hydrocarbon insecticide which
has broad-spectrum insecticidal activity. DDT residues have been
detected in a wide variety of fruits, vegetables, meat, fish, and
poultry, and will probably continue to be present in agricultural
produce indefinitely as a consequence of the persistence of DDT in
soil. DDT is absorbed completely after inhalation and ingestion
and absorbed poorly through skin. DDT has not been found to be
mutagenic in bacterial test systems, either with or without meta-
bolic activation. The evidence from mammalian test systems in
vitro and in vivo is inconclusive.
There is no epidemiological evidence relating to the carcino-
genicity of DDT, but there are a number of carcinogenicity studies
conducted by feeding DDT to animals. A number of chronic studies
have been reported in various species, but the major evidences for
tumorigenicity in mice and rats are described below. In mice, DDT
increased tumor incidence significantly in experimental groups as
compared to controls in liver (Innes, et al. 1969; Walker, et al.
1972; Turusov, et al. 1973; Terracini, et al. 1973; Thorpe and
Walker, 1973), lungs (Tarjan and Kemeny, 1969; Shabad, et al. 1963)
and lymphoreticular tissue tumors (Innes, et al. 1969; Tarjan and
Kemeny, 1969). In rats, liver tumors were significantly increased
in the experimental group as compared to controls in two studies
(Fitzhugh and Nelson, 1947; Rossi, et al. 1977).
*Thj.s summary has been prepared and approved by the Carcinogens
Assessment Group of EPA on June 20, 1980.
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The negative NCI mouse study might be explained on the basis
of shorter duration of exposure, low dose in male mice, and the use
of a strain different from the other positive studies. The nega-
tive NCI rat study might be explained on the basis of shorter dura-
tion of exposure and lower dose compared to that used in the Fitz-
hugh study. There are other negative carcinogenicity studies in
mice, rats, hamsters, dogs, and monkeys.
The water quality criterion for DDT is based on a six-genera-
tion study in CF, mice by Tarjan and Keraeny (1969). It is concluded
that if water alone is consumed, the water concentration should be
less than 42 ng/1 in order to keep the lifetime cancer risk below
10 . If fish and water are consumed, the water concentration
should be less than 0.24 ng/1 to achieve the same risk level.
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Summary of Pertinent for DDT
Of the five positive carcinogenicity studies with DDT, the one
showing the most potent response is the male mice group in the
Turusov study. This study, however, is not used for the water
quality criteria because the dose response curve was flat down to
the lowest dose tested, and the background rate of tumors was
abnormally large.
Instead, the data of Tarjan and Kemeny (1969) was used. Five
generations of mice were fed dietary DDT with an equivalent intake
of 0.55 mg/kg/day. Tumors were found in excess of controls in each
generation beyond the second. They were widely distributed in sev-
eral sites and consisted of adenocarcinomas as well as several
types of carcinomas. The parameters of the calculation are:
Dose Incidence
(mg/kg/day) (no. responding/no, tested)
0.0 13/406
0.55 196/683
le * 26 months1 w = 0.030 kg
Le » 26 months R * 53,600 I/kg
L * 26 months
With these parameters, the carcinogenic potency factor, q-i*/
for humans is 8.422 (mg/kg/day) . The result is that if fish and
water are consumed the water concentration should be less than 0.24
ng/1 in order to keep the individual lifetime risk below 10~ . If
only water were consumed, the corresponding concentration is 42
ng/1.
There was some confusion in the original article over the length
of DDT exposure. In a subsequent publication Tarjan clearly
stated that DDT exposure was from weaning to death (Fd. Cosmet.
Toxicol., August 1970, p. 478).
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