COLLECTED REPRINTS
1975 - 1976
ENVIRONMENTAL RESEARCH LABORATORY
GULF BREEZE
U.S. ENVIRONMENTAL PROTECTION AGENCY
OFFICE OF RESEARCH & DEVELOPMENT
ENVIRONMENTAL RESEARCH LABORATORY
GULF BREEZE, FLORIDA 32561
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COLLECTED REPRINTS
1975 - 1976
ENVIRONMENTAL RESEARCH LABORATORY
GULF BREEZE
U. S. ENVIRONMENTAL PROTECTION AGENCY
OFFICE OF RESEARCH & DEVELOPMENT
ENVIRONMENTAL RESEARCH LABORATORY
GULF BREEZE, FLORIDA 32561
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United States Environmental Research
Environmental Protection Laboratory
Agency Gulf Breeze FL 32561
Research and Development
&EPA REPRINTS
Gulf Breeze
Laboratory
1975 - 1976
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DISCLAIMER
These reports have been reviewed by the Gulf Breeze Environmental
Research Laboratory, U.S. Environmental Protection Agency, and approved for
publication. Mention of trade names or commercial products does not con-
stitute endorsement or recommendation for use.
11
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TABLE OF CONTENTS
Contrib. Page
Number Number
COUCH, JOHN A. Histopathological effects of pesti-
cides and related chemicals on the livers of
fishes. In: The Pathology of Fishes, William E.
Ribelin and George Migaki, editors. University
of Wisconsin Press, 1975, pp. 559-584 152 1
NIMMO, DEL WAYNE R., DAVID J. HANSEN, JOHN COUCH,
NELSON R. COOLEY, PATRICK R. PARRISH, AND JACK
I. LOWE. Toxicity of AroclorR 1254 and its
physiological activity in several estuarine
organisms. Arch. Environ. Contain. Toxicol. 1975,
Vol. 3(1), pp. 22-39 162 29
BORTHWICK, PATRICK W., MARLIN E. TAGATZ, AND
JERROLD FORESTER. A gravity-flow column to pro-
vide pesticide-laden water for aquatic bioassays.
Bull. Environ. Contain. Toxicol. 1975, Vol. 13(2),
pp. 183-187 189 49
DUKE, THOMAS W., AND DAVID P. DUMAS. Implications
of pesticide residues in the coastal environment.
In: Pollution and Physiology of Marine Organisms,
F. John Vernberg and Winona B. Vernberg, editors.
Academic Press, Inc. 1974, pp. 137-164 195 57
COPPAGE, DAVID L., AND EDWARD MATTHEWS. Brain-
acetylcholinesterase inhibition in a marine
teleost during lethal and sublethal exposures to
l,2-dibromo-2,2-dichloroethyl dimethyl phosphate
(naled) in seawater. Toxicol. Appl. Pharmacol.
1975, Vol. 31, pp. 128-133 199 87
SCHIMMEL, STEVEN C., AND DAVID J. HANSEN. Sheeps-
head minnow (Cyprinodon variegatus): An estuarine
fish suitable for chronic (entire life-cycle)
bioassays. Proc. 28th Annu. Southeast. Assoc. Game
Fish. Comm. , Nov. 17-20, 1974, pp. 392-398 205 95
ill
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Contrib.
Number Number
HANSEN, DAVID J., STEVEN C. SCHIMMEL, AND JERROLD
FORESTER. Effects of AroclorR 1016 on embryos,
fry, juveniles, and adults of sheepshead minnows
(Cyprinodon variegatus). Trans. Am. Fish. Soc.
1975, Vol. 104(3), pp. 584-588 206 105
SCHOOR, W.P. Problems associated with low-
solubility compounds in aquatic toxicity tests:
Theoretical model and solubility characteristics
of AroclorR 1254 in water. Water Res. 1975, Vol.
9, pp. 937-944 208a 113
BAHNER, LOWELL H., AND DEL WAYNE R. NIMMO. A
salinity controller for flow-through bioassays.
Trans. Am. Fish. Soc. 1975, Vol. 104(2),
pp. 388-389 214 123
BOURQUIN, AL W., AND S. CASSIDY. Effect of poly-
chlorinated biphenyl formulations on the growth
of estuarine bacteria. Appl. Microbiol. 1975,
Vol. 29(1), pp. 125-127 217 127
SCHIMMEL, STEVEN C., PATRICK R. PARRISH, DAVID J.
HANSEN, JAMES M. PATRICK, JR., AND JERROLD
FORESTER. Endrin: Effects on several estuarine
organisms. Proc. 28th Annu. Southeast. Assoc. Game
Fish. Comm. , Nov. 17-20, 1974, pp. 187-194 218 133
TAGATZ, M.E., P.W. BORTHWICK, AND J. FORESTER.
Seasonal effects of leached mirex on selected
estuarine animals. Arch. Environ. Contain. Toxicol.
1975, Vol. 3(3), pp. 371-383 222 143
SCHIMMEL, STEVEN C., AND DAVID J. HANSEN. An auto-
matic brine shrimp feeder for aquatic bioassays.
J. Fish. Res. Board Can. 1975, Vol. 32(2), pp.
314-316 224 159
PARRISH, PATRICK R., GARY H. COOK, AND JAMES M.
PATRICK, JR. Hexachlorobenzene: Effects on
several estuarine animals. Proc. 28th Annu. Conf.
Southeast. Assoc. Game Fish. Comm., Nov. 17-20,
1974, pp. 179-186 226 165
COPPAGE, D.L., AND T.E. BRAIDECH. River
pollution by anticholinesterase agents. Water Res.
1976, Vol. 10(1), pp. 19-24 227 177
IV
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Contrib. Page
Number Number
TAGATZ, M.E., P.W. BORTHWICK, J.M. IVEY, AND
J.KNIGHT. Effects of leached mirex on
experimental communities of estuarine
animals. Arch. Environ. Contam. Toxicol. 1976,
Vol. 4(4), pp. 435-442 229 185
BOURQUIN, A.W., L.A. KIEFER, N.H. BERNER, S. CROW,
AND D.G. AHEARN. Inhibition of estuarine micro-
organisms by polychlorinated biphenyls. Dev. Ind.
Microbiol. 1975, Vol. 16, pp. 256-261 230 195
CROW, S.A., D.G. AHEARN, W.L. COOK, AND A.W.
BOURQUIN. Densities of bacteria and fungi in
coastal surface films as determined by a membrane-
absorption procedure. Limnol. Oceanogr. 1975, Vol.
1, pp. 644-646 232 205
PARRISH, PATRICK R., STEVEN C. SCHIMMEL, DAVID J.
HANSEN, JAMES M. PATRICK, JR., AND JERROLD
FORESTER. Chlordane: Effects on several
estuarine organisms. J. Toxicol. Environ.
Health, 1976, Vol. 1, pp. 485-494 234 211
COPPAGE, DAVID L., EDWARD MATTHEWS, GARY H. COOK,
AND JOHNNIE KNIGHT. Brain acetylcholinesterase
inhibition in fish as a diagnosis of environmental
poisoning by malathion, 0,0-dimethyl S-(l,2,-
dicarbethoxyethyl) phosphorodithioate. Pest. Biochem.
Physiol. 1975, Vol. 5(6), pp. 536-542 237 223
BAHNER, LOWELL H., C.D. CRAFT, AND D.R. NIMMO. A
saltwater flow-through bioassay method with con-
trolled temperature emd salinity. Prog. Fish-Cult.
1975, Vol. 37(3), pp. 126-129 239 233
COUCH, JOHN A. Attempts to increase Baculovirus
prevalence in shrimp by chemical exposure. Prog.
Exp. Tumor Res. 1976, Vol. 20, pp. 304-314 240 239
BAHNER, LOWELL H., AND DEL WAYNE R. NIMMO. Precision
live-feeder for flow-through larval culture or food
chain bioassays. Prog. Fish-Cult. 1976, Vol. 38(1),
pp. 51-52 246 253
HOLLISTER, TERRENCE A., GERALD E. WALSH, AND JERROLD
FORESTER. Mirex and marine unicellular algae:
Accumulation, population growth, and oxygen evolu-
tion. Bull. Environ. Contam. Toxicol. 1975, Vol.
14(6), pp. 753-759 248 257
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Contrib.
Number Number
MIDDAUGH, D.P., AND P.W. LEMPESIS. Laboratory
spawning and rearing of a marine fish, the silver-
side Menidia menidia menidia. Mar. Biol. 1976, Vol.
35(4), pp. 295-300 252 257
COUCH, JOHN A., MAX D. SUMMERS, AND LEE COURTNEY.
Environmental significance of Baculovirus in-
fections in estuarine and marine shrimp. Ann. N.Y.
Acad. Sci. 1975, Vol. 266, pp. 528-536 253 275
CROW, S.A., W.L. COOK, D.G. AHEARN, AND A.W. BOURQUIN.
Microbial populations in coastal surface slicks. In:
Proc. 3rd Int. Biodegr. Symp., J.M. Sharpley and
A.M. Kaplan, editors. Applied Science Publ. Ltd.,
London, 1976, pp. 93-98 254 287
SMITH, N.G., A.W. BOURQUIN, S.A. CROW, AND D.G. AHEARN.
Effect of heptachlor on hexadecane utilization by
selected fungi. Dev. Ind. Microbiol. 1976, Vol. 17,
pp. 331-336 255 295
WILSON, ALFRED J., JR. Effects of suspended material
on measurement of DDT in estuarine water. Bull.
Environ. Contam. Toxicol. 1976, Vol. 15(5), pp.
515-532 258 303
BAHNER, LOWELL H., AND DEL WAYNE R. NIMMO. Methods to
assess effects of combinations of toxicants, salinity,
and temperature of estuarine animals. In: Trace
Substances of Environmental Health-IX. A Symposium.
D.D. Hemphill, editor, 1975, pp. 169-177 259 313
COUCH, JOHN A. Discussions from selected papers from
EPA-USDA working symposium, Bethesda, Maryland. In:
Baculovirus for Insect Pest Control: Safety
Considerations, Max D. Summers et al., editors,
1975, pp. 62 and 111-114 262 325
SCHIMMEL, STEVEN C., JAMES M. PATRICK, JR., AND JERROLD
FORESTER. Heptachlor: Uptake, depuration, retention,
and metabolism by spot, Leiostomus xanthurus. J.
Toxicol. Environ. Health, 1976, Vol. 2, pp. 169-178. . . 264 333
SCHIMMEL, STEVEN C., JAMES M. PATRICK, JR., AND
JERROLD FORESTER. Heptachlor: Toxicity to and uptake
by estuarine organisms. J. Toxicol. Environ. Health,
1976, Vol. 1, pp. 955-965 265 345
VI
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Contrib. Page
Number Number
SCHOOR, W.P., AND S.M. NEWMAN. The effect of mirex
on the burrowing activity of the lugworm, Arenicola
cristata. Trans. Am. Fish. Soc. 1976, Vol. 105(6),
pp. 700-703 268 359
NIMMO, DEL WAYNE R., AND LOWELL H. BAHNER. Metals,
pesticides, and PCB's: Toxicities to shrimp singly
and in combination. Estuarine Processes, Vol. 1,
Uses, Stresses, and Adaptation to the Estuary.,
Martin W. Wiley, editor, 1976, pp. 523-531 271 365
COOK, GARY H., AND JAMES C. MOORE. Determination
of malathion, malaoxon, and mono- and dicarboxylic
acids of malathion in fish, oyster, and shrimp
tissue. J. Agric. Food Chem. 1976, Vol 24(3),
pp. 631-634 273 377
COOK, GARY H., JAMES C. MOORE, AND DAVID L. COPPAGE.
The relationship of malathion and its metabolites
to fish poisoning. Bull. Environ. Contain. Toxicol.
1976, Vol. 16(3), pp. 283-290 275 383
TAGATZ, MARLIN E. Effect of mirex on predator-prey
interaction in an experimental estuarine ecosystem.
Trans. Am. Fish. Soc. 1976, Vol. 105(4), pp.
546-549 276 393
DUKE, THOMAS W. Cycling of pollutants. Estuarine
Processes, Vol. 1, Uses, Stresses, and Adaptation
to the Estuary., Martin W. Wiley, editor, 1976,
pp. 481-482 320 399
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Reprinted from The Pathology
of Fishes, pp. 559-584, 1975,
with permission of The Univ.
of Wisconsin Press , Madison
HlSTOPATHOLOGICAL EFFECTS OF PESTICIDES AND RELATED
CHEMICALS ON THE LIVERS OF FISHES
John A. Couch
Contribution No. 152
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Histopathological Effects of
Pesticides and Related Chemicals on
the Livers of Fishes
JOHN A. COUCH
Evidence for the accumulation of pesticides in aquatic ecosystems is abundant
(26). Nontarget species such as fishes from salt and fresh water have been moni-
tored for pesticide contamination (15, 16). Certain pesticides, e.g., organo-
chlorines and their metabolites, accumulate in wild fish, particularly in liver and
fatty tissues (7).
Results of controlled laboratory exposures of fishes to pesticides and related
chemicals reveal that the liver is often the organ with highest pesticide concentra-
tions (7, 14, 15), and greatest damage or impairment (8, 12). This information,
combined with the general knowledge that the liver of vertebrates is a chief meta-
bolic and detoxication organ, suggests that a review of the histopathology of the
livers of fishes in reference to pesticide exposure would be of value.
REVIEW OF PESTICIDE-RELATED LIVER LESIONS
Considerable bioassay and toxicological research concerning effects of pesti-
cides on fishes has been reported. Experimental pesticide-induced acute and
chronic mortalities have been well documented for many fresh- and saltwater
species (26). However, of over 900 commercial pesticide formulations in general
use (26), fewer than 30 have been reported to have been tested in the laboratory
for histological effects on the livers of fishes (Table 23.1).
559
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560 Part IV: Chemical and Physical Agents
Table 23.1. Pesticides and Related Chemicals That Have Induced Nonspecific or
Specific* Liver Changes in Fishes (see text for details)
Organochlorines
Organophosphates Carbamates Other
Insecticides Insecticides
Chlordane Abate
DDT Dursban
Endrin* Dylox
Heptachlor Malathion
Lindane* Parathion
Methoxychlor*
Mirex
Telodrin
Toxaphene
Herbicides
Dichlobenil
Dowicide G
2,4-D*
Silvex
Industrial chemicals
Polychlorinated biphenyls
(Aroclor 1248)
(Aroclor 1254)*
Insecticide Lampreycide
Sevin* TFM (3-trifluoromethyl-
4-nitrophenol)
Herbicide
Hydrothol 191 (N,N
dimethyl-alkylamine salt
of endothal)
Even though many species of fishes are inadvertently exposed to pesticides
every year, fewer than 20 species have been reported to have been examined for
liver changes following exposures to pesticides under controlled conditions (Table
23.2).
The following is an attempt to summarize available results of controlled field
and laboratory research. Included are recent results from the author's laboratory
concerning the pathogenic effects of pesticides and certain related chemicals on
the livers of estuarine and marine fishes.
The significance of microscopic effects reported by different authors should
be determined in relation to the concentration of pesticides and the methods of
exposure of fishes. Information has been included, when available, on the types
of exposure (e.g., bath, food, flowing water, ponds, aquaria, tanks) as well as on
the concentration of pesticide and the duration of exposure. Lethal concentra-
tion data (LCSO; LC100) for specific pesticides and fishes are given when cited in
relevant works.
Discussions of pesticides and their effects are placed under the following ma-
jor chemical groupings: (1) organochlorines, (2) Organophosphates, (3) carba-
mates, (4) other chemicals.
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Effects of Pesticides on the Livers of Fishes
561
Table 23.2. Fishes Examined for Liver Lesions Following Exposure to
Pesticide or Related Chemicals
Common name
Scientific name
Chemical and source data
Freshwater species
Bluegill
Brook lamprey
Brown trout
Bluntnose minnow
Cutthroat trout
Goldfish
Guppy
Lake trout
Rainbow trout
Redear sunfish
Marine species
Pinfish
Sheepshead minnow
Spot
Lepomis macrochirus
Entosphenus lamottei
Salmo trutta
Pimephales notatus
Salmo clarki
Carassius auratus
Poecilia reticulata
(Lebistes reticulatus)
Salvelinus namaycush
Salmo gairdneri
Lepomis microlophus
Lagodon rhomboides
Cyprinodon variegatus
Leiostomus xanthurus
Abate (11), heptachlor (1),
methoxychlor (16), mirex (27),
dichlobenil (Casoron) (4),
2,4-D (5, 11), silvex(29)
TFM (2)
DDT (17)
Endrin (25)
Endrin(12)
Mirex (27)
DDT (17), DowicideG(6)
Aroclor 1248 (11), chlordane (11)
DDT (29), endrin (29), heptachlor
(3,29),malathion(3,29),
methoxychlor (3, 29), parathion
(29), Toxaphene (29), TFM (2)
Hydrothol 191 (10)
Mirex (21)
Dursban (Lowe, Wood)
Aroclor 1254, endrin (19), Sevin
(Lowe), Telodrin (Lowe),
2,4-D (Lowe)
ORGANOCHLORINES
Organochlorine Insecticides
Chlordane.* Eller (11) examined lake troutt that were exposed to 1.2 to 12
ppm of chlordane for one year (from March 1970 to March 1971). He found that
early in the exposure period both control and exposed fish had liver damage con-
sisting of focal areas of parenchymal cell vacuolation and degeneration. From
April through June the incidence and severity of degeneration in the liver increased
in the chlordane-exposed fish to a level about twice as severe as that in control
trout. About 80% of the exposed fish had degenerative changes.
* Common names of pesticides are used here. Chemical names or structural formulas of
many pesticides can be found references 13, 24, and 26.
•}• All fishes examined in this and subsequent summaries were killed and fixed for histolog-
ical examination; i.e., fish were survivors of experimental exposures.
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562 Part IV: Chemical and Physical Agents
DDT. King (17) found liver lesions in brown trout fry and adult guppies ex-
posed in aerated aquaria for 14 days to 0.00032 to 3.2 ppm DDT. In the trout
fry, both control and experimental fish had many small vacuoles in hepatic cells.
This may have been related to resorption of fatty yolk by the young trout. Gup-
pies in 0.32 ppm DDT for one day presented entire liver sections with severe
vacuolation and necrosis.
Wood (29) examined rainbow trout exposed to 5 ppb DDT in tanks for 14
days (LC50 14 days = 5 ppb). He found no signs of liver changes.
Endrin. Mount (25) found lipid deposits in hepatocytes of bluntnose minnows
exposed for 291 days to 0.4 ppm or more of endrin in a continuous flow fresh-
water system. Wild fish did not have extensive lipid deposits but control fish did.
Mount attributed the lipid change in the liver of exposed fish and control fish to
high lipid content of the artificial diet.
Wood (in 19) found no liver changes in spot, an estuarine fish, exposed to sub-
lethal concentrations of endrin in flowing sea water. However, in spot surviving
near-lethal concentration (0.075 ppb) exposures he found focal necrosis and in-
flammation in the liver as well as loss of glycogen and lipid.
Rainbow trout surviving exposure to 0.269 ppb endrin in tanks (LC50 7 days =
0.269 ppb) had severe, nonspecific degenerative liver lesions (29).
The most detailed report available on endrin-induced changes in the liver of
fishes is that of Eller (12). He found that cellular changes occurred in the livers
of cutthroat trout following water or food exposures of 0.01 ppm or 0.01 mg/kg
respectively of endrin. Certain of the induced changes resembled prehepatoma-
tous lesions: (1) liver cord disarray, (2) presence of mitotic cells in liver, (3) bi-
nucleate cells, (4) swollen cells, (5) pleomorphic cells, (6) bizarre cells with en-
larged nuclei, (7) acidophilic, pigmented cells, and (8) intrazonal and periportal
inflammatory foci. However, no fully developed hepatoma occurred. Fish ex-
posed to lower levels of endrin also showed certain of the above changes, but
most had lesions that were intermediate in severity between those in fish exposed
to higher levels and those of the unexposed controls. Eller believed the above de-
generative changes in the liver suggested nutritional deficiency enhanced by en-
drin exposure.
Heptachlor. Bluegills exposed to 0.050 ppm and 0.037 ppm heptachlor in
ponds developed severe degenerative liver lesions (1). After 14 days' exposure,
lesions consisted of variation in staining intensity, early necrotic change, and loss
of glycogen and fat. Exposures to lower concentrations produced no liver
changes. Bluegills fed 25 mg of heptachlor per kg of body weight in small ponds
had vague liver changes (1) consisting of loss of liver cord pattern, cellular shrink-
age, and loss of glycogen and fat.
Wood (in 3; 29) reported that rainbow trout exposed to 7.5 to 17.5 ppb hep-
tachlor for 14 days (LC50 14 days = 7.5 ppb) in tanks had severe liver changes
consisting of deposition of bile pigment in parenchymal cells and degeneration of
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Effects of Pesticides on the Livers of Fishes 563
liver tissue. The changes were too nonspecific to be used to identify heptachlor
as the cause of the changes.
Lindane (BHC). Wood (in 3; 29) found lesions in the livers of rainbow trout
surviving exposure to 15 to 23 ppb lindane for 7 days (LCSO 7 days = 15 ppb).
The lesions were focal, necrotic areas mainly associated with the portal triads.
Fish having signs of toxicity during exposures had liver lesions of an "early coag-
ulative type." Cellular detail in these lesions was obliterated. Wood believed that
these lesions were specific for lindane in comparison to more general lesions pro-
duced by exposures of fish to other insecticides.
Methoxychlor. Rainbow trout exposed to 10.0 ppb methoxychlor in ponds
for one week (LC50 7 days = 10.0 ppb) had only nonspecific degenerative changes
in the liver (3, 29). These changes indicated damage caused by a toxic substance,
but were not specific enough to identify methoxychlor as the toxic agent.
Bluegills exposed to 0.01 ppm or 0.04 ppm methoxychlor in ponds for 13
weeks exhibited variation in liver condition over a period of time (16). After 3
days at 0.01 ppm or 14 days at 0.04 ppm, nonspecific degenerative changes char-
acterized by some liver parenchymal shrinkage, increased cytoplasmic granularity,
and partial loss of liver cord orientation occurred. By day 56 of exposure, these
vague changes were not obvious. At day 1 of the 0.04 ppm exposure, minute eo-
sinophilic globules appeared in blood vessels of the liver of all exposed fish. By
day 3, these globules had coalesced to form spherical masses of variable sizes in
the liver blood capillaries. By day 56 the spherical masses (possibly precipitated
serum proteins (16)) had disappeared.
Mirex. Bluegills exposed to 0.0013 ppm or 1.0 ppm mirex in ponds had no
liver changes (27).
Pinfish fed approximately 20 ppm mirex for five months in flowing sea water
had no liver changes (21).
Goldfish-exposed to 1.0 ppm or 0.1 ppm mirex underwent little change in liver
structure (27). However, Van Valin, Andrews, and Eller (27) did find foci of acid-
fast bacteria in the livers of exposed fish and stated that the stress of the pesticide
challenge in the presence of mycobacterial infection contributed to higher mortal-
ity in the exposed as compared to control fish.
Telodrin. J. I. Lowe (personal communication) reported that of several pesti-
cides tested on estuarine and marine fishes, Telodrin was one of the more toxic.
Wood examined the spot which Lowe had exposed to 0.01 ppb Telodrin (sub-
lethal level) for five months in flowing sea water. He found minimal degenerative
changes in the livers of surviving fish. These changes consisted of scattered eosino-
philic globules (remains of necrotic liver cells) adjacent to the intrahepatic pan-
creatic acinar tissues.
Toxaphene. Rainbow trout that survived exposure to 0.005 ppb toxaphene
for seven days (LC50 7 days = 0.005 ppb) had extensive liver damage, according
to Wood (29). Parenchymal cell necrosis and disruption of liver cord structure
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564 Part IV: Chemical and Physical Agents
were striking, but neither together nor alone were these signs sufficient for iden-
tification of toxaphene as the toxic agent.
OrganochlorineHerbicid.es
Dichlobenil (Casoron). Cope, McCraren, and Eller (4) studied changes in the
livers of bluegills exposed to single treatments of 10 ppm, 20 ppm, and 40 ppm
of dichlobenil in ponds. Livers of exposed fish had an increase in connective tis-
sue, and early adenomatous change. Also present was abundant nuclear pyknosis,
hepatocyte karyolysis, and focal and massive necrosis. All exposed bluegills had
these lesions through 59 days. Livers of fish exposed to 10 ppm dichlobenil re-
turned to normal by 59 days, but those treated with 20 ppm and 40 ppm contin-
ued to have the lesions through 112 days.
Dowicide G (Sodium pentachlorophenate). Guppies were exposed to 0.5 ppm
of this herbicide for 180 days in aerated aquaria by Crandall and Goodnight (6).
Fish examined following 20 to 30 days of exposure had enlarged liver sinusoids
and enlarged, hyperchromic hepatocyte nuclei. Controls were normal. At 60
days, the exposed fish possessed less liver fat than controls, but were histological-
ly similar to controls. At 88 to 100 days, one fish had a necrotic liver and at 131
days, one had a "fatty" liver. At 180 days, two surviving fish had abnormally
compact parenchyma with few visible sinusoids and their liver parenchyma tis-
sues appeared "coagulated."
2,4-D. Bluegills were exposed in ponds for 112 days to single treatments of
0.1 ppm to 10.0 ppm 2,4-D by Cope, Wood, and Wallen (5). Liver sections from
fish exposed 1 to 14 days had glycogen loss, irregular staining, and periodic acid-
Schiff-positive deposits in liver sinusoids. These signs were not found in fish ex-
posed longer than 112 days. The authors considered that the glycogen loss in the
fish from early exposure was possibly related to the appearance of the PAS-posi-
tive deposits in the liver sinusoids and to similar deposits found throughout the
vascular system of the exposed fish. These deposits appeared globular or spheri-
cal and ranged in diameter from 1 to 50 microns. Their PAS-positive material
was also diastase-resistant, iron-negative (Prussian blue reaction), was not acid-
fast, and was gram-negative. PAS-positive bodies in the liver sinusoids provided a
possible clue to the origin of the PAS bodies elsewhere. The authors suggested
that the concomitant losses of glycogen from the liver and the appearance of the
PAS-positive vascular deposits were highly specific signs of 2,4-D toxicity.
In an unrelated experiment, Eller (11) found histological evidence of effects
of 2,4-D on liver carbohydrate metabolism in bluegills. In exposed, moribund
fish he found abnormal PAS-positive material in the nuclei of hepatocytes along
with large clear intranuclear vacuoles. This, with other evidence, suggested a
2,4-D-induced hyperglycemic state. This state was only transitory, however, be-
cause in fish exposed for long periods the pathological liver changes disappeared.
Lowe (personal communication) exposed spot to 1.0 ppm of 2,4-D for 30
days in flowing sea water. 1 have examined livers of both control and exposed
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Effects of Pesticides on the Livers of Fishes 565
fish. In only one of five exposed spot was there a significant departure from nor-
mal. This fish had moderate-to-heavy congestion of the liver sinusoids and large
lytic cavities in the liver parenchyma. The cavities may have been the result of
tissue instability during the tissue processing, but the congestion seen in the
sinusoids appeared to be a genuine pathological condition.
Silvex (Kuron). Wood (29) examined bluegills exposed to 1.0 ppm, 3.0 ppm,
and 10.0 ppm silvex in ponds. In exposed fish he found generalized, degenerative
liver changes consisting of glycogen and lipid loss, staining variability, cell shrink-
age, and liver cord disorientation.
Poly chlorinated Biphenyls (PCS'S)
These chemicals have not been used extensively as pesticides, but are listed in
Pimentel's review (26) of ecological effects of pesticides on nontarget species.
Several polychlorinated biphenyls have been found in water, sediments, and ani-
mal tissues from around the world (14, 28). The apparent ubiquity of PCB's in
the natural environment has led to considerable effort to learn of their possible
toxic effects on many animal species (26, 28), including fishes (11, 14).
Chemically the PCB's are chlorinated biphenyls and thus share some charac-
teristics, such as fatty and liver tissue affinity and relative long life (persistence)
in the environment, with certain organochlorine pesticides, e.g., DDT.
To date, no published reports exist concerning effects of PCB's on livers of
fishes. However, the following information is from work recently completed by
Eller (11) at the U.S. Bureau of Sports Fisheries and Wildlife, Fish Pesticide Lab-
oratory, Columbia, Missouri, and from work recently completed by several inves-
tigators at the U.S. Environmental Protection Agency, Gulf Breeze, Florida,
Laboratory.
Two PCB's, Aroclor 1248 and Aroclor 1254, have been tested on fishes at the
above laboratories.
Aroclor 1248. Lake trout exposed to 1.2 to 12.0 ppm Aroclor 1248 for one
year (March 1970 to March 1971), and examined by Eller (11), had early, pro-
gressive, degenerative liver changes. Control fish also had some similar lesions ini-
tially. The changes were (1) focal degenerative regions of the liver, (2) cytoplas-
mic vacuolation, and (3) pleomorphism of parenchymal cells. As the exposure
period progressed, the liver lesions of exposed fish increased in severity. From
April through June, 80% of exposed fish examined had liver lesions two to three
times more severe than any found in control fish. By August all fish exposed to
high levels of Aroclor 1248 had extensive liver parenchymal cell vacuolation.
The final effects of this exposure are yet to be reported.
Aroclor 1254. Hansen et al. (14) at the Gulf Breeze Laboratory studied the
chronic toxicity, uptake, and retention of Aroclor 1254 in two estuarine fishes,
spot and pinfish. They determined the relative uptake of Aroclor 1254 in six
different tissues. The liver concentrated the greatest relative amount of this PCB.
Lowe and Hansen (personal communication) supplied me with fish samples for
-------
566 Part IV: Chemical and Physical Agents
I
t .
i.
•
Fig. 23.1. Normal liver parenchyma of spot. Note liver cords, 1 to 2 cells thick, and sinus-
oids in tubulo-sinusoidal pattern. X450.
10
-------
Effects of Pesticides on the Livers of Fishes 567
pathology from Aroclor 1254 chronic exposure experiments. Also, I have ex-
posed spot to Aroclor 1254 in order to acquire tissue for pathology. The follow-
ing is a report on liver changes in spot associated with Aroclor 1254 exposures.
Spot from wild populations in Pensacola Bay, Florida, and from control tanks
were used to establish histological and cytological patterns in normal liver, and
were directly compared to exposed fish.
Normal spot possess a tubulo-sinusoidal liver (9) containing disseminated pan-
creatic tissue (Figs. 23.1-23.6). The pancreas follows the course of the portal
vein and bile duct through the liver parenchyma. In wild and control fish the liver
parenchymal cells are arranged in cords (or muralia) 1 to 2 cells thick (Figs.
23.1, 23.2). Liver parenchymal cells of wild spot are usually laden with glycogen
(heavily PAS-positive, diastase-labile), whereas those of fish held for one week or
longer under control or experimental exposure conditions lose most of the detect-
able glycogen. Apparently these fish do poorly on artificial diets. This fact possi-
bly introduces a nutritional variable into any evaluation of effects of toxic sub-
stances. Wild spot have very little fat in their livers, as demonstrated by oil red 0
treatment of frozen sections.
Juvenile and adult spot were exposed to 5.0 ppb Aroclor 1254 in flowing sea
Fig. 23.2. Cross section of normal liver parenchyma of spot demonstrating the tubular nature
of sinusoids and their relationship to parenchymal cells. Note space of Disse (arrows). X1000.
11
-------
568 Part IV: Chemical and Physical Agents
** •-- s-l'
'&&•.
:,&"
t ^'VVi^''*
4' 23g?4/
*jd&
% .«v;vr •• -
•^ v; -*LU fcv
f^.*^/'
^^~i ^ » * j . —•• -«w . *>v9| I i^*
••^^e^p^
«,x-/--^
-•
+ ? ^
f
Fig. 23.3. Normal liver parenchyma, pancreatic acinar tissue, and branches of the portal vein
in spot. X450.
12
-------
Fig. 23.4. Longitudinal section of normal bile duct and parallel pancreatic acinar tissue in
liver of spot. X450.
if *
-
. •
«- -
.
V* V," •
i
St '***
Fig. 23.5. Central vein and branches in normal liver of spot. Note that lumen of vein is only
partially filled with red blood cells. X450.
569
13
-------
X^K>XV-"- - **i '•*»*» X
r ,
'--~^S;;>
%" • - . ^» ^
^
I
•
Fig. 23.6. Normal reticulin pattern of portal vein, adjacent pancreatic acinar tissue, and liver
parenchyma. Lillie's silver oxide method; X450.
. • -.-. *, , y . ,
Fig. 23.7. Liver of spot exposed for 2 weeks to 5 ppb Aroclor 1254. Note extensive vacuo-
lation of parenchyma! cells characteristic of intermediate pathological response. X450.
570
14
-------
f
i
* m
Fig. 23.8. Large, smooth-edged vacuoles in liver of spot (enlarged from Figure 23.7), X1000.
** ** ».*"•
•JP t
«' %
<
,1
*-t
'
, « *
. *»
\ '!•
Fig. 23.9. Frozen section of control, normal spot liver treated with oil red O. Note lack of
large lipid deposits. XI000.
571
15
-------
572 Part IV: Chemical and Physical Agents
• * <•$**- .v
Fig. 23.10. Frozen section of exposed spot liver treated with oil red O. Note extensive lipid
deposits which are oil red O-positive. Arrows indicate parenchymal cells. XI000.
16
-------
Effects of Pesticides on the Livers of Fishes 573
Fig. 23.11. Focus of early fibrosis and cholangiolar epithelial proliferation in liver of spot
exposed to Aroclor 1254 for 2 weeks. X450.
water for from 14 to 56 days. After one week's exposure, considerable glycogen
depletion had occurred in livers from both control and exposed spot, but there
were no significant morphological differences between the two groups.
Following two weeks' exposure, relative differences in histological patterns of
exposed vs. control spot livers were found. Parenchymal cell vacuolation in ex-
posed livers was two to three times as great as that in control livers (Figs. 23.1,
23.7, 23.8). Moderately large (15 to 20 microns), smooth-edged vacuoles in he-
patic cells indicated fatty accumulation in the exposed spot. Oil red O staining
of frozen sections from these fish confirmed that the content of the vacuoles was
lipid (Figs. 23.9, 23.10). One fish had abnormal fibrotic, cholangiolar epithelial
proliferative foci in the parenchyma (Fig. 23.11).
Usually, following the third week of exposure to 5.0 ppb, cumulative mortali-
ty reached 50% or greater (14). During this period, surviving fish that were exam-
ined demonstrated fatty change, as indicated by extensive vacuolation in liver
parenchymal cells (Fig. 23.12). Also, in some, pancreatic acinar tissue had under-
gone severe degeneration, consisting of vacuolation and necrosis (Fig. 23.12).
Moribund fish had the most striking changes in liver tissues. These changes con-
sisted of focal necrosis, sinusoidal congestion, extreme fatty change, and occur-
rence of PAS-positive, diastase-resistant, amorphous inclusions (probably ceroid)
17
-------
574 Part IV: Chemical and Physical Agents
•-
.•* -V.
Fig. 23.12. Liver from spot exposed for 3 weeks to 5 ppb of Aroclor 1254. Note extensive
vacuolation of parenchyma and degeneration of pancreatic acinar tissue characteristic of
advanced pathological response. XI00.
18
-------
Effects of Pesticides on the Livers of Fishes 575
'<•>'
>*M
*
*• ' * 1» W^ Ik 'I « \ / —• V -\
**^L ->- **J k^ i »s*F«f ..v,-,-.- ,f• J--.S; ?%v '•*• • ?^.
*••'' ' / • ' WW />».. •»
v;^^.^- :*l'f • *
f *r " »? • *•«*.' '#r* *» 0 • . % * *
A^— .<%. •?« ' -; <
•? -\ . -r
.•-*• v . j-^ JJH, " -•' • •;%.
•*f ^ ^>, k^ * *.af-
^ : *' ' ' - ; > y
*v . V :
;*.,*'
Fig. 23.13. Parenchyma of liver from moribund spot exposed to 5 ppb Aroclor 1254 for 3
weeks. Note congestion of sinusoids, vacuolation of cells, and presence of PAS-positive,
diastase-resistant, intracellular inclusions (arrows). Hematoxylin and PAS method; X450.
in parenchymal cells (Figs. 23.13, 23.14). A few moribund fish also had exten-
sive infiltrates of lymphocytes in and around the degenerate pancreatic acinar tis-
sues of the liver.
Control fish examined during the exposures had light-to-moderate parenchy-
mal cell vacuolation but none of the other signs described above. Vacuoles in
liver cells of controls were usually much smaller than those in exposed fish.
The great extent of the fatty change as indicated by oil red O retention in
frozen sections and extensive vacuolation in paraffin sections (Figs. 23.12-23.14)
and the occurrence of PAS-positive amorphous bodies in parenchymal cells of the
liver of moribund spot are indications, possibly specific, for Aroclor 1254 toxici-
ty. These two changes, in combination, have not been reported for fish exposed
to other toxic compounds.
ORGANOPHOSPHATES
Organophosphate Insecticides
Abate. Bluegills exposed to 1 Ib of Abate per acre in ponds for 63 days had
liver alterations, according to Eller (11). The changes were (1) atrophy and stain-
19
-------
576 Part IV: Chemical and Physical Agents
Fig. 23.14. Higher magnification of liver parenchyma from spot shown in Figure 22.13.
Note varying sizes of PAS-positive inclusions (arrows) and vacuoles. Hematoxylin and PAS
method; X1000.
20
-------
Effects of Pesticides on the Livers of Fishes
577
ing variability of liver parenchymal cells, (2) liver cord distortion, (3) massive foci
of edema, (4) hyperemia, and (5) necrosis of both liver parenchyma and pancreat-
ic tissue.
Dursban. Lowe (personal communication) exposed sheepshead minnows to 5
to 10 ppb Dursban in flowing sea water for five months. These control and ex-
posed fish were examined by Wood and by me for lesions. Both of us found ex-
tensive fatty change (Figs. 23.15, 23.16) in the livers of exposed fish. Wood sug-
gested that these changes may have been due not to a direct effect of Dursban,
but to a secondary effect of starvation caused by a change in food habits (per-
sonal communication). I found vascular stasis in the livers of several sheepshead
minnows that survived the exposure (Fig. 23.17). This condition was not appar-
ent in any control fish.
. None of these changes was considered specific for Dursban.
Dylox. Matton and LaHam (22) exposed rainbow trout to 10 to 100 ppm Dy-
lox for 16 hours. They found vacuolation of liver cells and suggested that the
cause was tissue hypoxia. No other liver changes were reported.
Malathion. Wood (in 3; 29) found degenerative lesions of an undescribed na-
ture in livers of rainbow trout exposed to sublethal 0.6 ppm and 1.0 ppm mala-
thion concentrations for up to 30 days in pond exposures. These lesions disap-
peared after 30 days.
fl*58lfe*
Fig. 23.15. Normal liver parenchyma of sheepshead minnow. X450.
21
-------
.»
-
*
f
*•{** *«J
-
\
« ;••• » « , ... . «' **
f • • -- ^r i^* •* «•>*
-.*» ^, - > *. *- ;ir.--* •. , *
^^JL. V
7 A 'I
% *A V *^ •» *
._ 5, «• » » "
^ ".,
of
fc>*"
. * &
2-
'•%
' ^A
•
3*' %
.. i tt . t/ ^>
Fig. 23.16. Region of parenchymal cell vacuolation in liver of sheepshead minnow exposed
to Dursban insecticide for 5 months. X450.
Tig. 23.17. Stasis in central vein of liver of sheepshead minnow exposed to Dursban. Com-
pare with normal condition of central vein in spot in Figure 23.5. X450.
578
-------
Effects of Pesticides on the Livers of Fishes
579
Parathion. Rainbow trout exposed in ponds to 950 ppb parathion for 7 days
7 days = 950 ppb) had nonspecific liver changes. Wood found liver paren-
chymal cell swelling and liver sinusoid congestion. He thought these changes were
probably secondary to either gill or kidney damage which he found in the exposed
fish (29).
CARBAMATES
Carbamate Insecticide
Sevin (carbaryl). The only carbamate pesticide tested for possible liver effects
was Sevin. Lowe (20) exposed spot to 0.1 ppm Sevin for five months (LCSO 12
days = 1 .0 ppm) in flowing sea water. Wood (in 20) reported that he was unable
to find changes in liver tissues of exposed fish. I have recently examined fish that
Lowe exposed to Sevin and have found a possible change in the intrahepatic pan-
creatic tissue in the liver. Four out of five exposed spot which were examined
had clusters of large vacuoles in the deep periportal pancreatic acinar tissue (Fig.
23.18). The vacuoles appeared to be intracellular, having caused hypertrophy of
the acinar cells. They did not appear to be hydropic in nature, but could have
Fig. 23.18. Foci of vacuolation (possible fatty change) in deep periportal pancreatic acinar
tissue in spot liver. Fish was exposed to 0.1 ppm Sevin for 5 months. X100.
23
-------
580 Part IV: Chemical and Physical Agents
contained lipids, which were extracted during tissue processing. Control fish did
not have these vacuoles and I have not found them in other pesticide-exposed
fishes or seen them described in reports in the literature.
OTHER CHEMICALS
Lamprey cides
TFM (3-trifluoromethyl-4-nitrophenol). TFM selectively kills lamprey larvae at
low concentrations (LC100 for 8 hours = 0.75 ppm), but is relatively nontoxic to
rainbow trout, according to Christie and Battle (2). Wild lamprey larvae and rain-
bow trout were exposed to 0.75 ppm, 3.00 ppm, and 6.00 ppm by these authors
and examined for liver lesions. At all three concentrations the brook lamprey
livers became excessively red in color, indicating considerable stasis in the super-
ficial hepatic vessels. Vasodilation was found involving the hepatic sinusoids.
Liver vascular/cellular ratios were greater in lampreys exposed to all three concen-
trations than in control lampreys or rainbow trout. Rainbow trout had no liver
vascular or cytological effects. The authors suggested that the toxic effect of
TFM was perhaps associated with a direct effect on the vascular endothelium (in
gills as well as liver), giving rise to increased permeability of vascular membranes
and to loss of plasma.
Herbicides
Hydrothol 191 (NJV-dimethylalkylamine salt of endothal). Eller (10) studied
redear sunfish that had been exposed to 0.3 ppm to 1.50 ppm Hydrothol 191 in
ponds (LC50 96 hours = 125 ppm). He sampled these fish over a period of 56
days following single, initial applications of the herbicide. Liver structure had
considerable change over this period in different fish samples. At sublethal doses
of 0.03 ppm and 0.30 ppm, between 7 to 14 days' exposure the liver showed
small lymphocyte infiltrates, small aggregates of pigmented liver cells, and occa-
sional swollen liver cells. At 0.3 ppm after 28 days' exposure, fish livers had
chronic changes consisting of many pleomorphic cells, hypertrophic bizarre cells,
large masses of pigmented cells containing hemosiderin and lipofuchsins, and liver
cord distortion. By 56 days of exposure, liver structure had returned to a condi-
tion similar to controls. Because of marked variation among replicate samples
taken simultaneously, Eller believed that these changes could not be associated
unequivocally with Hydrothol 191 exposure.
CRITICAL EVALUATION
The liver, gills, kidney, gonads, and brain have usually been the organs of
choice for histological studies of pesticide-induced changes in fishes. Although
the present review is restricted to pesticide-induced changes in the livers of fishes,
24
c
-------
Effects of Pesticides on the Livers of Fishes 581
it also serves as a representative index to the available or published histological
works on fishes experimentally exposed to pesticides and related chemicals.
Except for the work of Eller (4, 10-12) and Wood (in 3, 19; 29), few detailed
investigations have been reported. Most studies providing data on liver lesions
have involved freshwater fishes, particularly the rainbow trout and the bluegill
(see Table 23.2). Of special concern is the paucity of published studies on effects
of commercial biocides on estuarine and marine fishes. Information in this regard
is presently to be found for only three species (Table 23.2) (18-21), and this in-
formation is far from conclusive.
Most liver lesions in fishes which have been exposed to pesticides have been
general or nonspecific. Specific syndromes produced in the liver as responses to
particular pesticides have been few in number. This may reflect the lack of de-
tailed work and the relatively small number of pesticides and fish species tested.
The following are worthy of emphasis. Cutthroat trout exposed to different con-
centrations of endrin (12) had a large spectrum of lesions constituting a prehepa-
tomatous syndrome. The severity of the lesions was related to concentrations of
endrin. Early necrotic, coagulative lesions were associated with the portal triads
of rainbow trout exposed to critical levels of lindane (BHC) (3, 29). Bluegills ex-
posed to methoxychlor had unique, eosinophilic globular masses in the liver vas-
cular system (16). These globules increased in number and size up to several days'
exposure and then disappeared. The similar occurrence of PAS-positive vascular
deposits in liver vessels of bluegills was related to abnormal glycogen metabolism
following 2,4-D exposure (5). Extensive fatty change, necrosis, and deposition of
PAS-positive amorphous inclusions in liver cells appear to be characteristic of PCB
(Aroclor) exposures in spot. Foci of medium-to-large vacuoles in the pancreatic
acinar tissue in the livers of spot were associated with Sevin exposure.
The most commonly encountered nonspecific liver lesion reported for fish
following pesticide exposure was fatty change. Exposure of fishes to the follow-
ing pesticides and chemicals produced liver parenchyma! cell vacuolation or oil
red 0-positive frozen liver sections interpreted here as probably the result of ab-
normal accumulation of lipid in the liver: chlordane, DDT, endrin, Dowicide G,
Aroclor 1248, Aroclor 1254, Dursban, Dylox. This list includes both organo-
chlorines and organophosphates. In certain of these cases (DDT, endrin, Aroclor
1254, Dursban) nutritional factors, as well as pesticide exposure, were suspected
in the onset of fatty change.
Chemicals from each of the representative major classes of commercial bio-
cides (e.g., organochlorines, organophosphates, carbamates, etc.) have been tested
for effects on livers of fishes. No completely characteristic trend of histopatholog-
ic effects has been reported for any of the given classes of chemicals. That is, no
described liver histopathological syndromes are presently known for positive iden-
tification of particular classes of chemicals toxic to fishes. Pesticides within the
same major group, such as DDT and endrin, both organochlorines, may produce
25
-------
582 Part IV: Chemical and Physical Agents
different or noncomparable signs in livers of fishes. Thus, even they are difficult
to diagnose on histopathology alone.
It is pertinent to note, at this point, that the information available does dem-
onstrate unequivocally that liver damage occurs in both acutely lethal and chronic
sublethal exposures of fishes to certain pesticides.
From the above brief evaluation it is obvious that considerable histological and
cytological investigation is needed to further define and characterize effects of
pesticides on the livers of fishes. Particularly needed are studies of the possible
interaction of variables such as nutrition and pesticides and their effects on such
organs as the liver. Mawdesley-Thomas (23) recently stated: "Following the use
of the more persistent biocides, much concern has been expressed in recent years
as to their long term effects on wild species. The long term effects of sublethal
doses of even DDT are ill-defined and insufficiently documented and further
study is required." Unfortunately, even the histopathology of acute or lethal
pesticide poisoning is unknown or incomplete for most wild species, including
fishes.
In future studies, emphasis should be placed on histochemistry and electron
microscopy of pesticide-related lethal and chronic sublethal changes in order to
approach more closely actual mechanisms of pesticide-induced injury in fishes.
Acknowledgment. I thank Jack Lowe and Dave Hansen for providing fish tis-
sues from their experimental exposures. L. L. Eller and E. M. Wood were very
cooperative in supplying information. Darryl Christinsen prepared some of the
fish tissue used in the Aroclor 1254 study. Pat Borthwick aided in design and con-
struction for experimental exposures of fish and in collection of fish.
Use of trade names of pesticides and related chemicals does not constitute en-
dorsement of these products by the U.S. Environmental Protection Agency.
Conribution number 152, Gulf Breeze Environmental Research Laboratory.
REFERENCES
1. Andrews, A. K., Van Valin, C. C., and Stebbings, B. E. Some effects of hep-
tachlor on bluegills (Lepomis macrochirus). Trans. Amer. Fish. Soc. 95: 297.
1966.
2. Christie, R. M., and Battle, H. I. Histological effects of 3-trifluoromethyl-4-
nitrophenol (TFM) on larval lamprey and trout. Can. J. Zool. 41: 51. 1963.
3. Cope, O. B. Contamination of the freshwater ecosystem by pesticides.
/. Appl. Ecol. J(suppL): 33. 1966.
4. Cope, O. B., McCraren, J. P., and Eller, L. L. Effects of dichlobenil on two
fishpond environments. Weed Sci. 17: 158. 1969.
5. Cope, O. B., Wood, E. M., and Wallen, G. H. Some chronic effects of 2,4-D
on the bluegill (Lepomis macrochirus). Trans. Amer. Fish. Soc. 99: 1. 1970.
6. Crandall, C. A., and Goodnight, C. J. The effects of sublethal concentrations
of several toxicants to the common guppy, Lebistes reticulatus. Trans Amer
Microsc. Soc. 82: 59. 1963.
26
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Effects of Pesticides on the Livers of Fishes 583
7. Duke, T. W., and Wilson, A. J., Jr. Chlorinated hydrocarbons in livers of
fishes from the northeastern Pacific Ocean. Pestic. Monit. J. 5: 228. 1971.
8. Eisler, R., and Edmunds, P. H. Effects of endrin on blood and tissue chem-
istry of a marine fish. Trans. Amer. Fish. Soc. 95: 153. 1966.
9. Elias, H., and Bengelsdorf, H. The structure of the liver of vertebrates. Acta
anat. 14: 24. 1952.
10. Eller, L. L. Pathology in redear sunfish exposed to Hydrothol 191. Trans.
Amer. Fish. Soc. 98: 52. 1969.
11. Eller, L. L. Annual reports. U. S. Bur. Sport Fish. Wildlife, Fish Pestic. Lab.,
Columbia, Mo. Also unpublished reports, cited with permission of Eller.
1970, 1971.
12. Eller, L. L. Histopathologic lesions in cutthroat trout (Salmo clarki) exposed
chronically to the insecticide endrin. Amer. J. Pathol. 64: 321. 1971.
13. Frear, D. H., ed. Pesticide Index. 4th ed. State College, Pa., College Sci.
Publishers, 1969.
14. Hansen, D. J., Parrish, P. R., Lowe, J. I., Wilson, A. J., Jr. and Wilson, P. D.
Chronic toxicity, uptake, and retention of Aroclor 1254 in two estuarine
fishes. Bull. Environ. Contam. Toxicol. 6: 113. 1971.
15. Johnson, D. W. Pesticides and fishes—a review of selected literature. Trans.
Amer. Fish. Soc. 97: 398. 1968.
16. Kennedy, H. D., Eller, L. L., and Walsh, D. F. Chronic effects of methoxy-
chlor on bluegills and aquatic invertebrates. U. S. Bur. Sport Fish. Wildlife,
Tech, Pap. 53. 18pp. 1970.
17. King, S. F Some effects of DDT on the guppy and the brown trout. U. S.
Fish Wildlife Serv., Spec. Sci. Rep., Fish. 399. 20 pp. 1962.
18. Lowe, J. I. Chronic exposure of spot, Leiostomus xanthurus, to sublethal
concentrations of toxaphene in seawater. Trans. Amer. Fish. Soc. 93: 396.
1964.
19. Lowe, J. I. Some effects of endrin on estuarine fishes. Proc. Annu. Conf.
Southeast. Ass. Game Fish Comm. 19: 271. 1965.
20. Lowe, J. I. Effects of prolonged exposure to Sevin on an estuarine fish,
Leiostomus xanthurus, Lacepede. Bull. Environ. Contam. Toxicol. 2: 147.
1967.
21. Lowe, J. I., Parrish, P. R., Wilson, A. J., Jr., Wilson, P. D., and Duke, T. W.
Effects of mirex on selected estuarine organisms. N. Amer. Wildlife Natur.
Resour. Conf. Trans. 36. 1971.
22. Matton, P., and LaHam, Q. N. Effect of the organophosphate Dylox on rain-
bow trout larvae. /. Fish. Res. Board Can. 26: 2193. 1969.
23. Mawdesley-Thomas, L. E. Research into fish diseases. Nature (London) 235:
17. 1972.
24. Menzie, C. M. Metabolism of pesticides. U. S. Fish Wildlife Serv., Spec. Sci.
Rep., Wildlife 127. 487pp. 1969.
25. Mount, D. E. Chronic effects of endrin on bluntnose minnows and guppies.*
U. S. Fish Wildlife Serv., Res. Rep. 58. 38 pp. 1962.
26. Pimentel, D. Ecological effects of pesticides on non-target species. Exec. Off.
President, Off. Sci. Technol. U. S. Government Printing Off., Washington,
D.C., 1971.
27
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584 Part IV: Chemical and Physical Agents
27. Van Valin, C. C., Andrews, A. K., and Eller, L. L. Some effects of mirex on
two warm-water fishes. Trans. Amer. Fish Soc. 97: 185. 1968.
28. Vos, J. G. Toxicology of PCB's for mammals and for birds. Environ. Health
Perspect. 1: 105. 1972.
29. Wood, E. M. The pathology of pesticide toxicity in fish. Unpublished.
DISCUSSION OF HISTOPATHOLOGICAL EFFECTS OF PESTICIDES AND
RELATED CHEMICALS ON THE LIVERS OF FISHES
C. J. Dawe: This kind of work has been needed for a long time and you are
just getting into the studies. I wonder how the situation will be for long-term
study, particularly keeping in mind the idea of looking for neoplasms, either in the
liver or elsewhere?
J. A. Couch: The problem of estuarine fish that we chose to work with is that
most of them are not amenable to culture and to maintenance as satisfactorily as
many of the freshwater species. You can keep fish three months or so and think
you are doing everything right and then lose them overnight. That is the problem
we face. For example, dietary needs are not even known for most of the estuarine
fishes as compared to the cultured fishes such as the rainbow trout. As long as
you try hard, I think you can maintain spot for long-term sublethal exposures that
might lead to neoplastic involvement. It is still a pioneering field as far as con-
cerns maintenance of any of the very fastidious marine organisms.
Question: You mentioned fin rot. I believe it took place only in the experi-
mental fish.
J. A. Couch: Right; I didn't go into detail. The most striking effect grossly
to the fish exposed to PCB's was that in the control tank you could maintain the
fish for weeks and have a very low mortality, no fin rot, and the fish appeared
well, whereas in those fish exposed to PCB I would say that the primary cause of
death was a tremendous induced fin rot concurrent with exposure to PCB's. By
the second or third week of exposure, you might lose up to 90% of your fish with
fin rot disease. After seeing this repeated several times I can only believe that it
is related to the PCB action somehow on the mucous or protective coating of the
fish.
Question: Did you notice any damage to the hematopoietic system, particu-
larly the thymus?
J. A. Couch: I haven't looked at the thymus in that much detail.
28
-------
Reprinted from Archives of
Environmental Contamination
and Toxicology, Vol. 3(1):
22-39, 1975, with permission
of Springer-Verlag, New York
Inc.
TOXICITY OF AROCLORR 1254 AND ITS PHYSIOLOGICAL
ACTIVITY IN SEVERAL ESTUARINE ORGANISMS
Del Wayne R. Nimmo, David J. Hansen, John Couch, Nelson R. Cooley,
Patrick R. Parrish, and Jack I. Lowe
Contribution No. 162
29
-------
TOXICITY OF AROCLOR® 1254
AND ITS PHYSIOLOGICAL ACTIVITY
IN SEVERAL ESTUARINE ORGANISMS
D. R. NIMMO, D. J. HANSEN, J. A. COUCH, N. R. COOLEY,
P. R. PARRISH and J. I. LOWE
V. S. Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Sabine Island, Gulf Breeze, Florida 32561
(Associate Laboratory of the National Environmental
Research Center, Corvallis, Oregon)
The occurrence of high concentrations of a PCB (Aroclor 1254) in the Pensacola
estuary prompted field and laboratory studies by the Gulf Breeze Environmental
Research Laboratory (EPA). Monitoring of the estuary indicates the chemical is
present in all components—particularly in sediments and fishes. Residues appear to
be diminishing in sediments. Toxicity tests show estuarine species sensitive at ppb
concentrations in water, with a ciliate protozoan (Tetrahymena pyriformis W),
shrimps (Penaeus duorarum, P aztecus, and Palaemonetes pugio), and a fish
(Fundulus similis), affected at or near 1.0 ppb. Tissue concentrations of Aroclor
1254 similar to those found in natural populations of shrimps from the contaminated
estuary were successfully duplicated in laboratory experiments. Shrimps also con-
centrated the PCB from very low concentrations (0.04 ppb) in the water. Three
estuarine species demonstrated pathologic changes at tissue and cellular level after
chronic exposure to the chemical. Oysters (Crassostrea virginica) developed abnormal
infiltration of leukocytes in the connective tissue, spot (Leiostomus xanthurus)
developed fatty changes in their livers, and shrimp (Penaeus duorarum) developed
crystalloids in hepatopancreatic nuclei.
Polychlorinated biphenyl (PCB) residues were found in water, sediments and biota of
Escambia Bay, Florida, in April 1969 (Duke et al. 1970). Subsequently, investigation
into the effects of this chemical on estuarine organisms has been a major research goal of
the Gulf Breeze Environmental Research Laboratory. In this report, we present data on
the chemical in water, sediments and biota of Escambia Bay and adjacent areas, review
toxicological data and present some physiological-pathological information. Experimental
methods and materials used as well as chemical analyses are given elsewhere (Duke et al.
1910,Coo\ey etal. 1972, Hansen etal. 1971, Nummo et al. 197la, Lowe etal. 1972).
®In this paper, Aroclor and PCB are used interchangeably for Aroclor 1254. Aroclor® is a registered
trademark of the Monsanto Company, St. Louis, Mo. Reference to commercial products does not
constitute endorsement by the Environmental Protection Agency.
Contribution No. 162, Gulf Breeze Environmental Research Laboratory.
Archives of Environmental Contamination 22
and Toxicology, Vol. 3, No. 1, 1975, © 1975
by Springer-Verlag New York Inc. 01
-------
Toxic action of Aroclor 1 254 in estuarine organisms 23
PCB in Escambia Bay
Unlike other studies in which PCBs have been found in the environment, the chemical
in Escambia Bay apparently came from a single point-source and has been identified as
Aroclor 1254. Figure 1 is a comparison of chromatograms of (a) an Aroclor 1254
standard, (b) Aroclor 1254 in oysters from a 72-week chronic exposure to the chemical
in the laboratory, and (c) PCB isolated from oysters taken from Escambia Bay. PCB con-
centrations in both oyster samples were similar and both samples and standard were
analyzed on the same chromatograph. In laboratory studies with Aroclor 1254, the
oysters were continuously exposed to approximately ten ppt for 72 weeks and the
similarity of tissue residues to the PCB found in oysters from Escambia Bay indicates that
the chemical in the bay has changed little with time.
Monitoring data for the period September 1969 through December 1971 are presented
in Table 1. Average concentrations in water are given by assuming non-detectable levels as
Aroclor 1254 standard
Aroclor 1254 in oysters
exposed in laboratory
PCB in Escambia Bay
oysters
Fig. 1. Chromatograms of an Aroclor 1254 standard, Aroclor 1254 in oysters from a 72-
week chronic exposure in the laboratory and PCB in oysters from Escambia Bay.
32
-------
24 D. R. Nimmo et al.
zero. Concentrations of Aroclor in unfiltered water samples from Escambia River down-
stream averaged 0.6 ppb. Aroclor was found in 64% of the water samples from the river
and in 27% of the water samples from Escambia Bay. The average concentration in sediment
samples from the bay was 2.3 ppm. PCB in 101 samples of invertebrates, predominately
mollusks and crustaceans, averaged 0.8 ppm. Fishes from the bay had five times as much
PCB in their tissues as did invertebrates.
Aroclor in sediment samples from a survey taken in February 1970 were all in the ppm
range (Nimmo et al. 1971b). The samples were taken from the upper strata (e. g., upper
ten in.) by corer or by dredge. In this survey, the maximum residue (61 ppm) observed in
the river was found at the outfall from the industry; the maximum in Escambia Bay
(30 ppm) was found near the mouth of the river.
The amounts of PCB in sediment samples from several locations within the river and
bay have decreased in the ensuing months. The number of samples was inadequate for
absolute comparison, but the data indicate a trend. A decrease is especially noticeable in
the December 1970 and October 1971 surveys (Fig. 2), in which sediment samples
were taken with a corer at three locations in the bay. Generally, residues in the 1971
survey were about one-tenth the 1970 values, except one sample taken below the trestle
in the surface stream. PCB in the lower strata (4-12 inches) in the 1971 survey was
non-detectable, except at the outfall of the industry. Later, we examined a deeper
core—to 24 inches—taken above the trestle but found no residues. Cores taken in a 1972
survey generally indicated less PCB than in 1971.
Whole-body residues of Aroclor 1254 found in shrimps from Escambia Bay and con-
tiguous waters during 1969/1970 are shown in Figure 3. Each datum represents a com-
posite sample of at least five individuals. We show these data to indicate the dispersal of
Table I. Concentration of Aroclor® 1254 in water, sediment, and biota,
September 1969 through December 1971, in Escambia Bay, Florida
Type
Water
Water
Sediment
Invertebrates
Fishes
Samples
Total No.
67
37
56
101
17
Concentration
Positive %
64
27
78
92
100
Average (ppm)
0.0006
ND
2.33
0.81
3.99
Range (ppm)
ND- 0.0086
ND- 0.00007
ND-30.
ND- 6.9
0.29-20.
ND = Non-detectable: Water, < 0.00003 ppm; sediment or biota, < 0.01 ppm.
33
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Toxic action of Aroclor 1254 in estuarine organisms
25
the chemical in an estuarine environment from an apparent point source by either
biological or physical transport. Although the material was originally localized in the
sediments of upper Escambia Bay, shrimps captured in lower Pensacola Bay also contained
significant amounts of the PCB. In contrast, shrimps from adjacent bays such as Perdido
or Choctawhatchee did not have detectable levels of PCBs.
The amounts of Aroclor in biota from the estuary remain relatively high and the latest
survey showed amounts generally increasing at higher trophic levels (Fig. 4). In the
survey of October 1971, we found no detectable PCB in the sea grasses, Spartina sp. and
Zostera marina. The mollusk, Neritina reclivata, contained 0.49 ppm. Of two crustaceans,
blue crabs (Callinectes sapidus) had the greater residues (6.9 ppm). Among fishes, one
might expect sand seatrout (Cynoscion arenarius) and Atlantic cutlassfish (Trichiurus
lepturus) to have the highest residues because these species are predators; instead, the
highest residue (10 ppm) was found in silversides (Menidia beryllind), a species whose diet
consists mainly of plankton.
0-2 in.
2-4
4-6
6-8
Above trestle
'70 '71
0-2 in.
2-4
4-6
6-8
Below trestle
'70 '71
0.19 0.19
0.08 ND
0.002 ND
ND ND
Sampling dates: 12-14-70
10-26-71
11-22-72
'72
0.14
ND
ND
Fig. 2. Comparison of concentrations of Aroclor 1254 in cores taken 10 months apart in
upper Escambia Bay and River. Concentrations given in 1971 above and below trestle
are averages of two cores each.
34
-------
26
D. R. Nimmo ct al.
In Figure 4, PCB residues are compared in the same species or in species occupying
similar trophic levels and captured on the same day in Escambia and East Bays. The
collecting site in East Bay is about 35 km from the original source of the chemical. PCB
residues in species from Escambia Bay were five to ten times greater than those found in
East Bay but the data clearly show that the chemical was found in species captured dis-
tant from the original source of the PCB.
Toxicity of Aroclor 1254 to estuarine organisms
Laboratory research on the toxicity of PCB to estuarine organisms began immediately
after the chemical was discovered in the Bay. Animals from several trophic levels have
been tested.
Industrial plant
024 6 8 10
Big Lagoon
Fig. 3. Residues of Aroclor 1254 in shrimp (whole body) from Escambia Bay and con-
tiguous waters during 1969/1970. Each datum represents a composite sample of at least
five individuals.
35
-------
Toxic action of Aroclor 1254 in estuarine organisms
27
Population growth in test-tube cultures of the ciliate protozoan, Tetrahymena
pyriformis W, was reduced significantly by exposure to one ppb of Aroclor 1254
(Table II). Reduction was measured as effect on population growth rate and on popula-
tion density at 96 hr. Growth rate was estimated as the quantity b of the least squares
estimate of the line y = a + bx for the exponetial growth phase of the population growth
curve. These ciliates accumulated the PCB from the test media during the exposures. Cells
contained a maximum of 60 ppb (dry-weight basis) of PCB when grown for seven days in
medium that contained one ppb of Aroclor. Because the ciliates can accumulate PCB from
a culture medium, they could be a step in the transport of the chemical into aquatic food
webs under natural conditions.
Bioassays were conducted in flowing water to determine the toxicity of Aroclor 1254
to a mollusk, three crustaceans and three fishes (Table HI). Growth in oysters (Crassostrea
(ppm)
Escambia Bay
ND
ND
0.49
NO
0.98
6.90
3.00
3.80
10.00
4.50
1.50
(
1.80
1.60
1.30
2.90
Olive Nerite
Rangia
Penaeid Shrimp
Blue Crabs
Bay Anchovy
Catfish
Tidewater silversides
Silver perch
Sand seatrout
Spotted seatrout
Spot
Atlantic croaker
Hoflchoker
Atlantic cutlassfish
ND
ND
Trace
0.46
Fig. 4. Comparison of concentrations of Aroclor 1254 found in species collected in
Escambia (left side) and East Bays (right side): seagrasses, Spartina sp. and Zostera
marina; Rangia clams, Rangia cuneata; olive nerite, Neritina reclivata; brown and white
shrimp, Penaeus aztecus and P. setiferus; blue crabs. Callinectes spidtis; bay anchovy,
Anchoa mitchilli; sea catfish and gaff topsail catfish, Arius felis and Bagre marinus; tide-
water silversides, Menidia beryllina; silver perch Bairdiella chrysura; sand seatrout
Cynoscion arenarius; spotted seatrout, Cynoscion nebulosus;spot, Leiostomus xanthurus,
Atlantic croaker, Micropogon undulatus; hogchoker, Trinectes maculatus; and Atlantic
cutlassfish, Trichiurus lepturus.
36
-------
28 D. R. Nimmo et al.
virginicd), exposed to five ppb for 24 weeks was significantly reduced, but growth in
oysters exposed to one ppb for 30 weeks was not. Earlier work showed that hydrocarbon
compounds inhibited shell deposition significantly at concentrations of 0.1 to 0.5 ppm
in short-term tests (Butler 1966). In tests lasting about two weeks, various shrimps
Table II. Effect of Aroclor 1254 on population growth of
Tetrahymena pyriformis W a
Toxicant
(Mg/litef)
0
0.1
1.0
10.0
Mean growth
rateb
b
0.0212
0.0203
0.0195
0.0199
Mean population
Difference densityb
(%) (absorbance)
1.044
-4 0.984
- 8C 0.936
- 6C 0.954
Difference
-6
- 10d
_9d
aAfter Cooley et al. (1972).
b Means of 6 replicate experiments.
CF(3, 15) = 6.00 (P< 0.01).
dF(3, 15) = 23.001 (P< 0.005).
Table III. Chronic toxicity of Aroclor® 1254 to estuarine animals
in flowing water, 1 to 30 weeks a
Concentration, ppb (jxg/1)
Test animals
Pink shrimp
Longnose killifish
Grass shrimp
Brown shrimp
Pinfish
Spot
Eastern oyster
Range
0.6- 19.0
1.0-100.0
0.2- 12.5
0.1- 1.4
5.0
1.0- 5.0
1.0- 5.0
Minimum affecting
0.9
1.0
1.3
1.4
5.0
5.0
5.0
aControls did nbt exceed 25% mortality. Toxicity in oysters was measured by
reduced shell growth.
37
-------
29
Toxic action of Aroclor 1254 in estuarine organisms
(Penaeus duorarum, Penaeus aztecus, and Palaemonetes pugio) were killed by exposure to
0.9, 1.4, and 4.0 ppb, respectively. In tests lasting two weeks or longer this chemical was
lethal to longnose killifish(Fundulus similis) at 1.0 ppb, and pinfish (Lagodon rhomboides)
and spot (Leiostomus xanthurus) at 5.0 ppb. Test animals exposed for over one week
accumulated the PCB from the water; concentrations ranged from about 104 in
crustaceans and fishes to 10s in oysters over water concentrations.
Acute toxicity tests did not show the true sensitivity of marine species to this com-
pound. In comparison to short-term tests lasting 48 hr, Aroclor in chronic bioassays
lasting one week or more proved to be 100 times more toxic.
Mortalities were "delayed" as they usually did not begin until after one week of
exposure and continued to occur after the animals were removed from the toxicant. This
delayed mortality was similar to that observed with the insecticide mirex (Lowe et al.
1971). Gross signs of poisoning varied with species. Fishes typically developed hemorrhagic
lesions on the body, ragged fins, and stopped feeding. Shrimps became lethargic and also
stopped feeding, thereby mimicking the effects of low concentrations of DDT [1,1,1-
trichloro-2,2-&zXp-chlorophenyl)ethane] (Nimmo and Blackman 1972). Shrimps appear
to be most susceptible to the chemical during molting, as previously noted by Duke et al.
(1970) and Wildish( 1970).
Accumulation of Aroclor 1254 by shrimp
The pathway by which organisms obtain toxicants or the actual effects observed in the
laboratory under controlled conditions may or may not approximate those obtained
under field conditions. The pathway appears to be an open question in the field of
aquatic toxicology and the need for such research was stated by Sodergren ef a/. (1972)
after a study of the accumulation of DDT and PCB by a crustacean.
If we could produce tissue distributions of PCB in shrimp in the laboratory similar to
those found in the field, some insight might be gained into its mode of entry and con-
centrations in food or water in nature. We began a study by determining PCB residues of
shrimp from Escambia and Pensacola Bays (Fig. 5). We administered PCB at three con-
centrations (0.2, 0.68 and 43 ppm) in food in aquaria with flowing PCB-free seawater
(Fig. 6). PCB was added at 3.0 ppb to seawater filtered through gravel and charcoal and
3.0 ppb was added to unfiltered seawater (Fig. 7). PCB was added to unfiltered seawater
at 3.5 and 0.2 ppb (Fig. 8). Analytical methods for PCB were those of Nimmo etal. (197 la).
The results of the field surveys and laboratory studies are expressed as the "relative con-
centration" (Fig. 5, 6, 7, 8):
_ . . ppm in a single tissue or organ
Relative concentration = — x 100.
ppm in all tissues or organs
The proportion of Aroclor found in tissues of shrimp exposed in the laboratory to 0.2
ppb in water was nearest to that found in feral shrimp captured in the bays (Fig. 8) and
concentrations were within the ranges found in shrimp from the field. Thus, we believe
38
-------
30
D. R. Nimmo et al.
that shrimp from the laboratory exposures or feral shrimp from the bays probably ob-
tained most of the chemical from water. Shrimp could have absorbed PCB from the water
column near the sediment-water interface or directly from interstitial water of the sub-
strate. However, this suggestion does not exclude the possibility that shrimp obtained
some of the chemical from food, as was the case observed in Fig. 6.
We believe the rate of Aroclor loss from shrimp was low and insufficient to affect the
localization of Aroclor in the tissues of shrimp because in earlier experiments, we found
that in the hepatopancreas only half of the chemical was lost in 17 days and in the re-
maining tissues, Aroclor content remained constant for five weeks (Nimmo et al. 197la).
The shrimp used in the laboratory exposures and those captured in the bays are all
benthic species—either feeding on sediments or burrowing directly in them. It seems
reasonable to suppose that adsorption of PCB directly through the gills from contaminated
sediments would be greater because the animals are continually moving PCB-contaminated
wa'ter through them. The results of the laboratory studies lead us to believe that concen-
trations of PCB available to shrimp in Escambia and Pensacola Bays were low (e.g., <1.0
ppb in water; <1.0 ppm in food).
To determine if there was a concentration below which shrimp could not accumulate
the chemical, we tested several hundred grass shrimp (Palaemonetes pugio) at 0.04, 0.09
80
60
c
o
CO
b.
*•*
c
o
o
JO
o
EC
40
20
Natural populations
(range of 4 samples)
Hepato-
pancreas
Ventral
nerve
I
Digestive
tract
Heart
I
Gills
Abdominal |
muscle
Exo-
skeleton
Fig. 5. Distribution of Aroclor 1254 in tissues of shrimp expressed as relative concentra-
tion (%). The grey area represents the range of concentrations in four composite samples
of shrimp from different locations in the Pensacola estuary on different dates.
39
-------
Toxic action of Aroclor 1254 in estuarine organisms
31
80
60
c
o
CD
tr
20
Pink shrimp fed spot
(43 ppm whole body) for 16 days
Pink shrimp fed spot (field-captured,
0.2 ppm whole body) for 16 days
Pink shrimp fed croaker (0.68 ppm
in muscle) for 50 days
Ventral
nerve
Digestive
tract
| Heart |
Gills
Abdominal
muscle
Exo-
skeleton
| Hepato-
pancreas
Fig. 6. Distribution of Aroclor 1254 in tissues of shrimp fed Aroclor 1254-contaminated
diets. The grey area is as in Fig. 5.
80
c 60
o
u
c
o
u
a 40
01
DC
20
.. Pink shrimp exposed to 3.0 ppb
Aroclor in filtered water for 10 days
Pink shrimp exposed to 3.0 ppb
Aroclor in unfiltered water for 10 days
| Hepato- | Ventral | Digestive I Heart I
pancreas nerve tract
Gills
Abdominal | Exo- |
muscle skeleton
Fig. 7. Distribution of Aroclor 1254 in tissues of shrimp exposed to Aroclor 1254 in
filtered and unfiltered seawater. The grey area is as in Fig. 5.
40
-------
32
D. R. Nimmo et al.
80
60
c
o
c
0)
u
o
40
JO
CD
IT
20
Pink shrimp exposed to 0.2 ppb
Aroclor in water to 50 days
Pink shrimp exposed to 3.5 ppb
Aroclor in water for 35 days
| Hepato- | Ventral | Digestive | Heart | Gills | Abdominal | Exo- |
pancreas nerve tract muscle skeleton
Fig. 8. Distribution of Aroclor 1254 in tissues of shrimp exposed to the chemical in
flowing-water aquaria. The grey area is as in Fig. 5.
. ' Control residue level
8
Fig. 9. Aroclor 1254: Uptake and depuration in grass shrimp exposed to 0.04, 0.09 and
0.62 ppb in water.
41
-------
Toxic action of Aroclor 1254 in estuarine organisms
33
Fig. 10. (A) Normal vesicular connective tissue (parenchyma) from control oyster. Note
uniform cell patterns and distribution of leukocytes. XI00. (B) Vesicular connective
tissue from oyster exposed to PCS for six months. Note loss of uniform cell distribution
and infiltration by many leukocytes. XI00. (C) Normal vesicular connective tissue from
control oyster. X450. (D) Tissue from exposed oyster. Note many leukocytes and degener-
ation of vesicular connective tissue adjacent to gut epithelium. X450. (E) Normal digestive
gland tubules from control oyster. Note the thick epithelia which form normal
triradiate lumina. X450. (F) Digestive gland tubules of oyster exposed to PCB. Note
atrophy (thinning) of tubule epithelium and enlarged, abnormal lumen of tubule X450.
-------
34
D. R. Nimmo et al.
''$'~\£f'
•* • I** ^L
Fig. 11. (A) Liver parenchyma of normal spot (Lagodon rhomboides) from control tank.
Note uniform orientation of hepatocytes. X1000. (B) Liver parenchyma of spot exposed
to PCB for several weeks, intermediate pathogenesis. Note large, smooth-edged vacuoles
indicative of abnormal fatty-change in hepatocytes. XI000. (C) Liver parenchyma of
spot exposed to PCB until moribund, advanced pathogenesis. Note large vacuoles, amor-
43
-------
Toxic action of Aroclor 1254 in estuarine organisms
35
N
phous inclusions and sinusoidal congestion. XI000. (D) Hepatopancreas (digestive gland)
tubule from pink shrimp exposed to PCB. Note triangular crystalloids in some hyper-
trophied nuclei (H); also, normal nuclei (N) with large, prominent endosomes. (Feulgen
reaction; DNA appears black in photomicrograph). X1000.
44
-------
36
D. R. Nimmo et al.
Fig. 12. (A) Fresh squash of hepatopancreas from exposed shrimp. Note two crystalloids
in center. XI000. (B) Single, large crystalloid from fresh squash of hepatopancreas of
exposed shrimp. XI000. (C) Longitudinal section of hepatopancreatic duct with
branching tubules. In exposed shrimp, the crystalloids appear in the tubule epithelia
nearer the main hepatopancreatic ducts. XI000. (D) Cross-section of hepatopancreas
tubule from exposed shrimp. Note crystalloids in several epithelial nuclei; also, normal
nuclei with conspicuous endosomes. X450. (E) Intermediate size crystalloid within
hypertrophied epithelial nucleus. XI000. (F) Pathologic effect of crystalloid in PCB-
exposed shrimp. Note rupture of cells and nuclei releasing the crystalloids. X1000.
45
-------
Toxic action of Aroclor 1254 in estuarine organisms 37
and 0.62 ppb of Aroclor 1254. The tests were conducted in flowing-water aquaria as
before, with one modification. Each tank had a false floor of nylon screen to hold the an-
imals above the detritus. We believe the shrimp obtained more chemical through absorp-
tion from the water than in previous experiments.
Within the test concentrations, no threshold level existed below which shrimp did not
accumulate the chemical (Fig. 9). Concentrations produced in the shrimp (whole-body)
were about 0.2, 1.0 and 10 ppm, respectively, and these were reached between the third
and fifth weeks of exposure. Concentrations in the shrimp did not reach equilibrium
during the five-week exposure but the rate of accumulation decreased with time. When
transferred to PCB-free water, the shrimp lost most of the chemical in four weeks.
Pathology in estuarine organisms
To date, toxicology of PCBs with respect to histopathological effects in estuarine
organisms has been little studied. Published sources of information are from studies on
mammals and birds (Dahlgren etal. 1972, Vos 1972, Norback and Allen 1972) and a single
study on fishes (Couch 1972). Structural changes found in tissues of oysters, fish, and
shrimp exposed to Aroclor 1254 are characterized below.
Oysters exposed to 5.0 ppb of Aroclor 1254 for up to six months showed several
major tissue changes. Normal structural pattern of oyster vesicular connective tissue
(parenchyma) is seen in Figure 10A, and the irregular and broken pattern representative
of altered tissue from exposed animals is seen in Figure 10B. In exposed oysters, an
abnormal infiltration of leukocytes was found in the vesicular connective tissue
(Fig. IOC and 10D;. Sections of digestive glands of normal control oysters (Fig.
10E) can be compared to those in exposed oysters (Fig. 10F). The epithelia of distal
digestive tubules of exposed oysters have undergone atrophy and surround enlarged lumina.
Oysters that were removed from Aroclor-contaminated water and allowed to live in
natural water for several weeks demonstrated partial or complete tissue recovery.
Spot, an estuarine fish, exposed for two weeks or longer to 5.0 ppb of Aroclor, showed
fatty changes in their livers. Normal liver tissue has regular distribution of hepatic cells
and typical nuclei (Fig. 11 A). Note the regularity in the orientation of liver cords, cells
and uniform scattering of nuclei. In intermediate stages of liver pathogenesis in experi-
mental fish, there are extreme fatty changes characterized by the presence of large
vacuoles within hepatocytes and disorientation of liver cord distribution (Fig. 11B).
Figure 11C shows an advanced stage of pathogenesis in a moribund fish. Note the presence
of intracelJular PAS-positive bodies (ceroid) and congestion of blood sinuses, and severe
1 ucuolation.
Probably the most dramatic tissue change associated with chronic PCB exposure was
observed in shrimp. The normal hepatopancreas, or digestive gland, of shrimp is a
tightly packed organ of small elongated tubules (Fig. 12C). Figure 12D shows a cross
section through one of these tubules. In the hepatopancreas of exposed shrimp, pyramidal
-------
38 D. R. Wimmoet al.
crystalloids of various sizes were found as inclusion bodies in the nuclei of epithelial cells
(Fig. 12D and 12E). Free crystalloids are shown in Figure 12F We know of no other report
of the occurrence of precisely shaped crystalloids in hepatopancreatic tissue of crustaceans.
We have routinely studied unfixed, fresh exposed shrimp and have found crystalloids in
squashes of the tissue (See Fig. 12A and 12B).
Epithelial cells of the hepatopancreas from shrimp which were exposed to three ppb
Aroclor for at least 30 days are shown in Figure 11D. Exposed shrimp that do not have
crystalloids have no conspicuous pathologic tissue signs. In those that have the crystalloids,
hypertrophy of the affected nucleus results. Eventually, the growth of the crystalloid
inclusion distorts and ruptures the nuclear membrane. Several nuclear membranes and
inclosed crystalloids are indicated in Figure 1 ID. The crystalloids are histochemically
positive for protein. They occur in widely separated nuclei but may also appear in clusters
of adjacent nuclei and are most abundant in epithelial cells of tubules proximal to the
main hepatopancreatic ducts.
These crystalloids were found in individual shrimp before moribundity or death. In
certain exposures, crystalloids have been found in up to 80% of the survivors but the
incidence is very low over time until about 75% of the test animals have died. Crystalloids
were found more often in larger shrimp than in juveniles. At present, we are attempting to
establish whether crystalloids occur in individuals from other localities when exposed to
PCB, in other species of shrimp, or in shrimp from contaminated areas in nature.
Several possibilities exist as to the origin of the crystalloid inclusions. One suggestion
was that they may be the result of sequestering of some normal or abnormal metabolite.
Another possibility is that they represent a material produced by a virus2 and were pro-
duced under PCB stress. In reference to this suggestion, Friend and Trainer (1970) showed
that PCB enhanced the pathogenic effects of hepatitis virus in ducks.
References
Butler, P. A.: Pesticides in the marine environment: J. Appl. Ecol. 3 (suppl), 253 (1966).
Cooley, N. R., J. M. Keltner, Jr., and J. Forester: Mirex and Aroclor® 1254: Effect on
and accumulation by Tetrahymena pyriformis W. J. Protozool. 19, 636 (1972).
Couch, J. A.: Histopathologic effects of pesticides and related chemicals on the livers of
fishes. Proc. Fish Disease Symposium. Armed Forces Inst. Path., Univ. of Wisconsin
Press (In press) (1972).
Dahlgren, R. B., R. L. Linden, andC. W. Carlson: Polychlorinated biphenyls: Their effects
on penned pheasants. Environ. Health Perspect. 1, 89 (1972).
2One of us (J. C.) as a result of electron microscope studies has recently found rod-shaped virus-like
particles occluded within the crystalloid inclusion bodies. This matter is currently under study.
47
-------
Toxic action of Aroclor 1254 in estuarine organisms 39
Duke,T.W.,J. I. Lowe, and A. J. Wilson, Jr.: A polychlorinated biphenyl (Aroclor® 1254)
in the water, sediment, and biota of Escambia Bay, Florida. Bull. Environ. Contam.
Toxicol. 5, 171 (1970).
Friend, M., and D. 0. Trainer: Polychlorinated biphenyl: Interaction with duck hepatitis
virus. Science 17, 1314(1970).
Hansen, D. J., P R. Parrish, J. I. Lowe, A. J. Wilson, Jr., and P. D. Wilson: Chronic
toxicity, uptake, and retention of Aroclor® 1254 in two estuarine fishes. Bull.
Environ. Contam. Toxicol. 6, 113 (1971).
Lowe, J. I., P. R. Parrish, J. M. Patrick, Jr., and J. Forester: Effects of the polychlorinated
biphenyl Aroclor 1254 on the oyster Crassostrea virginica. Mar. Biol. 11, 209 (1972).
Lowe, J. I., P R. Parrish, A. J. Wilson, Jr., P D. Wilson, and T. W. Duke: Effects of mirex
on selected estuarine organisms. Trans. 36thN. Am. Wildl. Nat. Resour. Conf., p. 171
(1971).
Nimmo, D. R., and R. R. Blackman: Effects of DDT on cations in the hepatopancreas of
penaeid shrimp. Trans. Am. Fish. Soc. 101, 547 (1972).
Nimmo, D. R., R. R. Blackman, A. J. Wilson, Jr., and J. Forester: Toxicity and distribu-
tion of Aroclor® 1254 in the pink shrimp Penaeus duorarum. Mar. Biol. 11, 191
(1971a).
Nimmo, D. R., P D. Wilson, R. R. Blackman, and A. J. Wilson, Jr.: PolychJorinaied
biphenyl absorbed from sediments by fiddler crabs and pink shrimp. Nature 231,
50(1971b).
Norback.D. H.,and J. R. Allen: Chlorinated aromatic hydrocarbon induced modifications
of the hepatic endoplasmic reticuium: Concentric membrane arrays. Environ.
Health Perspect. 1, 137(1972).
Sodergren, A., Bj. Svensson, and S. Ulfstrand: DDT and PCB in South Swedish streams.
Environ. Pollut. 3, 25 (1972).
Vos, J. G.: Toxicology of PCBs for mammals and for birds. Environ. Health Perspect 1
105(1972).
Wildish, D. J.: The toxicity of polychlorinated biphenyls (PCB) in sea water to
Gammarus oceanicus. Bull. Environ. Contam. Toxicol. 5, 202 (1970).
Manuscript received December 12, 1973; accepted April 12, 1974
-------
Reprinted from Bulletin of
Environmental Contamination
and Toxicology, Vol. 13(2):
183-187, 1975, with permission
of Springer-Verlag, New York
Inc.
A GRAVITY-FLOW COLUMN TO PROVIDE PESTICIDE-LADEN
WATER FOR AQUATIC BIOASSAYS
Patrick W. Borthwick, Marlin E. Tagatz, and Jerrold Forester
Contribution No. 189
49
-------
A Gravity-Flow Column to Provide
Pesticide-Laden Water for Aquatic Bioassays
by PATRICK W. BORTHWICK, MARUN E. TAGATZ, and JERHOLD FORESTER
U.S. Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Sabine Island, Gulf Breeze, Fla. 32561
Associate Laboratory of the National Environmental
Research Center, Corvallis, Ore.
Traditionally, chemicals having a low solubility in water have
been dissolved in water miscible solvents (e.g. acetone, ethanol,
polyethylene glycol) before introduction into bioassay water. These
solvents act as carriers or dispersants, and allow exposure of test
animals to relatively high concentrations of the toxicant. Toxi-
city of the solvent must be considered as well as the toxic test
chemical and solvent. Concentration of solvent should be several
orders of magnitude less than that which is toxic to the test an-
imal (PARRISH, personal communication).
It is often desirable to avoid the use of a solvent, and a
pesticide typically applied as a dust, wettable powder, granule,
or bait could be tested as formulated using the method described
here.
A column containing granular pesticide, bait, or inert mater-
ial coated with pesticide may be utilized to achieve realistic con-
centrations of pesticides in assay water without using a solvent.
Column systems have produced good results in several pesticide bio-
assay experiments. CHADWICK and KIIGEMAGI (1968) showed that water
from a glass column packed with sand coated with technical dieldrin
contained fairly constant amounts of the toxicant for a period of
five months after an initial leaching period. GRAJCER (1968), used
a similar elution column that contained gravel charged with endrin.
JOHNSON (1967) utilized a glass elution column containing endrin
coated sand to provide concentrated stock solution to a special
serial-dilution apparatus. Our report shows that mirex can be in-
troduced into flow-through aquatic bioassay systems without a sol-
vent by means of a gravity-flow column containing mirex bait.
Materials and Methods
Mirex is a chlorinated hydrocarbon insecticide formulated in
bait which consists of corn cob grits (84.7 percent) impregnated
with soybean oil (15.0 percent) containing mirex (0.3 percent).
Columns (FIGURE 1) were designed to hold three layers of mi-
rex bait. An outer glass tube (50 cm length x 100 mm O.D.) with
Contribution No. 189, Gulf Breeze Environmental Research Laboratory
183
Bulletin of Environmental Contamination & Toxicology, £"1
Vol. 13, No. 2 ©1975 by Springer-Verlag New York Inc.
-------
FRESHWATER SIPHON (3OI/hrl,
CONSTANT-HEAD BOX
fFilt*r»d fruhwatvr)
NYLON SCREEN
(Retain! mir« bail)
FLOW INTO TANK
(9OI/hr
MIXING AREA
FIGURE 1. Gravity-flow Column
a glass powder funnel cemented to the bottom served as a holder
for three inner glass tubes (15 cm length x 90 mm O.D.). Nylon
monofilament screen (0.84 mm mesh) cemented to the bottom of each
inner tube retained the bait.
The bait was soaked for 24 hours in fresh water to allow
swelling before being placed in the columns. Filtered tap water
siphoned from a constant head box was percolated through three
layers of mirex bait (150 grams total). During a 4 to 8-day con-
ditioning period, column effluent bypassed tanks to avoid intro-
duction of excessive amounts of mirex into the bioassay system.
Mirex concentrations in column effluents, initially high, dimin-
ished and became more consistent after this conditioning period.
Each of three tanks received filtered tap water that had
passed through 150 grams of mirex bait. Three tanks received
effluent from columns that contained 150 grams of control bait
(with all components except insecticide). To simulate mixing in
an estuary, column effluent (30 liters/hour) and unfiltered sea-
water (60 liters/hour) were mixed in a V-shaped trough that
184
-------
emptied into each tank (FIGURE 1).
Biweekly water samples were taken directly from the tanks and
extracted with petroleum ether. Extracts were dried with anhydrous
sodium sulfate and evaporated to an appropriate volume for identi-
fication and measurement by electron-capture gas chromatography.
The limit of detection for mirex in water samples was 0.010 parts
per billion (ppb, micrograms/liter).
Range, mean, and standard error values for each experiment
are shown in Tables 1 to 3. Randomized block analysis of variance
and the Newman-Keuls range test (HICKS, 1973) were used to detect
significant differences in mirex concentrations among tanks and
experiments. Linear regression analysis was utilized to detect
significant variation within individual tanks during each 28-day
experiment.
Results and Discussion
In May 1973, a 28-day flowing-seawater experiment was com-
pleted in six 2.44 m - diameter fiberglass tanks. Mirex residues
(Table 1) in treated tank water varied between <0.010 and 0.125 ppb
over the 28-day period.
TABLE 1
Mirex concentration (parts per billion) in tank water
during first 28-day simulated estuary experiment, April-
May 1973 ( n - 9 for each tank).
TANK 123
Range <0.010-0.091 <0.010-0.110 0.014-0.125
Mean 0.03 0.04 0.04
Standard error 0.01 0.01 0.01
Tank water temperature 23.1°C (range: 19.3 to 25.2)
Salinity 13.2,1 (range: 10 to 18)
Conditioning period 8 days
185
53
-------
A second 28-day experiment was conducted in July-August 1973.
Concentrations of the pesticide in treated tanks ranged between
0.032 and 0.52 ug/fc. Data in Table 2 indicate that the columns de-
livsred higher concentrations of mirex than in the first experiment.
TABLE 2
Mirex concentration (parts per billion) in tank water during second
28-day experiment, July-August 1973 (n = 9 for each tank).
TANK 123
Range 0.032-0.36 0.043-0.23 0.053-0.52
Mean 0.10 0.10 0.16
Standard error 0.03 0.02 0.05
Tank water temperature 29.8°C (range: 28.0 to 30.8)
Salinity 15.7 X (range: 14 to 18)
Conditioning period 4 days
Although mirex concentrations in water samples fluctuated, re-
sidues differed by less than one order of magnitude.
In October-November 1973, a third experiment was conducted
(Table 3). Mirex residues (0.013 to 0.23 ug/S.) were somewhat lower
than for the summer experiment.
TABLE 3
Mirex concentration (parts per billion) in tank water during third
28-day simulated estuary experiment, October-November, 1973
(n = 9 for each tank).
TANK 123
Range 0.029-0.20 0.029-0.23 0.013-0.12
Mean 0.07 0.06 0.04
Standard error 0.02 0.02 0.01
Tank water temperature 23.4°C (range: 17.0 to 27.0)
Salinity 17.5 & (range: 15 to 19)
Conditioning period 6 days
186
54
-------
CONCLUSIONS
Concentrations of mirex among individual tanks in each test
were not statistically different at the 5-percent significance
level; whereas, differences in mirex concentrations in tank wa-
ter among experiments were significant. Paired comparisons in-
dicated statistical differences between the first and second,
and the second and third experiments, but not between the first
and third experiment. These differences in mean mirex concen-
trations in tank water may have been caused by seasonal varia-
tions in water temperature. Fluctuations in the mirex concen-
trations within individual tanks were not significant.
In its present state of development, the described gravity-
flow column is being utilized in seasonal tests to deliver mirex-
laden water to determine toxicity and uptake of mirex by several
animal species in an artificial estuarine ecosystem.
REFERENCES
CHADWICK, G. G., and U. KIIGEMAGI: J. Water Pollut. Control
Fed., 40, 76 (1968).
GRAJCER, D., Ph. D. Dissertation, Univ. Wash. 80 pp. (1968).
HICKS, C. R.: Fundamental Concepts in the Design of Experi-
ments. 2 ed. New York: Holt, Rinehart and Winston 1973.
JOHNSON, H. E., Ph. D. Dissertation, Univ. Wash. 136 pp. (1967),
PARRISH, P. R. Personal Communication, U. S. Environmental
Protection Agency, Gulf Breeze, Florida (1973).
187
55
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Reprint from Pollution and
Physiology of Marine Organisms
pp. 137-164, 1974, with
permission of the Academic
Press , New York, San Francis-
co , London
IMPLICATIONS OF PESTICIDE RESIDUES IN THE COASTAL ENVIRONMENT
Thomas W. Duke and David P. Dumas
Contribution No. 195
57
-------
Reprinted from:
POUtmON AND PHYSIOLOGY OF MARINE ORGANISMS
© 1974
ACADEMIC PRESS, INC.
Mew York Son Froncisco londoe
IMPLICATIONS OF PESTICIDE RESIDUES
IN THE COASTAL ENVIRONMENT
THOMAS W. DUKE and DAVID P. DUMAS
U. S. Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Sabine Island, Gulf Breeze, Florida 32561
Residues of pesticides occur in biological and
physical components of coastal and oceanic environ-
ments and some of the residues have been implicated
in degradation of portions of these environments.
The presence of many pesticides can be detected at
the parts-per-trillion level, but the effects of
such levels of pesticides on the organisms and sys-
tems in which they occur are not clear in many in-
stances. Knowledge of these effects is especially
important when the residues occur in the coastal en-
vironment—a dynamic, highly productive system where
fresh water from rivers mingles with salt water from
the sea. The coastal zone interfaces with man's
activities on land and, therefore, is especially sus-
ceptible to exposure to acute doses of degradable
pesticides, as well as chronic doses of persistent
ones.
This paper briefly reports the state-of-the-art
of research on the effects of pesticides on coastal
aquatic organisms. For1 a comprehensive review of
recent literature in this field, see Walsh (1972b);
137
59
-------
THOMAS W. DUKE AND DAVID P DUMAS
for a compilation of data, see the EPA Report to the
States (1973) .
Patterns of pesticide usage are changing in this
country and these changes are reflected in amounts of
various pesticides produced annually. Smaller amounts
of the organochlor^ne pesticides are being applied
because of their persistence in the environment, the
capability of organisms to concentrate them (biocon-
centration) and their adverse effects on nontarget
organisms. For many uses, organophosphates and
carbamates have replaced organochlorines because
organophosphates and carbamates hydrolyze rapidly in
water and, therefore, are not accumulated to the same
extent as organochlorines. Some of the organophos-
phates, however, are extremely toxic to aquatic
organisms on a short-time basis (Coppage, 1972).
Much effort is being devoted to developing biological
control measures that will introduce viruses and
juvenile insect hormones into the environment as part
of ark integrated pest control program. The integrated
pest control approach combines biological and chemical
methods to control pests in an effort to reduce the
amount of synthetic chemicals being added to the
environment. A list of several important pesticides
that are used currently or appear as residues in
marine organisms or both is presented in Table 1.
Samples collected in the National Estuarine
Monitoring Program and in other programs show that a
variety of pesticides occur in biota and nonliving
components of the marine environment. Pesticide
residues have been reported in .whales from the Pacific
Ocean (Wolman and Wilson, 1970), fish from southern
California (Modin, 1969) , invertebrates and fish from
the Gulf of Mexico (Giam et al., 1972), fish from
estuaries along the Gulf of Mexico (Hansen and Wilson,
1970), fauna in an Atlantic coast estuary (Woodwell
et al., 1967), zooplankton from the Atlantic Ocean
(Harvey et al., 1972), and shellfish from all three
coasts (Butler, 1973). These residues indicate that
pesticides can reach nontarget organisms in the
marine environment and give some indications of
138
60
-------
TABLE 1
Toxic Organics Used as Pesticides or Appearing as Residues in
Marine Organisms or Both
to
Organochlorines
(Insecticides)
Organophosphates
(Insecticides)
Carbamates
(Insecticides
Herbicides
Chlordane
DDT
Dieldrin
Endrin
Methoxychlor
Mi rex
Toxaphene
Diazinon
Guthion
Malathion
Naled
Parathion
Phorate
Carbaryl 2 , 4~D
Carbofuran Picloram
Triazines
Urea
-------
THOMAS W. DUKE AND DAVID P. DUMAS
biological reservoirs of pesticides in this environ-
ment. The information obtained in these monitoring
programs is invaluable to those interested in manag-
ing our natural resources, but care must be exercised
in interpreting monitoring data.
Biological problems that affect the interpreta-
tion of monitoring data were discussed recently by
Butler (1974). Factors affecting persistent organo-
chlorine residues include kind of species sampled,
age of individuals monitored, natural variations in
individuals, seasonal variation, and selection of
tissues to be analyzed. Laboratory experiments and
observations in the field have shown that filter-
feeding mollusks are good indicators of the presence
of organochlorine pesticides in estuarine waters.
These animals are sedentary, have the capacity to
concentrate the chemicals in their soft tissues many
times the concentration in the water and lose the
chemicals rather quickly when exposed to clean water.
Obviously.- mollusks would be helpful in locating the
source of a particuJar organochlorine. Conversely,-
pelagic fish might not be useful in locating a par-
ticular source because they could have accumulated a
residue some distance from the point of collection.
As patterns of pesticide usage change, techniques
for monitoring the occurrence of the pesticides ax:^
must change. Occurrences of organophosphates, carba-
mates and biological control agents cannot be moni-
tored in the same manner as occurrences of organo-
chlorine and other more persistent chemicals. To
help identify the presence of a pesticide it may be
necessary to utilize changes in biological systems,
as opposed to routine chemical analyses of organisms
or other components of the environment. Also required
is a concomitant effort to understand the effect of
residues on the organisms and systems in which they
occur.
140
62
-------
POLLUTION AND PHYSIOLOGY OF MARINE ORGANISMS
CONCEPT OF EFFECTS
The implications of pesticide residues in the
marine and other environments depends upon the effect
of the chemicals on the component in which they occur,
A conceptual model of possible effects of pesticides
and other toxic substances on biological systems is
shown in Figure 1 (Dr. John Couch, Gulf Breeze Envi-
ronmental Research Laboratory, Gulf Breeze, Florida,
unpublished personal communication). The possible
impact of a stressor on a biological system is de-
scribed as the system changes from (1) a normal
steady-state to (2) one of compensation to (3) decom-
pensation to death. Accordingly, a pesticide could
be considered to have an adverse effect if it tem-
porarily or permanently altered the normal steady-
state of a particular biological system to such a
degree as to render the homeostatic (compensating)
mechanism incapable of maintaining an acceptable
altered steady-state.
CONCEPT OF POSSIBLE EFFECTS OF
TOXIC SUBSTANCES
s
T
A
T
E
OF
B
t
O
L
O
G
C
S
Y
S
T
E
M
_
NORMAL STEADY
STATE
ALTERED STEADY
STATE
(COMPENSATION)
_
DECOMPENSATION
PT. OF NO RETURN"
DEATH-
POST-MORTEM
CHANGE
,v SUBSTANCE X
•""""l*1**!!?*!!--.. »*"""
-K*x
• » \ v% s
\ \ \ *
» I *
t * \
» *
* ' *
1
1 \
• \
* I \
i V
* * *
1 * *
1 * *
" 1 \
1 * %>l
1 4 *%
* * *
II - .1 ... . :
0
TIME
Fig. 1.
Concept of possible effects of toxic
substances.
141
63
-------
THOMAS W. DUKE AND DAVID P. DUMAS
NORMAL STEADY-STATE
It has been said that the most consistent trait
of biological systems is their inconsistency. The
normal steady-state of a particular biological system,
therefore, is difficult to define. Each system, from
an estuarine ecosystem to a system within individual
organisms, has a natural range of variability in such
factors as population density, species diversity,
community metabolism, oxygen consumption, enzyme pro-
duction, avoidance mechanisms, osmotic regulation,
natural pathogens, and others. Obviously, much must
be known about the normal or healthy system before an
evaluation can be made of the effect of a pesticide
on the system.
In relation to this, the impact of pesticides on
ecosystems is poorly understood because often the
"normal" system itself is poorly understood. An eco-
system can be considered a biological component that
consists of all of the plants and animals interacting
in a complex manner with their physical environment.
The "normal" state of a dynamic coastal ecosystem no
doubt depends upon the characteristics of a particu-
lar ecosystem, and changes as the system matures.
The importance of symbiosis, nutrient conservation,
and stability as a result of biological action in an
estuarine ecosystem is pointed out by Odum (1969) .
According to Odum, in many instances, biological con-
trol of population and nutrient cycles prevents
destructive oscillations within the system. There-
fore, a pollutant that interferes with these bio-
logical actions could adversely affect the ecosystem.
ALTERED STEADY-STATE (COMPENSATION)
An acute dose of a pesticide could cause a bio-
logical system to oscillate outside its normal range
of variation, yet with time, the system could return
to the normal state without suffering lasting effects.
An example of this phenomenon at the ecosystem level
was demonstrated by Walsh, Miller, and Heitmuller
142
64
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POLLUTION AND PHYSIOLOGY OF MARINE ORGANISMS
(1971), who introduced the herbicide dichlobenil into
a small pond on Santa Rosa Island. Applied as a wet-
table powder at a concentration of one part per
million, the herbicide eliminated the rooted plants
in the pond. As the benthic plants died, blooms of
phytoplankton and zooplankton occurred and a normal
oxygen regime was maintained. As benthic plants
returned, the number of plankters dropped. The pond
returned to a "normal" state in reference to the pri-
mary producers approximately 3 months after treatment.
A possible example of su9h compensation in an indi-
vidual organism was shown recently when spot,
Leiostomus xanthurus, were exposed to Aroclor®3 1254
under laboratory conditions (Couch, 1974). Even
though in many fish no outward signs of stress were
present, the livers of the fish accumulated excess
fat during the tests. For a period of time, the
liver evidently was able to contend with excessive
fat accumulation, but eventually chronic damage lead-
ing to necrosis occurred; therefore, the fish entered
another biological state.
DECOMPENSATION TO DEATH
The effect of a stress can eventually reach the
point where the biological system can no longer com-
pensate and death results. In the instance in which
Aroclor 1254 was related to fat globules in the liver
of fish, continued exposure to the chemical caused a
necrotic liver. Eventually, the test organisms died
as a result of the exposure. In the past, most of
the data upon which criteria and standards were based
used death as the criterion for effect. Much time
and effort now are being devoted to developing other
criteria, such as effects of 'relative concentrations
of the chemicals on tissue and cell structure, enzyme
a v£V Registered trademark: Aroclor, Monsanto Co.
Mention of commercial products does not constitute en-
dorsement by the U. S. Environmental Protection Agency.
143
65
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THOMAS W. DUKE AND DAVID P. DUMAS
reaction, osmotic regulation, behavioral patterns,
growth and reproduction.
ASSESSMENT OF EFFECTS
The concept just presented is helpful in visual-
izing the manner in which pesticides can affect
coastal organisms and systems. However, quantitative
information must be developed in order to assess the
effect of a particular pesticide .on the environment
or on a. component of the environment. For example,
it is not enough to know that a pesticide causes an
altered steady-state in a fish and eventually causes
death. The level of pesticide in the environment
that causes the effect must be known and, perhaps
even more important, the level at which no effect
occurs must be known.
Much of the quantitative information available
on effects of pesticides on marine organisms is in
terms of acute mortality of individual organisms. In
many instances, these data were obtained through rou-
tine bioassay tests in which known amounts of pesti-
cides are administered to test organisms for a given
period of time. In routine bioassays, the test
organisms are examined periodically and compared with
control organisms. If conducted for a short time in
relation to the life span of the organisms, usually
96 hrs, the tests are considered acute. Longer tests
over some developmental stage or reproductive cycles
are termed chronic. (An excellent discussion of
bioassays and their usefulness is presented by
Sprague (1969, 1970).)
Often, it is necessary to estimate the effect of
a pesticide on the coastal environment from only a
minimum amount of data. Interim guidelines sometimes
must be issued on the basis of a few acute bioassays
while more meaningful data are being obtained. An ap-
plication factor is helpful in these instances. This
factor is a numerical ratio of a safe concentration
of a pesticide to the acutely lethal concentration
144
66
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POLLUTION AND PHYSIOLOGY OF MARINE ORGANISMS
An estimate can be made for an "acceptable"
level of a pesticide in marine waters by multiplying
the LC5Q determined in acute bioassays by the appro-
priate application factor. In many instances, an
arbitrary application factor of 0.01 is used when
necessary scientific data have not yet been developed.
For a discussion on obtaining the application factor
experimentally, see Mount (1968) and Brungs (1969).
Information obtained by various bioassay tests
on some toxic organics of current interest is shown
in Table 2. These results were compiled from the
literature and indicate the most sensitive organisms
tested against these pesticides and organochlorines.
The data give a general idea of the relative toxicity
of the various pollutants.
During the past few years, the need for data on
chronic or partial chronic exposures and on sublethal
effects of pesticides on marine organisms has become
evident. Chronic studies involve the exposure of
organisms to a pesticide over an entire life cycle,
and often are referred to as "egg-to-egg" studies. A
subacute chronic is conducted over part of a life
cycle. Sublethal studies are designed to determine
if a pesticide has an effect at concentrations less
than those that are lethal to the organisms and uti-
lize such criteria as growth, function of enzyme
systems, and behavior of populations of organisms.
EFFECT OF PESTICIDES ON GROWTH OF ORGANISMS
The effects of pesticides on marine phytoplank-
ton are often related to growth of the organisms.
The effects often vary according to the pesticide and
to the species of phytoplankton. For example, Menzel
et al. (1970) found that growth in cultures of marine
phytoplankton was affected .oy DDT, dieldrin and
endrin. Dunaliella apparently was not affected by
concentrations up to 1000 parts per billion. In
Cyclotella, cell division was completely inhibited by
dieldrin and endrin and DDT slowed division of the
cells. The authors suggested that estuarine species,
145
67
-------
TABLE 2
Toxicity of Selected Pesticides to Marine Organisms
inirrt icldes
Org tv .oct i lo r i nt- :. :
D2T C^npour.ds
L'.E '-DDTfl,3 ,1- T-c.l-nic^il
Ip-cKcruphenyU
Pi chlotv-2, ?-bis
t til a ne
p,i-'-aoF (3 ,i-
Dlch!orn~2,?-bis
(p-c'-'lorophenyl)
ethylene
Endrin 100\
100*
Kethoxychlor 89. 5%
Mir^x Technical
Toxaphcne 1CO\
Cone, (ppb
in Water
(HV )M1
! en-?'. :-;j.- d.ji , -IZU.T "inX shririp 0. ] 2
-------
TABLE 2—Continued
Toxicity of Selected Pesticides to Marine Organisms
Substance Tested
Insecticides
Grganophosphates :
Diazinon
Guthion
Foitnulation Organism Tested
Technical Cyprinodon varj-rjatus
Grade
«*
934 GastsrosSeus tculeatas
Technical Cyprinodan varieqatus
Grade
*r?:ade"Ca 9°U^n
'
Grade.
Grade "
Technical Lagodvn rhomboides
Grade
,
Grade
100* Thalassona bS.f64%)
Threespinc 4, ft run
Shespshead mi.-.nc-w J Mean inhibition
of brain AChfi
(Pusult: d4^)
of brain AChE
(Rpsult : 004)
Spot 20 M'lan inhibi t ion
of brain AChE
of brain AChE
(Result: •>d4'»}
Pijflish 238 Meun inhibition
of brain AChK
(Result: BSt)
of brain AChE
(Result: 701)
Bluohead 27 " LC-50
Tsst Procedure
St^ci-c bioassay,
4f?"hr LC 4O-60
S-5-ht static lab
Static bioassay,
72 -hr 1-C 40-GO
ossiy, ;-i-hr LC 40-GO
Fl'.'|Ji'i'-f J.;P.IV, ,1 1 cr bio-
a --3 ay, ^4-hr IJ3 40-50
24-hr LC 40-60
Flowing scawatPT bio-
asFr.y, 24-hr I/: 40~(-0
assay, 24-hr LC 40- CO
O'i-hr static lab
bioassay
Raft- rcnce
Ccppage, 1972
Kata, 1961
•,-oJ^aye, 1^72
Matthews, 1074
Copiiev/c ar.d
Matthews, 1974
Coppi-iqe and
Mat thews, I97J
Matthews, 1074
Kislcr, 1970
-------
TABLE 2—Continued
Toxicity of Selected Pesticides t() Marine Organisms
00
Insecticides
Organophosphates :
Haled
Para th ion
Methyl
Parathion
Phorata
Carbonates:
Carbaryl
Technical
Grade
Technical
Gr-^c
Technical
Grade
Technical
Grade
Technical
Grade
loot
Technical
Grade
lOOi
Technical
Grade
Lagodon rhomboides
Lf-iostonnis xanthurus
djprinodon variey&tus
Lagodon rhomboides
Leiostomus xanthurus
Crangon sept ems pi nos a
Ctfprlnodon varJ egatas
Palaeioon toacrodactylus
La godon rhooboS dea
Cone . (ppb
in Water
Pinfish 23
Spot 70
Shoepshead minnow 10
Pinfish 10
Spot 10
Sand shrjmp 2
Sheepshead minnow S
Korean shrimp 7.0
U.S-28)
Pinfish 1333
Method
i ) Of
Assessment
Mean inhibition
of brain r.ChE
(Pesult: fi9\>
Mean inhibition
of brain AChE
(Result: 051)
Mean inhibition
of brain AChE
(Resullt 84%)
Me*n inhibition
of brain AChE
Mean inhibition
of brain AChE
{Result: 90%)
LC-50
Mean inhibition
of brain AChE
(Result: >84\)
TL-50
Mean inhibition
of brain AChE
(Result: 81%)
Plowing suflwater bJo-
assay, 7,">-hr 1C 40-60
flowing senwater bio-
assay, 24-hr 1£ 40-60
Static bioafisay.
72-hr LC 40-60
Flowing seawDter bio-
assay, 24-hr 1C 40-60
Flowing seawater bio-
asaay, 24-hr LC 40-60
96-hr static lab
Static bioasstiy
72-hr LC 40-60
96-hr intermittent
flow lab bioasnAy
Flowing seawater bio-
assay, 24-hr LC 40-60
Coppiwie and
Motthovs, .1974
Copjiflqe and
Matthews, 1974
C'.ippau-. 1972
Coppng and
Matthews, 1974
Coppage ajid
Mdtthewc, 1074
Eisler, 1969
Coppage, 1972
Earnest,
unpublished
Coppage/
unpublished
-------
TABLE 2—Continued
Toxicity of Selected Pesticides to Marine Organisms
Substance Tested
Insecticides
Carbamates:
Carbofuran
Herbicides
2,4-D and
derivatives
PicloraiR
Tordon ® 101
(39.61. 2,4-D
14.3% piclrrAin)
Amctrync-
Atrazir.e
Formulation Organism Tested
Acetone Cyprinorfon variegatus
wash iirom
sand-coated
particle
formulation
Ester Crassostrca virtjinica
fsochrysis galbana
Technical ChlorococfuKi sp .
o rid
Technical Zsochrysis galbana
acid
Technical Honochrys.fs Jutheri
acid
Tpchnical Pnapo^actyJum
acid tricornutum
Technical Chlorococcura sp.
acid
TetTinical ChLaaiydomon&s sp.
. acid
acid
Cone, (ppb
Common Home Act. Tngrcd.
in Water
5ho(*pshci-'':l minnow Unknown
American oystGr 740
S * 10s
lo
10
10
10
100
60
77
Method
.) of
Assessment
Mean inhibition
of brain AChE
(Result: 84* )
TLH
50t decrease in
O2 pvolution3
growth
SO*, decrease in
O2 evolution3
50* decrease in
02 evolution0
50* decrease in
02 evolution3
50% decrease in
growth
50% decrease in
Oj evolution3
50% decrease in
02 evolution3
Test Procedure Reference
Static bioassay Coppage*
48-hr LC 40-60 unpublished
14-day static lab Davis and Hidu,
bioassay 1965
Waish, 1972a
t525mu) after 10 daysb
Walsh. 1972a
Walsh, 1972a
Walsh, 1972a
Measured as ABS. Walsh, 1972a
(525mp) after 10 daysb
— - Jtollister and
Walsh, 1973
Holl-ister and
Walsh, 1973
-------
TABLE 2—Continued
Toxicity of Selected Pesticides to Marine Organisms
N)
Cone - ( ppb He thod
Snbstance rested Formulation Organisms Tested Conroon Kane Act. Ingred, ) of
Iftbicides
Triazines -
»trazir.e
.
Urea:
Diuron
Technical
acid
Technical
"rid
arid
Technical
Technical
acid
1
Isachrysis galbana
Phaeod<*ctyJum
trico: nut am
Fhodacty]iJm
Frotococcus sp.
Honorf-rysis JutherJ
ChjorococcuFp sp.
Isochrysj's galb&na
Wonochrys J s 1 utheri
ICO 50* decrease in
O^ evolution3
100 50% decrease in
02 evolutiona
growth
growth
U.02 0.52 OPT. DEN.
expt/OPT. DEN.
cnntrolb
" 02 0.00 OPT. DEN.
expt/OPT. DEN.
10 501 decrease in
growth
10 50\ decrease in
growth
290 0.67 OPT. DEN.
expt/OPT. DEN.
controlb
Test Procedure
(525mp) after 10 daysb
(525tnu) after 10 days^1
10-day growth
10-day growth
10-day growth
10-day growth
10-day growth
Reference
Walsh, 197Ja
Walsh, 1972a
Walsh 1972a
Walsh, 197?a
UXelos, 1962
Ukolcs, 1962
Walsh, 1972a
Walsh, 197Ja
Ukeles, 1962
"Oj evolution measured by Gilson differential rcspirometer on 4 mt of culture in log phase. Length of test 90 min.
bABs. (525TBU) - Absorbance at 525 millimicrons wavelength. OPT. DEN. expt/OPT. DEN. control » Optical density of experimental culture/optical density
of control culture.
-------
POLLUTION AND PHYSIOLOGY OF MARINE ORGANISMS
such as Dunalislla, a.-c perhaps less susceptible than
are open ocean forms, such as CycJ.ote.LZa. Similar
studies on phytoplanktqn and PCBs by Fisher et al.
(1972) also suggested that coastal phytoplankton may
be more resistant to organochlorines than are those
found in open ocean. Isolates of diatoms from the
Sargasso Sea were more sensitive than clones from
estuaries and the continental shelf. Herbicides ap-
plied to four species of marine unicellular algae
adversely affected their growth (Walsh, 1972a). Urea
and triazine herbicides wore the most toxic of the
formulations tested. In some instances, smaller
amounts of herbicides were required to inhibit growth
than to inhibit oxygen evolution. Interestingly,
Dunaliella was most .resistant of the four species
tested, as occurred in Menzel's et al. studies (1970).
The effect cf m;; rex and a FOB, Aroclor 1254, on
growth of ciliate, T&tz&hymena p'jritornds, was stud-
ied by Ccoley et al. (lr*72) . Both chemicals caused
significant reduction'. i.n growth rate and population
density and the cilia'ce accumulated both toxicants
from the culture media, concentrating mirex up to
193 times and Aroolor to :-.pproxi:nately 60 times the
nominal concentration ir. the media. The authors pos-
tulate that if this ciliats encountered similar con-
centrations of these materials in nature, the results
would be a reduction of their availability as food
organisms and nutrient regenerators. Also, the
capacity of the organisms to concentrate mirex and
Aroclor could provide a ca^-hway for entry of these
chemicals into thi J;--d web.
Growth rates r-f young c-vRteis. Crassostrea
virginica, as indicated by he,'.-cat and in-water weight,
was significantly I'educad in individuals exposed to
5 micrograms of Aroclor 1254 uer lit^r (ppb) for
24 weeks, but grew--12 rate w- • not affected in indi-
viduals < ^posed to 1. part p-,,.- billion for 30 weeks
(Lowe et &'. . f 197^). Oysters exprsr-ic, to 1 part per
billion concentrated the chemical 101,000 times, but
less than 0.2 part per r.dilion remained after 12
weeks of depuration. Ine growth rate of the oyster
73
-------
THOMAS W. DUKE AND DAVID P. DUMAS
was a much more sensitive indicator/ since no sig-
nificant mortality occurred in oysters exposed to
5 ppb.
The effects of mirex on growth of crabs, as
measured by the duration of developmental stages of
crabs as an indicator of their growth, is illustrated
by the work of Bookhout et al. (1972). The duration
of developmental stages of zoea and the total time of
development was generally lengthened with an increase
in concentration of mirax from 0.01 to 10.0 parts per
billion. Menippe did not demonstrate this effect,
but the percentage of the extra ,6th zoeal stage in-
creased as concentrations of mirex increased. This
method of determining the effect of mirex on crabs
appears to bs more sensitive than previous tests with
juvenile blue crr^s rfcported by McKenzie (1970) and
Lowe et al. (1971).
EFFECTS OF P3ST7CIDES ON BEHAVIOR OF ORGANISMS
•rhe behaviors 1 activity c;_ organisms is a sensi-
tive criterion for determining the effect of pesti-
cides on morii.3 org-an^iuS. Dr. H. G. Kleerekoper has
successfully studied t.'ie interactions of temperature
and a heavy mot'-? I on the locotnotor behavior of fish
in the laboratory (Klperekoper and Waxman, 1973) and
will present data on Lhe effect of pescicides on
marine fish later in this volume. Hansen (1969)
showed that the estuarine fish, Cyprlnodon variegatusf
avoided water__ containing DDT, enarin, Dursban ®^ or
2,4-D in control]eI laboratory experiments, but the
fish did not avoid t 3t concentrations of malathion
or Sevin ™ .c Likewise grass shrimp, Pslaemonetes
pugio, an important forage food for estuarine organ-
isms, avoided 1.0 c.ad 10.0 ppm of 2,4«D by seeking
©Registered trademark; Dursban, Pow Chemical
Company.
c® Registered trademark: Sevin, Union Carbide
Company.
152
74
-------
POLLUTION AND PHYSIOLOGY OF MARINE ORGANISMS
water free of this herbicide, but did not avoid the
five insecticides tested (Hansen et al., 1973). The
capacity of coastal organisms to avoid water contain-
ing pesticides may enhance their survival by causing
them to move to an area free of pesticides. Avoid-
ance could be disastrous to a population if, by
avoiding the pesticides, the population is unable to
reach an area where spawning normally occurs.
EFFECTS OF PESTICIDES ON ENZYME SYSTEMS
Inhibition of the hydrolyzing enzyme, acetyl-
cholinesterase (AChE), by organophosphate and carba-
mate pesticides can be used as an indication of the
effect of these chemicals on estuarine fish (Coppage,
1972) . Evidently, esterase-inhibiting pesticides
bind active sites of the enzyme and block the break-
down of acetylcholine, which causes toxic accumula-
tion of acetylcholine. As a result, nerve impulse
transfers can be disrupted. Laboratory bioassays
with estuarine fish soot, Leiostomus xanthurus,
showed that lethal exposures of this fish to malathion
reduced the AChE activity level by 81%. Such informa-
tion developed in the laboratory is useful in evalu-
ating effects of pesticides applied in the field.
EFFECTS OF PESTICIDES ON ECOSYSTEMS AND COMMUNITIES
Few data are available concerning the effects of
pesticides at the ecosystem or community level of
organization. This is not surprising considering the
complexities of ecosystems and ou~ lack of knowledge
of the structure and function of coastal zones.
Effects of pesticides could .be masked by variations
in population densities and it would require several
years to evaluate such variations. However, it is
possible to design laboratory and field experiments
to yield information on this complex system.
An experimental community that received 10
micrograms per liter of a polychlorinated biphenyl,
Aroclor 1254, did not recover to a "normal" state in
153
75
-------
THOMAS W. DUKE AND DAVID P. DUMAS
terms of numbers of phyla and species after 4 months
(Hansen, 1974). Communities of planktonic larvae
were allowed to develop in "control" aquaria and
aquaria that received the Aroclor 1254. Communities
that received 10 micrcgrams per liter of the chemical
were dominated by tunicates, whereas controls were
dominated by arthropods. The Shannon-Weaver species
diversity index was not altered by Aroclor 1254, but
numbers of phyla, species and individuals decreased.
The capacity of a fish population to compensate
for the effect of a pesticide was suggested in a
recent study made in Louisiana, where malathion was
applied aerially to control mosquito vectors of
Venezuelan equine encephalomyelitis (Coppage and
Duke, 1972). Fish were collected from the coastal
area before, during and after the application of
malathion. Acetylcholinesterase (AChE) activity in
the brains of fish were used as an indicator of the
effect of malathion on the community of fish. Levels
of inhibition during and soon after spraying in one
lake approached levels that were associated with
death of fish in laboratory bioassay studies. The
AChE activity of the fish population returned to
normal within 40 days after application of the
chemical.
CONCENTRATION FACTORS
The capacity of organisms to concentrate a pesti-
cide is another factor that must be considered when
evaluating the impact of these chemicals on a coastal
system. Many of the persistent pesticides are passed
through the food web through accumulation and bio-
concentration. Some question exists about the mechan-
isms involved in trophic accumulation of fat-soluble
hydrocarbons from water by aquatic organisms (Hamelink
et al., 1971). Whatever the mechanisms for accumula-
tion, many coastal organisms have the capacity to
concentrate pesticides many times more than the con-
centration occurring in the water around them. Con-
centration factors, the ratio of the amount of
154
76
-------
POLLUTION AND PHYSIOLOGY OF MARINE ORGANISMS
pesticide in the animal to that in the water, for
some specific organisms and pesticides determined by
investigators at the Gulf Breeze Environmental
Research Laboratory are shown in Table 3.
STATE OF THE ART
Concern about the occurrence of pesticides in
the marine environment is continually emphasized
because surveillance and research on these chemicals
are given high priority by knowledgeable scientists.
The analytical capability for determining residues of
some pesticides in the parts per trillion range is
available, but we often do not understand the bio-
logical or ecological significance of these residues.
We need more information on chronic exposures of
sensitive marine organisms during complete reproduc-
tive cycles and on effects of sublethal levels of
exposure. Also, information is required on the
structure and function of coastal ecosystems and
criteria for evaluating the stress of pesticides on
these systems. Laboratory microcosms and other kinds
of experimental environments no doubt will be useful
in this evaluation.
As mentioned previously, use-patterns of pesti-
cides in this country are changing. We must be pre-
pared to evaluate possible effects, on the environ-
ment, of integrated pest control procedures, whereby
biological control may be just important as chemical
control of pests. Viruses and juvenile-hormone
mimics are being .tested for use as pesticides and
could inadvertently reach the coastal zone. The
research effort to evaluate the impact of these new
agents must take into account that the coastal
environment already contains residues of pesticides,
persistent organochlorines, and other pollutants.
155
77
-------
TABLE 3
Accumulation of Pesticides from Water by Marine Organisms9-
oo
tn
Substance Tested
Insecticides
Organochlorines :
Chlordanc
DDT
Dieldrin
Endrin
H
oxyc
Mirex
Organism Tested.
Psftudomonas spp.
Brachi don tes rccurvus
Mcrcenar-fa n*?rcc'nar.ia
Kya ir»«rja
Crassostrea gigas
Penaeus duoraru/n
Id got/on rfcomlxjides
Hsrcvn&ria /nercr/iflria
«fircenarja m.'rr.-snaria
rftrahymenj pyriformis H
penacus dunrarunt
Common Nome
llooKud mussel
Hard-shell clam
Soft-shell clam
Pacific oyster
Pink shrimp
Plnfish
Ilard-shpll clnm
Hard-shell clam
Pink shrimp
Exp. Cone.
10 ppnt
1 PPb
1 ppb
0.1 ppb
1.0 ppb
0.14 ppb
0.1, 1.0
ppb
0 . S ppb
0.5 ppb
PP
0.9 ppb
0.1 ppb
Cone. Factor Time
0.83
24,000
6,000
8.000
20,000
1,500
10,600
3B,000
760
400
470
193
2,600
10
1
1
5
7
J
2
S
S
5
1
3
days
week
week
days
days
weeks
weeks
days
days
ys
week
weeks
Special Uutails
Mixed culture of
four species
Whole body residues
(Meats)
Whole body residues
(Heats)
Whole body roaidues
(Moats)
Whole body residues
(Meats)
whole body residues
Whole body residues
Whole body residues
(Meats)
Whole body residues
(Heats)
(Meats)
ftxenic cultures
incubated at 26°Cj
concentration
factor on dry
weiqht basis
Whole body residues
Reference
Bourquin, unpublished
Bntler,
Butlsr,
Butler,
Butler,
1966
19&6
1971
1966
Nimmo ft aJ, , 1970
Hflnsen
1970
Butler,
Butler,
Butler
Cooley
and wilaon,
1971
1971
1971
A3 ' A
et al. . 1972
lowo st al. , 1971
-------
TABLE 3—Continued
Accumulation of Pesticides from Water by Marine Organismsa
Substance Tested Organism Tested Common Name Exp. Cone- Cone. Factor Time Special Details Reference
Insecticide?
Organochlori nes: •
Mirex Khi thropanopsus Mud crab (larvae) 0.1 ppb 1,000 7 weeks Static culture bowl Bookhout et al.f 1972
harrisii method with a
change to fresh
i, medium + chemical
-------
130
COPPAGE AND MATTHEWS
hr exposure in the laboratory to a nominal concentration of 100 /^g/liter (Weiss, 1961).
Brain-AChE of a fish species in an estuary sprayed with malathion was inhibited and
remained inhibited 50 days after spraying was discontinued (Coppage and Duke,
1971). Relatively irreversible AChE inhibition apparently occurs.
METHODS
The AChE of the pinfish brain was characterized and assayed with a pH-stat method
previously described (Coppage, 1971). Normal brain AChE activity was determined
with pinfish (65-125 mm total length) taken randomly from wild populations over a
2-year period. Normal AChE activity of 31 samples (each sample was an homogenate
of pooled brains taken from five pinfish) had a coefficient of variation of only ±9 %
for the period. Each assay sample for AChE inhibition consisted of pooled brains taken
from 4-6 fish that survived naled exposure for a designated time, and percentage
inhibition was determined by comparison with mean normal activity. Dead fish were
not used to interpret AChE inhibition because it cannot be applied in practical field
studies where it is not known how long fish have been dead and subject to loss of AChE
activity due to protein destruction.
In each test, 4-12 replicates of 10 fish each were exposed to technical grade pesticide
in 8-liter acrylic plastic aquaria that received a mixture of flowing seawater (400 ml/min)
and naled from a common source. The naled was dissolved in benzene and infused
into seawater by means of a syringe pump. Solvent infusion never exceeded 2.5 mg/liter
of water. Benzene did not significantly affect AChE activity offish exposed to 8 mg/liter
for 72 hr (Table 1). Pesticide concentration in the water was expressed as fig added per
TABLE 1
BRArN-ACETYLCHOLINESTERASE ACTIVITY IN CONTROL PlNFISH AND POPULATIONS SUBJECTED
TO SUBLETHAL AND LETHAL EXPOSURE TO NALED
Percent less
AChE
Nominal
concentration
(^g/liter)
Control
8000 benzene only
15
15
15
15
25
55
75
Hours
exposed
72
24
48
72
96
72
48
24
Percent
killed
0
0
0
0
0
40-60
40-60
40-60
Number
of
samples
31
3
4
4
4
4
4
4
3
Mean
AChE
activity"
2.03
2.166
0.84C
0.77°
0.72C
0.60C
0.23d
0.33"
0.24"
SD
0.18
0.10
0.11
0.19
0.10
0.13
0.03
0.10
0.00
activity
than
control
59
62
65
70
89
84
88
" Expressed as ^mol of acetylcholine hydrolyzed/hr/mg brain tissue.
6 Not significantly different from control.
c Significantly less activity than controls (p < 0.001).
d Significantly less activity than fish exposed to 15 ^g/liter for same period (p < 0.01).
91
-------
BRAIN-AChE INHIBITION BY NALED
131
liter. No attempt was made to chemically analyze naled or its transformation products
in the water. Temperature range was 18-23°C, and salinity was 23-29 parts per thousand
during the tests.
To determine the extent of AChE inhibition resulting from a near-median kill, we
assayed the survivors of test in which 40-60 % of the test populations were killed by
exposure to naled in 24, 48, and 72 hr. Brain-AChE activity was measured at 24 hr
for the 24-hr lethal exposure (75 ng naled/liter), 24 and 48 hr for the 48-hr lethal exposure
(55 [j.g naled/liter), and at 24, 48, and 72 hr for the 72-hr lethal exposure (25 /j,g naled/
liter). This was accomplished by exposing several groups of fish, in separate aquaria,
to the same source of naled in flowing seawater. At each specified time interval, 3-4
replicate groups of 4-6 fish each were taken from the replicate aquaria and their
brain-AChE was measured. Also, we exposed other groups of pinfish for 96 hr in
water to which 15 pg naled/liter were added and measured their brain-AChE activities
at 24,48,72, and 96 hr.
RESULTS AND DISCUSSION
Inhibition data for fish, expressed as percentage reduction of AChE activity when
compared with mean normal activity, are summarized in Fig. 1. Statistical comparisons
of AChE activities offish exposed to lethal concentrations were made with fish exposed
to the sublethal concentration (Student's t test, p < 0.01) (Table 1).
1OO
24 48 72
HOURS EXPOSURE
FIG. 1. Inhibition of brain-acetylcholinesterase activity by sublethal and lethal exposure to naled.
Each experimental point represents the mean of 3-4 replicate tests. The amounts of naled added to the
seawater, in //g/liter, for a particular test or test sequence are given in parentheses. One SD from normal
during a 2-year sampling of wild fish populations is indicated by dashed lines.
For the same length of exposure, lethal concentrations of naled always produced a
significantly greater inhibition of brain-AChE than sublethal concentrations (Table 1).
Mean reductions of AChE activity in all lethal exposures that killed 40-60% of the
test populations were similar (84-89 %), regardless of concentration of naled or length
5*
92
-------
of exposure. The mean inhibition caused by 15 jig/liter (sublethal) was greater in each
succeeding 24-hr period reaching a maximal inhibition of 65 % in 72 hr and 70 % in
96 hr. Thus, in exposures of up to 72 hr, brain-AChE inhibition in excess of 83 %
indicates a high probability of impending death in an exposed population. In the
"sublethal" exposure the increasing inhibition with time and the mean reduction of
70 % at 96 hr suggest that, if exposure continued, a lethal level of AChE inhibition may
occur. These findings strongly support earlier findings for other fish species (Coppage,
1972; Alsen et al, 1973; Coppage and Matthews, 1974) that brain-AChE inhibitions
of about 70-80% are critical in short-term lethal poisoning by organophosphate
insecticides. The relatively specific levels of AChE inhibition during "kills" show that it
is unnecessary to rely on the dubious interpretation of residues alone to determine
poisoning and cause of "kills" in the environment. Correlation of AChE inhibition
with insecticide usage or residue analysis should be sufficient to establish identity of the
compound or compounds causing poisoning and "kills" of fishes in aquatic systems.
The large AChE inhibitions (59-65%) caused by "sublethal" exposure indicate
pollution can be readily detected, possibly before acute poisoning occurs. However,
even if acute lethal poisoning does not occur during organophosphate insecticide
exposure, depression of AChE in vertebrates may cause physiological and behavioral
modifications (Koelle, 1963; Karczmar et al., 1970) that reduce animal survival ability.
Cumulative reduction of AChE by repetitive exposure to some organophosphate
pesticides has been demonstrated in some vertebrates (Heath, 1961; Koelle, 1963;
Karczmar, 1970), and this possibly happens in fish repetitively exposed in the environ-
ment. For example, repetitive exposures of fish to short-term "sublethal" concentra-
tions of azinphosmethyl has resulted in deaths (Lahav and Sarig, 1969). This may be of
significance because anticholinesterase mosquito control chemicals are often applied
to marshes several times a month.
REFERENCES
ALDRIDGE, W. N. (1971). The nature of the reaction of organophosphorus compounds and
carbamates with esterases. Bull WHO 44, 25-30.
ALSEN, C., HERRLINGER, A. AND OHNESORGE, F. K. (1973). Characterization of cholinesterases
of the cod (Gadus callarias) and their in vivo inhibition by paraoxon and tabun. Arch.
Toxikol. 30, 263-275.
BUTLER, P. A. (1963). Commercial fisheries investigations. In Pesticide-Wildlife Studies
during 1961 and 1962, pp. 11-25. U. S. Fish Wildlife Serv. Circ. 167, Washington, DC.
BUTLER, P. A. (1965). Commercial fisheries investigations, In The Effects of Pesticides on Fish
and Wildlife, pp. 65-77. U. S. Fish Wildlife Serv. Circ. 226, Washington, DC.
CARTER, F. L. (1971). In vivo studies of brain acetylcholinesterase inhibition by organo-
phosphate and carbamate insecticides in fish. Ph.D. dissertation, Louisiana State Univer-
sity, Baton Rouge, LA.
COPE, O. B. (1963). Sport fishery investigations. In Pesticide-Wildlife Studies during 1961
and 1962, pp. 26-42. U. S. Fish Wildlife Serv. Circ. 167, Washington, DC.
COPE, O. B. (1965). Sport fishery investigations. In The Effects of Pesticides on Fish and Wildlife,
pp. 51-63. U. S. Fish Wildlife Serv. Circ. 226, Washington, DC.
COPPAGE, D. L. (1971). Characterization offish brain acetylcholinesterase with an automated
pH stat for inhibition studies. Bull. Environ. Contam. Toxicol. 6, 304-310.
COPPAGE, D. L. (1972). Organophosphate pesticides: Specific level of brain AChE inhibition
related to death in sheepshead minnows. Trans. Amer. Fish. Soc. 101, 534-536.
93
-------
BRAIN-AChE INHIBITION BY NALED 133
COPPAGE, D. L. AND DUKE, T. W. (1971). Effects of pesticides in estuaries along the Gulf and
Southeast Atlantic Coasts. In Proceedings of the 2nd Gulf Coast Conference on Mosquito
Suppression and Wildlife Management (C. H. Schmidt, Ed.), pp. 24-31. National Mosquito
Control-Fish and Wildlife Management Coordinating Committee, Washington, DC.
COPPAGE, D. L. AND MATTHEWS, E. (1974). Short-term effects of organophosphate pesticides
on cholinesterases of estuarine fishes and pink shrimp. Bull. Environ. Contam. Toxicol.
11, 483^88.
DUPUY, A. J. AND SCHULZE, J. A. (1972). Selected Water-Quality Records for Texas Surface
Waters, 1970 Water Year. Texas Water Development Board, Report 149, Austin, TX.
EHRENPREIS, S. (Ed.) (1967). Cholinergic mechanisms. Ann. N. Y. Acad. Sci. 144, 385-935.
FOWLER, D. L. AND MAHAN, J. N. (1971). The Pesticide Review 1971. U.S. Department of
Agriculture, ASCS, Washington, DC.
HEATH, D. F. (1961). Organophosphorus Poisons, Anticholinesterases and Related Compounds.
Pergamon Press, New York.
HOLLAND, H. T., COPPAGE, D. L. AND BUTLER, P. A. (1967). Use offish brain acetylcholinesterase
to monitor pollution by Organophosphorus pesticides. Bull. Environ. Contam. Toxicol. 2,
156-162.
KARCZMAR, A. G. (Ed.) (1970). Anticholinesterase Agents. Pergamon Press, New York.
KARCZMAR, A. G., NISHI, S. AND BLABER, L. C. (1970). Investigations, particularly by means
of anticholinesterase agents, of the multiple peripheral and central cholinergic mechanisms
and of their behavioral implications. Acta. Vitaminol. Enzymol. 24, 131-189.
KOELLE, G. B. (Ed.) (1963). Cholinesterases and Anticholinesterase Agents, Springer-Verlag,
Berlin.
LAHAV, M. AND SARIG, S. (1969). Sensitivity of pond fish to cotnion (azinphosmethyl) and
parathion. Bamidgeh. Bull. Fish. Cult. Isr. 21, 67-74.
LAWLESS, E. W., VON RUMKER, R. AND FERGUSON, T. L. (1972). The Pollution Potential in
Pesticide Manufacturing. U. S. Environmental Protection Agency, Washington, DC.
MACEK, K. J., WALSH, D. F., HOGAN, J. W. AND HOLTZ, D. D. (1972). Toxicity of the insecticide
Dursban to fish and aquatic invertebrates in ponds. Trans. Amer. Fish. Soc. 101, 420-427.
MAYER, F. L., JR. AND WALSH, D. F. (1970). Multiple exposures of bluegills and aquatic
invertebrates to Abate. U. S. Bur. Sport Fish. Wildl. Resour. Publ. 106, 16.
MURPHY, S. D., LAUWERYS, R. R. AND CHEEVER, K. L. (1968). Comparative anticholinesterase
action of Organophosphorus insecticides in vertebrates. Toxicol. Appl. Pharmacol 12
22-35.
O'BRIEN, R. D. (1960). Toxic Phosphorus Esters, Chemistry, Metabolism and Biological
Effects. Academic Press, New York.
O'BRIEN, R. D. (1967). Insecticides, Action and Metabolism. Academic Press, New York.
PIMENTEL, D. (1971). Ecological Effects of Pesticides on Non-Target Species. Executive Office
of the President, Office of Science and Technology, Washinton, DC.
PINKOVSKY, D. D. (1972). United States Air Force aerial spray activities in operation combat
VEE. Mosq. News 32, 332-334.
SANDERS, H. O. AND COPE, O. B. (1966). Toxicities of several pesticides to two species of
cladocerans. Trans. Amer. Fish. Soc. 95, 165-169.
WEISS, C. M. (1961). Physiological effect of organic phosphorus insecticides on several species
offish. Trans. Amer. Fish. Soc. 90, 143-152.
WILLIAMS, A. K. AND SOVA, R. C. (1966). Acetylcholinesterase levels in brains of fishes from
polluted waters. Bull. Environ. Contam. Toxicol. 1, 198-204.
94
-------
Reprinted from Proceedings of
the 28th Annual Conf. of
Southeastern Assoc. of Game
and Fish Comm., Nov. 17-20,
1974, pp. 392-398, with per-
mission of the Southeastern
Assoc. of Game and Fish Comm.
SHEEPSHEAD MINNOW (CYPRINODON VARIEGATUS): AN ESTUARINE FISH SUITABLE
FOR CHRONIC (ENTIRE LIFE-CYCLE) BIOASSAYS
Steven C. Schimmel and David J. Hansen
Contribution No. 205
95
-------
Reprinted from the Proceedings of the 28th Annual Conference of the Southeastern Association of
Game and Fish Commissioners, 1974.
SHEEPSHEAD MINNOW (CYPRINODON VARIEGATUS):
AN ESTUARINE FISH SUITABLE FOR
CHRONIC (ENTIRE LIFE-CYCLE) BIOASSAYS1
by
Steven C. Schimmel and David J. Hansen
U.S. Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Sabine Island, Gulf Breeze, Florida 32561
(Associate Laboratory of the National
Environmental Research Center, Corvallis, Oregon)
ABSTRACT
The sheepshead minnow (Cypnnodon variegatus), an estuarine fish of the Atlantic and Gulf Coasts, is suitable for both partial
chronic and chronic (egg-to-egg) bioassays. The fish is easily held at high population densities in the laboratory and, at about 30
C, produces numerous eggs. The average 30-day survival of the fish from fertile egg to fry is 75% Generation time for this species
is short (3-4 months) and its small adult size (male average standard tength=48mm) provides for relatively inexpensive bioas-
says. This killifish's susceptibility to organochlorine toxicants is similar to that of other estuarine fishes tested and thus should
produce significant information on the effects of these toxicants on the estuanne community.
INTRODUCTION
Acute, partial-chronic and chronic bioassays are necessary for setting water quality
standards, according to Mount and Stephans' (1967) definition of maximum accep-
table toxicant concentration and experimental definition of application factor.
Partial-chronic bioassays have been accomplished on several fresh-water species such
as the bluegill (Eaton, 1970) and brook trout (McKim and Benoit, 1971) in which
effects of toxicants were observed on each life stage. In chronic bioassays, the test
organisms are exposed to a toxicant during their entire life cycle to measure effects on
survival, growth and reproduction. In this manner, the most susceptible life stage can
be ascertained and the survival potential of future generations of the organism es-
timated. Fresh-water chronic bioassays have been completed by several investigators
such as Brungs (1971), using the fathead minnow (Pimephales promelas). To our
knowledge, no marine or estuarine fish has been used in chronic or partial-chronic
bioassays.
There are several criteria to be considered when choosing a fish for a chronic bioas-
say:
1. The fish should be able to reproduce readily in close confinement, producing
large numbers of eggs.
2. Fertility as well as survival to adulthood should be high.
'Contribution No. 205, Gulf Breeze Environmental Research Laboratory.
392
97
-------
3. The organism should mature rapidly, yet be small enough at adult size to main-
tain large, statistically-valid numbers of fish in the bioassay.
4. The fish should be relatively sensitive to toxic pollutants.
LIFE HISTORY
The sheepshead minnow is an omnivorous killifish (Family Cyprinodontidae) that
occurs in estuaries from Massachusetts to northern South America (Moore, 1968). It is
important in estuarine food chains as food for commercially valuable fishes (Darnell,
1958). The adults are sexually dichromatic after attaining 27mm in standard length
and, according to Hildebrand (1917), adult males average 48mm standard length and
females 45mm.
Hildebrand stated that the spawning period for this species was from April to Oc-
tober in the Beaufort, North Carolina area. Kilby (1955) collected young fish of 15mm
or less during all months except January, February and March at Cedar Key, Florida,
indicating a spawning period in the warmer months.
In our laboratory studies, no spawning occured below 26 C. Eggs are approximately
1 mm in diameter, demersal and adhesive by means of minute threads. Under
laboratory conditions, fry hatch in approximately five days at 30 C (Figure 1). Newly-
hatched fry are 4mm in length and are able to feed on brine shrimp (Anemia salina)
nauplii within 48 hours of hatching. Generation-time in the field has been estimated at
4 months by Holland and Coppage (1970); however, we have cultured fertile eggs to
mature adults within three months in the laboratory.
24
x20
o
•o
ui 16
O
z
12
8
• NO
HATCH
•
15 2O 25 3O 35
TEMPERATURE (°C)
4O
Figure 1. Relation between hatching time and temperature with Cyprinodon vari-
egatus embryos and fry. Data taken from Schimmel et al., 1974 (In
press).
393
98
-------
ENVIRONMENTAL FACTORS
Field observations indicate that the sheepshead minnow can tolerate a wide range of
environmental stress. Populations of the fish have been observed at water
temperatures from 1 1 to 35 C and salinities from 0 to 120 o/oo (Copeland, 1967).
Population densities of the adult fish in shallow marsh ditches have exceeded 20 in-
dividuals per m2 despite aggressive territorial activities of the males during spawning.
Our laboratory studies corroborate field observations on tolerances to temperature,
salinity and population density. Embryos and fry can be efficiently cultured in water
from 24 C to 35 Cand 15 to 30 o/oo (Figs. 2 and 3). Our studies also indicate that hold-
ing fish at high population densities is not difficult. For example, twenty-five adult fish
have been held in 50-liter aquaria under acceptable flow-through conditions
(A.P.H.A., Standard Methods, 1971) without appreciable loss due to aggressiveness of
the fish or other factors.
TOO
~ 80
v^
< 60
>40
D
2O
18 26 34
TEMPERATURE (°C)
42
Figure 2. Relationship between temperature and survival with Cyprinodon vari-
egatus embryos and fry. Data taken from Schimmel el al., 1974 (In
press).
FERTILITY, GROWTH AND SURVIVAL
To determine the quantity of fertile eggs that would be produced by a pair of sheeps-
head minnows during a 28-day period (at 30 C), two fish were placed in a small spawn-
ing chamber (10 X 12 X 18cm) constructed of acrylic plastic. The spawning chamber
was small enough to be placed in an aquarium, but large enough to permit the female to
avoid the aggressiveness of the male. Two sides of the spawning chamber were made of
2mm-square mesh nylon screen that allowed water to exchange between the chamber
and the larger bioassay aquarium. The bottom of the chamber was made of 4mm-
square mesh nylon screen through which eggs could pass, thus reducing chance for
predation of eggs by adult fish. Eggs falling below the adult chamber landed on a
0.25mm-square mesh screen drawer which was removed daily for counting of eggs and
394
99
-------
1OO
N 80
v^
< 6O
4O
20
U)
1O
2O
SALINITY
3O
4O
Figure 3.
Relationships between salinity and survival with Cyrpinodon variegatus
embryos and fry. Data taken from Schimmel et ai, 1974 (In press).
confirmation of fertility. The reproduction of 34 pairs of fish was monitored for 28
days. Pairs of fish were selected so that males and females ranged in size from 23 to
52mm standard length.
The pairs survived well and spawned readily, most producing enough fertile eggs for
a chronic bioassay. All of 34 females produced eggs and 66% of the females survived
the full 28 day spawning period in the chamber. The number of eggs produced per
female ranged from 2 to 1,028 and averaged 186. Eighty-eight percent of the 6,339 eggs
produced were fertile. The number of eggs produced each day varied (Figure 4), and in-
creased with time. Fish produced an average of 22 eggs during the first week, 57 eggs
and 83 eggs in the next two weeks. These data indicate that there is a 1-2 week ac-
climation period prior to optimum production. Once the fish began spawning,
however, most spawned daily.
Total egg production was not related to the size of the fish but frequency of spawning
and egg fertility appeared size-dependent. Females began producing eggs at about
27mm standard length. Nineteen fish less than 35mm long produced an average of 8.2
eggs per day and 15 fish 35mm and larger produced 7.8 eggs per day. The smaller fish
produced eggs more consistently (50% of the days versus 31%) with greater fertility
than the larger fish (94% fertile versus 79%).
Data analyses of control and no-effect experimental groups in C. variegatus bioas-
says reveal that most deaths occur among embryos and newly-hatched fry, with
negligible deaths among juveniles and adults. Survival of embryos and fry averaged
75% over the first four weeks (Fig. 5). Most of this mortality occurred during em-
bryonic development. Survival of 250 juvenile fish held under laboratory conditions
for four weeks was 97%. Ninety-three percent of 250 adult fish, including territorial
males, survived the four-week bioassays.
A 5-month chronic bioassay using C. variegatus exposed to endrin has recently been
completed and the data are now being analyzed.
395
100
-------
60
50
4O
u
30
2O
1O
t I 1 3 I .J tssassa i .J
0 1-10 11-20 21-30 31-40 41-5O ^51
NUMBER OF EGGS PER DAY
Figure 4. Daily egg production by 34 female Cyprinodon variegatus during a 28-
day period.
100.
-SO* HATCH
EMBRYOS
DAYS
Figure 5. Survival of embryos and fry of Cyprinodon variegatus at 30 C (range 28
to 32 C).
396
101
-------
A fish for chronic bioassay should be at least as susceptible to toxicant poisoning as
other fishes in a similar ecosystem. Studies performed at the Gulf Breeze Laboratory
(Lowe, unpublished data2) showed that susceptibility of C. variegatus to endrin,
dieldrin and DDT was comparable to that of three other estuarine fishes - Fundulus
similis, Leiostomus xanthurus and Mugil cephalus (Table 1). Schimmel et al. (1974)
have found that embryos and fry'of the sheepshead minnow are more susceptible to
the polychlorinated biphenyl, Aroclor®1254 (effect at 0.32 ug/1, P 0.01) than other
juvenile or adult species in any taxonomic group studied at this laboratory (Nimmo et
al. 1974). Since the above-mentioned chemicals, being organochlorines and probably
similar in their activity, do not aptly reflect how all toxicant types will affect the fish,
our data do indicate that the animal is not prohibitively resistant to some major insec-
ticides and related compounds.
Table 1. Comparative 48-hour EC-50's (in ug/1) of four organochlorines for Cy-
prinodon variegatus and three other estuarine fishes. Jack I. Lowe, un-
published data, Gulf Breeze Environmental Research Laboratory, Gulf
Breeze, Florida.
Cyprinodon Mugil Leiostomus Fundulus
Chemical variegatus cephalus xanthurus similis
Endrin 0.32 2.00 0.32 0.23
Dieldrin 24. 0.66 - 5.5
DDT 3.2 0.50 1.8 5.5
CONCLUSION
In our view, Cyprinodon variegatus would fill the need for an estuarine fish suitable
for chronic bioassays, producing significant information on the effects of toxic com-
pounds on estuarine fishes. This small fish is ubiquitous on the Atlantic and Gulf
Coasts of the United States, survives well in the laboratory, is fecund and has a short
generation time. Susceptibility of this fish to some commonly used pesticides is com-
parable to that of some other fish species in the estuarine system.
ACKNOWLEDGEMENTS
We wish to thank Stephen Foss for preparing the Figures (1-5).
LITERATURE CITED
American Public Health Association. American Water Works Association, and Wat-
er Pollution Control Federation. 1971. Standard Methods for the Examination
of Water and Wastewater. 13th ed. American Public Health Assoc Inc New
York, N.Y. 874p,
Brungs, William A. 1971. Chronic effects of low dissolved oxygen concentrations on
the fathead minnow (Pimephales promelas) J. Fish. Res. Board Canada. 28(8):
1119-1123.
Copeland, B. J. 1967. Environmental characteristics of hypersaline lagoons. Publ.
Inst. Mar. Sci. Univ. Texas. 12:207-218.
Darnell, R, M. 1958. Food habits of fishes and larger invertebrates of Lake Pont-
chartrain, Louisiana, an estuarine community. Publ. Inst. Mar. Sci. Univ. Texas
5:354-416.
!J. I. Lowe, Gulf Breeze Environmental Research Laboratory, Gulf Breeze, Florida 32561.
^Registered Trademark, Monsanto Co., St. Louis, Mo. Mention of commercial products does not constitute endorsement by
the Environmental Protection Agency.
397
102
-------
Eaton, John G. 1970. Chronic malathion toxicity to the bluegill (Lepomis mac-
rochirus Rafinesque). Water Res. 4:673-684.
Hildrebrand, S. F. 1917. Notes on the life history of the minnows Gambusia affinis
and Cyprinodon variegatits. Rep. U. S. Comm. Fish., 1917:1-14.
Holland, Hugh T. and David L. Coppage. 1970. Sensitivity to pesticides in three
generations of sheepshead minnows. Bull. Environ. Contam. Toxicol. 5(4):
362-367.
Kilby, John D. 1955. The fishes of two Gulf Coastal marsh areas of Florida. Tulane
Stud. Zool. 2(8): 173-247.
McKim, J. M. and D. A. Benoit. 1971. Effects of long-term exposures to copper on
survival, growth and reproduction of brook trout (Salvelinus fontinalis). J.
Fish. Res. Board Canada. 28(5):655-662.
Moore, G. A. 1968. Fishes In. W. F. Blair, ed., Vertebrates of the United States.
(New York: McGraw-Hill) 616 p.
Mount, Donald I. and Charles E. Stephans. 1967. A method for establishing accep-
table toxicant limits for fish. Trans. Amer. Fish. Soc. 96(2):185-193.
Nimmo, D. R., D. J. Hansen, J. A. Couch, N. R. Cooley, P. R. Parrish and J. I. Lowe.
1974. Toxicity of Aroclor® 1254 and its physiological activity in several estuarine
organisms. Archives Environ. Contam. Toxicol. (In Press).
Schimmel, Steven C., Hansen, David J. and Jerrold Forester. 1974. Effects of Aro-
clor®1254 on laboratory-reared embryos and fry of sheepshead minnows (Cy-
prinodon variegatus). Trans. Amer. Fish. Soc. (In Press).
398
103
-------
Reprinted from Transactions
of the American Fisheries
Society, Vol. 104(3): 584-
588, 1975, with permission
of the American Fisheries
Society
EFFECTS OF AROCLORR 1016 ON EMBRYOS, FRY, JUVENILES, AND
ADULTS OF SHEEPSHEAD MINNOWS (CYPRINODQN VARIEGATUS)
David J. Hansen, Steven C. Schimmel, and Jerrold Forester
Contribution No. 206
105
-------
Effects of Aroclor® 1016 on Embryos, Fry, Juveniles,
and Adults of Sheepshead Minnows (Cyprinodon variegatus)
DAVID J. HANSEN, STEVEN C. SCHIMMEL, AND JERROLD FORESTER
Made in United States of America
Reprinted from TRANSACTIONS OF THE AMERICAN FISHERIES SOCIETY
Vol. 104, No. 3, July 1975
pp. 584-588
© Copyright by the American Fisheries Society, 1975
107
-------
Effects of Aroclor® 1016 011 Embryos, Fry, Juveniles,
and Adults of Sheepshead Minnows (Cyprinodon variegatus)1
DAVID J. HANSEN, STEVEN C. SCHIMMEL, AND JERROLD FORESTER
United States Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Sabine Island, Gulf Breeze, Florida 32561
ABSTRACT
We investigated the toxicity of Aroclor 1016 to, and uptake by, fry and juvenile and adult
sheepshead minnows (Cyprinodon variegatus) in intermittent-flow bioassays lasting 28 days.
Survival of eggs, of fry hatched from them, and of juvenile and adult fish apparently was not
affected by 0.1, 0.32, 1.0, 3.2, or 10 /ug/liter of Aroclor 1016 added to aquaria, but 32 and 100
Ag/liter killed newly hatched fry and juvenile and adult fish. Sheepshead minnows accumulated
the chemical in proportion to its concentration in the test water. Fry contained 2,500 to 8,100
X the concentration of Aroclor 1016 added to the test water, adults 4,700 to 14,000 X, and
juveniles 10,000 to 34,000 X- As much as 77 /tg/g of Aroclor 1016 in eggs from exposed adults
apparently did not affect survival of embryos and fry.
Polychlorinated biphenyls, PCB's, occur in
estuarine environments in the United States
(Butler 1973; Nimmo and Banner 1974). One
PCB, Aroclor 1254, is acutely toxic (48 or
96 hours) to estuarine animals, such as the
eastern oyster, Crassostrea virginica, and pink
shrimp, Penaeus duorarum (Duke, Lowe, and
Wilson 1970). In exposures lasting 14 or
more days it is toxic to pink shrimp (Nimmo
et al. 1971) and grass shrimp, Palaemonetes
pugio (Nimmo et al. 1974), to oysters (Lowe
et al. 1972) and to fishes, such as the pinfish,
Lagodon rhomboides, spot, Leiostomus xan-
tliurus (Hansen et al. 1971), and the sheeps-
head minnow Cyprinodon variegatus (Schim-
mel, Hansen, and Forester 1974). It is over
30 times more toxic to fry of sheepshead min-
nows than to juveniles or adults (Schimmel,
Hansen, and Forester 1974). Some fry from
sheepshead minnow eggs thct contained 7 /xg/g
or more of Aroclor 1254 died within the first
week following hatching (Hansen, Schimmel,
and Forester 1974).
A new PCB, Aroclor 1016, is now being
manufactured to replace certain PCB's that
are no longer produced. Aroclor 1016 has
©Registered trademark, Monsanto Company, St.
Louis, Missouri. Mention of commercial products
or trade names does not constitute endorsement by
the Environmental Protection Agency.
1 Contribution No. 206 from the Gulf Breeze En-
vironmental Research Laboratory, U. S. Environ-
mental Protection Agency, Gulf Breeze, Florida 32561
(Associate Laboratory of the National Environ-
mental Research Center, Corvallis, Oregon).
not been found in oysters or fishes from estu-
aries in the United States sampled by the En-
vironmental Protection Agency's National Es-
tuarine Monitoring program (P. A. Butler,
United States Environmental Protection
Agency, personal communication). In the
laboratory, however, it is as acutely toxic to
oysters, brown shrimp (Penaeus aztecus),
and pinfish as Aroclor 1242 and as toxic to
oysters and pinfish as Aroclor 1254 (Hansen,
Parrish, and Forester 1974). Its delayed tox-
icity to pinfish in exposures lasting 14 or more
days is similar to that found with Aroclor 1254
(Hansen, Parrish, and Forester 1974).
This experiment was conducted to deter-
mine the toxicity of Aroclor 1016 in water to
embryos, fry, and juvenile, and adult sheeps-
head minnows, and to determine its endog-
enous toxicity to embryos and fry reared from
eggs from exposed adults. Also, chemical
analyses were made to determine if this PCB
accumulates in eggs, fry, and juvenile and
adult fish.
MATERIALS AND METHODS
Test Fish
Adult and juvenile fish were collected near
the Gulf Breeze Laboratory and acclimated in
30 C seawater for at least 14 days before ex-
posure. Fish were not used if mortality ex-
ceeded \% or if abnormal behavior was ob-
served in the 48 hours immediately preceding
a test. Adult fish averaged 41 mm standard
584
108
-------
HANSEN ET AL.—AROCLOR 1016 TOXICITY
585
length (range 31-51 mm) and juveniles aver-
aged 20 mm standard length (range 16-27
mm). Fish were fed PCB-free dry commer-
cial fish food daily and PCB-free frozen adult
brine shrimp twice a week during acclimation
and testing.
Eggs were obtained from acclimated adult
females whose egg production was enhanced
by hormonal injection using the techniques
of Schimmel, Hansen, and Forester (1974).
Eggs, stripped manually from six or more
females, were placed in. 20-40 ml of filtered
30 C seawater and fertilized with excised
macerated testes from six or more males. Suc-
cess of fertilization was determined by micro-
scopic examination for cleavage 1.5 hours
after mixing with macerated testes. Fry from
the eggs were fed daily on PCB-free brine
shrimp nauplii.
Aroclor 1016 Exposure
Adult and juvenile fish and sheepshead min-
now eggs were exposed simultaneously to Aro-
clor 1016 in two intermittent-flow bioassays,
each lasting four weeks. In the first experi-
ment, we exposed 25 adults (13 females and
12 males), 25 juveniles, and 100 eggs in
aquaria receiving 0, 0.1, 0.32, 1.0, 3.2, or 10
//,g/liter of Aroclor 1016. In the second ex-
periment 0, 1.0, 3.2, 10, 32, or 100 ^g/liter
of Aroclor 1016 was metered into an aquar-
ium. The dosing apparatus (Schimmel, Han-
sen, and Forester 1974) was a modification
of that of Brungs and Mount (1970). Each
of approximately 150 daily cycles siphoned
1.5 liters of filtered 30 C seawater, 11 /xg of
carrier (polyethylene glycol 200) and appro-
priate amounts of PCB to each 80-liter aquar-
ium. Water and carrier without PCB were
delivered to the control aquaria. Salinity of
the water averaged 21.1%° (12.0&> to 32.0&0-
Aquaria were checked daily to determine sur-
vival of embryos, fry, and juvenile and adult
fish.
Aroclor 1016 in Eggs
The effect of Aroclor 1016 accumulated in
eggs was determined during the second bio-
assay by enhancing egg production with hor-
monal injections of exposed adults, fertilizing
the eggs artificially, and then monitoring their
development in flowing PCB-free seawater.
Female sheepshead minnows were injected in-
traperitoneally with 50 IU human chorionic
gonadotropic hormone on exposure days 25
and 27. On day 28 or 29, eggs were stripped
manually from five females from each concen-
tration of Aroclor and eggs from each female
were placed in individual beakers containing
20^10 ml of filtered 30 C seawater. Testes'
were removed from each of five males exposed
to the same concentration and macerated in
separate containers. Fertilization was per-
formed by pairing eggs from one female and
testes from one male and no male fertilized
more than the eggs from one female. Twenty-
five eggs from each fish were transferred to
a 10-cm Petri dish to which a 9 cm-high collar
of 500/x nylon mesh was attached. Dishes
were submerged 7 cm in the 80-liter con-
trol aquarium; average salinity was 24.0%"
(19.5%° to 28.0&*). Success of fertilization
was determined by microscopic examination
for cleavage 1.5 hours after mixing with mac-
erated testes. Thereafter, dishes were exam-
ined daily for 21 days to determine survival
of embryos and fry. Fry were fed PCB-free
live brine shrimp nauplii daily.
Chemical Analyses
Concentrations of Aroclor 1016 in water
and fish were determined by electron-capture
gas chromatography, as described by Han-
sen, Parrish, and Forester (1974). Unfiltered
water samples from each concentration were
obtained at least three times during each four-
week exposure and analyzed. Concentrations
of Aroclor were determined in fry and ju-
venile and adult fish surviving both bioassays
and in eggs from adult fish from the second
bioassay. All fish samples consisted of 14 or
more individuals. Aroclor 1016 was quanti-
tated by comparing the total height of all
peaks in a sample with the total height of all
peaks in a standard of known concentration.
Recoveries from spiked solutions were greater
than 80%; data were not adjusted for re-
covery. All tissue residues were determined
on a wet weight basis.
109
-------
586
TRANS. AM. FISH. SOC., 1975, NO. 3
TABLE 1.—Survival of sheepshead minnows (Cy-
prinodon variegatus) exposed to Aroclor 1016 for
28 days.
Concentration
28-day survival,
Desired
Control
0.10
032
1.0
3 2
10.
Control
1.0
3.2
10.
32.
100.
Measured
NDb
ND
ND
0.46
1.4
5.5
ND
0.29
1.0
3.4
15.
42.
Eggs"
Experiment 1
76
76
73
90
75
61
Experiment 2
77
78
67
71
14
0
Juveniles
100
100
100
100
100
100
100
96
100
96
12
0
Adults
96
96
100
100
96
84
92
88
88
88
8
0
* Eggs hatched from day 4 to 9 and their fry were ex-
posed until termination on day 28.
b ND = not detectable; < 0.2 /ig/liter.
Statistical Analyses
Concentrations of Aroclor 1016 lethal to
50% of the fish (95% confidence limits) were
estimated using the method of Litchfield and
Wilcoxon (1949). The ^2 test for independent
samples was used to compare the survival of
control and PCB-exposed fish. Differences
were considered significant at a = 0.05.
RESULTS AND DISCUSSION
Survivals of sheepshead minnow adults, ju-
veniles,'embryos and their fry exposed to Aro-
clor 1016 for 28 days apparently were not
affected by addition of 0.1, 0.32, 1.0, 3.2, or
10 jug/liter of the PCB to test aquaria; but,
when Aroclor was added at 32 or 100 /xg/
liter the fish died (Table 1). Signs of poison-
ing of juveniles and adults included darkened
body coloration, uncoordinated swimming,
cessation of feeding, and lesions on the body.
Juvenile pinfish, another estuarine fish, also
died in 32 /x,g/liter of Aroclor 1016, but sur-
vived lower concentrations (Hansen, Parrish,
and Forester 1974) ; signs of poisoning of
pinfish were similar to those listed for sheeps-
head minnows. Although fin rot was not ob-
served in Aroclor 1016-exposed fish (present
study; Hansen, Parrish, and Forester 1974),
it was common in Aroclor 1254-exposed
sheepshead minnows (Schimmel, Hansen, and
Forester 1974), spot, and pinfish (Hansen
et al. 1971).
Survivals of embryos and their fry, and of
juvenile and adult sheepshead minnows that
were exposed to Aroclor 1016 were similar
(Tables 1 and 2). Aroclor 1016 did not alter
the survival of embryos to hatching or the
hatching time. Eggs began to hatch on the
fourth day of exposure; one-half hatched by
day five and all hatched by day nine. Fry,
juveniles, and adults treated with Aroclor at
32 and 100 /xg/liter began to die during the
second week of exposure; over half died be-
fore the end of the second week. All fish
treated with 100 /xg/liter were dead by 3
weeks, but some fish of each life stage sur-
vived the four-week exposure to 32 /xg/liter.
The data show that the toxicity of Aroclor
1016 was similar to eggs, juveniles, and adults
(Table 2). In contrast, Schimmel, Hansen,
and Forester (1974) reported that Aroclor
1254 affected each of these life stages differ-
ently. Sheepshead minnow fry were much
more sensitive to Aroclor 1254 than were
embryos, juveniles, or adults. During three
weeks exposure, 0.32 /xg/liter of Aroctor 1254
was lethal to fry, 10 /j.g/liter was lethal to
juveniles, and adults survived 10 /xg/liter.
Sheepshead minnows exposed for 28 days
to concentrations of Aroclor 1016 ranging
from 0.1 and 100 /Ag/liter accumulated the
chemical in proportion to the concentration
TABLE 2.--Concentrations of Aroclor 1016 (/j.g/liter added) lethal to 50 percent of the eggs, juvenile, and
adu.lt sheepshead minnows (LC,a) treated continuously during a four-week bioassay. Ninety-five percent
confidence limits are in parentheses.
Week
Stage
Eggs'
Juveniles
Adults
1
>100
>100
> 100
2
31 (25-38)
47 (35-62)
42 (29-61)
3
24 (21-28)
22 (17-28)
28 (22-36)
4
21 (18-25)
20 (15-26)
19 (14-25)
1 Bioassay began with eggs, which hatched from the 4th to 9th days.
110
-------
HANSEN ET AL.—AROCLOR 1016 TOXICITY
587
TABLE 3.—Concentrations of Aroclor 1016 in fry, juvenile, and adult sheepshead minnows exposed to the PCB
in water. Except when indicated by footnote, each sample consisted of a minimum of 14 fish. Concentra-
tion factors in parentheses.
Concentration in water
(/ig/liter)
Concentration in whole fish (/ig/g wet weight)
Added
Control
0.10
0.32
1.0
3.2
10.
Control
1.0
3.2
10.
32.
100.
Measured
ND»
ND
ND
0.46
1.4
5.5
ND
0.29
1.0
3.4
15.
42.
0,
Fry"
Experiment 1
ND"
ND
,81 (2,500)
4.9 (4,900)
22.
38
5.
26,
57,
200,
(6,900)
(3,800)
Experiment 2
ND
,9 (5,900),
(8,100)
(5,700)
(6,200)
Juveniles"
ND
2.3
8.9
11.
79.
230.
(23,000)
(28,000)
(11,000)
(25,000)
(23,000)
ND
10.
54.
220.
1100.
(10,000)
(17,000)
(22,000)
(34,000) =
Adults"
ND
0.84
1.5
12.
46.
100.
( 8,400)
( 4,700)
(12,000)
(14,000)
(10,000)
ND
5.4
22.
110.
( 5,400)
( 6,900)
(11,000)
a Eggs and hatched fry were exposed for 33 days; juveniles and adults were exposed for 28 days.
b ND = not detectable; < 0.2 /tg/liter in water, < 0.2 /ig/g in tissue.
c Sample consisted of three juvenile fish.
in the test water (Table 3). In general, ju-
venile fish contained the greatest concentra-
tions of the PCB, followed by adults and fry
(fry were exposed 5 days as embryos followed
by 28 days as fry). Concentration factors
(concentration in whole fish divided by nomi-
nal concentration in the test water) ranged
from 2,500 to 8,100 for fry, 4,700 to 14,000
for adults, and 10,000 to 34,000 for juveniles.
Concentration factors in adult sheepshead min-
nows exposed identically to Aroclor 1254
were greater (15,000 to 30,000) (Hansen,
Schimmel, and Forester 1974). By compari-
son, concentration factors in fry and adult
sheepshead minnows exposed for 3 weeks to
0.1, 0.32, 1.0, 3.2, or 10 /^g/liter of Aroclor
1254 were similar, ranging from 11,000 to
32,000 (Schimmel, Hansen, and Forester
1974).
Eggs from sheepshead minnows exposed to
1.0, 3.2, or 10 /xg/liter of Aroclor 1016 for
4 weeks contained up to 77 ^g/g of the PCB
(Table 4). Fertilization success, survival of
embryos to hatching, and survival of fry for
two weeks after hatching did not appear to be
altered by this or lesser concentrations.
Sheepshead minnow eggs that contained Aro-
clor 1254 in concentrations of 7 /ig/g or more,
however, exhibited decreased fry survival in
the first week following hatching (Hansen,
Schimmel, and Forester 1974).
This study indicates that although Aroclor
1016 is readily accumulated by and is toxic
to fry and to juvenile and adult sheepshead
minnows, it is about 0.01 times as toxic as
Aroclor 1254 to sensitive fry. More impor-
tant, concentrations of Aroclor 1016 in eggs
TABLE 4.—Fertility of eggs from adult sheepshead minnows exposed to Aroclor 1016 for 28 to 29 days, sur-
vival of embryos from fertile eggs until hatching, and survival of hatched fry. Eggs are from five fish, 25
eggs each, per concentration.
Concentration
added to water
(/tg/liter)
Adult exposure
Control
1.0
3.2
10.
Concentration
in eggs
(/ig/g)
Average (range)
ND«
4.2 (3.1-5.1)
17. (12-19)
66. (51-77)
Tested
125
124
125
124
Eggs
Fertile
119 (95)"
120 (97)
122 (98)
117 (94)
Fry
Hatched
82 (69)=
76 (63)
102 ( 84 )
88 (75)
Week 2
81 (99)1
71 (93)
99 (97)
84 (96)
Week 3
60 (73)*
64 (84)
91 (89)
72 (82)
" ND = not detectable; < 0.2 /ig/g.
b Values in parentheses indicate percentage of eggs tested that were fertili
c Values in parentheses indicate percentage of fertile eggs that hatched.
d Values in parentheses indicate percentage of hatched fry that survived.
Ill
-------
588
TRANS. AM. FISH. SOC., 1975, NO. 3
as great as 77 jtg/g apparently did not affect
survival of embryos and fry for three weeks.
LITERATURE CITED
BRUNGS, W. A., AND D. I. MOUNT. 1970. A water
delivery system for small fish-holding tanks.
Trans. Am. Fish. Soc. 99(4): 799-802.
BUTLER, P. A. 1973. Organochlorine residues in
estuarine mollusks. 1965-1972. National Pesti-
cide Monitoring Program. Pestic Monit. J. 6(4) :
238-362.
DUKE, T. W., J. I. LOWE, AND A. J. WILSON, JR.
1970. A polychlorinated biphenyl (Aroclor
1254®) in the water, sediment and biota of
Escambia Bay, Florida. Bull. Environ. Contam.
Toxicol. 5(2): 171-180.
HANSEN, D. J., P. R. PARRISH, AND J. FORESTER.
1974. Aroclor® 1016: Toxicity to and uptake
by estuarine animals. Environ. Res. 7(3) : 363-
373.
, , J. I. LOWE, A. J. WILSON, JR., AND P.
D. WILSON. 1971. Chronic toxicity. uptake and
retention of Aroclor® 1254 in two estuarine
fishes. Bull. Environ. Contam. Toxicol. 6(2) :
113-119.
HANSEN, DAVID J., STEVEN C. SCHIMMEL, AND JER-
ROLD FORESTER. 1974. Aroclor® 1254 in eggs
of sheepshead minnows: Effect on fertilization
success and survival of embryos and fry. Proc.
Southeastern Assoc. Game Fish. Comm. pp. 805—
812.
LlTCHFIELD, J. T., JR., AND F. WlLCOXON. 1949. A
simplified method of evaluation dose-effect ex-
periments. J. Pharmacol. Exp. Ther. 96(2):
99-113.
LOWE, J. I., P. R. PARRISH, J. M. PATRICK, JR., AND
J. FORESTER. 1972. Effects of the polychlori-
nated biphenyl Aroclor® 1254 on the American
oyster Crassostrea virginica. Mar. Biol. (Berl.)
17(3): 209-214.
NIMMO, D. R., AND L. H. BAHNER. 1974. Physio-
logical consequences of polychlorinated biphenyl-
and salinity-stress in penaeid shrimp. Sympo-
sium "Pollution and the physiological ecology
of estuarine and coastal water organisms. In
press.
, R. R. BLACKMAN, A. J. WILSON, JR., AND J.
FORESTER. 1971. Toxicity and distribution of
Aroclor® 1254 in the pink shrimp, Penaeus
duoramm. Mar. Biol. (Berl.) 11(3): 191-197.
-, J. FORESTER, P. T. HEITMULLER, AND G. H.
COOK. 1974. Accumulation of Aroclor® 1254
in grass shrimp (Palaemonetes pugio) in lab-
oratory and field exposures. Bull. Environ.
Contam. Toxicol. 11 (4) : 303-308.
SCHIMMEL, STEVEN C., DAVID J. HANSEN, AND JER-
ROLD FORESTER. 1974. Effects of Aroclor® 1254
on laboratory-reared embryos and fry of sheeps-
head minnows (Cyprinodon variegatus). Trans
Am. Fish. Soc. 13 (3) : 182-186.
112
-------
Reprinted from Water Research,
Vol. 9: 937-944, 1975, with
permission of Pergamon Press,
Elmsford, New York
PROBLEMS ASSOCIATED WITH LOW-SOLUBILITY COMPOUNDS IN AQUATIC TOXICITY TESTS;
THEORETICAL MODEL AND SOLUBILITY CHARACTERISTICS OF AROCLORR 1254 IN WATER
W.P. Schoor
Contribution No. 208a
113
-------
Water Research Vol. 9. pp. 937 to 944. Pergamon Press 1975. Printed in Great BnUiin.
PROBLEMS ASSOCIATED WITH LOW-SOLUBILITY
COMPOUNDS IN AQUATIC TOXICITY TESTS:
THEORETICAL MODEL AND SOLUBILITY
CHARACTERISTICS OF AROCLOR® 1254 IN WATER*
W. P. SCHOOR
U.S. Environmental Protection Agency, Gulf Breeze Environmental Research Laboratory,
Sabine Island, Gulf Breeze, Florida 32561, U.S.A.t
(Received 9 January 1975)
Abstract—A theoretical model of the behavior of substances having low water-solubility is presented
and discussed with respect to aqueous bioassay. Ultracentrifugal techniques were used in an attempt
to study size distributions of Aroclor 1254 aggregates in aqueous emulsions. Results indicate strong
adsorption from emulsion by surfaces and a water-solubility at 20°C of less than 0.1 /ig T ' in distilled
water and approximately 40% of that value in water containing 30gl~' Nad. Implications with
regard to aqueous bioassay are discussed.
INTRODUCTION
Laboratory experiments designed to determine the
effects of chemicals on aquatic organisms require that
the tests be conducted under conditions which repro-
duce those present in nature as closely as possible.
In order to accomplish this in a precise and scientific
fashion, the physical state of a compound in an
aqueous dispersion must be known. Convenience,
time and other factors have in the past often led to
the use of techniques in the laboratory which do not
take into consideration that the solubility characteris-
tics of a compound may possibly affect the toxicity
necessitating extrapolation from an apparent toxicity
established in the laboratory to an expected toxicity
under field conditions. In many instances, the practice
of using extrapolation in scientific investigations is
necessary and has proven to be a valuable tool when
certain conditions cannot be met. However, the range
through which the extrapolation is carried out must
be chosen with great care, because without sufficient
experimental and theoretical justification, a resulting
extrapolation in this light may well prove to be unrea-
listic. Since natural water conditions represent a
multi-component system, any attempt to understand
it quantitatively must be preceded by a study of the
system under ideal conditions. While the knowledge
thus gained may or may not be of consequence in
direct application, it, nevertheless, provides a more
precise scientific basis for choosing valid limits for
extrapolation.
The physical state of a compound in water is not
a simple and straightforward phenomenon, even given
* Contribution No. 208-^Gulf Breeze Environmental
Research Laboratory.
t Associate Laboratory of the National Environmental
Research Center, Corvallis, Oregon.
® Registered trademark, Monsanto Company, St. Louis,
MO. Mention of trade names does not constitute endorse-
ment by the Environmental Protection Agency.
the idealized conditions of a two-component sys-
tem—a single solute and a single solvent. A definable
system should, however, be the starting point of any
scientific investigation aimed to arrive at data which
lead to a quantitative understanding of the behavior
of a compound in water. With these data a more
precise attempt can be made to extrapolate from a
system employed in the laboratory to the obviously
much more complex system present in natural waters.
The purpose of this work is to provide a working
theory on the behavior of substances of low water-
solubility and to test this theory by investigating the
solubility characteristics of Aroclor 1254.
THEORY
To explain and predict the characteristics of water-
insoluble substances at low concentrations, an
attempt is made here to redefine the basic principles
underlying a disperse system. No attempts have been
made to include in the definition the somewhat obso-
lete and often vague definitions of emulsions, suspen-
sion, colloids, etc. The characteristics ascribed to each
becoming readily apparent as the theoretical treat-
ment of the proposed model continues.
In this paper, an ideal or true solution is defined
as a solute dispersed in a solvent so that any single
molecule of solute is surrounded by enough solvent
molecules to ensure that at any instant all solute
molecules are distributed statistically equidistant,
assuming a dilution at which interactions between
solute molecules become negligible.
The ideal solution, under the conditions described,
is represented by the presence of single solute mole-
cules. Solute aggregates consisting of two or more
molecules may represent a deviation from the ideal
solution because, at least theoretically, these aggre-
gates could consist of any number of molecules whose
behavior would not necessarily coincide with that of
937
115
-------
938
W. P. SCHOOR
a single molecule. For each solute and a single sol-
vent there is assumed to exist amongst all aggregates
a maximally stable aggregate which, due to its nature,
remains statistically equidistant from all other aggre-
gates for at least a certain period of time. The stability
of this aggregate depends solely on the molecularly
characterized interactions at the solute-solvent inter-
phase and on temperature.
By definition, a single solute molecule in a disperse
system possesses a certain sphere of influence, the
nature of which governs the fate of the solvent mole-
cules that surround it, which in turn affects the behav-
ior of the solute molecule, and thus determines the
characteristics of the solute molecule in the system.
While precise information is lacking, it is known,
nevertheless, that the range of effect of a solute mole-
cule may extend through several layers of surround-
ing solvent molecules. This means, of course, an
orderly alignment involving either oppositely charged
polar regions or non-polar regions on the solute and
the solvent molecules. If this interaction between
solute and solvent molecules is of significance, the
above defined ideal solution can be visualized, pro-
vided also that there is no competition among the
solvent molecules belonging to respective spheres of
influence of two separate solute molecules.
The complexity of the situation is increased in cases
where the interactions between solute and solvent
molecules (solute-solvent interactions) become less
pronounced, and. as a result, the interactions between
solute and solute molecules (solute-solute interac-
tions) become more pronounced. This implies that the
sphere of influence around the solute molecule is
diminished with respect to the solvent molecules
which are now no longer attracted to the same degree.
As two or more solute molecules start to form aggre-
gates, the factor of size of aggregates versus their sta-
bility in a solvent becomes of utmost importance.
A generalized illustration of the size distribution
of aggregates that one might expect to find in a sus-
pension is shown in Fig. 1. Region "A" describes an
area in which the aggregates are too small to exist
independently because interactions in the sphere of
influence at that point are such that solute-solute in-
teractions, which have now become aggregate-aggre-
gate interactions, are more pronounced than the
aggregate-solvent interactions. Therefore, these aggre-
gates are expected to coalesce, moving them into
region "B". which describes a range of aggregate sizes
of maximum stability. The aggregate-aggregate inter-
actions in this range are weaker than in region "A"
for that size of aggregate. Region "C" described aggre-
gates which are too heavy to remain in suspension
for a given period of time and will settle out or break
into smaller, more stable aggregates. The exact shape
of this curve and especially that of region "B",
depends on how tightly the solvent is held within the
* The equations used are normally found in any text-
book on phxsicn! chemistry, and their reproduction here
is intended mereh for the convenience of the reader.
(a)
Region where small
aggregates coalesce
(c)
Region where large
aggregates precitote
Increase in aggregate diameter
Fig. 1. Theoretical relative stability of different sizes of
aggregates in an emulsion during a given time interval.
sphere of influence of the solute aggregate, which is
a function of the molecular interactions between
solute and solvent.
The distribution of different aggregate sizes in
terms of molecularly characterized interactions is
shown in Fig. 2. The actual equilibrium reaction tak-
ing place is described in a simplified manner at the
top of the figure. The two curves relate the hypotheti-
cal strength of interactions of solute-solvent (aggre-
gate-solvent) type and solute-solute (aggregate-
aggregate) type to aggregate size. The region where
the curves cross corresponds to a distribution of
aggregate sizes of maximum stability.
MODEL
Aroclor 1254 was chosen as a model compound because
it has been extensively used in bioassay at this laboratory
(Duke, Lowe and Wilson, 1970; Nimmo et al., 1971a;
Nimmo et al., 1971b; Hansen et al., 1971; Lowe et at.,
1972; Walsh, 1972; Cooley, Keltner and Forester, 1972).
One approach to estimate quantitatively the solubility
of Aroclor 1254 in water and the behavior of its aggregates
is to use ultracentrifugal analysis. This technique permits
the selective removal of particles of a certain size. For a
spherical particle having a density of (pj and a radius of
(/•) the molecular weight (mol. wt) is represented by:
mol. wt = 4/3nr3pN0
(1)
where N0 is Avogadfo's Number.*
Two opposing forces (/) which determine the fate of a
Equilibrium between
single molecule to)
and aggregates
(b)and(c)
(b)
•
Solute-solute
'(Aggregate-aggregate)
iter actions
Solute-sol vent
(aggregate-solvent}
interactions
Aggregate size
Fig. 2. Theoretical strength of interaction between solute
and solvent.
116
-------
Theoretical model and solubility characteristics of Aroclor® 1254 in water
939
particle in solution:
sedimentation f = 4/3m-3(p — p0)g
bouyancy / = 671/7;,
and
(3)
where (p0) is the density of the solvent, (g) is gravity, and
(r\) is the viscosity of the solvent.
To remove a small particle from an emulsion at a
reasonable rate, a force larger than gravity must be
applied. Using the ultracentrifuge. (g) in equation (2) is
replaced with (or.x), the angular velocity of the centrifuge
rotor (at) times the distance of travel (.\-j of the emulsified
particle.
The rate of sedimentation during centrifugation is
described by:
dx
df
2r2(p
(4)
where (t) is time in seconds to reach equilibrium. Integration
yields:
2r2(p - p0)orr
In x-, — In \, =
9fj
The radius of a spherical particle is then given by:
-------
940
W. P. SCHOOR
Aroctor 1254
Fig. 3. Typical gas chromatograms (see text for detailed
information).
mean that solubility is approached at that point, only
that perhaps a stable emulsion is reached at that
point.
The hexane extract of type II emulsion (chromato-
gram B) indicates a relative reduction in peak height
for the early eluting peaks. This phenomenon is better
described by the results shown in Table 2. For com-
parison peak 7 was arbitrarily assigned a relative
value of 100%. The results indicate that on standing
a type II emulsion shows a reduction of the individual
peaks, with the early eluting components, or less
chlorinated biphenyls (Zitko, 1970), being reduced
much more than the late eluting ones. The degree
of reduction depends somewhat on the preparation
and initial concentration of individual type II emul-
sions (Table 2). Type HI emulsions of comparable "to-
tal" concentration show a relative distribution of the
isomers identical to that of the standard.
The distribution of isomers in a hexane extract of
the gill tissue of a pink shrimp (Penaeus duorarum)
exposed to 2.5 jig I"1 Aroclor 1254 for 20 days is
shown in parentheses at the bottom of Table 2.
Because peaks 2, 4 and 7 showed obvious contamina-
tion, peak 6 was assigned the arbitrary, relative 100%
value. The "total" concentration of 3.4mgkg~' was
based on the total height of peaks 1, 3, 5 and 6, and
on the wet weight of gill tissue (blotted to remove
adhering water).
Filtration of type I emulsion through 450 nm
(0.45 fi) Millipore® filters revealed obstructed passage
of Aroclor 1254 aggregates smaller than 450 nm. Start-
ing with a Imgl"1 emulsion and changing filters
after each filtration, less than 0.01 p.g \~' of the mater-
ial remained in the water after 15 passages. Since
aggregates in the starting emulsion were most likely
smaller than 450 nm (calculations using equation (1)
Table 1. Effect of storage time on amount of Aroclor 1254 remaining
in the water phase
lig/J Aroclor 1254
Time (days)
0
2
5
6
8
9
13
15
19
20
21
23
26
28
33
34
41
43
Type
2300
502
483
428
355
350
280
I
301
115
113
112
97
87
78
Type
286
123
98.5
54.7
48.1
44.5
15.5
II
50.2
23.6
11.3
6.7
7.1
6.5
7.7
7.4
6.8
Table 2. Isomer distribution of Aroclor 1254 type II emulsion after stand-
ing for various periods of time in 3 1. glass bottle
Time
(days)
2
9
13
19
20
21
41
21
33
38
Total
cone.
(UK/D1
286
123
98.5
54.7
58.1
44.5
15.5
13.4
3.6
1.6
(3.4 ppm)
X Peak Height1
Peak Numbers2
1
76
79
79
72
64
61
37
16
12
9
(41)
2
93
78
79
75
70
65
41
27
21
10
3
95
89
93
85
80
82
56
44
39
(80)
4
95
94
98
93
89
90
70
55
45
31
5
98
98
99
91
92
99
77
76
64
46
(87)
6
104
99
96
95
94
93
88
87
82
63
(100)
7
100
100
100
100
100
100
100
100
100
100
'Calculations are based on the relative height of peak 7 (see below).
2Peak numbers are shovn on the chromatogram in Fig. 1.
118
-------
Theoretical model and solubility characteristics of Aroclor® 1254 in water
941
lead to roughly 10'° times the average molecular
weight of Aroclor 1254), the Aroclor 1254 must have
been adsorbed on the filter. This was also evidenced
by the fact that the filter paper turned slightly trans-
parent after the first passage during which about 95%
of the material was removed from the emulsion.
The first centrifugation experiments were carried
out by centrifuging 180ml of 42/jgT1 Aroclor 1254
type II emulsion in 60 ml polyacetate centrifuge tubes
for 60min at 107,000 x g (maximum).
At an 85% total recovery the following distribution
was found:
acetone extract of tubes 66%
hexane rinses of tubes 18%
top 50 ml water phase 5%
bottom 10ml water phase 11%
The low recovery (85%) was probably due to incom-
plete extraction of the tubes in spite of refluxing with
acetone.
Polyallomer® centrifuge tubes were tried next.
When 180ml of 286/igT' type II emulsion were cen-
trifuged in 60ml Polyallomer tubes for 60min at
107,000 x g (maximum) the following distribution
was found:
acetone extract of tubes —
hexane rinses of tubes 22%
top 25 ml water phase 0.5%
bottom 35 ml water phase 0.6%.
These percentages were based on the total amount
of starting material, i.e. assuming 100% recovery in-
stead of the 85% in the case of the polyacetate tubes.
Extraction of the Polyallomer tubes by refluxing with
acetone produced too many interfering peaks on the
chromatogram, making complete recovery calcula-
tions impossible. Direct adsorption on Polyallomer
Table 3. Adsorption of Aroclor 1254 type II emul-
sion on Polyallomer centrifuge tubes on standing
Time (hrs)
0
3
72
0
1
3
Aroclor 1254 (yg/i)
in water phase
125
86
3.
45
35
27
3
tubes was achieved by permitting type II emulsions
to sit undisturbed in the tubes. Table 3 shows the
outcome for two different concentrations.
To permit recovery and study of the material
adsorbed on surfaces, 34ml stainless steel centrifuge
tubes were used for static tests, as well as for ultracen-
trifugal analysis. Table 4 shows the amounts of Aroc-
lor 1254 adsorbed on the wall of a stainless steel cen-
trifuge tube in relation to starting concentration and
time. The amounts adsorbed from the 14 and 2 ^g 1~ '
emulsions were greater than that adsorbed from the
113 ^g I"1 emulsion during the same time period. It
should be pointed out that 0.100/ig of Aroclor 1254
adsorbed as a monomolecular layer per tube repre-
sents about 2% of the minimum area available. The
calculated inside area of a stainless steel centrifuge
tube was 60.8cm2 This area must be considered
minimum because the surface was assumed to be
ideally smooth, which certainly is not the case. How-
ever, for the approximations involved, this figure was
used.
A simple calculation using equation (1) yields
0.613 nm2 for the cross-sectional surface area of an
average Aroclor 1254 molecule using the average
molecular weight of 327 (Hutzinger et a/., 1972), and
p = 1.505 g cm"3 (W. B. Papageorge, Monsanto
Company, St. Louis, Missouri, personal communica-
tion). Utilizing a molecular model with the phenyl
groups at right-angles to each other and bond length
(Pauling, 1940) as the basis for calculations, a cross-
sectional area of 0.643 nm2 for the fully chlorinated
and 0.356 nm2 for the unchlorinated or biphenyl
molecule was obtained. Values falling between are not
linearly related to amount of chlorination. Using
0.613 nm2 as an approximate, average cross-sectional
area, 0.100 /ig of Aroclor 1254 occupies 1.13cm2 in
the form of a monomolecular layer. This corresponds
to approximately 3/jgT1 in a 34ml stainless steel
centrifuge tube. It can be seen that even at 50%
adsorption from a S^gl"1 emulsion only about 1%
(maximum) of the available surface area is occupied,
and surface saturation was not a factor.
The amounts of Aroclor 1254 in the form of emul-
sions of type II and type III adsorbed on the walls
of the stainless steel centrifuge tubes are shown in
Table 5. There is a difference in adsorption of the
two different types of emulsion in the absence of
NaCl. At least for type III emulsions, the introduction
Table 4. Adsorption of Aroclor 1254 on stainless steel centrifuge tubes as
a function of time and concentration
Aroclor 1254 type II
Time
(hrs)
0.5
1
2
16
1
2
Total
3.83
3.83
3.83
3.83
0.48
0.06
113
113
113
113
14
2
Water
3.63
3.31
3.20
3.14
0.35
0.03
107
97
94
92
10
1
emulsion
S.S. tube!
-------
942
W. P. SCHOOR
Table 5. Adsorption of Aroclor 1254 on stainless steel centrifuge tubes
Time
(hrs)
0 5
1 0
2.0
4.0
22
y/g Aroc
Type II Emulsion
0 K/i NaCl
0.19
0.30
0.33
0.42
lor 1254 I adsorbed
Type III Emulsion
30 K/i NaCl 0 K/i. NaCl
0.09
0.10
0.14
0.19
0.39
0.10
0.14
0.45
'Data adjusted to 4.00 pg total starting amount.
Table 6. Centrifugation of Aroclor 1254 in water of varying salinities at
69,000 x g (maximum)
pg/1 Aroclor 1254
remaining in water phast
Time (hrs)
0.5
1.0
2.0
0
13.9
12.5
7.2
g/t NaCl
15
7.1
6.6
4.6
:'
30
6.0
4.9
2.9
'Started with 50 ug/fc Type III emulsion.
of 30gl~' NaCl appears to have no effect on the
amount of Aroclor 1254 adsorbed. However, centrifu-
gation reveals a difference in the size of the aggregates
formed in the presence of NaCl, as shown in Table
6.
In comparison with an Aroclor 1254 standard, the
relative distribution of the isorners in emulsions of
type II and III is quite different, as shown in Tables
7 and 8. However, in all cases the adsorbed Aroclor
1254 had a higher percentage of early eluting (gas
chromatography) isomers than did that which
remained in solution.
DISCUSSION
The original intent for conducting the work de-
scribed was to find the absolute solubility of Aroclor
1254 in fresh and salt water. This, unfortunately, was
not completely accomplished to any accurate degree,
because a series of significant problems occurred at
the beginning of the Centrifugation experiments. Re-
covery of Aroclor 1254 after Centrifugation was low
and, hence, led to the discovery that adsorption
occurred on the walls of the polyacetate centrifuge
tubes as well as on Polyallomer and stainless steel
centrifuge tubes. Ultimately, only the stainless steel
centrifuge tubes were used in the adsorption and
ultracentrifugal studies.
The "apparent disappearance of early eluting
isomers, such as shown in Table 2, has been observed
by others. It was found to occur in the eggs of the
double-crested cormorant and regarded as possibly
due to metabolic breakdown (Hutzinger et ai, 1972).
Similar behavior in the carcasses of bobwhite quail
after exposure to Aroclor 1254 was observed and
believed to be because of isomeric transformations
(Bagley and Cromartie, 1973). Application of Aroclor
1254 to different types of soil showed a reduced recov-
ery of the early eluting, lower chlorinated biphenyls
(Iwata, Westlake and Gunther, 1973), and it was pos-
Table 7. Distribution of isomers of Aroclor 1254 type II emulsion on stand-
ing in stainless steel centrifuge tubes
Storage
1
5
8
13
Hrs In
tube
0
0
2
2
0
2
2
0
2
2
Pg/J
310
115
97
12
112
102
8.0
97
86
6.1
water
wat
wat
ads
wat
wat
ads
wat
wat
ads
phase
bed
phase
bed
phase
bed
1
93
53
49
96
51
48
69
47
43
47
%
2
90
71
67
106
67
66
82
64
59
68
Peak
Peak
3
98
73
69
103
71
68
85
68
66
77
heights'
number-^
4
99
91
83
127
82
79
104
81
78
94
5
98
98
100
119
96
98
107
97
92
101
6
100
98
100
100
97
98
100
98
96
98
7
100
100
100
100
100
100
100
100
100
100
'Compared to standard Aroclor 1254 (Fig. 1). Calculations are based on
the relative heights of peak 7.
2Peak numbers are shown on the chromatogram in Fig. 1.
120
-------
Theoretical model and solubility characteristics of Aroclor® 1254 in water
943
Table 8. Distribution of isomers in the adsorbed Fraction of Aroclor 1254 type
III emulsion on standing in stainless steel centrifuge tubes
7. Peak heights1
NaCl
hrs in water phase adsorbed
(f>/
-------
944
W. P. SCHOOR
(c) How can the solubility characteristics and field
conditions be best simulated in the laboratory?
Such information would undoubtedly result in more
precise and relevant data on acute toxicity as well
as long-term effects regarding aqueous bioassay of
water-insoluble test compounds.
Acknowledgements—The author thanks Messrs. D. Lamb
and W. Burgess for assistance with the analytical work,
Mr. A. J. Wilson, Jr. for the chromatographic column
packing. Dr. D. R. Nimmo for the shrimp exposed to
Aroclor 1254. and Dr. Ralph Birdwhistell, Dean, the School
of Chemistry, University of West Florida, for reviewing
the manuscript.
REFERENCES
Baglcy G. E. and Cromartie E. (1973) Elimination pattern
of Aroclor 1254® components in the bobwhite. J. Chro-
mal. 75, 219-226.
Bowman M. C, Acree F. Jr. and Corbett M. K. (I960)
Solubility of carbon-14 DDT in water. J. agric. Fd Chem.
8. 5. 406-408.
Cooley N. R., Keltner J. M. Jr. and Forester J. (1972)
Mirex and Aroclor 1254®: Effect on and accumulation
by Tetrahymena pyriformis Strain W. J. Protozool. 19,
4. 636-638.
Duke T. W., Lowe J. I. and Wilson A. J. Jr. (1970) A
polychlorinated biphenyl (Aroclor® 1254) in the water,
sediment, and biota of Escambia Bay, Florida. Bull.
Environ. Contain. Toxicol. 5. 2. 171-180.
Hansen D. J., Parrish P. R., Lowe J. I., Wilson A. J. Jr
and Wilson P. D. (1971) Chronic toxicity, uptake and
retention of Aroclor 1254® in two estuarine fishes. Bull.
Environ. Contain. Toxical. 6, 2, 113-119.
Hutzinger O., Safe S. and Zitko V. (1972) Polychlorinated
biphenyls. Analabs Res. Notes 12, 2, 1-11.
Iwata Y., Westlake W. E. and Gunther F. A. (1973) Vary-
ing persistence of polychlorinated biphenyls in six Cali-
fornia soils under laboratory conditions. Bull. Environ.
Contain. To.\icol. 9, 4, 204-211.
Lowe J. I., Parrish P. R., Patrick J. M. Jr. and Forester
J. (1972) Effects of the polychlorinated biphenyl Aroc-
lor® 1254 on the American oyster (Crassostrea virginicd).
Mar. Biol. 17, 3, 209-214.
Nimmo D. R., Blackman R. R., Wilson A. J. Jr. and Fores-
ter J. (1971a) Toxicity and distribution of Aroclor® 1254
in the pink shrimp (Penaeus duorarum). Mar. Biol. 11,
3, 191-197.
Nimmo D. R., Wilson P. D., Blackman R. R. and Wilson
A. J. Jr. (1971b) Polychlorinated biphenyl absorbed from
sediments by fiddler crabs and pink shrimp. Nature,
Land. 231, 50-52.
Pauling L. (1940) Nature of the Chemical Bond, p. 164.
Cornell University Press, Ithaca.
Schoor W. P. (1973) In vivo binding of p,p'-DDE to human
serum proteins. Bull. Environ. Contam. ToxicoL 9, 2, 70-
74.
Walsh G. E. (1972) Insecticides, herbicides and polychlor-
inated biphenyls in estuaries. J. Wash. Acad. Sci. 62, 2,
122-139.
Zitko V. (1970) Polychlorinated biphenyls solubilized in
water by nonionic surfactants for studies of toxicity to
aquatic animals. Bull. Environ. Contain. Toxicol 5, 3
219-226.
122
-------
Reprinted from the Trans-
actions of the American
Fisheries Society, Vol. 104
(2): 388-389, 1975, with per-
mission of the American
Fisheries Society
A SALINITY CONTROLLER FOR FLOW-THROUGH BIOASSAYS
Lowell H. Bahner and Del Wayne R. Nimmo
Contribution No. 214
123
-------
Made in United States of America
Reprinted from TRANSACTIONS OF THE AMERICAN FISHERIES SOCIETY
Vol. 104, No. 2, April 1975
pp. 388-389
© Copyright by the American Fisheries Society, 1975
A Salinity Controller for Flow-through Bioassays1
LOWELL H. BAHNER AND DELWAYNE R. NIMMO
U. S. Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Sabine Island, Gulf Breeze, Florida 325612
ABSTRACT
An electro-mechanical device has been constructed to monitor and dilute seawater to a con-
stant salinity for flowing-water bioassays. It has been used successfully in pesticide bioassays
and requires little maintenance.
Salinity is one variable that is most
difficult to control in estuarine flow-
through bioassays. Two possible ways
of controlling salinity are: (1) to adjust
large water reserves in a tank to a given
salinity; or (2) to adjust the incoming salin-
ity continuously by dilution or by adding
artificial salts. Currently our flow-through
bioassays require 75 liters of water/
hour/tank; therefore, the use of water re-
serves or artificial salt is impractical.
At the Gulf Breeze Environmental Re-
search Laboratory situated on Santa Rosa
Sound, near Pensacola, Florida, water
pumped from the estuary is normally
above 20 °/oo throughout the year. We
have designed and constructed a device to
monitor and adjust any volume of incoming
salt water, continuously, by dilution to 20
%o ± 1 °/oo. This device has been used
Contribution No. 214, Gulf Breeze Environmental
Research Laboratory.
"Associate Laboratory of the National Environmen-
tal 'Research Center, Corvallis, Oregon.
DETECTOR
FIGURE 1.—Block diagram of salinity control ap-
paratus.
successfully for several months in pesticide
bioassays and requires little maintenance.
The salinity control device (Fig. 1) con-
sists of a detector that monitors salinity
and a solid-state electronic amplifier and
relay (Fig. 2) that controls a washing-
machine solenoid valve to regulate fresh
water input. In our system, estuarine
water is pumped into a mixing tank, where
temperature and salinity are adjusted. It
then flows to a constant-head trough,
where the salinity is monitored before the
water is siphoned into the bioassay tanks.
The detector consists of a light source, a
POWER SUPPLY
DC AMPLIFIER
SOLID-STATE LAMP
FOR DETECTOR
FIGURE 2.—Schematic diagram of electronic
amplifier, solid-state lamp, and suitable power supply.
Rl-50,000, R2-3,300, R3-500, R4-680, R5-220, R6-270,
Cl-2000, C2-100, C3-0.1, C4-100, Dl-4 silicon rectifier
bridge, D5-12 volt zener diode, D6-germanium diode,
D7-silicon rectifier, Q1-HEP52, Q2-HEP230,
Q3-ECG123A, Q4-ECG128, LED-Radio Shack
276-026, PCl-Radio Shack 276-116, Kl-12 volt relay,
Tl-12.6 VCT secondary. All resistances in ohms; all
capacitances in microfarads.
388
125
-------
BAHNER AND NIMMO—SALINITY CONTROLLER
389
photocell, and a modified sea water
hydrometer. Small vanes of opaque tape
are attached to the stem of the hydrometer
to break the light beam to the photocell,
when the salinity rises above the desired
level. A signal from the detector activates
the solenoid valve, allowing fresh water to
enter. The detector assembly is housed in
an acrylic enclosure, large enough to per-
mit adjustment of the height of the light
source and photocell to obtain a desired
salinity. The detector assembly is mounted
on a floating styrofoam base so that water
level variations do not affect salinity con-
trol. The amplifier and relay are mounted
in a separate aluminum box.
126
-------
Reprinted from Applied Mic-
robiology, Vol. 29(1): 125-
127, 1975, with permission
of the American Society for
Microbiology
EFFECT OF POLYCHLORIMATED BIPHENYL FORMULATIONS
ON THE GROWTH OF ESTUARINE BACTERIA
«
Al W. Bourquin and S. Cassidy
Contribution No. 217
127
-------
APPLIED MICROBIOLOGY, Jan. 1975, p. 125-127
Copyright © 1975 American Society for Microbiology
Vol. 29, No. 1
Printed in U.S.A.
Effect of Polychlorinated Biphenyl Formulations on the
Growth of Estuarine Bacteria1
AL W. BOURQUIN* AND S. CASSIDY
U.S. Environmental Protection Agency, Gulf Breeze Environmental Research Laboratory, Sabine Island, Gulf
Breeze, Florida 32561
Received for publication 11 November 1974
Polychlorinated biphenyl formulations inhibited the growth of certain estua-
rine bacteria. The sensitive strains, although exhibiting some similar physiologi-
cal characteristics, contained both gram-positive and gram-negative bacteria.
Polychlorinated biphenyl formulations
(PBCs) called Aroclors (Monsanto Co., St.
Louis, Mo.) have been used commercially as
lubricants, plasticizers in paints, plastics, and
chlorinated rubbers, as heat-exchange fluids in
industrial heating systems, and dielectric com-
pounds in large electrical transformers and
capacitors (9). Aroclor formulations are identi-
fied by four-digit numbers, the first two indicat-
ing the type of molecule and the last two
indicating the weight percentage of chlorine in
the molecule. PBCs have been reported as
environmental pollutants from water, sediment,
and biota (2, 3). The ubiquity and significance
of various Aroclor formulations have been re-
viewed (3, 7, 9).
Despite their ubiquity (7) and effects on
estuarine and marine organisms (1, 6), only
cursory information has been accumulated on
the interactions of PCBs and microorganisms. A
commercial PCB formulation (Aroclor 1254) has
been reported to stimulate in vitro growth of
Escherichia coli, a bacterium from human in-
testinal microflora used as an indicator of water
quality (4). Inhibition of bacterial growth by
PCBs has not been previously reported. This
study employs a rapid sensitivity disk assay
method to determine the effects of Aroclor 1016
and 1242 on the growth of selected estuarine
bacteria.
Eighty-five bacterial isolates from various es-
tuarine environments near Pensacola, Florida,
were examined for ability to grow on solid
medium in the presence of varying concentra-
tions of Aroclor 1016 and 1242. Each liter of test
medium contained 1.0 g of yeast extract (Difco),
5.0 g of beef peptone (Difco) and 20.0 g of agar
(BBL), in aged, artificial seawater (Rila Marine
Mix, aged 1 month in dark), pH 7.4 and 2.0%
1 Gulf Breeze Environmental Research Laboratory Contri-
bution No. 217.
salinity. Cells for inocula were grown in this
medium without agar for 18 h at 28 C on a
rotary shaker. This culture was diluted 1:1 with
sterile 2.0% seawater, and 0.1 ml was spread on
the agar medium. Absorbant paper disks
(Schleicher and Schuell, Inc., no. 740-E, 12-mm
diameter) were saturated with 0.1 ml of acetone
solutions containing 1.0, 2.5, and 5.0 mg of PCB
formulation per ml and air-dried for 24 h at
room temperature before use. The PCB-treated
paper disks and control disks treated only with
acetone were arranged on the agar surface
within 2 h of inoculation. Duplicates of each test
were prepared. The cultures were incubated for
24 h at 28 C and examined for sensitivity.
Sensitivity was defined as a zone of inhibition
surrounding the paper disk (Fig. 1).
Of 85 different isolates tested, growth of 26
was inhibited to varying extents by 0.5 mg of
either PCB formulation. Zones of inhibition
ranged in size from 14 to 20 mm. Cultures
sensitive to Aroclor 1242 were also inhibited by
Aroclor 1016. Sixty-five percent of the cultures
sensitive to 0.5 mg of Aroclor 1242 were still
sensitive at 0.1 mg, and 58% of cultures sensi-
tive to Aroclor 1016 at 0.5 mg were sensitive at
0.1 mg.
Identified genera sensitive to PCBs included
Flauobacterium, Bacillus, Corynebacterium,
Pseudomonas, Achromobacter, Micrococcus,
and Serratia. Representative cultures of the
sensitive bacteria from the disk assay were
tested for growth inhibition in liquid culture.
Four bacteria—two gram-positive and two gram-
negative—were monitored for 30 h in nutrient
marine broth and nutrient marine broth plus
Aroclor 1241 (10 /xg/ml). Growth inhibition of
all four test bacteria by Aroclor 1242 is shown
in Fig. 2. The nature of the inhibition is un-
known; however, the greatly extended lag
phase of all test bacteria suggests bacterio-
stasis.
129
-------
126
NOTES
APPL. MICROBIOL.
Fi<;. 1. Effect of PCB-treated disk on growth of estuarine bacteria. Lower middle disk is acetone control' left
to right, 0.5 mg of 1242, 0.5 mg of 1016, 0.5 mg of 1242. 0.5 mg of 1016.
E
o
2
200
Klo
80
60
40
20
CULTURE NUMBER
0 ; 100 * = 31
= 60 • = 9
CONTROLS
16 20
TIME hour.
iti
12
Fir.. 12. Four sensitive bacteria were tested for growth in one-half strength 2216 Marine Broth containing 1(1
ng of Aroclor 1242 per ml. Symbols: (O) Achromobacter up.. (•) Bacillus sp.. (A) Cornyebacterium up., and (D)
Pseudomonas sp. Controls contain one-half strength 2216 Marine Broth only: all salinities were 20'7< .
130
-------
VOL. 29, 1975
NOTES
127
TABLE 1. Comparison of physiological activities of PCB-test bacteria"
Bacteria
tested
Sensitive
Total tested
Production of:
Ureaae
19
13
Amylase
76
33
Lipase
33
23
Gelatinase
86
42
H2S
5
6
Nitrate
reduction
29
37
Citrate
utilization
48
41
Gram
reaction
50
32
"Percentage of cultures showing positive reaction.
The PCB-sensitive bacteria included both
gram-positive and gram-negative isolates; how-
ever, compared to the test population, a slightly
greater percentage of the sensitive strains were
gram-positive (Table 1). Previous reports indi-
cated that the growth of gram-positive bacteria
was inhibited by organochlorine insecticides
(technical chlordane), whereas that of
gram-negative organisms was unaffected (8,
10). These results suggest that the growth of
some gram-negative estuarine bacteria (13 orga-
nisms) was sensitive to PCBs, a mixture of
organochlorine molecules. When PCB-sensitive
and nonsensitive strains were compared (Table
1), the biochemical activities of test bacteria
showed two different groups. Amylase was pro-
duced by 76% and gelatinase by 86% of the
sensitive strains, whereas in the total test popu-
lation only 33 and 42% showed these activities.
Twenty of the 28 amylase producers were also
sensitive to PCBs, as were 22 of the 36 gelatin-
ase producers. Further investigation is needed
to determine whether the observed prevalence
of PCB sensitivity among such amylase and
gelatinase producers is significant in relation to
total nutrient catabolism.
The sensitivity disk testing procedure com-
monly used in clinical bacteriology provides a
rapid method for detecting estuarine microorga-
nisms that are sensitive to commercial PCB
preparations. Although the procedure is quali-
tative only, it eliminates many hours of tedious
experimentation in selecting suitable test orga-
nisms for more detailed examinations. These
results indicate that a sizeable proportion of
estuarine microbes tested were sensitive to
commercial PCB formulations. If present in
estuaries at the concentrations tested, these
PCB formulations could disrupt microbial het-
erotrophic activity.
PCBs persist in estuarine sediments in rela-
tively high concentrations. Sediments from
Escambia Bay, Florida, contained concentra-
tions ranging from 0.6 to 61.0 mg/kg dry weight
(2) and sediments at an industrial outfall in the
same bay system up to 486 mg/kg dry weight
(6). When considering the microenvironment of
a sediment core, discrete organic particles,
potential microbial substrates, may contain a
much greater concentration of PCBs than that
reported for the whole core. Such PCB concen-
trations are similar to, or greater than, concen-
trations found inhibitory to heterotrophic
growth in our study. In nature, PCBs and
bacteria are adsorbed on particulate surfaces;
consequently, PCBs, by inhibition of specific
heterotrophic bacteria attached to organic de-
tritus, could inhibit normal turnover of carbon in
estuarine sediments.
We thank L. Keifer and C. Shanika for technical assist-
ance.
LITERATURE CITED
1. Cooley, N. R., J. M. Keltner, Jr., and J. Forester. 1973.
The polychlorinated biphenyl Aroclor 1248 and 1260;
effect on and accumulation by T. pyriformis. J. Proto-
zool. 20:443-445.
2. Duke, T. W., J. I. Lowe, and A. J. Wilson, Jr. 1970. A
polychlorinated biphenyl, Aroclor 1254, in water, sedi-
ment, and biota of Escambia Bay, Florida. Bull.
Environ. Contam. Toxicol. 5:171-180.
3. Dustman, E. H., L. F. Stickel, L. J. Blus, W. L. Reichel,
and S. N. Wiemeyer. 1971. The occurrence and signifi-
cance of polychlorinated biphenyls in the environment,
p. 118-133. Trans. 36th N. Amer. Wildl. Nat. Resour.
Conf.
4. Keil, J. E., S. J. Sandifer, C. D. Graber, and L. E.
Priester. 1972. DDT and polychlorinated biphenyl
Aroclor 1242. Effects of uptake on E. co/i growth. Water
Res. 6:837-841.
5. Lowe, J. E., P. R. Parrish, J. M. Patrick, Jr., and J.
Forester. 1972. Effects of polychlorinated biphenyl
Aroclor 1254 on the American oyster, Crassostrea
uirginica. Mar. Biol. 17:209-214.
6. Nimmo, D. R., P. D. Wilson, R. R. Blackman, and A. J.
Wilson, Jr. 1971. Polychlorinated biphenyl absorbed
from sediments by fiddler crabs and pink shrimp.
Nature (London) 231:50-52.
7. Risebrough, R. W., P. Reiche, C. B. Peakall, S. G.
Herman, and M. N. Kirven. 1968. Polychlorinated
biphenyls in the global ecosystem. Nature (London)
220:1098-1102.
8. Trudgill, P. W., R. Widdus, and J. S. Rees. 1971. Effects
of organochlorine insecticides on bacterial growth,
respiration and viability. J. Gen. Microbiol. 69:1-13.
9. U.S. Interdepartmental Task Force on PCB's. 1972.
Publication no. COM-72-10419. National Technical
Information Service, U.S. Department of Commerce,
Springfield, Va.
10. Widdus, R., P. W. Trudgill, and D. C. Turnell. 1971. The
effects of technical chlordane on growth and energy
metabolism of Streptococcus fecalis and Mycobacte-
rium phlei: a comparison with Bacillus subtilis. J. Gen.
Microbiol. 69:23-31.
131
-------
Reprinted from Proceedings of
the 28th Annual Conf. of
Southeastern Assoc. of Game
and Fish Comm., Nov 17-20,
1974, pp. 187-194, with
permission of the Southeastern
Assoc. of Game and Fish Comm.
ENDRIN: EFFECTS ON SEVERAL ESTUARINE ORGANISMS
Steven C. Schimmel, Patrick R. Parrish, David J. Hansen,
James M. Patrick, Jr., and Jerrold Forester
Contribution No. 218
133
-------
Reprinted from the Proceedings of the 28th Annual Conference of the Southeastern Association of
Came and Fish Commissioners, 1974.
ENDRIN: EFFECTS ON SEVERAL ESTUARINE ORGANISMS1
by
Steven C. Schimmel, Patrick R. Parrish2,
David J. Hansen, James M. Patrick, Jr. and Jerrold Forester
U. S. Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Sabine Island, Gulf Breeze, Florida 32561
(Associate Laboratory of the National Environmental
Research Center, Corvallis Oregon)
ABSTRACT
Acute (96-hour) bioassays were performed with endrin and the following estuarine organisms: American oyster (Crassoxtrea
virginica), pink shrimp(Penaeus duorarum). grass shrimp (Palaemonetespugio). sailfin molly {Poet-ilia lalipinna) and sheepshead
minnow (Cyprinodon variegatus), Endrin was acutely toxic to all organisms tested, except oysters, whose shell growth was ap-
preciably inhibited by 56 ug/1 (parts per billion) of the chemical. Pink shrimp were the most sensitive animal tested, but significant
numbers of both species of shrimps and fishes died when exposed to concentrations of one ug/1 or less. In'a separate test, embryos
and fry of the sheepshead minnow were exposed to concentrations of endrin ranging from 0.046 to 1.0 ug/1 (nominal) for 33 days in
an intermittant-flow bioassay. Embryos were not affected by the concentrations to which they were exposed, but the estimated
LC50 (probit analysis, a=.05) of fry was 0.27 ug/1.
INTRODUCTION
Widespread use of the organochlorine insecticide, endrin, has prompted numerous
investigations to determine the effects of this compound on aquatic organisms. Several
studies involving marine organisms have shown that endrin is acutely toxic at low
levels. Eisler (1969) found endrin acutely toxic to sand shrimp (Crangon
septemspinosa), grass shrimp (Palaemonetes vulgaris), and hermit crabs (Pagurus
longicarpus). The 96-hour LCSO's were 1.7 ug/1, 1.8 ug/1; and 12 ug/1, respectively.
Davis and Hidu (1969) assessed the.effects of endrin on oysters by (1) determining the
number of fertilized eggs that developed into normal larvae after 48-hours exposure to
a given concentration of endrin, and (2) observing survival and growth of larvae over a
period of 12 days. At concentrations greater than 0.025 mg/1, endrin reduced the
number of eggs developing, survival, and growth of larvae. Katz (1961) and Katz and
Chadwick (1961) found the 96-hour LC50 of endrin to threespine stickleback
(Gasterosteus aculeatus) ranges from 0.5 to 1.5 ug/1. Salinity had little effect on tox-
icity, but temperature markedly affected toxicity the higher the temperature the
greater the toxicity. Eisler and Edmunds (1966) found that acute exposure to sublethal
concentrations (1 ug/1 or less) of endrin impaired liver function in northern puffers
(Sphoeroides maculatus).
Few data have been published concerning the effects of endrin on larval marine
fishes. One such study by Johnson (1967) on the threespine stickleback (Gasterosteus
aculeatus) demonstrated that endrin immobilizes hatching fry at 15.0 ug/1 and
'Contribution No. 218 Gulf Breeze Environmental Research Laboratory.
2Present Address: Bionomics Marine Laboratory, Rt. 6, Box 1002, Pensacola, Florida 32507
187
135
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produced behavioral changes in fry reared in water containing 2 and 5 ug/1.
Our study was conducted to determine (1) the 96-hour LC50 of endrm to pink
shrimp (Penaeus duorarum), grass shrimp (Palaemonetes pugio), sheepshead min-
nows (Cyprinodon variegatus), and sailfin mollies (Poecilia latipinna), (2) the effect of
endrin on shell growth of American oysters (Crassostrea virginica), and (3) the effect of
endrin on egg fertility, hatching success of embryos, and survival of fry of the sheep-
shead minnow.
We thank Johnny Knight for his chemical analyses of water samples and Steven S.
Foss for his work on the illustration.
MATERIALS AND METHODS
Test Animals
All test animals were collected near the Gulf Breeze Environmental Research
Laboratory in Florida and acclimated to laboratory conditions for at least ten days
before exposure. If mortality in a specific lot of animals exceeded 1% in the 48 hours
immediately preceding the test or if abnormal behavior was observed during ac-
climation, the entire lot was discarded. Oysters tested were from 27 to 54 mm in height;
pink shrimp, 39 to 70 mm rostrum-telson length; grass shrimp, 19 to 36 mm rostrum-
telson length; sheepshead minnows, 15 to 25 mm standard length; and sailfin mollies,
36 to 49 mm standard length. Animals were not fed during acute toxicity tests, but they
could obtain food (plankton and other paniculate matter) from the unfiltered sea
water in which they were maintained. Adult sheepshead minnows, 35 to 50 mm stan-
dard length were used to produce eggs used in studies of fertility, hatching success, and
fry survival. Fry were fed daily, using live brine shrimp (Anemia salind) nauplii which
contained no pesticide or polychlorinated biphenyl detectable by our gas
chromatographic analysis.
96-Hour Test Conditions
Acute toxicity of endrin was determined by exposing 20 animals per aquarium (ex-
cept 15 sailfin mollies per aquarium) to logarithmic concentrations for 96 hours. A 30-
liter aquarium was used for each concentration. Technical grade endrin (96% active
ingredient) was dissolved in reagent grade acetone and metered by Milroyal® pumps
(Lowe, et al., 1972) at 60 ml/hr into unfiltered sea water that entered each aquarium at
150 1/hr. A control aquarium received the same quantity of water and solvent, but no
endrin.
Effect of endrin was assessed by., measuring reduction of shell growth of oysters
(Butler, 1962) and by determining mortality in shrimps arid fishes.
Sheepshead Minnow Embryo and Fry Test Conditions
The exposure apparatus used in the sheepshead minnow embryo ai.J fry test was a
modification of the dilutor designed by Mount and Brungs (1967). Each discharge of
our dilutor delivered 500 ml of salt water and a constant concentration of carrier
(polyethylene glycol 200) to control aquaria and to aquaria receiving each of five
concentrations of toxicant. Another control aquarium received 500 ml of salt water
without carrier. The Mount and Brungs apparatus consists of two tiers of dilution cells
and a mixing chamber; in our apparatus, (Figure 1), we added a tier of three cells (F)
situated above the W-cell series of Mount and Brungs. One of these cells (F-l) provided
500 ml of seawater without carrier directly to the exposure tank. One of the larger cells
(F-2) delivered 2250 ml of seawater to the toxicant-mixing box (M-2); the other 2250
ml cell (F-3) to the carrier-mixing box (M-l) which discharged into the W-cell series.
Two injectors with 50 ml syringes were used: one provided endrin in 0.018 ml of carrier
per cycle to the toxicant-mixing box; the second provided an identical quantity of
"Registered trademark, Milton Roy Co., Philadelphia, Pa. Mention of commercial products does not constitute endorsement
by the Environmental Protection Agency.
188
136
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carrier without endrin to the carrier-mixing box. Both injectors were activated
mechanically by the filling and draining of a bucket (not illustrated) situated below the
excess water drain in the reservoir (R). A float switch in the bucket shut off the water
pump (P) at the beginning of each cycle and opened it at the end of the cycle. Since
volumes of carrier were equal in each injection, the same amount of polyethylene glycol
200 was administered to each exposure tank, regardless of the dilution. Excess water
containing endrin was discarded prior to cycling of the dilutor.
VoMo
Figure 1. Exposure apparatus:
E Endrin and seawater cells; Ex-Excess water; F Filtered seawater
cells; 1-1 carrier injector; 1-2 - carrier and endrin injector; M-l - carrier
mixing box; M-2 - carrier and endrin mixing box; P - Seawater pump; R
Reservoir; Va Ma - vacuum manifolds; Va V Vacuum venturi; W - Sea-
water with carrier.
189
137
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Seawater used in this bioassay was pumped from Santa Rosa Sound, Florida
through a swimming-pool sand filter and a 1 u-pore polypropylene filter into a
constant head box in the laboratory. In the box, the water was heated to 30 C •* 1 C
while the salinity varied with that of the Sound water (15 to 28 o/oo, average 23.4
o/ oo). The water was then pumped to the toxicant-delivery apparatus. Our dilutor
cycled approximately 80 times each day, delivering 125 ml water to each of four 1200
ml exposure chambers per concentration per cycle.
Eggs of the sheepshead minnow were obtained and fertilized by procedures des-
cribed by Schimmel et al. (1974). Twenty eggs were placed in 10-cm Petri dishes to
which a 9 cm high nylon screen collar (0.5 mm mesh) was attached. This collar
permitted water exchange while preventing escape of fry. Exposure of the embryos
began one hour after fertilization and lasted for 33 days.
Concentrations of endrin were calculated to give 1.0, 0.46, 0.21, 0.1 and 0.046 ug of
endrin/liter of water. Concentrations, measured weekly were typically within 35-60%
of the intended concentration. Dissolved oxygen concentrations, determined weekly
by the modified Winkler method of Strickland and Parsons (1968), were above 50%
saturation and appeared adequate.
Chemical Analyses
Concentrations of endrin in water and animals were determined by electron-capture
gas chromatography, using a 182 cm x 2mm ID glass column packed with 2%OV-101
on 100-120 mesh Gas Chrom Q. Nitrogen flow rate was 25 ml/min, the oven
temperature was 190°C, and the injector and detector temperatures were 210°C.
Recovery exceeded 85%; data were not adjusted for recovery. Sensitivity of this
method was 0.004 ppb when using a 1-liter sample. Unfiltered water samples from each
concentration and control were analyzed once during the acute 96-hour bioassays. In
the study of sheepshead minnow fertility, hatching success, and survival, water samples
from each concentration and controls were analyzed weekly. Whole-body concen-
trations in surviving animals were determined on wet weight basis.
Tissue samples that weighed more than 5 g. were prepared using the methods des-
cribed by Lowe et al. (1972) except that endrin was eluted in the 15% (v/v) ethyl
ether-in-petroleum ether fraction. Samples from 1-5 g. were analyzed using the semi-
micro method described by Hansen et al. (1974) except that endrin was eluted in 20%
(v/v) ethyl ether/hexane fraction, while those less than 1 g were analyzed by the micro
method described in the Pesticide Analytical Manual, Volume III (U. S. Food and
Drug Administration, 1970). Sensitivity of each method wasO.OlOppm when usinga 1
gram sample.
Statistical Analyses
Data from the 96-hour oyster shell growth study were analyzed by linear regression;
shrimp and fish mortality data were analyzed by the probit method of Litchfield and
Wilcoxon (1949). None of the bioassay results were rejected by the Chi-square test for
variation.
Data from the study of sheepshead minnow fertility, hatching success and survival
were analyzed by the Chi-square test (a =0.01) to determine differences in mortality of
experimental and control animals and probit analysis to determine the LC50 (a =0.05).
RESULTS AND DISCUSSION
A cute (96-hour) Exposures
With the exception of oysters, endrin was acutely toxic to all organisms tested
(Tables 1 and 2). Pink shrimp were the most sensitive, but significant numbers of both
species of shrimps and fishes died when exposed to concentrations of 1 ug/1 (part per
billion) or less. Shell growth in oysters was appreciably inhibited by exposure to a
concentration of 56 ug/1 (32 ug/1 measured) for 96 hours. Although the LC50 of
Palemonetespugio (0.73 ug/1) in our tests compares favorably with Eisler's (1969) data
for P. vulgaris(\.S ug/1), we found that the shrimp, Penaeusduorarum had an LC50 of
0.049 ug/1. This concentration is about 9 times greater than the lower limit of analytical
detectibility of endrin in a J.O 1 water sample.
190
138
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Table 1. Acute toxicity of endrin to and uptake by American oysters (Crassostrea
vir'ginica), pink shrimp (Penaeus duorarum), grass shrimp (Palaemon-
etes pugio), sheepshead minnows (Cyprinodon variegatus), and sailfin
mollies (Poecilia latipinna) in relation to concentration of endrin in sea-
water during 96-hour exposures. Whole-body residues are from animals
alive at end of exposure.
SPECIES
C. virginica
CONCENTRATION IN WATER
(ug/I)
Intended Measured
P. duorarum
P. pugio
C. variegatus
P. latipinna
Control
1.8
5.6
18.
56.
180.
Control
0.010
0.032
0.10
0.32
1.0
Control
0.032
0.1
0.32
1.0
3.2
Control
0.010
0.032
0.10.
0.32
1.0
Control
0.0135
0.075
0.135
0.75
1.35
NDa
1.6
4.9
13.
32.
168.
NDa
0.009
0.023
0.077
0.28
0.88
NDa
0.024
0.081
0.24
0.96
2.4
NDa
0.019
0.026
0.10
0.33
0.95
NDa
0.012
0.073
0.150
0.60
1.20
EFFECTb
0
13
40
47
67
87
0
5
30
80
100
100
0
0
5
25
60
100
0
0
0
0
20
100
7
0
0
0
47
100
WHOLE-BODY
RESIDUE
(ug/g, wet weight)
NDa
1.4
5.8
20.
52.
124.
NDa
Trace
0.025
0.067
NDa
0.02
0.07
0.19
0.02
Trace
0.013
0.11
0.30
1.5
0.013
0.035
0.17
0.26
1.7
aND = 0.004 ug/1 in water, 0.010 ug/g in tissue.
bEffect is expressed as percentage reduction in shell growth for oysters and death
for shrimps and fishes.
191
139
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Table 2. Acute toxicity (EC50, 96-hr) of cndrin to American oysters (Crassosirea vir&nica), pink shrimp (Penaeits duorarum), grass
shrimp (Palaemoneiei. pugio), sheepshead minnows (Cyprinodon variegatus) and saillm mollies (Puecilia latipinna). Ef-
ect is expressed as percentage reduction in shell growth for oysters and death lor shrimps and fishes. Confidence limits
are in parentheses.
SPECIES 96-HOUR EC50 TEMPERATURE SALINITY
(°C) (o/oo)
C. virginica
P duorarum
P. pugio
C. variegatus
P latipinna
Intended
19.1
(5.3-68.78)
0.049
(0.034-0.070)
0.73
(0.40-1.32)
0.40
(0.30-0.53)
0.79
(0.60-1.05)
Measured
14.2
(3.99-50.49)
0.037
(0.025-0.053)
0.63
(0.35-1.15)
0.38
(0.31-0.45)
0.63
(0.47-0.84)
Mean
22.0
14.8
12.6
17.4
19.6
Range
19.0-24.0
12.0-16.0
10.0-14.5
16.5-19.0
18.0-21.0-
Mean
29.3
28.4
27.2
19.5
28.2
Range
27.0-30.5
26.0-31.0
25.5-31.0
14.0-22.0
23.5-30.5
-------
Endrin, at the concentrations tested, appeared to have no significant effect on
fertility of sheepshead minnow eggs or survival of embryos (Table 3). The LC50 for fry,
however, was estimated to be 0.267 ug/1, indicating that this is the most sensitive stage.
Chi-square analysis of fry-survival data showed significant mortality in experimental
groups exposed to 1.0 and 0.46 ug/1. In these concentrations, most fry died within one
week of the time required for a 50% hatch (approximately 5 days for all concen-
trations). Observed signs of endrin poisoning were: flared opercles, erratic swimming,
failure to feed, lethargic behavior and loss of equilibrium. Although the group exposed
to 0.21 ug/1 had no significant mortality, their lethargic behavior and loss of
equilibrium was evidence of effect. Survival of embryos and fry in the control aquaria
was 76% and that of the control with carrier was 75%. In whole fry from experimental
groups, concentration factors (concentration of endrin in the fry divided by the
concentration in the water) ranged from 3.3 to 4.8 x 103. No detectable concentrations
( 0.03 ug/g) were found in the controls.
Effects from the polyethylene glycol 200 carrier at the concentrations employed were
not anticipated because none was observed in preliminary bioassays of this carrier.
Static tests using a control and five concentrations of the carrier (11.2, 23.6, 51.8, and
112.6 g/liter of sea water) were conducted in separate tests on embryos and on newly-
hatched fry. Results of the 7-day embryo test (30° C, 25 o/ oo) indicate that the LC50 of
polyethylene glycol 200 was 33.0 g/liter (Probit analyses; a =0.05). The estimated 96-
hour LC50 for newly-hatched fry (24 hrs.) was 18.0 g/ liter. (The extended duration of
the embryo test allowed for observation of fry after hatching). Because high levels of
polyethylene glycol were required to produce an effect, the effect may be the result of
osmotic stress rather than direct toxicity of the carrier. The maximum concentration of
carrier (112.6 g/liter) when mixed with 25 o/oo seawater is osmotically equivalent to
70 o/oo seawater. Maximum solvent concentration used in the study of sheepshead
minnow fertilization, hatching success and survival was only 0.009 mg/ liter of water.
Table 3. Relative susceptibility of various life stages of sheepshead minnows (Cy-
prinodon variegatus) to endrin in flow-through systems. Criteria are fer-
tility of eggs and survival of embryos, fry and juveniles. Confidence limits
(95%) are in parentheses.
LIFE STAGE EXPOSURE LC50 (ug/1)
(days) Intended Measured
Egg fertilizationa 0.25 No observable effect at 1.0 ug/1
Embryos 5.0 No observable effect at 1.0 ug/1
Fry 33.0 0.267(0.08-0.62) 0.158(0.00-0.57)
Juveniles 4.0 0.40 (0.30-0.53) 0.38 (0.31-0.45)
aStatic test.
Results of this study indicate that addition of endrin to an estuarine system, even at
low concentrations, could adversely affect the estuarine communities. Of particular
concern in the southeastern United States is the potential effect of low concentrations
of endrin on penaeid shrimp populations since our study demonstrates that the LC50
measured for Penaeus duorarum (0.037 ug/liter) is only 9x the present limit of
analytical detectability in water. Long-term studies exposing the shrimp to ultra-low
concentrations of endrin (parts per trillion) could help establish "safe" levels of this
compound in estuaries.
193
141
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LITERATURE CITED
Butler, P. A. 1962. Reaction of some estuarine mollusks to environmental factors.
In: Biological problems in water pollution. Third Seminar. U. S. Public Health
Serv. Publ. No. 999-wp-25, 1965: 92-104.
Davis, H. C. and H. Hidu. 1969. Effects of pesticides on embryonic development of
clams and oysters and on survival and growth of the larvae. U. S. Fish and Wildl.
Serv. Fish. Bull. 67(2):393-404.
Eisler, R. 1969. Acute toxicities of insecticide to marine decapod crustaceans. Crus-
taceana 16:302-310.
and P. H. Edmunds. 1966. Effects of endrin on blood and tissue
chemistry of a marine fish. Trans. Am. Fish. Soc. 95(2):153-159.
Hansen, David J., P. R. Parrishand J. Forester. 1974. Aroclor® 1016:Toxicitytoand
uptake by Estuarine Animals. Environ. Res. (In Press)
Johnson, Howard Ernest, 1967. The effects of endrin on the reproduction of a fresh
water fish (Oryzias latipes). Ph.D. thesis, Univ. Wash. Wash.
Kartz, M. 1961. Acute toxicity of some organic insecticides to three species of salmon-
ids and to the threespine stickleback. Trans. Am. Fish. Soc. 90(3):264-268.
and G. C. Chadwick, 1961. Toxicity of endrin to some Pacific North-
west fishes. Trans. Am. Fish. Soc. 90(4):394-397.
Litchfield, J. T., Jr. and F Wilcoxon. 1949. A simplified method of evaluating dose-
effect experiments. J. Pharmacol. Exp. Ther. 96(2):99-l 13.
Lowe, J. I., P. R. Parrish, J. M. Patrick, Jr. and J. Forester. 1972. Effects of the poly-
chlorinated byphenyl Aroclor® 1254 on the American oyster, Crassostrea virgin-
ica. Mar. Biol. (Berl.) 17:209-214.
Malone, C. R. and Blaylock, B. G. 1970. Toxicity of insecticide formulations to carp
embryos reared in vitro. J. Wildl. Manage. 34(2):460-463.
Mount, Donald 1. and William A. Brungs. 1967. A simplified dosing apparatus for
fish toxicology studies. Water Res. 1:21-29.
Schimmel, Steven C., Hansen, David J. and Jerrold Forester. 1974. Effects of Aro-
clor® 1254 on laboratory-reared embryos and fry of sheepshead minnows (Cy-
prinodon variegatus). Trans. Am. Fish Soc. 103(3):582-586.
Strickland, J. D. H. and Parsons, T. R. 1968. A practical handbook of seawater
analysis. Fish. Res. Board Can. Bull. 167:21-26.
U. S. Food and Drug Administration, 1970. Pesticide Analytical Manual Sect H212
U. S. Dep. Health, Educ., Welfare, Wash., D. C.
194
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Reprinted from Archives of
Environmental Contamination
and Toxicology, Vol. 3(3):
371-383, 1975, with permis-
sion of Springer-Verlag New
York Inc.
SEASONAL EFFECTS OF LEACHED MIREX ON
SELECTED ESTUARINE ANIMALS
M.E. Tagatz, P.W. Borthwick, and J. Forester
Contribution No. 222
143
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SEASONAL EFFECTS OF LEACHED MIREX ON
SELECTED ESTUARINE ANIMALS
M. E. TAGATZ, P. W. BORTHWICK and J. FORESTER
Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Gulf Breeze, Fla. 32561
Four 28-day seasonal experiments were conducted using selected estuarine animals in
outdoor tanks that received continuous flow of mirex-laden water. Mirex
(dodecachlorooctahydro-l,3,4-metheno-2//-cyclobuta [cd] pentalene) leached from
fire ant bait (0.3% mirex) by fresh water and then mixed with salt water was toxic to
blue crabs (Callinectes sapidus), pink shrimp (Penaeus duorarum), and grass shrimp
(Palaemonetes pugio) but not to sheepshead minnows (Cyprinodon variegatus), at
concentrations less than 0.53 jiig/L in water. The amount of leaching was greatest in
summer and least in spring. Greatest mortality occurred in summer at the highest water
temperature and concentration of mirex; least mortality occurred in spring at next to the
lowest temperature and at the lowest concentration. Earliest deaths of blue crabs
occurred after six days of exposure and shrimps after two days. Small juvenile crabs
were more sensitive to leached mirex than were large juveniles. Mirex did not appear
to affect growth or frequency of molting in crabs. All exposed animals concentrated
mirex. Among animals that survived for 28 days, sheepshead minnows concentrated
mirex 40,800X above the concentration in the water, blue crabs 2.300X, pink shrimp
10,OOOX, and grass shrimp 10,800X. Sand substrata contained mirex up to 770X that
in the water. Most control and exposed animals in samples examined histologically had
normal tissues, but alteration in gills of some exposed fish and natural pathogens in
some exposed and control crabs and shrimp were observed. The experiments de-
monstrated that mirex can be leached from bait by fresh water, concentrated by
estuarine organisms, and can be toxic .to crabs and shrimps.
This study was conducted to determine the seasonal effects, on various estuarine
organisms, of mirex1 leached from mirex fire and bait (84.7% corn cob grits, 15.0%
soybean oil, and 0.3% mirex) by fresh water. Mirex is a chlorinated hydrocarbon
insecticide applied in bait form (1.4 kg per ha) to control the imported fire ant,
Solenopsis richteri Forel, in the southeastern United States.
Field studies have shown translocation of mirex from treated land and high marsh
areas to estuarine biota (Borthwick et al. 1973). Possible routes of entry into the
estuarine environment include, but may not be limited to, biological transport, tidal
action, or fresh water runoff containing the bait or mirex leached from bait.
Low concentrations of mirex (ng/L or ppt range) have been detected in natural
waters. In three cases during periods of heavy runoff, a residue of 0.03 /ng/L (part
'Dodecachlorooctahydro-l.SA-metheno-ltf cyclobuta [cd] pentalene.
Archives of Environmental Contamination 371
and Toxicology Vol. 3, 371-383 (1975) 145
©1975 by Springer-Verlag New York Inc.
-------
372 M. E. Tagatz et al.
per billion) was found in samples of water from various streams of Mississippi after
application of mirex bait in the watershed (Alley, personal communication2).
In almost all studies on the effects of mirex on non-target aquatic organisms,
experimental animals were exposed directly to the bait or to the technical compound
dissolved in a water-miscible solvent (Butler 1963, Muncy and Oliver 1963, Van
Valin et al. 1968, McKenzie 1970, Lowe et al. 1971, Ludke et al. 1971, Bookhout
et al. 1972, Cooley et al. 1972, Collins et al. 1973, Redmann 1973). However, the
study by Ludke et al. (1971) included exposing crayfish in small aquaria to mirex
leached from bait enclosed in filter paper and screen wire.
Our study considers the possibility that in field applications of mirex, estuarine
organisms may not come into direct contact with the bait, but could be exposed to
mirex leached from the bait and carried from treated areas by rainwater runoff into
estuarine drainage systems. Because solvents (other than the constituent soybean oil)
are not used during application of mirex bait in the field, they were not used in the
experimental design of the present work.
Materials and methods
Four 28-day replicate experiments were conducted, using caged animals in out-
door tanks that received a continuous flow of mirex-laden water from gravity-flow
columns that contained mirex ant bait. Our experimental design (mixing of mirex-
laden tap water from the columns with salt water) was chosen to simulate conditions
in an estuary, where fresh water runoff from the watershed may contain mirex
leached from bait and the mirex could be introduced into the estuary when fresh water
mixes with salt water. The experiments were conducted seasonally: spring (April 25
May 23, 1973), summer (July 10 - August 7, 1973), fall (October 6 - November
13, 1973), and winter (January 15 - February 12, 1974).
Six 2.4-m diameter fiberglass tanks and six 10-cm diameter glass columns were
used (Figure 1). Three treated tanks received filtered (cartridges of one- /apore size)
tap water that had trickle-filtered through columns of mirex bait before being mixed
with unfiltered sea water. Similarly, three control tanks received water from col-
umns containing all components of the bait except mirex. Tap water contained no
detectable chlorine (< 0.1 mg/L). Tap water (0.5 L/min) and natural seawater (1.0
L/min) siphoned from constant-head boxes were mixed in glass troughs, positioned
below the columns, before entering the tanks. The water level in each tank was
maintained at 30.5 cm by a standpipe opposite the site where water entered from the
mixing trough. The tank area was partially enclosed by a fiberglass roof and back
wall for protection from severe weather.
Each column consisted of an outer cylinder containing three 15-cm-high and
9-cm-diameter cylinder inserts (Figure 2). Each insert contained a 50-g layer of bait
'E G. Alley, Mississippi State Chemical Laboratory, State College, Miss. 39762.
146
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Seasonal Effects of Mirex
373
(total of 150 g per column) which was soaked for 24 hr in fresh water to allow
swelling before being placed in the columns. The bottoms of the inserts were
covered by nylon screen (0.84 mm mesh) to retain the bait. A funnel attached to the
bottom of the column concentrated the flow of tap water leaving the column.
Because exploratory tests indicated relatively high concentrations (> 2 /ig/L) of mirex
in column runoff during the first few days, we passed tap water through the columns
(0.5 L/min.) for 4 to 8 days (spring, 8 days; summer, 4 days; fall and winter, 6
days) before the tanks were filled. Tanks were filled a day prior to the start of
animal exposure. The amount of bait used in the column and the water flow were
chosen because they produced low but detectable residues of mirex in tank water
(desired range, 0.01-0.50 /ig/L).
Temperatures (12:00 Noon, mercury thermometer) and salinities (8:00 AM,
temperature-compensated refractometer) of the tank water were measured at the start
of each test and four times weekly thereafter. Temperature and salinity changes were
gradual. Temperature responded to natural changes in air and sea water tempera-
tures; salinity, to natural fluctuations in salinity of the sea water source (Santa Rosa
Sound, Florida).
Fig. 1. Experimental system showing outdoor tanks and gravity-flow columns.
147
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374
M. E. Tagatz el al.
Caged animals were placed on two cm of beach sand covering the bottom of the
tanks. Each tank held four cages that contained, respectively, 25 adult sheepshead
minnows, Cyprinodon variegatus; 14 juvenile blue crabs, Callinectes sapidus; 25
juvenile pink shrimp, Penaeus duorarum; and 50 (35 in the fall experiment) adult
grass shrimp, Palaemonetes pugio. Fish ranged from 32 to 59 mm total length;
crabs, 21 to 75 mm carapace width; pink shrimp, 42 to 92 mm rostrum-telson
length; and grass shrimp, 22 to 32 mm rostrum-telson in the spring study and 33 to
36 mm in the summer, fall and winter studies. All animals were captured in local
waters and acclimated for 3 to 16 hr to the initial salinity and temperature of each
experiment by gradual addition of tap water to the salt water stock-aquaria.
Cages for crabs (115.5 cm x 33.0 cm x 23.0 cm deep) were made of stainless-
steel screening (7.1 mm mesh) over wooden frames, and cages for fish and shrimps
(76.0 cm x 76.0 cm x 30.5 cm deep) were made of nylon screening (3.3 mm
mesh) over wooden frames. To prevent cannibalism, crabs were confined in indi-
Freshwater siphon (301/hr)
Constant-head box
(Filtered freshwater)
Constant-head box
(Unfiltered seawater)
Seawater siphon (601/hr)
Nylon screen
(Retains mirex bait)
Flow into tank
(901/hr)
\
Mixing area
Fig. 2. Gravity-flow column
148
-------
Seasonal Effects of Mirex 375
vidual compartments (16.5 cm x 16.5 cm x 23.0 cm deep). Pink shrimp were
provided 7.5 cm of sand on the cage bottom for burrowing.
Crabs were fed small fish, and pink shrimp were fed cubes of fish meat, weekly.
Fish and grass shrimp were not given food but could consume plankton.
To determine growth of crabs, carapace width (in mm between the tips of the
lateral spines) was measured at the beginning, middle and end of the experiment.
Percentage survival of animals was determined at the end of the experiment. The
chi-square test was used to determine significant differences in numbers of dead and
living animals between treated and control tanks.
Samples of water, sand, or animals were analyzed by electron-capture gas
chromatography to determine mirex concentrations. Sensitivity was 0.01 /J-g/g (ppm)
for whole animals (wet-weight basis) and for sediment samples (air-dried weight
basis), and 0.01 /ng/L (ppb) for water samples. Samples to which known amounts of
mirex were added gave recovery rates greater than 85%, but concentrations were not
corrected for percentage recovery. Techniques for most residue analyses were those
of Lowe et al. (1971) and for tissue samples that weighed less than five g, those of
Hansen et al. (in press). Samples of water from each tank were obtained at the start
of the experiment (when the animals were placed in the tank) and twice a week
thereafter. A water sample consisted of a composite collection from two sites in
each tank. At 14 and 28 days, a composite sample of sand from four sites was
obtained from each tank. Surviving animals (composite samples) were analyzed for
mirex at the end of the experiment, and dead animals (individual or composite
samples) were analyzed as deaths occurred. No residues of mirex were detected in
pretest samples of sand and animals or in fish used as food.
Samples of sheepshead minnows, blue crabs, or pink shrimp surviving at the end
of the experiments, and samples of dead pink shrimp, were taken for histological
examination. Surviving pink shrimp were sampled in all experiments; fish and crab,
in all experiments except winter. Samples of survivors consisted of 10 or 15
individuals of a species from treated tanks and 10 or 15 individuals from control
tanks. Samples of dead pink shrimp consisted of seven treated shrimp in fall and
nine treated and three control shrimp in winter.
Results and discussion
Measured concentrations of mirex (/*g/L) in the three treated tanks averaged
approximately 0.04 in spring, 0.12 in summer, 0.06 in fall, and 0.09 in winter
(Table I). Analysis of variance showed no significant differences ( °c = 0.05)
among the three treated tanks in each test. Also, mirex residues did not increase or
decrease statistically ( °c = 0.05) with time within individual tanks.
Seasonal temperatures and salinities of tank water and average residues of mirex
are summarized in Table II. Temperatures averaged 23.1°C in spring, 29.8°C in
149
-------
-J
ON
Table I. Concentrations of Mirex (vg/L) in Treated Tank Water During Four 28-Day Seasonal Experiments.
Cn
o
Time
elapsed
Start
Week 1
Week 2
Week 3
Week 4
Tank 1
<0.01a
0.02
0.02
0.02
0.01
0.07
0.03
0.03
0.09
Spring
Tank 2
0.03
0.02
0.03
0.03
< O.Ola
0.07
0.06
0.04
0.11
Tank 3
0.02
0.12
0.02
0.04
0.01
0.05
0.02
0.03
0.08
Tank 1
0.36
0.14
0.05
0.04
0.04
0.03
0.05
0.05
0.09
Summer
Tank 2
0.17
0.08
0.04
0.04
0.08
0.09
0.10
0.10
0.23
Tank 3
0.52
0.20
0.06
0.05
0.07
0.10
0.19
0.11
0.11
Tank 1
0.20
0.14
0.07
0.03
0.04
0.05
0.04
0.04
0.05
Fall
Tank 2
0.23
0.09
0.04
0.03
0.03
0.03
0.04
0.03
0.04
Tank 3
0.12
0.07
0.04
0.01
0.02
0.02
0.02
0.02
0.02
Tank 1
0.25
0.11
0.07
0.03
0.05
0.05
0.03
0.04
0.05
Winter
Tank 2
0.24
0.09
0.05
0.03
0.02
0.04
0.03
0.05
0.04
Tank 3
0.52
0.12
0.04
0.04
0.08
0.05
0.03
0.03
0.04
2
m
H
(TO
N
0
Average 0.03 0.04 0.04
Range < 0.01-< 0.01- 0.01-
0.09 0.11 0.12
0.09 0.10 0.16
0.03- 0.04- 0.05-
•0.36 0.23 0.52
0.07 0.06 0.04
0.03- 0.03- 0.01-
0.20 0.23 0.12
0.08 0.07 0.11
0.03- 0.02- 0.03-
0.25 0.24 0.52
a0.005 used for calculating average.
-------
Seasonal Effects of Mirex 377
summer, 23.4°C in fall, and 19.TC in winter. Overall salinity range was from 7 to
19 parts per thousand.
In mirex-contaminated tanks, survival of crabs and shrimps usually was signific-
antly reduced; survival of fish was not affected (Table III). The number of deaths
was greatest in the summer, followed by fall, winter and spring. Observed deaths
among exposed shrimps occurred after 2 to 15 days in summer (all pink shrimp died
in 15 days), after 8 to 28 days in fall, and after 17 to 28 days in winter. During
summer and fall, significantly more exposed blue crabs died than did control crabs,
deaths oqcurring after 6 to 28 days of exposure. At the end of the summer experi-
ment, one-third of the crabs surviving in treated tanks were paralyzed or had lost
equilibrium. Most deaths occurred among the smaller crabs. Of 12 deaths among 42
crabs exposed in summer, 11 were among 19 crabs 28 to 40 mm carapace width, but
only one died among 23 crabs 41 to 64 mm. Survival of sheepshead minnows was
unaffected in all experiments. This species can spawn in the presence of mirex since
each tank contained 50 to 100 young sheepshead minnows (up to 14 mm total
length) at the end of the spring studies and 75 to 200 young (up to 26 mm) at the end
of the summer studies. Lowe et al. (1971) also found that fish were relatively
unaffected, juvenile pinfish (Lagodon rhomboides) living for five months on a diet
that contained approximately 20 /ig/g of technical mirex.
Mortality and delayed toxicity to crabs and shrimps exposed to mirex (technical
grade) occurred in other studies. Lowe et al. (1971) found that a significant number
of juvenile Penaeus duorarum died during a seven-day exposure to 1.0 /x.g/L of
mirex in sea water averaging 17°C, but few died during a 21-day exposure to 0.1
/ig/L in sea water averaging 14°C. All juvenile Callinectes sapidus died within three
weeks after a 96-hr exposure to 0.1 mg/L (Lowe et al. 1970). Redmann (1973)
reported a 40% mortality of Palaemonetes pugio in 12 days from a 48-hr .exposure to
0.01 /jig/g of mirex in sea water at 20°C. McKenzie (1970) and Lowe et al. (1971)
demonstrated that smaller blue crabs were more sensitive to mirex bait than were
larger crabs.
Table. II. Temperature, Salinity and Average Concentration of Mirex in
Tank Water During Four 28-Day Seasonal Experiments
Temperature (°C)
Spring
Summer
Fall
Winter
Av.
23.1
29.8
23.4
19.1
Range
19.3-25.2
28.0-30.8
17.0-27.0
13.8-22.9
Salinity
(Parts per thousand)
Av.
13.2
15.7
17.5
13.6
Range
10-18
14-18
15-19
7-16
Mirex
M8/L
0.04
0.12
0.06
0.09
151
-------
378
M. E. Tagatz et al.
The greater mortality observed in summer could be caused by increased leaching
and by increased toxicity of mirex in warmer water. Analysis of variance showed
significant differences ( <* = 0.05) in mirex residues in water between summer and
spring and between summer and fall; other paired treatments were not significant.
Most deaths of affected animals occurred in summer at the highest water temperature
and concentration of mirex; fewest in spring at next to the lowest temperature and at
the lowest concentration. McKenzie (1970) reported that survival time and rate in
juvenile blue crabs exposed to mirex bait decreased as temperatures increased from
20 to 27°C.
Mirex had no marked affect on growth of juvenile blue crabs exposed in these
studies (Table IV). The greatest difference in mean percentage increase in size per
Table III. Percentage Survival of Animals Exposed to Mirex and Chi-Square Values
After Each of Four 28-Day Seasonal Experiments. Data on Treated Tanks and on
Control Tanks were Combined
Percentage survival and chi-square
Animals
Number Spring Summer
Fall
Winter
Sheepshead minnows
Control
Treated
Chi-square
Blue crabs
Control
Treated
Chi-square
Pink shrimp
Control
Treated
Chi-square
Grass shrimp
Control
Treated
Chi-square
75 63
75 55
N.S.a
42 100
42 98
N.S.
75 86
75 76
N.S.
150b 95
150b 86
-c
96
99
N.S.
98
71
11.01**
87
0
114.70**
79
10
145.87**
91
92
N.S.
95
81
4.09*
91
19
78.44**
96
59
41.69**
99
100
N.S.
95
95
N.S.
91
49
30.51**
97
89
6.20*
aN.S. = non-significant, ** significant at 1% level (X2, l d.f. = 6.63), * significant at 5%
level (X2, 1 d.f. = 3.84).
b 105 used in the fall experiment.
cSome small grass shrimp escaped through the mesh of the cages, and all those free in the
tanks may not have been captured (values in table include those captured).
152
-------
Seasonal Effects of Mirex
379
molt between treated and control groups was less than 3%; not decisive in a species
normally characterized by variable growth among individuals (Tagatz 1968). Taking
into account deaths that occurred among treated crabs, frequency of molting did not
differ greatly between treated and control groups.
Animals exposed to leached mirex concentrated the chemical in body tissues
(Table V). During 28 days' exposure, depending on the experiment, surviving
sheepshead minnows accumulated from 10,500 to 40,800X the average concentra-
tion of mirex in the test water. Residues in young fish (<27 mm long) hatched in the
treated tanks in summer ranged from 1.2 to 2.7 /u,g/g, compared to a range of 1.2 to
4.9 Atg/g in adults. Maximum concentration factors in other animals surviving at the
end of the four seasonal tests ranged from 900 to 2,300X for blue crabs, from 3,000
to 10,OOOX for pink shrimp, and from 8,000 to 10,800X for grass shrimp. Lowe et
al. (1971) reported that a sample of live pink shrimp exposed to 0.1 /Ag/L of
technical mirex for three weeks contained 0.26 /u,g/g, a value within the range 0.09
to 0.40 /ng/g observed in our study. Residues of mirex (/ig/g) in various categories
of living animals sampled from estuaries near treated areas in South Carolina
(Borthwick et al. 1973) compared to those obtained in our study (in parentheses)
are: crabs, 0-0.60 (0.02-0.17); shrimps, 0-1.3 (0.09-1.30); and fish, 0-0.82 (0.35-
4.9).
Residues of mirex in dead animals generally were higher than those in living
animals, but the ranges of each group usually overlapped. Animals concentrated
mirex rapidly. For example, mirex in dead grass shrimp from one tank after two
Table IV. Growth of Blue Crabs in Control and Mirex-Treated Tanks During
Each of Four 28-Day Seasonal Experiments. Total of 42 Crabs in Control or
Treated Tanks Per Season
Season
Spring
Summer
Fall
Winter
Initial width
(mm)
Ave. (Range)
43.1
(25-75)
43.3
(25-63)
36.0
(25-46)
31.0
(21-44)
Control
No. of
molts
32
44
42
18
Treated
Mean %
increase
per molt
22.1
17.2
21.2
15.5
Initial width
(mm)
Av. (Range)
44.2
(25-72)
43.0
(28-64)
36.7
(24-46)
31.4
(23-45)
No. of
molts
37
33
38
14
Mean %
increase
per molt
22.4
14.4
20.3
18.1
153
-------
oo
o
Cn
Table V. Ranges of Mirex Residues (vg/g) in Living Animals After 28 Days Exposure (Composite Sample From Each of Three
Treated Tanks) and in Dead Animals (Individual or Composite Samples From One or More Tanks)
Animal
Sheepshead minnow
Blue crab
Pink shrimp
Grass shrimp
Spring
Living Dead
0.35-0.42
0.02-0.05
0.09-0.40
0.09-0.32
Summer
Living Dead
1.2-4.9
0.04-0.11 0.10-0.59
0.30-1.00
0.62-1.30 0.66-2.40
Fall
Living
0.94a
0.03-0.14
0.12-0.20
0.27-0.50
Dead
-
0.03-0.17
0.14-0.50
0.75-0.76
Winter
Living Dead
0.93-1.30
0.13-0.17
0.16-0.27 0.15-0.49
0.64-0.74 0.73-0.76
S
tn
H
OP
N
£.
a Residues in fish determined for only one tank.
-------
Seasonal Effects of Mirex
381
days of exposure was 2.4 £ig/g, about 7,OOOX the two-day average concentration of
mirex in the water.
Sand substrata adsorbed mirex from the test water. (Table VI). The greatest
concentration of mirex in sand relative to that in test water after 14 days was 220X
(0.028 jig/g in sand and 0.126 jttg/L in water); after 28 days, 770X (0.023 /ig/g in
sand and 0.03 ;u,g/L in water). No residues of mirex were detected in samples of
water, sand, or animals from control tanks.
Pathologist John A. Couch, of this laboratory, found gill alteration in some
exposed sheepshead minnows, and natural pathogens in some blue crabs and pink
shrimp from treated and control groups. Among exposed and surviving sheepshead
minnows, four of ten fish in a spring sample, none of ten fish in a summer sample,
and three of 15 fish in a fall sample showed slight gill lamellar endema; gills of
control fish were not altered. Other tissues (liver, pancreas, intestine and kidney) in
treated and control fish were normal. A parasitic dinoflagellate, Hematodinium sp.,
was present in samples of blue crabs surviving the fall experiment (in two of ten
treated crabs and in one of ten control crabs). It was not observed in spring and
summer samples. The parasite invades hemolymph sinuses and replaces hemocytes.
According to Dr. Couch, this is the first report of this dinoflagellate in crabs from
the Gulf of Mexico. A nuclear polyhedrosis virus of shrimp (Couch 1974) was
present in samples of pink shrimp surviving the fall experiment (in eight of ten
treated shrimp and in one of 15 control shrimp). It was also present in one of five
dead exposed shrimp examined. None of the surviving pink shrimp in samples from
treated and control groups from the other seasonal experiments were infected with
this virus. It was not present in samples of dead pink shrimp (nine treated and three
control shrimp) examined in the winter study. We do not know if the pathogens that
occurred in some crabs and shrimp interacted with mirex in producing mortality.
Table VI. Mirex Residues in Water and in Sand Substrata During Four 28-Day
Seasonal Experiments
Season
Spring
Summer
Fall
Winter
Av. concn.
in water
(M8/L)
0.04
0.12
0.06
0.09
Concn. in sand
(Mg/g) at 14 days
Tank 1
<0.01
0.028
<0.01
ND
Tank 2
NDa
0.017
<0.01
ND
Tank 3
<0.01
0.018
ND
ND
Concn. in sand
(Mg/g) at 28 days
Tankl
0.023
0.025
ND
0.022
Tank 2
0.019
0.038
ND
<0.01
Tank 3
0.021
0.016
ND
0.016
= non detectable, < 0.01
155
-------
38o M. E. Tagatz et al.
However, because deaths of both crabs and shrimp occurred in treated groups which
showed neither of these pathogens, mirex alone could be responsible for mortality.
The experimental system illustrates that mirex can be leached from bait by
freshwater, concentrated by estuarine organisms, and can be toxic to crabs and
shrimps.
Acknowledgment
The authors wish to express their appreciation to Dr. John A. Couch for histolog-
ical examination of samples of animals.
References
Bookhout, C. G., A. J. Wilson, Jr., T. W Duke, and J. I. Lowe: Effects of mirex
on the larval development of two crabs. Water, Air, Soil Pollut. 1, 165 (1972)
Borthwick, P W., T. W Duke, A. J. Wilson, Jr., J. I. Lowe, J. M. Patrick, Jr.,
and J. C. Oberheu: Accumulation and movement of mirex in selected estuaries
of South Carolina, 1969-71. Pestic. Monit. J. 7, 6 (1973).
Butler, P A.: Commercial Fisheries Investigations. In, Pesticide-Wildlife Studies:
A Review of Fish and Wildlife Service Investigations during 1961 and 1962.
U. S. Fish Wildl. Serv. Circ. 167, 11 (1963).
Collins, H. L., J. R. Davis, and G. P Markin: Residues of mirex in channel catfish
and other aquatic organisms. Bull. Environ. Contam. Toxicol. 10, 73 (1973).
Cooley, N. R., J. M. Keltner, Jr., and J. Forester: Mirex and Aroclor® 1254: Effect
on and accumulation by Tetrahymena pyriformis Strain W. J. Protozool. 19,
636 (1972).
Couch, J. A.: Free and occluded virus, similar to Baculovirus, in hepatopancreas of
pink shrimp. Nature 247, 229 (1974).
Hansen, D. J., P R. Parrish, and J. Forester: Aroclor® 1016: Toxicity to and uptake
by estuarine animals. Environ. Res. (In press).
Lowe, J. I., P. R. Parish, A. J. Wilson, Jr., P D. Wilson, and T. W. Duke: Effects
of mirex on selected estuarine organisms. Trans. 36th N. Am. Wildl. Nat.
Resour. Conf., p. 171 (1971).
Lowe, J. I., P D. Wilson, and R. B. Davison: Effects of mirex on crabs, shrimp,
and fish. In, Progress Report of the Bureau of Commerical Fisheries, Center
for Estuarine and Menhaden Research, Pesticide Field Station, Gulf Breeze,
Fla., 1969. U. S. Fish Wildl. Serv. Circ. 335, 22 (1970).
Ludke, J. L., M. T. Finley, and C. Lusk: Toxicity of mirex to crayfish, Procam-
barus blandingi. Bull. Environ. Contam. Toxicol. 6, 89 (1971).
156
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Seasonal Effects of Mirex 383
McKenzie, M. D.: Fluctuations in abundance of blue crab and factors affecting
mortalities. S. C. Wildl. Resour. Dept., Mar. Resour. Div., Tech. Rep. No.
1, 45p (1970).
Muncy, R. J., and A. D. Oliver, Jr.: Toxicity often insecticides to the red crawfish,
Procambarus clarki (Girard). Trans. Am. Fish. Soc. 92, 428 (1963).
Redmann, G.: Studies on the toxicity of mirex to the estuarine grass shrimp,
Palaemonetes pugio. Gulf Res. Rep. 4, 272 (1973).
Tagatz, M. E.: Growth of juvenile blue crabs, Callinectes sapidus Rathbun, in the
St. Johns River,Florida. U. S. Fish Wildl. Serv. Fish. Bull. 67, 281 (1968).
Van Valin, C. C., A. K. Andrews, and L. L. Eller: Some effects of mirex on two
warm-water fishes. Trans. Am. Fish. Soc. 97, 185 (1968).
Manuscript received August 8, 1974; accepted October 22, 1974.
157
-------
Reprinted from Journal of the
Fisheries Research Board of
Canada, Vol. 32(2): 314-316,
1975, with permission of the
Ministry of Supply of Canada
AN AUTOMATIC BRINE SHRIMP FEEDER FOR AQUATIC BIOASSAYS
Steven C. Schimmel and David J. Hansen
Contribution No. 224
159
-------
An Automatic Brine Shrimp Feeder for Aquatic Bioassays1
STEVEN C. SCHIMMEL AND DAVID J. HANSEN
U.S. Environmental Protection Agency, Gulf Breeze Environmental Research Laboratory
Sabine Island, Gulf Breeze, Fla. 32561, USA
SCHIMMEL, S. C., AND D. J. HANSEN. 1975. An automatic brine shrimp feeder for aquatic
bioassays. J. Fish. Res. Board Can. 32: 314-316.
An electrically operated brine shrimp feeder is described. The device may be set to cycle
1-12 times each day for tests in fish and invertebrate culture and bioassay. Major advantages of
the feeder are that it is readily adapted to flow-through bioassay and culture apparatuses which
require that equal quantities of food be delivered to animals in two or more test aquaria and that
the number of feedings be recorded. The components, all readily available, cost approximately
$190.
SCHIMMEL, S. C., AND D. J. HANSEN. 1975. An automatic brine shrimp feeder for aquatic
bioassays. J. Fish. Res. Board Can. 32: 314-316.
Les auteurs decrivent un distributeur d'artemies actionne a 1'electricite. L'appareil peut etre
regie selon un cycle de 1-12 fois par jour et servir a nourrir des poissons et des invertebres lors
d'essais de culture ou d'analyse biologique. II a comme avantages principaux d'etre facilement
adaptable a des appareils d'analyse biologique ou de culture traverses par un debit d'eau, de
pouvoir donner aux animaux maintenus dans deux ou plusieurs aquariums une meme quantite de
nourriture et d'enregistrer le nombre de repas. Les composantes, toutesfaciles atrouver, coutent
environ $190.
Received September 3, 1974 Recu le 3 septembre 1974
Accepted November 13,1974 Accepte le 13 novembre 1974
IN this report, we describe an automatic brine or bioassay tanks and records the number of
shrimp feeder that is more useful than others feedings.
(Benoitet al. 1969; Anderson and Smith 1971) in The feeder (Fig. 1) consists of a compart-
that it delivers equal quantities of food to culture mented, all-glass container, two 115-V Gems®
'Contribution No. 224 Gulf Breeze Environmental ©Registered trademarks, Gems Sensors Div. Farm-
Research Laboratory. ington, Conn.; Intermatic timer, Model T185, Inter-
national Register Co., Chicago, 111. Mention of corn-
Printed in Canada (J 3548) mercial products does not constitute endorsement by
Imprime au Canada (J3548) the Environmental Protection Agency.
161
-------
NOTES
315
FIG. 1. Automatic brine shrimp feeder. A. a, electric timer; b, electric
counter; c, solenoid valve; d, vacuum venturi; e and f, float switches;
g, vacuum manifold; h, oscillating pump; i, float switch compartment;
j, siphon tubes to aquaria. B. Timer schematic. L-l and N, circuit to float
switch f and to oscillating pump (h); L-2 and N, circuit to float switch e,
solenoid valve (c), and counter (b).
float switch systems, an oscillating pump, a stain-
less steel solenoid valve, and an electrical counter
and Intermatic® timing mechanism. The glass
container, 32 X 14 X 4 cm, is divided into seven
compartments, six of which hold 100 ml each.
The six compartments provide equal numbers of
nauplii to culture or bioassay aquaria (compart-
ment i may also be used if carefully calibrated).
Although we selected the compartment number
and size to fit our needs, the dimensions may be
varied to suit other requirements.
The brine shrimp are incubated and hatched
in aerated, 2-liter separatory funnels apart from
the feeder (24—48 h at room temperature). The
nauplii are separated from the eggs by shutting of
air supply and allowing differential settling.
Nauplii are placed in a reservoir (approximately
20-30 liters of 1% salt solution, depending on
volume and rate of feedings). The reservoir (not
illustrated) is also aerated to provide oxygen and
disperse the animals uniformly throughout the
seawater medium. The nauplii remain in the
reservoir until pumped out at intervals dictated
by settings on the timer. In tests, nauplii remained
in the reservoir for up to 4 days without appre-
ciable mortality or bacterial contamination. Using
the feeder and reservoir volumes we describe, the
reservoir should be refilled every 48 h. Population
density of brine shrimp may be calibrated by a
photometric device of Sedgwick-Rafter counting
cell to ensure uniform density when refilling the
reservoir.
The operating sequence is: the timer mecha-
nism (a), through one of twelve possible "trip-
pers" that cover a 24-h period, energizes the
oscillating pump (h) and float switch f through
electrical connection L-l (Fig. IB). The pump
fills each compartment with nauplii and water
from the reservoir; the mixture finally cascades
into the largest compartment (i), raising float
switches e and f. On rising, float switch f opens,
deactivating the oscillating pump, while float
switch e closes. The closing of switch e does not
cause the feeder to empty until the timer's tripper
mechanism releases (duration can be adjusted
from 5 to 60 min). When the tripper mechanism
releases, circuit L-2 is energized and activates
the solenoid valve (c) and counter (b) through
float switch e. The purpose of the counter is to
number the feedings, since the device is updated
each cycle. The solenoid valve now opens, allow-
ing water in compartment i to flow, creating a
vacuum at the venturi d. The vacuum, applied
along the entire manifold (g), starts the siphons
in the small compartments. The nauplii in these
compartments are channelled to culture or bio-
assay aquaria through siphon tubes (j). As the
largest compartment (i) empties, float switch e
162
-------
316
J. FISH. RES. BOARD CAN., VOL. 32(2), 1975
opens and switch f closes. The system is now
ready to recycle when the timing mechanism
trips again.
Our feeder has been used (Schimmel et al.
1974) in several 30-day bioassays utilizing a
modified Mount-Brungs diluter. Since the total
daily volume from the feeder was 600 ml and
the daily diluter volume was 40 liters, concentra-
tion of toxicant in these bioassays was not appre-
ciably affected (note: feeder siphons empty di-
rectly into diluter delivery tubes but the connec-
tions must not be airtight as the two devices cycle
independently). During our bioassays no mechani-
cal or electrical difficulties occurred. If freshwater
cultures or bioassay experiments are attempted,
salt content of the reservoir should not be a
problem, provided a sufficient ratio of freshwater
volume to feeding volume is used. Reservoir
water of 1% salt solution will easily support brine
shrimp cultures.
Cost of the individual components of our feeder
is approximately $190, and the items described
here are available in most electronic catalogs.
Acknowledgments — We thank Mr S. Foss for
preparing Fig. 1.
ANDERSON, E. EX., AND L. L. SMITH, JR. 1971. An auto-
matic brine shrimp feeder. Prog. Fish-Cult. 33:
118-119.
BENOIT, D., R. SYRETT, AND J. HALE. 1969. Automatic
live brine shrimp feeder. Trans. Am. Fish. Soc. 98:
532-533.
SCHIMMEL, S. C., P. R. PARRISH, D. J. HANSEN, J. M.
PATRICK, JR. AND J. FORESTER. 1975. Endrin: effects
on several estuarine organisms. Proc. Assoc. South-
east. Game Fish Comm. (In press).
163
-------
Reprinted from Proceedings of
the 28th Annual Conf. of
Southeastern Assoc. of Game
and Fish Comm., Nov 17-20,
1974, pp. 179-186, with per-
mission of the Southeastern
Assoc. of Game and Fish
Comm.
HEXACHLOROBENZENE: EFFECTS ON SEVERAL ESTUARINE ANIMALS
Patrick R. Parrish, Gary H. Cook, and James M. Patrick, Jr.
Contribution No. 226
165
-------
Reprinted from the Proceedings of the 28th Annual Conference of the Southeastern Association of
Came and Fish Commissioners, 1974.
HEXACHLOROBENZENE:
EFFECTS ON SEVERAL ESTUARINE ANIMALS1
by
Patrick R. Parrish2, Gary H. Cook
James M. Patrick, Jr.
U. S. Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Sabine Island, Gulf Breeze, Florida 32561
(Associate Laboratory of the National Environmental
Research Center, Corvallis, Oregon)
ABSTRACT
Tests were conducted to determine (1) the acute (96-hour) toxicity of hexachlorobenzene (HCB) to pink shrimp (Penaeus
duorarum), grass shrimp (Palaemoneies pugio), sheepshead minnows (Cyprinodon variegalus) and pinfish (Lagodon rhom-
boides) and (2) the rate of HCB uptake and depuration by pinfish. Hexachlorobenzene was not acutely toxic to any of the animals
tested at measured concentrations in sea water to 25 ug/1. However, both species of shrimps in the highest HCB concentration
were lethargic as compared to controls and exhibited an uncharacteristically white hepatopancreas at the end of the 96-hour ex-
posure. Pinfish exposed to average measured HCB concentrations of 0.06,0.15,0.65,1.87, or 5.2 ug/1 for 42 days accumulated the
compound throughout the exposure. Maximum residue in muscle (wet-weight) was 34.000X the measured concentration in test
water. Pinfish retained most (>50%) of the HCB after a 28-day depuration period in HCB-free water.
'Contribution No. 226, Gulf Breeze Environmental Research Laboratory.
^Present address: Bionomics - EG&G, Inc., Marine Research Laboratory, Route 6, Box 1002, Pensacola, Florida 32507.
179
167
-------
INTRODUCTION
We initiated research into the effects of hexachlorobenzene, a fungicide and in-
dustrial chemical, in mid-1973 at the request of Dr. John Buckley, U. S. Environmental
Protection Agency, Office of Program Integration. High residues of hex-
achlorobenzene (HCB) had been detected in cattle in southern Louisiana, and it was
feared that the compound was present in the highly productive estuaries of this Gulf
state. Thus, we began to study its acute effects on four estuarine animals and to
determine the rate of HCB uptake and depuration in an estuarine fish.
METHODS AND MATERIALS
Test animals
All animals except pink shrimp were collected near the Gulf Breeze Laboratory and
acclimated to laboratory conditions for at least ten days before exposure. Pink shrimp
(Penaeus duorarum), purchased from a local bait dealer, were acclimated similarly. If
mortality in a specific lot of animals exceeded 1% in the 48 hours immediately preced-
ing the test or if abnormal behavior was observed during acclimation, those animals
were discarded. Pink shrimp tested were from 48 to 69 mm rostrum-telson length; grass
shrimp (Palaemonetes pugio), 22 to 33 mm rostrum-telson length; sheepshead min-
nows (Cyprinodon variegatus), 18 to 35 mm standard length; and pinfish (Lagodon
rhomboides), 52 to 83 mm standard length. Animals were not fed during acute toxicity
tests, but they could obtain food (plankton and other particulate matter) from the un-
filtered sea water in which they were maintained. In the uptake and depuration study,
pinfish were fed commercial fish food that contained no pesticide or polychlorinated
biphenyl contaminants detectable by gas-liquid chromatographic (GLC) analysis.
Test conditions
Acute toxicity of HCB was determined by exposing twenty animals for 96 hours to
different concentrations in 68-1 glass aquaria. Technical grade HCB (99.5% active
ingredient) was dissolved in reagent grade acetone and metered by pump (Lowe et al.,
1972) at 60 ml/hr into unfiltered sea water which entered each aquarium at 150 1/hr.
Control aquaria received the same quantities of water and solvent, but no HCB.
The rate of uptake and depuration of HCB by pinfish was determined by exposing 40
animals to 0.1, 0.32, 1.0, 3.2, or 10 ug/1 for 42 days, then placing them in HCB-free
water for 28 days. Test conditions were the same as for acute tests. Five fish were sam-
pled from each concentration and control after 3, 7, 14, 28, and 42 days exposure and
after 14 and 28 days depuration. HCB residue in pooled samples of edible flesh (all
muscle above lateral line on left side of fish and overlying skin without scales), liver and
remainder of fish was determined.
Chemical Analyses
Water samples. The approximate dilution volume for GLC analysis was
determined from the nominal concentration, and each sample was spiked with o,p'-
DDE as an internal standard at approximately the same concentration as that ex-
pected for HCB. One-liter samples were extracted twice with 100 ml of methylene
chloride and the methylene chloride drained through sodium sulfate pre-washed with
methylene chloride into a Kuderna-Danish(K-D)'evaporator. The methylene chloride
was concentrated on a steam-bath to approximately 3 ml, 50 ml of petroleum ether was
added and the extract again concentrated to approximately 1 ml to remove any
remaining methylene chloride. The 1-ml extract was transferred to a7-mmChromoflex
chromatographic column containing 1.6 g of Florisil activated at 130°C for 24 hours,
then diluted with 20 ml of 1% ethyl ether in hexane. Each sample was then concentrated
or diluted, as needed, to the pre-determined dilution volume for GLC analysis
Recovery ot o,p'-DDE was calculated for each sample. When less than 70% of the o,p'-
UUE was recovered, the sample was re-analyzed if possible. HCB concentrations were
not corrected for recoveries.
180
168
-------
Tissue samples. Tissue samples weighing less than 5 g were weighed into a Kontes
tissue grinder and spiked with o,p'-DDE at approximately the same concentration as
expected for HCB. The tissue was extracted 3 times by grinding with 5 ml of
acetonitrile, centrifuging, and decanting into a 150 mm screw-capped test tube. The
acetonitrile extract was then flooded with 75 ml of 2% sodium sulfate and extracted 3
times with 5 ml of hexane. The hexane extracts were concentrated and cleaned up in the
same manner as were the water samples.
Other tissue samples (5-30g) were blended with sodium sulfate, spiked with o,p'-
DDE at approximately the same concentration as that expected for HCB and ex-
tracted in a Soxhlet apparatus with petroleum ether for 4 hours. The extract was ab-
sorbed onto 10 cm of unactivated Florisil contained in a 400-mm X 20-mm chromato-
graphic column, then eluted with 20% methylene chloride in hexane. The hexane
was concentrated, adsorbed onto 10 cm of activated Florisil contained in a 400-mm
X 20-mm chromatographic column, and eluted with 6% ethyl ether in petroleum
ether.
GLC analyses were performed on a Varian model 1400 gas chromatograph equipped
with an electron capture detector and a 185-cm X 2-mm ID glass column packed with
2% OV-101 on 80/100 mesh GAS CHROM Q. Operating parameters were: Nitrogen
carrier gas 25 ml/min, detector temperature 210°C, injector temperature 250°C, and
column temperature 160°C. Confirmation was performed on a!83-cm X 2-mm glass
column packed with 5% OV-210 on 80/100 mesh GAS CHROM Q.
RESULTS AND DISCUSSION
Chemical analyses
Hexachlorobenzene, because of its nonpolar character, is "insoluble in water"
(Frear, 1969) and practically insoluble in polar organic solvents, suc,h as acetone. We
found that after filtering a saturated solution of HCB in acetone through a #1 What-
man filter, only 2.9 g/1 of HCB remained in solution. The insolubility of HCB is
reflected in the measured concentrations in test waters, which ranged from 26% to 96%
of nominal concentrations (Tables 1 and 2).
Table 1.
Acute (96-hour) toxicity of hexachlorobenzene to and uptake by pink
shrimp (Penaeus duorarum), grass shrimp (Palaemonetes pugio), sheeps-
head minnows (Cyprinodon variegatus) and pinfish (Lagodon rhomboi-
des). Whole-body residues are from animals alive at the end of the expo-
sure.
PINK SHRIMPa 11-14 Sep 1973
Water concentrations (ug/1) Mortality
Nominal Measured (%)
Control
0.1
1.0
3.2
10.0
32.0
NDb
0.08
0.87
2.3
7.0
25.0
0
0
0
0
13
33
a!5 shrimp per test container.
bNot detectable; 0.01 ug/1.
Temperature (°C)
Salinity (o/oo)
Minimum
29.0
24.0
Residue
(ug/g)
0.009
0.27
1.7
4.1
13.0
21.0
Concentration factor
Nominal Measured
30.0
26.5
2700
1700
1281
1300
656
3375
1954
1783
1857
840
Maximum Mean
29.8
25.0
181
169
-------
GRASS SHRIMP 13-17 Aug 1973
Nominal
Control
0.1
1.0
3.2
10.0
32.0
Temperature (°C)
Salinity (o/oo)
SHEEPSHEAD MINNOWS 20-24 Aug 1973
Water concentrations (ug/1) Mortality Residue
Nominal Measured (%) (ug/g)
ons(ug/l) Mortality Residue
easuret
ND
0.096
0.56
1.8
6.1
17.0
1
J (%)
0
0
0
0
0
10
Minimum
30.0
22.0
(ug/g)
0.017
1.1
1.9
4.5
10.0
27.0
Maximum
32.0
26.0
Concentration factor
Nominal
—
11000
1900
1406
1000
844
Mean
31.4
23.6
Measured
—
11458
3393
2500
1639
1588
Concentration factor
Nominal Measured
Control ND 0
0.1 0.072 0
1.0 0.33 0
3.2 1.49 0
10.0 4.06 0
32.0 13.3 0
Minimum
Temperature (°C) 29.5
Salinity (o/oo) 15.0
P1NF1SH 30Jul-2 Aug 1973
ND
0.024 240
0.028 28
0.69 216
15.0 1500
89.0 2781
Maximum Mean
32.0 30.5
28.0 21.6
—
333
85
463
3695
6692
Water concentrations (ug/1) Mortality Residue Concentration factor
Nominal Measured (%)
Control ND 0
1.0 0.3 0
3.2 -a 0
10.0 -a 0
32.0 8.4 0
100.0 -a 0
aNo chemical analyses performed.
Minimum
Temperature (°C) 30.5
Salinity (o/oo) 18.0
(ug/g) Nominal
0.021
6.3 6300
44.0 13750
47.0 4700
79.0 2469
120.0 12000
Maximum Mean
32.0 31 0
25.0 22.6
Measured
21000
—a
—a
9405
—a
182
l'70
-------
Table 2. Nominal and measured hexachlorobenzene concentrations in test water during an uptake and depuration study with pinflsh
(Lagodon rhomboides).
Mean Measured
Concentration
0:06
0.15
0.65
1.87
5.2
aNot detectable; 0.01 ug/1.
bNo chemical analyses performed.
Nominal
Concentration
(ug/1, ppb)
Control
0.1
0.32
1.0
3.2
10.0
Measured Concentration (ug/1, ppb)
26 Sep 2 Oct 10 Oct 16 Oct 24 Oct
NDa
0.05
0.17
0.65
1.73
5.8
ND
0.04
0.16
0.63
1.93
5.4
ND
0.09
0.18
0,77
2.04
3.3
ND
0.06
0.10
0.5JO
1.90
4.4
ND
0.08
0.13
0.67
1.62
8.9
6Nov
ND
-b
-b
0.69
2.00
3.3
-------
A number of methods have been reported for the determination and confirmation of
HCB in various substrates (Taylor and Keenan, 1970; Collins et al., 1972; Smyth, 1972;
Baker, 1973; Holdnnet, 1974 and Johnson et al., 1974). These authors point out that
because of the solubility, volatility, and nonpolar character of HCB and its elution time
on most standard pesticide GLC columns, the analysis and quantitation of HCB at low
concentrations is hampered by the presence of the electron capturing co-extractives,
polychlorinated biphenyls (PCBs) and chlorinated hydrocarbon pesticides. On the
other hand, because of the high percentage of chlorine in the HCB molecule, this com-
pound is very sensitive to electron capture detection.
In this study, HCB measured in both water (1 ug/1 nominal concentration) and tis-
sue samples from animals exposed at this concentration was well above any
background electron capturing compound. At higher concentrations, all interfering
peaks were completely eliminated at the dilution volumes required to bring the HCB
peak on scale. O,p'-DDE was an ideal internal standard. It eluted in the same Florisil
fraction as did HCB and elutedjust after HCB on the GLC column. Recoveries of HCB
and o,p'-DDE in samples spiked with both compounds were greater than 85% for each.
Acute toxicity tests
Hexachlorobenzene, at the concentrations tested, was not acutely toxic to the
animals tested (Table 1). Pink shrimp exposed to a measured concentration of 25 ug/1
experienced the greatest mortality, 33%. Both pink shrimp and grass shrimp exposed
to measured concentrations of 25 ug/1 and 17 ug/1, respectively, were lethargic as com-
pared to controls and exhibited uncharacteristically white hepatopancreases.
Uptake and depuration study
Pinfish exposed to average measured HCB concentrations of 0.06, 0.15, 0.65, 1.87,
or 5.2 ug/1 for 42 days accumulated the compound throughout the exposure (Table 3).
Maximum residue in muscle was 34,OOOX greater than the measured concentration in
test water, maximum residue in liver was 47,OOOX greater, and maximum residue in
remainder of fish was 56,OOOX greater (Table 4). Considering previous experience with
uptake of other organochlorine compounds by estuarine fish at this laboratory
(Hansen et al., 1971; Parish et al., 1974), it is interesting that HCB was accumulated
chiefly in the remainder of the fish, the liver being the usual site of greatest accumu-
lation for the other compounds.
184
172
-------
Table 3. Concentrations of hexachlorobenzene in tissues of pinfish (Lagodon rhomboides) during a 70-day uptake and depuration
study.
Nominal
Water
Concentration
(ug/1, ppb)
Control
0.1
0.32
1.0
3.2
10.0
Control
0.1
0.32
1.0
3.2
10.0
Control
0.1
0.32
1.0
3.2
10.0
aOnly four fish sampled.
bOnly two fish sampled.
LIVER
MUSCLE
REMAINDER
Tissue Concentration (ug/g, ppm)
Days Exposure
4
0.01
0.39
1.9
8.8
22.7
6J.4
J}.02
p.33
1.19
3.3
9.4
38.3
0.03
0.45
1.9
5.6
16.3
52.8
7
0.06
0.98
3.4
11.2
22.2
90.3
0.02
0.28
1.24
3.5
10.0
28.5
0.04
0.78
3.5
9.9
32.4
141.7
14
0.1
2.8
4.9
22.4
63.5
202.5
0.04
1.75
5.1
16.5
36.9
128.0
0.05
2.39
8.3
20.7
67.4
265.0
28
0.15
2.8
5.3
20.2
54.7
202.6
0.05
1.31
3.7
11.5
38.0
95.3
0.17
2.5
4.3
22.6
73.1
202.1
42
0.06
1.5
6.6
28.0
73.0
245.0
0.02
1.5
3.4
15.0
39.0
117.0
0.08
2.6
7.8
31.0
85.0
274.0
Days Depuration
14
0.03
1.6
5.1
19.0
45.0
131.0
0.03
1.4
4.3
10.0
36.0
79.0
0.05
3.1
8.7
23.0
78.0
236.0
28
0.05
0.75a
4.4b
17.0
48.0
234.0
0.02
l.la
3.0b
12.0
34.0a
104.0
0.05
1.75a
6.65b
23.0
60.0a
184.0
-------
Table 4. Ranges of concentration factors (based on nominal and measured water
concentrations) in an uptake and depuration study with pinfish (Lagodon
rhomboides).
LIVER MUSCLE REMAINDER
Nominal Measured Nominal Measured Nominal Measured
21,000 39,000 12,000 21,000 26,000 41,000
to to to to to to
28,000 47,000 18,000 34,000 31,000 56,000
The pattern of loss of HCB by pinfish was erratic. After 14 days of depuration in
HCB-free water, the loss rate appeared high (Table 3), but samples taken after 28 days
of depuration showed a much lower loss rate. Decrease in HCB residue in liver ranged
from 4% to 50% after 28 days of depuration; in muscle, from 11% to 27%; and in
remainder of fish, from 15% to 33%.
In another study (Parrish el a/., 1974), we found that spot (Leiostomus xanthurus)
lost all detectable dieldrin residues after a 13-day depuration period. Hansen and
Wilson (1970) found that after 56 days of depuration pinfish lost 87% of DDT residues
and Atlantic croaker (Micropogon undulatus) lost 78% of accumulated DDT. Thus,
the rate of loss of HCB by pinfish appears similar to that of DDT.
It has been shown that HCB is a widespread aquatic contaminant. HCB residues
have been reported in fish eggs, fish fry and fish oil from the U. S. (Johnson et al.,
1974), in fish from Canada (Zitko, 1971), and in fish from Europe (Holden, 1970).
Although our study showed that HCB is not acutely toxic to four estuarine animals,
the compound is accumulated by an estuarine fish. Further work is needed to
determine chronic effects of HCB on estuarine animals, particularly fishes.
LITERATURE CITED
Baker, B. E. 1973. Confirmation of hexachlorobenzene by chemical reaction. Bull.
Environ. Contam. Toxicol. 10(5):279-284.
Collins, G. B., D, C. Holmes, and M. Walden. 1972. Identification of hexachloroben-
zene residues by gas-liquid chromatography. J. Chromatogr. 69(1): 198-200.
Frear, D. E. H. 1969. Pesticide Index. 4th ed. College Science Publishers, State Col-
lege, Pennsylvania 16801, 399 p.
Hansen, D. J. and A. J. Wilson, Jr. 1970. Significance of DDT residues from the es-
tuary near Pensacola, Fla. Pestic. Monit. J. 4(2):51-56.
, P R. Parrish, J. I. Lowe, A. J. Wilson, Jr., and P D. Wilson. 1971.
Chronic toxicity, uptake, and retention of Aroclor®1254 in two estuarine fishes.
Bull. Environ. Contam. Toxicol. 6(2): 113-119.
Holden, A. V 1970. International cooperative study of organochlorine pesticide resi-
dues in terrestrial and aquatic wildlife. Pestic. Monit. J. 4(3):117-135.
Holdrinet, M. V. H. 1974. Determination and confirmation of hexachlorobenzene
in fatty samples in the presence of other residual halogenated hydrocarbon pes-
ticides and polychlorinated biphenyls. J. Assoc. Off. Anal. Chem. 57 (3):580-
Johnson, J. L., D. L. Stalling, and J. W. Hogan. 1974. Hexachlorobenzene (HCB)
residues in fish. Bull. Environ. Contam. Toxicol. 11(5):393-398.
Lowe, J. I.,P R. Parrish, J. M. Patrick, Jr. and J. Forester. 1972. Effects of the poly-
chlorinated biphenyl Aroclor®1254 on the American oyster, Crassostrea virgin-
ica. Mar. Biol. (Berl.) 17:209-214.
'Mention of commercial products or trade names does not constitute endorsement by the Environmental Proteclion Agency.
186
174
-------
Parrish, P. R., J. A. Couch, J. Forester, J. M. Patrick, Jr. and G. H. Cook. 1974.
Dieldrin: effects on several estuarine organisms. Proc. 27th Ann. Conf. SE As-
soc. Game Fish Comm., in press.
Smyth, R. J. 1972. Detection of hexachlorobenzene residues in dairy products, meat
fat, and eggs. J. Assoc. Off. Anal. Chem. 55(4):806-808.
Taylor, I. S. and F. P. Keenan. 1970. Studies on the analysis of hexachlorobenzene
residues in foodstuffs. J. Assoc. Off. Anal. Chem. 53(6):1293-1295.
Zitko, V. 1971. Polychlorinated biphenyls and organo-chlorine pesticides in some
freshwater and marine fishes. Bull. Environ. Contam. Toxicol. 6(5):464-470.
175
-------
Reprinted from Water Research
Vol. 10(1): 19-24, 1976, with
permission of Pergamon Press,
Elmsford, New York
RIVER POLLUTION BY ANTICHOLINESTERASE AGENTS
D.L. Coppage and I.E. Braidech
Contribution No. 227
177
-------
Water Research Vol. 10. pp. 19 to 24. Pergamon Press 1976. Printed in Great Britain
RIVER POLLUTION BY ANTICHOLINESTERASE
AGENTS*
D. L. COPPAGE
U.S. Environmental Protection Agency, t Gulf Breeze Environmental Research Laboratory,
Sabine Island, Gulf Breeze, Florida 32561, U.S.A.
and
T. E. BRAIDECH
U.S. Environmental Protection Agency, National Field Investigations Center,
5555 Ridge Avenue, Cincinnati, Ohio 42568, U.S.A.
(Received 12 May 1975)
Abstract—The effects of effluent discharged into the Blue River, near its confluence with the Missouri
River in Kansas City, Missouri, by a manufacturer of organophosphate and carbamate pesticides were
investigated. Since these pesticides act as nerve poisons by inhibiting the neurotransmitter modulating
enzyme acetylcholinesterase (AChE) in the nervous system, poisoning of fishes was diagnosed by
measurement of brain-AChE in fishes collected from the Missouri River upstream and downstream
from the mouth of the Blue River. Other fish were exposed to diluted effluent in glass jars and their
brain-AChE measured to determine combined poisoning potential of compounds present. Fishes im-
mediately downstream repeatedly had lower brain-AChE activity than fishes upstream, and fish exposed
to diluted effluent had lower brain-AChE activity than unexposed fish. Chemical analyses showed
substantial amounts of AChE-inhibiting pesticides in the effluent relative to their toxicities. These
data indicate the effluent is a contributing factor in the reduced brain-AChE activity of Missouri
River fishes, and that brain-AChE measurement in fishes is a sensitive and reliable indicator of such
pollution.
INTRODUCTION
The major quantities of highly toxic pesticides pro-
duced are organophosphate and carbamate esterase
inhibitors (Lawless, Von Rumker and Ferguson,
1972), and they present special water pollution eva-
luation problems. Poisoning by these pesticides
requires that they be converted to compounds for
which no practical means of extraction and analytical
chemical analysis from environmental samples is
available (Aldridge, 1971; Fukuto, 1971; Metcalf,
1971). The mode of action of the organophosphate
and carbamate pesticides in animals is disruption of
nerve impulse transmission by metabolites that "irre-
versibly" inhibit acetylcholinesterase (AChE), the
enzyme that modulates levels of the neurotransmitter
acetylcholine (O'Brien, 1960, 1967; Heath, 1961;
Koelle, 1963; Ehrenpreis, 1967; Karczmar, 1970;
Aldridge, 1971; Fukuto, 1971; Metcalf, 1971). Labora-
tory and field studies of fishes have shown brain-
AChE is a good indicator of whether anti-AChE
agents are present and biologically active in water
(Weiss, 1959, 1961; Williams and Sova, 1966; Hol-
land, Coppage and Butler, 1967; Carter, 1971; Cop-
page and Duke, 1971; Coppage, 1972; Macek et al,
1972; Alsen, Herrlinger and Ohnesorge, 1973; Cop-
*Gulf Breeze Environmental Research Laboratory, Con-
tribution No. 227.
f Associate Laboratory of the National Environmental
Research Center. Corvallis, Oregon.
page and Matthews, 1974, 1975). This report concerns
use of fish AChE to investigate possible anti-AChE
poisoning resulting from discharge of effluent into a
river system by a manufacturer of organophosphate
and carbamate pesticides.
MATERIALS AND METHODS
Brain-AChE was measured in fishes collected upstream
and downstream from the outfall in the river system and
in fish exposed in the laboratory to diluted effluent. Chemi-
cal analyses were made for some of the pesticides present.
The manufacturer, situated on the east side of Kansas
City, Missouri (Fig. 1), discharges wastes from manufacture
of organophosphate and carbamate pesticides into the Blue
River, approximately one mile upstream from its con-
fluence with the Missouri River. Only the Missouri and
Kansas rivers were sampled for fish. The water quality
of the lower reach of the Blue River was so degraded
that it was uninhabitable for fish. Although other possible
sources of anti-AChE pollution existed in the area, only
one manufacturers' effluent was studied at this time.
Sampling river fish
Fish were first sampled during July 1972 at one location
on the Missouri River upstream from the mouth of the
Blue River and one location downstream. To determine
if pollution was continuous and to obtain a larger number
of samples, a second sampling was undertaken in October
1972. The second sampling was from five stations. Stations
2 and 3 were located on the Missouri River, upstream
about 6 and 13 miles respectively from the mouth of the
Blue River (Fig. 1). Station 1 was located immediately
downstream from the mouth of the Blue River and Station
5 was located about 40 miles downstream from the mouth
of the Blue River. Station 4 was located on the Kansas
19
179
-------
20
D. L. COPPAGE and T. E. BRAIDECH
W94° 45'
N 39° 25'-\-
30'
15
94 00 W
-1-39°25'N
.exington
15'
-I-39° 00 N
94°00'W
Fig. 1 Location of pesticide manufacturer and fish sampling stations on Missouri and Kansas rivers.
River about 6 miles upstream from where it joins the Mis-
souri River.
Brain-AChE activity of river fishes was expressed as
mean percentage of brain-AChE activity of fishes collected
at Station 1 near the mouth of the Blue River, and statisti-
cal comparisons were made (Student's t-test, P < 0-05)
to determine if fishes from other stations had significantly
greater brain-AChE activities than fishes from Station 1.
Laboratory exposure offish
"Young of the Year" channel catfish, Ictalurus punctatus
(50-100 mm total length), were exposed in static tests to
river water to which effluent obtained from the manufac-
turer had been added. Fish were exposed for 96 h in 8
1. of water in 10-1. wide-mouthed glass jars. Four groups
of exposures were made from dilutions of 1 18-h and 2
24-h composite samples of the final effluent, collected on
three consecutive days. Statistical comparisons of AChE
activities of fish exposed in jars were made with unexposed
populations (Student's t-test).
The catfish used in the tests were obtained in the Kansas
City area from a commercial grower. Upon arrival at the
test site, a number of fish to be used as a background
sample were collected and frozen. The remainder of the
test fish were placed in 100% dilution water (Missouri
River water collected upstream from Kansas City, Kansas).
No fish died in the holding tank for the duration of the
testing period.
One test was set up on each of three consecutive days,
using composite samples of final effluent collected each
day. Using an effluent discharge rate of 400,000 gal (about
1,514,000 1.) per day and a low Missouri River flow of
7860 ft3 s~' (about 19,277,995,000 1. day'1), dilutions of
1:1300 and 1:650 were formulated (about 0-1 and 0-05 the
dilution of effluent that would occur in the Missouri River
if mixed completely). Limited facilities precluded long-term
exposures to greater dilutions. However, organophosphate
pesticide concentrations that cause brain-AChE inhibition
in fish in short-term exposures usually cause inhibition at
100-fold greater dilutions in 15-30 day exposures (Weiss,
1959; 1961; Weiss and Gakstatter, 1964). The controls and
each dilution were run in triplicate for each test. Five fish
were exposed in e;ich test chamber.
Another lest, consisting of dilutions of 1:55, 1:30 and
1:17, was performed, using the composite sample collected
during the second day of sampling. This test was a "back-
up" to estimate potential long-term anti-AChE activity in
case no short-term AChE inhibition was found in the high
dilution tests. There was no replication in these tests, but
each lest chamber contained 10 fish. For analytical pur-
poses, the exposed fish were divided into two groups of
five fish each.
It was necessary to aerate the individual containers dur-
ing the course of the testing because of the high oxygen
demand of the dilution water and the loading in the test
chambers. This was accomplished by slowly bubbling pure
oxygen into the test containers twice a day at approxi-
mately 12-h intervals.
AChE measurements
The AChE of the brains of fishes was assayed with a
pH-stat method previously described (Coppage, 1971). In
the case of the larger river fishes, each assay sample was
a single brain. Three to five brains were pooled for each
assay sample of "young of the year" fish tested in the labor-
atory. Specific activity is expressed as micromoles of acetyl-
choline hydrolyzed per hour per mg of brain tissue.
Chemical analyses of bioassay water for pesticides
Pesticides were extracted from samples of effluent,
effluent diluted 1:650, and the dilution water with hexane.
Gas chromatographic-mass spectrometric analyses were
performed on these extracts with methods described by
Webb et al. (1973). No analyses were made on other dilu-
tions. Pesticides in production at the plant during the sam-
pling period are given in Table 1.
RESULTS AND DISCUSSION
Carp (Cyprinus carpio) caught upstream from the
mouth of the Blue River in July, 1972 had significantly
(P < 0-05) greater brain-AChE activity than carp
caught downstream from the mouth of the Blue River
(Table 2). The brain-AChE activity of the upstream
fish was 205% of those caught downstream.
Three species of fish were obtained in the October,
1972 sampling in quantities sufficient for comparison
of brain-AChE activity to that of fish of the same
species at Station 1. Carp at all upstream stations
(Stations 2-4) had significantly greater brain-AChE
activity than carp immediately downstream from the
mouth of the Blue River (Station 1) (Table 3). Carp
from Station 5, downstream from Station 1, also had
significantly greater brain-AChE activity than carp
from Station 1. The AChE activity patterns in carp-
suckers, Carpiodes sp., at the upstream and down-
stream stations were similar to carp at the same
180
-------
River pollution by anticholinesterase agents
21
Table 1. Common and chemical names of pesticides in production
during testing
Azinphosraethyl
(Guthion*)
CoumaphoB
(Co-rala)
Disulfoton
(Di-Svstona)
Fensulfothlon
(Dasanlta)
Methamidophos
(Monitor8)
Propoxur
(Baygona)
Dyrene3
Chemical name
J3,0_-d line thy 1 £ (4-oxo-l,2,3 benzotriazin-
3 (4H)-ylraethyl)phoBphorodithioate
()(3-chloro-4-methyl-2-oxo-2H-l-benzopyran-
7-yl) 0.0-diethyl phosphorothioate
t>,£-diethyl ^-2(ethylthio)ethyl)phoaphoro-
dithloate
0,£-diethyl 0-p-(methylsulf inyl)phenyl)
phosphorothioate
0,S-dimethyl phosphoroamldothioate
o-iaopropoxyphenyl methylcarbamate
2.4-dichloro-6-(o-chloroanilino)-s-trlazine
"Trade name. Use of trade names does not constitute endorsement
by the U.S. Environmental Protection Agency. Dyrene is not known
to be an AChE inhibitor.
stations. Gizzard shad, Dorosoma cepedianum, from
upstream stations did not have significantly greater
brain-AChE activity than gizzard shad from Station
1. Gizzard shad are the least valuable of the three
species as an indicator of "point-source'7 pollution by
anti-AChE agents because they are more migratory
and may move into and out of a polluted area
rapidly. During the second sampling, all three species
had the greatest AChE activity at Station 5. This
apparent loss of biological activity of anti-AChE
agents against fish 40 miles downstream may be due
to breakdown or sorption by particles, biota, and
sediment.
In laboratory tests, fish exposed to high con-
centrations of effluent from the second composite
sample showed marked inhibition of brain-AChE,
Table 2. AChE activity in Missouri River fish collected upstream and downstream from
mouth of Blue River, July 1972
Species No.
Carp 4
Carp 3
Total length
(range mm)
260-410
400-475
Location from mouth
of Blue River
Downstream
Upstream
Mean brain-AChE
activity + SD
0.40 + 0.06
0.82 + 0.09
Percent of
downstream
AChE
100
205
Significantly
greater AChE
than downstream
at t
0.05
—
Yes
Table 3. AChE activity in Missouri River fishes collected upstream and downstream from mouth
of Blue River, October 1972
Species No.
Carp 5
3
5
2
6
Carp- 5
Sucker
2
4
2
Shad 6
9
4
4
Total length
(range mm)
300-601
308-451
308-525
510-582
305-449
351-508
228-345
315-391
215-345
251-292
280-345
310-356
225-295
Station Location from mouth
of Blue River
1 Immediately downstream
2 5.8 miles upstream
3 13.2 miles upstream
4 Kansas River
5 39.8 miles downstream
1 Immediately downstream
2 5.8 miles upstream
3 13.2 miles upstream
5 39.8 miles downstream
1 Immediately downstream
2 5.8 miles upstream
3 13.2 miles upstream
5 39.8 miles downstream
Mean
AChE
0.59
11.87
0.82
0.96
1.02
1.22
2.07
1.89
2.26
0.78
0.78
0.70
1.04
brain-AChE
activity +
+ 0.17
+ 0.09
+ 0.11
+ 0.00
+ 0.04
+ 0.23
+ 0.04
+ 0.38
+ 0.23
+ 0.10
+ 0.07
+ 0.13
+ 0.06
Percentage of Significantly
SD station 1 AChE greater AChE
than station 1
at t
0.05
100
147
139
163
173
100
170
155
185
100
100
90
133
Yes
Yes
Yes
Yes
Yes
Yes
Yes
—
No
No
Yes
181
-------
22 D. L. COPPAGE and T. E. BRAIDECH
Table 4. Inhibition of catfish, Ictalurus punctatus, brain-AChE by in vivo exposure to various dilutions
of effluent from pesticide manufacturer sampled on different days
Totml length Sa»ple sequence So. Effluent Mean AChE Mean percent
offLroS anZcreatLnt dilution activity + 3D Inhibited
50-100 Ho. 1 Control 8 1-41 ±
" Exposed 96 b 2 1:55 0.38 +
» » 2 1:30 0.20 ±
2 1:17 0.18 +
So. 2 Control 3 1.50 +
" &cpo««d 96 h 3 1:1300 1.36 +
" 3 1:650 1.17 +
Bo. 3 Control 3 1.57 +
" Exposed 96 h 3 1:1300 1.44 +
" " 3 1:650 1.22 +
No. 4 Control 3 1.44 +
" Exposed 96 h 3 1:1300 1.14 +
3 1:650 0.75 +
0.05
0.06 73
0.02 86
0.00 87
0.00
0.07 9
0.08 22
0.02
0.11 8
0.07 22
0.11
0.06 21
0.05 48
Significantly
inhibited at
f < 0.001
f < 0.001
P < 0.001
—
P < 0.05
P < 0.01
—
P < 0.20
P < 0.01
P < -0.025
P < 0.001
when compared with control fish (Table 4). A dilution
of 1 part effluent to 17 parts river water produced
87% inhibition, 1 part effluent to 30 parts river water
produced 86% inhibition, and 1 part effluent to 55
parts river water produced 73% inhibition of AChE.
The fish exposed to high dilutions showed less
AChE inhibition than those exposed to lower dilu-
tions (Table 4), when compared with control fish. The
first and second effluent samples taken on succeeding
days produced about the same inhibitions at the same
dilutions. Both produced 22% brain-AChE inhibition
at the 1:650 dilutions, and 9 and 8% brain-AChE
inhibition, respectively, at the 1:1300 dilutions. The
22% AChE inhibition values were significantly below
brain-AChE activity of control fish (P < 0-01). The
third sample of effluent produced an even greater in-
hibition of brain-AChE, 48% inhibition (significant at
P < 0-001) at the 1:650 dilution and 21% inhibition
(significant at P < 0-025) at the 1:1300 dilution. Since
these laboratory exposures to effluent produced sig-
nificant brain-AChE inhibition in only 96 h and the
effluent probably enters the river continuously, we
conclude that the effluent is probably a contributing
factor in the lower brain-AChE activity found in fish
immediately downstream from the mouth of the Blue
River.
Table 5. Chemical analyses of effluent water, diluted effluent and Missouri River water for
pesticide
Ssnple source
Miaaouri River water upstreaa
frtm the Blue River
Blink
lot effluent aaaple
lit 1:630 dilution
2nd affluent sample
2nd 1:650 dilution
3rd effluent anple
3rd 1:650 dilution
Compounds found*
None
None
Dtsulfoton (Di-Syston)
Penaulfothion (Daaanit)
AzinphoBmethyl (Guthlon)
None
Diaulfoton
Fena ulf othion
Az inphonaethyl
Propoxur (Baygon)
None
Diaulfoton
Pen rulf othion
Ar inphoraetby 1
Propoxur
None
Range of quantity of pestlcidea
found (ppb)a
400-800
400-800
400-800
300-600
300-600
300-600
300-600
2000-4000
2000-4000
2000-4000
2000-4000
—
"Lo»e«t detectable ijuantity for analytical nethoda used waa 50 ppb.
182
-------
River pollution by anticholinesterase agents
23
The results of chemical analyses of water samples
are given in Table 5. The data show substantial
amounts of the anti-AChE agents disulfoton, fensul-
fothion, azinphosmethyl, and propoxur in the effluent
at various times. The maximum concentrations after
1:650 dilution of the effluent were below the detect-
able limits (50 ppb, \>% 1"') for the analytical methods
used. However, residues ranging as high as 2000-4000
ppb were detected for each of the above pesticides
in the undiluted effluent. A dilution of 1:650 would
result in a theoretical concentration of about 3-6 ppb
each or a combined concentration of 12-24 ppb. If
uniformly distributed in Missouri River water, com-
bined concentrations as high as 0-6-1-2 ppb could
result.
Some relevant studies have been made on the
detected pesticides. Previous reports have indicated
that about 3-2 kg of azinphosmethyl enter the river
each day (about 0-16 ppb at low flow) (Lawless et
al., 1972). Concentrations of azinphosmethyl as low as
0-05 ppb can produce brain-AChE inhibition in gold-
fish, Carassius auratus, and bluegill sunfish, Lepomis
macrochirus, continuously exposed for 30 days (Weiss
and Gakstatter, 1964). Also, azinphosmethyl can
cause about 45% brain-AChE inhibition in channel
catfish after only 12 h exposure to a concentration
of 10 ppb (Carter, 1971). Static-test 96-h LC50 values
of azinphosmethyl for fresh-water fishes have been
reported in concentrations as low as 3-2 ppb for rain-
bow trout, Salmo gairdneri, (Katz, 1961) and as high
as 3290 ppb for channel catfish (Macek and McAllis-
ter, 1970). The 24-h LC50 of disulfoton for bluegill
sunfish has been reported as 40 ppb and the 48-h
LC50 of propoxur to fathead minnows, Pimephales
promelas, 25 ppb (Fed. Wat. Pollut. Control Adm.,
1968). Butler (1963) reported the 48-h LC50 of fensul-
fothion for longnose killifish, Fundulus similis, as 55
ppb. Frequent or long-term exposure to pesticide con-
centrations much lower than the acute LC50 can
result in harmful AChE inhibition and deaths (Post
and Leasure; 1974; Eaton, 1970; Lahav and Sarig,
1969; Weiss and Gakstatter, 1964).
It is apparent from these data that substantial harm
could occur to the fishes studied—and to more sensitive
species—at concentrations much lower than the limits
of detectability (50 ppb) for chemical analyses used
in this study. Chronic effects should be studied and
more sensitive chemical analyses should be made for
pesticides present in water.
Acknowledgements—We thank Dr. Richard Endrione and
his staff, National Field Investigations Center (NFIC), for
chemical analysis of water for pesticides, Mr. Ernest Kar-
velis and the biology staff of NFIC for collecting field sam-
ples, and Mr. Edward Matthews, Gulf Breeze Environmen-
tal Research Laboratory, for assistance in AChE assays.
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Holland H. T., Coppage D. L. and Butler P. A. (1967)
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184
-------
Reprinted from Archives of
Environmental Contamination
and Toxicology, Vol. 4(4):
435-442, 1976, with permis-
sion of Springer-Verlag
New York Inc.
EFFECTS OF LEACHED MIREX ON EXPERIMENTAL COMMUNITIES
OF ESTUARINE ANIMALS
M.E. Tagatz, P.M. Borthwick, J.M. Ivey, and J. Knight
Contribution No. 229
185
-------
EFFECTS OF LEACHED MIREX ON
EXPERIMENTAL COMMUNITIES OF ESTUARINE
ANIMALS1
M. E. TAGATZ, P w. BORTHWICK, J. M. IVEY, and J. KNIGHT
Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Gulf Breeze. Florida 32561
Experimental communities of various estuarine animals in outdoor tanks were
exposed to a continuous flow of water containing mirex for 10 weeks. The mirex was
leached from fire ant bait (0.3% active ingredient) by fresh water which was then
mixed with salt water to yield exposure concentrations averaging 0.038 /Ag/L. The
experiment simulated runoff from treated land into estuarine areas. Mortality of grass
shrimp (Palaemonetes vulgaris), pink shrimp (Penaeus duorarum), common mud crabs
(Panopeus herbstil), and striped hermit crabs (Clibanarius vittatus) was significantly
higher in tanks containing the toxicant. Mortality of ribbed mussels (Modiolus demis-
sus) and American oysters (Crassostrea virginica) was significantly lower in treated
tanks, probably because numbers of both species of crabs, which ate the bivalves, were
reduced. Sheepshead minnows (Cyprinodon variegatus) were least affected by mirex.
Almost all deaths occurred after 10 or more days of exposure. All exposed animals
accumulated mirex, with maximum concentrations ranging from 5,500X (pink shrimp)
to 73.700X (soft tissues of oysters) above the concentration in the water. Sand
substratum contained mirex up to 1,500X that in the water. The study demonstrated that
mirex can be leached from bait by fresh water and concentrated by and affect survival
of members in an experimental estuarine community.
The purpose of this study was to determine the effects of mirex2 leached from fire
ant bait3 by fresh water on communities of estuarine organisms. Mirex is a chlori-
nated hydrocarbon insecticide applied in bait form to control the imported fire ant,
Solenopsis richteri Forel, in the southeastern United States.
Field studies have demonstrated translocation of mirex from treated land to
estuarine biota (Borthwick et al. 1973), and in three cases during periods of heavy
runoff, 0.03 /u-g/L was found in water from streams in Mississippi after application
of mirex bait to the watershed (Alley, personal communication4). Some possible
'Contribution No. 229 from the Gulf Breeze Environmental Research Laboratory.
2Dodecachlorooctahydro-l,3,4-metheno-2H-cycIobuta [cd] pentalene.
384.7% corn cob grits, 15.0% soybean oil, and 0.3% mirex.
*E. G. Alley, Mississippi State Chemical Laboratory, State College, Miss. 39762.
Archives of Environmental Contamination 435
and Toxicology Vol. 4, No. 4, 435-442 (1976)
®1976 by Springer-Verlag New York Inc. •*• ° '
-------
436 M- E. Tagatz et al.
routes of entry into the estuarine environment are biological transport, tidal action,
or fresh water runoff containing the bait or mirex leached from bait.
Our study considers the possibility that organisms may not come into direct
contact with the bait, but could be exposed to leached mirex carried from treated areas
by runoff into drainage systems. Two other studies concerned exposure of aquatic
organisms to mirex leached from bait rather than to the bait or to technical mirex
introduced into water via a solubilizer. Ludke et al. (1971) exposed fresh water
crayfish in small aquaria where the bait was enclosed in filter paper and screen wire.
Tagatz et al. (1975) held various estuarine animals caged in outdoor tanks that
received water containing mirex from gravity-flow columns that contained the bait.
The present study is an attempt to determine effects of relatively long exposure to
low concentrations of mirex on experimental communities of estuarine animals.
Materials and methods
An experiment was conducted for a period of 10 weeks (March 26 to June 4,
1974) on estuarine animals in six fiberglass tanks of 2.4-m diameter. Depth of the
water, controlled by standpipes, was 30.5 cm. The tank area was protected by a
fiberglass roof and backwall.
Of 6 tanks, 3 tanks each received a continuous flow of water containing mirex
from a gravity-flow column that contained 150 g of mirex bait. Three control tanks
were treated in the same fashion with bait that did not contain mirex (Borthwick
et al. 1975). Tap water (0.5 L/min), one-half of which trickel-filtered through
the column, and natural sea water (1.0 L/min) were siphoned from constant head
boxes and mixed in a glass trough (positioned below the column) before entering
the tank. Tap water (chlorine content <0.1 mg/L) to the freshwater constant
head box was filtered through 1 /A-pore size cartridges; sea water (from Santa
Rosa Sound, Florida) was unfiltered. The amount of bait in the column and the rate
of siphon flows were chosen after preliminary testing indicated that they produced
low but detectable residues of mirex in tank water. The desired range was trace
amounts (<0.01 /ig/L) to 0.09 /xg/L. Tap water containing mirex from the columns
was mixed with salt water to simulate conditions in an estuary, where fresh water
runoff from the watershed may contain leached mirex.
Temperatures (12:00 Noon) and salinities (8:00 AM) of the tank water were
measured at the start of the test and five times weekly thereafter. Temperature and
salinity changes were gradual. Temperature responded to natural changes in air and
sea water temperatures; salinity, to natural fluctuations in the salinity of the sea
water source. Dissolved oxygen and pH were measured at the beginning, middle,
and end of the study.
Tank floors were covered with 2.5 cm of beach sand over 3 mm-bar netting, and a
188
-------
Effects of Mirex on Estuarine Animals 437
pile of 6 rocks was placed on the sand. The netting was used to capture small
experimental animals at the end of the test. Granite rocks (about 15 x 15 x 2 cm
size) provided an additional type of habitat.
All animals were captured in local waters and held 3 or 4 days in stock tanks
before transfer to experimental tanks. Prior to transfer, they were acclimated for 4 to
6 hr to the initial salinity and temperature of the experiment. Animals in each
experimental tank were 100 grass shrimp, Palaemonetes vulgaris (22 to 34 mm
rostrum-telson); 25 pink shrimp, Penaeus duorarum (83 to 105 mm rostrum-telson);
15 sheepshead minnows, Cyprinodon variegatus (39 to 63 mm total length); 15
common mud crabs, Panopeus herbstii (15 to 26 mm carapace width); 15 striped
hermit crabs, Clibanarius vittatus (size of univalve habitats, 41 to 125 mm); 25
ribbed mussles, Modiolus demissus (55 to 80 mm high); and 17 American oysters,
Crassostrea virginica (55 to 85 mm high). Size distributions of the various species
were similar in all tanks. Each community was fed 150 cubes (cm3) of fish meat
weekly. Members could also consume plankton from unfiltered seawater and at-
tached algae.
The percentage mortality of animals was determined at the end of the experiment.
The analysis of variance was used to determine significant difference between
treated and control tanks. To satisfy the assumption of normality, the percentage
data were first transformed to arc sins and then put into a format for a two-factor
analysis of variance (Steel and Torrie 1960). Tukey's multiple comparison test was
used to show significant difference in means between control and treated tanks for
all species (Snedecor and Cochran 1967).
Samples of water, sand, and animals were analyzed by electron-capture gas
chromatography to determine mirex concentrations. Limits of detection were 0.01
/u,g/g for whole animals (wet-weight basis) and for sediment (air-dried weight basis),
and 0.01 /u-g/L for water. Only the soft tissues of mussels or oysters were analyzed.
Samples to which known amounts of mirex were added gave recovery rates greater
than 85%, but concentrations were not corrected for percentage recovery.
Techniques for residue analyses were those of Lowe et al. (1971) except for tissue
samples that weighed less than 5 g for which the method of Hansen et al. (1974) was
used. Samples of water from each tank were obtained at the start of the experiment
and twice a week thereafter. A water sample consisted of a composite collection
from 2 sites in each tank. At 21, 42, 63 and 70 days, a composite sample of sand
from 4 sites was obtained from each tank. Surviving animals (composite samples)
were analyzed for mirex at the end of the experiment, and dead animals (individual
or composite samples) were analyzed as death occurred. In addition, mirex in live
samples (3 animals per tank) of mussels or oysters were determined at 21 and 42
days. Mirex was not detected in pretest samples of sand and animals or in fish
muscle used as food.
Samples of hermit crabs surviving at the end of the experiment were fixed for
histological examination. These consisted of 8 treated and 8 control crabs.
189
-------
438 M. E. Tagatzer al.
Results and discussion
Averages and ranges of rairex in water (/tg/L) in the 3 treated tanks were 0.033
(trace to 0 072) 0.037 (trace to 0.080) and 0.043 (trace to 0.083). The value 0.005
was substituted for trace residues (<0.010) for calculating averages. Analysis of
variance tests showed no significant differences (oc = 0.05) in residues among the 3
treated tanks, and no significant increase or decrease with time within individual
tanks.
Temperatures of tank water averaged 23.6°C (range, 15.8 to 28.5°C) and
salinities 13.5 parts per thousand (range, 8 to 18 ppt). Dissolved oxygen measure-
ments were from 9 to 11 ppm; pH, from 8 to 9-
Analysis of variance (using arc sin data) show that species difference, toxicant
effect, and interactions were significant (Table I). Some species are more tolerant
than others, and mirex does influence percentage mortality.
Mean percentage mortality values observed in control and treated tanks are
summarized in Table II. Differences between control and toxicant were significant
for all seven species (Table III). These analyses show that in the tanks containing the
toxicant, mortality of grass shrimp, pink shrimp, mud crabs and hermit crabs was
higher while the mortality of minnows, mussels and oysters was lower.
The first observed death among exposed grass shrimp was at 27 days and among
exposed pink shrimp at 1 day. Four exposed and seven control pink shrimp died the
first 2 days (cause of mortality not known), but subsequent deaths of exposed shrimp
were after 10 days and control shrimp after 27 days. All pink shrimp in treated tanks
were dead at 41 days. Distressed pink shrimp were of darker coloration, lost
equilibrium, and no longer burrowed into the sand. After the initial mortality of
control pink shrimp, high mortality of unknown cause occurred from 27 to 70 days.
Although fish and crabs ate dead shrimp, they were not observed to attack living
Table I. Analysis of variance for mortality of animals after a 70-day exposure to
mirex. Two-factor, completely randomized
Source of
variation df SS MS Fobs F,,.9S
Total
Treatments
Species
Toxicant
Interaction
Error
41
13
6
1
6
28
21778,32
20027.78
16609.17
403.06
3015.55
1750.54
1540.59
2768.19
403.06
502.59
62.51
24.65
44.28
6.45
8.04
—
2.12
2.45
4.20
2.45
—
190
-------
Effects of Mirex on Estuarine Animals 439
Table II. Summary of mean percentage mortality of animals after a 70-day exposure
to mirex. Data is on three control and three treated tanks
Animals
Grass shrimp
Pink shrimp
Sheepshead minnows
Mud crabs'
Hermit crabs
Ribbed mussels
American oysters
Total number
Control Treated
300
75
45
45
45
75a
51a
300
75
45
45
45
75"
51a
Mean %
Control
39
81
29
22
11
25
6
mortality
Treated
53
100
20
47
56
9
2
Includes living animals (18 mussels or 18 oysters) removed before 70 days for residue
analyses.
individuals. In shrimp, as in all animals, it is probable that some deaths were due to
mirex alone, to interactions with mirex, or to some other factor such as predation.
Sheepshead minnows were least affected by mirex. They exhibited natural ac-
tivities such as defending territories, burrowing in sand, and consuming algae. This
species spawned in the presence of mirex since each tank contained 100 to 600
young from < 10 to 32 mm TL at the end of the study.
Toxicity to crabs, as to shrimp, was not evident until after 10 or more days of
exposure. First death among exposed mud crabs was at 24 days, and deaths occurred
Table III. Multiple comparisons showing significant difference at 5% level
(>5.06) between control and treated animals after 70 days of exposure to mirex.
Means are arrived by converting percentage mortality data to arc sin V mean
mortality data (p. 158 of Steel and Torrie 1960)
AnimaJs
Grass shrimp
Pink shrimp
Sheepshead minnows
Mud crabs
Hermit crabs
Ribbed mussels
American oysters
Mean
control
38.81
64.61
32.22
28.14
18.81
29.25
14.18
Mean
treated
46.91
90.00
25.77
43.03
48.31
10.65
4.72
Difference
8.10
33.49
6.45
14.89
29.50
18.60
9.46
191
-------
440 M. E. Tagatz et al.
until the end of the study. Distress among mud crabs, noted only in treated tanks,
consisted of paralysis and/or loss of equilibrium. Those affected were often on their
backs for 1 to 1 1 days before dying. Most mud crabs in treated tanks did not seek
concealment, but almost all control crabs hid among shells or rocks. We observed
that some mud crabs molted in all tanks. At the end of the study, 1 exposed and 6
control crabs had grown larger than 26 mm, the maximum initial size.
One hermit crab died after 10 days of exposure, and the other deaths occurred
after from 21 to 70 days of exposure. All dead animals were out of their shell
habitats. However, we noted 2 crabs in treated tanks that seemed partially paralyzed
while within their protective shells. We also noted that those hermit crabs in control
and treated tanks that did not occupy other shells after molting, either retained the
same shells or left their shells for shelter, usually under oysters. Crabs without shells
in treated tanks moved awkwardly compared to those in controls. Dr. John A.
Couch, pathologist at this laboratory, found no significant histological differences in
samples of hermit crabs from treated and control groups.
More exposed than control mussles and oysters may have survived because
numbers of both species of crabs, which ate the bivalves, were significantly reduced
in treated tanks. In treated and control groups, valves of oysters and particularly
mussels were cracked and the meats consumed. Empty bivalve shells would often
provide habitats for mud crabs.
Deaths of estuarine crabs and shrimp exposed to mirex occurred in other studies
using this and other methods of exposure. Using the same experimental design as in
the present study, Tagatz et al. (1975) found significant mortality among caged
Callinectes sapidus, Palaemonetes pugio , orPenaeus duorarum exposed to less than
0.53/xg/L of leached mirex for 4 weeks. The toxicity of mirex wh'en consumed as
bait has been shown by McKenzie (1970) and Lowe et al. (1971). Toxicity using
technical grade mirex has been demonstrated by the following experiments. Lov/eet
al. (1971) found that a significant number of juvenile P. duorarum died during a
7-day exposure to 1.0 ^g/L in sea water. Redmann (1973) reported a 40% mortality
of P. pugio in 12 days from a 48-hour exposure to 0.01 mg/L of technical mirex.
Bookhout et al. (1972) showed that survival or development of larval Rhit-
hropanopeus harrisii and Menippe mercenaria was affected by concentrations from
0.01 to 10.0
Animals exposed to leached mirex concentrated the chemical (Table IV). Con-
centrations were not consistently higher in either surviving or dead animals.
Maximum concentration factors of residues by animal category (times concentraton
in water) were 21.100X for shrimp; 50.000X for fish; 71.100X for crabs; and
73.700X for molluscs. Residues in young fish hatched in the treated tanks ranged
from 0.34 to 0.74 fj,g/g, compared to a range of 1.1 to 1.9 /u,g/g in adults. Mirex was
not detected in live mussels at 21 days, but samples contained up to 1.2 jtig/g at 42
days and 2.0 /ig/g at 70 days. The range of mirex residues in live oysters at 21 and
42 days was 1.2 to 1.7 jtg/g; at 70 days, 1.3 to 2.8 /tg/g. Residues of mirex (/*g/g)
192
-------
Effects of Mirex on Estuarine Animals 441
Table IV. Mirex residues in live animals after 70 days exposure (composite sample
from each of three treated tanks) and in dead animals (individual or composite
samples from one or more tanks)
Animals
Grass shrimp
Pink shrimp
Sheepshead minnows
Mud crabs
Hermit crabs
Ribbed mussels
American oysters
Mirex in live
animals (ju.g/g)
0.50-0.66
—
1.1 -1.9
0.57-0.71
1.7 -2.7
1.6 -2.0
1.3 -2.8
Maximum
concn. factor"
17.400X
—
50,OOOX
18.700X
71.100X
52.600X
73.700X
Mirex in dead
animals (ju-g/g)
0.47-0.80
0.02-0.21
0.35-0.78
0.22-1.2
0.23-1.9
—
—
Maximum
concn. factor3
21,100X .
5,500X
20,500X
31.600X
50,OOOX
—
—
Maximum residue in living or dead animal compared to the average concentration in water
(0.038 ,ug/L).
in various categories of living animals from estuaries near treated areas in South
Carolina (Borthwick et al. 1973) compared to those in our study (in parentheses)
are: shrimp, <0.01 to 1.3 (0.02 to 0.80); fish, <0.01 to 0.82 (0.35 to 1.9); crabs,
<0.01 to 0.60 (0.22 to 2.7); and bivalve molluscs, <^0.01 (<0.01 to 2.8).
Sand substrata adsorbed mirex from the test water, concentrations increasing with
time. Ranges of residues (pg/g) in the three treated tanks were <0.010 to 0.012 at
21 days, 0.019 to 0.029 at 42 days, 0.031 to 0.042 at 63 days, and 0.031 to 0.050 at
70 days. The greatest concentration of mirex in sand relative to that in test water
after 70 days was 1,500X (0.050 /u,g/g in sand and 0.033 /ug/L in water).
No mirex was detected in water, sand, or animals from control tanks.
Our study illustrates that mirex can be leached from bait by fresh water and can
be concentrated by and affect survival of members of an experimental estuarine
community.
Acknowledgments
The authors wish to tank Dr. John A. Couch of the Gulf Breeze EPA laboratory
for histological examination of samples of animals, and Dr. Alvin L. Jensen, School
of Natural Resources, University of Michigan, for statistical assistance.
193
-------
442 M. E. Tagatz et al,
References
Bookhout, C. G., A. J. Wilson, Jr., T. W. Duke, and J. I. Lowe: Effects of mirex
on the larval development of two crabs. Water, Air, Soil Pollut. 1, 165 (1972).
Borthwick, P. W., T. W. Duke, A. J. Wilson, Jr., J. I. Lowe, J. M. Patrick, Jr.,
and J. C. Oberheu: Accumulation and movement of mirex in selected estuaries
of South Carolina, 1969-71. Pestic, Monit. J. 7, 6 (1973).
Borthwick, P. W., M. E. Tagatz, and J. Forester: A gravity-flow column to provide
pesticide-laden water for aquatic bioassays. Bull. Environ. Contam. Toxicol.
13, 183 (1975).
Hansen, D. J., P. R. Parrish, and J. Forester: Aroclor® 1016: Toxicity to and uptake
by estuarine animals. Environ. Res. 7, 363 (1974).
Lowe, J. I., P. R. Parrish, A. J. Wilson, Jr., P. D. Wilson, and T. W. Duke:
Effects of mirex on selected estuarine organisms. Trans. 36th N. Am. Wildl.
Nat. Resour. Conf. p. 171 (1971).
Ludke, J. L., M. T. Finley, and C. Lusk: Toxicity of mirex to crayfish, Procam-
barus blandingi. Bull. Environ. Contam. Toxicol. 6, 89 (1971).
McKenzie, M. D.: Fluctuations in abundance of blue crab and factors affecting
mortalities. S. C. Wildl. Resour. Dept., Marine Resour. Div., Tech. Rep. No.
1, 45 p. (1970).
Redmann, C.: Studies on the toxicity of mirex to the estuarine grass shrimp,
Palaemonetes pugio. Gulf Res. Rep. 4, 272 (1973).
Snedecor, G. W., and W. G. Cochran: Statistical methods, 6th ed. Ames: Iowa
State Univ. Press (1967).
Steel, R. G. D., and J. H. Torrie: Principles and procedures of statistics. New York:
McGraw-Hill (1960).
Tagatz, M. E., P. W. Borthwick, and J. Forester: Seasonal effects of leached mirex
on selected estuarine animals. Arch. Environ. Contam. Toxicol 3, 371
(1975).
Manuscript received February 3, 1975; accepted April 15, 1975
194
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Reprinted from Developments
in Industrial Microbiology,
Vol. 16: 256-261, 1975, with
permission of the Society for
Industrial Microbiology
INHIBITION OF ESTUARINE MICROORGANISMS
BY POLYCHLORINATED BlPHENYLS
A.M. Bourquin, L.A. Kiefer, N.H. Berner, S. Crow, and D.G. Ahearn
Contribution No. 230
195
-------
Reprinted from Volume 16 of DEVELOPMENT IN INDUSTRIAL MICROBIOLOGY
A Publication of the Society for Industrial Microbiology
AMERICAN INSTITUTE OF BIOLOGICAL SCIENCES • WASHINGTON, D.C. • 1975
CHAPTER 25
Inhibition of Estuarine Microorganisms by
Polychlorinated Biphenyls*
A. W. BOURQUIN AND L. A. KlEFER
U.S. Environmental Protection Agency, Gulf Breeze Environmental Research
Laboratory, ** Sabine Island, Gulf Breeze, Florida 32561
N. H. BERNER, S. CROW, AND D. G. AHEARN
Department of Biology, Georgia State University, Atlanta, Georgia 30303
Over 100 isolates of representative estuarine bacteria and fungi were screened for their
ability to grow in the presence of commercial preparations of polychlorinated biphenyls
(PCB). Super absorbant sensitivity discs impregnated with up to 0.5 mg of PCB were placed
on the surface of freshly inoculated solid media. Twenty-six bacteria, representing both
gram-positive and gram-negative strains of varying morphology, showed varying degrees of
sensitivity to PCB. In contrast to insensitive isolates, sensitive strains were mainly amylolytic
and proteolytic. PCB had negligible effect on the growth of fungi. The sensitivity of select
cultures of heterotrophic bacteria to PCB may be of considerable importance to nutrient
turnover in estuarine ecosystems.
197
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INTRODUCTION
Polychlorinated biphenyl formulations (PCB's) are chemically and thermally stable and
possess high dielectric constants. Because of these properties, PCB's have been important
commercially as coolant-insulation fluids in capacitors and transformers, hydraulic fluids,
plasticizers, lubricants, and fire'retardants. Jensen et al. (1969) were among the first to
note the magnification of PCB's in the food chain. Subsequent studies have shown their
environmental effects to be similar to those of DDT. The chemistry and persistence of
PCB's in the environment and their chronic toxicity for various animals have been
reviewed by Peakall and Lincer (1970) and Gustafson (1970). One report (Keil et al.
1972) describes a commercial PCB formulation in concentrations of 0.1 Mg/ml which
stimulated the growth of Escherichia colt. Ahmed and Focht (1973) reported
biodegradation of PCB isomers 2 to 5 chlorines by Achromobacter pCB. Little
information on the interactions of PCB's with heterotrophic microorganisms is available.
Our investigation examines the effects of two commercial PCB formulations (Aroclor®
1016 and 1242) on selected estuarine bacteria and fungi.
*GBERL Contribution No. 230.
••Associate Laboratory of National Environmental Research Center, Corvallis, Oregon 97330.
®Registered trademark, Monsanto Company, St. Louis, Mo. Mention of commercial products or trade names does not
constitute endorsement by the U.S. Environmental Protection Agency.
256
198
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CONTRIBUTED PAPERS
257
MATERIALS AND METHODS
Organisms. Bacterial isolates were obtained from estuarine waters and sediments of
Pensacola Bay, near Gulf Breeze, Florida. Yeast isolates were obtained from waters and
sediments of Barataria Bay, Louisiana. Biochemical analyses for bacteria were performed
by the methods of Colwell and Wiebe (1970) and identifications by means of Breed et al.
(1957).
Media. The medium for bacteria contained 1.0 g yeast extract (Difco) and 5.0 g peptone
(Difco)/liter of aged artificial seawater (Rila Marine Mix, aged 1 mo in dark) at 20%P
salinity (adjusted with distilled water) and pH 7.4. For solid medium, 20 g agar (Fisher)
were added per liter of medium. In growth curve studies, the above medium was diluted
to one-half nutrient strength and the desired salinity.
Mycological agar (Difco) prepared with distilled water was used for the growth of
fungi. For phosphatase studies, the yeasts were grown on this medium plus 0.02%
phenolphthalein diphosphate. The yeasts were also grown in a broth with this
formulation.
Test chemicals. Aroclors are commercial PCB formulations containing many isomers.
Two Aroclor formulations, 1242 and 1016, containing 42% chlorine, were examined in
this study. Concentrations (w/v) of PCB's are based on the Aroclor formulation as
received from The Monsanto Company, considered as being 100% PCB.
Sensitivity studies. Bacterial cells for inocula were grown in broth for 18 h at 28 C on a
rotary shaker. The culture was diluted 1:1 with sterile 20%o seawater and 0.1 ml of the
dilution was spread on the agar medium. Yeast cultures were grown for 48 h on
mycological agar and colonies were suspended in distilled water to produce a cell
suspension detectable by sight. The cell suspension was swabbed onto agar and prepared
absorbant paper discs were positioned on the surface of the agar.
MIREX
HEPTACHLOR
1242
1O16
1O16
1242
O.5 mg
O.25 mg O.I mg
FIG. 1. Growth inhibition of an estuarine bacterium by Aroclor 1242 and 1016 on solid marine medium.
Mirex and heptachlor (0.5 mg/disc), chlorinated hydrocarbon insecticides were included for screening
purposes only. Bacterial growth appears white on dark medium and the zones of inhibition appear
dark surrounding the white disc due to the negative reproduction of the photographic plate. The
culture dish served as the negative, placed directly on the photographic paper.
199
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258 A- w- BOURQUIN AND L. A. KIEFER
The absorbant paper discs (Schleicher and Schuell, Inc., No. 7-10-E) were saturated
with 0.1 ml of acetone solution containing 1.0, 2.5, and 5.0 mg/ml of PCB formulation.
The discs were air-dried for 24 h at room temperature before use. PCB-treated discs and
control discs treated only with acetone were placed on the agar plates within 2 h of
inoculation. All tests were performed in duplicate and examined for possible inhibition
after 24-48 h incubation at 24-28 C. Only isolates sensitive to 0.5 mg of either PCB
formulation were tested further for response to PCB's. Sensitivity was defined as a zone
of inhibition surrounding the paper disc (Fig. 1).
Phosphatase studies. To test the effects of PCB's on extracellular phosphatases, yeasts
were grown in liquid medium and on PCB-impregnated membranes placed on agar plates.
After 2-24 h incubation, the membranes were removed and the plates were exposed to
vapors of concentrated NH4OH. Occurrence of reddish zones on the plates demonstrated
phosphatase activity.
Growth curve studies. Cells for inocula were grown overnight in liquid medium and
inoculated into 50-100 ml of the same medium in a 500-ml Nephelo-culture flask (Bellco,
515-A) to make a final cell concentration of 1%. Cell density was monitored as
absorbance using a Bausch & Lomb Spectronic 20 at 660 ran or a Klett-Summerson
Photoelectric Colorimeter with a No. 66 red filter. Test chemicals were added in acetone
to facilitate dispersion and medium plus acetone cultures were monitored as checks for
acetone effects. Salinities were adjusted with distilled water prior to addition of nutrient
and the pH was adjusted to 7.4.
RESULTS AND DISCUSSION
Of 106 bacterial isolates, growth of 28 was inhibited in varying degrees by the PCB
formulations. Sensitive bacteria reacted similarly to both Aroclor formulations (Table 1).
The PCB-sensitive bacteria included both gram-positive and gram-negative isolates. Of all
strains tested, a slightly greater percentage of the sensitive bacteria were gram-positive
(Table 2). These results differ from previous reports of sensitivity of gram-positive
bacteria to other chlorinated compounds (Trudgill et al. 1971). Cyclodiene insecticides,
shown to inhibit a range of gram-positive bacteria, had no effect on gram-negative
bacteria tested (Widdus et al. 1971; Trudgill et al. 1971). Differences in toxicity of PCB's
and cyclodiene pesticides to gram-negative bacteria may be related to type of molecule
rather than to degree of chlorination.
Biochemical activities of sensitive and nonsensitive bacteria are compared in Table 2.
The majority of sensitive strains produced both amylase (75%) and gelatinase (89%),
whereas of all strains tested, only 37% were amylase-producers and only 45% were
gelatinase-producers. The significance of these results in relation to total nutrient
catabolism must await further investigation.
Fig. 2 shows the effect of Aroclor 1242 on the growth of four estuarine bacteria in
liquid culture. Two bacteria were completely inhibited for 18 h. Since the PCB's were
added in acetone solution, we believe that after volatilization (or degradation) of the
acetone, PCB's were adsorbed to the cells and glass, allowing cells with no adsorbed PCB
to attain logarithmic growth after 18-20 h incubation (not shown). However, in nature, if
the PCB's were adsorbed to the microbial substrate at inhibitory concentrations, no
growth would occur.
200
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CONTRIBUTED PAPERS
259
TABLE 1. Inhibition of growth of estuarine bacteria on nutrient seawater medium by PCB's
GBERL
Culture
No.
3
21
35
39
53
54
7
9
31
60
86
100
8
11
42
44
93
43
5
13
28
32
41
67
69
Gram
Reaction &
Morphology
+ ROD
-ROD
-ROD
-COCCOID
-ROD
+ ROD
+ ROD
+ ROD
-ROD
+ ROD
-ROD
-ROD
+ ROD
-ROD
+ COCCUS
+ COCCUS
+ ROD
+ COCCUS
-COCCOID
-ROD
-ROD
+ ROD
- COCCOID
-ROD
-ROD
Aroclor® 1242 (mg) Aroclor® 1016 (mg)
Genus 0.1 0.25 0.5 0.1 0.25 0.5
Unknown ++ ++ +++ ++ ++ +++
Unknown ++ ++ +++ ++ ++ +++
Flavobacterium sp. ++ ++ ++ + +++ +++
Unknown ++ +++ +++ ++ +++ +++
Unknown ++ +++ +++ +++ +++ +++
Bacillus sp. +++ +++ +++ + +++ +++
Bacillus sp. + + + x + ++
Bacillus sp. + ++ +++ + ++ +++
Unknown + + +++ + + +++
Bacillus sp. + + + ++++++
Flavobacterium sp. + + + x + +
Pseudomonas sp. +++++ + ++ ++
Corynebacterium sp. x + ++ x + ++
Achromobacter sp. x + ++ x + ++
Micrococcus sp. x + ++ x + ++•
Micrococcus sp. x + + — x +
Unknown x + + — x +
Micrococcus sp. — + ++ — + ++
Serratia sp. __++ — — ++
Achromobacter sp. — — ++ — — ++
Achromobacter sp. — — ++ — — ++
Corynebacterium sp. — + + — x +
Unknown — — ++ — — ++
Achromobacter sp. — * + — — +
Unknown — — ++ — — ++
Degree of sensitivity: +++ (18-20 mm zone), ++ (16-18 mm), + (14-16 mm), x (slightly), - (not sensitive).
TABLE 2. Biochemical activities ofPCB-test bacteria (percent of cultures shotuing positive reaction)
Bacteria Tested
Production of
Urease Amylase Lipase Gelatinase
Citrate Gram
Utilization Reaction
Sensitive cultures 25 75 29 89
Total test cultures 14 37 19 45
4
5
43
37
54
40
E .6
c
O
1.4
11 44 M> Jl
FIG. 2. Growth of estuarine bacteria in liquid marine medium
(20%0) containing 10 jUg/ml Aroclor 1242 of cultures
no. 31 (unknown, gram-negative) and no. 60 (Bacillus sp.)
were sensitive to PCB's and cultures no. 12 (unknown,
gram-negative) and 47 (Pseudomonas sp.) were not sensitive.
Average data points given for the latter two bacteria, two
curves (experimental , and control ), are not sig-
nificantly different.
TIMI (houri)
201
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260
A. W. BOURQUIN AND L. A. KIEFER
(FFfCT OF VARYING CONCENTtATIONS OF AROCLOft 1242 ON
GROWTH OF ESTUARINE BACTIRIA
Ciritun no. WO
Vr
f
f
.//
•
•F
TIMI (hourl)
FIG. 3. Dose-response curve showing growth (O.D.) response of estuarine bacterial isolates no. 9 (Bacillus
sp.) and no. 100 (Pseudomonas sp.) to varied concentrations of Aroclor 1242 (0-5 Mg/ml). The PCB
was added in acetone to facilitate dispersion. Check cultures contained marine broth at 20700salinity
only and with acetone (0.5 ml/flask).
Due to the insoluble nature of PCB's, the minimum inhibitory level of PCB
formulation could not be determined by the paper disc method. The lowest inhibitory
concentration of PCB's for two selected sensitive organisms -is demonstrated in Fig. 3.
Growth of the gram-positive isolate (culture No. 9, Bacillus sp.) was inhibited at 1.0
Mg/ml, whereas the gram-negative isolate (culture No. 100, Pseudomonas sp.) was sensitive
at 1.0 Mg/ml and completely inhibited at 5.0 Mg/ml.
Inhibition of growth of fungi (Cladosporium sp., Cephalosporium sp., Saccharomyces
sp., Candida lipolytica, C. subtropicalis, Pichia spartinae, and Kluyveromyces droso-
pkilarum) by PCB's on paper discs was negligible. Although yeasts failed to grow on
membranes completely saturated with PCB's, they grew on membrane areas that were
free of PCB's and demonstrated phosphatase activity. In liquid media with PCB's added in
petroleum ether (which was evaporated), yeast growth and phosphatase activity were
similar to those of controls.
Sensitivity of estuarine bacteria to PCB formulations was greater in liquid than on
solid medium. No inhibition of growth of estuarine fungi or phosphatase activity was
noted except on PCB-saturated membranes.
Inhibition of bacterial growth by PCB formulations adsorbed onto paper discs is a
simple technique which may be used "to detect, in the laboratory, the possibility of
inhibition in the environment. Although PCB's are almost insoluble in water, they are
202
-------
CONTRIBUTED PAPERS 261
readily adsorbed to solid surfaces (Rizwanul et al. 1974). In naturerif the solid surface to
which PCB's are adsorbed is a potential microbial substrate, inhibition of microbial
catabolism could occur. Such inhibition could account for unexplained increases in BOD
of effluents of sewage treatment facilities that receive industrial wastes containing large
amounts of PCB's. The nature of the inhibiting substance in the PCB formulation that
acts to interfere with the rate of nutrient turnover is uncertain.
LITERATURE CITED
Ahmed, M., and D. D. Focht. 1973. Oxidation of polychlorinated biphenyls by Achromobacter pCB.
Bull. Environ. Contam. ToxicoL 10:70-72.
Breed, R. S., E. G. D. Murray, and N. R. Smith. 1957. Sergey's Manual of Determinative Bacteriology.
Williams & Wilkins Co., Baltimore, Md.
Colwell, R. R., and W. J. Wiebe. 1970. "Core" characteristics for use in classifying aerobic, heterotrophic
bacteria by numerical taxonomy. Bull. Georgia A cad. Sci. 28:165-185.
Gustafson, C. G. 1970. PCB "s-prevalent and persist ant. Environ. Sci. Technol. 10:814-819.
Jensen, S., A. G. Johnels, S. Olson, and G. Ottcrlind. 1969. DDT and PCB in marine animals from
Swedish waters. Nature 224:247-250.
Keil, J. E., S. H. Sandifer, C. D. Graber, and L. E. Priester. 1972. DDT and polychlorinated biphenyl
(Aroclor® 1242) effects of uptake on E. coli growth. Water Res. 6:837-841.
Peakall, D. B., and J. L. Lincer. 1970. Polychlorinated biphenyls. Another long-life widespread chemical
in the environment. BioScience 20:958-964.
Rizwanul, H., D. W. Schmedding, and N. H. Freed. 1974. Aqueous solubility, adsorption, and vapor
behavior of polychlorinated biphenyl Aroclor 1254. Environ. Sci. Technol. 8:139-142.
Trudgill, P. W., R. Widdus, and J. S. Rees. 1971. Effects of organochlorine insecticides on bacterial
growth, respiration and viability. /. Gen. Microbial. 69:1-13.
Widdus, R., P. W. Trudgill, and D. C. Turnell. 1971. The effects of technical chlordane on growth and
energy metabolism of Streptococcus faecalis and Mycobacterium phlei: a comparison with Bacillus
subtilis.J. Gen. Microbiol. 69:23-31.
203
-------
Reprinted from Limnology and
Oceanography, Vol. 20(4):
644-646, 1975, with permission
of the American Society of Lim-
nology and Oceanography, Inc.
DENSITIES OF BACTERIA AND FUNGI IN COASTAL SURFACE FILMS AS
DETERMINED BY A MEMBRANE-ABSORPTION PROCEDURE
S.A. Crow, D.G. Ahearn, W.L. Cook, and A.W. Bourquin
Contribution No. 232
205
-------
DENSITIES OF BACTERIA AND FUNGI IN
COASTAL SURFACE FILMS AS DETERMINED BY A
MEMBRANE-ADSORPTION PROCEDURE
BY S. A. CROW, D. G. AHEARN, W. L. COOK AND A. W. BOURQUIN
Reprinted from LIMNOLOGY AND OCEANOGRAPHY
Vol. 20, No. 4, July 1975
pp. 644-646
Made in the United States of America
207
-------
644
Notes
Densities of bacteria and fungi in coastal surface films
as determined by a membrane-adsorption procedure1'2
Abstract—A membrane-adsorption tech-
nique for counting surface slick microbial
populations was evaluated. The simple pro-
cedure gave bacterial and fungal populations
several orders of magnitude greater than those
previously reported for surface slicks.
Ewing (1950. p. 161) noted that "slicks
or calm streaks on rippled seas are often
seen on coastal waters and lakes when the
wind is light." Surface slicks frequently as-
sociated with biologically productive wa-
ters are due to the ripple-damping action
of naturally occurring organic surface films.
Many investigators have reported bacterial
populations in such films several times
greater than those at a depth of a few centi-
meters (ZoBell 1946; Gunkel 1973; Parsons
and Takahashi 1973). To collect slicks,
various sampling procedures, including
wire screen (Garrett 1965), bubble collec-
tion (Bezdek and Carlucci 1972), and hy-
drophilic drums (Harvey 1966), have been
used. With these techniques there is diffi-
culty in collecting only the surface films
and, therefore, in determining the densities
of surface film microbial populations. We
describe a simple, rapid method for count-
ing microorganisms in surface slicks.
Surface slicks and microlayers were ad-
sorbed on sterile Nuclepore membranes
(47-mm diam, 0.4-/u,m pore size, Nuclepore
Corp.) floated on the water surface. The
membranes, with adhering surface film,
were retrieved by submerging sterile plas-
tic dishes under them and gently removing
the filter and the underlying water; the
membrane was then removed from the dish
with forceps. In calm water with little
wind we could often retrieve filters di-
rectly with forceps. Other membranes
(cellulose, cellulose acetate ester) were
not suitable because they sank when com-
1 Supported in part by Office of Naval Research
contract OXR XOOO-14-71-C-0145 and Environ-
mental Protection Agency contract R 803141-01-0.
= Gulf Breeze Environmental Research Labora-
tory Contribution Xo. 232.
pletely wetted. The membranes were
placed in 100 ml of sterile seawater and
transported to the laboratory, usually
within 20 min. The bottles were agitated
vigorously on a wrist-action shaker for 3
min and the water was diluted serially.
One-tenth milliliter of the required dilution
was inoculated onto the appropriate me-
dium. For counting fungi, membranes
often could be placed directly on nutrient
medium.
Subsurface samples (10-cm depth) were
also taken at each site with sterile 30-ml
disposable syringes fitted with sterile ex-
tension tubes. These samples were ex-
pelled into sterile containers for microbio-
logical analysis.
Marine agar 2216 (Difco) was used for
bacteria and mycological agar (Difco) pre-
pared with seawater and adjusted to pH
4.5 for fungi. All media were incubated at
20-25°C until colonies became visible.
The estuarine regions selected for field
study were a sheltered cove (site 1) adja-
cent to the Environmental Protection
Agency Laboratory at Sabine Island (Es-
cambia County, Florida) and a small salt-
water pond (site 2) that receives waste
water from the laboratory complex; and in
Louisiana (in the Barataria Bay estuarine
system) at Airplane Lake (site 3), a shel-
tered saline bay surrounded by Spartina
marsh, and Bayou Fer Blanc (site 4), a
shallow bayou bounded by Spartina-domi-
nated marsh.
Microbial populations of the surface
layer were at least 100 times greater than
those in waters from 10 cm (Table 1).
Membranes floated on subsurface waters
placed in containers gave populations simi-
lar to those obtained by direct dilution.
The sample volume of the Nuclepore mem-
brane as determined gravimetrically with
water of similar salinity was about 5.9 /zl.
Total pore volume within the membrane
calculated from standard physical param-
eters supplied by the Nuclepore literature
was smaller (2.8 /JL\), suggesting that some
208
-------
Notes
645
Table 1. Densities of microorganisms in surface
layers and subsurface marine waters.
Samples '
Site 1
bacteria
yeasts
molds
Surface
ml"1
io5-io8 :
104
103-104
layer
cm-2
^
2
0.2-2
Subsurfacet
ml'1
102-106
<102
<102
Site 2
bacteria
yeasts
molds 103-104
105-107
103-104
101-103
0.2-2
0.2-2
10
106-107
10
102-103 <104
Site 3
bacteria
fungi
Site 4
bacteria 107-108 103- >104
fungi 104-105 2-28
"l and 2 Sabine Island, 6 samples each;
3 and 4 = Barataria Bay, 2 samples each.
t 10-cm depth.
water adhered to the surface of the mem-
branes as well as filling the pore spaces.
Concentrations of microorganisms were
calculated on a per milliliter basis, using
the sampling volume of 5.9 p\, and on a
surface area basis, using the filter area of
17.3 cm2.
The number of microorganisms adsorbed
to the Nuclepore membrane was greater
than those for open-ocean surface films col-
lected by the wire screen method: Sie-
burth (1965) found as many as 103 organ-
isms ml'1 and Gunkel (1973) up to 9.3 X
105 liter1. The most common bacteria of
surface films collected with membranes
were nonchromogenic, motile, gram-nega-
tive rods. In contrast to bacterial isolates
from the subsurface water, a large number
of surface isolates were able to grow on a
freshwater medium. In addition, fungi
were markedly more numerous than previ-
ously noted for inshore marine waters by
Ahearn and Meyers (1972), who did not
sample the surface films selectively but
collected samples with a bottle at the air-
water interface. The populations we re-
port here per unit area are probably mini-
mal, since the surface slicks appeared dis-
continuous as seen by patches of sheen on
the membranes.
S. A. Crow
D. G. Ahearn
W. L. Cook
Department of Biology
Georgia State University
Atlanta 30303
A. W. Bourquin
United States Environmental Protection
Agency
Gulf Breeze Environmental Research
Laboratory
Gulf Breeze, Florida 32561
and
National Environmental Research Center
Corvallis, Oregon 97330
References
AHEARN, D. G., AND S. P. MEYERS. 1972. The
role of fungi in the decomposition of hydro-
carbons in the marine environment, p. 12-19,
In A. H. Walters and E. H. Haeck van Plas
[eds.], Biodeterioration of materials, v. 2.
Wiley.
BEZDEK, H. F., AND A. F. CARLUCCI. 1972.
Surface concentrations of marine bacteria.
Limnol. Oceanogr. 17: 566-569.
EWING, G. 1950. Slicks, surface films and inter-
nal waves. J. Mar. Res. 9: 161-187.
GARHETT, W. D. 1965. CoUection of slick-
forming materials from the sea surface. Lim-
nol. Oceanogr. 10: 602-605.
GUNKEL, W. 1973. Distribution and abundance
of oil-oxidizing bacteria in the North Sea, p.
127-139. In D. G. Ahearn and S. P. Meyers
[eds.], The microbial degradation of oil pollu-
tants. Center Wetland Resour., Louisiana
State Univ. Publ. LSU-SG-73-01.
HARVEY, G. W. 1966. Microlayer collection
209
-------
646 Notes
horn the sea surface: A new method and ini- Eng. Mar. Technol. Soc. Am. Soc. Limnol.
tial results. Limnol.Oceanogr.il: 608-613. Oceanogr., p. 1064-1067.
PARSONS, T. R., AND M. TAKAHASHI. 1973. Bio- ZoBELL, C. E. 1946. Marine microbiology.
logical oceanographjc processes. Pergamon. Chronica Botanica.
SIEBURTH, J. McN. 1965. Bacteriological sam-
piers for air-water and water-sediment inter- Submitted: 18 September 1974
faces. Trans. Jt. Conf. Ocean Sci. Ocean Accepted: 20 February 1975
210
-------
Reprinted from Journal of
Toxicology and Environmental
Health, Vol. 1: 485-494, 1976
with permission of the
Hemisphere Publishing Corp.,
Washington
CHLORDANE: EFFECTS ON SEVERAL ESTUARINE ORGANISMS
Patrick R. Parrish, Steven C. Schimmel, David J. Hansen,
James M. Patrick, Jr., and Jerrold Forester
Contribution No. 234
211
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CHLORDANE: EFFECTS ON SEVERAL
ESTUARINE ORGANISMS
Patrick R. Parrish, Steven C. Schimmel, David J. Hansen,
James M. Patrick, Jr., Jerrold Forester
U. S. Environmental Protection Agency, Gulf Breeze Environmental
Research Laboratory, Sabine Island, Gulf Breeze, Florida
(Associate Laboratory of the National Environmental Research
Center, Corvallis, Oregon)
Dynamic marine toxicity tests were performed with technical grade chlordane and
eastern oysters (Crassostrea virginicaj, pink shrimp (Penaeus duorarumj, grass shrimp
fPalaemonetes pugioj, sheepshead minnows (Cyprinodon variegatusj, and pinfish
fLagodon rhomboides/ The 96-hr LC50s (and 95% confidence limits) based on
measured concentrations of chlordane (in tig/liter) are: pink shrimp, 0.4 (0.3-0.6);
grass shrimp, 4.8 (4.0-6.0); sheepshead minnows, 24.5 (19.9-28.6); and pinfish, 6.4
(5.0-7.3). The 96-hr ECi0 for eastern oysters was 6.2 (4.8-7.9). In a flow-through test,
embryos and fry of sheepshead minnows were exposed to average measured
concentrations of chlordane from 1.3 to 36.0 tig/liter, for 28 days. Neith'er fertilization
success nor embryo survival was affected by the concentrations of chlordane to which
these life stages were exposed. However, sheepshead minnow fry did not survive for
more than 10 days in chlordane concentrations greater than 7.1 ^g/liter.
INTRODUCTION
Chlordane, a persistent organochlorine pesticide, is used primarily as a
soil insecticide. Approximately 50% of the chlordane used in the United
States is for structural termite protection; about 30% is used for pest
control in agricultural applications (DHEW, 1969).
Like other organochlorine pesticides not originally intended to be
dispersed into aquatic environments, chlordane has been found in major
river basins, the Great Lakes, and estuaries of the United States. For
example, Henderson et al. (1969) found chlordane residues in 22% of
nearly 600 composite fish samples collected from 50 sites in the Great
Lakes and certain major river basins. Bugg et al. (1967) found chlordane
in oysters (Crassostrea virginica] from estuaries of five South Atlantic and
Gulf of Mexico states.
This paper is contribution no. 234, Gulf Breeze Environmental Research Laboratory.
Patrick R. Parrish's present address is Bionomics—EG&G, Inc., Marine Research Laboratory,
Route 6, Box 1002, Pensacola, Florida 32507. Requests for reprints should be sent to this address.
485
Journal of Toxicology and Environmental Health, 1:485-494,1976
Copyright © 1976 by Hemisphere Publishing Corporation
213
-------
486 P. R.PARR1SH ETAL.
Preliminary bioassays conducted at this laboratory showed chlordane
to be acutely toxic to several estuarine animals (Butler, 1963). For
example, the calculated 48-hr LCSO (the concentration lethal to 50% of
the test animals) for adult brown shrimp (Penaeus aztecus] exposed to
chordane in flowing sea water was 4.4 Mg chlordane/liter seawater. Several
of the organochlorines, including chlordane, are currently being considered
by the U.S. Environmental Protection Agency for re-registration. For these
reasons we began this definitive study of the effects of chlordane on
estuarine animals.
This study was conducted to determine the acute (96-hr) toxicity of
technical grade (99.9%) chlordane to eastern oysters (Crassostrea virginica],
pink shrimp (Penaeus duorarum], grass shrimp (Palaemonetes pugio],
sheepshead minnows (Cyprinodon variegatus), and pinfish (Lagodon
rhomboides] and the effect of chlordane on fertility, hatching success, and
survival of sheepshead minnow fry.
Effects of chlordane were assessed by measuring reduction of shell
deposition of oysters (Butler, 1965; Butler et al., 1960) and by
determining mortality in shrimps and fishes from acute toxicity tests,
fertilization and hatching success, and survival of embryos and fry of
sheepshead minnows.
MATERIALS AND METHODS
Test Animals
All test animals except pink shrimp were collected near the Gulf
Breeze Environmental Research Laboratory and acclimated to laboratory
conditions for at least 10 days before exposure. Pink shrimp were
purchased from a local bait dealer and acclimated similarly. Mortality of
animals did not exceed 1% in the 48 hr immediately preceding the test,
nor was any abnormal behavior observed during the acclimation period.
Oysters tested were 29-53 mm umbo to distal valve edge in size;
pink shrimp, 50-65 mm rostrum-telson length; grass shrimp, 20-29 mm
rostrum-telson length; sheepshead minnows, 19-27 mm standard length;
and pinfish, 34-62 mm standard length. Animals were not fed during
acute toxicity tests, but they could obtain food (plankton and other
paniculate matter) from the unfiltered seawater in which they were
maintained. Adult sheepshead minnows over 30-mm standard length were
used to produce eggs used in the fertility, hatching success, and survival
study. An automatic feeder (Schimmel an'd Hansen, 1975) was used to
feed live brine shrimp (Artemia salina] nauplii ad libitum to the fry six
times daily. The eggs from which the nauplii were hatched contained no
organochlorine or polychlorinated biphenyl contaminants detectable by gas
chromatographic analysis.
214
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CHLORDANE EFFECTS ON ESTUARINE ORGANISMS 487
Acute (96-hr) Test Conditions
Acute toxicity of chlordane was determined by exposing 10 animals
per aquarium to different concentrations for 96 hr. Two 20-liter aquaria
were used for each concentration. Unfiltered seawater was pumped from
Santa Rosa Sound, Florida, into a constant-head trough in the laboratory.
Seawater was delivered to a mixing box by a siphon calibrated to deliver
150 liter/hr. Technical grade chlordane (99.9%), dissolved in reagent grade
acetone, was metered at the rate of 30 ml/hr into water entering the
mixing box and was split .equally to two replicate aquaria for each
treatment (Lowe et al., 1972). Two control aquaria received the same
quantities of water and solvent but no chlordane.
Sheepshead Minnow Embryo and Fry Test Conditions
The exposure apparatus used in the sheepshead minnow embryo and
fry test was that described by Schimmel et al. (1975), except that the
toxicant and carrier injectors were those described by Mount and Warner
(1965). A stock solution of chlordane, dissolved in polyethylene glycol
(average molecular weight 200), was prepared and delivered at appropriate
rates to give 100, 46, 21, 10, and 4.6 Mg/liter of chlordane in the test
aquaria. Polyethylene glycol was present at the same concentration (0.9
ml/liter) in all aquaria that received chlordane. One control aquarium
received the same concentration of polyethylene glycol but no chlordane;
a second control received seawater only.
Seawater used in this bioassay was pumped from Santa Rosa Sound
through a sand filter, then through a 5-jum pore polypropylene filter and
into a constant-head box in the laboratory where it was heated to
30 ± 1°C. Salinity was that of Sound water, averaging 17.4 ppt (range was
8.0-28.5 ppt). The water was then pumped to the toxicant delivery
apparatus. Our diluter cycled approximately 80 times each day, delivering
125 ml water (and appropriate amounts of chlordane/polyethylene glycol)
to each of four 1.2-liter exposure chambers per concentration per cycle.
For the embryo-fry test, eggs of C. variegatus were obtained and
fertilized using procedures described by Schimmel et al. (1974). Twenty
embryos were placed in 10-cm Petri dishes to which a 9-cm high screen
collar was - attached. This collar permitted water exchange while
preventing escape of fry. The exposure began 1 hr after microscopic
examination confirmed fertility and lasted 28 days. Dissolved oxygen
concentrations, determined weekly by the modified Winkler method of
Strickland and Parsons (1968), were above 50% of saturation and
appeared adequate.
To determine effects on fertility, eggs from six females were pooled,
and 20 eggs were placed in each of seven Petri dishes. The dishes
contained water from the embryo and fry exposure aquaria. One milliliter
of milt and seawater, pooled from eight males, was deposited in each Petri
215
-------
488 P. R.PARRISH ETAL.
dish and incubated at 30°C for 24 hr. At the end of this incubation
period, eggs were microscopically examined for fertility.
Chemical Analyses
Concentrations of chlordane in water and animals were determined by
electron-capture gas chromatography. Unfiltered water samples from each
concentration were analyzed once during the 96-hr exposures and weekly
during the sheepshead minnow embryo and fry test. Concentrations of
chlordane in animals that survived the 96-hr embryo and fry exposures
were determined as whole-body residues.
Tissue samples that weighed more than 5 g were prepared for analysis
by mixing with anhydrous sodium sulfate in a blender. The mixture was
extracted for 4 hr with petroleum ether in a Soxhlet apparatus. Extracts
were concentrated by evaporation on a steam bath in a Kuderna-Danish
flask to approximately 10 ml and transferred in 3- to 4-ml portions to a
400 X 20-mm chromatographic column that contained 76 mm of
unactivated Florisil. After each portion settled in the column, vacuum was
applied until all solvent was evaporated. This was repeated with three 5-ml
petroleum ether rinses. The residue was eluted from the column with 70
ml of a 9:1 mixture (v/v) of acetonitrile and distilled water. The eluate
was evaporated to dryness and the residue transferred to a Florisil column
(Mills et al., 1963) with petroleum ether. Chlordane was eluted in the 6%
ethyl ether-in-petroleum ether fraction.
Tissue samples that weighed less than 1 g were analyzed by the
micromethod described in the Pesticide Analytical Manual, vol. Ill (FDA,
1970).
Water samples were extracted with petroleum ether; the extracts were
dried with anhydrous sodium sulfate and evaporated to approximately 1
ml. The concentrates were transferred to a size 7 Chromaflex1 column
containing 1.6 g Florisil topped with 1.6 g anhydrous sodium sulfate.
Chlordane was eluted with 20 ml of 1% ether-in-hexane, and the eluates
were adjusted to an appropriate volume for analysis.
All samples were analyzed by electron-capture gas chromatography
using a 182-cm X 2-mm id glass column packed with 2% OV-101 on
100-120 mesh Gas Chrom Q. The nitrogen flow rate was 25 ml/min, the
oven temperature was 190°C, and the injector and detector temperature
was 210°C. Recovery of chlordane from fortified water and tissue samples
exceeded 85%; data were not adjusted for recovery. All tissue residues
were determined on a wet-weight basis.
This multiple peak compound was quantitated by comparing the total
peak heights of the samples with the total peak heights of a standard of
known concentration.
1 Mention of commercial products or trade names does not constitute endorsement by the U.S.
Environmental Protection Agency.
216
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CHLORDANE EFFECTS ON ESTUARINE ORGANISMS 489
Statistical Analyses
Data from the acute (96-hr) exposures were analyzed statistically.
Oyster shell deposition data were analyzed by linear regression (with
probit transformation) to determine an EC50 (the concentration of
chlordane effective in reducing shell deposition of exposed oysters,
compared with controls, by 50%) and 95% confidence limits. Shrimp and
fish mortality data were analyzed by maximum likelihood probit analysis
(Finney, 1971) to determine LC50s (the concentrations lethal to 50% of
the test animals) and 95% confidence limits.
Data from the sheepshead minnow embryo-fry test and fertility test
were analyzed using the chi-square method of determining significant
differences (« = 0.01) between experimental groups and the controls.
RESULTS AND DISCUSSION
Acute (96-hr) Tests
Chlordane was acutely toxic to the estuarine organisms tested (Tables
1 and 2). Shell deposition in oysters was appreciably inhibited by
exposure to measured concentrations >4.7 jug/liter for 96 hr. Pink shrimp
were the most sensitive animal tested, significant numbers dying at
concentrations less than 1 ppb (^g/liter). Grass shrimp died when exposed
to concentrations in the low parts per billion. Pinfish were about four
times more sensitive to chlordane than were sheepshead minnows.
Chlordane appeared to be slightly more toxic to the marine organisms
that we tested than it is to freshwater organisms. Direct comparisons are
difficult because of different test conditions (duration and temperature, in
particular). For freshwater invertebrates, Sanders and Cope (1966) found
that Daphnia pulex were immobilized at a chlordane concentration of 29
Mg/liter (48 hr, 15.6°C). The water flea, Simocephalus serrulatus, exhibited
a similar response at 20 Mg/'iter (48 hr, 15.6°C) and at 24 jug/liter (48 hr,
21.1°C). For freshwater fish, Henderson et al. (1959) found the following
to be the 96-hr TLm (or LCSO) for several freshwater fishes tested under
static conditions at 25°C in soft water: fathead minnows (Pimephales
promelas], 52 Mg/'iter; bluegills (Lepomis macrochirus], 22 Mg/'iter; goldfish
(Carassiusauratus], 82 ptg/liter; and guppies (Lebistesreticulatus], 190 jug/liter.
The acute toxicity of chlordane to marine organisms is similar to the
toxicity of dieldrin, another organochlorine pesticide. For example, Parrish
et al. (1974) found the 96-hr ECSO for eastern oysters exposed to dieldrin
in flowing seawater to be 31.2 Mg/l'ter with 95% confidence limits
between 0.60 and 61.8 //g/liter, based on measured concentrations. The
96-hr LCSO for pink shrimp was 0.7 Mg/liter with 95% confidence limits of
0.39-1.15 Mg/liter; for grass shrimp, 8.6 Mg/liter with 95% confidence
limits of 5.92-12.1 /ig/liter; and for sheepshead minnows, 10 jug/liter (no
confidence limits)—all based on measured concentrations of dieldrin.
217
-------
490 P. R.PARRISH ETAL.
TABLE 1. Toxicity of Chlordane to and Uptake by Selected Estuarine Organisms0
Species
C. virginlca
P. duorarum
P. pug/o
C variegotus
L rhomboides
Water
Nominal
control
4.2
7.5
13.5
24.0
42.0
control
0.075
0.24
0.42
0.75
2.4
control
2.4
4.2
7.5
13.5
24.0
control
28.0
37.0
49.0
65.0
87.0
control
8.5
11.5
15.5
21.0
28.0
concentration
(Mg/liter)
Measured
96-hr tests
ND*
2.2
4.7
8.2
9.6
12.9
ND
0.12
0.17
0.43
0.43
1.73
ND
2.1
2.1
4.2
7.3
17.0
ND
15.0
27.0
28.0
44.0
51.0
ND
5.4
7.3
8.7
9.6
15.2
Effect
(%)
0
8
41
46
76
84
0
5
10
55
60
90
0
0
15
45
70
100
0
25
40
60
95
90
0
30
70
70
85
95
Whole-body
residue
(Mg/g wet weight)
0.09
11.0
27.0
68.0
31.0
69.0
ND
0.49
0.71
1.7
2.6
—
ND
4.8
4.5
9.1
13.7
-
0.6
281.0
33'AO
405.0
-
ND
16.6
55.0
61.0
70.0
_
C var/egatus
control
4.6
10.0
21.0
46.0
100.0
28-day test
ND
1.3
3.3
7.1
17.0
36.0
0
0
3.7
100
100
ND
11.0
34.0
87.0
Effect is expressed as percentage reduction in shell deposition for oysters and death for
shrimps and fishes. Whole-body residues are from animals alive at end of exposure.
ND, not detectable; <0.05 Mg/Mter in water, <0.03 j/g/g in tissue.
218
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CHLORDANE EFFECTS ON ESTUARINE ORGANISMS 491
TABLE 2. Acute Toxicity of Chlordane to Estuarine Organisms
Species
96-hr EC50
(Mg/liter)
Temperature
Salinity
(ppt)
Nominal
Measured
Mean
Range
Mean
Range
C virginica
P. tiuorarum
P. pugio
C. variegatus
L rhomboides
11.6
(8.4-16.0)
0.5
(0.4-0.8)
8.4
(6.9-10.3)
39.8
(34.0-45.2)
10.4
(7.5-12.4)
6.2
(4.8-7.9)
0.4
(0.3-0.6)
4.8
(4.0-6.0)
24.5
(19.9-28.6)
6.4
(5.0-7.3)
31.6
28.4
30.0
30.7
31.3
31.0-32.0
27:5-30.0
27.5-32.0
29.0-32.0
30.0-32.5
24.3
21.8
22.7
25.0
24.6
20.0-28.0
20.0-25.0
15.0-30.0
1 8.0-28.0
22.0-28.0
"Effect is expressed as percentage reduction in shell deposition for oysters and death for shrimps and
fishes. Confidence limits (95%) are in parentheses.
Chlordane is not as acutely toxic as endrin, however, except for oysters.
Schimmel et al. (1975) found the 96-hr ECSO for oysters exposed to
endrin in flowing seawater to be 14.2 jug/liter, based on measured
concentrations. The 96-hr LCSO for pink shrimp was 0.037 jug/liter; for
grass shrimp, 0.63 jug/liter; and for sheepshead minnows, 0.38 pig/liter—all
based on measured concentrations.
Sheepshead Minnow Fertilization Success
and Embryo and Fry Tests
Fertility of sheepshead minnow eggs was not significantly affected by
the concentrations tested. Fertilization success ranged from 80 to 95% for
all concentrations and controls.
In the embryo and fry test, survival of embryos was not significantly
reduced in any of the concentrations of chlordane tested. Embryo
mortality -ranged from 10 to 24% for all concentrations and controls.
Fry died in significant numbers (Table 1, 28-day test) in the, 46.0 and
100.0 jug/liter experimental aquaria (average measured concentrations of
17.0 and 36.0 pig/liter). No fry survived longer than 10 days after hatching
in these two concentrations. Fry exposed to 21.0 Mg/liter chlordane (7.1
Mg/liter measured concentration) did not suffer significant mortality,
compared with controls, but they lost equilibrium and swam erratically.
Comparison of the sensitivity to chlordane of sheepshead minnow
juveniles in a 4-day test with the sensitivity to chlordane of sheepshead
minnow embryo and fry in a 28-day test is difficult given the differences
219
-------
492 P. R. PARRISH ETAL.
in duration of the tests and the lack of a statistically valid estimate of an
LC50 in the embryo and fry test. However, comparisons of percent
mortality and measured concentration of chlordane in test water in both
tests show that all fry died in average measured concentrations of 17 and
36 Mg/liter; similar measured concentrations between 15 and 28 /zg/liter
killed fewer juveniles (25-60%).
Bioaccumulation
All animals accumulated chlordane; the quantities depended upon the
species and the exposure concentration (Table 3). Grass shrimp
accumulated chlordane least, and sheepshead minnows accumulated more
chlordane than did the other animals tested.
The range of concentration factors for chlordane in sheepshead
minnows (Table 3) exposed for 96 hr in flowing seawater was greater than
that for sheepshead minnows exposed to dieldrin or endrin under similar
conditions. Dieldrin was accumulated 3,500-7,300X (Parrish et al., 1974)
and endrin was accumulated 684-4,545X (Schimmel et al., 1975) the
measured concentrations in test water.
TABLE 3. Range of Chlordane Concentration Factors" from Live
Marine Animals Exposed in Flowing Seawater
Concentration factors
Nominal
Measured
Crassostreo vlrglnica
(eastern oyster)
Penaeus duorarum
(pink shrimp)
Palaemonetes pugio
(grass shrimp)
Cyprinodon variegatus
(sheepshead minnow)
Lagodon rhomboides
(pinfish)
Cyprinodon variegatus
(sh£jpshead minnow)
96-hr tests
1,300-5,000
3,200-8,300
3,000-6,500 4,000-6,000
1,000-2,000
1,900-2,300
8,300-10,000 12,600-18,700
2,000-4,800 3,000-7,500
28-day test
2,400-4,100
8,500-12,300
Concentration factors (concentration in tissue divided by
concentration in test water) were calculated according to both nominal
and measured concentrations in the test water.
220
-------
CHLORDANE EFFECTS ON ESTUARINE ORGANISMS 493
ECOLOGICAL SIGNIFICANCE
Chlordane can be bioaccumulated by several marine animals. Also,
chlordane, like other chlorinated hydrocarbon insecticides, may be
transferred from one trophic level to another.
Chlordane could cause environmental damage because of its sublethal
effects. For example, although sheepshead minnow fry exposed to
measured concentrations of 7.1 Mg/'iter chlordane did not suffer significant
mortality, they lost equilibrium and swam erratically. In a,n estuary, fish
or invertebrates so affected could be subject to increased predation.
REFERENCES
Bugg, J. C., Jr., Higgins, J. E. and Robertson, E. A., Jr. 1967. Chlorinated pesticide levels in the
eastern oyster (Crossostrea -virginica) from selected areas of the South Atlantic and Gulf of
Mexico. Pestic. Monit. J. 1(3):9-12.
Butler, P. A. 1963. Commercial fisheries investigations. In Pesticide-wildlife studies. A review of fish
dnd wildlife service investigations during 1961 and 1962, pp. 11-25. U.S. Fish Wildl. Serv.
Circ. 167.
Butler, P. A. 1965. Reaction of some estuarine mollusks to environmental factors. USPHS Publ. No.
999-P-25, pp. 92-104. Washington, D.C: Department of Health, Education, and Welfare.
Butler, P. A., Wilson, A. J., Jr. and Rick, A. J. 1960. Effect of pesticides on oysters. Proc. Shellfish
Assoc. 51:23-32.
Finney, D. J. 1971. Probit analysis, 3d ed., 33 p. Cambridge: Cambridge University Press.
Henderson, D., Pickering, Q. H. and Tarzwell, C. M. 1959. Relative toxicity of ten chlorinated
hydrocarbon insecticides to four species of fish. Trans. Am. Fish. Soc. 88:23-32.
Henderson, D., Johnson, W. L and Inglis, A. 1969. Organochlorine insecticide residues in fish.
(National Pesticide Monitoring Program) Pestic. Monit. J. 3(3):145-171.
Lowe, J. I., Parrish, P. R., Patrick, J. M., Jr. and Forester, J. 1972. Effects of the polychlorinated
biphenyl Aroclor® 1254 on the American oyster, Crassostrea virginica. Mar. Biol. (Berlin)
17:209-214.
Mills, P. A., Onley, J. F. and Gaither, R. A. 1963. Rapid method for chlorinated pesticide residues
in non-fatty foods. J. Assoc. Off. Agric. Chem. 46(2):186-191.
Mount, D. I. and Warner, R. E. 1965. A serial dilution apparatus for continuous delivery of various
concentrations of materials in water. USPHS Publ. No. 999-WP-23. 16 pp.
Parrish, P. R., Couch, J. A., Forester, J., Patrick, J. M., Jr. and Cook, G. H. 1974. Dieldrin: Effects
on several estuarine organisms. Proc. Annu. Conf. Southeast Assoc. Gome Fish Comm., pp.
427-434.
Sanders, H. O. and Cope, O. B. 1966. Toxicities of several pesticides to two species of cladocerans.
Trans. Am. Fish. Soc. 95:165-169.
Schimmel, S. C. and Hansen, D. J. 1975. An automatic brine shrimp feeder for aquatic bioassays. /.
Fish. Res. Board Can. 32(2):314-316.
Schimmel, S. C., Hansen, D. J. and Forester, J. 1974. Effects of Aroclor® 1254 on laboratory-reared
embryos and fry of sheepshead minnows (Cyprinodon variegatus). Trans. Am. Fish. Soc.
103(3):582-586.
Schimmel, S. C., Parrish, P. R., Hansen, D. J., Patrick, J. M., Jr. and Forester, J. 1975. Endrin:
Effects on several estuarine organisms. Proc. Annu. Conf. Southeast. Game Fish Comm. In
press.
Strickland, J. D. H. and Parsons, R. R. 1968. A practical handbook of seawater analysis. F/5/7. Res.
Board Can. Bull. 167:21-26.
221
-------
494 P. R.PARRISH ETAL.
U.S. Department of Health, Education, and Welfare. 1969. Report of the secretary's commission on
pesticides and their relationship to environmental health. 677 pp. Washington, D.C
U.S. Food and Drug Administration. 1970. Pesticide analytical manual, Sec. H212. Washington,
D.C.: Department of Health, Education, and Welfare.
Received April 23, 1975
Accepted May 30, 1975
222
-------
Reprinted from Pesticide
Biochemistry and Physiology,
Vol. 5(6): 536-542, 1975,
with permission of the
Academic Press Inc.,
New York, San Francisco,
London
BRAIN ACETYLCHOLINESTERASE INHIBITION IN FISH AS A DIAGNOSIS OF
ENVIRONMENTAL POISONING BY MALATHION, O,O-DIMETHYL S-(l,2,-DicAR-
BETHOXYETHYL) PHOSPHORODITHIOATE
David L. Coppage, Edward Matthews, Gary H. Cook, and Johnnie Knight
Contribution No. 237
223
-------
Brain Acetylcholinesterase Inhibition in Fish as a Diagnosis
of Environmental Poisoning by Malathion, O,O-Dimethyl
S-(l,2-Dicarbethoxyethyl) Phosphorodithioate1
DAVID L. COPPAGE, EDWARD MATTHEWS, GARY H. COOK, AND JOHNNIE KNIGHT
U. S. Environmental Protection Agency, Gulf Breeze Environmental Research Laboratory, Sabine Island,
Gulf Breeze, Florida 32661 (Associate Laboratory of the National Environmental
Research Center, Corvallis, Oregon)
Received January 20, 1975; accepted April 23, 1975
Brain acetylcholinestrase (EC 3.1.1.7) activities were compared in groups of an
estuarine fish Lagodon rhomboides (pinfish) exposed in sea-water to sublethal and lethal
concentrations of malathion (0,0-dimethyl jS-(l,2-dicarbethoxyethyl) phosphorodi-
thioate) to determine enzyme inhibition values for diagnosis of poisoning. Lethal ex-
posures caused greater enzyme inhibition than sublethal exposures through 72 h. Con-
sistent levels of enzyme inhibition (72-79% inhibition) occurred when 40-60% of
replicate exposed groups were killed at 3.5, 24, 48 and 72 h at mean concentrations of 575,
142, 92 and 58 Mg/liter, respectively. A mean concentration of 31 /jg/liter was sublethal
through 72 h exposure and caused a maximum enzyme inhibition of only 34%. The
correlation of brain acetylcholinesterase inhibition with exposure and deaths is of value
in diagnosing poisoning in fish populations and has been applied to actual environmental
situations. Enzyme inhibition in fishes is positively correlated with spraying of an estuary
with malathion.
INTRODUCTION
Malathion, 0,0-dimethyl $-(l,2-dicarb-
ethoxyethyl) phosphorodithioate, is prob-
ably the most widely used organophosphate
insecticide in the United States, and it
presents problems for determining effects
on nontarget species. Production of mal-
athion in the United States in 1971 was
estimated to be more than 1 X 107 kg (1)
and its wide use provides many occasions
for entry into the aquatic environment. In
addition to possibly entering waters from
surface runoff (2-4), malathion is applied
directly to inland and coastal marshes for
mosquito control (5-8). Although residues
of malathion in water have rarely been
investigated, recent studies in Texas have
1 Contribution Number 237, Gulf Breeze En-
vironmental Research Laboratory.
shown residues ranging from 0.08 to 500
/zg/liter (2, 7, 9). However, presence or
absence of malathion residues in water or
animals cannot necessarily establish or
eliminate possibility of poisoning of aquatic
animals without continuous knowledge of
sources, distribution, physicochemical in-
teractions, frequency, and duration of
residues. It is complicated because the
ultimate effects of malathion require that it
be converted to another compound that
may not be detectable by chemical analysis
of environmental samples and it may
have biological effects after malathion
disappears.
The mode of action of organophosphate
insecticides in vertebrates is generally re-
garded as disruption of nerve impulse
transmission in the central and peripheral
nervous systems by inhibition of acetyl-
Copyright © 1975 by Academic Press Inc.
All rights of reproduction in any form reser
225
-------
ACETYLCHOLINESTERASE INHIBITION IN FISH BY MALATHION
537
cholinesterase (EC 3.1.1.7), the enzyme
that modulates the amounts of the neuro-
transmitter acetylcholine (10-19). It has
been shown with radioactively labeled
organophosphate anticholinesterase agents
in vitro, that the actual toxic agents are
deacylated metabolites which "irrevers-
ibly" phosphorylate 0-serine in the ester-
atic site of "purified" cholinesterase from
several sources (including the teleost fish,
Electrophorus electricus) and that further
alteration of the organophosphate may
occur by dealkylation (10, 12-14, 16-20).
The covalent phosphorus-serine bond is
maintained long after parent compound
has disappeared. Additional exposure would
increase the number of these bonds and
enzyme inhibition. However, organophos-
phates applied to the environment have no
radioactive label to aid in detecting the
enzyme bound forms and, because of tech-
nical problems, it is unlikely that a method
of extraction and chemical analysis of the
enzyme bound forms from environmental
samples will be developed in the foresee-
able future. Thus, acetylcholinesterase mea-
surements hi animals from the environ-
ment are probably the best direct measure
of poisoning when a complete history of
residues is lacking.
There is strong evidence that malathion
is metabolically altered before it inhibits
acetylcholinesterase in vivo in fishes.
Murphy (21) and Murphy et al. (22) have
shown that malathion has little or no direct
capacity to inhibit acetylcholinesterase but
is converted to active inhibitor in the fish
liver in vitro. This active inhibitor is
believed to be malaoxon, the oxygen analog
of malathion created by desulfuration
(P = S —» P = O), which reacts with ace-
tylcholinesterase to form dimethyl phos-
phorylated enzyme:
CH30 S
\ /
P
CH3O SCHCOOC2H6 CH3O SCHCOOC2H5 CH3O Enzyme
CH2COOC2H5 CI
The metabolic conversion of 99.5% pure
malathion resulted in more than a 1000-fold
increase in fish brain acetylcholinesterase
inhibitory potency in vitro (22). It was also
shown that malaoxon is much more toxic
and a more potent brain acetylcholin-
esterase inhibitor in fish than malathion
in vivo (22). It is probably malaoxon or a
related P = O metabolite, rather than
malathion, that is deacylated and phos-
phorylates brain acetylcholinesterase in
fishes. Oxygen, being more electronegative
than sulfur, has a stronger capacity to
withdraw electrons from phosphorus. This
decreased electron density of the phos-
phorus is necessary for rapid reaction of the
agent with an electron dense area in the
active site of the enzyme (13,17,18).
Bender (23) showed carp (Cyprinus carpio)
exposed 96 h to 5 mg malathion/1 water
had residues in flesh of about 28 /ig/g but
only 80 ng/g remained in flesh 96 h after
exposure was discontinued. If this rate of
loss is typical of fishes, malathion would
not be expected to remain in the body for
periods of several weeks. However, brain
acetylcholinesterase remains inhibited in
fishes for several weeks after substantial
inhibition in vivo from malathion exposure
(5, 24-28). A metabolite, not indicated by
presence or absence of parent malathion,
must be responsible for relatively irrevers-
ible acetylcholinesterase inhibition. The
continued inhibition may be due to essen-
tially irreversible phosphorylation of brain
226
-------
538
COPPAWE ET AL.
TABLE 1
Brain-AChE Activity in Pinfish Subjected to Sublethal and Lethal Exposure to Malathion
Concentration (Mg/1)
Nominal
Control
8,000
acetone only
25
25
25
70
125
250
500
Mea-
sured"
—
31
31
31
58
92
142
575
SD Hours
exposed
72
23 24
48
72
14 72
18 48
2 24
30 3.5
Percentage
killed
0
0
0
0
40-60
40-60
40-60
40-60
Number
of
AChE
samples
13
3
3
3
3
3
5
4
3
Mean
AChE
activity"
2.29
2.20"
2.10
2.08
1.50
0.64
0.48
0.64
0.58
SD
0.21
0.30
0.06
0.03
0.00
0.03
0.15
0.06
0.03
Percentage
less AChE
activity
than
control
—
8
9
34<*
72'
79"
72 '
75d
" Chemical analyses for malathion were performed on 2-13 water samples during a test.
6 Acetylcholinesterase (AChE) activity is expressed as ,umol of acetylcholine hydrolyzed/h/mg brain
tissxie.
cNot significantly different from control (P = 0.05).
d Significantly less AChE activity than control (P < 0.001).
' Significantly less activity than control (P < 0.001) and fish exposed to 25 /xg/liter for same period
(P < 0.001).
acetylcholinesterase with very slow syn-
thesis of new enzyme.
Information is needed on the relationship
of acetylcholinesterase inhibition to poison-
ing and deaths of aquatic animals to
diagnose poisoning by malathion when it
is suspected. This information will also be
of value in eliminating malathion as a
cause of poisoning when aquatic animal
"kills" result from other causes in areas
where malathion is used. In this report,
we define levels of reduction of brain
acetylcholinesterase activity that are asso-
ciated with short-term sublethal and lethal
exposures of an estuarine fish Lagodon
rhomboides (pinfish) to malathion.
MATERIALS AND METHODS
Determination of Enzyme Activity
The acetylcholinesterase of the pinfish
brain was characterized and assayed with a
recording pH stat as previously described
(29). Each assay sample for acetylcholin-
esterase consisted of pooled brains taken
from four to six fish. Normal acetylcholin-
esterase activity was determined from 13
samples of unexposed fish (55-100 mm
total length) from the same populations
as fish exposed to malathion. Inhibition
was determined by assay of samples of fish
that survived in replicate aquaria at a
designated time, and percentage of in-
hibition was determined by comparison
with mean normal enzyme activity. Dead
fish were not used because data would not
apply to field studies where it is not known
how long fish have been dead and subject
to loss in enzyme activity due to protein
destructuon.
Test Procedure
Fish were obtained from wild fish popu-
lations and acclimated to laboratory con-
ditions at least 2 wk before testing. In
each test, 3-12 replicates of 10 fish each
were exposed to technical grade malathion
(95% pure) in 8-liter acrylic plastic aquaria
that received a mixture of flowing seawater
227
-------
ACETYLCHOLINESTERASE INHIBITION IN FISH BY MALATHION
539
(400 ml/min) and malathion. The mala-
thion was dissolved in acetone and infused
into seawater by means of a syringe pump.
Solvent infusion never exceeded 2.5 mg/liter
of water. Acetone did not significantly
affect acetylcholinesterase activity of fish
exposed to 8 mg/liter for 72 h (Table 1).
Temperature ranged from 24-29°C and
salinity from 23-29 parts/thousand during
the tests.
To determine the extent of acetyl-
cholinesterase inhibition resulting from a
near-median kill, we assayed the survivors
of tests in which 40-60% of the test popula-
tions were killed by exposure to malathion
in 3.5, 24, 48, and 72 h. Brain acetyl-
cholinestrase activity was measured at 3.5 h
for the 3.5-h lethal exposure (575 ± 30
Mg/liter), 24 h for the 24-h lethal exposure
(142 ± 2 Mg/liter), 24 and 48 h for the 48-h
lethal exposure (92 ± 18 Mg/liter), and at
24, 48, and 72 h for the 72-h lethal exposure
(58 ± 14 Mg/liter). This was accomplished
by exposing several groups of fish, in sepa-
rate aquaria, to the same source of mala-
thion in seawater. At each specified time
interval, three to five replicate groups of
four to six fish each were taken from repli-
cate aquaria and their brain acetylcholin-
estrase was measured. Other groups of pin-
fish were exposed to sublethal malathion
concentration (31 ±23 Mg/liter) for 72 h
and their enzyme activity was measured at
24, 48, and 72 h.
Ckromatographic Analysis of Water for
MalaLhion
One liter water samples were spiked with
methyl parathion as an internal standard
at approximately the same concentration
as expected for malathion and extracted
twice with 100 ml petroleum ether. The
extracts were dried by eluting through
sodium sulfate and concentrated to the
desired volume in a Kuderna-Danish con-
centrator. .Malathion and the recovery of
the internal standard were determined
without further cleanup on a Tracer
MALATHION - PINFISH
24
TIMI EXPOSfP
48
(houri)
FIG. 1. Reduction of brain acetylcholinesterase
activity by sublethal and lethal exposure to malathion.
Each experimental point represents the mean of three
to five replicate tests. The mean measured amounts of
malathion in /j.g/1 for a particular test or test sequence
are shown by the open circles representing 40-60%
deaths and by the closed circle at the end of the sublethal
test.
MT-220 gas chromatograph using a flame
photometric detector operating in the
phosphorus mode. The glass column (182
X 0.32 cm) was packed with. 2% OV-101 on
80/120 mesh Chromsorb Q. The operating
conditions were, temperatures: column
180°C, injector 230°C, and detector
160°C; gas flows: nitrogen 60 ml/min,
hydrogen 200 ml/min, oxygen 20 ml/min,
and air 40 ml/min. Recovery of the internal
standard was greater than 90% for all
samples.
RESULTS AND DISCUSSION
Inhibition data for fish, expressed as
percentage of reduction of enzyme activity
when compared with mean normal activity,
are summarized in Fig. 1. Specific enzyme
activities and statistical comparisons (Stu-
dent's t-test, P < 0.001) are shown in
Table 1.
Lethal exposure always produced a signif-
icantly greater inhibition of enzyme activity
than sublethal exposure (Table 1). Mean
reductions of enzyme activity in lethal
exposures that killed 40-60% of the test
populations were similar (72-79%), at all
228
-------
540
COPPAGE ET AL,
20
o> 40
Z 60
o
£
^ 80
O
Ul
* 100
.
<
• 0.
-
\
I/
"i
; .* Spot • — «
i / Croakers 4 — <
\/ Mullet •
7
*
0 20 40 60
DAYS AFTER PRESPRAY SAMPLE
FIG. 2. Relationship of reduction of brain acetyl-
cholinesterase in three species of fish in a Louisiana
lake to two sprays with malathion.
the selected concentrations and exposure
times. The mean reductions caused by the
sublethal concentration (31 ± 23 jug/liter)
did not exceed 34% in the 72 h period.
These data indicate that brain acetyl-
cholinesterase inhibition in the 70-80%
range is associated with some impending
deaths from short-term exposures of pinfish
populations. Similar critical enzyme inhibi-
tion levels have been found, in less rigorous
tests with malathion and other organo-
phosphate insecticides, in spot (Leiostomus
xanihurus), Atlantic croaker (Micropogon
undulatus) and sheepshead minnows (Cy-
prinodon variegatus) (30, 31). Although
brain acetylcholinesterase is inhibited by ex-
posure to organophosphate pesticides other
than malathion, the level of enzyme inhibi-
tion in live fish during "kills" caused by
metabolites of these agents is relatively
specific. Therefore it is unnecessary to rely
solely on the dubious interpretation of
residues to determine poisoning and cause
of "kills" by anticholinesterase agents in
the aquatic environment. Correlation of
inhibition with malathion usage or presence
of residues and metabolites should be suffi-
cient to establish the cause of poisoning
should it occur in fish. Lack of enzyme
inhibition may exonerate malathion even if
residues are present because fish cannot be
killed by acute exposure to malathion with-
out substantial inhibition of brain acetyl-
cholinesterase.
The measurement of fish brain acetyl-
cholinesterase for diagnosis of anticholin-
esterase poisoning has been applied in
actual environmental situations. Coppage
and Duke (5) found that brain acetyl-
cholinesterase inhibition in fishes in a
Louisiana lake (estuarine because of con-
nection to the Gulf of Mexico by a ship
channel) was correlated with large scale
aerial spraying with malathion (approx.
250 g active ingredient/hectare) for mos-
quito control. The brain acetylcholin-
esterase reductions in three estuarine fishes
(spot, Atlantic croaker, and striped mullet
Mugil cephalus) during two sprays are
shown in Fig. 2. Substantial enzyme
inhibition was found in spot and croakers
relative to prespray levels and "fish kills"
were reported during the spraying period.
Enzyme activity in spot brains remained
significantly below prespray levels more
than 40 days after spraying was discon-
tinued. The mullet were moribund when
collected and mean reduction of enzyme
activity was 97.5%, relative to mullet
from an unsprayed area, which is in agree-
ment with 70-80% enzyme reduction being
critical levels at and below which some
deaths are likely to occur. However, in
another study malathion was applied to
a salt marsh with ground equipment at
rates of 57 and 420 g/hectare, but no
mortality or brain acetylcholinesterase in-
hibition was observed in sheepshead min-
nows in the marsh even though malathion
residues were found in the water (32).
A laboratory study showed spot and
croakers were not more sensitive to mala-
thion than sheepshead minnows (31). Thus,
poisoning in the environment by malathion
may depend on particular circumstances
not readily definable solely in terms of
residues or theoretical amounts of pesticide
applied, and ground applications may give
more suitable control of desired levels of
pesticides than some aerial applications.
229
-------
ACBTYLCHOLINESTERASB INHIBITION IN PISH BY MALATHION
541
Although our studies indicate brain
acetylcholinesterase inhibition of 70-80%
or more is associated with some impending
deaths in fish populations exposed to mala-
thion, lesser inhibition probably has effects.
Cholinergic responses have been demon-
strated pharmacologically in the central
nervous system of fish (33), and depression
of brain acetylcholinesterase in vertebrates
may cause physiological and behavioral
modifications that reduce animal survival
ability (13-16,27,34-38). It has been
shown that short-term reduction of brain
acetylcholinesterase in salmonid fishes by
only 50% during malathion exposure is
associated with pronounced reduction of
their activity index which would probably
reduce their survival ability (27).
We conclude that mechanisms related to
injury of fishes by malathion are quantifi-
able in the environment and should be
measured in additior to chemical residues.
ACKNOWLEDGMENT
We thank Mr. Steven Foss for preparing the
figures.
REFERENCES
1. U. S. Environmental Protection Agency, The
pollution potential in pesticide manufactur-
ing, Pesticide Study Series No. 5, Washing-
ton, DC, 1972.
2. U. S. Environmental Protection Agency, The
use of pesticides in suburban homes and
gardens and their impact on the aquatic
environment, Pesticide Study Series No. 2,
Washington, DC, 1972.
3. U. S. Environmental Protection Agency,
Pesticide usage and its impact on the aquatic
environment in the Southeast, Pesticide
Study Series No. 8, Washington, DC, 1972.
4. U. S. Environmental Protection Agency,
Patterns of pesticide usage and reduction in
use as related to social and enconomic factors.
Pesticide Study Series No. 10, Washington,
DC, 1972.
5. D. L. Coppage and T. W. Duke, Effects of
pesticides in estuaries along the Gulf and
Southeast Atlantic Coasts, in "Proceedings
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(C. H. Schmidt, Ed.), pp. 24-31. National
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Management Coordinating Committee,
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7. G. O. Gtierrant, L. E. Fetzer, Jr., and J. W.
Miles, Pesticide residues in Hale County,
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aerial application of malathion, Pestic.
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14. A. G. Karczmar, Ed. "Anticholinesterase
Agents," Pergamon Press, New York, 1970.
15. A. G. Karczmar, S. Nishi, and L. C. Blaber,
Investigations, particularly by means of
anticholinesterase agents, of the multiple
peripheral and central Cholinergic mechanisms
and of their 'behavioral implications, Acta
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16. G. B. Koelle, Ed. "Cholinesterases and Anti-
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17. R. D. O'Brien, "Toxic Phosphorus Esters,"
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18. R. D. O'Brien, "Insecticides," Academic Press,
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19. R. D. O'Brien, Phosphorylation and carbamyla-
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20. N. K. Schaffer, S. C. May, and W. H. Sum-
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22. S. D. Murphy, R. L. Lauwerys, and K. L.
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24. F. L. Carter, "In vivo studies of brain acetyl-
cholinesterase inhibition by organophosphate
and earbamate insecticides in fish,1' Ph.D.
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Baton Rouge, Louisiana, 1971.
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26. O. B. Cope, Sport fishery investigations, in
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226, Washington, DC, 1965.
27. G. Post and R. A. Leasure, Sublethal effect of
malathion to three salmonid species, Bull.
Environ. Contam. Toxicol. 12, 312 (1974).
28. C. M. Weiss, Physiological effect of organic
phosphorus insecticides on several species of
fish, Trans. Amer. Fish. Soc. 90, 143 (1961).
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acetylcholinesterase with an automated pH
stat for inhibition studies, Bull. Environ.
Contam. Toxicol. 6, 304 (1971).
30. D. L. Coppage, Organophosphate pesticides:
Specific level of brain AChE inhibition related
to death inrsheepshead minnows. Trans. Amer.
Fish. Soc. 101, 534 (1972).
31. D. L. Coppage and E. Matthews, Short-term
effects of organophosphate pesticides on
cholinesterases of estuarine fishes and pink
shrimp, Bull. Environ. Contam. Toxicol. 11,
483 (1974).
32. M. E. Tagatz, P. W. Borthwick, G. H. Cook,
and D. L. Coppage, Effects of ground ap-
plications of malathion on salt-marsh en-
vironments in northwestern Florida, Mosq.
News 34, 309 (1974).
33. G. B. Leslie, J. D. Ireson, and M. L. Tattersall,
Some central actions of a potent muscarinic
agent in lower vertebrates, Comp. Biochem.
Physiol. 31, 571 (1969).
34. P. L. Carl ton, Brain-acetylcholine and inhibi-
tion, in "Reinforcement and Behavior" (J.
T. Tapp, Ed.), pp. 286-327. Academic Press,
New York, 1969.
35. A. J. Deutsch, The cholinergic synapse and
memory, in "The Physiological Basis of
Memory" (A. J. Deutsch, Ed.), pp. 59-76.
Academic Press, New York, 1973.
36. P. H. Glow and A. J. Richardson, Chronic reduc-
tion of cholinesterase and the extinction of an
operant response, Psychopharmacologia (Ber-
lin) 11, 430 (1967).
37. D. L. Margules and A. S. Margules, The de-
velopment of operant responses by nor-
adrenergic activation and cholinergic sup-
pression of movements, in "Efferent Organi-
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(J. D. Masser, Ed.), pp. 203-228. Academic
Press, New York, 1973.
38. A. J. Richardson and P. H. Glow, Post criterion
discrimination behavior in rats with reduced
cholinesterase activity, Psychopharmacologia
(Berlin) 11, 435 (1967).
231
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Reprint from Progressive
Fish-Culturist, Vol. 37
(3): 126-129, 1975
A SALTWATER FLOW-THROUGH BIOASSAY METHOD WITH
CONTROLLED TEMPERATURE AND SALINITY
L.H. Bahner, C.D. Craft, and D.R. Nimmo
Contribution No. 239
233
-------
A SALTWATER FLOW-THROUGH BIOASSAY
METHOD WITH CONTROLLED TEMPERATURE
AND SALINITY
L. H. BAHNER, C. D. CRAFT, and D. R. NIMMO
U. S. Environmental Protection Agency
GidfBrefZt Environmental Research Laboratory
Sabine Island, Gulf Breeze, Florida 32SS1
For several years, researchers at the Gulf
Breeze Environmental Research Laboratory
(GBERL) have been refining techniques for
the flow-through bioassay, a testing method
in which a continuous supply of natural sea-
water flows through experimental tanks. The
flow-through bioassay offers many advan-
tages over static exposure methods. Continu-
ously flowing seawater simulates more
closely the natural estuarine or marine envi-
ronment, substantially reducing problems as-
sociated with static methods such as poor
mixing of toxicants, death of experimental
animals from anoxia, adsorption of toxicant
to sediments and to walls of exposure tanks,
and excess growth of microorganisms.
Temperature and salinity affect bioassays.
For example, in a freshwater study, the com-
bination of the pesticide dieldrin and thermal
changes reduced survival of the darter,
Etheostoma nigntm [£]. Temperature and sa-
linity stress increased mortality in fiddler
crabs, Uea pugilator, exposed to mercury [5].
Similarly, pink shrimp, Penaeus duorarum,
previously exposed to sublethal concentra-
tions of the polychlorinated biphenyl Aroclor*
NOTE.—Contribution No, 239, Gulf Breeze Environmen-
tal Research Laboratory.
The Gulf Breeze Laboratory is an associate laboratory
of the National Environmental Research Center, Corval-
lis, Oreg.
"Ki'jfiftie-rpi! trademark, Mmisanlo Co.. t>l. IsuuJR, Mn. Montiun of comm«'r-
eial iirndurta or trade name* ilora nm ftirtKtifute rndorsemont by the U. S.
Environmental Protection Agency.
126
1254, died when the salinity was gradually
lowered from 20 %o to approximately 12 c'/o° [S .
If results of toxicity tests are to be
confirmed, identical test conditions must be
repeated. It is important, therefore, to control
temperature and salinity in a flow-through
bioassay. Finally, the ability to control tem-
perature and salinity facilitates studies of
toxicant and environmental stress interac-
tions.
Our flow-through system has been used ex-
tensively in bioassays with the pink shrimp,
Penaeus dnorarnm,\.& well as grass shrimp,
Palaemonetes rutgaris and P. pugio. The pink
shrimp is a valuable commercial species and
both pink and grass shrimp are integral parts
of both estuarine and marine food webs. Al-
though the method described here deals with
shrimp, this flow-through bioassay method,
with minor modifications, is readily adaptable
to a wide variety of estuarine and marine
macroinvertebrates and fishes. The cost of
this system for a laboratory with flowing
seawater would be approximately $1,500, an
amount within the means of many research
budgets.
ACCLIMATION AND LOADING
Different periods of acclimation for shrimp
may be necessary depending on the desired
test conditions. Handling in the field and ac-
climation to ambient laboratory conditions
are the most severe stresses the animals
THE PROGRESSIVE FISH-CULTUR1ST
235
-------
should encounter before toxicant exposure.
The shrimp, obtained from bait dealers or
seined locally, were initially transferred in
the laboratory to water closely matching- the
temperature and salinity of the transporting
medium. They were then acclimated to am-
bient laboratory conditions for 4-7 days; ac-
climation was considered successful if less
than 10% of the animals died. After this ini-
tial acclimation, the shrimp were transferred
to 30-liter glass experimental flow-through
tanks containing water of ambient tempera-
ture and salinity. Each experimental tank
had 15-25 animals of uniform length and the
weight of each test shrimp did not exceed 6 g.
The shrimp were provided with 2-cm of sand
in which to burrow and the tanks were cov-
ered with screens to prevent shrimp from es-
caping. Water temperature and salinity were
changed gradually from ambient to test con-
ditions of 25' ± 2° C and 20 ± 2°/00 (change
did not exceed 2°C and 2%>o in 4 h). The shrimp
were allowed to acclimate in water with con-
trolled temperature and salinity an addi-
tional 3-7 days before toxicant exposure. This
laboratory acclimation procedure maintains
apparently healthy bioassay shrimp, pro-
vided that the animals were not diseased be-
fore acclimation.
Loading, in flow-through bioassays, is a
function of water exchange with time. The
ratio of waterflow rate to shrimp weight must
not be so low as to affect test results \1]. It is
important, therefore, that the ratio be such
that:
(a) The concentration of dissolved oxygen
in the tank does not fall below 60% sat-
uration;
(b) the concentration of metabolic products
does not. become toxic;
(c) the concentration of toxicant in the
water remains constant throughout the
test; and
(d) the shrimp are not stressed by crowd-
ing, since crowding evokes cannibalism.
Loading ratios in our bioassays range from
7.5 1/g per day to 22.5 1/g per day.
EXPERIMENTAL APPARATUS
The apparatus used to control temperature
and salinity is shown below. Raw seawater
LCCL.VD
P'JMP
SA.VD F1L1TR
VALVE
FINAL HLTKR
SKAUA7KR SOUND! D
FRESHWATER SOI EKOID
RFfRiiiERATioN mr
MIXING »<1X
Lf.FT CHAMBER 0? NIXINC BOX
Mirwu CHAMBER OF MIXIM: mvv
Rli.HT CHAMBER UK MIXING BOX
SAL IN I Tl tomoUEX nETKTOR
r" n\- PIPE
CXWSTAWT HEAD TROUim
FLOAT SWITCH
WUY CONTROL
TFWJ.RATIIRE RECOSUER
HIXIW. TUBF.
VOL. 37, NO. 3, JULY 1975
Diagram of Saltwater Flow-Through Bioassay System.
236
127
-------
was pumped (A) from Santa Rosa Sound
through a high-capacity Seablue sand fV.ter
(B) and an AMF 20 micron final filter (D). The
salinity controller detector (/) and relay con-
trol (M) described by Rahner and Nimmo \2\
activate the seawater (E) and freshwater (F)
solenoids, adjusting the salinity of the incom-
ing water. The float switch (L) regulates the
head of water in the constant-head trough
(K). Incoming saltwater and freshwater enter
the left chamber of the mixing box (Hi) and
pass through a 5-em hole into the middle
chamber (H2), where it is stirred and cooled
by a refrigeration unit (G). The mixed water
then flows via another 5-cm hole through the
right chamber of the mixing box (H3), where
the salinity is monitored, and then into the
constant-head trough. When incoming water
temperature is less than 25°C, the water is
heated with a 7,000-W tubular heater placed
in the constant-head trough. Temperature
and salinity stratifications in the constant-
head trough are eliminated by the action of
air stones placed at half-meter intervals.
Tempei*ature is recorded by a recording
thermograph and salinity is checked daily to
insure proper operation of the salinity con-
troller. All fittings are made of PVC or SS 316
stainless steel and, to insure reliable perfor-
mance, the sand filter is backflushed periodi-
cally.
TOXICANT DELIVERY
The flow-through system permits adminis-
tration of a variety of toxicants in a wide
range of concentrations. The method of toxic-
ant delivery varies according to the amount
of toxicant needed. For pesticide and metal
bioassays, we use Harvard syringe-pumps
equipped with Hamilton Gastight Luer-Lok
glass syringes. The toxicant is delivered
through 0.12-cm (0.047-inch) ID polyethylene
tubing into a glass mixing tube (see diagram)
held at the effluent end of a calibrated siphon
which delivers 50 to 75 liters of water/hour
from the constant-head trough. These flow
rates produce a simple water turnover (tank
volume) in 45-60 minutes, and a complete ex-
change in 2 to 3 hours. The mixing tube in-
sures thorough mixing of toxicant with water
entering the test tank (see O in diagram).
Choice of carriers or toxicant solubilizers (if
128
237
needed) depends primarily on the solubility ( f
the toxicant under study. For example, larga
amounts Of cadium chloride can be dissolve 1
in de-ionized water to make stock solutions
for cadmium bioassays. Aroclor 1254, DDT,
and malathion are readily soluble i i
triethylene glycol whereas methoxychlor is
not. The low solubility of methoxychlor nece-c -
sitates relatively greater carrier to toxicant
ratios and greater toxicant-carrier flow rate?
to achieve the desired concentrations of
methoxychlor in the bioassay. For extremely
insoluble organic compounds, acetone may b >
used as a earner; however, it should be note 1
that acetone is toxic and, as with all carrier*
used in bioassays, a control receiving th>
same concentration of carrier as the experi-
mentals, as well as a seawater control with-
out carrier, must be included in each test.
Conclusion
Increasing demand for toxicological re-
search makes it imperative that experimental
conditions be controlled during bioassays so
that test conditions can be repeated as accu-
rately as possible. We believe, therefore, that
the system described here will prove useful to
other investigators who perform bioassays.
ACKNOWLEDGMENT
We thank Steven Foss of GBERL for draw-
ing and photographing the illustration.
REFERENCES
>. ALABASTER. J. S., ami F. S. H. ABHAM.
l!«i!>. Development and use of a direct method of
•vv;»hiatinj? toxicity to fish. Advances in Wattr Pol-
lutiiin RfHfftrch. Proceedings 2nd International
Conference. Tokyo 19*5-1, Vol. 1, p. 41-54, Perjjamon
Press, Oxford.
2. BAHSKH, L. H., and I). R. NIMMO.
1975. A salinity controller for flowing-water bioas-
says. Transactions of the American Fisheries Soc-
iety. 10-t: :!KH-.'J,KJI.
3. NIMMO, D, R., and L. H. BAHSKR.
1973. Physiological consequences of polychlorinated
hiphenyl and salinity-stress in penaeid shrimp.
Presented at symposium "Pollution and
Physiological Ecology of Estuarine and Coastal
Water Organisms": Nov. 14-17, 1973, Hobcaw
Barony, South Carolina.
THE PROGRESSIVE F1SII-CULTURIST
-------
4 SlLBERCELH K K *>• VERXBKKl!. \V. B. and VKRXHFI'.'i. J-
1973. Dieldnn: Effect? of chronic suhlethnl exposure 1972, The synorKi«tic effects of temperature. sal'ni<-y
on adaption to thormal stress in freshwater li.sh. and mercury on survivsl and metabolism of the
Environmental Sc.ence and Technolopy, 7: 846- adull fia.ller crab, tVa puj/i/alor. L. b. l> ish and
849 Wildlife Service Fisheries Bulletin, 70: 415-420.
VOL 37, NO. 3, JULY 1975
238 129
-------
Reprinted from Progress in
Experimental Tumor Research,
VoL 20: 304-314, 1976, with
permission of S. Karger
AG, Basel
ATTEMPTS TO INCREASE BACULOVIRUS PREVALENCE IN SHRIMP BY CHEMICAL EXPOSURE
John A. Couch
Contribution No. 240
239
-------
Progress in Experimental Tumor Research
Series Editor: F. HOMBURGER, Cambridge, Mass.
Publishers: S. KARGER, Basel
REPRINT (Printed in Switzerland)
Prog. exp. Tumor Res., vol. 20, pp. 304-314 (Karger, Basel 1976)
Attempts to Increase Baculovirus Prevalence in
Shrimp by Chemical Exposure1
JOHN A. COUCH
US Environmental Protection Agency, Gulf Breeze Environmental Research Labora-
tory, Sabine Island, Gulf Breeze, Fla.
Introduction
Relatively little attention has been given to interactions - synergistic or
additive - possible among natural infectious pathogens, chemical pollutants
and aquatic animal species. In mammalian and avian systems, studies have
shown that unrelated chemical agents may have similar enhancing and in-
ducing effects on viruses. The fact that certain oncogenic viruses (i.e., RNA
C-type viruses such as Rauscher leukemia virus, murine sarcoma virus) can
be induced and activated by chemical inhibitors of protein synthesis [1],
heavy metals [2], and synthetic steroids [3], indicates that model systems are
needed to study further the interactions among host, chemical agent, and
infectious agent. Nononcogenic viruses also have shown enhanced positive
responses, deleterious to their hosts, in the presence of pesticides [4], and
polychlorinated biphenyls (PCBs) [5].
Invertebrates and lower vertebrates are known to be hosts for certain
viruses and to have viral diseases. Most viral diseases of invertebrates have
been reported to occur in insects [6]. Of the many viral groups found in
insects, only members of the Baculovirus (nuclear polyhedrosis viruses) group
have been studied in relation to interactions of viruses, chemicals, and other
environmental factors [7]. In this regard, hydrogen peroxide, potassium
nitrate, hydroxylamine and unusually cold temperature have been reported
to induce or enhance Baculovirus in the silkworm [7, 8]. Most interesting is
the recent report of increase, up to 9-fold, of virulence of a Baculovirus in the
insect Spodoptera frugiperda after exposure of the virus in vivo to 3-methyl-
1 Contribution No. 240, Gulf Breeze Environmental Research Laboratory.
241
-------
COUCH
305
cholanthrene [9]. Mechanisms of enhancement by most of these chemical
agents are not understood.
Aquatic animals have been largely ignored in past studies of possible
virus-chemical interactions. However, recent studies at the Gulf Breeze En-
vironmental Research Laboratory have revealed a new virus, probably a
Baculovirus, in penaeid shrimp from the northern Gulf of Mexico [10-12].
COUCH [12] placed this rod-shaped virus in the Baculovirus group [15] be-
cause of its close similarities to the baculoviruses found commonly in insects.
The shrimp-virus system is presently being studied, particularly with regard
to the physical, chemical, and biological characterization of the virus and
interactive effects of the virus and chemical agents, such as pesticides and
PCBs.
The purpose of this paper is: (a) to describe the virus-shrimp relation-
ship, and (b) to present results from recent tests of exposure of samples of
shrimp to several pesticides and industrial chemicals found as stressing polr
lutants in aquatic ecosystems. The specific chemicals tested were: Aroclor
1254®2(PCB); Mirex (insecticide); methoxychlor (insecticide); and cadmium
(metal).
Materials and Methods
Viral, cellular and tissue studies. Virus material has been obtained since 1970 from
pink shrimp (Penaeus duorarum) taken directly from nature. To date, laboratory attempts
to transmit the virus from shrimp to shrimp by feeding infected tissues have been only
minimally successful. Naturally infected shrimp have been relatively easy to obtain from
samples caught in Apalachee Bay near Keaton Beach, Florida [12]. The shrimp from these
samples used for virus morphology and study of virus-host cellular relationships have been
patently infected (large crystalline polyhedral inclusion bodies present in nuclei of infected ,
cells), thus permitting diagnosis by light microscopy [12]. The organ chiefly affected by the
virus is the hepatopancreas. Whole hepatopancreas was fixed in neutral buffered, 10%
formalin or in Davidson's fixative for light microscopy and general histopathology. Sec-
tions were processed and stained with Harris hernatoxylin and eosin, Feulgen reaction,
mercury bromophenol blue, periodic acid-Schiff reaction, and methyl green-pyronin. Fresh
squashes of hepatopancreas were studied with both bright-field and phase-contrast micros-
copy. Patent infections were quickly diagnosed by finding characteristic viral crystalline
inclusion bodies referred to herein as polyhedral inclusion bodies (PIBs) [12]. Tissues with
heavy patent infections were diced in 2.5% glutaraldehyde, and cubes less than 0.5 mm8
were fixed in fresh 2.5% glutarajdehyde, postfixed in 1% OsO4, processed, and finally
2 ® Trademark, Monsanto Company, St. Louis, Mo. (The use of a trade name does
not imply endorsement of a product by the United States Environment Protection Agency).
242
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COUCH 306
embedded in Epon 812. Thin sections were stained in uranyl acetate and lead citrate, and
examined with a Zeiss EM 9S-2 electron microscope.
Chemical enhancement experiments. Pink shrimp used in chemical exposures were
taken from Apalachee Bay, Florida, by local bait dealers. The feral population of pink
shrimp in Apalachee Bay appears to be enzootic for Baculovirus. Levels of prevalence of
patent virus infections have ranged from 0 to 80% in samples of feral shrimp taken over
a 4-year period (over 2,000 shrimp examined). Approximately 20% of shrimp have light
to heavy patent infections.
Samples of 100-150 shrimp were brought to the laboratory. From these, base sam-
ples of 50 were examined for presence of PIBs by study of fresh squashes of their hepato-
pancreas tissues. Prevalence of PIBs in this base sample was designated the base prevalence
of the whole sample (percentage of shrimp with PIBs). Tests in which base samples were
not examined are so designated in the text of this paper. The remaining shrimp were
acclimated for 5-7 days to flow-through sea water in laboratory tanks, then exposed to
low concentrations of selected toxic chemicals for various periods of time. Hypodermic
syringes and plastic tubing mounted on syringe pumps were used to inject the test chemi-
cals into tanks at controlled flow-through rates which maintained a desired range of chemical
concentrations in the test-tank water throughout the duration of the test exposure [13].
Similarly maintained control tanks contained equal numbers of shrimp from the same
stock sample as the test shrimp (presumably with the same base prevalence of virus). In
tests with PCB, syringe pumps delivered into the control tanks concentrations of carrier,
polyethylene glycol, equal in amount to that used to deliver low-solubility polychlorinated
biphenyls into test tanks.
Dead or dying shrimp in test or control tanks, when recovered during exposures,
were examined for patent infections in hepatopancreatic tissue. At the end of the exposure
periods, surviving shrimp from both test and control tanks were killed and examined and
final prevalence of patent virus infections was determined.
Gas chromatographic analyses of water samples, taken weekly by Gulf Breeze Labora-
tory chemists, provided data on ambient concentration levels of PCB and Mirex in test
and control tanks.
Shrimp-Virus Relationship
The shrimp virus, Baculovirus penaei, has as its major target organ the
hepatopancreas, and as its preferred cellular site, the epithelial absorptive
cell nucleus of the hepatopancreatic acinus (fig. 2-4). Occasionally, in heavy
patent infection (PIBs present), midgut-cell nuclei contain PIBs, indicating
that infections can spread to midgut. In light to moderate infections, midgut
has not been found to be affected.
At the organismic level, there are no gross signs that indicate infection
in shrimp. Lethargic and moribund shrimps in the same sample have been
found to be heavily to lightly infected, as well as uninfected. Therefore,
presence of the virus is diagnosed only after light microscopical finding of
243
-------
Baculovirus Prevalence in Shrimp by Chemical Exposure
307
. ..• -.
//^. y. Fresh squash preparation of heavily infected shrimp. Note many pyramidal-
shaped inclusion bodies free in squash (arrows). * 688.
Fig. 2. Electron micrograph of cross-section of normal hepatopancreatic epithelial
cells from pink shrimp. Note normal nucleolus, nuclear membranes and cytoplasm. Com-
pare with figures 3-5. x 10,224.
244
-------
COUCH
308
Fig. 3. Two adjacent hepatopancreatic epithelial cells infected with rod-shaped virus
(arrows). Early infection: Note absence of inclusion bodies in nuclei and nuclear and cyto-
plasmic abnormalities, such as nuclear hypertrophy, chromatin loss, nuclear membrane
proliferation into cytoplasm, x 7,920.
intranuclear PIBs (0.5-20 um) and is confirmed by electron microscopical
finding of rod-shaped virions (288 nm long x 59 nm diameter, fig. 5) and
characteristic cytopathic effects associated with presence of virions and PIBs.
Attempts to isolate the virus and infect insect cell lines are presently under
way. At present, there are no crustacean cell lines in culture.
Fig. 4. Infected nucleus containing triangular-shaped section of .polyhedral inclusion
body (PIB). Note virions (rods) free in nucleoplasm and cytopathic changes evident in
cell profile, x 15,565.
Fig. 5. Higher magnification of rod-shaped virions (arrows) in cross section and longi-
tudinal section near point of PIP in shrimp cell nucleus, x 38,500.
245
-------
Baculovirus Prevalence in Shrimp by Chemical Exposure
309
•-
x
246
-------
COUCH 310
The PIB associated with shrimp Baculovirus shows strong pyronin stain-
ing, is negative for methyl green and the Feulgen reaction, but is weakly to
strongly positive for mercury bromophenol blue. These cytochemical reac-
tions suggest that the PIB consists of protein and RNA. PIBs of insect
baculoviruses have been found with biochemical tests to contain protein and
RNA [6]. Further, electron microscopical studies reveal that the shrimp
virus PIB is made up of spherical subunits, up to 20 nm in diameter, arranged
in linear lattice-arrays constituting the PIB matrix [12].
Some virions are occluded within the growing PIB in the infected nucleus
(fig. 4). However, many virions are not occluded, but remain free in the
nucleoplasm, and when the nucleus and cell rupture they are liberated into
the lumen of the hepatopancreatic acinus. This series of events may lead to
autoinfection of other hepatopancreatic epithelial cells.
COUCH [12] has described in detail the cytopathic effects of the virus
in the pink shrimp. Most of the pertinent morphological reactions of cells
to advanced infections are demonstrated in figures 3 and 4.
Though no proliferative cellular component has been observed in the
shrimp-virus system, there are remarkable morphological similarities be-
tween certain infected shrimp cells and undifferentiated tumor cells from
other systems, particularly at the ultrastructural level. Undifferentiated tumor
cells (e.g., Ehrlich ascites tumor cells) generally have been characterized by:
(a) a high ratio of nuclear to cytoplasmic volume; (b) many ribosomes, hence,
much RNA in cytoplasm; (c) relatively few profiles of endoplasmic reticulum;
(d) inconspicuous mitochondria; (e) lack of digestive vacuoles and lyso-
somes, and (f) particulate ribosomal precursors (usually-associated with nu-
cleoli) abundant in nucleoplasm [14]. Virus-infected shrimp hepatopancreatic
cells also possess these features (fig. 3). The infected shrimp cell, however,
usually retains certain of its differentiated features such as microvilli and
columnar shape (fig. 3). The partial resemblance of these cells to undiffer-
entiated tumor cells in unrelated systems may merely reflect a progressive
dedifferentiating effect of the virus on shrimp hepatopancreatic cells as the
cells are converted to 'virus factories'.
Chemical Enhancement Experiments
The staff of the US EPA Laboratory, Gulf Breeze, Florida, has done
considerable research on the effects of various chemical pesticides and indus-
trial chemicals on marine organisms. Among the chemicals studied recently
247
-------
Baculovirus Prevalence in Shrimp by Chemical Exposure 311
have been the PCB (polychlorinated biphenyl), Aroclor 1254, Mirex (organo-
chlorine pesticide), methoxychlor (organochlorine pesticide) and cadmium
(heavy metal), all potential chemical pollutants of estuaries and the sea. Pink
shrimp from populations enzootic for Baculovirus were exposed to low con-
centrations of these chemicals in preliminary experiments designed to deter-
mine effects of the chemicals. These experiments provided an opportunity
to determine if natural prevalence of virus in pink shrimp could be increased
in the laboratory by chemically stressing the shrimp.
In an initial experiment, surviving shrimp exposed for 30 days to 3 jig/
liter Aroclor 1254 had a much higher final prevalence of patent Baculovirus
infection than did nonexposed controls (table I, No. 1). In a second com-
parable test, prevalence of Baculovirus apparently was little increased over
that of the nonexposed control shrimp (table I, No. 2). In a third exposure
to lower levels of Aroclor 1254, survivors of the exposed sample of shrimp,
which had a high base prevalence, showed no final patent infections (table I,
No. 3). Mortality in both control and experimental shrimp was high (50%)
and some of the dead shrimp were cannibalized before they could be exam-
ined. Therefore, it is possible that infected shrimp weakened, died, and
were eaten before they could be examined. Shrimp which survived and were
examined may have been part of the sample that was uninfected at the onset
of the test (zero base prevalence). These data show no consistent trend favor-
ing Aroclor 1254 enhancement of Baculovirus infections in two of the three
samples of shrimp.
A small sample of pink shrimp taken from a population enzootic for
Baculovirus penaei showed a higher prevalence of patent virus infections
after 28 days' exposure to 0.01-0.23 ug/liter Mirex than did nonexposed
control shrimp taken from the same population (table II). Mortality was
much higher in the Mirex-exposed shrimp than in the nonexposed control
shrimp (table II).
Samples of shrimp from the Baculovirus enzootic population were .also
exposed to cadmium, cadmium plus methoxychlor, and methoxychlor (all
at 1.0 ug/liter). Base samples for Baculovirus prevalence were not examined
from the original stock of shrimp used in these tests. No patent virus infec-
tions were found in the shrimp that survived or died during the test exposures.
None of the control shrimp had patent virus infections. Exposed shrimp suf-
fered much higher mortality than did control shrimp in all tests (table III).
In this case, it is highly probable that the original stock of shrimp used was
not infected or, if infected at all, sparsely and lightly.
Of necessity, the preceding tests were done on an incompletely defined
248
-------
COUCH
Table I
Chemical
(cone.)
312
Base Exposure Final virus Mortality, %
prevalence, time, days prevalence, %
exposed control exposed control
1. Aroclor 1254 - 30
(3 ug/liter)
2. Aroclor 1254 10 (50) 10
(3.5 ug/liter)
3. Aroclor 1254 36 (50) 25
(0.7 ug/liter)
60 (20) 0 (20) data not available
12(34) 9(34) 100(70) 58.5(70)
0 (20) 0 (20) 50 (40) 47.5 (40)
Number of shrimp in sample given in parentheses.
Table II
Chemical
(cone.)
Mirex
(0.01-0.23 ug/liter)
Exposure Final virus Mortality, %
time, days prevalence, %
exposed control exposed control
28
40(15) 6.7(15) 80(40) 5(40)
Number of shrimp in sample given in parentheses. Base sample not examined.
Table HI
Chemical
(cone.)
Cadmium
(1.0 ug/liter)
Cadmium +
Exposure
time, days
25
25
Final virus
prevalence, %
exposed control
0 (12) 0 (14)
0 (8) 0 (14)
Mortality,
exposed
75 (20)
55 (20)
%
control
0(20)
0(20)
methoxychlor
(1.0 ug/liter each)
Methoxychlor
(1.0 jag/liter)
25
0(11)
0 (14) 85 (20) 0 (20)
Number of shrimp in sample given in parentheses. Base sample not examined.
249
-------
Baculovirus Prevalence in Shrimp by Chemical Exposure 313
virus-host system, namely, an in vivo system. Therefore, it is not surprising
that the data collected to date indicate a mixed response of the in vivo virus-
shrimp system to stress by the chemical agent. There is some indication that
low concentrations of Aroclor 1254 and Mirex may enhance naturally occur-
ring Baculovirus infections in pink shrimp. To control the shrimp-virus sys-
tem better in future enhancement tests, we are presently engaged in the fol-
lowing tasks: (a) isolation and biochemical characterization of the shrimp
virus; (b) transmission and infectivity studies of the virus in shrimps in the
laboratory, and (c) attempts to culture the virus in insect cell-lines that are
presently available (in vitro studies).
Ideally, a crustacean cell line would be best for in vitro studies of the
shrimp-virus system. To my knowledge, however, there are no reports of
successful continuous culture of crustacean tissues. This remains a challenge
to those engaged in invertebrate pathology.
There are no current reports of active neoplasms occurring in Crustacea.
Several suspect tumor-like growths have been reported, but their histories
are uncertain. If aquatic pollution continues, however, both viral diseases
and neoplasia of Crustacea may become factors in future research.
Summary
Little information is available concerning interactions between pollu-
tant chemicals and viruses in aquatic animals. Samples of pink shrimp
(Penaeus duorarum) with various enzootic levels of a natural Baculovirus
infection were experimentally exposed to low levels of Aroclor 1254®, a
polychlorinated biphenyl (PCB), Mirex, cadmium, and methoxychlor in the
laboratory. No consistent pattern of increase in prevalence of virus was
found, and no indication of tumor induction was detected.
Acknowledgments
Dr, DEL NIMMO and Mr. MARLIN E. TAGATZ graciously provided shrimp and data
from their toxicant-exposure studies. Miss GWYNDOLYN WALDORF supplied much techni-
cal assistance, and Mr. LEE COURTNEY examined many feral shrimp. Gas chromatographic
analyses of water samples were performed by Mr. DENNIS KNIGHT.
250
-------
COUCH 314
References
1 AARONSON, S. A. and DUNN, C. Y.: High frequency C-type virus induction by inhibi-
tors of protein synthesis. Science, N.Y. 183: 422-424 (1974).
2 GAINER, J.H.: Effects of heavy metals on viral infections in mice. Envir. Hlth
Perspect. 4: 98 (1973).
3 PARKS, W. P.; SCOLNICK, E. M., and KOZIKOWSKI, E. H.: Dexamethasone stimulation
of murine mammary tumor virus expression: A tissue culture source of virus.
Science, N.Y. 184: 158-160 (1974).
4 CROCKER, J.F.S., et al.: Insecticide and viral interaction as a cause of fatty visceral
changes and encephalopathy in the mouse. Lancet 1974: 22-24.
5 FRIEND, M. and TRAINER, D.O.: Polychlorinated biphenyl: Interaction with duck
hepatitis virus. Science, N.Y. 170: 1314-1316 (1970).
6 VAGO, C. and BERGOIN, M.: Viruses of invertebrates. Adv. Virus Res. 13: 247-303
(1968).
7 HIMENO, M.; MATSUBANA, F., and HAYASHIYA, K.: The occult virus of nuclear poly-
hedrosis of the silkworm larvae. J. invertebr. Path. 22: 292-295 (1973).
8 YAMAFUGI, K.: Metabolic virogens having mutagenic action and chromosomal pre-
viruses. Enzymologia 27: 217-274 (1964).
9 REICHELDERFER, C. F. and BENTON, C.V.: The effect of 3-methylcholanthrene treat-
ment on the virulence of a nuclear polyhedrosis virus of Spodoptera frugiperda.
J. invertebr. Path. 22: 38-41 (1973).
10 COUCH, J.: Free and occluded virus, similar to Baculovirus, in hepatopancreas of
pink shrimp. Nature, Lond. 247: 229-231 (1974).
11 COUCH, J. and NIMMO, D.: Ultrastructural studies of shrimp exposed to the pollutant
chemical polychlorinated biphenyl (Aroclor 1254). Bull. Soc. pharm. envir. Path. 2:
17-20 (1974).
12 COUCH, J.: An enzootic nuclear polyhedrosis of pink shrimp: Ultrastructure, preva-
lence, and enhancement. J. invertebr. Path. 24: 311-331 (1974).
13 NIMMO, D. R.; BLACKMAN, R. R.; WILSON, A. J., jr., and FORESTER, J.: Toxicity and
distribution of Aroclor® 1254 in the pink shrimp Penaeus duorarum. Marine Biol. 11:
191-197 (1974).
14 TRUMP, B.F. and ARSTILA, A.U.: Cell injury and cell death; in LAViA and HILL
Principles of pathobiology, pp. 9-95 (Oxford University Press, New York 1971).
15 WILDY, P.: Classification and nomenclature of viruses. Monogr. Virol. (Karger,
Basel 1971).
Dr. J.A. COUCH, US Environmental Protection Agency, Gulf Breeze Environmental
Research Laboratory, Sabine Island, Gulf Breeze, FL 32561 (USA)
251
-------
Reprinted from Progressive
Fish-Culturist, Vol. 38 (1)
51-52, 1976
PRECISION LIVE-FEEDER FOR FLOW-THROUGH LARVAL
CULTURE OR FOOD CHAIN BIOASSAYS
Lowell H. Bahner and Del Wayne R. Nimmo
Contribution No. 246
253
-------
Precision Live-Feeder for Flow-Through Larval Culture
or Food Chain Bioassays1
Lowell H. Bahner and DelWayne R. Nimmo
C7.S. Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory, Sabine Island, Gulf Breeze, Florida 32561
This report describes an inexpensive automatic feeder
that features precise timing of a wide choice of food
delivery periods and time intervals between food deliv-
ery. The feeder can also control simultaneous delivery
of a variety of foods, is suitable for delivering toxicant-
laden live foods, and is compatible for use in flow-
through water systems. Although several automatic
feeders have been described for culturing fish and
crustaceans (Benoit et al. 1969; Serfling et al. 1974;
Schimmel and Hansen 1975), we believe that ours is
more versatile.
A study at the Gulf Breeze Environmental Research
Laboratory was undertaken to quantitate the ac-
cumulation of a toxicant by animals feeding on nauplii
of brine shrimp (Anemia salina) containing toxicant in
a flow-through system. Since the study required fre-
quent and precise feeding with brine shrimp that con-
tained a range of toxicant residues, an electronic timer
and switch circuit (Fig. 1) was designed to control the
feeding cycle.
Brine shrimp were hatched during 48 h in 2-liter
separatory funnels that contained aerated seawater
(salinity 20°/oo) to which different amounts of toxicant
had been added to produce the desired whole-body
residues in the brine shrimp. The total hatch in each
funnel was collected on a fine nylon screen (150 /tin)
and rinsed with clean seawater to remove loosely
sorbed toxicant. The brine shrimp were then placed in
the appropriate feeder reservoirs containing clean
seawater. Aeration of the reservoirs dispersed the
shrimp throughout the water. Oscillating pumps, acti-
vated by the timer, pumped the brine shrimp-seawater
mixtures to the test aquariums (Fig. 2).
The timer provides a 5-s food-delivery period each
8 min, when fitted with the components shown in
Fig. 1. Feeding cycles can be adjusted by the selection
of appropriate components. An electric counter is used
1 Contribution No. 246, Gulf Breeze Environmental Research
Laboratory.
VOL. 38, NO. 1, JANUARY 1976
CONTROUID AC OUTLITS
ILECTRONIC SWITCH
TIMER
Fig. 1. Schematic diagram of electronic timer and switch, Rl, 2.2
megaohms; R2, 15,000 ohms; R3, 150 ohms; R4, 100 ohms; Cl,
330 microfarads (mfd), 10-V, tantalum; C2, 0.01 mfd; C3, 0.1
mfd; Dl, germanium diode; Kl, 12-V DC relay; 1C1, 555 in-
tegrated circuit, Radio Shack 276-1723; Ql, Sylvania ECG5693.
Aquaria
Fig. 2. Illustration of timer-controlled automatic feeder.
to count the timer cycles as a check for proper system
operation.
The controlling system (Fig. 1) consists of a 555 in-
tegrated-circuit timer, a triac (a high-current electric
switch), and a low-voltage DC power supply. The 555
timer is capable of timing periods that are continuously
variable from microseconds to 2 h, thus providing pre-
cise timing control not possible with clock-driven
51
255
-------
timers. The duration of on/off time periods of the timer
for any given application can be derived from formulas
given by Jung (1973). The triac (Ql) is capable of con-.
trolling several pumps and counters simultaneously
without mechanical failure or arcing. Power for the
switch and pumps is standard 117-V AC, but the 555
timer requires 5- to 15-V DC; a suitable power supply
was described by Bahner and Nimmo (1975). The cost
of the components, enclosed in an aluminum minibox,
was about $40.00.
The feeder described here could be used as an aid in
rearing vertebrate and invertebrate larvae in the
laboratory and could provide deli very of food in studies
on animal nutrition, toxicology, or food chains. It is also
suitable for flow-through toxicity or accumulation ex-
periments, which have rarely been reported for small
aquatic animals when the mode of toxicant delivery
was through live food rather than through the water.
References
Banner, L. H., and D. R. Nimmo, 1975. A salinity controller for
flow-through bioassays. Trans. Am. Fish. Soc.
104(2):388-389.
Benoit, D., R. Syrett, and J. Hale, 1969. Automatic live brine
shrimp feeder. Trans. Am. Fish. Soc. 98(3):532-633.
Jung, W. G. 1973. The 1C time machine. Pop. Electr.
4(5):54-57.
Schimmel, S. C., and D. J. Hansen, 1975. An automatic brine
shrimp feeder for aquatic bioassays. J. Fish. Res. Board
Can. 32(2):314-316.
Serfling, S. A., J. C. Van Olst, and R. F. Ford, 1974. An
automatic feeding device and the use of live and frozen Ar-
temia for culturing larval stages of the American lobster,
Homarus americanus. Aquaculture 3:311-314.
52
256
THE PROGRESSIVE FISH-CULTURIST
-------
Reprinted from Bulletin of
Environmental Contamination
and Toxicology, Vol. 14(6):
753-759, 1975, with permis-
sion of Springer-Verlag
New York Inc.
AND MARINE UNICELLULAR ALGAE: ACCUMULATION, POPULATION
GROWTH, AND OXYGEN EVOLUTION
Terrence A. Hoi lister, Gerald E. Walsh, and Jerrold Forester
Contribution No. 248
257
-------
Mirex and Marine Unicellular Algae: Accumulation,
Population Growth and Oxygen Evolution1
by TERRENCE A. HoLLiSTER2, GERALD E. WALSH, and JERROLD FORESTER
U.S. Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Sabine Island, Gulf Breeze, Fla. 32561
(Associate Laboratory of the National Environmental
Research Center, Corvallis, Ore.)
Many organochlorine compounds are toxic to algae.
SODERGREN (1967) demonstrated that less than 0.3 parts
per billion (ppb) of DDT inhibited growth of a fresh
water species of Chlorella. WURSTER (1968) reported
that DDT reduced the rate of photosynthesis in five
species of marine algae. de la CRUZ and NAQVI (1973)
showed that one part per million (ppm) of mirex re-
duced net photosynthesis by 557o in a fresh water spe-
cies of Chlamydomonas.
Uptake of organochlorine compounds by algae is
well documented. VANCE and DRUMMOND (1969) reported
that selected cultures of green and blue-green algae
concentrated DDT an average of 210 x, aldrin 188 x,
and endrin 215 x the exposure concentration of 1 ppm.
RICE and SIKKA (1973) showed that various marine algae
accumulated dieldren from 1,000 to 16,000 x the expo-
sure concentration of 1.7 ppb.
Mirex (dodecachlorooctahydro-1,3,4-metheno-2H-
cylobuta (cd) pentalene) is a persistant organochlorine
insecticide used to control the imported fire ant,
Solenopsis richteri Forel, in the southeastern United
States. Fire ant infestations are often located in
areas that drain into estuarine marshlands and embay-
ments. Mirex applied to coastal areas and upland
watersheds can enter estuaries (BORTHWICK et al. 1973).
This study was initiated (1) to determine effects
of mirex, if any, on population growth and oxygen evo-
lution by selected estuarine unicellular algae under
various conditions of salinity and nutrient concentra-
tion and (2) to determine if mirex can be accumulated
by the algae.
^ Contribution No. 248, Gulf Breeze Environmental
Research Laboratory.
2
Present address: Bionomics Marine Laboratory,
Route 6, Box 1002, Pensacola, Florida 32507.
753
Bulletin of Environmental Contamination & Toxicology,
Vol. 14, No. 6 © 1975 by Springer-Verlag New York Inc. _ Q
-------
METHODS
Population growth. Algae were exposed to mirex in
artificial sea water£ supplemented with trace elements
and vitamins to determine its effects on population
growth. Three nutrient concentrations were tested:
one-tenth, one-half, and full-strength. Full-strength
medium was that which yielded maximal growth in untreat-
ed cultures and it contained, per liter, 30 mg Na2EDTA,
14 mg FeCl2-6H20, 34 mg H3Bo3, 4 mg MnCl2-4H20, 2 mg
ZnS04-7H20, 6 mg K3P04, 100 mg NaN03, 40 mg Na2Si03-
9H20, 5 yg CuS04, 12 yg CoCl2, 50 yg thiamine hydro-
chloride, 1 yg vitamin B12, and 0.01 yg biotin.
Salinities were 5, 15, and 30 parts per thousand (ppt)
and pH ranged between 7.9 and 8.1. The medium was
sterilized by autoclaving at 121 C for 15 minutes.
WALSH (unpubl.) showed mirex to have low solubi-
lity in seawater, 0.2 ppb being the highest concentra-
tion obtainable. This concentration was used in
growth studies.
Algae tested were the chlorophytes Chlorococcum
sp., Dunaliella tertiplecta Butcher, and Chlamydomonas
sp.; the bacillariophytes -Nitzschia sp-. (Indiana strain
684) and Thallasiosira pseudonana Hasle and Hundal, and
the rhodophyte Porphyridium cruentum (Ag.) Naeg. Algae
were obtained from the culture collections of the Woods
Hole Oceanographic Institution, Scripps Institution of
Oceanography, and Indiana University.
Stock cultures, grown in and acclimated to various
salinities and nutrient concentrations for one week,
were diluted with appropriate medium to the absorbance
of 0.100 at 525 nm on a Fisher electrophotometer. The
diluted algal suspensions were used as inocula for
growth tests. One ml of the appropriate suspension was
added to culture flasks that contained 49 ml of test
medium. Treated cultures contained 0.2 ppb technical
mirex that- was added in acetone carrier. Untreated
control flasks contained an identical amount of acetone
(0.01% of the volume).
All cultures were grown on rotary shakers at 20 C
under 6,000 lux illumination with alternating 12-hr
periods of light and darkness. Triplicate flasks were
3 From Rila Products, Teaneck, New Jersey. Mention
of commercial products does not constitute endorsement
by the Environmental Protection Agency.
754
260
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analyzed at each salinity and nutrient concentration
and each test was performed twice. After seven days,
growth was measured spectrophotometrically at 525 run
on a Fisher electrophotometer.
Oxygen evolution. To determine effects of mirex
on oxygen evolution,10 ml samples of Chlorococcum sp.
and Chlamydomonas sp. from growth studies were centri-
fuged gently and resuspended in fresh medium to the
absorbance of 0.100 at 525 nm. Four ml of each cell
suspension were placed in reaction vessels of a Gilson
photosynthesis-model differential respirometer. The
vessels contained C02 buffer in the wells (UMBRIET et
al. 1964). After equilibration at 20 C for 20 min,
oxygen evolution was measured for 60 min. Flasks
were analyzed in triplicate at each salinity and nutri-
ent concentration and each test was performed twice.
Accumulation. Fifty ml of stock algal cultures,
in the logarithmic phase of growth and diluted to the
absorbance of 0.100 at 525 nm, were added to 950 ml of
sterile medium in 2,800' ml Erlenmeyer flasks. The algae
were Chlorococcum sp., Chlamydomonas sp., D. tertiolecta,
and T. pseudonana. The test medium contained the full-
strength concentration of nutrients and' salinity was 15
ppt. Stock solutions of mirex were prepared in acetone
and added to the medium to give a concentration of 10,
25, or 50 parts per trillion (pptr) in accumulation
studies. Incubation was similar to that for growth
studies except the flasks were not shaken. After seven
days exposure, the cells were harvested by centrifuga-
tion at 4,200 x g for 10 min, resuspended in mirex-
free medium, and centrifuged again. This procedure was
repeated three times to remove mirex in interstitial
water of the algal pellet or bound loosely to the cells.
Samples were then stored in a dessicator until analyzed.
To determine the amount of mirex accumulated, each
sample was weighed and placed in a Duall® 1;issue grinder
and extracted with three 2.0 ml portions of aeetonitrile.
The combined aeetonitrile extract was diluted with 6 ml
of 2% (w/v) ,Na2S04 in distilled water, shaken and extrac-
ted with three 2.0 ml portions of hexane. The combined
hexane extract, concentrated to 0.5 ml by evaporation
with a Snyder column, was transferred to a size "B"
Chromaflex® column containing 1.5 g of Florisil and 1.5
g of anhydrous Na2SC>4. The extract was eluted from
the column with 20 ml of 1% (v/v) ethyl ether in hexane.
The eluate was adjusted to an appropriate volume for
Registered trademark, Kontes Glass Company, Vineland,
New Jersey.
755
26}
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analysis by gas chromatography using a Varian
Aerography model 1400 gas chromatograph equipped
with an electron capture detector and a 182 cm x 2
mm (ID) glass column packed with 270 OV-101 on Gas
Chrom Q. Operating parameters were: injector tem-
perature, 210 C; column temperature, 192 C; detector
temperature, 210 C; and gas flow, 25 ml/min. Mirex
was quantitated by comparison with the peak height
of a known-concentration standard.
Student's "t" test was used to analyze differences
between means of treated and control cultures.
BESULTS
Population growth. Figure 1 shows population den-
si tiesoFTEesTxaTgal species after growth for seven
days in 0.2 ppb mirex and three nutrient concentrations,
Chhimydomona* *p.
T.. pi«udonar
Nltxtchlo tp.
.200
.100
M
s
<
«n
a
0
P. criKntum
Chtorococcum »p.
D. tartlotectc
.300r
.100
0.1 0.5 1.0 0.1 0.5 1.0
NUTRIENT CONCINTRATION
as 1.0
Figure 1. Comparison of cell densities at three
nutrient,concentrations of untreated cultures
(solid bar) and those grown in 0.2 ppb mirex
(hatched bar).
756
262
-------
There were no statistically significant differences
(p = 0.05) in final population density when any species
was grown in 5, 15 or 30 ppt salinity. Therefore,
population densities for all salinities were combined
and only differences in density between nutrient concen-
trations are shown in Fig. 1.
In all cases, except Chlorococcum sp. at 1/10
strength nutrients, treated cultures exhibited higher
absorbances than did control cultures, regardless of
nutrient concentration, but the differences were not
statistically significant at the 0.05 level.
Oxygen evolution. Tables 1 and 2 compare oxygen
evolution by control and treated cultures of Chlorococcum
sp. and Chlamydomonas sp. in the growth study"! Cultures
grown at 1/10 strength nutrient concentration did not
contain a sufficient number of cells to make an adequate
comparison. No significant differences in rates of
oxygen evolution were found between control and treated
cultures of either species at any salinity or nutrient
concentration (p = 0.05).
TABLE 1
Oxygen evolved by Chlorococcum sp. in control cultures and in
cultures exposed to 0.2 ppb of mirex for seven days at three
salinities and tvo nutrient concentrations.
Control
Exposed
Control
Exposed
Control
Exposed
Salinity
ppt
T C
15
o n
30
0.5 nutrient
strength, yl/hr
17.8
20.it
23.6
18.2
19.8
16.8
Full nutrient
strength, yl/hr
18.9
18.6
19.2
19.8
20.it
20.8
Table 2
Oxygen evolved "by Chlamydomonas sp. in control cultures and in
cultures exposed to 0.2 ppb of mirex for seven days at three
salinities and two nutrient concentrations .
Salinity 0.5 nutrient Full nutrient
ppt strength, yl/hr strength, yl/hr
Control «- itl.O U2.6
Exposed 1*5.8 it2.*t
Control ,<- 39-5 ^5-6
Exposed ? Mt.it U3.2
Control
on
^
Exposed _ ^ it. 2 _ it6.8
757
263
-------
Accumulation. Figure 2 shows accumulation of
mirex by algae exposed to three low concentrations of
the pesticide for seven days at a salinity of 15 ppt.
Chlorococcum sp., D. tertiolecta and Chlamydomonas sp.
showed a significant linear relationship between amounts
accumulated and mirex concentrations in the medium,
Chlorococcum sp. was most efficient in uptake, accumu-
lating 88% of the mirex present in the medium.
Dunaliella tertiolecta and T. pseudonana removed approx-
imately 79% whereas Chlamydomonas sp. took up 55%.
Chlorococcum sp. concentrated the pesticide 7,300 x,
300
i
200
z
o
z
Ml
u
Z
8 100
10 25 50
CONCINTRATION IN MIDIUM (pptr)
Figure 2. Uptake of mirex by algal populations
after seven days exposure.
758
264
-------
D. tertiolecta 4,100 x, Ch1amydomonas sp. 3,200 x
and T. pseudonana 5,000 x the concentration in the
medium.
DISCUSSION
Accumulation of mirex by the algae was evident.
Within seven days, for example, Chlorococcum sp. accu-
mulated mirex from the nearly non-detectable concentra-
tion of 10 pptr in the medium and concentrated it to
more than 100 ppb on cells, a concentration factor of
10,000 x. If a similar condition existed in nature,
marine unicellular algae could accumulate mirex and,
when grazed upon, act as passive transporters of the
toxicant to consumers in the food chain.
In summary, these studies show that under our lab-
oratory conditions, mirex had no significant effect on
either population growth or oxygen evolution of selected
species of marine algae. It was however, accumulated
from the water by the algae.
REFERENCES
BORTHWICK, P.W., T.W. DUKE, A.J. WILSON, JR., J.I. LOWE,
J.M. PATRICK, JR. and J.C. OBERHEY. Pestic. Monit. J. ]_,
6 (1973) .
de la CRUZ, A.A. and S.M. NAQVI. Arch. Environ. Contam.
Toxicol. 1, 255 (1973) .
RICE, C.P. and H.C. SIKKA. Bull. Environ. Contam.
Toxicol. 9, 116 (1973) .
SODERGREN, A. Oikos 19, 126 (1967).
UMBREIT, W.W., R.H. BURRIS and J.F. STAUFFER.
Manometric Techniques, 4th ed., Burgess Publ. Co.,
Minneapolis, Minn., 305 pp. (1964).
VANCE, B.D. and W. DRUMMOND. J. Am. Water Works Ass.
61, 360 (1969).
WALSH, G.E. In Progress Report, Bureau of Commercial
Fisheries Center for Estuarine and Menhaden Research,
U.S. Dept. Interior Circular (Unoubl.).
759
265
-------
Reprinted from Marine Biology,
Vol. 35(4): 295-300, 1976,
with permission of Springer-
Verlag New York Inc.
LABORATORY SPAWNING AND REARING OF A I^RINE FISH,
THE SILVERSIDE I^NIDIA MENIDIA MENIDIA
D.P. Middaugh and P.W. Lempesis
Contribution No. 252
267
-------
Marine Biology 35, 295-3OO (1976)
© by Springer-Verlag 1976
Laboratory Spawning and Rearing of a Marine Fish, the
Silverside Menidia menidia menidia*
D. P. Middaugh and P. W. Lempesis
U.S. Environmental Protection Agency, Gulf Breeze Environmental Research Laboratory;** Bears Bluff Field Station, John's Island,
South Carolina, USA
Abstract
Adult silversid.es, Menidia menidia menidia (Linnaeus) , were collected in early March,
1974 and maintained in 3 recirculating seawater tanks in the laboratory. Respec-
tive groups were fed Moore-Clark Fry Fine at 3, 7 and 10% of their body weight per
day. The photoperiod (light intensity approximately 2OOO lux) was increased in in-
crements of 1O min/day from 12 h light to 14 h light. The water temperature was
increased by 1C°/day from the ambient collection temperature, 14°C, to 22°C.
Twenty-four days after beginning laboratory conditioning, fish in each tank were
stripped. There was a significant increase (x2^ a = 0.05) in the number of ripe
males at all three feeding levels, compared to an initial field-collected group,
that was checked at the beginning of the conditioning period. Females also showed
significant increases in ripeness at the 7 and 1O% but not at the 3% feeding level.
The gonadal indices (gonad weight expressed as percentage of body weight) of both
sexes were significantly greater than those measured for the initial field-col-
lected group, but did not differ from those of adults collected from the field at
the time laboratory conditioning was terminated. Techniques for maintaining eggs
from field-ripened adults in the laboratory have been developed, and the effect
of salinity on the percentage emergence of larvae determined. The highest emer-
gence rate of larvae was 61% when eggs were maintained at 3O& S . Emergence was 56%
at 2O& Sand 47% at 10£ S. The effect of delayed feeding on survival and growth of
larvae was determined at 2O and 30& Sand 25°C. Survival and growth was best for lar-
vae fed Artemia sp. nauplii immediately after emergence at 30%, S.
Introduction Females attain a maximum length of
123 mm; males, 112 mm. Spawning occurs
The southern subspecies of the silver- from March to August for the southern
side Menidia menidia menidia, an atherinid subspecies, with females releasing up to
fish, ranges from Florida to the Chesa- 500 eggs, 0.9 to 1.2 mm in diameter
peake Bay and infrequently as far north (Hildebrand, 1922).
as Woods Hole, Massachusetts (Kendall, Developing zygotes are found attached
1901; Bigelow and Schroeder, 1953; Rob- fcQ variou£ substrates by gelatinous
bins, 1969). The adults frequent estu- threads which form at ^e vegetal pole
arine areas through most of the year and Qf ^ following fertilization (Hil-
are usually found within a few meters of debrandf 1922). Development requires ap-
beaches and marshes. proximately 9 days and newly hatched
The silverside is omnivorous, feeding larvae are 5 mm in total length (Rubi-
on mysid shrimp, copepods, molluscan noff, 1958). Ryder (1883), Kuntz and
larvae and annelid worms (Kendall, 1901). Radcliffe (1917) and Hildebrand (1922)
have briefly described the embryology
of the silverside. Methods for labora-
* Contribution No. 252, Gulf Breeze Environ- tory spawning and rearing of field-
mental Research Laboratory. ripened adults have been reported by
**Associate Laboratory of the National Environ- Rubinoff (1958) and Rubinoff and Shaw
mental Research Center, Corvallis, Oregon, (196O) . These authors maintained devel-
USA. oping eggs in finger bowls at ambient
269
-------
296
D.P. Middaugh and P.W. Lempesis: Spawning and Rearing Menidia menidia menidia
room temperature (ca. 22°C). Hatched
larvae were transferred to flowing sea-
water aquaria and reared for 48 days.
Our study was conducted to develop
procedures for ripening and spawning
adults in the laboratory by altering tem-
perature and photoperiod. Methods were
also developed for routine spawning and
rearing of naturally-ripened field
stocks.
Materials and Methods
Laboratory Ripening and Spawning
Adult Menidia menidia menidia (Linnaeus) for
laboratory conditioning were collected
on March 9, 1974 from the North Edisto
River estuary in South Carolina, USA.
The water temperature was 13.8°C and
salinity 3O&. Three groups of fish, each
consisting of approximately 1O males and
10 females, selected according to size,
were placed in 1-ra diameter fiberglass
tanks in a recirculating system in the
laboratory. An activated charcoal-
crushed oyster-shell filter removed me-
tabolites and maintained a pH of 7.8 to
8.1. Water temperature was controlled
with a titanium-coil water-chiller and a
20OO W quartz immersion heater. Two
banks of 4 "cool white" fluorescent
lamps , suspended 1 00 cm above the tanks
provided approximately 2OOO lux illumi-
nation at the water surface. A Tork®
timer was used to control the photoperi-
od. Water depth was approximately 45 cm
and volume 31O 1 in each tank. A magnet-
ic-drive pump circulated 4OO to 46O 1 of
seawater per hour through each tank. The
salinity ranged from 24 to 26%,. Silver-
sides in respective tanks were fed
Moore-Clark Fry Fine at approximately
3, 7 and 1O% of their body weight each
morning. Excess food was siphoned from
each tank at the end of the day.
Silversides were laboratory-accli-
mated for 2 days at 1 4°C and 12 h light:
12 h dark. The water temperature was
then increased to 22°C (1C°/day for 8
days) and the photoperiod was increased
by 1O min per day until a 14 h light:
1O h dark regime was obtained. These
conditions remained constant until April
4, 1974 when fish were removed from
their respective conditioning tanks and
stripped.
Eggs were stripped into 200 mm diam-
eter, glass culture dishes and fertil-
ized with milt. The eggs were placed in
Mention of trade namss in this paper does not
imply endorsement by the U.S. Environmental
Protection Agency.
constant temperature boxes and incubated
at 22°C and 25£ S . Water in the culture
dishes was changed daily. Larvae began
to emerge 7 days after fertilization.
They were transferred to 2O 1 glass
aquaria and fed recently hatched Artenda
sp. nauplii each day.
To determine if feeding levels had
an effect on the gonadal index (gonad
weight expressed as percentage of body
weight), adults were sacrificed after
stripping. Each individual was weighed.
The gonads were then removed and
weighed. Gonadal indices were also de-
termined for initial and final groups
of fish collected from the field just
prior to and immediately after the lab-
oratory-conditioning interval. These
fish were sacrificed when collected.
Their gonadal indices were used to de-
termine the effectiveness of our lab-
oratory conditioning regime. Gametes
stripped from positive-response fish in
the laboratory study and final field-
collected fish were not included in
computing the gonadal index for these
two groups.
Field Spawning and Laboratory Rearing
Adult silversides were collected from
the same area as those used for the lab-
oratory ripening and spawning test. Col-
lections were made from April 15 to June
3O, 1974-(water temperature 19° to 30°C,
salinity 2O to 28%,). Immediately after
seining (while still on the beach) , 3 to
5 ripe females were dipped into a bucket
of seawater to remove sand and detritus.
Their eggs were stripped into 200 mm
glass culture dishes that contained sea-
water of 1O, 2O or 3O& S at ambient tem-
peratures (from 21° to 27°C) . After the
demersal eggs had been concentrated in
one area of the dish, excess water was
carefully decanted off, so that only
enough (approximately 1 mm deep) re-
mained to keep the eggs moist.
Milt from 2 or 3 males was then
stripped into the culture dish and gent-
ly mixed with the eggs using a glass rod.
Immediately after fertilization the eggs
of the silverside form gelatinous
strands which bind them together in a
mass (Hildebrand, 1922). Fertilized eggs
were removed from the culture dish by
rolling a 25 cm length of nylon string
around the gelatinous mass (1,5OO to
2,50O eggs), which adhered to the string.
The eggs were suspended in 4 1 wide-
mouth glass jars containing filtered sea-
water of 1O, 2O or 30%, S and transferred
to the laboratory. Each egg mass was
then suspended in a 4O 1 aerated glass
aquarium containing filtered seawater at
270
-------
D.P. Middaugh and P.W. Lempesis: Spawning and Rearing Menidia menidia menidia
297
the fertilization salinity and 21°C ±
1C0. A Dyna-Flo® Filter was used to main-
tain a gentle flow of water over the
eggs during development.
To assess the effect of starvation on
survival and growth, 7O newly emerged
larvae (less than 8 h old), from eggs
maintained at 2O and 30& S , were placed
in each of eight 2O-1 aerated glass
tanks containing water of 2O or 3O& S at
25°C ± 1C0- Recently hatched Artemia sp.
nauplii (approximately 5OO/1) were added
to the first two tanks immediately after
silverside larvae were introduced. Arte-
mia sp. were added to the second pair of
tanks after 48 h, and to the third pair
after 96 h. Larvae in the fourth pair of
tanks were not fed.
Concentrated aliquots of Artemia sp.
were then added to the first three re-
spective pairs of tanks each day to main-
tain a density of 2OO to 8OO nauplii/1.
Illumination at the surface of each tank
was approximately 2OOO lux for 14 h/day.
Observations were made for 14 days to
determine percentage survival in one of
the paired tanks. Subsamples of 6 to 1O
larvae, sacrificed at 48-h intervals,
from the other tank were measured to
determine total length.
Results
Laboratory Ripening and Spawning
Adult Menidia menidia menidia held in the
laboratory had respective survival rates
of 10O, 75 and 85% for feeding levels of
3, 7 and 10% of each group's body weight
per day. There were significantly more
(X2, a = O.O5) ripe males in all groups
held and conditioned in the laboratory
compared to males in the initial field-
collected group. However, there was no
significant difference between the num-
ber of ripe males conditioned in the
laboratory and those in the final field-
collected group. The mean gonadal index
of laboratory-held males also was sig-
nificantly greater (t-test, a = O.O5)
than that of males in the initial field-
collected group, but no significant dif-
ference was observed between the index
of males in the final field-collected
group and that of males held in the lab-
oratory (Table 1).
Significant increases in the number
of ripe females occurred among labora-
tory-conditioned groups fed at 7 and 1O%
of their body weight per day compared
to females in the initial field-collect-
ed group. None of the females fed at the
3% level were ripe when stripped at the
end of the conditioning period. There
was no significant difference in the num-
Table 1. Menidia menidia menidia. Summary of response of
adult silversides conditioned in laboratory and of groups
collected from field prior to (initial) and after (final)
laboratory conditioning interval. (Standard deviations of
mean gonadal indices are given in parentheses)
Feeding
level
Initial
field-
collected
3%,
7%
10%
Final
field-
collected
r>M-!^^al 1 nH
Sex
M
F
M
F
M
F
M
F
M
F
Number
Ripe
O
0
5
O
7
2
6
5
12
7
Non-ripe
12
12
2
13
2
4
1
5
3
11
Percentage
ripe
0
O
71
0
78
33
86
50
80
39
Mean
gonadal
index?
2.8
3.7
5.4
4.4
6.0
5.2
6.9
5.9
5.8
7.2
(0.72)
(0.96)
(1.8)
(1.5)
(1.2)
(1.2)
(1.4)
(1.6)
(1.5)
(1.7)
gonad weight
total body weight
U
Ul
O
1
III
OL
LU
a.
70
60
50
40
30
20
10
10 ppt
8 9 10 11
DAYS AFTER FERTILIZATION
12
Fig. 1. Menidia menidia menidia. Percentage
emergence of larvae from eggs maintained at 3
salinities and 21°C ± 1C°
271
-------
298
D.P. Middaugh and P.W. Lempesis: Spawning and Rearing Menidia menidia menidia
no
75
50
25
Ml *-*
DAYS AFTER EMERGENCE
DAYS AFTER EMERGENCE
100
7S
50
§2,
0
10
_ 9
£
I «
» 7
I *
5
41
24 6 8 10 12
DAYS AFTER EMERGENCE
14
100
75
§ 50
I 25
(9
4 6 8 10
DAYS AFTER EMERGENCE
12
14
Fig. 2. Menidia menidia menidia. Survival and growth of larvae (A) fed immediately after emergence,
(B) fed 48 h after emergence, (C) fed 96 h after emergence, (D) unfed. Pilled circles: 2C& S ,- open
circles: 3C& S ; vertical bars: ± 1 standard deviation
272
-------
D.P. Middaugh and P.w. Lempesis: Spawning and Rearing Menidia menidia menidia 299
her of ripe females at the 7 and 1O%
feeding levels, compared to females in
the final field-collected group. The
gonadal index of females (t-test, a =
O.O5) was significantly greater for all
three groups conditioned in the labora-
tory than that of females in the initial
field-collected group. No significant
difference was observed between the in-
dex of females in the final field-col-
lected group and that of those held in
the laboratory (Table 1) .
Field Spawning and Laboratory Rearing
Both the time of emergence and the per-
centage emergence of larval silversides
appears to be related to salinity (Fig.
1). We observed that many eggs developed
to an embryonic stage just prior to
emergence but often failed to emerge, or
emergence was delayed. This is partic-
ularly obvious if the times to 4O% emer-
gence for each salinity are compared
(Fig. 1). A delay of about 18 h at 20%, S
(relative to 30%,) and of about 42 h at
10% S (relative to 30%) occurred. Op-
timal emergence was observed for eggs
maintained at 3O% S .
The effect of delayed feeding on the
survival of silverside larvae was de-
termined at 2O and 30%, S . Optimal sur-
vival and growth was observed for larvae
fed immediately after emergence at 30%, S .
Survival and growth were not as good at
20% S (Fig. 2A) .
Larvae fed 48 h after emergence
showed poor survival compared to those
fed immediately after emergence. Some
larvae survived the 14-day post-hatch
interval at 30%, S , but at 20%, S all died
by the eighth day (Fig. 2B). Growth of
larvae fed 48 h after emergence and
those fed immediately after emergence
was similar.
Larvae fed 96 h after emergence and
unfed controls had similar survival and
growth rates. None of these larvae lived
more than 6 days after hatching (Fig.
2C, D).
Discussion
The maturation and spawning of the
silverside Menidia menidia menidia apparent-
ly is related to water temperature,
photoperiod and subspecific variations
(Kendall, 19O1; Gosline, 1948; Robbins,
1969). Kendall (19O1) collected maturing
silversides in April from Woods Hole,
Massachusetts; water temperature 9° to
12°C. Ripe fish were taken at water tem-
peratures from 13° to 21°C during May
through July. Adults collected at water
temperatures above 22°C were usually
spent. It is likely that these fish were
the northern subspecies, M. menidia notata.
Kendall (19O1) also reported an inter-
grading of M. menidia notata, the northern
form, and M. menidia menidia, the southern
subspecies, along the coast from Cape
May, New Jersey to Cape Hatteras, North
Carolina.
In the Chesapeake Bay the spawning
peak occurs in May (Bayliff, 195O), and
in North Carolina Hildebrand (1922) col-
lected ripe silversides from March
through August. In South Carolina, we
have collected ripe Menidia menidia menidia
from March through July at water temper-
atures from 16° to 30°C. However, the
mean gonadal indices of adults collected
during February, 1973 (water temperature
17° to 21°C) were low, (male gonadal in-
dex, 0.81 ± 0.69; female 1.33 ± O.32)
(Middaugh and Lempesis, Bears Bluff
Field Station, unpublished data).
These observations suggest that water
temperature and photoperiod may be im-
portant for maturation and ripening of
each subspecies of the silverside. In
higher latitudes, Woods Hole and the
Chesapeake Bay, water temperature ap-
parently is a limiting factor for Menidia
menidia notata. Our observations during
the winter of 1973 showed that photo-
period is probably limiting for the
southern silverside, M. menidia menidia.
Water temperatures warm enough for
spawning in the spring (17° to 21°C) oc-
curred during February (photoperiod ap-
proximately 10 h light), but the gonadal
index of both sexes remained low. During
late March and early April (photoperiod
approximately 12 h light), when the wa-
ter temperature was 16° to 2O°C, ripe
males and females were abundant.
Adult silversides maintained under in-
creasing water temperature and day-
length regimes in the laboratory ripened
and we were able to strip viable gametes
from some of them. Since the final field-
collected group showed the same kind of
response (percentage ripeness and gonad-
al indices), the specific effect of mod-
ifying environmental parameters in the
laboratory is unclear.
Additional tests with field-collected
fish held under constant winter environ-
mental regimes and experimental fish
subjected to combinations of constant
and increasing water temperature and
photoperiod will be necessary to deter-
mine the relationship of these factors
in the maturation process.
Adequate nutrition may also be essen-
tial for maturation and ripening of the
silverside. Males and females condi-
tioned in the laboratory showed in-
creased gonadal indices with increased
273
-------
300 D.P. Middaugh and P.W. Lempesis: Spawning and Rearing Menidia tnenidia menidia
food availability. Females in the tank
receiving food at the 3% level failed to
ripen. Kendall (1901) collected adults
during April and observed that the go-
nads were developing rapidly. The stom-
achs of these fish were often distended
with copepods, mysid shrimp, molluscan
larvae and annelid worms.
Optimal emergence of larvae from
eggs maintained at 30% S suggests that
the silversides used in our study were
adapted to high salinity. This may be
attributable to indigent population char-
acteristics since our fish normally en-
counter ambient salinities ranging from
24 to 31&. Silversides which live and
spawn in areas with lower salinity re-
gimes may therefore show optimal emer-
gence and survival at salinities (and
temperatures) nearest those typically
encountered during the time of spawning.
The effect of starvation on the sur-
vival and growth of other larval marine
fishes has been studied. Lasker et al.
(197O) found that irreversible starva-
tion of anchovy larvae (Engraulis mordax)
occurred if food was withheld for 1.5
days after yolk-sac absorption. Houde
(1974) learned that the elapsed time
from development of eye-pigmentation to
starvation in the bay anchovy Anchoa
mitchillif the sea bream Archosargus rhom-
boidalis and the lined sole Achirus lineatus
was useful in estimating the interval in
which the larvae of each species had to
begin feeding. Decreases in the time to
starvation were observed with increased
water temperature. Eye-pigmentation is
well developed at the time of emergence
of silverside larvae and they absorb
their yolk-sac 24 to 36 h after hatching
at 25°C. Optimal survival was observed
for larvae fed immediately after hatch-
ing in 20 and 30% S. Some larvae fed 48 h
after emergence, 12 to 24 h after yolk-
sac absorption, survived in 3O% S, but
all held in 20% S died by the eighth day
after emergence. This indicates that it
is essential to feed silverside larvae
immediately after emergence for maximum
survival. Rubinoff (1958) also demon-
strated a relationship between delayed
feeding and survival of silversides (Me-
nidia spp.). Larvae fed Artemia sp. on
the second day after emergence showed
good survival. If feeding was delayed
until the fifth day, no larvae survived
beyond the eighth day after emergence.
Literature Cited
Bayliff, W.H.: The life history of the silver-
side, Menidia menidia Linnaeus. Contr. Chesa-
peake biol. Lab. 9O, 1-25 (195O)
Bigelow, H.B. and W.C. Schroeder: Fishes of the
Gulf of Maine. Bull. Bur. Fish., Wash.-53,
1-577 (1953)
Gosline, W.A.: Speciation in the fishes of the
genus Menidia. Evolution, Lancaster, Pa. 2,
306-313 (1948)
Hildebrand, S.F.: Notes on habits and develop-
ment of eggs and larvae of the silversides,
Menidia menidia and Menidia beryllina. Bull.
Bur. Fish., Wash. 38, 113-12O (1922)
Houde, E.D.: Effects of temperature and delayed
feeding on growth and survival of larvae of
three species of subtropical marine fishes.
Mar. Biol. 26, 271-285 (1974)
Kendall, W.C.: Notes on the silversides of the
genus Menidia of the East coast of the
United States, with descriptions of two new
subspecies. Rep. U.S. Commnr Fish. 16,
241-267 (1901)
Kuntz, A. and L. Radclif fe: Notes on the embry-
ology and larval development of twelve
teleostean fishes. Bull. Bur. Fish., Wash.
35, 88-134 (1917)
Lasker, R., H.M. Feder, G.H. Theilacker and R.C,
May: Feeding, growth and survival of Engrau-
lis mordax larvae reared in the laboratory.
Mar. Biol. 5, 345-353 (197O)
Robbins, T.W.: A systematic study of the silver-
sides Membras Bonaparte and Menidia (Linnaeus)
(Atherinidae, Teleostei), 282 pp. Ph.D. Dis-
sertation, Cornell University 1969
Rubinoff, i.s Raising the atherinid fish, Meni-
dia menidia, in the laboratory. Copeia
1958(2}, 146-147 (1958)
- and E. Shaw: Hybridization in two sympatric
species of atherinid fishes, Menidia menidia
Linnaeus and Menidia beryllina, Cope. Bull.
Am. Mus. nat. Hist. 1999, 1-12 (1960)
Ryder, J.A.: On the thread-bearing eggs of the
silverside, Menidia. Bull. U.S. Fish Commn
3, 193-196 (1883)
Douglas P. Middaugh
U.S. Environmental Protection
Agency
Gulf Breeze Environmental
Research Laboratory
P.O. Box 368
John's Island,
South Carolina 29455
USA
Date of final manuscript acceptance: January 16, 1976. Communicated by M.R. Tripp, Newark
274
-------
Reprinted from the Annals of
the New York Academy of
Sciences, Vol. 266: 528-536,
1975, with permission of the
New York Academy of Sciences
ENVIRONMENTAL SIGNIFICANCE OF BACULOVIRUS INFECTIONS IN
ESTUARINE AND MARINE SHRIMP
John A. Couch, Max D. Summers, and Lee Courtney
Contribution No. 253
275
-------
ENVIRONMENTAL SIGNIFICANCE OF BACULOVIRUS
INFECTIONS IN ESTUARINE AND MARINE SHRIMP *
John A. Couch,t Max D. Summers,}: and Lee Courtney t
t Gulf Breeze Environmental Research Laboratory
United States Environmental Protection Agency
Sabine Island, Gulf Breeze, Florida 32561
and t Cell Research Institute
University of Texas
Austin, Texas 78712
INTRODUCTION
Certain enveloped, rod-shaped DNA viruses have long been known as
pathogens of insects under the descriptive term "nuclear polyhedrosis viruses." x
These viruses have been extensively and intensively studied since Berghold's2
early reports in 1947. Subsequent to Berghold's classic early studies, many
rod-shaped viruses associated with polyhedral inclusion bodies of a crystalline
nature have been described from different species of insects that represent
several orders of Insecta. At present, The International Committee on Nomen-
clature of Viruses places the nuclear polyhedrosis viruses of arthropods in
subgroup A under the genus or group name Baculovirus.* Prior to 1973, there
were no reports of viruses that resemble baculoviruses in animals other than
insects or mites. In 1973 and 1974, the first reports3- 4 were made of baculo-
virus-like particles and associated polyhedral inclusion bodies in a noninsect
arthropod host. The new host was the pink shrimp, Penaeus duorarum, from
Florida waters of the northern Gulf of Mexico. These reports indicated for the
Baculovirus group a host range extension into the arthropod class Crustacea.
In regard to specific characterization and identification of the shrimp virus,
it is pertinent to report that not all of Koch's postulates have been satisfied.
Koch's postulates, however, were meant to be used to show specificity of a
microorganism as an etiologic agent for a disease condition and not specifically
to determine phylogenetic affinity or identity of the microorganism. The latter
task (identification) includes determination of biologic, morphologic, chemical,
and physical characteristics. Much of our effort has gone into these determina-
tions for the shrimp virus. The first of Koch's postulates (that of association
or presence of a microorganism with a disease condition) has been satisfied
for patent virus infections in shrimp; that is, inclusion bodies and virions are
present in all patent infections that exhibit cytopathologic characteristics. The
second of Koch's postulates (that of isolation and pure culture of the micro-
organism) has not been satisfied for the shrimp virus and poses a severe problem
because of the lack of continuous cell cultures of crustacean tissues in which to
isolate and grow the virus. At present, we are attempting to use established
insect cell lines in which to grow the shrimp virus.
The baculoviruses have attracted much attention in recent years largely
because some microbiologists and entomologists consider these viruses to be
* Contribution 253, Gulf Breeze Environmental Research Laboratory.
528
277
-------
Couch et al: Baculovirus Infections 529
promising biologic control agents for numerous insect pests.5"7 The insect
baculoviruses have shown narrow host specificity,8 and all experimental attempts
so far to infect noninsect species with insect baculoviruses have failed.9
The purpose of the present paper is to consider, in light of our present
knowledge, the significance of the shrimp virus in regard to the ecology of
its crustacean host.
VIRAL EFFECTS
The capability for recognition of patent virus infections with light micros-
copy has made possible the harvesting of viral material from feral shrimp.
Patently infected shrimp are those in which hepatopancreatic cell nuclei show
hypertrophy (FIGURE 2) and in which many of these nuclei possess charac-
teristic virus-associated polyhedral inclusion bodies (PIBs) (FIGURE 3). In
producing these effects and in its fine structure, the shrimp Baculovirus is similar
to other well-described baculoviruses.
Aspects of the cytopathologic effects in shrimp have been described in detail
elsewhere.3' ' Here, however, it is of value to review certain changes induced
by the virus that reveal the extent of impact of the virus on shrimp hepato-
pancreatic cells. Nuclear hypertrophy, chromatin diminution and margination,
and nucleolar loss are the obvious signs of infection prior to the appearance of
the PIB in the nucleus. These signs are apparent in heavily infected shrimp
with both bright-field and phase-contrast microscopy (FIGURES 1—4). As cellular
infections progress in lightly to heavily infected shrimp, tetrahedra (PIBs) from
0.5 to 20 p.m in width appear in nuclei in numbers relative to infection intensity
(FIGURES 3 & 4).
Electron microscopy reveals both the striking fine structural changes that
occur in host cells and the structure of the PIBs and associated virions (FIGURES
5-7). The ultimate cytopathologic effect of the virus is destruction of the host
cell. Damage is obvious in loss of the cell's structural and functional integrity
and growth of the PIB to a size too great for cellular accommodation (FIGURE
5). TABLE 1 gives a list of cytopathologic alterations of infected hepato-
pancreatic cells visible with light and electron microscopy.
BIOCHEMICAL CHARACTERISTICS
The shrimp virus has an enveloped nucleocapsid that appears similar to
those of insect baculoviruses that have been characterized biochemically (FIGURE
7). Although not yet determined, nucleic acid of the virus is probably double-
stranded DNA, as is the case with other baculoviruses.1 Biochemical and
serologic investigations are underway to compare the nucleic acid, virus struc-
tural proteins, and inclusion body proteins of the shrimp virus to several species
of insect baculoviruses.
CELL CULTURE
A preliminary attempt has been made to introduce virus via ultrafiltrates
of infected hepatopancreas into tissue culture cells of Trichoplusia ni, Spodoptera
278
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530
Annals New York Academy of Sciences
FIGURES 1-4. Light photomicrographs of fresh squash preparation of shrimp
hepatopancreas. FIGURE 1 shows uninfected, normal cells nuclei (arrows) of hepato-
pancreas; note conspicuous nucleoli and chromatin; phase-contrast microscopy
(X275). FIGURE 2 illustrates early patent infection (black arrows); note nuclear
hypertrophy and loss of chromatin and nucleoli; white arrow points to early PIB
formation in a nucleus adjacent to basement membrane of hepatopancreatic acinus;
phase-contrast microscopy (X275). FIGURE 3 shows advanced patent infection with
many PIBs in hypertrophied nuclei and released from nuclei (black arrows); white
arrow points to normal nucleus; phase-contrast microscopy (X275). FIGURE 4 shows
PIBs free of cells; bright-field microscopy (x 250).
279
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Couch et al.: Baculovirus Infections
531
frugiperda, Aedes albopietus, and Culex solinarius. Unfortunately, results are
uncertain, but cytopathologic effects have been elicited in the Spodoptera and in
mosquito cells. The question whether the virus or a toxic effect from shrimp
protein caused the effect must be answered. No cytopathologic effects have been
observed in T. ni cells.
LABORATORY ENHANCEMENT OF VIRAL INFECTIONS
Though laboratory transmission of virus from shrimp to shrimp by feeding
has been somewhat successful, we are not yet able to depend consistently upon
FIGURE 5. Electron micrograph of patently infected shrimp hepatopancreatic cell.
PIB, polyhedral inclusion body; NE, nuclear envelope proliferation; VS, virogenic
stromata; V, rod-shrimp virion in edge of nucleoplasm. X4000.
feeding of infected tissues to shrimp as a major method to obtain large amounts
of virus. On several occasions, we have apparently increased the prevalence of
patent infections artificially by holding large samples of shrimp (with initial
low prevalence of virus = 0-10% ) in small aquaria for up to 40 days. Under
crowded, stressful conditions, shrimp with latent infections may develop patent
infections, and uninfected shrimp probably become infected by feeding upon
carcasses of infected shrimp in these aquaria.
Chemical stress of pink shrimp by laboratory exposures to low levels of
organochlorines» may increase prevalence of patent infections. This effect
280
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532
Annals New York Academy of Sciences
.** k*V
:f.-i,:i/$*-i**S>i-.~y*.
FIGURE 6. Virions, virogenic stages, and PIB in nucleus of patently infected cell.
N, nucleoplasm; V, virions in cross and longitudinal sections; Cy, cytoplasm of cell;
note many free ribosomes in cytoplasm, x 38,500.
FIGURE 7. Higher magnification of virions at edge of and partially occluded in
PIB. X 67,800.
281
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Couch et al: Baculovirus Infections
533
can only really be confirmed, however, in highly controlled in vivo and in vitro
shrimp-virus systems.
DISTRIBUTION AND PREVALENCE OF VIRUS IN NATURE
To date, shrimp have been sampled from waters of the northern Gulf of
Mexico and' estuaries from Pensacola, Florida eastward to Apalachee Bay near
Keaton Beach, Florida. FIGURE 8 shows approximate regions in which virus-
infected shrimp have been taken. Only the pink shrimp has shown consistently
recoverable natural infections. A single, adult brown shrimp (Penaeus aztecus),
taken from Escambia Bay, near Pensacola, Florida, was found moderately
infected in 1974. White shrimp (Penaeus setiferus) and grass shrimp (Paleo-
TABLE 1
CYTOPATHOLOGJC EFFECTS OF VIRUS ON SHRIMP HEPATOPANCREATIC CELLS
AS REVEALED BY LIGHT AND ELECTRON MICROSCOPY
Cytopathologic
Effect
Light
Microscopy
Electron
Microscopy
Nuclear hypertrophy
Chromatin diminution
Chromatin margination
Nucleolar loss
Inclusion body
Nuclear membrane proliferation
(membranous labyrinth)
Myeloid bodies in cytoplasm
Increase in free ribosomes
Reduction in number of mitochondria
Changes in nucleoplasm
Fibrillar stroma
Granular stroma
Rod-shaped virions in nucleoplasm
monetes spp.) have not yet been found infected. Other Crustacea, such as blue
crabs (Callinectes sapidus), stone crabs (Menippe mercenaria), and mud crabs
(Panopeus sp. and Neopanope sp.), have been found not to harbor the virus.
The prevalences of patently infected pink shrimp in samples taken peri-
odically from various locales along the northern Gulf Coast of Florida since
1970 are given in TABLE 2. Samples taken from Gulf waters near Keaton
Beach (Apalachee Bay) have shown the highest prevalence of virus. Our data,
to date, indicate no particular seasonal intensification of virus prevalence,
although October and January have been productive months for obtaining
larger numbers of infected shrimp. Original sample sizes may have influenced
this abundance (see TABLE 2).
It is noteworthy that we have not examined specimens from the epicenter
of the pink shrimp's geographic distribution in the Gulf of Mexico near the
Dry Tortugas and Key West, Florida. These waters maintain the highest known
282
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Annals New York Academy of Sciences
TABLE 2
NUMBER OF PINK SHRIMP EXAMINED AND PATENT VIRUS INFECTIONS SINCE 1970
Year
1970
1971
1973
1974
1975
Totals
Number
Examined
40
1
42
14
14
10
20
20
42
40
28
30
53
20
50
88
23
15
350
145
460
298
98
435
90
2426
Number
Patently
Infected
12
1
0
7
0
0
0
0
12
6
10
0
14
4
11
9
3
0
55
4
0
0
0
62
0
210
Month
June
August
July
August
September
October
June
August
Ocotber
November
November
January
January
February
March
May
August
September
October
October
November
December
December
January
January
Source
(all in Florida)
Keaton Beach
Pensacola
Pensacola
Keaton Beach
Keaton Beach
Keaton Beach
Pensacola
Pensacola
Keaton Beach
Pensacola
Port St. Joe
Pensacola
Keaton Beach
Keaton Beach
Keaton Beach
Keaton Beach
Keaton Beach
Keaton Beach
Keaton Beach
Port St. Joe
Pensacola
Apalachicola
Pensacola
Keaton Beach
Pensacola
pink shrimp densities according to catch per unit of effort of the shrimp fishery.10
Though pink shrimp sustain a fishery in northern Gulf waters, a study of the
virus in more dense populations off southwest Florida should be additionally
informative as to the epizootic behavior of the virus in nature. Presence of the
virus in southwest Florida pink shrimp is probable because that population
merges with the northern Gulf Coast population.
ENVIRONMENTAL SIGNIFICANCE OF SHRIMP VIRUS
Three questions should be considered in regard to discovery of a Baculovirus
in a marine arthropod. The first is obvious: What is the direct effect of the
virus on its natural, feral shrimp host? Thus far, laboratory studies of the
shrimp-virus system have not given results that would allow us to predict the
effect on pink shrimp populations in nature. There is little doubt that severe
cytopathologic effects occur in hepatopancreas of infected individuals. We have
some evidence at this time that the virus may cause epizootic mortalities in
feral shrimp. Dying shrimp have been found to be heavily infected in laboratory
283
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Couch et al: Baculovirus Infections
535
aquaria and large holding tanks. However, other shrimp in the same samples
have been found dying with no signs of patent Baculovirus infections. Fishery
reports10 indicate that unexplained fluctuations in pink shrimp abundance occur
regularly in the northern Gulf of Mexico waters from which virus-infected
shrimp have been taken. These fluctuations may be due to any number of
causes, but certainly the shrimp virus may be considered as one of the candi-
dates.
The second question is: Are there interactions between stress factors, such
as chemical pollutants, and virus infections in shrimp? Further studies must
be completed to answer this question. Tests completed to date suggest that the
polychlorinated biphenyls may increase prevalence of patent virus infections
in test shrimp, whereas other chemicals (methoxychlor) may not.11 Interactions
between natural pathogens and pollutant chemicals may become more apparent
in aquatic animals as further studies are completed on chronically polluted
estuaries and marine waters. The concept that pollutants act as stressors to
lower natural resistance to disease should be explored further with such systems
as the shrimp-virus.
The third question is: What are the risks, if any, of not better understanding
host specificity in regard to developing viral groups, such as the Baculovirus
group, for insecticidal uses? Though there appears to be little danger of arti-
ficially introducing insect viruses into nontarget species, this question may not
have been answered satisfactorily at this time. The discovery of a Baculovirus
in a marine crustacean suggests that host limitations for this virus group have
not been determined absolutely.
FIGURE 8. Chart that shows areas in northern Gulf of Mexico where virus-infected
shrimp have been found; stars indicate approximate sites of collection.
284
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536 Annals New York Academy of Sciences
SUMMARY
Pertinent questions must be answered concerning the significance of the
discovery of a new Baculovirus enzootic in populations of penaeid shrimp in
the northern Gulf of Mexico. The virus is rod shaped, both free and occluded
in polyhedral inclusion bodies in the nuclei of host hepatopancreatic cells, and
is associated with striking cytopathologic effects, but induces no specific gross
signs. Samples of pink shrimp taken since 1971 have shown prevalences of
from 0 to 50% (shrimp with patent infections/total number of shrimp in
sample). The virus has been found in samples of shrimp taken from waters
of Apalachee Bay, Port St. Joe, and Pensacola, all in Florida. Attempts to
culture the shrimp virus in established insect cell lines are underway; therefore,
not all of Koch's postulates have been satisfied for the virus. In the laboratory,
virus prevalence in samples of shrimp has been increased by holding the shrimp
in large numbers in small aquaria. The major questions that we are attempting
to answer about the crustacean Baculovirus are:
What direct effect does the virus have on populations of shrimp in nature?
Do pollutant chemicals found in coastal waters enhance the virus effect in
shrimp?
What relationship, if any, does the occurrence of a Baculovirus in a crus-
tacean have with the development of insect baculoviruses as potential biopesti-
cides?
Other important avenues of investigation have opened. The opportunity has
appeared for virologists working with insect baculoviruses to compare these
viruses with a Baculovirus from a noninsect host.
REFERENCES
1. WILDY, P. 1971. Classification and nomenclature of viruses. First report of the
international committee on nomenclature of viruses. Mon. Virol. 5: 1—81.
2. BERGHOLD, G. H. 1947. Die Isolierung des Polyeder-Virus und die Natur der
Polyeder. Z. Naturforsch. 2b: 122-143.
3. COUCH, J. A. 1974. Free and occluded virus, similar to Baculovirus, in hepa-
topancreas of pink shrimp. Nature (London) 247: 229-231.
4. COUCH, J. A. 1974. An enzootic nuclear polyhedrosis virus of pink shrimp:
ultrastructure, prevalence, and enhancement. J. Invert. Pathol. 24: 311-331.
5. JAQUES, R. P. 1970. Application of viruses to soil and foliage for control of the
cabbage looper and imported cabbage worm. J. Invert. Pathol. 15: 328-340.
6. HALL, I. M. 1963. Microbial control. In Insect Pathology, An Advanced
Treatise. E. A. Steinhaus, Ed. Vol. 2: 477-517. Academic Press, Inc. New
York, N.Y.
7. TANADA, Y. 1956. Microbial control of some lepidopterous pests of crucifers.
I. Econ. Entomol. 49: 320-329.
8. IGNOFFO, C. M. 1968. Specificity of insect viruses. Bull. Entomol. Soc. Amer.
14: 265-276.
9. LIGHTNER, D. V., R. R. PROCTOR, A. K. SPARKS, J. R. ADAMS & A. M. HEIMPEL.
1973. Testing penaeid shrimp for susceptibility to an insect nuclear polyhe-
drosis virus. Environ. Entomol. 2: 611-613.
10. ANONYMOUS. 1969. Gulf of Mexico shrimp atlas. Circular 312. U.S. Depart-
ment of the Interior, Bureau of Commercial Fish.
11. COUCH, J. A. 1975. Attempts to increase Baculovirus prevalences in shrimp by
chemical exposure. In Progress in Experimental Tumor Research. F. Horn-
burger, Ed. Vol.19. S. Karger. Geneva, Switzerland.
23976
285
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Reprinted from Proceedings of
the Third International Bio-
degradation Symposium, J.M.
Sharpley and A.M. Kaplan, eds.
pp. 93-98, with permission of
Applied Science Publishers,
Ltd., London, 1976
MICROBIAL POPULATIONS IN COASTAL SURFACE SLICKS
S.A. Crow, W.L. Cook, D.G. Ahearn, and A.W. Bourquin
Contribution No. 254
287
-------
MICROBIAL POPULATIONS IN COASTAL SURFACE SLICKS *f
S. A. CROW, W. L. COOK, D. G. AHEARN
Department of Biology, Georgia State University,
Atlanta, Georgia 30303, U.S.A.
and
A. W. BOURQUIN
U.S. Environmental Protection Agency, Gulf Breeze Environmental
Research Laboratory, Sabine Island, Gulf Breeze, Florida 32561, U.S.A.
Summary
Samples of the upper 10 /mi of inshore surface films obtained by adsorption
to membranes yielded microbial populations up to 108 ml"1 or 105 cm~2.
These populations were typically 10-100 times greater than those in under-
lying waters at a depth of 10 cm. Predominant bacteria in surface films were
motile, nonpigmented, gram-negative rods. Colony-forming units of yeasts
and moulds were found in concentrations to 104 ml"1 or 28 cm"2. The
predominant species in surface films were proteolytic and amylolytic but
exhibited only weak to negligible hydrocarbonoclastic and lipolytic activities.
A greater proportion of the surface film bacteria, as compared to those at
10 cm depth, were capable of growth on fresh-water media.
INTRODUCTION
Surface films are unique microbial habitats occurring at the air-water interface
of aquatic systems. They are a common phenomenon in coastal waters and
probably occur in ocean waters at most times, depending on the state of the
seas. The production of most surface slicks appears to be related to the decay
of naturally occurring aquatic organisms (Babkov, 1965). Numerous
investigators (Baier, 1970, 1972; Ewing, 1950; Garret, 1965; Sutcliffe et al,
1963) have examined the physical properties of surface slicks but their
chemical nature is unclear. Surface films have been shown to contain high
concentrations of organic carbon, nitrogen, phosphorus (Williams, 1967),
alkanes (Ledet and Laseter, 1974), and chlorinated hydrocarbons (Seba and
Corcoran, 1969).
The few studies of microbial populations in surface films have indicated
that they contain high densities of bacteria relative to underlying waters.
Sieburth (1965) reported bacterial populations up to 4 x 104 ml"1 in
surface films. The predominant bacteria were pseudomonads which ex-
pressed lipolytic activity. Harvey (1966) found that bacteria, small algae, and
colourless flagellates were concentrated in the upper 60 /j.m of surface water.
In a preliminary study, Crow et al. (1975) found concentrations of total
* Supported in part by Office of Naval Research contract ONR NOOO-14-71-C-0145 and
Environmental Protection Agency contract R 803141-01-0.
t Gulf Breeze Environmental Research Laboratory Contribution No. 254.
93
289
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94 5. A. Crow, W. L. Cook, D. G. Ahearn and A. W. Bourquin
heterotrophs up to 108 ml"1 in coastal surface films. This report examines the
microbial populations of surface films occurring in coastal waters of the
Gulf of Mexico.
MATERIALS AND METHODS
Surface slick samples were collected by a membrane adsorption technique
(Crow et al, 1975). Sterile polycarbonate membranes (Nuclepore®) 47 mm
in diameter with a porosity of 0-4 /mi were floated on surface waters near
Barataria Bay, Louisiana, and along the western coast of Florida near Sabine
Island. The membranes with adhering surface film and underlying water
were collected with sterile plastic dishes. The membrane was removed from
the dish with sterile forceps and either implanted directly on suitable media
or placed in 100 ml of sterile seawater. The sea water was agitated vigorously,
serially diluted, and various dilutions plated on suitable media. Numbers of
microorganisms were calculated on a per-ml basis using a sampling volume
of 5-9 n\ per membrane and for surface area using 17-3 cm2 per membrane
(Crow et al., 1975). Subsurface samples were collected from a depth of 10 cm
with sterile 30 ml disposable syringes fitted with extension tubes.
Total heterotrophic bacterial populations were enumerated using Marine
agar 2216 (Difco). Fungal populations were determined with Mycological
agar (Difco) prepared with natural seawater and adjusted to pH 4-5 with
0-1 N HC1. For select samples, populations of proteolytic bacteria were
enumerated on yeast extract (0-01 %)-skim milk (2-0%) agar prepared in
artificial seawater of 20°/00 salinity. Hydrocarbonoclastic bacteria were
enumerated according to the method of Gunkel (1973). In this method a
basal medium containing NH4C1 0-005%, K2HPO4 0-005%, Na2HPO4
0-01 %, and 2% Louisiana crude oil is inoculated with tenfold dilutions of
seawater. Utilization of the hydrocarbon is based on the occurrence of
turbidity following acidification for the dissolution of inorganic precipitate.
All cultures were incubated at 22-25°C; broth cultures were incubated with
agitation.
Representative microorganisms, unless otherwise stated, were characterized
physiologically with media prepared according to the formula of Colwell and
Wiebe (1970). Proteolysis was determined using 2-0% skim milk and 0-1%
yeast extract in 1-7% agar and with Thioglycollate gelatin medium (Difco)
prepared with artificial seawater. Oxidative or fermentative carbohydrate
metabolism was determined with MOF medium (Difco). Lipase and urease
activities were determined on Spirit Blue agar (Difco) and Urea agar (Difco),
respectively. Both were prepared with artificial seawater at 20°/00 salinity.
Microorganisms were tested for pristane, decane, tetradecane, ethanol, and
crude oil utilization according to previously described methods (Ahearn et al.,
1971). Studies of pesticide interference with ethanol and hexadecane metab-
olism were conducted using selected ethanol-positive bacteria, yeasts, and
filamentous fungi. Methoxychlor dissolved in ethanol was added to tubes of
yeast extract broth (0-01 %) to obtain a final ethanol concentration of 2%
® Registered Trademark. Mention of commercial products or trade names does not
constitute endorsement by the U.S. Environmental Protection Agency.
290
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Microbial Populations in Coastal Surface Slicks 95
and a concentration of 100 ng methoxychlor ml~1 of medium. Heptachlor
was dissolved in hexadecane to give a final concentration of about 0-1 /ig
ml"1. Tubes were incubated on roller drums at 20-25°C for 5 days. Cultures
were examined at 24-hr intervals and turbidity was compared to controls to
determine any interference with ethanol metabolism.
To ascertain the presence of pesticides in surface slicks, large membranes
(293 mm in diameter with a 0-4 jum porosity) were used to adsorb surface
slick material. These membranes were then placed in 250-ml Erlenmeyer
flasks and extracted with 100 ml of pesticide-grade petroleum ether; 10 ml of
this material was concentrated 10 times and used for gas chromatographic
(GC) analysis. Analyses were performed on a Varian Aerograph 2100 fitted
with an electron-capture detector on a 0-64 cm x 1-8 m glass column
containing 2% OV-101 on Gas Chrom Q 100/120.
RESULTS
The densities of microorganisms in coastal and estuarine surface slicks
ranged as high as 2-5 x 10s cm~2. Typically, underlying waters contained
substantially lower populations (Table I). The predominant bacteria were
Table I
CONCENTRATION RANGE OF MICROORGANISMS IN SURFACE SLICKS AND
UNDERLYING WATERS
Organisms
Bacteria
Yeasts
Moulds
Surface slick
(w/-i) (cm -2)
105-1Q8 28-2-5 x 105
102-1Q4 0-5
103-1Q4 0-28
Subsurface
(ml -i)
102-106
102
102
No.
samples
24
16
16
motile, nonpigmented, gram-negative rods, presumptively identified as
pseudomonads. Samples from surface films and from 10 cm depths appeared
to contain similar proportions of chromogenic bacteria. The most common
fungi were isolates of the genera Aureobasidium, Candida, Cephalosporium,
and Cladosporium. The proportions of hydrocarbon-utilizing and proteolytic
bacteria in the total populations at two geographically adjacent stations are
presented in Table II. Neither of these stations were subjected to heavy
Table II
AVERAGE NUMBER OF PROTEOLYTIC AND HYDROCARBONOCLASTIC
BACTERIA IN SURFACE SLICKS IN FOUR SAMPLES FROM FLORIDA
COASTAL WATERS
„ . Colony-forming unit si cm2
aactena Station 1 Station 2
Proteolytic
Hydrocarbonoclastic
Total heterotrophic
2-0 x 103
3-0 x 102
2-5 x 105
1-0 x 10
2-5 x 102
1-0 x 104
hydrocarbon pollution; however, station 2, an enclosed saline pond, received
laboratory wastes containing pesticides including methoxychlor and hepta-
291
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96 S. A. Crow, W. L. Cook, D. G. Ahearn and A. W. Bourquin
chlor. GC analyses of extracts of membranes retrieved from surface slicks at
both of these stations indicated substances compatible with chlorinated
pesticides. Surface slicks from the enclosed pond yielded fewer total hetero-
trophic bacteria and markedly reduced numbers of proteolytic bacteria.
Hydrocarbonoclastic bacteria were present at both stations, but fewer
organisms from station 2 exhibited intense growth and none significantly
emulsified the oil. Only five isolates of 20 selected cultures from both stations,
giving an indication of growth on crude oil, grew well in yeast extract broth,
with decane and tetradecane as sole carbon sources. Of these five isolates,
four also utilized pristane. Virtually all the yeast isolates representative of
Candida and Rhodotorula grew on the hydrocarbons.
Twenty-one bacteria, representative of the predominant morphological
types found at all collection sites, were physiologically characterized (Table
III). All organisms were initially isolated on and produced good growth on
Table III
PHYSIOLOGICAL CHARACTERISTICS OF 21 MORPHOLOGICALLY DIFFERENT
BACTERIA FROM SURFACE SLICKS
_, No. bacteria
Character
Proteo lysis
Skim milk
Gelatin
HaS production
Amylolysis
Lipolysis
Citrate utilization
Urease activity
Ethanol utilization
Crude oil utilization
MOF reactions
oxidative
fermentative
no change
Gram reaction
Motility
Growth on freshwater medium
Morphology
10
21
0
11
0
3
2
15
1
14
5
3
8
15
13
rods
8
0
13
8
21
16
18
6
20
4
13
0
11
6
8
3
0
8
1
0
2
1
0
0
0
0
0
2
0
0
seawater medium (20°/00 salinity); nevertheless, 13 of the 21 strains grew well
in yeast extract broth prepared with fresh water. None of the isolates showed
lipolysis and only a single isolate grew on crude oil.
Neither methoxychlor nor heptachlor were found to alter the metabolism
of ethanol or hexadecane by representative surface slick microorganisms.
The concentrations of these pesticides which could be reclaimed by hexane
extraction from several yeast and bacterial cultures were reduced in several
instances in comparison with uninoculated control flasks by about 25%.
Preliminary experiments with C1 ^labelled heptachlor indicated the radio-
label was bound to the cells and not extractable with the procedure employed.
Known biodegradation products of both pesticides were not demonstrated in
the culture systems.
292
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Microbial Populations in Coastal Surface Slicks 97
DISCUSSION
Microbial populations associated with surface films on coastal waters of the
Gulf of Mexico were found to be higher than those reported for films at other
oceanic sites collected with either a wire screen or a hydrophilic drum
(Sieburth, 1965; Harvey, 1966; Gunkel, T973). The reports of these investi-
gators that microorganisms were concentrated in the surface films as con-
trasted to underlying waters were substantiated. Fungi, moulds in particular,
occurred mainly in the surface film. In slicks from the eastern Gulf of Mexico
coastal waters, hydrocarbonoclastic and lipolytic bacteria appeared un-
common. Sieburth (1965) reported that 95 % of surface slick isolates, mainly
pseudomonads, from the Pacific Ocean were lipolytic. This anomaly is most
likely a reflection of the diverse nature of surface films (Maclntyre, 1974).
The ability of most of our isolates to grow on fresh-water media suggests
they are of terrestrial origin. Preliminary analysis with culture systems has
not demonstrated an effect of heptachlor or methoxychlor on the metabolism
of hexadecane or ethanol by representative slick microorganisms. In other
studies (Smith et al., 1976), heptachlor was shown to enhance or inhibit
hexadecane utilization by Candida maltosa from a fresh-water oil slick
dependent upon aeration and pesticide concentration. The binding of pesti-
cides to cells in culture systems suggests that the presence of chlorinated
hydrocarbons in surface films (Seba and Corcoran, 1969) may be related to
the microbial densities of these slicks.
REFERENCES
Ahearn, D. G., Meyers, S. P., and Standard, P. G. (1971). The role of yeasts in the
decomposition of oils in marine environments. Dev. Ind. MicrobioL, 12, 126-134.
Babkov, A. I. (1965). The causes of slicks on the surface. Oceanologiya, 5, 102-104.
Baier, R. E. (1970). Surface quality assessment of natural bodies of water. Proc.
Great Lakes Res. 13th, pp. 114-127.
Baier, R. E. (1972). Organic films on natural waters: their retrieval, identification,
and modes of elimination. /. Geophys. Res., 77, 5062-5-75.
Colwell, R. R., and Wiebe, W. J. (1970). 'Core' characteristics for use in classifying
aerobic, heterotrophic bacteria by numerical taxonomy. Bull. Georgia Acad. Sci.,
28, 165-185.
Crow, S. A., Ahearn, D. G., Cook, W. L., and Bourquin, A. W. (1975). Densities of
bacteria and fungi in coastal surface films as determined by a membrane-
adsorption procedure. Limnol. Oceanogr., 20, 644-645.
Ewing, G. (1950). Slicks, surface films and internal waves. /. Mar. Res., 9, 161-187.
Garret, W. D. (1965). Collection of slick-forming materials from the sea surface.
Limnol. Oceanogr., 10, 602-605.
Gunkel, W. (1968). Bacteriological investigations of oil-polluted sediments from the
Cornish coast following the 'Torrey Canyon' disaster. In The Biological Effects of
OilPollution onLittoral Communities, pp. 151-158 (J. D. Carthy and D. R. Arthur,
eds.), Field Studies Council, London.
Gunkel, W. (1973). Distribution and abundance of oil-oxidizing bacteria in the
North Sea. In The Microbial Degradation of Oil Pollutants, pp. 127-139 (D. G.
Ahearn and S. P. Meyers, eds.), Center Wetland Resources, Louisiana State
University, Publication LSU-SG-73-01.
Harvey, G. W. (1966). Microlayer collection from the sea surface: a new method and
initial results. Limnol. Oceanogr., 11, 608-613.
293
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98 S. A. Crow, W. L. Cook, D. G. Ahearn and A. W. Bourquin
Ledet, E. J., and Laseter, J. L. (1974). AJkanes at the air-sea interface from offshore
Louisiana and Florida. Science, 186, 261-263.
Maclntyre, F. (1974). The top millimeter of the ocean. Sci. Am., 230, 62-77.
Seba, D. B., and Corcoran, E. F. (1969). Surface slicks as concentrators of pesticides
in the marine environment. Pestic. Monit. J., 3, 190-193.
Sieburth, J. McN. (1965). Bacteriological samplers for air-water and water-sediment
interfaces. Trans. Jt. Conf. Ocean Sci. Ocean Eng. Mar. Technol. Soc. Am. Soc.
Limnol. Oceanogr., pp. 1064-1067.
Smith, G. N., Bourguin, A. W., Crow, S. A., and Ahearn, D. G. (1976). The effect
of heptachlor on hexadecane utilization by selected fungi. Dev. Ind. Microbiol.,
17, 331-336.
SutclifTe, W. H. Jr., Baylor, E. R., and Menzel, D. W. (1963). Sea surface chemistry
and Langmuir circulation. Deep-Sea Res., 10, 233-243.
Williams, P. M. (1967). Sea surface chemistry: organic carbon and organic and
inorganic nitrogen and phosphorus in surface films and subsurface waters.
Deep-Sea Res., 14, 791-800.
294
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Reprinted from Developments
in Industrial Microbiology,
Vol. 17: 331-336, 1976,
with permission of the Society
for Industrial Microbiology
EFFECT OF HEPTACHLOR ON HEXADECANE UTILIZATION
BY SELECTED FUNGI
N.G. Smith, A.W. Bourquin, S.A. Crow, and D.G. Ahearn
Contribution No. 255
295
-------
Reprinted from Volume 17 of DEVELOPMENTS IN INDUSTRIAL MICROBIOLOGY
A Publication of the Society for Industrial Microbiology
CHAPTER 36
Effect of Heptachlor on Hexadecane Utilization by Selected Fungi*
N. G. SMITH, A. W. BOURQUIN,** S. A. CROW, AND D. G. AHEARN
Department of Biology, Georgia State University, Atlanta, Georgia 30303 and
**U.S. Environmental Protection Agency, Gulf Breeze Environmental Research Laboratory,
Sabine Island, Gulf Breeze, Florida 32561
Various concn of heptachlor dissolved in hexadecane were added to cultures of fungi grown in
yeast-nitrogen base prepared with synthetic seawater and with deionized water. Candida maltosa
and Candida lipolytica showed greatest utilization of hexadecane (20-91%) whether heptachlor
was present or absent. Isolates of Pichia spartinae, Cladosporium sp., Cephalosporium sp., and
PenicUlium sp. also utilized the hydrocarbon, but to a lesser extent. Species of Kluyveromyces
failed to grow with hexadecane as a carbon source. Compared with low concn, high concn of
heptachlor appeared to have a slight stimulating effect on utilization of hexadecane by
C. maltosa, but had no effect with C. lipolytica.
INTRODUCTION
Microorganisms have been shown to be present in high densities in marine surface slicks
(Sieburth 1965; Maclntyre 1974; Crow et al. 1975), a habitat also known to contain relatively
high concn of pesticides (Seba and Coccoran 1969; Parker and Barson 1970). Numerous
investigations have been concerned with the biodegradation of pesticide molecules (Brooks
1974), but little effort has been directed toward understanding the effect of pesticides on
microbial metabolism. The insolubility of chlorinated hydrocarbons in aqueous systems has
hampered this type of investigation. To establish base line data on the interactions of
pesticides on microbial activities in surface slicks, we examined the effect of the chlorinated
hydrocarbon pesticide, heptachlor, on hexadecane metabolism by selected fungi, using the
carbon source as the solvent for the pesticide.
MATERIALS AND METHODS
The fungi and their sources are listed in Table 1. Inoculum for all tests was 0.1 ml of a cell
suspension grown on 0.1% (w/v) dextrose and 0.67% (w/v) yeast-nitrogen base (YNB, Difco1)
in deionized water for 48 h at 25 C.
Concentrations of heptachlor (analytical reference standard, 99.8% Velsicol) to 1 mg/ml
were dissolved in reagent grade hexadecane (Aldrich Chem. Co.). Nitrogen and other essential
nutrients were supplied in YNB. Media were prepared with deionized water and with artificial
seawater (30 °/0o salinity) (Instant Ocean, Aquarium Systems Inc.). In all cases, hexadecane
*GBERL Contribution No. 25S
Mention of trade names or commercial products does not constitute endorsement by the U.S.
Environmental Protection Agency.
297
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332
TABLE 1. Source of isolates
N. G. SMITH ET AL.
Species
Candida lipolytica
Candida maltosa
Candida maltosa
Candida maltosa
Cephalosporium sp.
CJadosporium sp.
Pichia spartinae
Pichia spartinae
Penicillium sp.
Culture No.
37-1
AJ4476
CBS5611
R-42
F2
Fl
PI
N18
P2
Source
Frankfurter
Air
Monosodium glutamate fermentation
Fresh-water oil slick
Estuarine water
Estuarine water
Estuarine water
Estuarine water
Air
was present at a final concn of 2%. Tubes with 5 ml of medium were inoculated, capped with
steel closures, and incubated at 25 C for 2 wk on a roller drum at 75 rpm. Flasks containing
25, 250, or 500 ml of media were inoculated and shaken at 200 rpm for 2 or 4 wk on a rotary
shaker at 25 C. Relative growth of selected cultures was determined spectrophotometrically
with a Bausch and Lomb Spectronic 20 at 580 nm.
Samples for gas chromatographic analysis (Varian Aerograph 2100) were extracted in their
culture container with two volumes of petroleum ether (pesticide grade) to one volume of
culture. Hexadecane and high concn of heptachlor were determined with a flame ionization
detector with a 0.64 cm x 1.8 m glass column packed with 2% SE-30 Gas Chrom Q 100/120.
Low concn of heptachlor and its metabolites were determined by electron capture detection
with a 0.64 cm x 1.8 m glass column packed with 2% OV-101 on Gas Chrom Q 100/120.
Cultures labeled with 1 fid of 14C -heptachlor (2.5 mg/5 ml hexadecane in 500ml
medium) were sampled for determination of cellular incorporation, residual pesticide, and
14C02. Samples were counted with a Beckman LS 250 scintillation counter. Protosol
(1.0 ml/15 ml scintillation cocktail, Beckman Inc.) was added to samples containing
particulate matter. Labile C02 was trapped in 40% KOH (Atlas and Bartha 1972).
RESULTS
All fungi remained viable in the hexadecane medium and, with the exception of
Kluyveromyces spp., produced discernible growth in both fresh- and salt-water media
(Table 2). The isolates of Cladosporium and Cephalosporium first produced sparse yeast-like
growth in the hexadecane medium and then, similar to the Penicillium sp., formed surface
mats which precluded accurate growth determinations by OD.
The higher concn of heptachlof appeared to stimulate hexadecane utilization (Table 3) by
C maltosa in test-tube experiments with medium prepared with deionized water. Such
stimulatory effect was not observed in the salt-water medium. Between 71 and 100% of
heptachlor was recovered from these systems. Additional strains of C. maltosa were examined
for the effect of heptachlor on hexadecane utilization (Table 4). In experiments employing
500 ml of medium incubated 14 days with constant agitation, there was a marked increase in
hexadecane utilization as compared to test-tube experiments. However, there was no apparent
effect of heptachlor on hexadecane utilization. Significant hexadecane utilization (> 60%)
was accompanied by a marked decrease in percentage of heptachlor recovered. Analysis of
culture broths of C lipolytica and C. maltosa for known degradation products of heptachlor
demonstrated the presence of increased concn of 1-hydroxychlordene (Table 5). Formation of
this degradation product occurs by hydrolysis (Bourquin et al. 1972) and may be mediated by
bacteria (Miles et al. 1969).
298
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CONTRIBUTED PAPERS 333
TABLE 2. Average growth (O.D. ± 0.11 at 580 nm) of selected fungi on hexadecane in yeast-nitrogen base
broth in four tests
Organism
Deionized Water:
Candida maltosa R42
Candida lipolytica
Pichia spartinae PI
Pichia spartinae N18
Cephalosporium sp.
Cladosporium sp. .
Seawater:
Candida maltosa
Candida lipolytica
Pichia spartinae PI
Pichia spartinae N18
3
1.25
0.96
0.00
0.00
0.00
0.00
0.84
0.67
0.00
0.00
Incubation (days)
8
2.00
2.00
0.73
1.25
0.00
0.00
2.00
2.00
0.00
0.00
17
2.00
2.00
2.00
2.00
0.19
0.01
2.00
2.00
0.67
0.28
TABLE 3. Effect ofheptachlor on hexadecane utilization
Heptachlor, mg/ml Hexadecane
Species
Deionized water :
Candida maltosa R42
Candida lipolytica 37-1
Pichia spartinae PI
Pichia spartinae N18
Control
Seawater:
Candida maltosa R42
Candida lipolytica 37-1
Pichia spartinae PI
Pichia spartinae N18
Control
0
71a
78
86
88
96
75
79
94
95
100
0.096
58a
75
84
87
100
81
76
96
89
99
0.96
65a
89
85
93
93
79
82
90
92
99
aPercentage recovery of hexadecane after 14 days' growth in capped test tubes.
TABLE 4. Effect ofheptachlor on hexadecane utilization by different strains of Candida maltosa
Species
Deionized water:
Control
Candida maltosa R42
Candida maltosa AJ4476
Candida maltosa CBS5611
Seawater :
Control
Candida maltosa R42
Candida maltosa AJ4476
Candida maltosa CBS5611
Hexadecane
Pesticide
100
10
64
93
100
64
77
89
Recovery3
No Pesticide
99
13
45
78
100
57
76
86
Heptachlor
Recoverya
100
33
91
117
100
72
120
120
Percentage recovery as compared to control.
299
ko.96 mg heptachlor/ml hexadecane.
-------
334 N. G. SMITH ET AL.
TABLE 5. Average recovery ([Jg/mty of l-hydroxychlordene from cultures of yeasts grown on
hexadecane-heptachlor media in duplicate tests
Organism
Sea water:
Control
Candida maltosa R42
Candida lipolytica 37-1
Deionized water:
Control
Candida maltosa R42
Candida lipolytica 37-1
Heptachlor,
0.096 mg/ml
Hexadecane
0.011
0.022
0.055
0.017
0.038
0.134
Heptachlor,
0.96 mg/ml
Hexadecane
0.103a
0.069
0.295
0.087
0.162
0.667
One sample only; no metabolite in duplicate.
TABLE 6. Effect of varied concn ofheptachlor on hexadecane utilization by species of Candida
Heptachlor Concn
(mg/ml)
0.00
0.01
0.10
1.00
Hexadecane Recovery3
C. maltosa (R-42) C.
23
73
65
55
lipolytica (37-1)
56
57
58
55
Percentage recovered from 25-ml cultures incubated 14 days in 125-ml flasks.
Preliminary experimentation indicated that labeled CO2 was not released by C. lipolytica
37-1 and C. maltosa R-42 during growth on hexadecane in the presence of radioactive
heptachlor. In experiments in which 50% of the hexadecane was utilized, approximately 80%
of the residual label was detected in the petroleum ether extracts and was approximately the
percentage of heptachlor recovered. The remaining radiolabel (approximately 15%) was found
to be associated with the cell slurry which had remained in the aqueous "phase. Both
C maltosa and C lipolytica were grown in shake cultures in the presence of increasing concn
of heptachlor (Table 6). Best utilization of hexadecane by C. maltosa was achieved in the
absence of heptachlor; however, better utilization of hexadecane was found in the presence of
heptachlor, at the higher concn. As shown in the previous experiments (Table 3) heptachlor
had no effect on C lipolytica.
DISCUSSION
In previous investigations of the effect ofheptachlor on microorganisms, the contact between
the microorganisms, the substrate, and the pesticide did not reach the level afforded by our
culture system. In most studies, heptachlor is added as a suspension, with a solvent such as
ethanol (Miles et al. 1969) or adsorbed to a solid medium (Shamiyeh and Johnson 1973). In
such systems, there is some question concerning (a) the fate of the pesticide as the solvent
evaporates or 1S diluted by the medium, and (b) how homogeneously it is distributed in an
agar medium. In the heptachlor-hexadecane system, the pesticide is uniformly distributed
throughout the substrate and the organisms must come in contact with the pesticide during
300
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CONTRIBUTED PAPERS 335
degradation of the carbon source. Another asset of this system is that the maximum concn of
the pesticide is greatly increased without fear of precipitation or unequal distribution.
With incubation of 2 to 4wk, fungi utilized from 5% to 91.6% of the hexadecane,
depending on cultural conditions. The volume of the culture flask, surface area, depth of
hydrocarbon surface layer, and aeration all affected the degree of hexadecane utilization.
Species of Candida most readily oxidized and emulsified hexadecane. Isolates of C. maltosa
utilized from 22% to 32% of the hexadecane when grown in test-tube culture and up to 91.6%
when grown in flasks. C. lipolytica 37-1 utilized approximately 22% of the hexadecane in test
tubes and up to 54% in flasks. Pichia species utilized only a small amount of hexadecane
(5-16%) and the filamentous fungi consistently utilized less than 10% during a 14-day period.
The species of Kluyveromyces did not metabolize hexadecane.
Degradation of heptachlor by C lipolytica and C. maltosa was suggested by the recovery of
1-hydroxychlordene from culture systems in concn greater than those found in controls. This
compound is produced in the environment by certain plants, by vertebrates, and by hydrolysis
(Brooks 1974). It has not previously been shown to be produced in conjunction with fungal
metabolism.
Through the use of l4C-heptachlor, preliminary findings suggest that heptachlor is not
oxidized to CO2. Of the 20% radioactivity which was not extractable, 15% was found to be
closely associated with cellular material. The exact location, in or on the cell, is not known.
However, it does not seem unlikely that it is dissolved in minute hexadecane droplets that are
attached to the cell wall or within hydrocarbon inclusions within the cell. (Munk 1970;
Finnerty et al. 1973; Hug et al. 1974).
The unusual inhibition spectra of heptachlor on hexadecane utilization by C. maltosa
contrasted to the lack of an effect on C. lipolytica. Evidence supporting the hypothesis of
nonidentical enzyme systems for utilization of alkanes by different species of Candida has
been indicated. Volfova et al. (1967) found that upon removal of the cell wall of C. lipolytica,
the protoplasts were unable to oxidize hexadecane, whereas Lebeault et al. (1969, 1970)
demonstrated that protoplasts of C. tropicalis degraded hexadecane and implicated mitochon-
drial function in hexadecane oxidation.
If it is assumed that n-alkanes enter yeast cells by passive diffusion and that active
transport is not involved (Prokop and Sobotka 1975), it may be speculated that heptachlor
inhibits hexadecane utilization by C. maltosa by interfering with mitochondrial enzyme
activity, whereas the cell membrane-centered oxidation by C. lipolytica is unaffected by
heptachlor. C. maltosa is similar to C. tropicalis in many physiological properties (Meyer et
al. 1975). The peculiar inverse relationship of heptachlor concn to inhibition was observed in
culture systems which were in an oxygen-limited state. In test-tube experiments, or in the
smaller Erlenmeyer flasks, the surface layer of hexadecane was thicker and less disturbed by
agitation than in culture flask systems. Growth and emulsification were also less under such
conditions. The overall effect of these conditions may be to increase the cellular content of
hexadecane and, hence, of heptachlor as well. In more actively metabolizing cells, intracellular
pools of hydrocarbon may decrease and accumulating heptachlor may be stored in lipids,
rendering it unavailable for immediate oxidation and thus not interfering with alkane
oxidation. Determination of the exact location of heptachlor in or on the cell and its precise
effect on the enzymatic processes will require further investigation.
The effect of heptachlor on metabolism of hexadecane by certain yeasts appears complex,
and, depending upon concn and culture conditions, is either stimulatory or inhibitory. This
emphasizes the complex nature of problems associated with determining the effects of
301
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336 N. G. SMITH ET AL.
chlorinated pesticides on the metabolism of yeasts growing in natural surface slicks.
Extrapolation of our preliminary results to the natural environment is not feasible.
ACKNOWLEDGMENT
Research performed at Georgia State University was supported in part by Grant No.
R803141-02 from the U.S. Environmental Protection Agency, Gulf Breeze Environmental
Research Laboratory.
LITERATURE CITED
Atlas, R. M,, and R. Bartha. 1972. Degradation and mineralization of petroleum by two bacteria isolated
from coastal waters. Biotechnol. Bioeng, 14:297-308.
Bourquin, A. W., S. K. Alexander, H. K. Speidel, J. E. Mann, and J. F. Fair. 1972. Microbial interactions
with cyclodiene pesticides. Dev. Ind. MicrobioL 13:264-276.
Brooks, G. T. 1974. Chlorinated Insecticides. Vol. 2. CRC Press, Inc. Cleveland, Ohio.
Crow, S. A., A. W. Bourquin, G. N. Smith, and W. L. Cook. 1975. Metabolic activity of microorganisms
from estuarine slicks. Abstr. Am. Soc. MicrobioL 191.
Finnerty, W. R., R. S. Kennedy, B. O. Spurlock, and R. A. Young. 1973. Microbes and petroleum:
Perspectives and implications. Pages 105-126 in D. G. Ahearn and S. P. Meyers, eds. The Microbial
Degradation Of Oil Pollutants. PubL No. LSU-SG-73-01 Center for Wetlands Resources, Louisiana State
Univ., Baton Rouge.
Hug, H., H. W. Blarch, and A. Fiechter. 1974. The functional role of lipids in hydrocarbon assimilation.
Biotechnol. Bioeng. 16:965-985.
Lebeault, J. M., B. Roche, Z. Duvnjak, and E. Azoulay. 1969. Protoplasts obtained from Candida tropicalis
grown on alkanes. /. Bacterial. 100:1218-1221.
1970. Isolation and study of the enzymes involved in the metabolism of hydrocarbons by
Candida tropicalis. Arch. Mikrobiol 72:140-153.
Maclntyre, F. 1974. The top millimeter of the ocean. Sci. Am. 230-62-77.
Meyer, S. A., K. Anderson, R. E. Brown, M. Th. Smith, D. Yarrow, G. Mitchell, D. G. Ahearn. 1975. The
physiological and DNA characterization of Candida maltosa, a hydrocarbon-utilizing yeast. Arch.
Microbiol. 104:225-231.
Miles, J. W. R., C. M. Tu, and C. R. Harris. 1969. Metabolism of heptachlor and its degradation products by
soil microorganisms. J. Econ. EntomoL 62:1334-1338.
Munk, V. 1970. Growth of yeasts on hydrocarbons. Pages 125-136 in D. G. Aheain, ed.Recent Trendsln
Yeast Research. Spectrum, Georgia State Univ., Atlanta.
Parker, B., and G. Barson. 1970. Biological and chemical significance on surface microlayers in aquatic
ecosystems. BioScience 20:87-93.
Prokop, A., and M. Sobotka. 1975. Insoluble substrate and oxygen transport in hydrocarbon fermentation.
Pages 127-157 in S. R. Tannenbaum and D. I. C. Wang, eds. Single Cell Protein II. The MIT Press,
Cambridge, Mass.
Seba, D. B., and E. F. Coccoran. 1969. Surface slicks as concentrators of pesticides in the marine
environment Pest. Monit. J. 3:190-193.
Shamiyeh, N. B., and L. F. Johnson. 1973. Effects of heptachlor on numbers on bacteria, actinomycetes
and fungi in soil. SoU Biol Biochem. 5:309-314.
Sieburth, J. M. 1965. Bacteriological samplers for air-water and water-sediment interfaces. Trans. Joint
Conf. Ocean Sci., Ocean Eng. Mar. Technol., Am. Soc. Limnol. Oceanogr., pp. 1064-1067
Volfova, O., V. Munk, and M. Dostalek. 1967. Loss of the ability to oxidize hydrocarbons in protoplasts of
Candida lipolytica. Experentia 23:1005-1006.
302
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Reprint from Bulletin of
Environmental Contamination
and Toxicology, Vol. 15(5):
515-521, 1976, with permis-
sion of Springer-Verlag
New York Inc.
EFFECTS OF SUSPENDED MATERIAL ON MEASUREMENT OF
DDT IN ESTUARINE WATER
Alfred 0. Wilson
Contribution No. 258
303
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Effects of Suspended Material on Measurement
of DDT in Estuarine Water
by ALFRED J. WILSON
Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Sabine Island, Gulf Breeze, Fla.
Accumulation of pesticides by suspended material in sea water
has been well documented. COX (1971) noted that 90% of DDT
residues recoverable from whole sea water was not bound to par-
ticulate material greater than one to two microns in diameter.
RICE and SIKKA (1973) showed that approximately 90% of the
maximum uptake by six species of marine algae occurred within
2 hours after exposure to C-DDT. PIERCE et al. (1974) re-
ported the humic acid fraction of sea water to have a greater
adsorption capacity than clay and sediment. OLOFFS et al.
(1973) showed that all detectable DDT residues moved into sed-
iments after 6 weeks of incubation in the laboratory. Other
studies have shown that phytoplankton and suspended particulate
material have relatively large capacities for sorption of
pesticides (COX, 1970; GREGORY et al., 1969; HILL and MCCARTY,
1967; KIEL and PRIESTER, 1969; POIRRIER et al., 1972; SODERGREN,
1968, 1973; SUFFET, 1973b; VANCE and DRUMMOND, 1969; WARE et al.,
1968; WHEELER, 1970). Recent monitoring studies show residues
of DDT in plankton from the open ocean (GIAM et al., 1973;
HARVEY et al., 1974; WILLIAMS and HOLDEN, 1973).
The present study was undertaken to investigate the effect of
suspended material on measurement of DDT in estuarine water.
Preliminary studies (WILSON et al., 1970) reported a decrease
in extractable DDT from estuarine water with time. These results
were observed in laboratory studies in which DDT was added to
natural estuarine water and incubated in sealed glass containers.
The results were interpreted by many workers as indication of DDT
breakdown. The objective of the studies described herein was
to learn if suspended material accumulated DDT under the con-
ditions of this experiment and prevented its complete recovery.
No attempt was made to evaluate the separate effects of biotic
and abiotic suspended material.
There are various methods for measuring pesticides in water. Due
to their simplicity, many analysts employ liquid-liquid extrac-
tion (LLE) methods of the batch or serial type. This method con-
sists of extracting up to about four liters of water with an or-
ganic solvent (KALLMAN et al., 1962). Other LLE methods are the
continuous multichamber systems such as that described by KAHN
and WAYMAN (1964). Carbon has been used to adsorb, pesticides
(MIDDLETON and LICHTENBERG 1960). Recently, other adsorptive
techniques have been described. HARVEY (1972) described a
515
Bulletin of Environmental Contamination & Toxicology
Vol. 15, No. 5 © 1976 by Springer-Verlag New York
-------
synthetic resin for the analyses of pesticides in sea water.
AHLING and JENSEN (1970) used reversed liquid-liquid partition
methods for extraction of chlorinated pesticides from water.
Once a method of extraction has been selected, it is usually
evaluated by fortification of a water sample with the compound
and determining if the method produces acceptable accuracy and
precision (MCFAKREN et al., 1970). Isotope tracers are also used
in determining acceptability of a method.
The following experiments were designed to determine the effi-
ciency of serial LLE of DDT fortified estuarine water and evalu-
ate the recovery rate from fortified samples by extracting sus-
pended material and water separately.
RECOVERY STUDIES OF LLE OF DDT FORTIFIED ESTUARINE WATER
Methods. Duplicate 4 liter clear glass bottles, containing 3.5 1
of estuarine water or distilled water, were fortified with 10.5
pg of p,p' -DDT in 350 pi of acetone to yield a concentration of
3.0 ppb (parts per billion). Five hundred ml samples were taken
from each bottle and extracted by shaking vigorously for one
minute in a separatory funnel as follows: three times with 50 ml
of petroleum ether, two times with 50 ml of 15% ethyl ether in
hexane followed by 50 ml hexane, or three times with 50 ml of
methlyene chloride. All solvents were dried with sodium sulfate,
concentrated to an appropriate volume and analyzed by electron
capture gas chromatographs. equipped with 2.0% OV-101 and 0.75%
OV-17: 0.97% OV-210 in glass columns. Methylene chloride ex-
tracts were transferred to petroleum ether by concentrating the
methylene chloride to about 10 ml, adding 50 ml of petroleum
ether and reconcentrating to about 10 ml. This removed the
methylene chloride which cannot be used in electron capture gas
chromatography. Just prior to extraction, all samples were
fortified with o,p' -DDE as an internal standard. The recovery
rate of o,p' -DDE in all tests was greater than 89%, indicating
no significant loss during analyses. The estuarine water was
collected in Santa Rosa Sound, Florida. The salinity ranged from
16 to 24 ppt.
After initial sampling, the bottles were sealed and incubated at
20 C under controlled light conditions (5000 lux; 12 hours light,
12 hours dark). Duplicate samples of 500 ml were extracted at
various time intervals.
Results. Table 1 and 2 show the average percentage recovery of
P»P? -DDT extracted from duplicate estuarine water or distilled
water samples up to 14 days after initiation of the experiment.
p,p' -DDE was the only metabolite measured, but since it never
exceeded 2% of the parent compound it is not included in the
percent recoveries.
Table 1 shows that immediately after the estuarine water (21 ppt
salinity) was fortified with 3.0 ppb of DDT, all solvent systems
removed 93% of the DDT. After six days of incubation, less was
516
306
-------
recovered with all solvents.
most efficient.
However, methylene chloride was the
TABLE 1
PERCENTAGE RECOVERY OF P,P' -DDT FROM ESTUARINE WATER
BY DIFFERENT EXTRACTION SOLVENTS
Day
0
6
Extraction Solvent
Petroleum
Ether
93
67
15% Ethyl Ether
In Hexane
93
66
Methylene
Chloride
93
76
An experiment was performed with estuarine water (21 ppt salin-
ity) and distilled water using petroleum ether and methylene chlo-
ride. Table 2 shows that immediately after fortification, recov-
eries were 90% or greater. After 14 days, similar recoveries were
obtained only from distilled water. In estuarine water however,
there were 49% and 28% reduction in recovery from zero day with
petroleum ether and methylene chloride respectively. Since dis-
tilled water is devoid of particulate matter, this suggests that
DDT may be absorbed or adsorbed to particulate matter found in
estuarine water and the DDT sorbed onto this matter was not re-
moved, resulting in low recoveries. This explains the initially
high extraction efficiency of DDT followed by decline in recovery
as DDT became associated with the particulate phase. Since meth-
ylene chloride was the most polar solvent used, it had the great-
est affinity for removing sorbed DDT.
TABLE 2
PERCENTAGE RECOVERY OF P,P' -DDT FROM ESTUARINE WATER AND
DISTILLED WATER BY PETROLEUM ETHER AND METHYLENE CHLORIDE
Estuarine Water
Day
0
7
14
Petroleum
Ether
90
58
46
Methylene
Chloride
94
78
68
Distilled Water
Petroleum
Ether
90
90
94
Methylene
Chloride
91
91
92
The above data are typical of results obtained from several tests
517
3,07
-------
with estuarine water of varying salinity (16 - 24 ppt). These
experiments suggest that LLE methods are not efficient for the
extraction of DDT from suspended material. To test this hypothe-
sis, experiments were done in which the suspended material was
separated from the water and both constituents analysed separa-
tely.
RECOVERY OF DDT FROM WHOLE SEA WATER AND SUSPENDED MATERIAL
Methods. Duplicate 4 liter bottles of estuarine water was forti-
fied and incubated under controlled temperature and lighting condi-
tions as described above. Samples were removed at various time
intervals and analysed as follows: 500 ml samples were taken from
each bottle and extracted in a separatory funnel three times with
50 ml of methylene chloride. In addition, 500 ml samples from
each bottle were filtered through 47 mm, 0.4 micron porosity
Nucleopore filters. The filter was placed in a 200 mm x 25 mm
(O.D.) screw top test tube containing 10 ml acetonitrile. The
test tube was then placed in a Varian Ultrasonic Cleaner and
sonicated for 30 minutes at 45 C. Twenty five ml of 2.0% aqueous
sodium sulfate and 5 ml of hexane were added. The test tube was
sealed with a teflon-lined cap and shaken for one minute. After
the solvent phases separated, the upper hexane layer was trans-
ferred with a dropping pipet to a clean 25 ml graduated cylinder.
This was repeated three times with 5 ml of hexane. The combined
extracts were analysed by electron capture gas chromatography.
The filtrate was extracted as follows: the filtering apparatus
and the graduated cylinder used to measure the sample were rinsed
with acetonitrile and the rinses added to the 500 ml filtrate.
The combined filtrate and acetonitrile rinses were extracted in a
separatory funnel three times with 50 ml of petroleum ether. The
extracts were dried with sodium sulfate and concentrated to an
appropriate volume for analyses by electron capture gas chroma-
tography. Before extraction, the filter and filtrate were forti-
fied with o,p'-DDE as an internal standard.
Estuarine water for these tests was collected from Santa Rosa
Sound, Florida (salinity range 22 - 28 ppt). However, in one
test, artificial estuarine water was used. This was prepared by
dissolving 210 grams of Rila salts in 7 liters of distilled water.
The resulting solution (23 ppt salinity) was filtered through a
glass filter. A mixed algal culture consisting of Chlorella sp.,
Dunaliella tertiolecta and Chlamydomonas sp. was added to the
artificial estuarine water.
Results. Table 3 compares the percentage recovery of DDT from
methylene chloride extractions of the entire water sample with
percentage recovery after extraction of the suspended material and
water separately (filter + filtrate). In all tests there was a
significant increase in the recovery of DDT when the suspended
material was analysed separately. The greatest increase was in
fortified artificial estuarine water containing the algal culture.
518
308
-------
TABLE 3
PERCENTAGE RECOVERY OF DDT FROM EXTRACTION OF WHOLE WATER
AND SEPARATE EXTRACTION OF SUSPENDED MATERIAL AND FILTRATE
Day
Exp. 1
(Estuarine water)
0
4
9
Exp. 2
(Estuarine water)
0
4
Exp. 3
(Estuarine water)
0
6
Exp. 4
(Estuarine water)
0
7
8
Exp. 5
(Artificial
estuarine water
+ algae)
0
5
Percentage Recovery (Std dev)
Whole water
85 (1.8)
75 (1.8)
77 (2.6)
84 (4.4)
84 (0)
72 (8.6)
76 (3.6)
91 (0)
56 (.89)
85 (0)
54 (5.9)
suspended Material
+ Filtrate
89 (4.4)
101 (6.2)
90 (12)
91 (3.5)
100 (.89)
89 (1.5)
99 (1.8)
83 (5.6)
88 (8.2)
84 (5.4)
85 (2.7)
Percentage
Increase
4.7
35
17
8.3
19
25
30
48
0
57
DISCUSSION
These experiments show the pitfalls of sample fortification.
Liquid-liquid extraction of estuarine water immediately after
fortification yielded acceptable recovery levels with all solvent
systems tested. However, analyses several days later gave only
partial recovery. Field residues may be subject to similar phys-
ical and chemical transformations and therefore complete recovery
of DDT may not be possible by these methods. Similar studies
519
309
-------
reported by OLOFFS et al. (1972) showed up to 60% decrease in re-
covery of DDT from estuarine water fortified with 25 ppb DDT and
incubated for 12 weeks in flasks stoppered with glass wool. The
mechanism suggested for this loss was evaporation and co-distill-
ation. SUFFET (1973a) suggested that the recovery from fortified
laboratory water samples approach actual recovery from field sam-
ples if a pesticide is completely dissolved and not associated
with suspended matter and the water properties are similar to
natural water. However, estuarine water has a pH above 7.0, is of
high ionic strength, and contains suspended material.
A more exhaustive extraction was required to remove the sorbed
DDT from the suspended material (Table 3). Experiment 5 shows
that phytoplankton will accumulate DDT and LLE methods will not
remove all the sorbed DDT.
Different results may be observed if experiments are conducted
below the solubility of DDT (1.2 ppb in distilled water; reported
by BOWMAN et al. , 1960). EICHELBERGER and LICHTENBERG (1971)
fortified natural river water with 1.0 ppb DDT and incubated rep-
licate samples in sealed glass containers for eight weeks. No
loss of DDT was observed by LLE methods. However, COX (1970)
studied C-DDT uptake in three species of marine phytoplankton
and noted concentration factors ranging from 2.5 to 8.0 x 10 at
0.8 to 3.0 parts per trillion. RICE and SIKKA (1973) also ob-
served rapid uptake of DDT by marine algae at 1.0 ppb.
Interaction of pesticides between water and particulate matter is
complex. Evaluating LLE methods of herbicides from river water,
SUFFET (1973b) observed that the isopropyl ester of 2,4-D was
adsorbed to particulate material in river water and that the
amount changed by alteration of the pH of the water. HUANG and
LIAO (1970) found that adsorption of DDT to clays was rapid but
the amount differed with the type of clay. A mixed culture of
algae consisting mainly of Vauchenia had a greater adsorption for
DDT than bentonite according to HILL and MCCARTY (1967). The
experiments reported in this paper support the work of other
investigators in that DDT is extremely hydrophobic and can easily
be adsorbed or abosrbed by suspended matter in liquid solutions.
These studies indicate the observed loss of DDT by this author's
earlier studies (WILSON et al., 1970) was due to sorption of DDT
to suspended material which prevented complete recovery of DDT by
the methods used in the experiment.
It is difficult to relate laboratory findings directly to that of
the estuary or open ocean. However, the laboratory data described
here illustrate clearly some problems that could be encountered
in monitoring estuarine water for pesticide pollution. Data
obtained by analysis of water by liquid-liquid extraction methods
or other methods which do not efficiently extract sorbed pollu-
tants from suspended material may be misleading.
520
310
-------
REFERENCES
AHLING, P., and JENSEN, S.: Anal. Chem. 42, 1483 (1970).
BOWMAN, M.C., AGREE, F. JR., and CORBETT, M.K.: J. Agric. Food
Chem. J3, 406 (1960).
COX, J.L.: Bull. Environ. Contain. Toxlcol. .5, 218 (1970).
COX, J.L.: U. S. Fish. Wildl. Serv. Fish. Bull. j>9, 443 (1971).
EICHELBERGER, J.W. and LICHTENBERG, J.J.: Environ. Sci. Technol.
5., 541 (1971).
GIAM, C.S., WONG, M.K., HANKS, A.R., SACKETT, W.M. and RICHARDSON,
R.L.: Bull. Environ. Contain. Toxicol. 9_, 376 (1973)
GREGORY, W.W. JR., REED, J.K. and PRIESTER, L.E. JR.: J.
Protozool. 16, 69 (1969).
HARVEY, G.R.: Woods Hole Oceanogr, Inst. Tech. Rep. 72-87 (1972).
HARVEY, G.R., MIKLAS, H.P., BOWEN, V.T. and STEINHAUER, W.G.: J.
Mar. Res. 32_, 103 (1974).
HILL, D.W. and MCCARTY, P.L.: J. Water Pollut. Control Fed. J39,
1259 (1967).
HUANG, J.C. and LIAO, C.S.: J. Sanit. Eng. Div., Proc. Am. Soc.
Civ. Eng. 96_, 1057 (1970).
KAHN, L. and WAYMAN, C.H.: Anal. Chem. 36, 1340 (1964).
KALLMAN, B.J., COPE, O.B. and NAVARRE, R.J.: Trans. Am. Fish.
Soc. 91, 14 (1962).
KEIL, J.E. and PRIESTER, L.E.: Bull. Environ. Contam. Toxicol.
4., 169 (1969).
MCFARREN, E.F., LISHKA, R.J. and PARKER, J.H.: Anal. Chem. 42,
358 (1970).
MIDDLETON, F.M. and LICHTENBERG, J.: Ind. Eng. Chem. .52, 99A
(1960).
OLOFFS, P.C., ALBRIGHT, L.J. and SZETO, S.Y.: Can. J. Microbiol.
18, 1393 (1972).
OLOFFS, P.C., ALBRIGHT, L.J., SZETO, S.Y. and LAU, J.: J. Fish.
Res. Board Can. 30, 1619 (1973)
PIERCE, R.H. JR., OLNEY, C.E. and FELBECK, G.T. JR.: Geochim.
Cosmochim. Acta 38, 1061 (1974).
POIRRIER, M.A., BORDELON, B.R. and LASETER, J.L.: Environ. Sci.
Technol. 6^, 1033 (1972).
RICE, C.P. and SIKKA, H.C.: J. Agric. Food Chem. 21, 148 (1973).
SODERGREN, A.: Oikos 19, 126 (1968).
SODERGREN, A.: Oikos ^4, 30 (1973).
SUFFET, I.H.: J. Agric. Food Chem. 21, 288 (1973a).
SUFFET, I.H.: J. Agric. Food Chem. 21, 591 (1973b).
VANCE, B.D. and DRUMMOND, W.: J. Amer. Water Work Assoc. 61,
361 (1969).
WARE, G.W., DEE, M.K. and Cahill, W.P.: Bull. Environ. Contain.
Toxicol. 1, 333 (1968)
WHEELER, W.B.: J. Agric. Food Chem. 18, 416 (1970).
WILLIAMS, R. and HOLDEN, A.V.: Mar. Pollut. Bull. $., 109 (1973).
WILSON, A.J., FORESTER, J. and KNIGHT, J.: U.S. Fish Wildl. Serv.
Circ. 335, 18 (1970).
521
311
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Reprinted from Trace Sub-
stances in Environmental
Health-IX. A Symposium, D.D.
Hemphill, editor, 1975, pp. 169-
177, with permission by the
Curators of the University of
Missouri
METHODS TO ASSESS EFFECTS OF COMBINATIONS OF TOXICANTS,
SALINITY AND TEMPERATURE OF ESTUARINE ANIMALS
Lowell H. Bahner and Del Wayne R. Nimmo
Contribution No. 259
313
-------
Reprinted from Trace Substances In Environmental Health-IX. 1975 A symposium
D. D. Hemphill. Ed., © University of Missouri, Columbia.
Methods to Assess Effects of
Combinations of Toxicants,
Salinity and Temperature on
Estuarine Animals
Lowell H. Bahner and DelWayne R. Nimmo
U.S. Environmental Protection Agency
Gulf Breeze Environmental Research Laboratory
Sabine Island, Gulf Breeze, Florida
(Associate Laboratory of the National Environment Research Center,
Corvallis, Oregon)
ABSTRACT
Aquatic species are exposed to toxicants singly, but more
often in combinations, under varying environmental regimes.
Consequently, an experimental flowing-water bioassay system was
developed that controls salinity and temperature while testing
toxicants either singly or in combination. Obvious advantages
of this control were that rates of toxicant accumulation,
translocation, loss or acute and chronic toxicity to animals
could be better assessed and repeated. Our bioassays were
conducted with pink shrimp (Penasus duoraYwn) exposed to the
following toxicant combinations: cadmium-malathion, cadmium-
methoxychlor, cadmium-methoxychlor-Aroclor® 1254 and a complex
industrial waste which contained both inorganic and organic
constituents. The toxicities of the pesticide-metal combi-
nations, when compared to that of each constituent singly,
appeared to be independent of each other.
INTRODUCTION
Components of an outfall may interact to exert toxic effects that
differ from those of single toxic components. Some examples in which
combined effects were greater (synergistic) than those of effects of
single components added together have been reported (2, 12, 13). Also,
it is known that environmental factors affect the toxicity of some
chemicals (5-8, 15). It is difficult to determine if toxicity is from a
waste as a whole, a single component in the waste or from combinations of
the waste interacting with environmental factors.
One of the major requirements in estuarine bioassays is that of
maintaining constant bioassay conditions, e.g. salinity, temperature,
adequate oxygen levels, etc. Of course, some of these factors can be
controlled with static tests, but disadvantages encountered with static
systems include: metabolism of toxicant, loss of toxicant to absorptive
surfaces and buildup of metabolic products from test organisms.
Therefore an investigation was conducted in which we developed a flow-
through system that controlled salinity and temperature and incorporated
a method to 1) deliver toxicants, singly or in combination and 2) to
alter the salinity after the manner of a partial tide cycle. To our
knowledge, this is the first report of a capability to control salinity
fluctuations during a flow-through bioassay.
We report the results of the short-term bioassays (48-hr, 96-hr,
10-day) utilizing the following combinations: cadmium-malathion,
169
31S
-------
cadmium-methoxychlor, cadmium-methoxychlor-Aroclor1 1254, as well as
preliminary results of a 13-day bioassay of a complex industrial waste.
MATERIALS AND METHODS
All flow-through bioassays employed our constant-temperature constant-
salinity seawater system and the acclimation procedures of Bahner et al.
(1). Sixty-five liters of filtered seawater per hr were delivered to
each 30-1 glass aquarium to provide sufficient flow for minimal loading;
12 1 per hr per aquarium were supplied during the industrial waste
bioassay. The temperature was 25 ± 2°C throughout each test; salinity
was maintained at 20 ± 2 %o during each exposure to toxicants and oxygen
concentrations remained at or near saturation in all test aquaria.
Test procedures were those described by Bahner et al. (1); pink
shrimp (Penaeus duorarum) were collected and handled in the manner given
by Nimmo et al. (9). Volumes of 2 ml to 4 1 of toxicant stock solutions
were delivered daily to achieve desired toxicant concentrations in the
bioassay aquaria. Animals were tested in a range of measured toxicant
concentrations to determine the LC50; thereafter, all combinations of
toxicants were tested at the respective LC50 concentration. LCSO's were
calculated singly on the range of toxicants using probit analysis (4).
Combinations of toxicants were obtained by simultaneously metering indi-
vidual toxicants from separate syringes or flasks into the aquaria,
except during the malathion-cadmium combination. Shrimp were exposed to
Cd singly for 96 hr and then to malathion singly for 48 hr. Bioassays of
toxicants singly were performed concurrently with each combinatorial test.
In the 10-day cadmium-methoxychlor-Aroclor 1254 combination (Figure 6),
test concentrations were the approximate LCSO's calculated from separate
30-day tests conducted with Cd and methoxychlor, and the LC50 from a 15-
day bioassay for Aroclor 1254. At the conclusion of each exposure time,
the salinity was gradually lowered to 2 °foo within 4-8 hr to determine
the effect of salinity reduction on survival.
Malathion in water was analyzed by gas chromatography with a flame
photometric detector in the phosphorus mode (14) and methoxychlor and
Aroclor 1254 were analyzed by electron-capture gas chromatography (9).
Cadmium was analyzed in water by flameless atomic absorption spectroscopy,
using methods of Segar (10).
RESULTS
Acute 48- or 96-hr LCSO's were: malathion, 12.5 ug/1; Cd, 4.6 mg/1;
methoxychlor, 3.5 yg/1 (Figures 1-3 and Table I). Combinations of these
toxicants caused mortalities equal those of the arithmetical sums of the
toxicants (Figures 4-6, Table I).
There was no difference in survival as a result of lowered salinity
between shrimp that had been exposed to toxicants and controls.
The 13-day bioassay of the industrial waste showed it to be rela-
tively innocuous to the shrimp compared to some of the toxicants (Table I).
There is no doubt that this waste should be considered "complex" because
11 organics and 14 metals were detected. Three glycols and an alcohol
were present at g/1 concentrations; 9 metals were present at mg/1
concentrations.'
DISCUSSION
The problem of assessing the toxicity of complex wastes or predicting
interactions of one toxicant with those present in the receiving water has
Registered Trademark, Monsanto Co., St. Louis, MO. Mention of
commercial products does not constitute endorsement by the U.S. EPA.
170
316
-------
TABLE I. SUMMARY OF BIOASSAYS WITH PINK SHRIMP, Penaeus duoi-arum
(25 ± 2 C, 20 ± 2 °/oo)
Toxicant (s)
Malathion
Cd
Methoxychlor
PCB*
Cd + Malathion
Cd + Methoxychlor
Cd (singly)
Methoxychlor (singly)
Cd + Methoxychlor + PCB
Cd (singly)
Methoxychlor (singly)
PCB (singly)
Industrial Waste
LC-50 for
Toxicant
12.5 yg/1
4.6 mg/1
3.5 yg/1
1.0 yg/1
**
**
**
0.03%
Length of
Exposure
48 hr
96 hr
96 hr
15 da
Cd 96 hr, then
Malathion 48 hr
120 hr
M
n
10 da
11
ir
ti
13 da
Toxicant
Concentrations
Texted
7.3-50 yg/1
1.0 - 10.3 mg/1
1.0 - 7.1 yg/1
0.57 - 19.0 yg/1
Cd: 3.5 - 10.3 mg/1
Malathion: 5.7 yg/1
5.4 mg Cd + 3.5 yg Methox. /I
5.6 mg Cd/1
3.3 yg Methox. /I
0.83 mg Cd + 0.9 yg Methox.
+ 0.82 yg PCB/1
0.64 mg Cd/1
1.0 yg Methox. /I
0.73 yg PCB/1
0.015% - 1.0%
*Aroclor 1254 data from Nimmo et at. (9)
**Not Determined
-------
CONTROL
MALATHION-ACUTE
1C 50 12.5 MQ/I
24 36
TIME (hours)
FIGURE 1—ACUTE TOXICITY (LC50) OF THE INSECTICIDE
MALATHION TO SHRIMP, Penaeua duororum, IN 48 HOURS.
•^
<
3
Z
CADMIUM-ACUTE
LC50 4.6mg/l
CONTROL
10.3 mo/I
24
48
TIME (hours)
72
96
FIGURE 2—ACUTE TOXICITY (LC50) OF CADMIUM TO SHRIMP
Penaeua duororum, IN 96 HOURS.
-------
been discussed by Cairns and Scheier (2). Although their discussion dealt
with fresh water, the same concepts apply equally in estuarine or marine
waters. They note, "Materials from several such industries entering a
stream within a short distance often create regulatory problems of incred-
ible complexity".
In the first combination test, we attempted to determine whether the
toxicity threshold to malathion was increased by previous exposure to Cd.
Also, we believed it unrealistic to expose shrimp to malathion for longer
than 48 hr, since Tagatz et al. (14) had not detected malathion in estu-
arine water 24 hr after treatment of an estuarine marsh. There was
evidence that Cd accumulated and remained in tissues of marine Crustacea
for some time (3). Our results showed that there was no substantial
increase in toxicity of malathion after previous exposure of shrimp to Cd.
8
at
3
METHOXYCHLOR - ACUTE
1C 50 3.5«g/<
48
TIME (hour*)
FIGURE 3—ACUTE TOXICJTY (LC50) OF THE INSECTICIDE METHOXYCHLOR
TO SHRIMP, Penaeus duorarum, IN 96 HOURS.
173
319
-------
301
234
TIME (days)
FIGURE *»—TOXICITY OF A 4-DAY EXPOSURE TO SEVERAL CONCENTRATIONS OF
CADMIUM FOLLOWED BY A SINGLE CONCENTRATION OF MALATHION FOR
2 DAYS TO Penaeus duorarwn.
In the second combination test, our results indicate that toxicity of
the methoxychlor-cadmium combination was almost identical to that of the
sum of each toxicant tested singly (predicted additive effect).
In the third series of tests, we again observed only additive toxi-
city of any 2-way or 3-way combination. For example, when Cd, methoxy-
chlor and Aroclor 1254 were administered In combination, the total toxi-
city was equal to that of the sum of each- tested singly. This rela-
tionship can be seen by equating the observed percentage killed (ordinate)
to that expected (abscissa) for each toxicant tested singly. Unity would
be denoted by the litie constructed through these points and the origin.
If any 2- or 3-way combination were exactly additive, the result would
fall on unity. If synergistic effects were observed with any combi-
nation, the percentage would lie to the left of unity; if antagonistic,
to the right of unity. In view of our previous work with Aroclor 1254,
we were surprised to find no effects of lowered salinity.
174
320
-------
20
IS
Ul
z
< 10
u.
o
CK
IU
CO
CONTROL
METHOXYCHLOR
•3.3ju9/l-
CADMIUM
PREDICTED
ADDITIVE s
EFFECT V
CD, 5.4 m0/1
METHOX., 3.5>i9/l
40 80
TIME (hours)
120
.FIGURE 5—TOXICITY OF CADMIUM AND METHOXYCHLOR ADMINISTERED SINGLY
AND IN COMBINATION TO Penaeus duorarum.
Industrial waste at a 3300-fold dilution (0.03%) was toxic to 50%
of the shrimp (LC50) in 13 days. A 6600-fold dilution of the industrial
waste was also lethal to 10% of the shrimp. The only component of the
waste for which we have toxicological data is Cd (Table I). We may not
have observed acute toxic effects because only minute quantities of Cd
were present in the waste. By our calculations, concentrations of Cd in
the test aquaria were lower than that occurring naturally (background)
in seawater (16).
175
32J
-------
• PCB (AROCLOR 1254)
O CADMIUM
-+- METHOXYCHLOR
20 30 40 50
PERCENTAGE KILLED IN 10 DAYS
{EXPECTED)
60
FIGURE 6—COMPARISON OF TOXICITIES TO Penaeus duora&m OF SINGLE VS.
COMBINED CONSTITUENTS. Measured concentrations in micrograms
per liter ranged from 6^0-829 cadmium; 0.9 .- 1.0 methoxychlor;
0.7 - 1.1 PCB (Aroclor 125*0. Each symbol represents la single
toxicant per'aquarium; superimposed symbols represent the
toxicants combined.
CONCLUSIONS
No dramatic Interactions among various components of mixtures were
evident. The toxicity of each component was independent and additive.
There was no apparent toxic effect of lowered salinity after exposures to
the various components or mixtures. We are not implying that synergistic
effects of toxicant combinations or the deleterious influence of envi-
ronmental factors do not exist. Rather, the results of the combinations
indicate additivity, and for this reason background concentrations of
multiple pollutants in receiving waters, or inherent in the aquatic
ecosystems, should also be considered before discharging, dredging or
compositing a specific waste.
ACKNOWLEDGMENTS
We thank Steven Foss for preparing and photographing the figures
included in the manuscript.
176
322
-------
LITERATURE CITED
1. Bahner, L. H., C. D. Craft and D. R. Nimmo. 1975. A saltwater flow-
through bioassay method with controlled temperature and salinity.
Prog. Fish-Cult. (In press).
2. Cairns, J., Jr. and A. Scheier. 1968. A comparison of the toxicity
of some common industrial waste components tested individually and
combined. Prog. Fish-Cult. 30:3-8.
3. Eisler, R., G. E. Zaroogian and R. J. Hennekey. 1972. Cadmium
uptake by marine organisms. J. Fish Res. Bd. Can. 29:1367-1369.
4. Finney, D. J. 1971. Probit Analysis, Cambridge University Press,
Cambridge, pp. 1-333.
5. Liang, T. T. and E. P. Lichtenstein. 1974. Synergism of insec-
ticides by herbicides: Effect of environmental factors. Science
186:1128-1130.
6. MacLeod, J. C. and E. Pessah. 1973. Temperature effects on mercury
accumulation, toxicity, and metabolic rate in rainbow trout (Salmo
gairdneri). J. Fish. Res. Bd. Can. 30:485-492.
7. Mosser, J. L., Tzu-Chiu Teng, W. G. Walther and C. F. Wurster. 1974.
Interactions of PCBs, DDT and DDE in a marine diatom. 1974. Bull.
Env. Contam. Toxicol. 12:665-668.
8. Nimmo, D. R. and L. H. Bahner. 1974. Some physiological conse-
quences of polychlorinated biphenyl- and salinity-stress in penaeid
shrimp. In: Pollution and Physiology of Marine Organisms, F. J. and
W. B. Vernberg, Eds., Academic Press, New York. pp. 427-443.
9. Nimmo, D. R., R. R. Blackman, A. J. Wilson, Jr. and J. Forester.
1974. Toxicity and distribution of Aroclor® 1254 in the pink shrimp
Penaeus duorarum. Mar. Biol. 11:191-197.
10. Segar, D. A. 1971. The use of the heated graphite atomizer in
marine sciences. Proa. 3rd Int. Cong. Atomic Abs. Atomic Fluores.
Spect., Adam Hilger, London, pp. 523-531.
11. Silbergeld, E. K. 1973. Dieldrin: effects of chronic sublethal
exposure on adaption to thermal stress in freshwater fish. Env. Sai.
Teahnol. 7:846-849.
12. Sprague, J. B. 1964. Lethal concentrations of copper and zinc for
young Atlantic salmon. J. Fish. Res. Bd. Can. 21:17-26.
13. Sprague, J. B. and B. A. Ramsay. 1965. Lethal levels of mixed
copper-zinc solutions for juvenile salmon. J. Fish. Res. Bd. Can.
22:425-432.
14. Tagatz, M. E., P. W. Borthwick, G. H. Cook and D. L. Coppage. 1974.
Effects of ground applications of malathion on salt-marsh envi-1
ronments in Northwestern Florida. Mosq. News 34:309-315.
15. Vernberg, W. B. and J. Vernberg. 1972. The synergistic effects of
temperature, salinity, and mercury on survival and metabolism of the
adult fiddler crabs, Uca pugilator. Fish. Bull. 70:415-420.
16. Vernberg, W. B. and J. Vernberg. 1972. Environmental Physiology of
Marine Animals, Springer-Verlag, Berlin, pp. 1-346.
DISCUSSION
Inquirer: Edward Groth III, National Research Council, Washington,
D.C.
Q. Have you looked at sub-lethal toxic effects, e.g., effects on repro-
duction?
A. The effects of Cd on the reproductive stages of Penaeus duorarum
were not studied since this species of shrimp spawns in the open
Gulf of Mexico. However, sub-lethal effects have been observed in
gill tissue as necrosis of the gill epithelium, and imbalances in
serum amino acids have been observed in Cd-poisoned shrimp.
177
323
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Reprinted from Baculoviruses
for Insect Pest Control:
Safety Considerations, pp.
62 and 111-114, 1975, with
permission of the American
Society for Microbiology
DISCUSSIONS FROM SELECTED PAPERS FROM EPA-USDA
WORKING SYMPOSIUM, BETHESDA, MARYLAND
John A. Couch
Contribution No. 262
325
-------
Discussant: JOHN COUCH, Environmental Protection Agency, Gulf Breeze Environmental Research
Laboratory, Sabine Island, Gulf Breeze, Florida 32651
In a positive sense, I would like to emphasize
several concepts that Dr. Ignoffo raised. In
respect to specificity, he made an important
distinction between specificity in nature versus
specificity in the laboratory. We all may hatch
theoretical schemes wherein parasites of various
kinds under certain conditions could become
infectious to new hosts. But I think that we
have to make the distinction between nature
and the laboratory in regard to the amount
of work that has been done with the insect
viruses.
I would like to acknowledge Dr. Ignoffo's
testing and the reporting of testing of insect
viruses in non-insect invertebrates because I
am particularly interested in aquatic and marine
invertebrates, especially Crustacea. However, I
think there is a great need for publication of the
methods and results that were used in these
tests so that those in EPA and other people
working in this area can have the benefit of
these data and make decisions and plan our
own tests based on them. I would call for quick
publication of those tests that have been com-
pleted where data are available presently. In my
search for effects on aquatic invertebrates I
have not found very many related publications,
and I would like to obtain any available. I am
sure other people would.
I think Dr. Ignoffo also pointed out the need
to define the level of specificity. Now as far as
EPA is concerned I can see that this would
pose a problem in regard to regulatory meas-
ures. The levels of specificity are going to have
to be examined and certain levels are going to
have to be determined to be acceptable.
I would like to turn my attention briefly to
what may perhaps be considered a negative
point of view in regard to insect virus pes-
ticides, and this concerns aquatic invertebrates.
The range of possible organisms for specificity
testing includes practically all animals here,
protozoa through man. Human and mammalian
safety is of paramount importance, of course.
But I would suggest that, because of closer
phylogenetic proximities of Crustacea to in-
sects, perhaps we should emphasize continu-
ally, as new agents are developed, the examina-
tion of specificity of NPVs and GVs within the
phylum Arthropoda.
It is staggering to consider the numbers of
Crustacea, aquatic and marine, that may come
in contact with the formulation of these viruses.
I would point out here there is a great economic
investment in Crustacea. The world shrimp
fishery is worth over a billion dollars a year,
so it is no small matter even in terms of
agricultural values.
Our recent finding of an apparent new NPV
in shrimp may suggest the potential extension
of the host range into non-insect species, and
the presence of this virus-like entity in the pink
shrimp is not a theoretical matter any longer, as
far as I am concerned; it is a fact.
Now, one last word on kinds of tests to
determine specificity. Most tests in the past
have employed, as far as I can determine, a
bioassay method using the mean lethal dose as
the chief criterion of the effect. Is this sensitive
enough? Should we look for latent infections
and sublethal effects in greater detail? As new
agents are developed I think this will become a
continuing effort, and I do not mean this to be
in any way negative toward biological control
agents and their development. I think it is part
of the scheme, the over-all scheme of continued
safety testing.
In conclusion, I would like to point out the
difficulty of working with some of these non-
insect invertebrates. It is very hard to work
with some of the marine invertebrates, for
which there are no cell lines; they are not even
amenable to culturing of the whole organism.
One therefore can anticipate running into prob-
lems, extreme problems, in testing or applying
tests of the NPVs and GVs to these organisms.
This is a pioneering field with regard to aquatic
organisms, but I think it has great promise.
I would emphasize my original point that we
should start close to the source from a concep-
tual point of view in testing some of the NPVs
and GVs and look critically at the effect on
other arthropods, particularly Crustacea.
327
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Reprinted from BACULOVIRUSES FOR INSECT PEST CONTROL: SAFETY CONSIDERATIONS
Copyright © 1975
American Society for Microbiology
Discussion
Discussant: JOHN COUCH, Environmental Protection Agency, Gulf Breeze Environmental Research
Laboratory, Sabin Island, Gulf Breeze, Florida 32651
Two points that come out of Dr. Wolf's
discussion interest me. I think he was speaking
largely of vertebrates as far as fish and wildlife
are concerned. If we go to the invertebrates
and include them in his group, then we
have, as far as percentages go, a much
larger group of nontarget species. When you
consider the relative values of fishery products,
commercial fish do not compare to the marine
arthropods—the shrimp, lobsters, and crabs—
as far as money value goes, that is, value
per pound. I think we should consider the
many hundreds of nontarget invertebrate species
in that group.
Also, it is quite evident from this meeting
that we need attempts by qualified people to
establish a non-insect invertebrate cell line in
culture for extending our capabilities of testing.
There has been some work in New York (Dr.
Nigrelli's laboratory) on culturing echinoderm
tissues. I think the claim has been made there
that there is a cell line for echinoderms. For
Crustacea there are no cell lines established, but
it is probably a matter of doing the work.
I had the privilege of looking at some of Dr.
WestaJl's reports in which oysters, shrimp, and
fish had been exposed in test situations to
certain of the insect viruses that were candi-
date viruses for control. The results of these
experiments were negative. Oysters, shrimp,
and fish showed no adverse effects under the
test conditions used.
Some of these tests were 96-h acute ex-
posures, 10-day maximum exposures, and 7-day
exposures to various dosages. A single criterion
for effect, for example, was the oyster's shell
growth over a 96-h period after exposure—but
one thing about oysters is that shell growth may
not reflect viral infection at all. I think the time
element and criteria of tests here are important.
Sample sizes are also important in these tests:
if you use 5 or 10 animals, is this statistically
large enough a test to be valid. I think we have
to consider these minimal numbers of test
animals in setting up safety tests.
Dr. Heimpel pointed out that he, Dr.
Sparks, and Dr. Lightner, through the Gal-
veston Laboratory, have tested brown shrimp
against NPV with no adverse effect, in both
feeding and inoculation studies.
I would like to give a brief review of the
interesting new developments in virus research
in aquatic invertebrates. Some of this work
probably has no direct bearing on NPV orGV,
but it is presented to illustrate that new finds
are broadening the frontiers of exploratory
virology, particularly in regard to lower or-
ganisms as hosts. The viruses I am going to talk
about have been reported as parasites or patho-
gens in their aquatic hosts and not as con-
taminants.
The first rod-shaped virus particle reported
from non-insect invertebrates was that of
Dougherty et al. in 1963 (5). It was found in
111
328
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112
BACULOVIRUS INSECT PEST CONTROL
microannelids that were being maintained in cul-
ture for various experimental purposes. These
microannelids were reported to contain various
rodlike particles (no inclusion bodies), and these
particles were associated with a lytic disease of
the worms that resulted in mortality of the
worms.
Vago (8) and Bonami and Vago (2) reported
icosahedral viruses from the crab Macropipus
depurator. These viruses caused disease in the
crabs and were transmitted in hemolymph from
crab to crab.
Bang (1) reported an icosahedral virus in the
shore crab, Carcinus, near Roscoff, France.
This virus was transmitted serially from crab to
crab and was associated with loss of cellular
clumping in crab hemolymph.
More recently, Kazama and Schornstein at the
Virginia Institute of Marine Sciences (7) have re-
ported and described the first herpes-type virus
from a lower organism, an estuarine fungus.
They were able to manipulate the fungus cells in
such a way in culture as to produce permissive
and nonpermissive host cells in reference to the
virus.
Farley and others (6) reported a herpes-type
virus from oysters that were exposed to ab-
normally high temperature in the effluent water
of a power plant. There was considerable
mortality in these oysters concurrent with the
finding of herpes inclusion bodies with electron
microscopy and the finding of the virus particles
in the infected cells.
Finally, I would like to talk briefly about the
presumed NPV of pink shrimp which we have
found recently in shrimp from the Gulf of
Mexico. The pink shrimp is one of three com-
mercially valuable shrimp in the Gulf and in the
Atlantic. The shrimp hepatopancreas is a large
yellowish organ beneath the dorsal carapace of
the shrimp. This is the organ that is infected
by a rod-shaped virus (3,4).
Fresh squash preparations of tubules from the
hepatopancreas of the shrimp demonstrate the
polyhedral inclusion body (PIB) in situ in one
of the epithelial cells. This is the way the virus
first was found with light microscopy.
The tetrahedral inclusion body grows from a
size that is imperceptible with a light micro-
scope until it reaches a size eventually at which
it ruptures and destroys the cell affected.
With an electron microscope, one sees the
characteristic triangular form of the PIB in two
dimensions. If you section a tetrahedron, the
only thing you can get is a triangle, any way you
look at it, and this is what you get with thin
sections. The virus particles occluded within
the matrix of the occlusion body are apparent.
If shrimp with from 0 to 10% prevalence in
the initial sample were held under abnormally
crowded conditions in a closed system for 30 to
40 days, then an increase in virus prevalence
occurred, what we call an increase of prevalence
of PIBs in that sample. This may be explained
in several ways. The stress of crowding plus the
proximity of shrimp to one another in a closed
system, as well as cannibalism, would enhance
transfer of the virus from shrimp to shrimp. In
nature shrimp are distributed over the bottom
of even fertile fishing grounds in less density
than they would be in an aquacultural or an
aquarium system. Therefore, in nature trans-
mission might not be facilitated or might not
occur rapidly because of the dilution problems
of larger volumes of water for infective stages
and also because of the presence of predators
which eat dying, dead, and weakened shrimp
quickly in a natural environment, thus taking
them out of the presence of other shrimp that
might feed on them and get the virus.
One must consider the laboratory situation
versus the natural situation. This requires an
understanding of the ecology of the organism
one is working with, in this case, shrimp. The
adult shrimp migrates into the ocean and deposits
its eggs. The larvae hatch, and metamorphosis
occurs similar to that of insects in many respects.
Larval stages are instars equivalent to those
in insects. In the past, we have worked only
with the adult shrimp. Recently, we have
found larval shrimp (protozoea and mysid stages)
to be heavily infected with the virus. In any
test situation (for the safety of viruses), I
think these larval stages should be considered
because they are feeding stages and may be
more susceptible. They do feed upon detritus
and organisms therein.
In conclusion, the major new evidence that
emerges here is that many groups of inverte-
brates are capable of harboring viruses that
formerly were studied only in more obvious
insect hosts. Thus, we may need to broaden
our views on virus-host concepts and seek more
widely for host-virus interactions.
Perhaps what I am trying to say is summed up
very generally in a paraphrased Shakesperean
quote: There are more things beneath the sun
than are dreamt of in our philosophies.
LITERATURE CITED
1. Bang, F. B. 1971. Transmissible disease, probably
viral in origin, affecting the amebocytes of the
European shore crab, Carcinus maenas. Infect.
Immun. 3:617-623.
2. Bonami, J. R., and C. Vago. 1971. A virus of a new
type pathogenic to Crustacea. Experientia 27:1363.
3. Couch, J. A. 1974. Free and occluded virus, similar
to baculovirus, in hepatopancreas of pink shrimp.
Nature (London) 247:229-231.
329
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DISCUSSION
113
4. Couch, J. A. 1974. An enzootic nuclear polyhedrosis
virus of pink shrimp: ultrastructure, prevalence, and
enhancement. J. Invertebr. Pathol. 24:311-331.
5. Dougherty, E. C., D. J. Ferral, B. Brody, and M. L.
Gotthold. 1963. A growth anomaly and lysis with
production of virus-like particles in an axenically
reared microannelid. Nature (London) 198:973-975.
6. Farley, C. A., W. G. Banfield, G. Kasnic, Jr., and W. S.
Foster. 1972. Oyster herpestype virus. Science 178:
759-776.
7. Kazama, F. Y., and K. L. Schorastein. 1972. Herpestype
virus particles associated with a fungus. Science
177:696-697.
8. Vago, C. 1966. A virus disease in Crustacea. Nature
(London) 209:1290.
General discussion
DR. HEIMPEL: I wanted to ask Dr. Couch if the
shrimp died.
DR. COUCH: Yes, they died from the virus
disease. So far, in our short-term experience with
this virus, we found that we can bring shrimp to the
laboratory and hold them and we get a certain pro-
portion of the sample dying. We look at these for
the virus. We find that certain ones that have died
or are moribund are heavily infected with the virus.
We find others exhibiting similar symptoms that
show no patent infections at all. Mostly, we used
the light microscope for determining infection and the
presence of PIBs.
DR. HEIMPEL: I would like to ask Dr. Wolf, when
you were doing your experiments with fish tissue
cultures did you back them up with bioassays?
DR. WOLF: I did not do that. Other people have
done it. There are no latent viruses in the fish and
amphibian cell lines that were used.
DR. HEIMPEL: When you applied the insect
viruses, did you do a back up with electron micros-
copy?
DR. WOLF: No.
DR. ENGLER: I would like to make a comment on
Dr. Couch's paper. One point which we should stress:
in animals which have a larval stage, we should look
at the larval stages as well and determine what
viruses do to them. To do this, could you maintain
in the laboratory the whole life cycle of the shrimp
in order to use it as a test system for viruses?
DR. COUCH: The shrimp can be cultured. In the
Galveston Laboratory of the National Oceanic and
Atmospheric Administration they are doing this in
attempts to establish aquaculture methods from egg
to egg. They are not easy to maintain, and there are
a lot of problems involved. There are other disease
factors and nutritional problems. There is another
candidate perhaps, the grass shrimp, which is a
small estuarian species (Paleonetes pugio). These
really do not leave the estuary to go into the open
ocean to spawn like the larger penaeid shrimps;
therefore, they can be grown more easily in cultural
closed laboratory systems. The female bears the eggs
rather than releasing them directly into the environ-
ment. The grass shrimp therefore would be an ideal
small experimental animal, and possibly some fresh-
water shrimps, too.
DR. IGNOFFO: What has been done to survey
these aquatic animals for natural presence and inci-
dence of virus? I think the question of disease-free
has been brought up in terms of mammalian host
systems, tissue cells, and everything else. Does the
grass shrimp contain any known or described viruses?
DR. COUCH: None, and I think there has been
very little research or pathology done on grass
shrimp. In our laboratory we are concerned chiefly
with the toxic effect of chemicals and pesticides,
and we came across this virus indirectly working on
this. It is something you have to consider when you use
animals in experiments for toxicity tests; one should
know that kind of natural diseases they may have.
We have epizootics occurring with fungi and proto-
zoa in our test animals a good portion of the time. We
have to be able to distinguish between the effects of a
natural disease complex in these test animals and the
effects of the toxicant or whatever variable we are
testing at the time.
DR. JACQUES: Have you any idea, or would you
like to hazard a guess, as to how this virus might be
transmitted within the population of shrimp?
DR. COUCH: It would only be a guess because
the most logical explanation from my experience is
that it is direct transmission. Shrimp are notoriously
cannabalistic. There is no hesitation on their part to
feed on any organ or the body of a dying comrade and
they can easily consume the hepatopancreas. There
is also the possibility that PIBs that are extruded from
the hepatopancreas nuclei may pass into the midgut
cells or into the feces and be deposited on the ocean
bottom, and this provides another possible source
for transmission of virus.
Autoinfection may occur within shrimp themselves
that are infected because a fairly large number of
these cells, with electron microscopy, that do not
have inclusion bodies in them do have these virus
particles, the rods, so there may be an autoinfective
cycle within the shrimp for other cells as well as pro-
duction of polyhedral bodies.
But I think probably direct transmission through
feeding is most likely. We have no evidence that the
larval stages become infected and then maintain
the virus in a latent form for a long period and
suffer mortality in different stages in the larval cycle.
DR. WOLF: I would like to comment on the pres-
ence or absence of pathogens as assayed by any
method. In North America it is difficult to buy a
commercial amphibian that has not had prior antigenic
experience or an active infection with, or the carrier
state of, the FV 3 virus.
So this fact also plays an important role in any
assay system. If the animals have had prior antigen
experience, they have mounted an immune response,
and, no matter what a person puts into the culture
in the way of the same agent, one may get no response,
not because there is any resistance but because there
has been an immune response. I suppose this may
apply to invertebrates as well.
DR. COUCH: Invertebrates have resistance mech-
anisms of various kinds, but not an antibody response
that we know of.
DR. HEIMPEL: I would like to put into the record
that, while we were doing the brown shrimp testing at
Dr. Spark's laboratory, Sam1 Ray, who is an oyster
330
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114
BACULOVIRUS INSECT PEST CONTROL
pathologist at the Texas A&M Marine Laboratory,
did some injection and feeding tests on oysters as
well with negative results. Most of these tests
were backed up by electron microscopy, and we
found no evidence of virus particles in the cells of
either the oyster or the shrimp.
DR. SUMMERS: I would like to ask Dr. Heimpel
what was the nature of the inoculum that they used.
If I remember correctly, it was the purified viruses from
the inclusion bodies. Perhaps you should have also
tested the inoculum that has been shown to be the
most effective and efficient in infecting another
system, which would be infectious hemolymph or tissue
culture-derived virus.
DR. HEIMPEL: Dr. Summers is right. We used
only freed virus and the polyhedra. To use the
infectious hemolymph, of course, gets into the area
of tricky laboratory experiments. Actually, the
animal is most likely to encounter polyhedra in
nature.
DR. IGNOFFO: That is a good point! But re-
member, we can also get infection in vivo in a
normal host using the degraded inclusion body mix-
ture. The use of infectious units other than inclu-
sion bodies has more implication for tissue culture
studies. Infection of cell culture was only obtained
with infectious hemolymph, infected tissue culture
supernatant, and possibly "treated" disrupted virions,
but never with intact inclusion bodies. In tests in
nontarget vertebrate systems, intact inclusion bodies,
degraded inclusion body mixtures, virions, and in-
fectious hemolymph have been employed in evaluations
of their safeness.
DR. SUMMERS: Normally one uses the natural
route for exposure, and that includes per os; it may
also include inhalation. But you might want to be even
more critical. I think injection could put the virus in a
more favorable site of infection in order to test for
"ability" to infect. That is why that approach should
be used.
DR. HEIMPEL: I think Dr. Summers is right, even
if the test is a sort of laboratory curiosity. If we
use infectious virus, we should perform safety tests to
cover all possibilities of infection.
331
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Reprinted from Journal of Toxi-
cology and Environmental Health,
Vol. 2: 169-178, 1976, with
permission of the Hemisphere
Publ. Corp., Washington
HEPTACHLOR: UPTAKE, DEPURATION, RETENTION, AND METABOLISM
BY SPOT, LEIOSTQMUS XANTHURUS
Steven C. Schimmel, James M. Patrick, Jr., and Jerrold Forester
Contribution No. 264
333
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HEPTACHLOR: UPTAKE, DEPURATION, RETENTION,
AND METABOLISM BY SPOT, Leiostomus xanthurus
Steven C. Schimmel, James M. Patrick, Jr., Jerrold Forester
U.S. Environmental Protection Agency, Environmental
Research Laboratory, Sabine Island, Gulf Breeze, Florida
The estuarine fish, spot fLeiostomus xanthurusj, was exposed to 0.27, 0.52, 1.01, 1.99,
and 3.87 ng/liter technical grade heptachlor (65% heptachlor, 22% trzns-ch/ordane, 2%
cis-ch/ordane, 2% nonachlor, and 9% unidentified compounds) for 24 days in a
flowthrough bioassay, followed by 28 days in heptachlor-free seawater. Concentrations
of heptachlor, heptachlor epoxide, and trans- and c\s-ch/ordane in edible tissues were
monitored at day 3 and weekly thereafter throughout the bioassay and at the end of
the postexposure period. All four chemicals were accumulated by spot. Maximum
concentrations of heptachlor were observed on day 3; maximum concentrations of the
other three compounds were observed on day 17. The average bioconcentration
factors for heptachlor and trzns-ch/ordane were 3,600 and 4,600, respectively. Only
10% or less of the maximum concentrations of heptachlor, heptachlor epoxide, and
trzns-chlordane accumulated during the exposure period remained after 28 days in
pesticide-free seawater; an average of 35% of the cis-chlordane remained. Relative total
amounts of heptachlor and c\s-chlordane changed during the exposure and post-
exposure periods. Nearly all of the heptachlor was eliminated or metabolized to its
epoxide. Cis-chlordane, which averaged 4-7% of the total residues (chlordanes and
heptachlors) in edible tissues during the exposure, increased to 18-23% of the total
residues by the end of the postexposure period.
INTRODUCTION
Although not intended for use in the aquatic environment, the
organochlorine insecticide heptachlor has been reported in fresh waters
and estuaries of the United States. Heptachlor has been found in water,
sediment, and biota of a lake (Hannon et al., 1970) and in creek and river
sediments (Barthel et al., 1969). Freshwater fishes accumulated the
heptachlor metabolite heptachlor epoxide in their eggs (Johnson and
Morris, 1974), muscle, and whole-body tissues (Hannon et al., 1970) in
excess of 0.8 Mg/g- 'n estuaries heptachlor has been found in water (Casper
et al., 1969; Fay and Newland, 1972), sediment. (Casper et al., 1969; Farb
and Moore, 1971), oysters (Casper et al., 1969), and fish (Smith and Cole,
1970). Heptachlor residues in muscle of winter flounder (Pseudo-
This is contribution no. 264, Environmental Research Laboratory, Gulf Breeze.
We thank Johnny Knight for conducting routine heptachlor water analyses and Steven Foss
for preparing the illustrations.
Requests for reprints should be sent to Steven C. Schimmel, EPA, Environmental Research
Laboratory, Sabine Island, Gulf Breeze, Florida 32561.
169
journal of Toxicology and Environmental Health, 2:169-178, 1976
Copyright © 1976 by Hemisphere Publishing Corporation
335
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17Q S. C. SCHIMMELET AL.
pleuronectes americanus] have exceeded 1.5 Mg/g (Smith and Cole, 1970).
The U.S. Food and Drug Administration (1973) established 0.3 Mg/g as the
maximum allowable concentration in edible fish and shellfish.
Little is known of the rate of uptake, depuration, or retention of
heptachlor in marine fish and shellfish. Wilson (1965) reported that
oysters exposed to 0.01 mg/liter of the chemical for 10 days accumulated
176 Mg/g, a concentration factor of 17,600 (concentration of chemical in
tissues divided by desired concentration in water). Schimmel et al. (1976)
found that concentration factors in sheepshead minnows (Cyprinodon
variegatus), pinfish (Lagodon rhomb/odes), and spot (Leiostomus xan-
thurus] exposed for 96 hr ranged from 2,800 to 21,300. No information
has been reported on the rate of depuration, retention, or metabolism of
heptachlor for an estuarine fish in bioassays that exceeded 96 hr.
In this paper we discuss experiments to determine (a) the rate of
uptake in spot of several compounds within technical heptachlor (includ-
ing heptachlor, c/s-chlordane, and fram-chlordane) during a 24-day expo-
sure, (b) the rate of metabolism of heptachlor to heptachlor epoxide, and
(c) the rate of depuration or retention of these compounds during a
28-day postexposure holding period.
METHODS AND MATERIALS
Spot were collected near the Gulf Breeze Laboratory. The fish (25-40
mm standard length; x = 32 mm) were held for 10 days in the laboratory
for acclimation and observation prior to exposure. Spot were fed frozen
adult brine shrimp (Artemia sa/ina) once daily during acclimation and the
bioassay. Those not eaten after 15 min were removed. The brine shrimp
contained no detectable (gas chromatographic analysis) organochlorine
pesticides or polychlorinated biphenyls.
Juvenile spot were exposed to technical grade heptachlor for 24 days,
using a modification of the Mount and Brungs (1967) delivery apparatus.
Our apparatus had an additional control cell that delivered seawater with
the acetone carrier. Raw seawater was pumped from Santa Rosa Sound,
Florida, through a gravel-filled swimming pool filter and a 15Mm pore
cartridge filter into a reservoir in the laboratory. In the reservoir the water
temperature was adjusted to 25°C (±1.5°C) and salinity to 20 ppt (±1.5
ppt), using a controller described by Bahner and Nimmo (1975). Water
was then pumped to the delivery apparatus, which cycled approximately
250 times each day. Each cycle delivered 1 liter of seawater to each of
five 30-liter experimental and two control aquaria. A stock solution of
heptachlor dissolved in reagent-grade acetone provided the following
desired concentrations of heptachlor in seawater: 0.27, 0.52, 1.01, 1.99,
and 3.87 Mg/liter. Two 50-ml syringes, activated by a mechanical injector,
delivered 0.026 ml acetone and heptachlor or acetone only during each
cycle. Thirty fish were exposed in each of seven aquaria.
Five fish were removed from each aquarium for chemical analyses on
336
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SPOT EXPOSED TO HEPTACHLOR
171
day 3 and weekly thereafter for 3 wk. Each fish was rinsed with acetone
to remove any adsorbed pesticide, measured, and weighed. The head and
viscera (offal) were separated from the associated muscle, vertebrae, fins,
and scales ("edible tissues"). Offal and edible tissue of fish from a single
concentration were each pooled for separate chemical analysis. Data on
pesticide content of edible portion and offal were summed to determine
the total body burden of the chemicals.
At the end of the exposure the animals were held for 28 days longer
in the same aquaria containing flowing seawater without heptachlor.
Hereafter the postexposure period is referred to as the "depuration"
period, which is the time in which the toxicant was no longer added to
the exposure water and there was a gradual, but not necessarily total, loss
of toxicant stored in tissues of the test animal. At termination the fish
were prepared as before and analyzed chemically.
Chemical analyses of heptachlor in water and tissues were conducted
by electron-capture gas chromatography. Samples were analyzed using
methods described by Schimmel et al. (1976). All samples were fortified
with an internal standard (2,3,4,5,6,2',5'-heptachlorobiphenyl) before
analysis to evaluate the integrity of the method. Extracts of tissues
fortified with heptachlor, heptachlor epoxide, c/s-chlordane, and trans-
chlordane gave recoveries greater than 90%. Concentrations were calculated
on a wet-weight basis without correction for percentage recovery. Gas
chromatographic analyses of technical heptachlor used in these experi-
ments showed heptachlor, 65%; fr-o/7s-chlordane, 22%; c/s-chlordane, 2%;
nonachlor, less than 2%; and unidentified compounds, 9%. The identified
compounds were confirmed by mass spectrometry.
RESULTS
Heptachlor, heptachlor epoxide, fraws-chlordane, and c/s-chlordane
were accumulated in offal and edible tissues of spot during the 24-day
exposure (Table 1). Offal contained approximately three times as much of
TABLE 1. Concentrations of Heptachlor (Hept.), Heptachlor Epoxide (H.E.), frans-Chlordane
(&ww-Chlor.), and c/s-Chlordane (c/s-Chlor.) in Edible Tissues and Total Body Burden (Edible Tissues
and Offal) in Spot, Leiostomus xanthurus, Exposed to Technical Heptachlor for 24 Days in
Flowthrough Bioassay
Heptachlor
concentration in
water (/jg/liter)
Concentrations of chemicals in edible tissues
(Mg/g)
Total body burden
-------
5.0
[HEPTACHLOR EPOXIDE
Depuration
10 17 24
TIME (days)
52
FIGURE 1. The bioconcentration of heptachlor and heptachlor epoxide in edible tissue of spot
(Leiostomus xanthurus) exposed in a 24-day bioassay followed by a 28-day depuration period.
Each plotted line represents bioaccumulation in a single exposure concentration (measured
heptachlor in water, /ug/liter).
172
338
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SPOT EXPOSED TO HEPTACHLOR
173
the four compounds as the edible portion. Because the U.S. Food and
Drug Administration (1973) has set maximum allowable limits of hepta-
chlor in edible tissues, we believed that we should emphasize evaluation of
pesticide concentrations in edible tissues of the spot in relation to
movement of the chemical through the food web to humans.
Death occurred only in the 3.87 jug/liter (2.55 jug/liter measured)
concentration. In this concentration no fish survived beyond 6 days.
Concentrations of heptachlor in edible tissues did not increase signifi-
cantly after the third day of exposure to any concentration tested. After
the 28-day depuration period, approximately 10% of the accumulated
heptachlor remained in edible tissue (Fig. 1). The average heptachlor
concentration factor (based on measured' heptachlor concentrations in
water) in edible tissues of spot was 1,300 and ranged from 1,000 to 1,500
in all concentrations over the exposure period (Fig. 2).
10 17
TIME (DAYS)
24
FIGURE 2. Average concentration factors (average of the concentrations of chemical in tissues
divided by measured concentrations of chemical in water) of heptachlor and trans-Mordant in
muscle tissue of spot (Leiostomus xanthurus) continuously exposed to technical grade heptachlor
for 24 days.
339
-------
2.0
TRANS -CHLORDANE^
Depuration
FIGURE 3. The bioconcentration of fro/75-chIordane and c/s-chlordane in edible tissue of spot
(Leiostomus xanthurus) exposed in a 24-day bioassay followed by a 28-day depuration (measured
r/ww-chlordane in water for fro/7s-chlordane data, measured heptachlor in water for c/s-chlordane
data).
174
340
-------
SPOT EXPOSED TO HEPTACHLOR 175
Maximum concentrations of c/s-chlordane in spot were detected on
day 17 of exposure. At the end of the depuration period 35% of the
maximum concentration of c/s-chlordane remained in edible tissues,
whereas only 10% of the maximum concentrations of the other com-
pounds remained (Fig. 3).
Heptachlor was metabolized to heptachlor epoxide at all concentra-
tions tested (Fig. 1). Davidow and Radomski established heptachlor
epoxide as a biodegradation product of heptachlor in 1953; this metab-
olite is commonly found .in animals after environmental application of
heptachlor (Bonderman and Slach, 1972; Hannon et al., 1970; Johnson
and Morris, 1974; Lichtenstein et al., 1970; Miles et al., 1969; Oberheu,
1970). We found maximum concentrations of heptachlor epoxide after 17
days of exposure, but less than 10% of the maximum after 28 days of
depuration.
7><7/7s-chlordane was accumulated within edible tissues of spot; highest
concentrations were detected on day 17 of exposure (Fig. 2 and 3). The
average concentration factor for fraws-chlordane in edible tissues was two
to four times greater than that for heptachlor. Concentration factors of
fra/75-chlordane in whole-body tissues, however, were only 1.3 times that
of heptachlor (Table 2).
Relative quantities of heptachlor, heptachlor epoxide, fra/?s-chlordane,
and c/s-chlordane in edible tissues changed during the entire test period
(Fig. 4). After 3 days of exposure heptachlor concentrations averaged 52%
of total residues, heptachlor epoxide 25%, Zram-chlordane 18%, and
c/s-chlordane 4%. After 24 days heptachlor averaged 39%, heptachlor
epoxide 35%, mws-chlordane 20%, and c/s-chlordane 6%. After 28 days of
depuration the relative amount of heptachlor concentrations decreased to
10% of the total residues, while heptachlor epoxide increased to 44%,
TABLE 2. Concentration Factors0 for Heptachlor and trans-
Chlordane in Spot, Leiostomus xanthurus, Exposed to Technical
Heptachlor for 72 hr in a Flowthrough Bioassay
Chemical concentration
in water (jug/liter) Concentration factor
Heptachlor mws-Chlordane Heptachlor ttws-Chlordane
0.14
0.26
0.58
1.03
2.55
0.04
0.07
0.15
0.24
0.67
2,200
2,800
3,900
4,800
4,500
x = 3,600
3,300
3,600
4,800
6,400
5,100
x = 4,600
"Concentrations of exposure chemical in whole-body tissues
divided by the amount of chemical measured in water.
341
-------
176
S. C. SCHIMMEL ET AL.
HEPTACHIOR
TRANS-CHIORDANE
3-DAY EXPOSURE
KMa HEPT. EPOXIDE
| | CIS-CHLORDANE
24-DAY EXPOSURE 28-DAY DEPURATION
0.27 0.52 1.01 1.99 3.87 0.27 0.52 1.01 1.99 0.27 0.52 1.01 1.99
CONCENTRATION IN WATER Ug/1)
FIGURE 4. Relative amounts of heptachlor, heptachlor epoxide, f/ws-chlordane, and cw-chlordane
in edible tissues of spot (Leiostomus xanthurus] exposed to technical grade heptachlor concentra-
tions (desired) for 24 days, followed by a 28-day depuration period.
mym-chlordane to 25%, and c/s-chlordane to 20%. The amount of
heptachlor epoxide, compared .with heptachlor, was expected to increase
with time because of metabolism of the parent compound.
DISCUSSION
Heptachlor bioconcentration. factors in whole-body tissues of spot in
this study were slightly lower than those of spot reported by Schimmel et
al. (1976). In the latter study spot were exposed to technical heptachlor
(0.5-1.25 jug/liter) for 96" hr in a different apparatus. The average
bioconcentration factor was 7,400 (range 3,000-13,800). In the present
72-hr study the average' bioconcentration.'factor was 3,600 (range 2,200-
4,500; Table 2).,Greater whole-body concentrations in spot exposed for
96 hr may be due to (a) differences in fat content of the fish in both
tests, (b) greater relative amount of solvent in the 96-hr test, or (c) the
24-hr difference in the duration of exposure. It should be noted, however,
that the ranges in bioconcentration factors of both tests did overlap.
7>tfA7s-chlordane bioconcentration factors in whole-body tissues of spot
in this study were also lower than those in spot reported by Schimmel et
al. (1976). In the latter bioassay, spot were exposed to mws-chlordane
342
-------
SPOT EXPOSED TO HEPTACHLOR 177
(0.15-0.52 jug/liter) for 96 hr. The average bioconcentration factor was
8,100 (range 3,700-14,800). In the present 72-hr study the average
bioconcentration factor was 4,600 (range 3,600-6,400; Table 2). Again,
lesser accumulation in our test may be due to differences in the fishes' fat
content, shorter exposure time (72 hr vs. 96 hr), or a higher pesticide-to-
solvent ratio. Once again it should be noted that there was considerable
overlap in ranges of bioconcentration factors in the two studies, particu-
larly in the lower concentrations. Parrish et al. (1976) exposed pinfish
(Lagodon rhomboides] and sheepshead minnows (Cyprinodon variegatus)
to technical chlordane (a mixture including trans- and c/s-chlordane) and
found that chlordane in whole-body tissues of pinfish was accumulated to
an average of 6,200 times that measured in exposure water; in sheepshead
minnows this averaged 15,100 times.
The metabolism of heptachlor to its epoxide shown in our study and
the environmental occurrences of heptachlor epoxide are significant
because heptachlor epoxide is as toxic as heptachlor to some estuarine
organisms (Schimmel et al., 1976). The 96-hr LCSO of analytical-grade
heptachlor to pink shrimp (Penaeus duorarum] was 0.03 /zg/liter (95%
Cl = 0.02-0.04); the 96-hr LCSO value for heptachlor epoxide was 0.04
(95% Cl = 0.001-0.1).
Heptachlor in water, at measured concentrations > 0.14 /ug/liter, was
accumulated with its epoxide to > 0.3 jug/g in edible tissues of spot;
maximum allowable concentration of heptachlor and its epoxide in fish
and shellfish for human consumption is 0.3 Mg/g- Concentration of
heptachlor in edible tissues some fish collected from the estuarine
environment has exceeded this allowable level (Smith and Cole, 1970).
REFERENCES
Bahner, L. H. and Nimmo, D. W. 1975. A salinity controller for flowthrough bioassays. Trans. Am.
Fish. Soc. 104(2):388-389.
Barthel, W. E., Hawthorne, J. C., Ford, J. H., Bolton, G. C., McDowell, L. L., Grissinger, E. H. and
Parsons, D. A. 1969. Pesticide residues in sediments of the lower Mississippi River and its
tributaries. Pest. Monit. ]. 3(1):8-66.
Bonderman, D. P. and Slach, E. 1972. Appearance of l-hydroxychlordene in soil, crops and fish. J.
Agric. Food Chem. 20(2):328-331.
Casper, V. L., Hammerstrom, R. J., Robertson, E. A., Jr., Bugg, J. C., Jr., and Gaines, J. L. 1969.
Study of chlorinated pesticides in oysters and estuarine environment of the Mobile Bay area.
Gulf Coast Marine Health Sciences Laboratory. Ala. Water Improve. Comm., Ala. State Dep.
Public Health, and Ala. Dep. Conserv.
Davidow, B. and Radomski, J. L. 1953. Isolation of an epoxide metabolite from fat tissues of dogs
fed heptachlor. J. Pharmacol. Exp. Ther. 107:259-265.
Farb, R. and Moore, B. G. 1971. A preliminary investigation of pollution and its distribution in
Mobile Bay. /. Ala. Acad. Sci. 42(3):138.
Fay, R. R. and Newland, L. W. (1972). Organochlorine insecticide residues in water, sediment, and
organisms, Aransas Bay, Texas-September 1969-June 1970. Pest. Monit. J. 6(2):97-102.
Hannon, M. R., Greichus, Y. A., Applegate, R. L. and Fox, A. C. 1970. Ecological distribution of
pesticides in Lake Poinsett, South Dakota. Trans. Am. Fish. Soc. 99(3):496-500.
343
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178 S. C. SCHIMMEL ETAL.
Johnson, L. G. and Morris, R. L. 1974. Chlorinated insecticide residues in the eggs of some
freshwater fish. Bull. Environ. Contam. Toxicol. 11(6):503-510.
Lichtenstein, E. P., Schultz, K. R., Fuhremann, T. W. and Liang, T. T. 1970. Degradation of aldrin
and heptachlor in field soils during a ten-year period. J. Agric. Food Chem. 18(1):100-106.
Miles, J. R. W., Tu, C. M. and Harris, C. R. 1969. Metabolism of heptachlor and its degradation
products by soil microorganisms./. Econ. Entomol. 62(6): 1334-1 338.
Mount, D. I. and Brungs, W. A. 1967. A simplified dosing apparatus for fish toxicological studies.
Water Res. 1:21-29.
Oberheu, J. C. 1970. Effects on fish and wildlife of heptachlor applied to eradicate the sugarcane root
weevil in Apopka, Florida. Proc. 24th Annu. Conf. Southeast Assoc. Game Fish Commun. p.
194-200.
Parrish, P. R., Schimmel, S. C., Hansen, D. J., Patrick, J. M., Patrick, J. M., Jr. and Forester, J.
1976. Chlordane: Effects on several estuarine organisms. J. Toxicol. Environ. Health.
1:485-494.
Schimmel, S. C., Patrick, J. M. Jr. and Forester, J. 1976. Heptachlor: Toxicity to and uptake by
several estuarine organisms./. Toxicol. Environ. Health 1:955-965.
Smith, R. M. and Cole, C. F. 1970. Chlorinated hydrocarbon insecticide residues in winter
flounder, Pseudopleuronectes americanus, from the Weweantic River Estuary, Massachusetts. /.
Fish. Res. Board Can. 27(12):2374-2380.
U.S. Food and Drug Administration. 1973. Administrative guidelines manual 7420.09. 1 January,
1973.
Wilson, A. J. 1965. Ann. Rep. Bur. Comm. Fish. Biol. Lab., U.S. Fish Wildl. Serv. Circ. 247:6-7.
Received March 4, 1976
Accepted April 22, 1976
344
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Reprinted from Journal of Toxi-
cology and Environmental Health,
Vol. 1: 955-965, 1976, with
permission of the Hemisphere
Publ. Corp., Washington
HEPTACHLOR: TOXICITY TO AND UPTAKE BY ESTUARINE ORGANISMS
Steven C. Schimmel, James M. Patrick, Jr., and Jerrold Forester
Contribution No. 265
345
-------
HEPTACHLOR: TOXICITY TO AND UPTAKE
BY SEVERAL ESTUARINE ORGANISMS
Steven C. Schimmel, James M. Patrick, Jr., Jerrold Forester
U. S. Environmental Protection Agency, Environmental Research
Laboratory, Sabine Island, Gulf Breeze, Florida
Technical-grade heptachlor (65% heptachlor, 22% trans-chlordane, 2% c\s-chlordane, and
2% nonachlor) was tested in 96-hr bioassays to determine its toxicity to estuarine
animals. The test organisms and the 96-hr LCSO or ECSO s (based on measured
concentrations in water) are as follows: American oyster fCrassostrea virginicaj, 7.5
tig/liter, pink shrimp (Penaeus duorarumj, 0.77 ^g/liter; grass shrimp fPalaemonetes
vulgarisj, 7.06 m/liter; sheepshead minnow fCyprinodon variegatus,), 3,68 tig/liter;
pinfish ('Lagodon rhomboides^, 3.77 ^filter; and spot fLeiostomus xanthurus,), 0.55
ng/liter. Analytical-grade heptachlor (99.8% heptachlor) and heptachlor epoxide (99%)
were also studied. The analytical-grade heptachlor 96-hr LC50 for pink shrimp and spot
was 0.03 vy/liter and 0.86 ng/'liter, respectively, while that for pink shrimp exposed to
heptachlor epoxide was 0.04 tig/liter. Heptachlor was accumulated and some
metabolized to its epoxide by all animals tested. Fish and oysters accumulated
heptachlor in their tissues 2,800-21,300 times the measured concentration in water;
shrimp, only 200-700 times.
INTRODUCTION
Heptachlor, a persistent organochlorine pesticide, has been used
primarily as a crop insecticide. Over 500,000 kg were applied to agri-
cultural fields in 1971 (Andrilenas, 1974).
Heptachlor and its metabolite, heptachlor epoxide (Davidow and
Radomski, 1953), have been found in freshwater, estuarine, and marine
systems. Barthel et al. (1969) found 2.4 jug/liter heptachlor in river water
and 11 Mg/g in river sediments near Memphis, Tennessee. Heptachlor and
heptachlor epoxide were reported in most aquatic animals of Lake
Poinsett, South Dakota (Hannon et al., 1970). In the estuarine and marine
environments, oysters, water, and sediments in Mobile Bay, Alabama,
contained detectable levels of heptachlor (Casper et al., 1969). Smith and
Cole (1970) reported over 1.5 jug/g heptachlor-and 0.5 jug/g heptachlor
epoxide (wet weight) in muscle of the winter flounder, Pseudopleuronectes
americanus, collected in Massachusetts.
Heptachlor was reported highly toxic to nontarget marine organisms.
The 96-hr LCSO (the amount of heptachlor in water estimated to kill 50%
This paper is contribution no. 265, Gulf Breeze Environmental Research Laboratory.
Requests for reprints should be sent to Steven C. Schimmel, U.S. Environmental Protection
Agency, Environmental Research Laboratory, Sabine Island, Gulf Breeze, Florida 32561.
955
journal of Toxicology and Environmental Health, 1:955-965, 1976
Copyright © 1976 by Hemisphere Publishing Corporation
347
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956 S. C. SCHIMMEL ET AL.
of the test organisms in 96 hr) in static bioassays with estuarine species
was 8 Mg/'iter for sand shrimp, Crangon septemspinosa (Eisler, 1969), and
0.8 Mg/'iter for the bluehead, Thallassoma bifasciatum (Eisler, 1970). For
juvenile striped bass (Morone saxati/is) in flowthrough bioassays, the
96-hr LC5o was 3 jug/liter (Korn and Earnest, 1974).
Additional data on bioaccumulation and toxicity of heptachlor and its
epoxide are needed to evaluate better the effects of heptachlor on marine
and estuarine environments. Also, several pesticides, including heptachlor,
are being considered for re-registration by the U.S. Environmental Protec-
tion Agency. In this paper we report (a) the 96-hr EC50 (amount of
chemical estimated to reduce shell deposition by 50%) of technical-grade
heptachlor to the American oyster (Crassostrea virginica], (b) the 96-hr
LC50 of technical-grade heptachlor to pink shrimp (Penaeus duorarum],
grass shrimp (Palaemonetes vulgaris], sheepshead minnow (Cyprinodon
variegatus], pinfish (Lagodon rhomboides], and spot (Leiostomus
xanthurus), (c) the 96-hr LC50 of analytical-grade heptachlor to pink
shrimp and spot, and (d)'the 96-hr LC50 of heptachlor epoxide to pink
shrimp. We also report the amount of chemical and metabolites accumu-
lated by these organisms in 96 hr.
METHODS AND MATERIALS
Test Animals
All test animals except pink shrimp were collected in estuarine waters
near the Gulf Breeze Environmental Research Laboratory, Gulf Breeze,
Florida, and acclimated to laboratory test conditions for at least 10 days.
Pink shrimp were purchased from a local bait dealer and acclimated
similarly. Mortality of animals did not exceed 1% of the stock in the 48
hr immediately preceding the test, nor did any animals exhibit disease or
abnormal behavior during the acclimation period. The size of test animals,
the concentrations tested, and the temperature and salinity for each test
are listed in Table 1.
Test Conditions
Acute toxicity of technical-grade heptachlor (65% heptachlor, 22%
fra/7s-chlordane, 2% c/s-chlordane, and 2% nonachlor), analytical-grade
heptachlor (99.8%), and heptachlor epoxide (99%) was determined by
exposing 20 animals per aquarium to different concentrations for 96 hr in
flowthrough bioassays similar to those of Lowe et al. (1972). One 30-liter
aquarium was used for each concentration. Animals were not fed during
the 96-hr tests; however, animals used in the technical-grade heptachlor
tests could obtain food (plankton and other paniculate matter) from the
unfiltered seawater. Animals exposed to analytical-grade heptachlor and
heptachlor epoxide were maintained in filtered seawater.
348
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HEPTACHLOR TOXICITY IN ESTUARINE ORGANISMS 957
TABLE 1. Acute (96-hr) Test Conditions for Estuarine Organisms Exposed to Technical-Grade
Heptachlor, Analytical-Grade Heptachlor, and Heptachlor Epoxide
Chemical
Technical-
grade
heptachlor
(65%)
Analytical-
grade
heptachlor
(99.8%)
Heptachlor
epoxide
(99%)
Species
Crassostrea
virginica
Penaeus
duorarum
Palaemontes
vulgaris
Cyprinodon
variegatus
Lagodon
rhomboides
Leiostomus
xanthurus
Penaeus
duorarum
Leiostomus
xanthurus
Penaeus
duorarum
Length0
(range, mm)
30-32
44-72
11-16
22-29
34-51
22-35
71-97
20-30
62-81
Concentration
(jug/liter)
0.32,1.0,3.2,
10.0,32.0
0.1,0.32,1.0,
3.2,10.0
0.32, 1.0,3.2,
10.0,32.0
6.5, 8.7, 11.5
15.5, 21.0
0.32,1.0,3.2
10.0,32.0
1.15, 1.55, 2.1,
2.8,3.7
0.0046,0.01,
0.021,0.046,
0.1
1.55,2.1,2.8,
3.7,4.9
0.021,0.046,
0.1,0.21,0.46
Temperature
(range, °C)
30.0-32.0
27.5-30.0
28.7-30.0
26.0-27.0
27.5-30.0
23.0-26.0
24.0-26.0
24.5-25.5
24.2-26.5
Salinity
(range, ppt)
24.5-27.0
25.5-29.5
24.5-28.0
20.5-24.5
25.0-31.0
20.0-21.0
20.0-22.0
20.0-22.0
20.0
"Oysters, umbo-to-distal valve edge length;shrimp, rostrum-to-telson length;and fish, standard length.
For the technical-grade heptachlor tests, unfiltered seawater was
pumped from Santa Rosa Sound, Florida, into a constant-head trough in
the laboratory. Seawater was delivered to each aquarium by a calibrated
siphon that delivered 100 liters/hr. Two control and five experimental
aquaria were used in each test. Stock solutions of technical-grade
heptachlor, in reagent-grade acetone, were metered into experimental
aquaria at the rate of 60 ml/hr. Acetone was required as a carrier because
of the extreme insolubility of heptachlor in water (Burchfield et al.,
1965). One control aquarium was provided acetone at the rate of 60
ml/hr; the other, seawater only. Stock solutions of technical-grade
heptachlor were prepared by weight of the chemical in acetone. Although
heptachlor consisted of only 65% of the total, we felt that the biologically
active nature of trans- and c/s-chlordane should be considered and they
were therefore combined with heptachlor as 100% active ingredients. Wide
differences between desired and measured heptachlor in water were due, in
part, to these considerations.
Filtered seawater was used in bioassays with analytical-grade
heptachlor and heptachlor epoxide. Raw seawater was pumped from Santa
Rosa Sound through a sand filter and a 15-/zm polypropylene filter into a
constant-head trough in the laboratory. Water was supplied to each
349
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958 S. C. SCHIMMEL ETAL.
aquarium by a calibrated siphon that delivered 25 liters/hr. Analytical-grade
heptachlor in reagent-grade acetone was metered into experimental aquaria at
the rate of 30 ml/hr; heptachlor epoxide was delivered similarly. Two control
aquaria, one with and one without acetone, were provided for each bioassay.
The 96-hr LCSO and EC50s were determined for both desired and measured
concentrations in water. Desired concentrations were those calculated to be
in water, based on the concentration of the stock solution, plus stock solu-
tion and seawater flow rates. The LC50s, based on measured concentrations,
are those derived by direct chemical analyses of exposure water.
At the end of each 96-hr test surviving animals were sacrified and
removed for residue analyses. Whole fish and shrimp from each concen-
tration were washed with acetone and pooled as single sample. Oyster
meats were removed from their shells and analyzed for toxicant residues.
Statistical Analyses
To determine the concentration of heptachlor required to reduce shell
deposition of exposed oysters by 50% as compared with controls (ECSO),
oyster shell deposition data were analyzed by straight-line graphical
interpolation (APHA, 1971) on log/probit paper. Shrimp and fish mortality
data were analyzed by probit analysis to determine LCSO using the
method of Finney (1971).
Chemical Analyses
Tissues of fish, oysters, and shrimp were weighed in 150 mm X 25
mm (o.d.) screw top test tubes and extracted twice with 5 ml volumes of
acetonitrile for 1 min with a model PT10-ST Willems Polytron1 (Brinkman
Instruments, Westbury, New York). The test tube was centrifuged and the
acetonitrile transfered to a clean 200 mm X 25 mm test tube. After the
second extraction the tissue was rinsed with 5 ml of acetonitrile by
agitation on a Vortex mixer for 30 sec. The test tube was centrifuged and
the acetonitrile supernate was added to the above extracts. This process
was repeated a second time. To the combined extracts 25 ml 2.0%
aqueous sodium sulfate and 5 ml hexane were added. The test tube was
sealed with a Teflon-lined cap and shaken for 1 min. After the solvent
phases separated the upper hexane layer was transferred with a dropping
pipette to a 25-ml Kuderna-Danish concentrator tube. The hexane ex-
traction was repeated 3 times with 5 ml hexane. The combine extracts,
concentrated to about 0.5 ml by evaporation, were transferred to a 200
mm X 9 mm (i.d.) chromatographic corumn containing 3.2 g Florisil
topped with 3.2 g anhydrous sodium sulfate. Heptachlor, heptachlor
epoxide, c/5-chlordane, and f/ww-chlordane were eluted from the column
'Mention of commercial products does not constitute endorsement by the U.S. Environmental
Protection Agency.
350
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HEPTACHLOR TOXICITY IN ESTUARINE ORGANISMS 959
with 20 ml 5.0% ethyl ether in hexane. Extracts were concentrated or diluted
to appropriate volumes for analyses by electron-capture gas chromatography.
Water samples were analyzed by extracting 1 liter with two 100-ml
portions of petroleum ether.
The operating parameters of model 5713 Hewlett-Packard gas
chromatographs were as follows: 183 cm X 2 mm (i.d.) glass columns
packed with 2.0% OV-101 on 100/120 Gas Chrom Q and 0.75% OV-
17:0.97% OV-210 on 100/120 Gas Chrom Q; oven temperature, 200°C;
injector temperature, 200°C; detector (63Ni) temperature, 300°C; argon/
methane carrier gas flow rate, 25 ml/min.
All samples were fortified with an internal standard (2,3,4,5,6,2',5'-
heptachlorobiphenyl) prior to analysis in order to evaluate the integrity
of the results. Extracts of tissues fortified with heptachlor, heptachlor
epoxide, c/s-chlordane, and frW75-chlordane gave recoveries greater than 90%.
Residue concentrations were calculated on a wet weight basis without a
correction factor for percentage recovery. Technical heptachlor used in
these experiments analyzed by gas chromatography contained heptachlor
(65%), f/w7s-chlordane (22%), c/s-chlordane (2%), and nonachlor (<2%).
Identities of these compounds were confirmed by mass spectrometry.
RESULTS AND DISCUSSION
Toxicity
Technical-grade heptachlor, analytical-grade heptachlor, and heptachlor
epoxide were acutely toxic to the estuarine organisms tested (Tables 2-5).
Shell deposition of oysters exposed for 96 hr was greatly retarded at
measured concentrations of heptachlor >4.0 jug/liter. Wide differences in
heptachlor measured in lower concentrations of this test, compared with
those measured in identical desired concentrations of shrimp and fish
bioassays, may have been due to the filter-feeding of the oyster. Organo-
chlorines are readily adsorbed on plankton and other particulates,
especially at low levels in water; therefore removal of these particulates
could alter the concentrations detected in water. Oysters were sensitive to
two other organochlorine insecticides at similar levels. Shell deposition was
appreciably inhibited in chlordane concentrations >4.7 jug/liter (Parrish et
al., 1976) and at measured endrin concentrations >4.9 (Schimmel et al.,
1975).
Pink shrimp was the most sensitive of all species exposed to technical-
grade heptachlor; the 96-hr LCSO was 0.11 jug/liter. Pinfish and sheephead
minnows were the least sensitive, giving 96-hr LC50s of 3.77 and 3.68
/Kg/liter, respectively. Grass shrimp were sensitive to technical-grade
heptachlor, exhibiting an LC50 of 1.06 /zg/liter. This value is extremely
small when compared with Eisler's (1969) value of 440 Mg/liter for this
species. The greater sensitivity observed in our test could have been
351
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TABLE 2 Toxicity of Technical-Grade Heptachlor (65%) and Uptake by American Oysters
(Crassosirea virginica), Pink Shrimp (Penaeus duorarum), Grass Shrimp (Palaemonetes vulgaris),
Sheepshead Minnows (Cyprinodon variegatus), Spot (Leiostomus xanthurus), and Pinfish (Lagodon
rhomboides] Exposed for 96 hr
Water concentration (jig/liter)
Heptachlor
Species
C. I'irginica
P. duororuin
P. vulquris
C. vurk'galns
L. \unlhurus
/.. rtintllhuidt's
Desired*
Control
Control and
carrier
0.32
1.0
3.2
10.0
32.0
Control
Control and
carrier
0.1
0.32
1.0
3.2
10.0
Control
Control and
carrier
0.32
1.0
3.2
10.0
32.0
Control
Control and
carrier
6.5
8.7
11.5
15.5
21.0
Control
Control and
carrier
1.15
1.55
2.1
2.8
3.7
Control
Control and
carrier
0.32
1.0
3.2
10.0
32.0
Measured
NDf
NDr
0.083
0.40
0.91
4.0
14.0
NDf
NDC
0.04
0.20
0.43
2.0
5.0
NDC
NO'
0.13
0.44
2.0
5.0
12.2
ND(
NDf
2.7
3.3
3.6
4.0
8.8
ND^
ND'
0.5
0.65
0.77
1.25
1.4
ND'
NO'
0.20
0.44
1.10
4.4
1 1.
Measured
d
d
d
d
d
d
d
d
d
d
d
d
d
d
d
d
d
d
d
d
d
ND'
ND'
1.1
1.5
1.6
2.4
2.8
ND'
NDf
0.15
0.20
0.30
0.48
0.52
d
,/
c/
d
d
d
J
Effect
0
13
30
28
33
78
95
0
5
5
82
100
100
100
0
0
6
13
70
95
100
5
5
35
50
50
60
85
0
5
25
35
40
70
85
0
0
0
0
5
50
100
H
0.016
0.021
0.43
3.1
7.7
18.
55.
NDC
NDf
0.01
0.033
-
-
—
ND'
NDC
0.062
0.2
0.97
3.6
—
0.020
0.022
20.
33.
34.
85.
133.
NDC
0.01
1.5
2.3
7.6 -
17.3
9.8
ND'
ND'
0.55
1.8
5.7
34.
-
Whole-body .residue
(Mg/g, wet weight)
HE
0.01
ND
0.14
0.48
0.78
1.9
8.
ND'
NDC
0.054
0.18
-
-
-
0.014
0.012
0.26
0.55
2.5
6.1
_
0.016
0.020
6.7
9.2
9.9
18.
26.
0.011
0.016
0.58
0.72
2.5
4.0
2.1
0.015
0.013
0.39
1. 2
3.2
11.
-
trans-ct\
-
-
0.60
2.2
6.5
14.0
47.0
d
d
d
d
—
-
—
d
d
d
d
d
d
_
0.019
0.019
9.9
17.
18.
32.
47.
NDr
NDC
0.55
0.89
3.3
7.1
3.5
d
d
d
d
d
d
-
cis-ch
0.022
0.020
0.075
0.30
0.78
1.9
5.6
d
d
d
d
—
-
—
d
d
d
d
d
d
—
NDf
ND'
1.2
1.8
2.1
3.9
6.1
NDf
NDC
0.16
0.22
0.94
1.6
0.7
d
d
d
c'
d
d
-
"EHcct is expressed as percentage reduction in shell deposition for oysters and death for shrimp and fish. Whole-body
residues are for animals still alive at end of exposure. H, heptachlor; HE, heptachlor epoxidc; lrans-ch, /ron.v-chlord.inc;
i-A-ch, r/Vchlordanc.
'Calculated on 1 00% weight of technical heptachlor; heptachlor = 65% of the technical compound.
'ND. nondctcclable; 0.01 j/g/liler in walcr; 0.01 /jg/n in tissue.
Not analv/cd.
960
352
-------
HEPTACHLOR TOXICITY IN ESTUARINE ORGANISMS 961
TABLE 3. Toxicity of Analytical-Grade Heptachlor (99.8%) to Pink Shrimp (Penaeus
duorarum) and Spot (Leiostomus xanthurus) in 96-hr Flowthrough Bioassays
Heptachlor concentration
in water (jag/liter)
Species Desired
P. duorariim Control
Control and
carrier
0.0046
0.01
0.021
0.046
0.1
/.. xanihtirub Control
Control and
carrier
1.55
2.1
2.8
3.7
4.9
Measured
ND°
ND
ND
0.014
0.032
0.030
0.062
ND
ND
0.47
0.93
1.1
2.0
2.3
Mortality
(%)
0
0
IS
40
40
45
90
5
5
0
85
70
100
100
Whole-body residue
-------
962 S. C. SCHIMMEL ETAL.
TABLE 5. 96-hr LC50s for Several Estuarine Organisms Exposed to Technical-Grade Heptachlor,
Analytical-Grade Heptachlor, and Heptachlor Epoxide
Chemical
Species
Desired concentration.
LC50 (jug/liter)
(95% confidence limits)
Measured concentration
LCSO (Mg/liter)
(95% confidence limits)
Technical-
grade
heptachlor
(65%)
Analytical-
grade
heptachlor
(99.8%)
Heptachlor
epoxide
(99%)
Crassoitrea
virginica
Palaemonetes
vulgaris
Penaeus
duorarum
Cyprinodon
variegatus
Lagodon
rhomboides
Leiostomus
xanthurus
Penaeus
duorarum
Leiostomus
xanthurus
Penaeus
duorarum
4.0b
2.08
(1.39-3.02)
0.21
(0.16-0.28)
10.53
(7.39-13.71)
9.29
(6.98-12.59)
2.18
(1.86-2.58)
0.03
(0.02-0.05)
2.14
(1.74-3.00)
0.04
(0.03-0.13)
^.5b
1.06
(0.46-2.07)
0.11
(0.07-0.15)
3.68
(2.74-4.67)
3.77
(2.02-8.80)
0.85
(0.72-1.00)
0.03
(0.02-0.04)
0.86
(0.75-1.33)
0.04
(0.001-0.10)
"Measured concentration is for heptachlor content only.
°96-hr EC50; criterion is reduction of shell deposition.
aeration. Further, Earnest and Benville (1971) found LCsos in static
bioassays were greater than those in dynamic bioassays..
Juvenile spot were 4 times more sensitive to technical-grade heptachlor
than were juveniles of sheepshead minnows and pinfish; the 96-hr LCSO for
spot (based on measured concentrations) was 0.85 Mg/!iter (Table 5). Korn
and Earnest (1974) reported a lower sensitivity of striped bass juveniles (3
Mg/liter 96-hr LCSO at 13°C); however, their values were based on desired
concentrations. Katz (1961) in aerated, static bioassays found the
threespine stickleback relatively insensitive (112 Mg/liter 96-hr LCSO at
Z\j T^f*
Comparison of the relative toxicity of technical-grade heptachlor to
marine or estuarine fish and freshwater fish is difficult because of different
test conditions. Toxicity values may be underestimated because of in-
creased volatilization of heptachlor due to aeration. Henderson et al.
(1959) exposed Lepomis macrochirus to technical-grade heptachlor for
96-hr in static bioassays at 25°C; they reported a 19 ju/liter LCSO . Katz
(1961) tested three salmonids at 20°C for 96 hr and reported that chinook
salmon (Oncorhynchus tshawytscha) was the most sensitive (LC50 = 17.3
Mg/liter). Henderson implied that aeration was used in some of his
354
-------
HEPTACHLOR TOXICITY IN ESTUARINE ORGANISMS 963
bioassays and Katz aerated all exposure aquaria; therefore toxicity value in
these tests may be underestimated.
Pink shrimp was also the most sensitive species to analytical-grade
heptachlor (Tables 3 and 5). The 96-hr LC50 was 0.03 jug/liter for the
shrimp and 0.86 jug/liter for spot.
Heptachlor epoxide was also very toxic to pink shrimp; the 96-hr
LC50 was 0.04 /ig/liter (Tables 4 and 5).
Bioaccumulation
Heptachlor was concentrated in the tissues of all estuarine animals
exposed to the chemical (Tables 2-4). Fish accumulated heptachlor in
greater quantity than did crustaceans. The concentration factor (con-
centration of chemical in tissues divided by concentration measured in
water) for sheepshead minnows ranged from 7,400 to 21,300 while that
for two shrimp species was from 200 to 700 times (Table 6). This
difference may be due to the differences in structure and permeability
between fish and shrimp gill membranes. In numerous studies gills were
found to be a major organ of pesticide uptake. Also, fish have relatively
high amounts of fat in their tissues; organochlorine pesticides are highly
lipophilic.
Concentrations of heptachlor and its epoxide in fish collected in
estuarine waters were comparable with those in fish in our tests. In our
bioassays spot exposed to 0.5 Mg/liter heptachlor for 96 hr exhibited 25%
TABLE 6. Range of Concentration Factors"for Heptachlor, Heptachlor Epoxide, and fro/75-Chlordane
Concentration factor
Species Heptachlor Heptachlor epoxide fra/75-Chlordane
Heptachlor (technical)
Crassostrea virginica
Penaeus duorarum
Palaemonetes vulgar is
Cyprinodon variegatus
Leiostomus xanthurus
Lagodon rhomboides
Heptachlor (analytical)
Penaeus duorarum
Leiostomus xantnurus
Heptachlor epoxide
Penaeus duorarum
3,900-8,500
200-300
500-700
7,400-21,300
3,000-13,800
2,800-7,700
300-600
3,600-10,000
—
b
b
b
b
b
b
b
b
200-1,700
c
c
c
9,000-16
3,700-14
c
-
-
—
,800
,800
Concentration of chemical in tissue divided by measured concentration in water.
^Concentration factor for heptachlor epoxide could not be determined; although present in tissues,
it was not found in water.
c7r
-------
964 S. C. SCHIMMEL ET AL.
mortality; whole body residues of heptachlor were 1.5 Mg/g and of
heptachlor epoxide, 0.58 jug/g. Nearly identical concentrations of the two
chemicals have been reported in muscle of winter flounder in
Massachusetts waters (Smith and Cole, 1970). High residues such as these,
relative to toxicity data generated in our study, are disturbing.
Although fram-chlordane concentrations in water were not determined
in all tests, this chemical was concentrated in tissues of sheepshead
minnows and spot (Table 6) and concentration factors for ?/W7S-chlordane
derived from whole-body residues in these fish were comparable with
those for heptachlor. Since /ra/?s-chlordane constituted 22% of the tech-
nical heptachlor mixture and chlordane was readily accumulated in
tissues of estuarine animals in other tests (Parrish et al., 1976), we
anticipated high residues of this chemical in the animals we exposed.
Parrish et al. (1976) reported that chlordane was concentrated in tissues of
pink shrimp and grass shrimp 1,000-2,300 times that measured in water.
Unfortunately, tows-chlordane residues in shrimp from our tests were not
analyzed; thus a direct comparison cannot be made.
Heptachlor epoxide readily accumulated within the tissues of pink
shrimp. The highest concentration factor was nearly 3 times that of
heptachlor (Table 6).
CONCLUSION
The presence of heptachlor and heptachlor epoxide in estuarine
environments as well as their extreme toxicity to estuarine animals, in
particular the concentration in pink shrimp, is a cause for concern. We are
unaware of studies reporting residues of heptachlor or heptachlor epoxide
in tissues of marine or estuarine crustaceans. From our studies, however, it
appears possible that significant mortality can occur in pink shrimp
populations without the compound's being detected in water or in tissues.
Fifteen percent of the pink shrimp died in an experimental aquarium
when exposed to 0.0046 pig/liter heptachlor (desired concentration); no
detectable concentrations of the insecticide were measured in the water
(Table 3). Residues of heptachlor were not detected in surviving animals
from this same aquarium. Heptachlor epoxide, the metabolite of
heptachlor, was detected but the residue (0.023 jug/g) was only twice the
lower detectable limit of heptachlor epoxide (0.01 jug/g) in tissues.
The extreme toxicity of heptachlor to estuarine organisms shown in
our tests, coupled with its occurrence in aquatic organisms from estuarine
waters, represents a potentially dangerous situation. Possible subtle effects
of heptachlor on these organisms, such as reduced reproductive potential,
behavior modification, pathologic and physiologic changes, may occur
undetected. Long-term bioassays, involving the reproductive phase of the
life cycle, are required to determine these effects. Such studies are
essential to evaluate overall effects of toxic organic chemicals on aquatic
biota.
356
-------
HEPTACHLOR TOXICITY IN ESTUARINE ORGANISMS 965
REFERENCES
Andrilenas, P. A. 1974. Farmers' use of pesticides in 1971 ...quantities. Econ. Res. Serv. Agric.
Econ. Rep. 252. U.S. Food Drug Admin. 56 pp.
APHA. 1971. Standard methods for the examination of water and wastewater, 13th ed. New York:
American Public Health Association. 874 pp.
Barthel, W. F., Hawthorne, J. H., Bolton, G. C., McDowell, L. L., Gressinger, E. H. and Parsons,
D. A. 1969. Pesticide residues in sediments of the lower Mississippi River and its tributaries.
Pestic. Monit. J. 3:8-66.
Burchfield, H. P., Johnson, D. E. and Storrs, E. E. 1965. Guide to the analysis of pesticide
residues, 1 vol. Washington D.C.: U.S. Department Health, Education, and Welfare, Public
Health Service.
Casper, V. L., Hammerstrom, R. J., Robertson, E. A., Jr., Bugg, J. C. and Gaines, J. L. 1969.
Study of chlorinated pesticides in oysters and estuarine environment of the Mobile Bay area.
Unpublished report. Gulf Coast Mar. Health Sci. Lab., Ala. Water Improvement Comm. Ala.
State Dep. Public Health and Ala. Dep. Conserv.
Davidow, B. and Radomski, J. L. 1953. Isolation of an epoxide metabolite from fat tissues of dogs
fed heptachlor. J. Pharmacol. Exp. Ther. 107:259-265.
Earnest, R. D. and Benville, P. E. 1971. Correlation of DDT and lipid levels for certain San
Francisco Bay fish. Pestic. Monit. J. 5:235-241.
Eisler, R. 1969. Acute toxicities of insecticides to marine decapod crustaceans. Crustaceana
(Leiden) 16:302-310.
Eisler, R. 1970. Acute toxicities of organochlorine and organophosphorus insecticides to estuarine
fishes. U.S. Dep. Interior Fish. Wild. Ser. Bur. Sport Fish. Wild. Tech. Pap. No. 46, 11 pp.
Finney, D. J. 1971. Probit analysis, 3rd ed. Cambridge: Cambridge Univ. Press.
Hannon, M. R., Greichus, Y. A., Applegate, R. L. and Fox, A. C. 1970. Ecological distribution of
pesticides in Lake Poinsett, South Dakota. Trans. Am. Fish. Soc. 99:496-500.
Henderson, C., Pickering, Q. H. and Tarzwell, C. M. 1959. Relative toxicity of ten chlorinated
hydrocarbon insecticides to four species of fish. Trans. Am. Fish. Soc. 88:23-32.
Katz, M. 1961. Acute toxicity of some organic insecticides to three species of salmonids and to the
threespine stickback. Trans. Am. Fish. Soc. 90:264-268.
Korn, S. and Earnest, R. 1974. Acute toxicity of twenty insecticides to striped bass, Morone
saxatilis. Calif. Fish. Game 60:128-131.
Lowe, J. I., Parrish, P. R., Patrick, J. M., Jr. and Forester, J. 1972. Effects of the polychlorinated
biphenyl AroclorR 1254 on the American oyster, Crassostrea virginica. Mar. Bio/. (Berlin)
17:209-214.
Parrish, P. R., Schimmel, S. C., Hansen, D. J., Patrick, J. M., Jr. and Forester, J. 1976. Chlordane:
Effects on several estuarine organisms. /. Toxicol. Environ. Health 1:485-494.
Schimmel, S. C., Parrish, P. R., Hansen, D. )., Patrick, J. M., Jr. and Forester J. 1975. Endrin:
Effects on several estuarine organisms. Proc. 28th Annu. :Conf. SE Assoc. Game Fish.
Commun. (In press).
Smith, R. M. and Cole, C. F. 1970. Chlorinated hydrocarbon insecticide residues in winter
flounder, Pseudopleuronectes americanus, from Weweantic River Estuary, Massachusetts. /.
Fish. Res. Board. Can. 27:2374-2380.
Received October 10, 1975
Accepted November JO, 1975
357
-------
Reprinted from Transactions of
the American Fisheries Society
Vol. 105(6): 700-703, 1976,
with permission of the American
Fisheries Society
THE EFFECT OF MIREX ON THE BURROWING ACTIVITY OF THE
LUGWORM, ARENICOLA CRISTATA
W.P. Schoor and S.M. Newman
Contribution No. 268
359
-------
The Effect of Mirex on the Burrowing Activity
of the Lugworm (Arenicola cristata)1
W. P. SCHOOR AND S. M. NEWMAN2
U.S. Environmental Protection Agency, Environmental Research Laboratory
Gulf Breeze, Florida 32561
ABSTRACT
An inexpensive bioassay system was developed to estimate pollutant effects on a benthic animal.
Mirex, a fire ant toxicant, was taken into the substrate by the burrowing and feeding activity of
the lugworm, Arenicola cristata, and significantly affected this activity. Mirex was present in the
adult worm as well as in its juvenile stage.
The binding of organic compounds to the
aquatic substratum is generally realized
(Oloffs et al. 1972). Flow from there to the
water column above is subject to equilibrium
considerations due to leaching from the sub-
strate, which in turn could be altered consider-
ably by biological activity (Rhoads and Young
1971). With regard to the distribution of a
pollutant in an aqueous environment, the
following situations may be encountered: (1)
substratum, water, and biota are in quasi-
equilibrium (steady state) ; (2) substratum
(to include associated biota) acts as the only
source; and (3) water (to include associated
biota) acts as the only source. We have taken
the latter case in an attempt to demonstrate
whether or not: (1) the substratum can act
as a sink for the chlorinated hydrocarbon
mirex (a fire ant toxicant) ; (2) changes can
occur in the feeding and burrowing behavior
of the lugworm, Arenicola cristata, in the
presence of mirex; and (3) mirex is taken up
by the lugworm and can be channeled back
into the epibenthic system.
As an example of a macrobenthic species,
we chose the lugworm, Arenicola cristata, for
our studies because of its relative abundance
in the shallow estuaries of northwest Florida.
Natural density in salt marshes ranges between
6 and 14 adults per square meter. In addition,
the following characteristics (D'Asaro 1975;
Rubinstein 1975) make it a suitable species
for our studies: Arenicola cristata adjusts
well to laboratory conditions, reproduces
1Gulf Breeze Environmental Research Laboratory
Contribution No. 268.
2 Present address: Exxon Corporation, 100 South
5th Street, Kingsville, Texas 78363.
readily, and is easily maintained if fed a
compost of ground turtle grass, Thalassia tes-
tudinum, or algae, Ectocarpus sp.
METHODS
Two 180-liter covered glass aquaria (120 cm
X 30 cm X 50 cm) were filled to a height of
30 cm with alternating layers of sand (25%
coarse particles: #35 standard sieve; 75%
medium particles: #120 standard sieve) and
organic silt from a pristine Spartina alterni-
flora marsh (Fig. 1). A photoperiod of 10 h
light and 14 h dark was established using four
40-watt fluorescent tubes mounted 30 cm above
each aquarium. Filtered seawater (26%«) was
added to the aquaria and maintained at 20 ±
2 C by controlling the temperature in the
room. An alga, Ectocarpus sp., was chopped
in a blender and introduced into the aquaria,
where it formed a growing mat on the sub-
stratum. Four adult worms were then intro-
duced into each aquarium and allowed to
adjust for 3 days, after which one aquarium
was exposed to mirex3 by means of a modified
air-lift column (Tagatz 1976). The column
was constructed to contain mirex bait granules
equivalent to five times the field rate applica-
tion (1.40 kg/hectare of 0.3% mirex, 15%
soybean oil adsorbed on 84.7% corncob grit)
on the basis of surface area. Air introduced
in the bottom of the column swept water past
the compartment containing the mirex flushing
the leached mirex into the aquarium. One-
liter water samples were taken 5 days per week
and analyzed for mirex by gas chromatog-
3Dodecachlorooctahydro-l,3,4-metheno-2H-cyclo-
buta [C, D] pentalene.
700
361
-------
SCHOOR AND NEWMAN-MIREX EFFECTS ON BENTHOS
701
TABLE 1.—Mirex concentrations in water.
a
FICUHE 1.—Undisturbed habitat before addition of
raphy. The quantitation limit for mirex was
set at 20-mm peak height for the highest
obtainable resolution which yielded 0.003 fig/
liter for a one-liter water sample. Each water
sample removed was replaced by an equal
volume of filtered 26%' seawater.
In order to facilitate observations of surface
activity, vertical lines, 10 cm apart, ending at
the surface of the substratum, were drawn on
the outside of the aquaria, dividing each into
12 equal transects perpendicular to the longi-
tudinal axis. Daily observations (8a.m.) were
made along each transect and the following
surface features were graphically recorded and
counted: (1) active head shafts (feeding fun-
nels) ; (2) inactive head shafts; and (3) tail
shafts (respiratory holes). The exposure to
mirex was discontinued at day 30, the columns
removed, and the aquaria carefully flushed
twice and then refilled with filtered seawater.
The detrital material and other debris were
siphoned off and the algae removed as much
as physically possible.
RESULTS AND DISCUSSION
Table 1 shows the concentration of mirex
in the water of the exposed aquarium. The
variation seen is most likely due to adsorption
of mirex to particulate matter, possibly the
alga. Adsorption plays a major role in the
disposition of some compounds in the water
column (Schoor 1975) and ultracentrifugal
treatment of the water samples and analysis
of the water phase showed as much as 80%
of the mirex adsorbed to particulate matter.
Table 2 shows the total of the surface
features apearing per day during exposure to
mirex. In the absence of precise knowledge
Days elapsed
2
5
7
8
9
12
13
14
15
16
19
20
21
22
23
26
28
44
Mirex concentration
Ag/Uter
0.062
0.039
0.024
0.026
0.025
0.013
0.011
0.032
0.022
0.020
0.022
0.024
0.036
0.048
0.025
0.013
0.016
< 0.003
of the relative importance of each feature and
to quantitate behavior, we assigned to each
feature a value of one. Statistically significant
changes in the feeding behavior occurred (t
test; a = 0.05) between exposed and non-
exposed tanks. In the exposed tank three egg
masses appeared on day 4 and two on day 23,
while in the control tank only one egg mass
occurred on day 22. The exposure to mirex
was terminated at day 30, and both aquaria
observed non-quantitatively for another 45
days. On day 37 more alga was introduced
and the activity of the lugworms began to
increase in both aquaria, the activity in the
mirex-exposed aquarium being considerably
less. A water sample taken at day 44 showed
mirex present but below our 0.003 /ig/liter
quantitation limit. Similar levels were found
by Spence and Markin (1974) in ponds after
mirex treatment. On day 55, alga was again
added to both aquaria. Feeding behavior
appeared normal in the control aquarium, but
was still considerably less in the exposed
aquarium. Movement of small lugworms (1-
2.5 cm total length) was seen in the control
aquarium starting on day 57, but swimming
juveniles were not observed. On day 62, free-
swimming juveniles (approximately 1 cm)
were seen in the exposed tank and larger
juveniles (1-2.5 cm) were observed burrow-
ing in the upper layers of both aquaria. In
the exposed aquarium swimming declined on
day 63; and the last twelve swimming juvenile
worms were caught and analyzed for mirex.
Their whole-body residue was 60 fig/kg of
live weight. On day 75 there was reduced
362
-------
702
TRANS. AM. FISH. SOC, 1976, NO. 6
TABLE 2.—Appearance of surface features. See text for details.
Exposed
Length of
exposure
(days)
Worms added
2
Mirex added
4
5
6
7
8
9
10
11
12
1 I
1 I
15
16
17
18
19
20
21
22
2:;
21
2r,
26
2:
28
211
30
Head
shafts
Active Inactive
2
4
3
3
7
1
7
.1
2
I
(i
S
3
1
1
1
1
3
2
1
o
7
2
2
2
2
3
1
!
Tail
shafts
5
2
2
2
1
2
1
Daily Dailv
total Total avg.'
5
6 25 6.3
4
10
16a
1 35 5.0
9
7
2
4»
9 22 3.7
2
7
5a
3 12 1.7
1
1
2
2« 3 0.6
1
0
Head
shafts
Active Inactive
1
4
•f
2
III
|
2
3
.".
(i
3
S
1
5
2
1
2
3
3
2
2
3
12
1
3
I
1
4
2
•1
2
t
1
1
1
2
2
1
1
3
Control
Tail
shafts
1 1
1
!
1
1
I
1
1
2
1
Dafly
total Total
4
16 30
5
5
25»
3 45
7
5
5
11"
5 30
9
5
9«
4 29
6
4
(i
7« 18
5
6
Daily
avg.
7.5
6.4
5.0
4.1
3.6
1 Represents total counted over period of 3 days.
activity in both aquaria, that of the exposed
aquarium being much lower. The experiment
was terminated on day 75 and the original
adult lugworms were recovered. The whole-
body residue of mirex of the adults was 500
p,g/kg of dry weight. Sand samples taken
from the top of the substratum contained 0.54
fjig mirex/kg of dried sand; from the middle,
0.34 /ug/kg; and from the bottom, 0.08 /ng/kg.
Figure 2, taken on day 75, shows considerable
FIGURE 2.—Habitat at day 75. Note destruction and
mixing of layers caused by the burrowing activity
of Arenicola, and air-lift column containing mirex
bait.
destruction in the layered substratum, espe-
cially in the upper half and indicates mixing
occurred due to the burrowing activity of the
worms.
Mirex residues in the same substratum, in
the absence of worms, were approximately 1.5
/tig/kg of dried sand in the top 2 cm after a
17-day exposure to similar amounts of mirex
in the water (0.025 jug/liter average). None
was detected in the middle and the bottom of
the substratum.
Sand and lugworm samples for mirex resi-
due analyses were not taken at the end of the
exposure to mirex (30 days) because of the
ensuing destruction of the habitat. It was,
therefore, not established how much depu-
ration of mirex occurred during the 45 days
following the exposure. In reality, the expo-
sure to mirex continued past day 30 because
of its presence in the substratum. The amount
of mirex in the water at day 44 could not be
quantitated precisely because of our 0.003 /ig/
liter quantitation limit, but represents evidence
that a quasi-equilibrium was established with
mirex reappearing in the water column.
In summary, feeding and burrowing activity
363
-------
SCHOOR AND NEWMAN—MIREX EFFECTS ON BENTHOS
703
of Arenicola can affect distribution of mirex
in the substrate, and low concentrations of
mirex in water decreased bebavior activity as
measured by surface activity. Aside from
predation on adult worms, swimming juvenile
lugworms could transmit mirex to predators,
representing an example of biological feed-
back.
LITERATURE CITED
D'ASABO, C. N. 1975. A preliminary plan for a
commercial bait-worm hatchery to produce the
lugworm Arenicola cristata. Florida Sea Grant
Marine Advisory Program. (In press.)
OLOFFS, P. C., L. J. ALLRIFHT, AND S. Y. SZETO. 1972.
Fate and behavior of five chlorinated hydro-
carbons in two natural waters. Can. J. Microbiol.
18:1393-1398.
RHOADS, D. C., AND D. K. YOUNG. 1971. Animal-
sediment relations in Cape Cod Bay, Massachu-
setts. 11. Reworking by Molpadia Oolictica
(Holothuroidea). Mar. Biol. 11:255-261.
RUBINSTEIN, N. I. 1975. Thermal and haline optima
and lethal temperature limits for the culture of
Arenicola cristata (Polychaeta—Arenicolidae).
Master's Thesis. University of West Florida.
SCHOOR, W. P. 1975. Problems associated with low
solubility compounds in aquatic toxicity tests:
theoretical model and solubility characteristics
of Aroclor® 1254 in water. Water Res. 9: 937-
944.
SPENCE, J. H., AND G. P. MARKIN. 1974. Mirex
residues in the physical environment following a
single bait application. Pestic. Monit. T. 8 (2):
135-139.
TAGATZ, M. E. 1976. Effect of mirex on predator-
prey interaction in an experimental estuarine
ecosystem. Trans. Am. Fish. Soc. 105(4) : 546-
549.
364
-------
Reprinted from Estuarine Pro-
cesses, Vol. 1, Uses,
Stresses, and Adaptation to
the Estuary, Martin W. Wiley,
editor, pp. 523-531, 1976,
with permission of the Aca-
demic Press Inc. New York,
San Francisco, London
METALS, PESTICIDES, AND PCB's: TOXICITIES TO SHRIMP SINGLY
AND IN COMBINATION
Del Wayne R. Nimmo and Lowell H. Bahner
Contribution No. 271
365
-------
Reprinted from:
ESTUARINE PROCESSES, Vol. I
Uses, Stresses, and Adaptation to the Estuary
® 1976'
ACADEMIC CttSS. »NC
N*w York San Francisco londoa
METALS, PESTICIDES AND PCBs: TOXICITIES TO SHRIMP SINGLY
AND IN COMBINATION1
Del Wayne R. Nimmo and Lowell H. Banner
U.S. Environmental Protection Agency
Environmental Research Laboratory
Sabine Island, Gulf Breeze, Florida 32561
ABSTRACT: The objective of this study was to assess potential deleterious
effects of certain toxicants, singly and in combination, to penaeid shrimp. In
nature, these shrimp are exposed to combinations of toxicants from industrial
and municipal outfalls, from agricultural runoff or from dredge-and-fill opera-
tions.
The combined toxicities of methoxychlor and cadmium to penaeid shrimp,
Penaeus duorarum, were either independent or additive, and varied with the
method(s) of bioassay. Conclusions were based on the results of 10-, 25- and
30-day bioassays conducted with the toxicants added singly or in combination
to flowing water of constant salinity and temperature.
Cadmium, but not methoxychlor, was accumulated by shrimp and methoxy-
chlor appears to influence the processes of accumulation or loss of cadmium
from tissues of shrimp.
INTRODUCTION
Water quality criteria or effluent guidelines for heavy metals usually do not
take into account other toxicants which might exist in the effluent or receiving
waters. However, it is well accepted that aquatic species are subjected to com-
binations of toxicants rather than to single toxicants in the environment. An
example of this situation is the Southern California Bight, in which municipal
wastewaters, storm runoff and aerial fallout are responsible for the occurrence of
mercury, copper, DDT and other chlorinated hydrocarbons, such as PCBs in
estuarine and marine waters (12).
'Contribution No. 271, Gulf Breeze Environmental Research Laboratory
523
367
-------
524 D. R. NIMMO AND L. H. BAHNER
Synergistic effects of toxic agents have been demonstrated in fresh water tests
by Sprague (10), Sprague and Ramsay (11), and Cairns and Scheirer (3). Re-
cently, Roales and Perlmutter (7, 8), found that combinations of methylmercury
with copper or cygon (pesticide) and zinc appeared to have antagonistic (less
than additive) effects to freshwater fish or fish embryos. In a saltwater system,
Bahner and Nimmo (2) found that in short-term flow-through tests (48-hr and
96-hr), the combination of malathion-cadmium or of methoxychlor-cadmium
appeared to be independent and additive. In this report, results of 10-, 25- and
30-day tests of the toxicity of Cd, methoxychlor and Aroclor® 1254 (a PCB),
given singly and in combination to penaeid shrimp are discussed.
METHODS AND MATERIALS
All bioassays were conducted in flowing seawater at constant salinity and
temperature (1). Sixty-five liters/hr of filtered water (25±2° C and 20±2°/00
salinity) were delivered to each 30 - 1 glass aquarium. Single toxicants were
added to the water with metering pumps and combinations were obtained by
simultaneously metering individual toxicants from separate syringes or flasks
into the aquaria. Initially, shrimp were tested in a range of individual toxicants
to determine the LC50, a mathematical expression referring to a calculated
concentration of toxicant in which 50% of the experimental animals died within
a prescribed time interval. Thereafter, combinations tested were conducted at or
near the LCSO's of the individual toxicants. These LCSO's were calculated by
probit analysis (5).
The procedures for collecting and acclimation of test animals are described in
Bahner et al. and Nimmcvet al. (1, 6).
Methoxychlor and Aroclor 1254 were analyzed by electron-capture gas chro-
matography, as outlined by Nimmo et al. (6). Cadmium in water and tissues of
shrimp was analyzed by flameless atomic absorption spectroscopy, using the
methods of Segar (9).
RESULTS AND DISCUSSION
Bioassays with single toxicants. Initially, bioassays were conducted with each
toxicant singly, using different intervals of time, to determine LCSO's. Earlier
Bahner and Nimmo (2) reported the 96-hr acute toxicity of methoxychlor to be
about 1000X more toxic to shrimp than was cadmium (Table 1). In 30-day tests,
the toxicities were similar but lower (Table 1, Figs. 1 and 2). The 15-day LC50
for Aroclor 1254 was calculated from data reported by Nimmo et al. (6), and
shown in Table 1.
©Registered trademark, Monsanto Co., St. Louis, Mo. Mention of commercial products
does not constitute endorsement by the U.S. Environmental Protection Agency.
368
-------
METALS, PESTICIDES AND PCBs
525
Bioassays with two toxicants. The 25-day combination test, conducted near
the 30-day LC50 of each, indicated that the toxicities of methoxychlor and
cadmium were additive (Fig. 3) and these results are consistent with those found
in the 96-hr combination (2).
Table 1. Summary of bioassays of single toxicants to pink shrimp, Penaeus duorarum.
Toxicant
Cadmium
Cadmium
Methoxychlor
Methoxychlor
Aroclor 1254s
LC50
(Measured Concentrations)
4.6 mg/£
0.718 mg/8
3.5 «g/£
1.3Mg/«
1.0 «/«
Length of Exposure
96-hr1
30-day s
96-hr1
3 0-day s
1 5-day s
1 Data from Banner and Nimmo (2)
a Data from Nimmo et al., (6)
METHOXYCHLOR - 30DAY
IC50 1.3M8/I
K) 15 20
TIME (dpys)
Figure 1. Acute toxicity (LC50) of the insecticide, methoxychlor, to pink shrimp, Penaeus
duorarum during 30 day exposures.
569
-------
526
D. R. NIMMO AND L. H. BAHNER
Because toxicant concentrations vary in the environment, additivity at the
LC50 concentration may not represent environmental effects in the field. There-
fore, we attempted to distinguish the type of interaction between methoxychlor
and cadmium using differing combinations of concentrations in 10-day tests
(Fig. 4). Thus, the combinations of toxicants were prepared so as to have equal
toxicity, but each was tested in increasing concentrations and counter to the
other. The pairs of concentrations were arranged so that as the concentration of
one toxicant increased, that of the other decreased. The combinations of con-
centrations were selected so that if the toxicities of methjoxychlor and cadmium
were additive, the sums of the toxicities of each pair of concentrations would be
equal. Fig. 4 suggests that the two toxicants exert their effect independently; or,
as Warren (12) has suggested there is "no interaction" (Fig. 5). Lack of inter-
action in this test does not invalidate the conclusion drawn from the previous
CAOMIUM-3ODAY
LC50 718MQ/I
10 20
TIME (days)
Figure 2. Acuic toxicity (LC50) of cadmium to shrimp, Penaeus duorarum in 30 days.
3-70
-------
METALS, PESTICIDES AND PCBs
527
test, which showed additivity, rather the difference is probably due to the rela-
tive concentration of each toxicant. Also, other researchers have shown the lack
of interaction between two chemicals. The dosage-mortality curve, shown in Fig.
4 is similar to that given for zebrafish embtyos that were exposed to combina-
tions of Varying percentages of the 72-hr Tlm concentrations for Cygon and zinc,
and to that shown for the blue gouramis that were exposed for 96 hrs to
combinations of capper and methylmercury (7, 8).
Bioassays with three toxicants. Polychlorinated biphenyls (PCBs) are indus-
trial chemicals found in estuaries that possess properties similar to chlorinated
20
3
CONTROL
METHOXYCHLOR
0.85 iig/l
ADDITIVE
EFFECT
20
21 22
TIME (days)
23
24
25
Figure 3. Toxicity of Cadmium and methoxychlor administered singly and in combination
to PenaeuS duorarum. The predicted additive effect shown refers to the sum of
the numbers of living shrimp in aquaria containing individual toxicants. (CD +
Methox = 750 Mg Cd/X + 0.80 Mg Methoxychlor /£).
3,71
-------
528
D. R. NIMMO AND L. H. BAHNER
hydrocarbon pesticides (4). To determine the effect of one PCB (Aroclor 1254)
in combination with cadmium and methoxychlor, we conducted bioassays using
all possible single, 2-, and 3-way combinations of the toxicants. Cadmium and
methoxychlor were administered at the 30-day LC50; Aroclor 1254, at the
15-day LC50. Again, the results showed that the toxicity of any combination
was equal to the sum of each chemical tested singly (Fig. 6).
Accumulation of cadmium. After the exposures, cadmium, but not methoxy-
chlor was detected in shrimp tissues. In the 25-day tests, with methoxychlor and
cadmium singly, and in combination, there was no difference in cadmium
accumulation in the muscle. There was less cadmium in the muscle of shrimp
that had been exposed to methoxychlor alone than in controls (Table 2). Like-
wise, when cadmium was administered in combination with methoxychlor in the
second test (Table 3), a statistically lower cadmium concentration was found
when methoxychlor was present.
Pathology. Shrimp exposed to cadmium alone developed pronounced
blackened areas on the gills, but this condition was reduced or absent when
methoxychlor and cadmium were tested in combination.
100
>95
X
80
o
in
S
20
*
/
70
55
o
TOXICANT COUPUT
'CADMIUM
)400A
10V4A
'MITMOX.
*M«c»uf«d
0.29 A
AO
A 1.9
Figure 4. Do sage-mortality curves for toxicant combinations after 10 days. The toxicants,
cadmium and methoxychlor were prepared and administered so as to have equal
toxicity when tested separately. Along the abscissa we show each toxicant as a
small triangle, and the combination in each aquarium, as a toxicant couplet.
372
-------
METALS, PESTICIDES AND PCBs
529
Dover sole, (Microstomuspacificus) collected near the outfall of Palos Verdes
Peninsula, California (an area in which the biota contained high DDT concentra-
tions) did not contain as much arsenic, cadmium and selenium as specimens
taken near Santa Catalina Island (13). Concentrations of these elements in sedi-
ments from the Palos Verdes outfall were greater by a factor of 15 for arsenic,
Table 2. Accumulation of cadmium in the pink shrimp when tested singly and in combina-
tion with methoxychlor.
Test Concentrations
(Mg/D
Nominal Measured
Control -
Cadmium
1000 860
Methoxychlor
1.0 0.85
Cadmium/Methoxychlor
1000 730
1.0 0.80
Percentage Mortality
In 25 Days1
0
25
15
55
Cadmium in muscle
mg/kg ± 2 S.E.M.
0.4 ± 0.2
89.4 ± 17.4
0.1 ± 0.0
94.1 ± 26.1
1 Bioassay conducted at 20 v/oo salinity and 25° C.
Tolerance of Lethal Conditions
K
N
3
O
Supro-additivt interaction
o
Ptrcfntoff of solution A : IOO
Percentage of solution B: O
Strictly additive interaction
Solutibn combinations
Figure 5. Possible kinds of interactions between two hypothetical toxicants tested in com-
bination. After Warren (12).
.373
-------
530
D. R. NIMMO AND L. H. BAHNER
160 for cadmium and 14 for selenium than those from Catalina Island. The
apparent discrepancy in concentrations of metals in the tissues of biota may be
due to the presence of organochlorine pesticides similar to DDT which influ-
enced heavy metal accumulation.
CONCLUSIONS
No dramatic toxic interaction of the combination of methoxychlor and cad-
mium, or of the combination of methoxychlor-cadmium-PCB, to shrimp was
70
60
o
9
2 2
30
20
10
• PCB (AROCLOR 1254)
OCADMIUM
-|- METHOXYCHLOR
• CONTROL
10 20 30 40 50
PERCENTAGE KILLED IN 10 DAYS
(EXPECTED)
60
Figure 6. Comparison of toxicities of single vs. combined constituents to Penaeus duo-
rarum. Measured concentrations in micrograms per liter ranged from 640-829 for
cadmium, 0.9-1.0 for methoxychlor, 0.7-1.1 for Aroclor 1254. Each symbol
represents the effect of a single toxicant, per aquarium; superimposed symbols
represent the toxicants combined. Unity is denoted as the line constructed
through the points equaling the observed percentage killed (ordinate) to that
expected (abscissa) for each toxicant singly, and the origin. If any 2- or 3-way
combination has elicited an exactly additive effect, the datum would fall on
unity. If synergistic effects had been observed with any combination, the datum
would lie to the left of unity; if antagonistic, to the right of unity. By construct-
ing the Working-Hotelling confidence bands (regression analyses) on these data,
they indicated that all toxicant combinations were probably additive (a = 0.05).
374
-------
METALS, PESTICIDES AND PCBs 531
evident. We have presented evidence that methoxychlor influences the accumula-
tion in, or loss of cadmium from, the tissues of shrimp. The toxicities of the
combinations, when compared to each toxicant tested singly, were independent
and additive. Distinguishing between these relationships depends upon the
method of assessment and the concentrations of each of the toxicants. We
observed additivity, but no synergistic effects of the toxicants in the laboratory,
although synergism might exist in the environment. Before allowable limits for
effluents are established, therefore, background concentrations of all toxicants
in receiving waters used in experimental systems must be known in order to
properly evaluate the toxicity of the compound under study.
Table 3. Accumulation of cadmium in tissues of shrimp in the presence of other chemicals.
Measured Concentrations
in test
Cadmium3
640
774
746
829
media
PCB3
0.9
0.9
(Mg/D1
Methoxy.4
1.1
0.8
Control
0
0.7
1.1
1.0
0
1.0
Cadmium in muscle
mg/kg ± 2 S.E.M.
15.58 ±2.90
9.90 ± 2.79
13.99 ±3.05
16.28 ± 5.44
0.25 ± 0.1
0.25 ± 0.05
0.28 ±0.1
0.26 ± 0.07
Concentration
Factor
24
13
19
20
0
0
0
0
110-day flowing water bioassays; 20° /oo, 25° C
2 Cadmium chloride
3Aroclor 1254
41,1,1 - Trichloro-2, 2-bis-[p-methoxyphenyl] ethane
LITERATURE CITED
1. Banner, L. H., C. D. Craft and D. R. Nimmo. 1975. A saltwater flow-
through bioassay method with controlled temperature and salinity. Prog.
Fish-Cult. 37:126-129.
2. . , and D. R. Nimmo. In Press. Methods to assess effects of combinations
of toxicants, salinity and temperature on estuarine animals. Presented at
the 9th Annual Conference on Trace Substances in Environmental Health,
10-12 June 1975, Columbia, Mo.
3. Cairns, J. Jr., and A. Scheier. 1968. A comparison of the toxicity of some
common industrial waste components tested individually and combined.
Prog. Fish-Cult. 30:3-8.
4. Duke, T. W., J. I. Lowe, and A. J. Wilson, Jr. 1970. A polychlorinated
biphenyl (Aroclor 1254®) in the water, sediment, and biota of Escambia
Bay, Florida. Bull. Environ. Contam. Toxicol. 5:171-180.
5. Finney, D. J. 1971. Probit analysis. Cambridge University Press. 333p.
375.
-------
532 D. R. NIMMO AND L. H. BAHNER
6. Nimmo, D. R., R. R. Blackman, A. J. Wilson, Jr. and J. Forester. 1971.
Toxicity and distribution of Aroclor® 1254 in pink shrimp Penaeus
duorarum. Mar. Biol. (Berlin) 11:191-197.
7. Roales, R. R., and A. Perlmutter. 1974. Toxicity of zinc and cygon, applied
singly and jointly, to zebrafish embryos. Bull Environ. Contain. Toxicol.
12:475-480.
8. , and 1974. Toxicity of methyl-mercury and copper, applied
singly and jointly, to the blue gourami, Trichogaster trichopterus. Bull.
Environ. Contam. Toxicol. 12:633-639.
9. Segar, D. A. 1971. The use of the heated graphite atomizer in marine
sciences. Proc. 3rd International Congress of Atomic Absorption and
Atomic Fluorescence Spectrometry. Adam Hilger, London, pp. 523-532.
10. Sprague, J. B. 1964. Lethal concentrations of copper and zinc for young
Atlantic salmon. J. Fish. Res. Board Can. 21:17-26.
11. , and B. A. Ramsay. 1965. Lethal levels of mixed copper-zinc solutions
for juvenile salmon. J. Fish. Res. Bd. Can. 22:425-432.
12. Warren, C. E. 1971. Biology and Water Pollution Control. W. B. Saunders
Co. Philadelphia and London. 434 p.
13. Young, D. R., D. J. McDermott, T. C. Heesen and Tsu-Kai Jan. In Press.
Presented at the American Chemical Society Symposium on Marine Chem-
istry in the Coastal Environment, 8-10 April 1975, Philadelphia, Pa.
376
-------
Reprinted from Journal of
Agriculture and Food Chemistry/
Vol. 24(3): 631-634, 1976,
with permission of the Amer-
ican chemical Society
DETERMINATION OF MALATHION, MALAOXON, AND MONO- AND DICARBOXYLIC
ACIDS OF MALATHION IN FISH, OYSTER, AND SHRIMP TISSUE
Gary H. Cook and James C. Moore
Contribution No. 273
377
-------
Determination of Malathion, Malaoxon, and Mono- and Dicarboxylic Acids of
Malathion in Fish, Oyster, and Shrimp Tissue
Gary H. Cook and James C. Moore*
A method is described for monitoring the presence of malathion and its metabolites in the aquatic
environment. Malathion, malaoxon, malathion monoacid, and malathion diacid were determined in
fish, oyster, and shrimp tissues by gas-liquid chromatography (GLC) using phenthoate and phenthoate
acid as internal standards. GLC analyses were performed without cleanup, using a flame photometric
detector operating in the phosphorus mode. Acid compounds were methylated with diazomethane.
Pinfish exposed to 75 jtg/1. of malathion in flowing seawater for 24 h contained no residues of malathion
or malaoxon, although the concentration of the malathion monoacid in the gut was 31.4 Mg/g- The data
illustrate that pinfish rapidly convert malathion to the mono- and dicarboxylic acids of malathion.
Malathion (0,0-dimethyl dithiophosphate of diethyl various substrates have been studied extensively (Krueger
mercaptosuccinate) has a broad spectrum of effectiveness and O'Brien, 1959; Corley and Beroza, 1968; Shafik and
against insects and is widely used along coastal areas for Enos, 1969; Shafik et al, 1971; El-Refai and Hopkins, 1972;
control of mosquitoes, flies, and other noxious pests. Wolf et al., 1975) practically no residue data have been
Although the chemistry and metabolism of malathion in reported for malathion or its degradation products in
aquatic species.
Binder (1969) studied the uptake of malathion in carp
U.S. Environmental Protection Agency, Environmental exposed to 5 mg/1. malathion for 4 days and found that
Research Laboratory, Sabine Island, Gulf Breeze, Florida residues in the flesh had an average half-life of 12 h, the
32561. liver concentrating the greatest amount. At our laboratory,
379 J. Agric. Food Chem., Vol. 24, No. 3, 1976 631
-------
COOK, MOORE
o
C-OC2H5
CHj
0
II
I
0 CHs
II I
^P-S-CH
CHjO |
0
MALAOXON
CHjO
0
C-OC2H5
CH2
C-OH
II
0
CMjO
0
II
C-OH
C-OH
II
0
MCA
DCA.
S
I
P-S-CH
0
II
C-OCjH5
CHjO
PHENTHOATE
CHjO
I I
P-S-CH
PHA
0
II
C-OH
Figure 1. Malathion, malaoxon, monocarboxylic acid of
malathion (MCA), dicarboxylic acid of malathion (DCA),
phenthoate (PHE), and the carboxylic acid of phenthoate
(PHA).
no residues of malathion have been found in the tissue of
fish at exposures to 300 ^g/L in the water (Coppage et al,
1975; Tagatz et al., 1974).
Because of the different properties of malathion and its
hydrolytic products, separate methods of analysis are
usually performed for their extraction and cleanup. ShafLk
et al (Shafik and Enos, 1969; Shafik et aL, 1971) developed
methods for monitoring human beings exposed to mala-
thion by analyzing the urine for malathion, the mono-
carboxylic acid of malathion (MCA), the dicarboxylic acid
of malathion (DCA), as well as the alkyl phosphate me-
tabolites. Kadoum (1969) described a method of analysis
for malathion and its hydrolytic products in stored grain.
In this investigation, a method was developed for
analysis of malathion, malaoxon, MCA, and DCA in fish,
oysters, and shrimp, using Phenthoate (PHE; 0,0-di-
methyl phosphorodithioate of ethyl mercaptophenyl-
acetate) and its acid degradation product (PHA; 0,0-
dimethyl phosphorodithioate of mercaptophenyl acetic
acid) as internal standards (Figure 1).
Each sample was spiked with PHE and PHA to permit
evaluation of the integrity of the analysis. Recoveries of
malathion were based on recoveries of PHE. MCA and
DCA recoveries were based on recoveries of PHA. All
residues were adjusted for recovery. The method is
suitable for (1) monitoring the presence of malathion in
the aquatic environment and species, (2) pointing out its
path of degradation, and (3) explaining the lack of reported
residues for malathion in fish.
EXPERIMENTAL SECTION
Apparatus. We employed a Tracor MT-220 gas-liquid
chromatograph equipped with a flame photometric de-
tector operating in the phosphorus mode and a 63Ni
electron-capture detector. The 180 cm X 3 mm i.d. glass
column was packed with 2% OV-101 on Gas-Chrom Q
100/120 mesh. Operating conditions for the flame pho-
tometric detector were: column, 175 °C; inlet, 225 °C;
detector, 165 °C; nitrogen carrier flow, 65 ml/min; hy-
drogen, 200 ml/min; air, 50 ml/min; oxygen, 15 ml/min.
Operating conditions for the electron capture detector
were: column, 175 °C; inlet, 225 °C; detector, 250 °C;
nitrogen carrier flow, 65 ml/min.
New columns were conditioned by placing a small piece
of glass wool in the inlet end of the column and adding
approximately 10 cm of 10% Carbowax (Chemical Re-
search Services, Inc.,) on Chromosorb W, acid-washed
80/100 mesh. The column was heated overnight at 230
°C with a nitrogen flow of 20 ml/min. (During the con-
ditioning period, the column was not connected to the
detector.) The Carbowax and the glass wool separator plug
were replaced with 2% OV-101, the carrier flow was ad-
justed to 50 ml/min, and the column was conditioned an
additional hour at 230 °C, when sensitivity and efficiency
of the column became unacceptable because of peak
tailing, lack of peak separations, or the lack of reprodu-
cibility. Replacement of the glass wool and approximately
5 cm of the OV-101 at the injector end usually returned
the column to its original efficiency.
Reagents. All solvents were Nanograde, distilled in
glass (Mallinckrodt Inc., or equivalent).
Standards. Malathion, malaoxon, and phenthoate were
obtained from the Pesticide Reference Standard Section,
Environmental Protection Agency, Washington, D.C.,
MCA and DCA were obtained from the American
Cyanamid Co., Princeton, N.J., and PHE was obtained
from Thomson-Hayward Chemical Co., Kansas City, Kan.
Primary standard solutions of malathion, malaoxon, and
PHE were prepared by diluting 100 mg to 100 ml with
benzene. DCA, MCA, and PHA primary standard solu-
tions were prepared by diluting 100 mg to 100 ml with
benzene-acetone (3:1). Working standards and spike
solutions were prepared by diluting the primary standard
solutions with petroleum ether to the desired concen-
trations. Primary standards were kept refrigerated at 3
°C in amber bottles closed with Teflon-lined screw caps.
Acidified Sodium Sulfate. Anhydrous sodium sulfate
(Baker Chemical Co.) was stirred into a smooth slurry with
9 N sulfuric acid and the excess acid removed by vacuum
filtration. The sodium sulfate was washed twice with
methyl alcohol, then with ethyl acetate, the mixture being
stirred each time into a smooth slurry before vacuum
filtration. The sodium sulfate was allowed to air-dry for
2 h and then heated at 130 °C overnight. All lumps were
removed by blending in a Waring blender.
Procedure. A 0.5 to 5.0 g (wet weight) sample of tissue
was placed in a 25 X 150 mm culture tube and fortified
with 20 Mg of PHE and PHA. Ten milliliters of aceto-
nitrile, acidified by adding 2% (v/v) 2 N HC1, was added
to the tube and the tissue extracted at 20 000 rpm for 30
s with a Willems Polytron. The culture tube was cen-
trifuged and the acetonitrile extract decanted into a 150-ml
beaker containing 100 ml of 2% aqueous sodium sulfate.
The tissue was extracted a second time with 5 ml of
acidified acetonitrile and centrifuged and the extract
decanted into the 150-ml beaker. The aqueous solution
was adjusted to pH 8.5 with 5% aqueous sodium carbonate
and transferred to a 250-ml separatory funnel,
Malathion and PHE were removed by extracting the
aqueous solution with two 25-ml portions of petroleum
ether. The petroleum ether extracts were dried by eluting
through a 15-g plug of anhydrous sodium sulfate, in a
30-mm powder funnel The sodium sulfate was rinsed with
15 ml of petroleum ether. The combined extracts (col-
lected in a 70 X 50 mm crystallizing dish) were placed on
a 50 °C slide warmer in a hood and evaporated to about
5 ml by pulling a gentle stream of air over the dish. The
extracts were then diluted to a standard volume of 10 ml
for determining the percentage recovery of PHE. The
sample volume was then further adjusted as required for
632 J. Agrtc. Food Chem., VoL 24, No. 3, 1976
380
-------
MALATHION IN AQUATIC ENVIRONMENT
Table I. Recoveries of Malathion, Phenthoate, Malaoxon,
MCA, DCA, and PHA from Fortified Fish, Oyster,
and Shrimp Tissue
Add-
Com- ed,
pound0 Mg
Percentage recovery (X ± SD)
Fish
Oyster
Shrimp
Malathion 20 93.0 ± 3.8 91.7 ± 3.2 91.4 ± 5.0
Phenthoate
(PHE)
Malaoxon
MCA
DCA
PHA
40
20
20
20
40
92.0
68.9
81.0
74.5
73.6
±
+
±
±
±
4.6
5.8
3.6
5.2
5.6
94.1 ±
71
86
91
81
.0±
.5 ±
.8±
.7±
3.5
11.8
9.4
9.6
13.4
91.
80.
88.
84.
84.
.6
.6
.2
.4
.8
+
±
±
+
±
4.8
3.8
3.7
4.3
3.8
0 Tissue samples were 1-5 g (percentage recovery based
on total micrograms recovered). Fish tissue are averages
of 10 samples; shrimp tissue, 5 samples; oyster tissue, 8
samples.
determination of malathion.
Malaoxon was removed by extracting the aqueous so-
lution twice with 25-ml portions of methylene chloride.
The extracts were dried by eluting through sodium sulfate
and then evaporated to about 1 ml on a slide warmer as
described for malathion. The volume was adjusted with
acetone as necessary for GLC analysis.
To extract MCA, DCA, and PHA, the aqueous solution
was transferred to a 150-ml beaker and 5% (w/v) solid
sodium chloride was added. The pH was adjusted to pH
2 with 6 N HC1. The solution was returned to the 250-ml
separatory funnel and extracted twice with 50-ml portions
of ethyl acetate. The ethyl acetate extracts were dried by
eluting through a plug of acidified sodium sulfate, collected
in a crystallizing dish, and esterified with diazomethane.
Esterification of the acidic compounds was carried out
in a diazomethane generating apparatus described by
Schlink and Gellerman (1960). Diazomethane was bubbled
into each sample until a slight yellow color of excess
diazomethane persisted. Excess diazomethane was re-
moved during concentration on the slide warmer. The
samples were diluted to 10 ml for the analysis of PHA and
then adjusted as required, for the analysis of MCA and
DCA. (Caution: Diazomethane is explosive, carcinogenic,
and extremely toxic.)
Samples to be analyzed by electron-capture gas chro-
matography were cleaned up by the procedure of Hansen
et al. (1974) before analysis. Malathion, PHE, and the
methyl esters of MCA, DCA, and PHA were eluted from
the Florisil column with 20 ml of 50% ethyl ether in
hexane. Malaoxon was eluted with 20 ml of acetone. The
eluate was concentrated to 3-5 ml under a gentle stream
of N2 in a Kontes concentrator apparatus at 45 °C and
diluted to the desired volume with petroleum ether.
Acetone was removed from the malaoxon fraction by
evaporation to 1-2 ml, adding 10 ml of hexane, and re-
evaporating before adjusting to the desired volume.
RESULTS AND DISCUSSION
The chromatograms in Figure 2 were from the extract
of a 2-g sample of fish tissue which had been fortified with
20 Mg of malathion, malaoxon, MCA, and DCA and 40 ng
of PHE and PHA and extracted by the procedure de-
scribed above. The extract was concentrated to 10 ml and
analyzed on the flame photometric detector without
cleanup Table I shows the average percentage recovery
of malathion, malaoxon, MCA, DCA, and the internal
standards, PHE and PHA, from fortified tissue samples.
The petroleum ether extract contained 10-15% of the
malaoxon present. This malaoxon was added to the
malaoxon recorded in the methylene chlonde extract to
obtain total malaoxon recovery.
| STANDARD |
Ph*nthoat«
Malathion
SAalaoxon
r
v/s.
^-
PETROLEUM
ETHER
Ph*nthoat«
Malathion
P"-1
^
^-
METHYLENE
CHLORIDE
Malaoxon
ETHYL
ACETATE
OCA
r
MCA
'*— (
PHA
Sn
3&-1
Figure 2. Flame photometric chromatograms of the ex-
tracts of fish tissues fortified at 10 Mg/g of tissue with mal-
athion, MCA, and DCA and 20 Mg/g of tissue with PHE
and PHA.
A single methylene chloride extraction which will remove
malathion, malaoxon, and PHE may be used instead of
separate petroleum ether and methylene chloride ex-
tractions. This has the advantage of eliminating one
injection; however, methylene chloride also removes ap-
proximately 80% of the acetonitrile from the aqueous
phase. Because of the higher boiling point of acetonitrile
and the volatility of malathion and PHE, evaporation was
slow and low recoveries of malathion and PHE were ob-
tained. This accounts for the lower recovery obtained for
malaoxon. If the analysis for malaoxon is eliminated, the
methylene chloride extraction should be made at pH 8 and
discarded; otherwise, the acetonitrile will be removed in
the ethyl acetate extract and will interfere with methyl-
ation of the acids.
The aqueous phase must be kept to a minimum due to
the solubility of the acid compounds in water. The dis-
tribution coefficients of the acids are made favorable for
extraction of the compounds with ethyl acetate by ad-
justing to pH 2 and adding enough NaCl to reach 5% w/v.
To check the applicability of the method to real samples,
5 to 7 cm long pinfish (Lagodon rhomboides) were exposed
for 24 h to 75 /*g/l. malathion in flowing seawater as
described by Coppage et al. (1975). The pinfish were then
rinsed with distilled water and whole-body analysis was
performed. Table II shows the concentrations of mala-
thion, and its degradation products and the percentage
recovery of the internal standards. Concentrations were
adjusted for percentage recovery of the internal standard.
The FPD chromatogram of each extract is shown in Figure
3. The internal standards peaks, phenthoate and PHA,
represent 20 ^g/g of tissue diluted to 10 ml for analysis.
For routine analysis, the sensitivity of the FPD is 0.1
ppm for a 1-g sample compared to 0.01 ppm for the
J. Agric. Food Chem., Vol. 24, No. 3, 1976 633
-------
COOK, MOORE
Table II. Whole-Body Residues of Malathion and Its Metabolites in Pinfish Exposed to 75
Flowing Seawater for 24 h
- Malathion in
Sample
no.
1
2
3
4
5
Phenthoate (PHE),
% recovery
88
87
93
93
100
Malathi-
on, Mg/g
ND°
ND
ND
ND
ND
Malaoxon,
Mg/g
ND°
ND
ND
ND
ND
MCA,
Mg/g
3.8
3.0
6.0
4.1
4.3
DCA,
Mg/g
0.28
0.22
0.43
0.25
0.34
PHA,
% recovery
76
78
81
73
79
X± SD
92 ± 5
ND
ND
" ND = not detected; <0.10 Mg/g.
Table III. Concentration of Malathion, Malaoxon, MCA
and DCA in Various Organs of Pinfish Exposed to 75
of Malathion in Flowing Seawater for 24 h
Organ
Brain
Liver
Gills
Flesh
Gut
Malathi-
on, Mg/g
ND°
ND
ND
ND
ND
Malaoxon,
Mg/g
ND°
ND
ND
ND
ND
MCA,
Mg/g
1.7
6.0
2.5
3.9
31.4
DCA,
Mg/g
0.22
0.25
0.36
0.34
0.7
1 ND = not detected; <0.10 Mg/g.
PETROLEUM
ETHER
Ph*nfhoot*
^J
^~
T
Figure 3. Flame photometric chromatograms from
whole-body residue extracts of pinfish exposed to 75 Mg/1-
malathion for 24 h in flowing sea water.
electron-capture detector. However, the flame photometric
detector offers advantages of specificity for phosphorus
and a larger linear range of detection (three or more orders
of magnitude). Also, the latter detector is not flooded with
solvents, such as ethyl acetate, acetone, and small amounts
of acetonitrile and methylene chlorine, and the extracts
may be analyzed without cleanup.
Knowledge of the location of the pesticide residues in
various tissues is important for understanding the route
4.2 ± 1
0.30 ± 0.08
77 ± 3
of detoxification and degradation, as well as for attaining
increased analytical sensitivity by analyzing the organ of
highest concentration. Table HI shows the concentrations
of malathion and its metabolites found in the various
organs of pinfish exposed to 75 ppb of malathion for 24
h. Pinfish very rapidly convert malathion to mono- and
diacids, whose greatest concentrations were found in the
gut. Malathion itself was not found in any organ.
ACKNOWLEDGMENT
We thank P. R. Parrish and Edward Matthews for
assistance in collecting and exposing the test animals used
in this study and Steven S. Foss for preparation of tables
and illustrations.
LITERATURE CITED
Binder, M. E., Prog. Fish Cult. 31, 155-159 (1969).
Coppage, D. L., Matthews, E., Cook, G. H., Knight, J., Pestic.
Biochem. Physiol. 5(6), 536-542 (1975).
Corley, C., Beroza, M., J. Agric. Food Chem. 16, 361 (1968).
El-Refai, Hopkins, T. L., J. Assoc. Off. Anal. Chem. 55(3), 526-531
(1972).
Hansen, D. L., Parrish, P. R., Forester, J., Environ. Res. 7,363-373
(1974).
Kadoum, A. M., J. Agric. Food Chem. 17(6), 1178-1180 (1969).
Krueger, H. R., O'Brien, R. D., J:Econ. Entomol. 52,1063-1067
(1959).
Schlink, H., Gellerman, J. L., Anal. Chem. 32(11), 1412-1414
(1960).
Shafik, M. T., Bradway, D., Enos, H. F., J. Agric. Food Chem.
19(5), 885-889 (1971).
Shafik, M. T., Enos, H. F., J. Agric. Food Chem. 17,1186 (1969).
Tagatz, M. E., Borthwick, P. W., Cook, G. H., Coppage, D. L.,
Mosq. News 34(3), 390-315 (1974).
Wolf, N. L., Zepp, R. G., Baughman, G. L., Gordon, J. A., Bull.
Environ. Contain. Toxicol. 13(6), 707-713 (1975).
Received for review November 3, 1975. Accepted February 17,
1976. Contribution No. 273 from the Gulf Breeze Environmental
Research Laboratory. Mention of commercial products does not
constitute endorsement by the U.S. Environmental Protection
Agency.
634 J Agric Food Chem.. Vol. 24. No. 3. 1976
382
-------
Reprinted from Bulletin of
Environmental Contamination
and Toxicology, Vol. 16(3):
283-290, 1976, with permis-
sion of Springer-Verlag New
York Inc.
THE RELATIONSHIP OF MALATHION AND ITS METABOLITES
TO FISH POISONING
Gary H. Cook, James C. Moore, and David L. Coppage
Contribution No. 275
383
-------
The Relationship of Malathion
and Its Metabolities to Fish Poisoning1
Gary H. Cook, James C. Moore, and David L. Coppage
U.S. Environmental Protection Agency
Environmental Research Laboratory
Sabine Island, Gulf Breeze, Fla. 32561
INTRODUCTION
Malathion is a widely used organophosphate insecticide with
an annual production estimated to be in excess of 1 X 10^ kg in
the United States (ENVIRONMENTAL PROTECTION AGENCY 1972a). It
may enter surface water through surface runoff (ENVIRONMENTAL
PROTECTION AGENCY 1972b,c,d) or through direct spray for mosquito
control (GUERRANT et al. 1970, COPPAGE and DUKE 1971, PINKOVSKI
1972). Concentrations ranging from 0.08 to 500vg malathion/£in
some surface waters have been reported (GUERRANT et al. 1970,
DUPUY and SCHULZE 1972, ENVIRONMENTAL PROTECTION AGENCY 1972b)
but interpretation of effects of residues on non-target species
is difficult because the toxic agent during poisoning is a "per-
sistent" metabolite bound to an enzyme in a form not identifi-
able by analytical chemical analysis of animal tissue (ALDRIDGE
1971, FUKUTO 1971). Poisoning results from accumulation of a
neurotransmitter substance (acetylcholine) because the active
site of its hydrolyzing enzyme (acetylcholinesterase) of nerve
cells is phosphorylated by dimethyl or methyl phosphate after
conversion of malathion to its oxygen analog (O'BRIEN 1960,
KOELLE 1963, KARCZMAR 1970, ALDRIDGE 1971, FUKUTO 1971). In ani-
mals from the natural environment, enzyme inhibition is measurable
in nerve tissue and indicates poisoning even though chemical resi-
dues of the enzyme-bound pesticide metabolites are not measurable.
After inhibition of acetylcholinesterase, the enzyme-bound
metabolites of malathion may cause inhibition for several weeks
after exposure is discontinued (WEISS 1961, CARTER 1971, POST and
LEASURE 1974) and the parent compound disappears from water.
Although acetylcholinesterase inhibition in animals from the
environment indicates poisoning, numerous anticholinesterase
pesticides are applied to the environment and the specific agent
or agents causing poisoning need to be identified. To infer what
parent compound caused poisoning, it may be necessary to find
metabolites that are readily measurable during poisoning. In this
report, we determine the relation of short-term measurability of
malathion and some of its metabolites in fish to poisoning of
fish in the laboratory. Degree of poisoning is determined by brain
acetylcholinesterase inhibition and deaths in exposed populations.
Gulf Breeze Contribution Number 275
283
Bulletin of Environmental Contamination & Toxicology,
Vol. 16, No. 3 ® 1976 by Springer-Verlag New York Inc.
-------
MATERIALS AND METHODS
Exposure of fish in the laboratory; Three laboratory exposures
of fish were made. The first exposure was to 75 yg malathion/£.
of flowing seawater. The second exposure was to 30 yg malathion/£
and the third to 20 yg malathion/£. Pinfish, Lagodon rhomboides,
(52-101 mm total length) were obtained from wild fish populations
and acclimated to laboratory conditions at least 2 weeks before
testing. In each test, 8 replicates of 10 fish each were exposed
to technical grade malathion (95% pure) in 8-liter acrylic plas-
tic aquaria. The malathion was dissolved in acetone and infused
into the water by means of a Lambda--'pump or syringe pump. Solvent
infusion never exceeded 5 mg/£ of water and did not affect acetyl-
cholinesterase activity (COPPAGE et al. 1975). Temperature ranged
from 23-29°C and salinity from 11-29 parts per thousand during the
tests. The fish exposed to 75 yg malathion/£ were sacrificed
after 24 h for residue analyses of whole body, brain, liver, flesh
(muscle) and gut (whole gut with contents). In exposure to 30
yg malathion/£, 3 replicate samples of fish were removed from the
replicate aquaria at 0.5, 1, 4, 8, 24, 48 and 72 h for analyses
of brain acetylcholinesterase and residues in gut. In the 20
yg malathion/^ exposure, fish samples were removed for analyses
at 1, 6, and 24 h exposure and at 24, 48, 120 and 192 h after
exposure was discontinued (depuration).
Determination of enzyme activity in laboratory exposures: The
acetylcholinesterase of the pinfish brain was characterized and
assayed with a recording pH stat (COPPAGE 1971). Each assay
sample consisted of pooled brains taken from 3 fish. Normal
enzyme activity was determined from 27 samples of unexposed
fish taken throughout the testing period from the same popu-
lations as fish exposed to malathion. Inhibition was deter-
mined by assay of fish that survived a designated time, and
percentage inhibition was determined by comparison with mean
normal activity. Specific enzyme activities of exposed fish
were statistically compared to control activity by Student's
t-test (p <0.005).
Residue analyses: Residues for malathion, malaoxon, malathion
monoacid (MCA), and malathion di-acid (DCA) in fish were deter-
mined (COOK and MOORE). Pooled tissue from selected organs was
placed in a 25 x 150 mm culture tube, spiked with 20 yg phen-
thoate and its acid degradation product PHA, and extracted twice
with 10 ml acetonitrile which had been acidified with 2% 2 N HC1.
The extraction was carried out on a Willems polytron for 30 sec
at 20,000 rpm. After extraction, the culture tubes were cen-
trifuged and the acidified acetonitrile decanted into 100 ml of
2% aqueous sodium sulfate. The pH was adjusted to 8 and
.rvard Apparatus Co., Millis, Mass. Mention of commercial
products does not constitute endorsement by the Environmental
Protection Agency.
284
386
-------
malathion and phenthoate separated by extraction twice with 20 ml
petroleum ether. Malaoxon was removed by extracting the aqueous
solution twice with 20 ml methylene chloride. MCA, DCA, and PHA
were removed by adjusting the pH of the aqueous solution to 2,
adding 5% (wt/vol) solid sodium chloride, and extracting twice
with 50 ml ethyl acetate. The petroleum ether and methylene
chloride extracts were dried by eluting through a plug of anhydrous
sodium sulfate and collected in a crystallizing dish. The ethyl
acetate extracts were dried by eluting through a plug of acidified
sodium sulfate, collected in a crystallizing dish, and methylated
with diazomethane as described by SCHINK and GELLERMAN (1960).
The extracts were concentrated by placing the crystallizing
dish in a hood on a slide warmer at 50 C and passing a gentle
stream of air over them. Recovery of the internal standard,
phenthoate, was determined as a measure of integrity of the
malathion analysis and recovery of the acid internal standard,
PHA, was determined as a measure of the integrity of the MCA and
DCA analyses.
All residue analyses were performed without cleanup on a
Tracer MT-220 gas chromatograph equipped with a flame photometric
detector operating in the phosphorus mode. The column was a 182
cm x 3 mm ID glass column packed with 2% OV-101 on Gas Chrora
Q 100/120 mesh. Operating conditions were: column 175 C, inlet
225 C, detector 165 C; nitrogen 65 ml/min, hydrogen 200 ml/min,
oxygen 15 ml/min, and air 50 ml/min.
Water samples were analysed as previously described (COPPAGE
et al. 1975).
RESULTS AND DISCUSSION
Residues found in whole body and organs after exposure of
pinfish to 75 yg malathion/-£ are shown in Table I.
Residues of malathion or malaoxon were not found in any body
tissue in concentrations greater than 0.10 yg/g after 24 hr expo-
sure to 75 yg malathionAC. However, malathion monoacid and di-
acid were found in all tissues with the greatest residues in gut
and liver. This indicates rapid conversion of malathion to other
compounds with the major portion of the malathion being hydrolyzed
by carboxyesterase enzyme to monoacid and di-acid (O'BRIEN 1960).
Assuming malaoxon is the active enzyme inhibitor, the fraction of
malathion converted to malaoxon must be small (below 0.10 yg/g) or
malaoxon reacts very rapidly to form other compounds and to phos-
phorylate proteins such as acetylcholinesterase.
285
387
-------
TABLE I
Malathion, malaoxon, malathion -monoacid (MCA), and malathion di-acid
(DCA) residues in pinfish after 24 h exposure to 75 yg malathion/£
Organ
Whole body
Brain
Liver
Gills
Flesh
Gut
Malathion
NDa
ND
ND
ND
ND
ND
Residue (yg/g)
Malaoxon
ND
ND
ND
ND
ND
ND
MCA
4.2
1.7
6.0
2.5
3.9
31.4
DCA
0.30
0.22
1.25
0.36
0.34
3.70
indicates less than 0.10 Ug/g
The degree of brain acetylcholinesterase inhibition resulting
from lethal poisoning of pinfish by exposure to 30 yg malathion/-£
and accumulation of malathion monoacid and di-acid in the gut are
shown in Table II.
These data show accumulation of monoacid in the gut that coin-
sides with phosphorylation of acetylcholinesterase in brain and
poisoning in short-term continuous exposure to relatively con-
stant levels of malathion in seawater. The brain acetylcholines-
terase was inhibited 79% after 60% mortality and monoacid in gut
had accumulated to 19 yg/g which indicates measurement of mono-
acid in gut is useful for determining short-term poisoning by
malathion.
286
388
-------
TABLE II
Brain acetylcholinesterase (AChE) inhibition and malathion mono-
acid (MCA) and di-acid (DCA) residues in gut of pinfish exposed
to 30 yg malathion/£ for 72 ha
Hours
exposed
Control
0.5
1
4
8
24
48
72
Percent
killed
-
0
0
0
0
0
0-10
60
AChEb Inhibition
activity (%)
2.01 ± 0.19(27)°
1.91 ± 0.23(3)
1.85 ± 0.12(3)
1.83 ± 0.13(3)
1.49 ± 0.14(3)
0.90 ± 0.30(3)
1.05 ± 0.16(3)
0.43 ± 0.10(3)
-
5
8
9
26d
55d
48d
79d
Residues
MCA
0
0.95
1.9
2.1
4.8
11.1
10.9
19.0
(ug/g)
DCA
0
0.29
0.19
0.48
0.50
0.70
1.00
0.92
aThe measured concentration of malathion in water was 26.9 ± 0.57
yg/£; no residues of malathion or malaoxon ( 0.10 yg/g) were
detected in gut during the test.
AChE activity = ymoles acetylcholine hydrolyzed/h/mg brain tissue.
CNumbers in parentheses indicate replicate samples assayed.
Significantly inhibited at p < 0.005 (Student's t-test).
Data from the exposure of pinfish to 20 yg malathion/£ for
24 h followed by 192 h depuration are shown in Table III.
Exposure of pinfish to the lower concentration of malathion
(20 yg/£) produced correspondingly less brain acetylcholinester-
ase inhibition and monoacid residues in gut than the exposure to
greater concentration. Only 29% brain acetylcholinesterase inhi-
bition was produced in 24 h and a maximum residue of 7.2 yg mono-
acid/g of gut was found after 24 h - the greatest residue found.
Monoacid concentration decreased sharply to 2.2 yg/g of gut 24 h
after exposure was discontinued but brain acetylcholinesterase re-
mained inhibited. Residues and enzyme inhibition were lower 48 h
287
3-89
-------
TABLE III
Brain acetylcholinesterase (AChE) inhibition and malathion mono-
acid (MCA) and di-acid (DCA) residues in gut of pinfish exposed
to 20 ug malathion/-^ for 24 h and through 192 h after exposure
Hours AChE Inhibition Residues (Ug/g)
Exposure
Control
1
6
24
-
-
-
-
Post-exposure activity (%)
2.01 ± 0.19(27)° -
1.94 ± 0.15(3) 3
1.88 ± 0.03(3) 6
1.42 ± 0.12(3) 29d
24 1.35 ± 0.15(3) 33d
48 1.65 ± 0.05(3) 18d
120 1.84 ± 0.15(3) 8
192
MCA
0
1.2
3.0
7.2
2.2
0.94
0.14
ND
DCA
0
0.13
0.17
0.37
0.28
0.29
0.09
ND
a
The measured concentration of malathion was 19.8 ± 0.77 yg/£; no
residues of malathion or malaoxon (>0.10 ug/g) were detectable
in gut during the test.
AChE activity = ymoles of acetylcholine hydrolyzed/h/mg brain
tissue.
c
Numbers in parentheses indicate replicate samples assayed.
Significantly inhibited at p <0.005 (Student's t-test)
after exposure was discontinued. However, measurable concen-
trations of monoacid and di-acid remained in gut 120 h after
exposure but brain acetylcholinesterase inhibition had returned
to normal range. No residues were found in the gut 192 h after
exposure was discontinued.
288
390
-------
These findings suggest that, although monoacid and di-acid
residues in gut are probably not causally related to poisoning,
they are the most readily measurable metabolites produced during
short-term acute poisoning by malathion. Since malathion and
malaoxon are converted to other compounds during poisoning and
are not readily measurable, monoacid residues in gut concomitant
with brain acetylcholinesterase inhibition should confirm mala-
thion as a cause of poisoning in fish.
It has been shown that some members of the Order Cyprini-
formes can accumulate malathion in body tissues. KANAZAWA (1975)
exposed "Motsugo", Pseudorasbora parva, to 1.2 mg malathion/£
freshwater in static tests and'reported about 2.5 pg/g in whole-
body on day 1 with rapid decrease to 0.001 mg/£ water and about
0.1 yg/g in fish by day 7. BENDER (1969) exposed carp, Cyprinus
carpio, to 5 mg malathion/£ freshwater (a concentration not
likely to occur in natural water) for 96 h and reported concen-
trations of 2.58 pg/g brain, 4.97 pg/g blood, 3.23 pg/g gills,
66.59 pg/g liver and 28.43 yg/g flesh.
Concentrations in flesh of carp exposed to 1.0, 2.5, 5.0,
and 7.5 mg/£ water for 96 h were progressively greater with
greater concentration (low = 1 pg/g, high = 42 yg/g). Carp
exposed for 8 days to 5 mg malathion/^ attained relatively con-
stant concentrations in flesh at day 4 (3 yg/g in 1 day, 39 pg/g
in 4 days and 32 pg/g in 8 days). Carp weighing about 1 kg were
fed capsules containing 200 mg malathion and their flesh was
analyzed for malathion residues 24 h later. Flesh contained only
0.82 pg/g and 0.87 pg/g in duplicate tests indicating low uptake
of malathion through gut, rapid conversion in gut, or rapid con-
version elsewhere in the body after absorption through gut. Carp
and other fishes of the Order Cyprinifonnes are many-fold more
tolerant to malathion than other fishes (MACEK and McALLISTER
1970) which allows them to survive in concentrations that would
kill more sensitive species such as pinfish before malathion
reached detectable levels in tissues.
We conclude that periodic environmental monitoring by chem-
ical analyses of sensitive fishes or water for malathion or
malaoxon would not show poisoning caused by enzyme-bound metabol-
ites because malathion and malaoxon are rapidly absorbed and
metabolically altered in fish. Parent pesticide remaining in
water is probably dispersed or altered (GUERRANT et al. 1970).
Our data indicate analysis for malathion monoacid in gut and
measurement of brain acetylcholinesterase activity in fish from
the natural environmental are a practical measure of poisoning
caused by malathion in sensitive fish. Measurable concentra-
tions of parent compound need not be present, but acetylcho-
linesterase inhibition in brain must occur for fish to be
poisoned.
ACKNOWLEDGMENT
We thank Mr. Edward Matthews for assistance in assays.
289
391
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REFERENCES
ALDRIDGE, W. N. : Bull. W. H. 0. 44, 25 (1971).
BENDER, M. E.: Prog. Fish Cult. 31, 155 (1969).
CARTER, F. L.: Ph. D. dissertation, Louisiana State Univ. Baton
Rouge, LA. (1971).
COOK, G. H. and J. C. MOORE: To be published.
COPPAGE, D. L.: Bull. Environ. Contain. Toxicol. ^, 304 (1971).
COPPAGE, D. L. and T. W. DUKE: Proc. 2nd Gulf Coast Conf. Mosq.
Suppr. Wildl. Manage. New Orleans, LA, pp. 24-31 (1971).
COPPAGE, D. L., E. MATTHEWS, G. COOK and J. KNIGHT: Pestic.
Biochem. Physipl. £, (1975).
DUPUY, A. J. and J. A. SCHULZE: Texas,Water Development Board,
Report 149, Austin, TX. (1972).
FUKUTO, R.: Bull. W. H. 0. 44, 31 (1971).
GUERRANT, G. 0., L. E. FETZER, Jr., and J. W. MILES: Pestic.
Monit. J. 4-, 14 (1970).
KANAZAWA, J.: Bull. Environ. Contain. Toxicol. 14, 346 (1975).
KARCZMAR, A. G., Ed.: Anticholinesterase Agents. New York:
Pergamon Press 1970.
KOELLE, G. B., Ed.: Cholinesterases and Anticholinesterase Agents.
Berlin: Springer-Verlag 1963.
MACEK, K. J. and W. A. McALLISTER: Trans. Amer. Fish. Soc. 99,
20 (1970).
O'BRIEN, R. D. : Toxic Phosphorus Esters. New York: Academic
Press 1960.
PINKOVSKI, D. D.: Mosq. News. 32, 332 (1972).
POST, G. and R. A. LEASURE: Bull. Environ. Contam. Toxicol. 12,
312 (1974).
SCHINK, H., and J. L. GELLERMAN: Anal. Chem. 32, 1412 (1960).
ENVIRONMENTAL PROTECTON AGENCY: The Pollution Potential in Pesticide
Manufacturing, Pesticide Study Series No. 5. U.S. EPA.
Washington, D. C. 1972a.
ENVIRONMENTAL PROTECTION AGENCY: The Use of Pesticides in Suburban
Homes and Gardens and their Impact on the Aquatic Environment,
Pesticides Study Series No. 2. U. S. EPA. Washington, D. C.
1972b.
ENVIRONMENTAL PROTECTION AGENCY: Pesticide Usage and its Impact on
the Aquatic Environment in the Southeast, Pesticide Study Series
No. 8. U. S. EPA. Washington, D. C. 1972c.
ENVIRONMENTAL PROTECTION AGENCY:' Patterns of Pesticide Usage and
Reduction in Use as Related to Social and Economic Factors,
Pesticide Study Series No. 10. U. S. EPA. Washington, D. C.
1972d.
WEISS, C. M.: Trans. Amer. Fish. Soc. 90, 143 (1961).
290
392
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Reprinted from Transactions of
the American Fisheries Society.
Vol. 105(4): 546-549, 1976,
with permission of the American
Fisheries Society
EFFECT OF MIREX ON PREDATOR-PREY INTERACTION IN AN
EXPERIMENTAL ESTUARINE ECOSYSTEM
Marl in E. Tagatz
Contribution No. 276
393
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Effect of Mirex on Predator-Prey Interaction
in an Experimental Estuarine Ecosystem1
MARLIN E. TAGATZ
U.S. Environmental Protection Agency
Environmental Research Laboratory, Gulf Breeze
Gulf Breeze, Florida 32561
ABSTRACT
Tests of 14- to 16-days duration were conducted to determine the distribution and sublethal
effects of mirex in an experimental estuarine ecosystem. The insecticide was translocated from
water at concentrations of 0.011 to 0.13 /us/liter to sand, plant, and animal components. An
alteration of predator-prey interaction due to mirex was manifested by a significant difference
(x2 test, a = 0.05) in survival of grass shrimp, Palaemonetes vulgaris, in control and treated
tanks after one, two, or three days of predation by pinfish, Lagodon rhomboides.
A simple experimental ecosystem was de-
signed to investigate the distribution and bio-
logical effects of pesticides in estuaries. The
objective was a system that can be used to
screen pesticides in estuarine waters in a way
comparable to the Metcalf et al. (1971) system,
developed for evaluating biodegradability and
ecological magnification of toxicants in fresh-
water microcosms. Modifications of that sys-
tem include those used by Isensee et al. (1973)
and Booth et al. (1973) to study various
herbicides.
Components of our system are representative
of those forming the turtle grass (Thalassia
testudinum) communities which are abundant
along the northern Gulf of Mexico from depths
of about 0.5 m to more than 4.5 m (Humm
1956). The dominant food web consists of
turtle grass, grass shrimp (Palaemonetes vul-
garis), and pinfish (Lagodon rhomboides)
(Hansen 1969; Adams and Angelovic 1970).
Tests were conducted to determine if the
system could be used to obtain data on the
distribution and sublethal impact of mirex2,
an organochlorine insecticide applied in the
form of mirex bait to control the imported fire
ant, Solenopsis richteri Forel, in the south-
eastern United States. Mirex was chosen for
this study because of its chemical stability,
which minimizes the additional complication
of toxic metabolites.
1 Contribution No. 276, Environmental Research
Laboratory, Gulf Breeze.
2Dodecachlorooctahydro-l,3,4-metheno-2H-cyclobuta
[cd] pentalene; bait form consists of 84.7% corn cob
grits, 15.0% soybean oil, and 0.3% mirex.
Sublethal concentration of mirex were used
in these tests. Earlier studies (Tagatz et al.
1975) indicated that neither the concentration
of mirex used in the present tests (average
< 0.05 /ig/liter) nor the duration of exposure
(14 to 16 days) would cause significant
mortality of grass shrimp. For example, in
these earlier studies, no shrimp deaths were
observed before day 17 at exposures averaging
0.09 jug/liter; however, significant mortality
did occur at higher concentrations, or longer
exposure times, or at combinations of both.
My criterion of effect was alteration of
predator-prey interaction. The concept that
toxicants or other stresses affect predation is
not new. A simple technique using a fresh-
water system was described by Goodyear
(1972). Experiments with freshwater fishes
have shown that predator avoidance is im-
paired by radiation (Goodyear 1972), thermal
stress (Sylvester 1972; Coutant et al. 1974),
and mercury (Kania and O'Hara 1974). Re-
sults of studies by B. L. Howes (Cook College,
Rutgers University, personal communication,
July 8, 1975) indicate that a population de-
cline of the fiddler crab, Uca pugnax, in a New
Jersey salt marsh was the result of increased
avian predation due to deterioration of the
crab's escape-response caused by exposure to
the mosquito larvicide, abate.
METHODS
Static tests were conducted in glass-covered
180-liter glass aquaria in a constant-tempera-
ture room that maintained water temperature
at 20 ± 1 C. Illumination was provided by
546
395
-------
TAGATZ—MIREX AND PREDATOR-PREY RELATIONS
547
fluorescent tubes, using alternating 12-hour
periods of light and darkness. Dissolved oxy-
gen was maintained by aeration with an air-
stone at each end of the aquarium.
The system consisted of:
1. 160 liters of artificial seawater (distilled
water and marine mix3) maintained at a
salinity of 20 parts per thousand.
2. 4 cm of sand dredged from Santa Rosa
Sound, Florida, consisting of 25% coarse
particles (#35 standard sieve) and 75%
medium particles (#120 sieve).
3. 75 turtle grass plants, Thcdassia testu-
dinum (175-210 g total), planted over %
(1,500 cm2) of the bottom.
4. 75 adult grass shrimp, Palaemonetes vul-
garis (30-35-mm rostrum-telson length, 0.20-
0.25 g).
5. 2 juvenile pinfish, Lagodon rhomboides
(90-95 mm total length, 9-12 g).
Mirex was introduced into the water by air-
induced water flow through a glass column
(20 mm diameter, 85 mm long with screened
ends) that contained 0.630 g of mirex bait
(P. W. Borthwick, unpublished data, Environ-
mental Research Laboratory, Gulf Breeze).
The column provided for the continuous leach-
ing of mirex from bait and was positioned in
the middle of the aquarium. Control aquaria
had similar columns that contained control
bait (corn cob grits and soybean oil, but no
mirex).
Stocks of turtle grass, grass shrimp, and
pinfish were collected from Thalassia beds in
Santa Rosa Sound. Shrimp and fish were held
at the temperature, salinity, and spatial con-
ditions of the study for one week prior to
being used in an experiment. Only apparently
healthy animals were tested.
Three experiments were conducted: Experi-
ment 1 provided data on survival of shrimp
after three days predation by pinfish; Experi-
ments 2 and 3 provided data on survival after
two and one days predation, respectively.
In the first experiment, the ecosystem was
replicated in three treated and three control
tanks. The system, without pinfish, was equi-
3 Rila Products, Teaneck, New Jersey. Mention of
commercial products does not constitute endorsement
by the U.S. Environmental Protection Agency.
librated for four days, after which it was
exposed to mirex for 16 days. Pinfish were
added for the last three days of exposure.
In each of the second and third experiments,
one exposure and one control tank were used.
The system in each tank was equilibrated for
six days. In the second experiment, the expo-
sure tank was treated with mirex for 15 days;
pinfish were added to the exposure tank and
to the control tank for the last two days of
exposure. In the third experiment, the expo-
sure tank was treated with mirex for 14 days
and pinfish were added to it and to the control
tank for the last day of exposure.
Percentage survival of grass shrimp was
determined (a) before pinfish were intro-
duced and (b) one, two, or three days after
the fish were introduced. The chi-square test
was used to determine significant differences
(a = 0.05) in numbers of dead and living
animals in treated and control tanks.
Components were sampled to determine
mirex concentrations. Samples of water (1
liter), sand (35-45 g), turtle grass (composite
of six plants; leaves, 7-8 g per sample, and
roots, 5-6 g per sample), and shrimp (com-
posite of three shrimp, 0.6-0.8 g per sample)
were obtained from each tank after one day of
exposure and every third day thereafter. The
two pinfish (9-12 g each) per tank were
collected at the end of the exposure period for
individual analysis. A water sample was ob-
tained by siphoning from mid-depth along the
length of the aquarium and was replaced by
an equal volume of water. The top 1 cm of
sand was sampled by filling a 30-ml glass vial
by scraping the vial across the surface of the
sand.
Samples were analyzed by electron-capture
gas chromatography, using a modification of
the method of Schimmel et al. (1976). Except
for residues in water, the present report gives
residue data only for Experiment 1, in which
whole animals were analyzed for mirex. In
Experiments 2 and 3, individual tissues of the
animals were analyzed, detail beyond the scope
of this paper.. Analytical sensitivity for Ex-
periment 1 was: grass and whole animal (wet-
weight basis) and sand (air-dried weight
basis), 0.02 mgAg; and water samples, 0.01
/xg/liter. Samples to which known amounts
396
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548
TRANS. AM. FISH. SOC., 1976, NO. 4
TABLE 1.—Survival of grass shrimp after 13 days exposure to mirex in the absence of predation and after an
additional one to three days of exposure and predation by two pinfish per tank. (Numbers in parentheses
indicate number of shrimp in tanks before deaths occurred.)
No, tanks
( control/
treated )
1/1
1/1
3/3
Average
concentration
mirex
(/tg/liter)
0.046
0.044
0.025
Survival after 13 days
Control
82% (63)
78% (63)
94% (189)
Treated
68%
81%
88%
(63)
(63)
(189)
Chi-
square"
NS
NS
NS
No.
days of
predation
1
2
3
Survival after predation
Control
44%
16%
24%
(52)
(49)
(177)
Tre
23%
0%
4%
;ated
(43)
(51)
"(115)
Chi-
square"
4.57*
9.05**
19.39**
»NS = non-significant; * significant at 5% level (x2, 1 d.f. = 3.84); ** significant at 1% level (x2, 1 d.f. = 6.63).
11 Based on two, instead of three, treated tanks due to death of a pinfish in the third tank.
of mirex were added gave recoveries greater
than 85%, but concentrations were not cor-
rected for percentage recovery.
Measurements of pH, turbidity, and dis-
solved oxygen of the water were obtained with
appropriate meters twice during equilibration
and whenever components were sampled for
mirex residue analysis.
RESULTS AND DISCUSSION
Observations of the system and water data
indicated a healthy community that appeared
to be of sufficient size not to be stressed by
removal of replicate samples. The water re-
mained clear, turbidity averaging 0.6 nephelo-
metric unit (range 0.2 to 1.8). No algal
growth was observed in the tanks during
Experiment 1, but small amounts were visible
near the end of Experiments 2 and 3. Dis-
solved oxygen ranged from 7.2 to 8.5 parts
per million, but was usually at the saturation
level (8.1 ppm). Overall pH of the water
averaged 7.2 (range 7.0 to 7.4). All plants
survived, some showed new growth, and most
of their leaves remained green. Shrimp were
closely associated with the plants, eating leaf
detritus or epiphytic material on the leaves or
both.
Concentration of mirex in the water aver-
aged 0.025 /tg/liter (range 0.015-0.050 ^g/
liter) for Experiment 1, 0.044 /Ag/liter (0.011-
0.13 /ig/liter) for Experiment 2, and 0.046
jug/liter (0.017-0.11 /ig/liter) for Experiment
3.
Mirex was translocated from water to sand
and biota. All components sampled during 16
days of 'exposure in Experiment 1 contained
mirex. Only trace amounts (< 0.02 mgAg)
occurred in sand. Mirex was not detected in
Thalassia at day 1, but subsequent concen-
trations in leaves (trace to 0.033 mg/kg) were
as great as 1,300 X the average concentration
in the water. Roots had less mirex (not de-
tected, < 0.020 mg/kg, to 0.024 mgAg) than
did leaves (trace to 0.033 mg/kg). Mirex in
shrimp ranged from trace amounts to 0.20 mg/
kg. Concentrations in shrimp increased with
time and were as great as 8,000 X the average
concentration in the water. After three days
exposure, pinfish contained 0.050 to 0.063
mgAg mirex, up to 3,800X that in the water.
In all experiments, an alteration erf pred-
ator-prey interaction due to the effect of mirex
was evident. There was no significant differ-
ence (a = 0.05) in survival of grass shrimp
in control and treated tanks after 13 days
exposure in the absence of predation by pin-
fish. However, there was a significant differ-
ence (a = 0.05) in survival after one, two, or
three days of predation by pinfish (Table 1).
Survival of shrimp after three days of preda-
tion was based on two, instead of three, treated
tanks due to death of a pinfish in the third
tank.
That more deaths due to predation occurred
in the treated tanks than in the control tanks
could be interpreted as an alteration of the
normal behavior of either shrimp or pinfish
by mirex; I believe, however, that the behavior
of only the grass shrimp was altered. This is
supported by the experiments of Tagatz et al.
(1975), who found that higher concentrations
of mirex than used here caused deaths among
grass shrimp, and those of Lowe et al. (1971),
who found that on the basis of mortality,
pathology, and observations for symptoms of
pesticide poisoning, mirex had no apparent
effect on pinfish.
397
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TAGATZ—MIREX AND PREDATOR-PREY RELATIONS
549
Comparison of deleterious effects of toxi-
cants consisting of nonadaptive behavioral
changes in animals in experimental ecosystems
with mortality data from single-species bio-
assays should be useful in better defining safe
concentrations in the environment. The con-
centration of mirex that increased predation
is, to my knowledge, the lowest concentration
of mirex in water that has been reported to
cause death of an estuarine animal. Thus, for
mirex, it would be useful to determine the
threshold concentration for this selective pre-
dation. Behavioral indices should also serve
as measures of stress for other toxicants.
ACKNOWLEDGMENTS
P. W. Borthwick, T. A. Hollister, Dr. W. P.
Schoor, and Dr. G. E. Walsh participated in
the design of the experimental ecosystem. P.
W. Borthwick provided the mirex columns,
J. M. Ivey helped with the bioassays, and J.
Forester and J. Knight performed the chemical
analyses.
LITERATURE CITED
ADAMS, S. M., AND J. W. ANCELOVIC. 1970. Assimi-
lation of detritus and its associated bacteria by
three species of estuarine animals. Chesapeake
Sci. 11: 249-254.
BOOTH, G. M., C. Yu, AND D. J. HANSEN. 1973.
Fate, metabolism, and toxicity of 3-isopropyl-lH-
2,l,3-benzothiadiazin-4 (3H)-1,2, 2-dioxide in a
model ecosystem. J. Environ. Qual. 2: 408-411.
COUTANT, C. C., H. M. DUCHARME, JR., AND J. R.
FISHER. 1974. Effects of cold shock on vulner-
ability of juvenile channel catfish (IctaluTus
punctatus) and largemouth bass (Micropterus
salmoides) to predation. J. Fish. Res. Board
Can. 31:351-354.
GOODYEAR, C. P. 1972. A simple technique for
detecting effects of toxicants or other stresses^ on
a predator-prey interaction. Trans. Am. Fish.
Soc. 101:367-370.
HANSEN, D. J. 1969. Food, growth, migration, re-
production, and abundance of pinfish, Lagodon
rhomboides, and Atlantic croaker, Micropogon
undulatus, near Pensacola, Florida, 1963-65. U.S.
Fish Wildl. Serv., Fish. Bull. 68: 135-146.
HUMM, H. J. 1956. Sea grasses of the northern
Gulf coast. Bull. Mar. Sci. Gulf Caribb. 6: SOS-
SOS.
ISENSEE, A. R., P. C. KEARNEY, E. A. WOOLSON, G. E.
JONES, AND V. P. WILLIAMS. 1973. Distribution
of alkyl arsenicals in model ecosystems. Environ.
Sci. Technol. 7: 841-845.
KANIA, H. J., AND J. O'HARA. 1974. Behavioral
alterations in a simple predator-prey system due
to sublethal exposure to mercury. Trans. Am.
Fish. Soc. 103: 134-136.
LOWE, J. I., P. R. PARRISH, A. J. WILSON, JR., P. D.
WILSON, AND T. W. DUKE. 1971. Effects of
mirex on selected estuarine organisms. Trans.
36th N. Am. Wildl. Nat. Resour. Conf., 171-186.
METCALF, R. L., G. K. SANGHA, AND I. P. KAPOOR.
1971. Model ecosystem for the evaluation of
pesticide biodegradability and ecological mag-
nification. Environ. Sci. Technol. 5: 709-713.
SCHIMMEL, S. C., J. M. PATRICK, JR., AND J. FORESTER.
1976. Heptachlor: Toxicity to and uptake by
several estuarine organisms. J. Toxicol. Environ.
Health. In press.
SYLVESTER, J. R. 1972. Effect of thermal stress on
predator avoidance in sockeye salmon. J. Fish.
Res. Board Can. 29: 601-603.
TAGATZ, M. E., P. W. BORTHWICK, AND J. FORESTER.
1975. Seasonal effects of leached mirex on
selected estuarine animals. Arch. Environ. Con-
tarn. Toxicol. 3: 371-383.
398
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Reprinted from Estuarine Pro-
cesses, Vol. I, Uses,
Stresses, and Adaptation to
the Estuary, Martin W. Wiley,
editor, pp. 481-482, 1976,
with permission of the Aca-
demic Press Inc. New York,
San Francisco, London
CYCLING OF POLLUTANTS
Thomas W. Duke
Contribution No. 320
399
-------
Reprinted from:
ESTUARINE PROCESSES, Vol. I
Uses, Stresses, and Adaptation to the Estuary
@ 1976
ACADEMIC PRESS, INC
N»w York San Francisco London
CYCLING OF POLLUTANTS
Convened by:
Thomas W. Duke
U.S. Environmental Protection Agency
Environmental Research Laboratory
Gulf Breeze, Florida 32561
Estuaries continue to receive pollutants such as oil, heavy metals, pesticides
and other toxic organics. It is fitting, therefore, that one session of this meeting
be devoted to the impact of these pollutants on this productive environment.
Because of their location, estuaries are susceptible to industrial, municipal, agri-
cultural and similar wastes, transmitted through freshwater streams, and other
pollutants derived from development of off-shore oil fields and wastes disposed
in oceans. It is impossible to discuss all of the important pollutants which enter
estuaries. However, for purposes of this session, we will discuss the impact of oil,
heavy metals, and pesticides on ecosystems and on biological systems ranging
from micro-organisms to fishes.
Studies designed to determine the impact of oil on the estuarine environment
are especially important with the increased interest in development and trans-
port of off-shore oil. Of particular interest is knowledge concerning the effect of
oil on estuarine microbial populations and the effect of the microbial popula-
tions on oil. Most marshes include a high percentage of cellulolytic bacteria, and
these bacteria are important in the breakdown or metabolism of oil. The concept
of seeding certain species of bacteria or yeast is also of concern at this time.
The effect of metals on estuarine organisms and their environment continues
to be investigated. The role of seagrass meadows in coastal ecosystems is just
beginning to be documented. Before the impact of metals on these and other
primary producers can be assessed, much baseline data must be developed.
Although there appears to be a decline in the level of residues of "hard"
organochlorine pesticides, such as DDT, in marine organisms, even low levels of
residue may affect the organisms and estuarine systems in which they occur. The
distribution of the pesticides, various pathways of transfer and bioaccumulation
are known in many instances, yet the ultimate effects of the pesticide on organ-
'isms and their environment are relatively unknown. Even less is understood
JJ-U.S GOVERNMENT PRINTING OFFICE 1979-645-370 4,ft 1
401'
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about the synergistic, antagonistic and additive effects of metals, pesticides and
toxic organics. The combined toxicities of methoxychlor, cadmium, and poly-
chlorinated biphenyls are discussed in this session.
482
402
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