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

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                           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

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                           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

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                     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

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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

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                                              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

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                 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

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                  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

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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

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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

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                  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

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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

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                                                                              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

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                  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

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                 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

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                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

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                 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

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                                               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

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                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.

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                                   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

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                         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

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                          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

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            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

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              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




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             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

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        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

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             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
-------
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

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       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

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             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

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        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

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                        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

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                          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

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                                                 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

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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
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Kilby, John D. 1955. The fishes of two Gulf Coastal marsh areas of Florida. Tulane
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    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

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                                               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

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    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

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          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

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                        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

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 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

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                          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

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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

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                                               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

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                               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

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                     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

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                                                 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

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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

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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

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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

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                                                 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

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  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

-------
   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

-------
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

-------
                           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
                                   142

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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.

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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

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                           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

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                                                                                                                                       -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.

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                            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

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                                                                                                                                   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

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                             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

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  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

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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

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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

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  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

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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

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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

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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

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                                              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

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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

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                                 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

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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

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                                  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.
                    REFERENCES

Aldridge W. N.  (1971) The nature of the reaction  of
   organophosphorus compounds  and  carbamates  with
   esterases. Bull W. H.  O. 44, 25-30.
Alsen C, Herrlinger A. and Ohnesorge F. K. (1973) Char-
  acterization  of cholinesterases of the cod  (Gadus  cal-
  larias)  and their in vivo  inhibition  by paraoxon  and
  tabun.  Arch. Toxicol. 30. 263-275.
Butler P. A. (1963) Commercial fisheries  investigations.
  Pesticide-Wildlife Studies during 1961 and 1962. pp. II-
  25. U.S. Fish. Wildl. Serv.  Circ.  167, Washington, D.C.
Carter F. L.  (1971). In vivo studies of brain acetylcholines-
  terasc inhibition by organophosphate and carbamate in-
  secticides  in  fish.  Unpublished  Ph.D.  dissertation,
  Louisiana  State University, Baton Rouge, Louisiana.
Coppage D. L. (1971) Characterization  offish brain acetyl-
  cholinesterase 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 sheep-
  shead minnows. Trans. Am. Fish. Soc. 101, 534-536.
Coppage D. L. and Duke T. W. (1971)  Effects of pesticides
  in estuaries along  the Gulf and  Southeast  Atlantic
  Coasts. Proc. of the  2nd Gulf Coast Conf. on Mosquito
  Suppression and Wildlife Management, pp. 24-31. (Edited
  by C.  H.  Schmidt).  National Mosquito Control—Fish
  and  Wildlife Management  Coordinating  Committee,
  Washington, D.C.
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. 438-488.
Coppage D.  L. and Matthews E. (1975) Brain-Acetylcho-
  linesterase  inhibition in a marine teleost during  lethal
  and  sublethal  exposures  to  l,2-dibromo-2,2-dichloro-
  ethyl dimethyl phosphate  (naled)  in seawater. Toxicol.
  Appl. Pharmacol. 31, 128-133.
Eaton J.  G. (1970) Chronic malathion toxicity to bluegill
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  684.
Ehrenpreis S.  (Editor)  (1967)  Cholinergic  Mechanisms.
  Ann. N. Y. Acad. Sci. 144. 385-935.
Fukuto T. R. (1971) Relationships between the structure
  of organophosphorus compounds and  their activity as
  acetylcholinesterase inhibitors. Bull. W. H. O. 44, 31-42.
Fed. Wat. Pollut. Confrol Adm. (1968) Water Quality  Cri-
  teria. Report of the National Tech. Adm. Comm. to  Sec.
  of the Interior, Fed.  Wat. Pollut. Control Adm., Wash-
  ington, D.C.
Heath  D. F. (1961) Organophosphorus Poisons, Anticho-
  linesterases and Related Compounds. Pergamon  Press,
  New York.
Holland  H. T., Coppage  D.  L. and Butler P. A. (1967)
  Use of fish brain acetylcholinesterase to monitor pollu-
  tion by organophosphorus pesticides. Bull Environ. Con-
  tam. Toxicol. 2, 156-162.
Karczmar A. G. (Editor) (1970) Anticholinesterase Agents.
  Pergamon  Press, New York.
Katz M. (1961) Acute toxicity of some  organic insecticides
  to three species of salmonids and to threespine  stickle-
  back. Trans. Am. Fish. Soc. 90. 264-268.
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  linesterase  Agents. Springer-Verlag, Berlin.
Lahav  M. and Sarig S. (1969) Sensitivity of pond fish to
  cotnion (azinphosmethyl) and parathion. Bamidgeh, Bull.
  Fish. Cult.  Israel 21. 67-74.
Lawless E. W., Von Rumker R. and Ferguson T. L. (1972)
  The Pollution Potential  in Pesticide Manufacturing.  U.S.
  Environ. Prot. Agency,  Washington, D.C.
Macek K. J.  and McAllister W. A. (1970) Insecticide  sus-
  ceptibility  of some common fish family representatives.
  Trans.  Am. Fish. Soc. 99, 20-27.
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. Am. Fish.  Soc.
  101, 420-427.
Metcalf R. L. (1971) Structure-activity relationships for in-
  secticidal carbamates. Bull. W. H. 0. 44. 43-62.
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24
D. L. COPPAOE and T. E. BRAIDECH
O'Brien R. D. (1960)  Toxic Phosphorus Esters. Academic
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O'Brien R. D. (1967)  Insecticides.  Academic Press, N.Y.
Post G.  and Leasure R. A. (1974) Sublethal effects  of
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Webb  R. G., Garrison A. W., Keith L. H. and McGuire
  J. M. (1973)  Current practice in  GC-MS  analysis  of
  organics in water. Environmental Protection  Technology
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                                                    184

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                                              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

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            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

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                       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

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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

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                       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

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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

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                       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

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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

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  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

-------
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

-------
           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

-------
                         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

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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

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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

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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

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                               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

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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

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                                             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

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         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

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               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

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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,
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  3. U.  S. Environmental   Protection  Agency,
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 4. U.  S. Environmental   Protection  Agency,
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      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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     Suppression  and  Wildlife   Management,"
     (C. H. Schmidt, Ed.), pp. 24-31. National
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      Management  Coordinating  Committee,
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 6.  Florida Bureau of Entomology. Annual Report,
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 7.  G. O. Gtierrant, L. E. Fetzer,  Jr.,  and J.  W.
      Miles, Pesticide  residues  in Hale County,
      Texas,  before  and after  ultra-low  volume
      aerial  application   of malathion,   Pestic.
      Monit. J. 4, 14 (1970).
 8.  D. D. Pinkovski, United States Air Force aerial
      spray activities  in  operation combat VEE,
      Mosq. News. 32,  332 (1972).
 9.  A. J. Dupuy and J. A. Schulze, Selected water-
      quality  records  for Texas  surface  waters,
      1970 Water Year, Texas Water  Development
      Board, Report 149,  Austin, TX  (1972).
10.  W. N.  Aldridge, The  nature of the reaction of
      organophosphorus compounds and carbamates
      with esterases,  Bull. W.H.O. 44,  25 (1971).
11.  S. Ehrenpreis, Ed.  "Cholinergic Mechanisms,"
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12.  T. R. Fukuto, Relationships  between the struc-
      ture  of  organosphosphorus  compounds  and
      then- activity as acetylcholinesterase inhibi-
      tors,  Bull. W.H.O. 44, 31 (1971).
13.  D.  F.  Heath,  "Organophosphorus  Poisons,
      Anticholinesterases and Related Compounds,"
      Pergamon Press, New York,  1961.
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
      Vitaminol. Enzymol. 24, 131 (1970).
16.  G. B. Koelle, Ed.  "Cholinesterases and Anti-
      cholinesterase Agents," Springer-Verlag, Ber-
      lin, 1963.
17.  R. D.  O'Brien, "Toxic  Phosphorus  Esters,"
      Academic Press,  New York, 1960.
18.  R. D. O'Brien, "Insecticides," Academic Press,
      New York, 1967.
19.  R. D. O'Brien, Phosphorylation and carbamyla-
      tion of cholinesterase, Ann. N.Y.  Acad.  Sci.
      160, 204 (1969).
20.  N. K. Schaffer, S.  C. May, and  W. H. Sum-
      merson.  Serine  phosphoric   acid  from  di-
      isopropylphosphoryl derivative  of  eel cholin-
      esterase, J. Biol. Chem. 206, 201 (1954).
21.  S. D. Murphy, Liver metabolism and toxicity
      of thiophosphate insecticides  in mammalian,
      avian and  piscine  species, Proc. Soc. Exp.
      Biol.  Med. 123, 392 (1966).
22.  S.  D. Murphy,  R. L. Lauwerys, and  K. L.
      Cheever,  Comparative  anticholinesterase  ac-
      tion  of   organophosphorus  insecticides  on
                                             230

-------
542
                                         COPPAGE ET AL.
      vertebrates,  Toxicol. Appl.  Pharmacol.  12,
      22 (1968).
23. M.   E.  Bender,  Uptake  and   retention  of
      malathion  by the carp,  Prog.  Fish Cult. 31,
      155 (1969).
24. F. L. Carter, "In vivo studies of brain  acetyl-
      cholinesterase inhibition by organophosphate
      and  earbamate insecticides  in fish,1'  Ph.D.
      dissertation,   Louisiana  State  University,
      Baton Rouge, Louisiana, 1971.
25. 0.  B.  Cope, Sport  fishery investigations,  in
      "Pesticide-Wildlife Studies, 1963," pp. 29^3.
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      ton,  DC, 1964.
26. O.  B.  Cope, Sport  fishery investigations,  in
      "The Effects of Pesticides on Fish  and Wild-
      life," pp. 51-63. U. S. Fish Wildl. Serv. Circ.
      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
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29. D.  L.  Coppage,  Characterization of  fish brain
      acetylcholinesterase with an automated  pH
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      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-
      zation  and the Integration  of Behavior"
      (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

-------
                                                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

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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

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    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

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   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

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     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

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   Baculovirus Prevalence in Shrimp by Chemical Exposure
                                                                     309
•-
   x

                                 246

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    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

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    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

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    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

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     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

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                                             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

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                                                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

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   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

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                        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

-------
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}

-------
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

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                                                 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

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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

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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

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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

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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

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                                                 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

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     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

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                   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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534
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

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  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

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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.

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  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

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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

-------
        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

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                             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

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                           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

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                             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

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     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

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                            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
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     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.
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10.  Segar, D. A. 1971.  The use of the heated graphite atomizer in
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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

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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
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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
    
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    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

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     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

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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

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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

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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

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           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

-------
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

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                          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

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                          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

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                                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

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                                            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

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               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

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               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.

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                      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

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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

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                            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

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                           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

-------
                 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

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                   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

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                                   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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