TOXICITY TESTING WITH FISH, ZOOPLANKTON AND MUSSELS — A COMPARISON OF SENSITIVITIES By ANNE E. KELLER A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY UNIVERSITY OF FLORIDA 1989 Copyright 1989 by Anne E. Keller To my parents, Robert and Lucille Keller, whose support and confidence in me made all the difference ACKNOWLE DGEMENTS I would like to thank Dr. Thomas L. Crisman, my committee chairman, for offering me many of the challenges that have been part of my education. My other committee members, Dr. Gabriel Bitton, Dr. Frank Nordlie, Dr. Stephen G. Zam and Dr. Clay Montague, broadened my horizons via the different perspectives they hold. I am grateful for that. Many other people in this department and elsewhere were instrumental in the completion of this work. I thank them for their generosity and care. Friendship often makes frustration and hard work less difficult. Though they deserve much more, I thank my family and friends for their faith and trust in me. IV TABLE OF CONTENTS page ACKNOWLEDGMENTS iv ABSTRACT viii CHAPTERS 1 INTRODUCTION 1 2 LITERATURE REVIEW 6 Toxicity Testing with Pelagic Biota 6 Rationale for the Use of Freshwater Molluscs in Acute Toxicity Tests 10 Distribution and Life History of Unionid Mussels • 11 Habitat Destruction and Faunal Decline 13 Propagation of Freshwater Mussels in Artificial Media 16 Sensitivity of Unionid Molluscs to Environmental Pollutants 18 THE SENSITIVITY OF THE FATHEAD MINNOW (PIMEPHALES PROMELAS) TO HYDROTHOL-191 AT 15° and 25" C 29 Introduction 29 Materials and Methods 33 Test Organism 33 Test Organism food 33 Dilution Water 33 Test Chemical 34 Reference Toxicant 34 Range-finding Test 35 Seven-day Survival and Growth Toxicity Test 35 Statistical Analysis 36 Results 37 Dilution Water Quality 37 Survival and Growth of Fathead Minnow Larvae 37 Discussion ^^ AN ASSESSMENT OF THE CHRONIC TOXICITY OF HYDROTHOL-191 TO THE ZOOPLANKTER CERIODAPHNIA DUBIA USING A 7 -DAY SURVIVAL AND REPRODUCTION TEST 56 Introduction 56 Materials and Methods 58 Test Organism 58 Test Chemical 58 Dilution Water 58 Test Organism Food 60 Reference Toxicant Tests 60 Range-finding Test 61 Preparation for Chronic Toxicity Tests.. 61 Chronic Toxicity Tests 62 Statistical Analysis 63 Results 64 Reference Toxicant 64 Acute Toxicity 64 Chronic Toxicity 67 Discussion 74 SIMPLIFICATION OF IN VITRO CULTURE TECHNIQUES FOR FRESHWATER MUSSELS 79 Introduction 79 Materials and Methods 81 Test Organism 81 Plasma Substitutes 82 CO2 Incubator 84 Use of Commercial Media 85 Species Collected 86 Results 87 Discussion 95 A TEST PROTOCOL FOR DETERMINING THE ACUTE TOXICITY OF POLLUTANTS TO JUVENILE FRESHWATER MUSSELS 98 Introduction 98 Materials and Methods 104 General Conditions 104 Physical Conditions 105 Water Quality 106 Feeding Tests 108 Test Organisms 109 Results 114 Discussion 118 THE TOXICITY OF SELECTED METALS TO THE VI FRESHWATER MUSSEL, ANODONTA IMBECILIS AND THE ZOOPLANKTER, CERIODAPHNIA DUBIA 122 Introduction 12 2 Materials and Methods 127 Test Organisms 127 Test Methodology 127 Dissolved metals 127 Metal mixtures 128 Sediment tests 129 Metal effluent 130 Test chemicals 131 Data Analysis 131 Results 133 Dissolved Metals 133 Metal Mixtures 143 Sediment Tests 147 Effluent Toxicity 148 Discussion 150 8 THE TOXICITY OF SEVERAL PESTICIDES, ORGANIC COMPOUNDS AND A WASTEWATER EFFLUENT TO THE FRESHWATER MUSSEL, ANODONTA IMBECILIS. THE ZOOPLANKTER, CERIODAPHNIA DUBIA AND THE FATHEAD MINNOW, PIMEPHALES PROMELAS 156 Introduction 156 Materials and Methods 162 Test Organisms 162 Test Conditions 163 Aqueous exposures 163 Karate, atrazine and carbaryl 164 Toxaphene and chlordane tests 165 Effluent toxicity test 167 Data Analysis 168 Results 168 Aqueous Tests 168 Karate, Atrazine and Carbaryl 174 Toxaphene and Chlordane Tests 174 Effluent Toxicity Test 178 Discussion 178 9 CONCLUSIONS 184 REFERENCES 190 BIOGRAPHICAL SKETCH 212 Vll Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy TOXICITY TESTING WITH FISH, ZOOPLANKTON AND MUSSELS — A COMPARISON OF SENSITIVITIES By ANNE E. KELLER December 1989 Chairman: Thomas L. Crisman Major Department: Environmental Engineering Sciences Toxicity testing with aquatic organisms is common- place today. Prior to their manufacture and use, the effects of pesticides, herbicides and toxic substances on biota and the environment must be assessed. However, the focus has been primarily on the fate of pelagic fauna, particularly fish and zooplankton. Little attention has been focussed on the responses of invertebrates, other than insects, to pollutants. In recent years, there has been a significant decline in the once abundant freshwater mussel fauna, purportedly due to dam-building and pollution. Since little is known about the sensitivity of freshwater mussels to metals and pesticides, there has been no way to establish protective measures. Currently, the United States Environmental Protection Agency is using zooplankton as surrogates for freshwater mussels in Vlll toxicity tests with no verification that the two are comparably sensitive. This dissertation was designed (1) to determine how sensitive mussels are to metals and organics, (2) to compare the sensitivity to freshwater fish such as the fathead minnow and (3) to determine by comparison whether zooplankton are good substitutes for mussels in toxicity tests. Anodonta imbecilis was chosen as the test species because it was locally available, has a relatively long reproductive period and has been previously cultured in the laboratory. Acute toxicity tests were performed with juvenile mussels in reconstituted freshwater. Copper, cadmium, chromium, mercury, zinc and nickel were the metals used. It was found that mussels were about as sensitive to metals as were zooplankton. Organic compounds assessed included lindane, toxaphene, chlordane, Hydrothol-191 , PCP, carbaryl, atrazine, an unregistered pyrethroid pesticide, acetone, methanol and SDS. Anodonta imbecilis was not sensitive to any of these substances except PCP. It appears that the use of zooplankton species, e.g. Daphnia magna or Ceriodaphnia dubia, as surrogates for freshwater mussels is appropriate in tests for metal toxicity, but may not be so for organic pollutants. IX CHAPTER 1 INTRODUCTION Although the use of aquatic organisms to test impacts of industrial, agricultural or wastewater effluents on the biota of streams and rivers is a common occurrence today, this is a relatively new development. Initial concern centered on the safety of chemicals to humans and domestic animals relative to their efficacy on target organisms (Casarett and Bruce 1980) . However, as concern for the environment has increased, so have the number and uses of aquatic toxicity tests. There are now test methods for many vertebrate and invertebrate aquatic animals (Peltier and Weber 1985) . Aquatic toxicology arose as an outgrowth of the chemical revolution of the 1940s. Biologists, seeing adverse changes in the biota of streams receiving human and industrial wastes, advocated the use of fish or other aquatic species as a means of predicting the response of stream organisms to industrial wastes (Buikema et al. 1982). Regulatory control of water quality was established in the U.S. with the passage of the Federal Insecticide, Fungicide and Rodenticide Act (FIFRA) in 1947 and the Federal Water Pollution Control Act (FWPCA) in 1948. These pieces of legislation regulated the input of both conventional municipal and industrial waste into lakes, rivers and the ocean. In 1970, when United States Environmental Protection Agency (USEPA) became the agency responsible for improving and protecting the nation's water resources, the major concern was still the impact on both drinking water supplies and harvestable aquatic species. While sewage treatment and receiving water quality improved over time, as recently as 1982 some 30 states were cited for water quality standards violations due to toxic pollutants (wise 1985) . Biomonitoring of effluents via acute toxicity tests became the standard method for assessing environmental impacts because such tests are inexpensive and directly measure biological response. Acute toxicity tests are also required for all new chemicals posing a potential risk to human health or the environment, and as supporting documentation for pesticide registration (Zucker 1985a, 1985b). To date, over 165 species of aquatic organisms have been used in acute toxicity tests (Buikema et al. 1982). Rapid methods to assess chronic toxicity have recently been developed for several species, including Pimephales promelas, Ceriodaohnia dubia and Selenastrum capricornutum (Horning and Weber 1985) . These new methods measure sublethal effects of toxicants on biota, but acute toxicity tests are still the more commonly used. While it would be convenient to use only one or a few species to assess the impacts of various pollutants on aquatic organisms, such a practice would not be adequate. Species specific tests are necessary because of differences in biological sensitivity and because organisms from different habitats may be exposed to a wide range of pollutant concentrations. The need for a variety of test methods has been demonstrated by results of testing with many pollutants (Johnson and Finley 1980, Mayer and Ellersieck 1986) . Most of the accepted test organisms are pelagic species (fish and zooplankton) which are well-known, easy to culture, economically important or of public concern (Buikema et al. 1982). However, benthic organisms are more appropriate for use in tests assessing the impacts of pollution in flowing waters because they typify the fauna of lotic systems. Pelagic fauna are more representative of lacustrine habitats. To date, toxicity data for benthic organisms including aquatic oligochaetes, turbellarians, pelecypods and gastropods are still extremely limited, and only one toxicity test protocol for these organisms exists, i.e., for the exotic Corbicula fluminea (Foster 1979) . Because its life cycle differs significantly from that of native North American freshwater clams (Unionidae) , the use of C. fluminea as a model for other mollusk species is questionable. The current dissertation research is divided into two large parts. The first part deals with the assessment of the chronic toxicity of Hydrothol-191 , an aquatic herbicide, to Pimephales promelas (the fathead minnow) and Ceriodaphnia dubia. using recently developed EPA protocols. The Florida Department of Environmental Regulation foresaw an increased demand for Hydrothol use in Florida and wanted to know what impact this might have on nontarget organisms. While both the fathead minnow and C. dubia are used to monitor the toxicity of wastewater effluents and assess the impact of pure compounds on aquatic organisms, neither of them is native to Florida (Lee et al . 1980). In addition, their responses have not been compared to those of benthos that inhabit canals, streams or rivers where Hydrothol is widely used for macrophyte control. Therefore, the second part of the dissertation contains results of test development work with a representative of native benthic fauna, the freshwater mussel, Anodonta imbecilis. The need for a toxicity test for native freshwater mollusks was apparent based on their importance in flowing waters, their taxonomic distinction from insects and other benthos that have been tested, and the need to corroborate the use by EPA of tests with Daphnia magna to estimate the sensitivity of mussels to pollutants. My research with A. imbecilis was designed to (1) simplify the culture techniques permitting easier production of test organisms, (2) develop an acute toxicity test protocol for use in assessing the sensitivity of freshwater mussels to pesticides, metals and wastewater effluents, and (3) determine the toxicity of a number of pure compounds and effluents to A. imbecilis. Results from this work could then be used to determine whether it is appropriate for EPA to use D. magna in toxicity tests as a surrogate for freshwater mussels. CHAPTER 2 LITERATURE REVIEW Toxicity Testing With Pelagic Biota The bulk of information on the toxicity of pollutants to aquatic biota was derived from tests with pelagic organisms. In particular, several species of economically important fish, e.g., Salmo gairdneri, Oncorhynchus tshawytscha, Lepomis macrochirus and Ictalurus punctatus, and a number of zooplankton species, e.g. Daphnia magna, Daphnia pulex, and Simocephalus spp. have been the most common test organisms (Johnson and Finley 1980, Mayer and Ellersieck 1986, Buikema et al. 1982). The latter group has been well studied because they are both easy to rear in the laboratory and important links in the aquatic food chain leading to fish. Since the late 1970s there has been increasing interest in the development of short term chronic toxicity tests for fish that combine the simplicity of acute methods with the estimation of sublethal effects provided by lifecycle toxicity tests (Horning and Weber 1985) . The latter can require months or years to complete, depending on the lifespan of individual species. Chronic exposures to low concentrations of pollutants can affect reproduction, growth, behavior or species interactions, any of which may alter the structure of the aquatic community (Alabaster and Lloyd 1982, Rand 1985) . Studies by McKim (1977) and Macek and Sleight (1977) proved that exposure of critical life-stages of fish (embryos or larvae) to toxicants for 30-60 days provided toxicity estimates comparable to full life cycle tests. These early life stage (ELS) tests were soon adopted as the standards for estimating water quality criteria because they were faster and cheaper, as well as accurate (Horning and Weber 1985) . Further simplification followed as data from ELS tests showed that larval growth could be as sensitive a measure of sublethal toxicity as larval survival (Benoit et al. 1982, Weltering 1984, Birge et al. 1981). As a result, a seven-day fathead minnow larval survival and growth test was developed for effluent and single- compound toxicity evaluations (Norberg and Mount 1985) . The method, published by EPA (Horning and Weber 1986) , is described as a static-renewal subchronic toxicity test that uses larval growth as a measure of sublethal response. Larval fathead minnows (< 24 h old) are exposed to a series of toxicant concentrations (usually 5) and a control comprised of dilution water. The test solutions are changed daily, after a count of survivors has been made. Larvae are fed rinsed brine shrimp. At the end of the test, all surviving larvae are preserved in formalin until their growth can be assessed based on weight gain compared to that of controls. An LC50 is calculated using survival data. The second subchronic toxicity test to be published by the EPA was the Ceriodaphnia dubia survival and reproduction test (Horning and Weber 1985) . The impetus behind the development of the 7-day Ceriodaphnia survival and reproduction was somewhat different from that of the test with fathead minnows. While test duration was a factor, it was perhaps more related to the length of the work week than to expense since a cladoceran life cycle test may be completed in about 3 0 days (Mount and Norberg 1984). If the test is begun on a Friday, little maintenance time is reguired over the weekend. More intense effort is required as the test progresses. Historically, Daphnia magna has been the most used species for the estimation of zooplankton acute sensitivity to pollutants (Mount and Norberg 1984, Buikema et al. 1982, Anderson 1980). Anderson (1980) described a series of 17 papers produced by Einar Naumann in 1933 and 1934 detailing various aspects of toxicity testing with D. magna. This was perhaps the real beginning of the use of D. magna in such tests. with movement toward the use of chronic test methods in aquatic toxicology, both a lifetime (Buikema 1973, Winner and Farrell 1976) and 21-day chronic test (Biesinger and Christensen 1972) were developed for D. magna . As designed, these methods provided estimates of sublethal effects based on changes in fecundity (Buikema et al. 1980), but they were still too lengthy. Under the auspices of the EPA, Mount and Norberg (1985) developed a 7-d subchronic toxicity test with a different species, Ceriodaphnia reticulata. C. reticulata was chosen (over D. magna) because it is widely distributed in North America, it was easier to culture than was D. magna and it produces three broods of young in seven days (Mount and Norberg 1984) . These characteristics facilitate the performance of many tests in a short time, virtually anywhere. Since the developmental work by Mount and Norberg (1984), EPA has suggested the use of C. dubia in their protocol manual (Horning and Weber 1985) . A Ceriodaphnia dubia survival and reproduction test is begun with the collection of neonates (<24 h old) from the adult culture. Neonates are placed in individual test vessels consisting of 30 ml plastic cups containing 15 ml of solution. Isolation of individuals is necessary so that separate tallies of fecundity can be maintained for each animal. A daily count of survivors is made. Beginning on Day 3 or 4 of the test when the first brood 10 is produced, offspring are also counted. Adults are then moved to new test vessels and fed. This daily counting and transfer to new solutions continues until the test is terminated at seven days. A 7-d LC50 for adults is calculated, and their fecundity is used to measure sublethal effects. Rationale For The Use of Freshwater Molluscs In Acute Toxicity Tests To date, over 165 species of aquatic organisms have been used in acute toxicity tests (Buikema et al . 1982). Species specific tests are necessary because of differences in biological sensitivity and because organisms from different habitats may be exposed to a wide range of pollutant concentrations. The need for a variety of test methods has been demonstrated by results of testing with many pollutants (Johnson and Finley 1980, Mayer and Ellersieck 1986) . Most of the accepted test organisms are pelagic species (fish and zooplankton) which are well-known, easy to culture, economically important or of public concern. Little attention has been given the response of benthic macroinvertebrates, other than insects, to pollutants. Benthic invertebrates are more appropriate test organisms for flowing waters than are zooplankton because the latter are not typically found in such systems, and are therefore not good indicators of the impact of pollutants on lotic invertebrates. However, toxicity data for 11 benthic organisms, including aquatic oligochaetes, turbellarians, pelecypods and gastropods, are still extremely limited, and only one toxicity test protocol for these organisms exists, i.e., for the exotic clam Corbicula fluminea (Foster 1979) . Non-insect benthos have been considered either unimportant or their responses have been extrapolated from those of the common test organisms. However, use of zooplankton or fish as surrogates for non-insect benthic fauna is questionable. Not only are conditions at the water-sediment interface different than those in the open water and difficult to assess with pelagic organisms, there ought to be specific information on the response of benthic organisms to pollutants since they represent distinctly different taxa (Buikema et al. 1982). One of the most widely distributed groups of macrobenthos native to streams in Florida is the unionid mussels. Little is known about their sensitivities to various pollutants entering their environments . Distribution and Life History of Unionid Mussels The vast majority (36 genera and 250 species) of bivalve mollusc species in North American continental waters belong to the family Unionidae (Burch 1973) . The group as a whole is endemic to North America, but many species have limited ranges. Unionid mussels generally prefer lotic habitats with stable substrates and some 12 silt. They are distributed from southern Ontario to Florida and west to Washington and Oregon. However, the best studied and perhaps richest mussel (or clam) fauna is found in the eastern United States between the Appalachian Mountains and Mississippi River. Several unique life history features were key to the evolutionary development of freshwater mussels from their marine ancestors (Stein 1971) . These included the production of a parasitic glochidia larva rather than the free-living veliger, incubation of the larvae in the marsupia (gills) of the female and the requirement for a fish host during the 9-30 day parasitic phase. Reproduction begins when male mussels shed sperm into the water. Sperm cells are drawn into the incurrent siphon of the female mussel and become lodged in the gills on each side of her body that are specifically modified for incubation. Development to the bivalved glochidia occurs inside the female. During this time, the larvae may become infected by any number of bacterial, fungal, protozoan or water mite species, which may reside in the mussel permanently or temporarily. When mature, glochidia are shed into the water via the excurrent siphon either directly onto fish hosts whose presence stimulates release or randomly broadcast into water currents (Buchanan 1980, Parmalee 1967). Complete development into juveniles requires a period of parasitism on fish during which the organ systems 13 develop. Distribution of unionid mussels is facilitated by their attachment to mobile hosts instead of by the production of mobile larvae as in marine bivalves (Fuller 1974) . After encystment periods of varying times, glochidia drop off of their hosts and become free-living filter feeders (Arey 1932) . Habitat Destruction and Faunal Decline Mussel fishing was a thriving business in the Illinois, Tennessee and Mississippi Rivers from the late 1800's to the mid-1960's (Isom 1969, van der Schalie and van der Schalie 1950, Starrett 1971) . They were a source of freshwater pearls for jewelry, and their shells supplied the button industry with raw materials. Later, shell slugs were used as seeds in the Japanese cultured pearl industry (Parmalee 1967). Ten thousand tons of mussel shells per year were harvested from the Tennessee River in the 1940 's and 1950's, but harvests declined steadily during ensuing years (Starrett 1971, Parmalee 1967) . Similar changes in abundance were observed in other eastern rivers during the same period (van der Schalie and van der Schalie 1971, Isom 1969, Starrett 1971). Several factors have been suggested as causes for the decline including overharvesting, habitat destruction by damming and pollution. 14 Overharvesting of mussels in the Tennessee, Illinois and Mississippi Rivers may have contributed significantly to the decreased abundance of some species (Starrett 1971, Forbes and Richardson 1919, Danglade 1914). In 1910, there were over 2,600 boats engaged in mussel fishing along the lower half of the Illinois River alone (Starrett 1971) . Similar intensive harvests were made by the well-developed mussel industry of the Tennessee and Mississippi Rivers, Not only did such harvesting reduce populations directly, but it also may have reduced breeding stock below the replacement capacity of remaining stock, destroyed stream habitat and resulted in the death of disturbed but uncollected animals (Fuller 1974). While the button industry switched from pearl to plastic in the 1930's and 1940's, another use for mussel shells was found. Spheres of mussel nacre (pearlized shell) were used as nuclei by the Japanese cultured pearl industry beginning in the late 1950 's. The use of SCUBA gear permitted the harvest of whole beds of mussels leading to localized extinction (Fuller 1974) . A second major contributor to the declining mussel fauna was the extensive habitat destruction resulting from damming activities of the Tennessee Valley Authority (TVA) beginning in the 1930's (Isom 1969) . While adult mussels of some species prefer quiet water (Wilson and Clark 1912, Danglade 1914), juveniles and adults of many species need riffle water. Thus, damming may have 15 provided more habitat for some species, but reduced the area suitable for others. Decreased water flow also hinders reproduction by limiting dispersal of sperm and later, glochidia. Further, impoundment changes fish species distributions which may affect mussel recruitment since they are briefly parasitic on fish (Fuller 1974) . Other consequences of damming include increased siltation which can lead to suffocation of mussels, as well as general loss of habitat and decreased recruitment success due to release of tailwaters into otherwise suitable stream reaches. The latter results from both low temperatures and low oxygen levels of tailwaters (Fuller 1974, Marking and Bills 1980, Ellis 1936). Finally, the effects of water pollution by human waste, industry and agricultural activities have damaged the mussel fauna in many areas. Pulp and paper mills which release sawdust and process effluents destroyed mussel populations in Minnesota (Danglade 1974) , the upper Tennessee River drainage (Ortman 1918) , panhandle Florida (Heard 1970) and in areas around Ottawa, Canada (Mackie and Qadi 1973) , As noted earlier, siltation is a problem for mussels and with its increase along with agricultural activities, the molluscs declined steadily. The effects on mussel populations of pesticides and herbicides used in farming and aquatic weed control, and metals released in acid mine drainage and industrial effluents have only recently been examined. Little 16 conclusive evidence is available since the specific sensitivity of mussels to such pollutants is difficult to determine from field data, and laboratory exposures have been limited to a few studies of adults (Imlay 1971, Imlay 1974, Foster and Bates 1978). Development of better culture techniques and toxicity test procedures have begun to make experimental work possible. Pr-opaqation Of Freshwater Mu'^-^pI s In Artificial Media The unusual mode of reproduction of unionid molluscs makes their culture in the lab more difficult than it is for other molluscs that have free-living veliger larvae. The life cycle of unionid mussels includes a parasitic larva (glochidia) which normally attaches to fish gills or fins during early development. This stage must have fish to parasitize or a culture medium that would provide the necessary nutrients. Earliest efforts to propagate freshwater mussels (LeFevre and Curtis 1912) in fish plasma were unsuccessful. Glochidia did not transform. By 1926, transformation of glochidia to juveniles had been accomplished with the use of an artificial medium (Ellis and Ellis 1926). However, glochidia were permitted to encyst on fish gills for 18-96 hours before being dissected out to incubate through transformation. The contents of their growth medium were described as including NaCl, KCl, CaCl2, NaHC03 , dextrose, a mixture 17 of amino acids, small quantities of phosphates and traces of magnesium salts (Ellis and Ellis 1926) . Research into mussel propagation lost its impetus as the button and mussel-fishing industries dwindled in the 1940's and 1950's. However, in the hope of replenishing declining natural populations, the Tennessee Valley Authority funded research to develop methods for in vitro propagation of these freshwater molluscs in the early 1980's (Isom and Hudson 1982, Hudson and Isom 1984). The goal was to develop a complete culture medium that eliminated the need for fish hosts during the larval stage and to produce a large number of juvenile mussels at one time. As a result, a culture medium containing a Ringers solution, vitamins, glucose, amino acids, antibiotics and fish plasma was developed as a substitute for live fish, (Isom and Hudson 1982). Glochidia were removed from the gills of ripe female mussels, rinsed several times in sterile water and put in the medium. Culture dishes were then placed in a temperature controlled CO2 incubator (23° ± 3° C) . The transformation of glochidia to juveniles takes 9-30 days (23° ± 3° C) depending on the species, culture temperature and degree of glochidia maturity at the start of incubation. While the Hudson and Isom method (1982, 1984) is far better than that of Ellis and Ellis (1926) which relied on the use of fish hosts for encystment of glochidia during transformation, further simplification 18 is desirable. The old method (Hudson and Isom 1982, 1984) is laborious, still requires the use of fish plasma which may not be readily available nor of consistent quality, and a CO2 incubator. A simplified method for mussel culture is necessary before they can be available in the numbers needed for replenishment of declining wild stocks or for other purposes, e.g. toxicity tests. Sensitivity of Unionid Molluscs to Environmental Pollutants Interest in the effects of toxic pollutants on mussels directly and their use as environmental indicators in general has increased following the decrease in population sizes. Freshwater mussels have been suggested for use as biological monitors in lotic environments for many years. Biomonitors are important because analysis of water does not reflect biologically available concentrations of toxicants (Leard et al. 1980) . Sometimes biota can be affected by concentrations below the detection limit of analytical instruments, and at other times, high ambient concentrations are benign because they are refractory or are adsorbed to particulate matter. The utility of freshwater mussels as biomonitors is enhanced by their sedentary lifestyle, their long lifespan compared to other invertebrate species and the fact that they live in the sediments while being filter-feeders. Thus, mussels are exposed to dissolved, particulate and sediment-sorbed contaminants 19 (Havlik and Marking 1987). However, current information on mussel sensitivities consists largely of species presence-absence data for locations impacted by toxics, and measurements of contaminants in shells or tissues (Havlik and Marking 1987). Experimental data on specific sensitivities are very limited. Simmons and Reed (1973) used the presence and abundance of freshwater molluscs as indicators of biological recovery in the North Anna River, Virginia. The North Anna River was receiving acid mine drainage from a defunct coal mine. While aquatic insects re- established quickly below the confluence of the river and an unpolluted creek, molluscan species were absent for another 50 miles downstream. In this example, molluscs were more sensitive indicators of biological recovery than were insects, traditionally regarded as good biomonitors. Simmons and Reed (1973), however, suggested that the lack of mussel fauna in the acidified river reach may have been caused by siltation and loss of host fish for the glochidia. Three species of mussels from the Illinois River near Peoria, Fusconaia flava, Amblema plicata and ouadrula auadrula. were analyzed for metals along with fish, tubificid worms, river sediment and water (Mathis and Cummings 1973). The portion of the river studied was highly industrialized and therefore received metallic effluents. The goal was to determine if there had been a 20 loss of species and if the loss could be related to metal concentrations. Mussels accumulated the metals to levels exceeding dissolved concentrations by 1-2 orders of magnitude, but had lower levels than those in the sediments. No attempt was made to determine species abundance, but the presence of a thriving mussel-fishing industry within the study area indicated that these three mussels were plentiful. Another species, Musculium transversum, the fingernail clam, was absent from polluted areas of the river where it had been common before 1954. While concentrations of metals in the tested adult clams may not have been lethal, this does not mean there was no impact on juveniles or on other species such as M. transversum which were absent. Anderson (1977) analyzed shells and tissues of freshwater clams from the Fox River, Wisconsin for cadmium, copper, lead and zinc. He found that body burdens generally paralleled sediment concentrations while being much higher than were found in water. Metal concentrations were much lower in the shells (Cd 16 cm). This is 1.5-4 times the normal stream water concentration in treated areas and similar to the sensitivities of the crayfish Orconectes (17.8 mg/L) and Gammarus pseudolimnaeus (22,3 mg/L) (Johnson and Finley 1980) . The lampricide Bayer 73 was lethal to 50 per cent of adult Elliptic dilatatus at 382 ug/L concentration (Rye and King 1976) . The LC50 for rotenone was 2.7 mg/L in soft water for adult Lampsilis sp. (Farringer 1972) . The response of adductor muscle activity in glochidia of Anodonta cygnea was used as a measure of toxicity by Varanka (1977, 1978). Spontaneous adductor muscle activity is crucial if glochidia are to attach to their hosts. In a series of experiments, Varanka applied the muscle-contraction inducer tryptamine to glochidia and recorded baseline contraction rates. He then exposed larvae to tryptamine + pesticide to determine the effect of the pesticide on this activity. Results of 30-minute exposures to malat.iion, 2,4-D and Shell-DD indicate that the EC50S (effective concentration for 50% of the sample) based on adductor muscle activity were considerably 28 higher than the usual environmental levels of these pesticides. However, the fact that muscle activity responded to such short exposures suggests that further study with longer exposure times may be worthwhile. Though in these and other studies the mussel bioaccumulation capacity for metals and pesticides is great, we have very few measures of the lethal or sub- lethal effects of such exposures. With such a limited database, it is premature to draw conclusions about the sensitivity of mussels to environmental exposures to metals or pesticides. The majority of studies to date consist of species surveys and measurements of tissue or shell toxicant levels. How the presence of these contaminants may affect mussel survival, growth or reproduction remains to be determined. Laboratory exposures of mussels of various species and age groups to pesticides, metals or organic pollutants would provide extremely valuable measures of direct effects. Such data would allow us to separate the effects of habitat destruction, siltation and competition with Corbicula sp. from response to pollutants. CHAPTER 3 THE SENSITIVITY OF THE FATHEAD MINNOW fPIMEPHALES PROMELAS) TO HYDROTHOL-191 AT 15° AND 25° C Introduction The United States Environmental Protection Agency (EPA) has recently advocated the use of short-term chronic toxicity tests for biological monitoring of water and wastewater (Horning and Weber 1985) , Such tests were developed by EPA on the basis of their cost- effectiveness, rapidity and requirement for low sample volumes over the course of the test. These features allow the methods to be implemented in on-site effluent toxicity evaluations, as well as in toxicity assessments of pure compounds and samples shipped to central laboratories. Implementation of short-term chronic toxicity methods was justified by evidence that these early life cycle tests reasonably approximated more complete, full life cycle chronic toxicity tests that had been the standard for many years (Norberg and Mount 1985) . One of these tests employs fathead minnow larvae (Pimephales promel.'.s) in a seven-day, static renewal, survival and growth test. Test results are based on the 29 30 survival and growth (weight gain) of larval fathead minnows over seven days in the presence of a range of toxicant concentrations. it is a new evaluative method for which few test results are available. Norberg and Mount (1985) determined the chronic toxicity of several industrial effluents and receiving waters, as well as zinc, copper and Dursban during their test development work. The 7-d test with fathead minnow larvae gave results similar to those from much longer (3-6 months) early life stage (ELS) tests. The current study was designed to evaluate the toxicity of the herbicide Hydrothol-191 to fathead minnows. Hydrothol (endothall) acts to decrease photosynthesis and cellular respiration in turfgrass and to decrease the production of amylase in germinating barley seeds (Ashton and Crafts 1981) . it has a half- life of 10 days, biodegradation by bacteria being the primary fate process leading to the decline of Hydrothol concentrations (Reinert and Rodgers 1987) . its effects, similar to those of Actinomycin D, are not reversible with benzyladenine (Penner and Ashton 1968) . Since actinomycin D selectively inhibits the synthesis of m- RNA, the translator of DNA messages during protein (enzymes) production, it is hypothesized that Hydrothol also interferes with m-RNA production (Ashton and Crafts 1981) . 31 Support for this assertion was given by the effect of Aquathol-K (dipotassium endothall, Pennwalt Corp.) on the smoltif ication of juvenile Chinook salmon ( Oncorhynchus tshawytscha) . When juvenile salmon were exposed to water with endothall levels as low as 3 mg/L for four or fourteen days prior to their transfer from freshwater to artificial seawater, they were unable to survive (Liguori et al. 1983). Fish that were allowed to recover in freshwater for 10 days prior to their transfer to seawater survived well. Histopathologic analyses indicated that hypertrophy of branchial epithelium occurred in fish exposed to 10 mg/L or more of endothall. The process of smoltif ication involves changes in plasma levels of thyroxine, triiodothryonine, and gill ATPase activity. Both triiodothryonine and ATPase are found in gill tissues. Liguori et al . (1983) suggested that in endothall damaged gill tissue, the levels of triiodothryonine and ATPase may be depressed, although neither was measured in their study. Since Hydrothol interferes with smoltif ication possibly related to levels of ATPAse, it may be the result of its inhibitory effects on m-RNA synthesis (Liguori et al. 1983). Therefore, the impact of Hydrothol on organisms should be temperature dependent as are most chemical reactions (Wilson 1972) . In Florida, Hydrothol is registered for use in the control of algae. Hydrilla verticillata , Myriophyllum spicatum. Utricularia spp. , Valisneria spp. and other 32 submersed macrophytes, as well as for several agricultural purposes (Dupes and Mahler 1982, Blackburn and Weldon 1963). Although there are some data on the acute toxicity of Hydrothol to non-target aquatic organisms, only limited data are available on its chronic toxicity to aquatic biota. The Florida Department of Environmental Regulation (FDER) , seeing the likelihood of increased Hydrothol use because of its efficacy in the long-term control of aquatic macrophytes, requested the evaluation of its potential impact on non-' target organisms. FDER was also interested in knowing what if any effect water temperature might have on the toxicity of Hydrothol to freshwater fish. There were almost no data indicating the effect of temperature on Hydrothol toxicity (Walker 1963). since many physiological processes are affected by temperature because of the pivotal role of enzymes in cellular respiration and the synthesis or degradation of organic compounds (Wilson 1972), it was important to determine whether water temperature would alter the toxicity of Hydrothol to organisms. The goals of this investigation were: (i) to determine the chronic toxicity of Hydrothol to the fathead minnow and (2) evaluate the impact of temperature on its toxicity. These data would be useful 33 in the development of better guidelines for the use of this herbicide in Florida. Materials and Methods Test Organism Fathead minnows (Pimephales promelas) were acquired from the EPA-Newton Laboratory in Cincinnati, Ohio. Fish embryos were sent in insulated containers by express mail and hatched in transit. Newly hatched fathead minnow larvae preferably less than 24 hours old were used to initiate a test. Test Organism Food Fathead minnow larvae were fed live brine shrimp nauplii (Artemia salina) raised from eggs in the laboratory. Brine shrimp nauplii were incubated at 25° C and harvested when nauplii were less than 24 h old (Peltier and Weber 1985, Horning and Weber 1985) . Fathead minnow larvae (10/500 ml test vessel) were fed 0.1 ml harvested brine shrimp (approximately 1000 nauplii) three times daily at 4-hour intervals. Dilution Water Moderately hard reconstituted freshwater was used as diluent throughout the test. It was prepared by adding the following constituents to 1 1 of deionized 34 water: 96 mg NaHC03, 60 mg CaSO^ * 2H2O, 60 mg MgSO^ and 8.0 mg KCl. This produced water with a pH of 7.4-7.8, a hardness of 80-100 mg/L CaC03 and an alkalinity of 60-70 mg/L CaC03 (Horning and Weber 1986) . Dilution water was made in bulk and stored in a plastic carboy at 15° or 25° C for the duration of each seven day test. Test Chemical The toxicity of Hydrothol-191 (Pennwalt Corp., Philadelphia, PA) , an agricultural and aquatic herbicide was evaluated in this study. Hydrothol, the trade name for the alkylamine salt formulation of endothall (7- oxabicyclo [2,2,1] heptane-2 , 3-dicarboxylic acid), is used extensively to control Hydrilla verticillata and Myriophyllum spicatum. Test solutions were made fresh daily by dilution of a stock solution of Hydrothol with moderately hard reconstituted freshwater (v/v) . All Hydrothol concentrations given are nominal. Reference Toxicant Cadmium chloride, obtained from EPA (Quality Assurance Branch, EMSL, United States Environmental Protection Agency, Cincinnati, OH 45268) was the reference toxicant for fathead minnow tests. Three tests were performed with CdCl2 to evaluate the consistency of test results. Later, CdCl2 was used as a check on test organism quality in each definitive test with Hydrothol. 35 Range Finding Test In order to obtain the appropriate range of Hydrothol concentrations to be used in 7-day tests an acute toxicity range finding test was conducted. It was determined that 0% larval mortality occurred at 100 ug/L and that 100% mortality occurred at 1000 ug/L. Consequently, the definitive 7-day Hydrothol concentration range was 50-1000 ug/L. Seven-day Survival and Growth Toxicity Test Toxicity tests were initiated by placing larvae in 1 liter borosilicate beakers (test chambers) containing 500 ml control or test water. Larvae were transferred into duplicate test chambers by a large-bore pipet until each test chamber contained 10 larvae, for a total of 20 larvae at each Hydrothol test concentration. Definitive tests were conducted at 15° C and 25° C in constant temperature rooms, with a photoperiod of 16 hours light and 8 hours of darkness. The test chambers were randomized at the beginning of the test. Ninety per cent of the test solution was renewed every day after a large-bore pipet was used to siphon dead brine shrimp and other debris from the bottom of the test chambers. Chemical and physical analyses of test water were conducted according to standard EPA methods (EPA 1983) . Dissolved oxygen, temperature, pH, conductivity, alkalinity and hardness were measured at the beginning of 36 each 24 hour exposure at all test concentrations and in the control . The numbers of live and dead larvae in each test chamber were recorded daily, and the dead larvae were removed. After seven days of exposure the test was terminated. The surviving larvae were removed and preserved as a group in 4% formalin. At a later date, the groups of preserved larvae were rinsed in distilled water and dried at 105° c for a minimum of 2 hours. Dry weights of each group of larvae were measured to the nearest 0.001 g. Statistical Analysjc; All LC50 and 95% confidence intervals were calculated using the TOX-Dat multimethod computer program (Peltier and Weber 1985, Horning and Weber 1985). survival data were arcsine-transformed and analyzed by Dunnett's Procedure which includes an analysis of variance (ANOVA) , followed by a comparison of each toxicant concentration with the control. From this analysis a No Observed Effect Concentration (NOEC) and a Lowest Observable Effect Concentration (LOEC) were calculated. In addition, the Chronic Value (ChV) was determined by calculating the geometric mean of the NOEC and LOEC. Growth (dry weight) data were also analyzed by Dunnett's Procedure r F = 0.9545 Pr > F = 0.0006 ■^Not significant at p<0.05. ^Significant at p<0.05. 49 Discussion The results of several acute toxicity studies (Table 3-9) with fish species using Hydrothol are available for comparison with LC50 values calculated from the current study. Johnson and Finley (1980) determined the 96-h LC50 of 0.75 mg/L for fathead minnows at 18° C. This value is close to my 96-h LC50 for the same species. Rainbow trout (Salmo gairdneri) and golden shiner (Notemigonus crysoleucas) were much less sensitive to Hydrothol as evidenced by their four day LC50s of 1.7 mg/L and 1.6 mg/L, respectively (Mudge et al. 1986, Finlayson 1980) . Other workers reported a Hydrothol no mortality range of from 3.0-55.0 mg/L (Holmberg and Lee 1976, Liguori et al. 1983, Berry 1984). Since Hydrothol- 191 is applied at a concentration of 1-5 ppmw (part per million water) and has a half-life of 10 d (Blackburn et al. 1971, Reinert et al. 1985), it poses a potential threat to fish. Its concentration can remain above the 96-h LC50 for 10-20 days. Only limited chronic toxicity data are available from other studies with endothall products (e.g. Hydrothol and Aquathol-K) (Liguori et al. 1983, Eller 1973). In a study of the impact of 10 mg/L Hydrothol on juvenile Chinook salmon (Oncorhynchus tshawvtscha) , Liguori et al. (1983) observed marked changes in branchial tissues. Effects included epithelial 50 Table 3-9. Comparison of literature LC50 values for Hydrothol to various freshwater fish. Test Organism Stage or Temp 96-h Reference Wet Wt. (C) LC50 (g) (mg/L) 0.62 18 1.6 1 25 15 1.7 2 1.2 13 0.56 3 - - 1.3 4 1.0 10 0.18 4 0.3 18 0.49 4 0.5 24 0.94 4 - - 1.2 4 0.6 18 0.75 3 larvae 25 0.39 5 larvae 15 0.47 5 Golden shiner^ Rainbow trout Rainbow trout Rainbow trout Cutthroat trout Channel catfish Bluegill Bluegill Fathead Minnow Fathead minnow Fathead minnow -^Finlayson (1980). ''Mudge et al. (1983) Finley (1980). "^Pennwalt Corp. (1980). Johnson and ^Current study. 51 hyperplasia and lamellar fusion. At lower exposure levels (<10 mg/L) , no histopathologic effects were detected. The 14-d LC50 was 62.5 mg/L of endothall. The most sensitive measure of Aquathol-K toxicity was the seawater test (Liguori et al. 1983), In this experiment, survival of juvenile Chinook salmon after placement in seawater was measured following their exposure to endothall concentrations of 10.1-105,7 mg/L for 14 days. Transfer to seawater simulated the migration of this species to the ocean during the smoltif ication process, a critical stage in their development. All fish died within three days of entry into seawater. Even at sublethal concentrations endothall exerts an effect that impairs important physiological processes related to osmoregulation (Liguori et L. 1983) . Eller (1973) followed histopathological changes in bluegill exposed to Hydrothol-191 for up to 112 days. He found significant but transitory changes in gill epithelium in fish exposed to 0.3 mg/L of the herbicide. Epithelial hyperplasia and lamellar fusion were noted in bluegill during the first 14 days of exposure. After that time, gill damage gradually reversed and gills were normal by the end of the study. Some abnormalities were noted in hepatic and testicular cells, but they were not conclusively related to Hydrothol concentration (Eller 1973) . 52 The limited data relating test temperature to the toxicity of Hydrothol to fish comes from a study by walker (1963). He found that 96-h LCSOs for bluegill sunfish (0.05 mg/L), redear sunfish (O.IO mg/L) , largemouth bass (0.14 mg/L) and yellow bullhead (0.31 mg/L) at 23.8° c were reduced by 43%, 44%, 36% and 13%, respectively, at 18.3° C. Reductions in Hydrothol toxicity due to test temperature were greater for three of these four species than was measured for the fathead minnow in the current study. The only exception was the yellow bullhead. Interspecific differences in sensitivity to herbicides is seen throughout the literature (Johnson and Finley 1980) . The greater effect of test temperature on Hydrothol toxicity to species tested by Walker (1963) than for the fathead minnow may be due simply to species differences. No other explanations are readily apparent. Hydrothol is relatively toxic to fish in comparison with other herbicides such as 2,4-D, dichlobenil, diquat, and PCP (Table 3-io) that may enter the aquatic environment. 2,4-D is applied to ponds and lakes for control of water hyacinth (Ag Consultant 1988) and to agricultural fields for control of broadleaf weeds (Ware 1978) . Its mode of action via a complex mixture of effects on cell division and nucleic acid metabolism is somewhat different than the impairment of m-RNA production caused by Hydrothol (Ware 1978, Ashton and 53 Table 3-10. Summary of acute toxicity data for selected herbicides that may enter the aquatic environment in Florida^. Herbicide and Animal Temp. 96-h LC50 test organism wt. (g) (°C) (mg/L) 2,4-D fathead minnow 0.5 17 18.0 bluegill 1.4 17 7.5 Dichlobenil fathead minnow 0.8 18 6.0 bluegill 1.5 18 8.3 Diquat bluegill 1.3 12 245 PCP fathead minnow 1.1 20 0.21 bluegill 0.4 15 0.03 channel catfish 0.3 20 0.07 Aquathol bluegill 1.3 22 343 channel catfish 0.4 12 >150 Hydrothol fathead minnow 0.6 18 0.75 fathead minnow^ 0.4 15 0.39 fathead minnow*-* 0.4 25 1^ 0.47 ^ata from Johnson and Finley (1980) current study. ^Data from the 54 crafts 1981). The 96-h LC50 for 2,4-D was 18.0 mg/L for fathead minnow and 7 . 5 mg/L for bluegill (Johnson and Finley 1980) . These levels are far below both the application rates for aquatic environments or the expected levels entering water from treated agricultural areas (Ag Consultant 1988) . Dichlobenil, an inhibitor of CO2 fixation and oxidative phosphorylation (Ware 1978), is used to eliminate Chara, Potamogeton spE- and mriojmU^ ^^ in marshes. Its 96-h LC50 for fathead minnows is 6.0 .g/L and 8.3 mg/L for bluegill (Johnson and Finley 1980). in normal use, dichlobenil is not toxic to fish in treated areas (Ag Consultant 1988). Diquat is the most widely used herbicide for control of broadleaf weeds along ditchbanks and irrigation canals (Gangstad 1986). In lakes and slow-moving waters, Diquat use controls coontail ( Ceratophyllum demersum) , bladderwort (Ut^-i-i^ ^•) -^ ^^^^^^^ iE^t^m^t^ SEE.). It is a contact herbicide that reduces photosynthetic activity (Ware 1978). Treatment of canal banks with 2,4-D at recommended doses results in a water concentration of only 0.025-0.061 mg/L (Gangstad 1986), far below the 96-h LC50 of 245 mg/L for bluegill. PCP (pentachlorophenol), on the other hand, is not used much anymore primarily because of its extreme toxicity to biota (Ware 1978). PCP is a non-selective herbicide and preharvest defoliant. It has multiple 55 routes of action including plasmolysis and protein precipitation and is destructive to all cells (Ware 1978) . Its toxicity to fish is evident from the 96-h LC50 of 0.21 mg/L for fathead minnow, 0.03 mg/L for bluegill and 0.07 mg/L for channel catfish (Johnson and Finley 1980) . Results of the current study using fathead minnow larvae indicate that the use of Hydrothol in the aquatic environment should be limited to properly trained professionals. It is a highly toxic herbicide for which there are a number of substitutes. The effect of temperature on the toxicity of Hydrothol can be substantial. It should be applied at the lowest water temperature at which it will control the particular macrophyte of interest. This temperature will vary based on the species of plant because Hydrothol is most effective when applied early in the growing season (Ag Consultant 1988) . CHAPTER 4 AN ASSESSMENT OF THE CHRONIC TOXICITY OF HYDROTHOL-191 TO THE ZOOPLANKTER CERIODAPHNIA DUBIA USING A 7-DAY SURVIVAL AND REPRODUCTION TEST Introduction Since zooplankton are an extremely important part of most aquatic ecosystems and contribute substantially to the food supply of fish (Horning and Weber 1985, Mount and Norberg 1984), these organisms have been used extensively in toxicity tests. Early studies used either Daphnia magna or Daphnia pulex as test organisms. However, each of these species had their shortcomings. D. magna has a limited distribution in aquatic systems and neither animal is easy to culture in the laboratory. Ceriodaphnia was chosen for use in a new subchronic toxicity test for several reasons (Horning and Weber 1985, Mount and Norberg 1984) . Ceriodaphnia reproduce more rapidly (3 broods in a week) than Daphnia, are ubiquitous, and are somewhat easier to culture under laboratory conditions (Horning and Weber 1985) . The static renewal Ceriodaphnia dubia survival and reproduction test (Horning and Weber 1985) was developed as a substitute for the 21- to 28-day Daphnia chronic toxicity test. Toxicity is based on survival and reproduction over a 7-day period in the newer 56 57 test. Thus, the toxic effects of chronic exposure to a substance may be more easily and rapidly assessed than methods using D. magna. The Florida Department of Environmental Regulation (FDER) requested the determination of the chronic effects of Hydrothol-191 on C. dubia using this new test method. Their concern stemmed from the growing use of Hydrothol in Florida aquatic systems for control of several species of macrophytes. Specifically, too little was known about its long-term impacts on non-target organisms. Since this herbicide has a half-life of 10 days (Reinert and Rodgers 1987), it can remain at potentially toxic levels in the environment for 10 days or more. During that time, zooplankton biomass could be seriously lowered if Hydrothol affected both survival of adults and their reproductive capacity. m that case, their role as a food source for fish would be impaired. The C. dubia survival and reproduction test was designed to measure the effects of toxicants on survival of adults and production of young (Mount and Norberg 1984, Horning and Weber 1985) . Tests were performed at 15° and 25° C to see if temperature at the time of Hydrothol application would affect its impact on zooplankton. If so, field use could be limited to times when water temperature and plant growth activities were compatible. 58 Materials and Methods Test Organism Ceriodaphnia dubia stock obtained from EPA-Newtown, Ohio was used to start a laboratory culture. The animals were maintained in 1 L beakers in a 25° C environmental chamber, with 16 hours of light and 8 hours of dark. Test Chemical Several formulations of endothall (7-oxabicyclo [2,2,1] heptane-2 , 3-dicarboxylic acid) are used in Florida for control of aquatic weeds and algae. However, the chronic toxicity of the alkylamine form of endothall, i.e. Hydrothol-191 (Pennwalt Corp., Philadelphia, PA), was assessed in this project based on the response of Ceriodaphnia dubia during a 7-day test. Hydrothol concentrations were not measured, but were calculated based on volume/volume dilutions of the 53% active ingredient (the alkylamine salt of endothall) indicated on the product label. Dilution Water Moderately hard reconstituted freshwater (Horning and Weber 1985) inoculated with bacteria-rich aerobically digested trout chow and aged for one week, was used as the culture medium (Table 4-1) , The addition of bacteria and aging of the dilution water has been suggested (De Graeve and Cooney 1987, FDER 1986, Mount and Norberg 1984) to stabilize water quality and increase ambient food levels. 59 Table 4-1. Dilution water quality parameters for Ceriodaphnia dubia Survival and Reproduction Tests, Parameter Mean S. D. pH 6.84 0.08 Alkalinity (as mg/L CaC03) 53.21 1.55 Hardness (as mg/L CaC03) 86.83 1.56 Conductivity (umhos/cm) 349.3 3.4 60 Bacteria are a major food source for Ceriodaphnia (Norberg and Mount 1985) . Thus, while cultures were fed daily, the presence of a high background bacterial population assured that food density was adequate to support high reproduction. Aeration was provided by a small air pump set at minimum output to prevent oxygen depletion by bacterial respiration. Test Organism Food Ceriodaphnia were fed a mixture of digested trout chow, Cerophyll, and yeast (Horning and Weber 1985) provided at a rate of 3 ml/L of water per day. Most cultures developed a lush algal growth which was allowed to remain even though water in the culture chambers was replaced weekly. The algae provided an extra food source. Reference Toxicant Tests At least once a month, a reference toxicant test using sodium dodecyl sulfate (SDS) was performed to verify that the in-house Ceriodaphnia cultures were healthy and nominally sensitive. That is, LC50s for SDS were compared to those in the literature to ensure that their responses to the test chemical were not due to an inherent sensitivity. The SDS was obtained from EPA-Cincinnati specifically for use as a reference toxicant. Several toxicity tests were also performed using CuSO^ . The results of tests with CUSO2 proved to provide more consistent results. 61 Range-Finding Test. A 48-hour range finding test was performed at the two test temperatures (15° and 25° C) before definitive testing began. Hydrothol concentrations ranged from 100-3200 ug/L based on percent active ingredient (ai) as indicated on the product label. Dilution and control water were moderately hard reconstituted freshwater "conditioned" with a bacterial inoculum and aerated for a week. Preparation Fnr chronir Toxicity T^c^-t-e Approximately one week prior to the start of a test, 20 brood animals were obtained as neonates and placed in separate 30 ml plastic cups containing 15 ml of culture medium. They were fed 0.2 ml of the TCY mixture and 0.2 ml Of an algal mixed culture (Chlamydomonas, Klebsomidium and Euglena ) each day. Algal supplements have been suggested for use in C^ dubia toxicity tests to promote high fecundity (Cowgill et al. 1985). Water was changed every other day. Neonates to be used as test organisms were harvested from these brood chambers during a 4- hour period on about the seventh day. They were held 12-24 hours prior to the start of each test. Chronic Toxicity Tests Toxicity tests were performed at 15° and 25° C in a constant temperature room with a 16L:8D lighting regime. To begin each test, five toxicant solutions were prepared from 62 a concentrated stock diluted with moderately hard reconstituted freshwater. Hydrothol concentrations of 2 5 ug/L to 400 ug/L were used. Test chambers were filled with 15 ml of toxicant or control water and the neonates were randomly distributed among them, one to each chamber. Each treatment consisted of 10 replicate chambers placed in a plywood rack. Test solutions were prepared and renewed daily. The presence and number of young were recorded for each chamber daily before transferring the adult organisms to fresh test solutions. Ceriodaphnia were fed 0.2 ml of TCY and 0.2 ml of algal culture following transfer to clean vessels. Temperature, pH, alkalinity, hardness and conductivity of the dilution water were measured each day. Since the dilution water was saturated with oxygen by aeration, dissolved oxygen measurements were not made. Each test was terminated after 7 days, and the mean young production per adult was calculated for each treatment and the control. Statistical Analysis Statistical analysis followed standard EPA protocol based on the original number of adult animals used per test chamber, i.e. if one died, it was still included in the calculation of mean brood number and size (Horning and Weber 1985) . LC50S for Hydrothol were calculated with the TOX-DAT Multi-method (Peltier and Weber 1985) . This series of computer programs calculates the LC50 and 95% confidence 63 intervals by 3 methods: moving-average angle, binomial and probit. Fisher's Exact Test was used to identify treatments in Which adult survival was significantly different from the control. No further analysis was performed on such treatments. However, reproduction data were analyzed for toxicant levels in which adult survival was not significantly different from the control using ANOVA and Dunnett's Procedure. This differentiation between treatments with and without significant adult mortality was necessary because average reproductive capacity would have been affected by the number of live adults. Based on the results of the Dunnett's Procedure, the No Observed Effect Concentration (NOEC) and Lowest Observed Effect concentration (LOEC) were determined and the Chronic Value (ChV) was calculated (Horning and Weber 1985) . The ChV is the geometric mean of the NOEC and LOEC. Results Reference Toxic^^ni- Results of reference toxicant tests using sodium dodecyl sulfate (SDS) indicated that the test organisms were nominally sensitive (Table 4-2). LC50s varied from 2.8-5.5 mg/L. Literature values for SDS 48-hour LC50s are 1.5-8.2 ing/L for Ceriodaphnia dubia (FDER 1986) and 7-13 mg/L for 64 Table 4-2. Results of the Ceriodaphnia dubia 48-hour reference toxicant tests using sodium dodecyl sulfate (SDS) calculated by the moving average angle. LC50 (mg/L SDS) 95% Confidence Interval 2.83 5.53 4.61 2.01-3.62 4.66-6.74 3.80-5.41 65 for Daphnia magna, substantially lower Daphnia magna. Reference toxicity tests with CuSO^ produced a 48-hour LC50 of 92.7 ± 36 ug/L. Acute Toxicity A 48-hour range-finding test was used to delineate the appropriate Hydrothol concentrations for the chronic toxicity tests (Table 4-3) . The Ceriodaphnia dubia 48-hour LC50 was 0.49 ± 0.03 mg/L Hydrothol at 2 5° C. This agrees well with the published LC50 for Daphnia sp. at 21° C, 0.36 mg/L (Pennwalt Corp. 1980) , but is substantially lower than values recorded for several algal species. Mudge et al. (1986) found 1.5 mg/L Hydrothol to be the LC50 for an algal mix (Cyclotella, Euglena, Fragilaria, Nitzschia and Pediastrum) after five days of exposure. No other acute data on plankton responses to Hydrothol are available. At 15° C, the C. dubia 48-hour LC50 was 1.43 ± 0.32 mg/L Hydrothol. This increased tolerance of C. dubia to Hydrothol compared to results at 2 5° C may be attributable to a lower metabolic rate at the lower temperature (Gophen 1976) . Since Hydrothol is membrane-active and apparently affects m-RNA production (Ashton and Crafts 1981) , its impacts may be dampened with decreased temperature because processes involved in protein synthesis would be slower. Such a response would be typical of chemical reactions in general, as well as those mediated by enzymes (Wilson 1972). 66 Table 4-3. Acute toxicity of Hydrothol to various aquatic organisms. LC50 iTig/L Organism Temp . -C (95 % C.I. ) Duration of Test (h^ C. dubia* 25 0.495^ 48 Daphnia sp. 25 0.360" 48 algal mix 20.5 1.50^ 120 C. dubia 15 1.43^ 48 0.495^ (0 .363-0.7.65) 0.360^ 1.50^ 1.43^ (1 09-2.00) Ceriodaphnia dubia. ^ Results of the current experiments, Pennwalt Corp. 1980. ^ Mudge et al. 1986. 67 Chronic Toxicity Survival . The Ceriodaphnia dubia survival and reproduction test permits the calculation of a 7-day LC50 and uses changes in reproductive capacity over a 7-day period as a measure of sub-lethal chronic toxicity (Horning and Weber 1985) . At 25° C, the 7-day LC50 was 190 ±6.2 ug/L (Table 4- 4) . This value is lower than the suggested Hydrothol field application rate of 1-5 mg/L (Pennwalt Corp. 1980) by over an order of magnitude and points to the potential hazards of Hydrothol use in aguatic systems. Since its half-life is approximately 10 days (Blackburn et al. 1971, Reinert et al. 1985) , the impact of normal field application on the food chain could be devastating. While fish may escape the treated areas providing there are refugia, widespread use of Hydrothol in a lake could markedly reduce the zooplankton populations, which are less mobile, for 1-2 weeks after its application. Consequently, during periods of high fish reproduction, fry could be adversely affected by low zooplankton availability. Based on the results of the 48-hour tests in which C. dubia had a higher LC50 at 15° C than at 25°, it was expected that the LC50 at seven days would also be higher for the 15° C test. However, this was not the case. The LC50 was significantly (p<0.05) lower at 15° C (143 +4.6 ug/L), than at 25° C (190 ± 6.2 ug/L) (Table 4-5). Why this reversal in sensitivity occurred is not clear. It is 68 Table 4-4. LC50s from three replicate Ceriodaphnia dubia 7- day Hydrothol toxicity tests at 15° and 2 5° C based on the moving average angle method. LC50 ug/L at 15° C (95 % C.I.) LC50 ug/L at 25° C (95 % C.I.) 149 (114-199) 192 (134-326) 141 (103-210) 195 (142-306) 141 (103-210) 183 (141-255) MEAN (s.d. ) 143.7 (4.6) 190 (6.2) 69 Table 4-5. Reproduction data for replicate tests of ceriodaohnia dubia exposed to various concentrations of Hydrothol at 25" C for 7 days. [Hydrothol] (ug/L) Final Survival % Mean (S.D.) young/ female Mean No. broods/female 0 100 11.6(3.2) 2.50 25 90 5.0(2.8) 1.33 50 90 4.1(2.6) 0.80 ^ 100 90 2.9(2.6) 0.80 0 0 200 70 0 400 0 0 0 100 11.8(3.7) 2.50 ^ 25 100 5.9(3.4) i'on * 50 80 4.3(4.1) 0.90 ^ 100 80 1.8(1.5) 0.70 0 0 200 80 0 400 0 0 0 100 23.6(4.4) ?•? * 25 100 5.3(2.8) I'l * 50 100 0.2 0.2 100 80 0 0 0 0 200 70 0 400 0 0 Indicates a sign p < 0.05. ificant difference from control at 70 opposite to the response of several fish species tested at 18.3° and 23.3° C by Walker (1963), while fathead minnows (Chapter 3) showed no significant change in sensitivity to Hydrothol with a 10° C increase in temperature (15°-25° C) . Over time, mortality at the two temperatures became equal. Such results demonstrate the utility of chronic studies in assessing the impact of a toxicant on aquatic biota. Chronic effects of Hydrothol on reproduction. Chronic toxicity tests are designed to measure more subtle (sublethal) responses of organisms to toxicants than are acute tests. The Ceriodaphnia dubia survival and reproduction test (Horning and Weber 1985) uses changes in reproductive rate over a 7-day period as a measure of sublethal toxicity. The effects of a toxicant on zooplankton reproduction is more subtle but no less significant than its lethality. Even if a population of Ceriodaphnia dubia survives the initial stress of toxicant input it is still possible that fecundity may decline or cease. The impact of such an occurrence could seriously alter trophic level interactions in the ecosystem. Seven-day reproduction data for C^ dubia exposed to Hydrothol at 2 5° C indicated that even at concentrations as low as 25 ug/L (the lowest test concentration) , Hydrothol affected fecundity (Table 4-6) . Control animals produced 2.5-2.9 broods of young each and an average of 11.6-23.6 young during the tests. At a Hydrothol concentration of 25 ug/L, fecundity was significantly (p<0.05) reduced to 1.10- 71 Table 4-6. Summary of the chronic toxicity of Hydrothol at 25° C to Ceriodaphnia dubia based on reproduction. [Hydrothol 1 10 uq/L Control Rep. 1 Rep. 2 Rep. 3 Final Survival 90 60 100 100 Mean No, (S.D. ) Young/ female 25.4(6.9) 14.8(9.9) 20.7(4.7) 18.3(4.2) Mean No. (S.D. ) Broods/ female 2.9(0.32) 2.7(0.67) 2.4(1.07) 2.4(0.52) NOEC (ug/L) LOEC (ug/L) ChV (ug/L) N/A N/A N/A <10 25 <15.8 10 25 15.8 <10 25 <15.8 Denotes significant difference from control at p < 0.05, 72 1.33 broods and an average of 5.0-5.9 offspring per adult female. This lowest observed effect concentration (LOEC) was almost eight times lower than the LC50 at 2 5° C, 190 + 6.2 ug/L Hydrothol . At higher toxicant concentrations, reproduction was reduced even more. Since Hydrothol affects the production of m-RNA (Ashton and Crafts 1981) , its impact on C. dubia reproduction is not surprising. Normal production and development of eggs is a process requiring adults to have healthy metabolic capacity. If protein synthesis is impaired by the limited availability of m-RNA, enzymes would become limiting factors. To determine the NOEC (no observed effect concentration) used to calculate a chronic value (ChV) , I ran 3 additional 7-day tests with controls and 10 ug/L Hydrothol test concentrations (Table 4-7) . At 2 5° C, the chronic value (ChV) for Hydrothol is less than or equal to 15 ug/L based on reproduction. Since there was no reproduction in seven days in the tests performed at 15° C, no statistical analysis of toxicant effects was possible. The fact that no reproduction occurred at this low temperature is no surprise. McNaught and Mount (1985) found that the 7-day C. dubia reproduction test became a 28-day test at 18° C. Even at 20° C it took nine days for Cj^ dubia to produce three broods of offspring (Cowgill et al. 1985). At 15° C, the 73 Table 4-7. Acute toxicities to zooplankton of several herbicides used to control submergent macrophytes in Florida lakes. Herbicide Organism 48- -hour LC50 mg/L Temperature Hydrothol-191 C. dubia 0.49 25 Aquathol-K D. magna 316^ 25 diquat D. magna 7.1*d 21 dichlobenil D. pulex Simocephalus D. magna 9.8*^ 15 15 21 2,4-D D. magna 100*^ 21 diuron Simocephalus D. pulex D. magna 2.0^ 47.0^ 15 15 21 ^ Pennwalt Corp. 1980. ^ Johnson and Finley 1980. ^ Water hardness 272 ppm CaC03 . * IC50 at 26-h. <^Crosby and Tucker 1966, 74 metabolic rate decreases significantly from that at 22 C (Gophen 1976) , and was reflected in a much slower reproductive rate. Discussion The acute toxicity of Hydrothol to Ceriodaphnia dubia has been determined based on survival at 48-h. My findings confirm previous conclusions that Hydrothol is considerably more toxic to zooplankton than some alternative compounds (Table 3-4). For example Aguathol-K, an inorganic salt of endothall, has a 48-h LC50 of 316 rag/L. That level is far above field use levels (5-10 mg/L) and should not pose a threat to zooplankton (Pennwalt Corp. 1980) . Hydrothol is often chosen over Aquathol because the former is better for control of algae and remains effective longer. Dichlobenil is another effective herbicide that is non- toxic to aquatic fauna at normal use concentration (Ag Consultant 1989). It acts to inhibit CO2 fixation and oxidative phosphorylation in plants. Johnson and Finley (1980) determined that the 48-h LC50 for D. pulex was 3.7 mg/L and for Simoceohalus spp. it was 5.8 mg/L, both at 15 C. A common herbicide used in water hyacinth (Eichhornia crassipes) control programs is 2,4-D (Ag Consultant 1989). It is also used to eliminate broadleaf weeds in sorghum, sugar cane and alfalfa fields. By a complex mixture of effects at the cellular level, this herbicide inhibits cell 75 division and impairs nucleic acid metabolism. At 21 C, the 48-h EC50 of 2,4-D was >100 mg/L for D. magna (Crosby and Tucker 1966) . Various algae, water hyacinth, coontail (Ceratophyllum demersum) , hydrilla (Hydrilla verticllata) , pondweeds ^Potamoaeton spp.) and several broadleaf weeds in agricultural fields can be controlled with the use of diuron. Diuron is a substituted urea herbicide (Weed society of America 1979) that inhibits photosynthesis by blocking electron transport (Ashton and Crafts 1981). The 48-h LC50 was 2.0 mg/L for .simocephalus and 1.4 mg/L diuron for D. pulex. These values represent levels exceeding those produced by proper weed control programs (Ware 1979). Because of their important role in the aguatic food web, the response of zooplankton to long-term treatments with a pesticide is important to know before it is widely used. However, until recently there were no accepted methods to assess impacts on zooplankton reproduction. Even now, only acute toxicity test (24-48 hours) data are required by EPA for pesticide registration (Zucker 1985a). Only two other studies have provided information on the chronic effects of endothall herbicides on zooplankton. Serns (1975) followed the response of zooplankton to a 5 mg/L Aquathol-K exposure from June through October. Plant control was effective, resulting in the increased presence of Chara, but no significant change in the structure or composition of the zooplankton community was noted. 76 Cladocerans and copepods exhibited their usual seasonal changes in density. Results from a field study of the efficacy and impacts of Aquathol-K and Hydout, a pelletized amine formulation used to control Hydrilla. found little effect on zooplankton populations (Westerdahl 1983). A movement of zooplankton into the water column as plant height decreased and an increase in naupliar size 49 days after treatment were noted. However, zooplankton community structure and composition remained constant throughout the post-treatment period. Results from the aforementioned studies contradict those of my laboratory study with Hydrothol-191 . I found a significant reduction in reproduction by Ceriodaphnia dubia in concentrations as low as 0.016 mg/L. There are several factors that may explain these differences. First of all, neither the study by Serns (1963) nor the work by Westerdahl (1983) used the same formulation as I did. Serns (1963) tested Aquathol-K, while Westerdahl (1983) used both Aquathol-K and Hydout. The toxicity of Aquathol-K to aquatic organisms is several orders of magnitude lower than that of Hydrothol-191 (Pennwalt Corp. 1980, Johnson and Finley 1980) . Hydout is an amine formulation, as is Hydrothol-191, but the former is a slow-release granular product, while Hydrothol 191 is a liquid. This difference may affect the amount of herbicide in solution at any time. 77 Second, in the laboratory study I renewed the test solutions each day, thereby maintaining a constant exposure level. The studies by both Serns (1963) and Westerdahl (1983) were conducted outdoors using one application of the herbicide. Therefore, concentrations of endothall began to decrease immediately due to microbial degradation and biotransformation (Reinert and Rodgers 1987) . Finally, adsorption of a herbicide may remove significant amounts from the pool of biologically active compound in waters containing macrophytes, algae and sediments. Both of the field studies were performed in the presence of natural flora and sediment (Serns 1963, Westerdahl 1983) . Thus, effective concentrations of endothall were likely reduced compared to those in the laboratory test vessels. The latter contained only the test organism and solution. This study showed that temperature had a measurable effect on the toxicity of Hydrothol to C. dubia. However, the relationship between temperature and survival after 2-d exposures was inverse to that noted in 7-d tests. At 48-h, the LC50 at 15° C was 1.43 mg/L Hydrothol, while at 25° C it was 0.49 mg/L. With 7-d exposures, the LC50s decreased at both test temperatures, but the survival rate was lower at 15° C than at the higher temperature. The reasons for this contradiction are unclear. A lower metabolic rate may have initially protected the animals in 15° C tests from the impact of Hydrothol on m-RNA production. However, it 78 appears that with longer exposure low temperature compounded the toxicity of Hydrothol. Because this herbicide is applied in late spring or early summer, zooplankton in Florida should not be concurrently exposed to both low temperature and Hydrothol toxicity. CHAPTER 5 SIMPLIFICATION OF IN VITRO CULTURE TECHNIQUES FOR FRESHWATER MUSSELS Introduction Recently, there has been growing concern over the loss of freshwater mussel species (Unionidae) and their declining densities in areas perturbed by pollution and installation of dams. These molluscs, historically abundant in most North American waters, inhabit both lakes and streams. The area with the greatest number of species and individuals was the Mississippi River and its tributaries, notably the Cumberland, Tennessee and Ohio rivers (Burch 1973). The unusual mode of reproduction of unionid molluscs makes their culture in the laboratory more difficult than other groups that have free-living veliger larvae. The life cycle of unionid mussels includes a parasitic larva (glochidium) that normally attaches to fish gills or fins during early development. To propagate these molluscs in vitro, a suitable culture medium is necessary to provide the nourishment usually obtained from the host fish. With the hope of replenishing declining natural populations, the Tennessee Valley Authority began funding research to develop methods for in vitro propagation of these freshwater molluscs in the early 1980s (1982, 79 80 1984) . The goal was to eliminate the need for fish hosts during the larval stage so that laboratory culture of mussels would be practical. In turn, such artificial propagation would produce a large number of juvenile mussels for use in restoration of lost natural populations. As a result, a culture medium containing vitamins, glucose, amino acids, antibiotics and fish plasma in place of live fish was developed (Isom and Hudson 1982, 1984; Isom 1986). The transformation in culture of glochidia to juveniles takes 9-30 days (23° + 3° C) depending on the species, culture temperature and glochidia maturity at the start of incubation (Isom and Hudson 1982) . While the Hudson and Isom method (1982) is far better than that of Ellis and Ellis (1926) which relied on the use of fish hosts for encystment of glochidia during transformation, further simplification is desirable. The old method (Isom and Hudson 1982, 1984) is laborious, still requires the use of fish plasma which may not be readily available nor of consistent quality, and a CO2 incubator. A simplified method for mussel culture is necessary before juveniles can be available in the numbers needed for replenishment of declining wild stocks or for other purposes, e.g. toxicity tests. The objectives of the work described here were: (1) to use standard tissue culture media and plasma available from commercial 81 suppliers to propagate unionid mussels in vitro as a means of simplifying the culture technique, (2) to determine if the use of non-bicarbonate organic buffers, i.e. N-2-HydroxYethylpiperazine-N'-2-ethanesulfonic acid (HEPES) or 3-[N-morpholino] propanesulf onic acid (MOPS), would circumvent the need for a CO^ incubator to maintain pH, and (3) to test the efficacy of these methods in the culture of several species of mussels. Material'^ ^"^ Methods Test Organisms Glochidia of Anodonta imbecilis, the feeble mussel, was used in culture experiments. Since longterm propagation and culture of unionid mussels has not been achieved to date, gravid females must be collected when they are naturally available. In northern Florida, females carrying mature glochidia can be found from April through June. Most of the mussels used in the development of these culture techniques were collected from the Suwannee River, Florida. Several specimens of A. imbecilis were obtained from Dr. Paul Yokely, of the University of Northern Alabama. Anodonta imbecilis was chosen because it had been successfully cultured by Isom and Hudson (1982), it is a widely distributed mussel and has a relatively long reproductive period (two to three months depending on the location) . These characteristics are important when 82 choosing an organism as a potential bioassay animal, one of the proposed uses of juvenile mussels produced by these in vitro techniques. Plasma substitutes Culture techniques developed by Isom (1986) and Isom and Hudson (1982) were used as the starting point for simplification. Their culture medium, modified from Ellis and Ellis (1926) and Eagle (1959) , uses vacuum sterilized fish plasma as a nutrient source during culturing in place of the fish themselves. It contains a mixture of amino acids, salts, glucose, vitamins, antibiotics (carbenicillin, rifampin, gentamycin, amphotericin B) and phenol red as a pH indicator. A typical 15 X 60 mm culture dish would contain 2 ml of medium, 1 ml of serum and 0.5 ml of the antibiotic/antimycotic agents as described in Isom (1986) . Glochidia are removed from the gills of a female mussel, washed and added (500-1000) to the culture medium under a sterile hood. The plates are placed in a CO2 incubator (5% CO2) at 23° + 3° C. Isom and Hudson (1982) found 23° C to be the incubation temperature that allowed transformation but kept bacterial and fungal growth to a minimum. Cultures are monitored daily under a microscope to follow the process of organogenesis. When the foot becomes active and other parts are developed, juveniles are said to be transformed. They are then placed in 83 water where they can begin siphoning water for oxygen and food. While standard tissue culture methods require the use of plasma or serum (Ham and McKeehan 1979) because they contain essential proteins, growth factors and hormones that enhance cell division, there has been no indication of their specific role in glochidia transformation. Isom and Hudson (1982) determined that fish plasma was an absolute necessity for successful transformation of glochidia for all mussel species they cultured. However, verification of their findings was desirable since the simplification of procedures afforded by the substitution of a more readily available protein source would be substantial. Therefore, two modifications of the culture medium studied were first, the substitution of other protein sources for fish plasma at 5% w/v and second, the use of other sera (33 % v/v) readily available from tissue culture supply houses in place of fish plasma. Protein sources were acetone precipitates of trout liver, salmon liver and rabbit pancreas, and bovine casein (Sigma Chemical Co.). For each protein, 3 g of powdered extract were mixed in 60 ml of distilled water for three minutes by vortexing. The resulting slurry was centrifuged at 1500 g for 5 minutes to remove undissolved materials. One ml of the supernatant was used per three ml of culture medium. Alternate sera used were bovine. 84 neonatal calf and horse. These sera were used at the same final culture concentration as was fish plasma (1 ml/2ml growth medium) . Glochidia were also cultured in medium with no protein source or plasma. Culture medium with fish plasma was used as the control. Per cent transformation of glochidia in each culture medium was used as the measure of success of the modification, ANOVA and Duncan's Procedure were used to analyze results from a total of 6 trials with 2 plates per treatment. Three microscope fields (40X) were counted per treatment to determine the number of glochidia that had transformed vs those that had not. The untransformed glochidia included those that did not begin to develop at all due to lack of maturity, those that did not complete transformation and those that were non-viable after 24 h. In cases where glochidia transformed but the juveniles were lethargic and survived only 24 hours, such a response was taken as an indication of morbidity and the medium judged unsuccessful in producing juveniles for field or laboratory use. C02_Incubator Once the necessity for plasma was tested, the use of organic buffers in place of the CO2 incubator as a pH- stat was investigated. In the Isom and Hudson (1982) method, culture pH is maintained in the optimal range (7.3-7.4) by a HCO3-CO3 buffer system based on NaHC03 and 85 C02. The CO2 atmosphere is provided by a CO^ incubator. MOPS (3-[N-morpholino] propanesulfonic acid) and HEPES (hydroxyethylpiperazine-N'-2-ethanesulfonic acid), non- bicarbonate organic buffers, are widely used in tissue culture methods for many cell lines (Ham and McKeehan 1979) . To test their efficacy in pH maintenance in a non- CO2 environment, either MOPS or HEPES were added to the complete medium (0.22 %) in addition to NaHC03 , and the pH was adjusted to 7.3-7.4 by titration with NaOH. Five hundred to a thousand glochidia were cultured at 23° ± 3° C in pairs of culture dishes containing standard medium with bicarbonate, or medium fortified with MOPS or HEPES (0.22% w/v) in addition to NaHC03 . Incubation temperature was set at 23° ± 3° C based on results of Isom and Hudson's work. One dish was then placed in an incubator with 5% CO2 at 100% relative humidity, while its duplicate was incubated in ambient air at 24° + 3° c. Again, per cent transformation was the parameter used for statistical analysis. Use of Commercial Media A third series of experiments was designed to see if standard tissue culture media could be used in place of the medium developed by Isom and Hudson (1982). Their medium must be made from many separate reagents that are components of commercial media, e.g. Medium 199 (M199) 86 and Dulbecco's Modified Eagle's Medium with high glucose (DME) . The advantages of using commercial media are that they are: (1) easier to use, (2) readily available, and (3) manufactured under consistent conditions with quality control that may not be possible in all research laboratories. Glochidia were cultured in the Isom and Hudson (1982) medium, M199 or DME (with added antibiotics) , and horse serum (1 ml/2ml medium) . DME and M199 were hydrated in distilled water, adjusted to pH 7.3-7.4 with NaOH and filter-sterilized prior to their use, just as was the Isom and Hudson medium. Per cent transformation was compared among the three media as a measure of media suitability for mussel culture. Species Cultured Finally, Anodonta imbecilis , Lampsilis teres and Villosa lienosa were cultured using M199, DME or Isom and Hudson's (1982) medium and horse serum. As mentioned before, A. imbecilis has been cultured in vitro for several years by Hudson and Isom (1982). Hudson and Isom (1982) have also had success culturing V. lienosa and L. teres using fish plasma and their own culture medium. These species are less widely distributed than A. imbecilis but are common in northern Florida and were collected in the Suwannee River (Burch 1973) . The usefulness of simpler culture methods would be greatly 87 enhanced if a number of species could be transformed using them. Transformation of Villosa lienosa and LamEsilis teres was also attempted using horse serum and the commercial media. Results in the first group of tests, transformation success ranged from 0% with casein to a mean of 95.5% for neonatal calf serum based on six trials (Table 5-1) . While they did develop, juvenile mussels transformed in the salmon and trout media were not as healthy (inactive, lethargic) as those from the neonatal calf and horse media although the transformation success for these four groups were not significantly different based on ANOVA and Duncan's procedure (p - 0.05). Since in vitro propagation of mussels is designed to provide stock either for replenishment of declining wild populations or for toxicity testing, survivability past transformation is important. Therefore, salmon and trout acetone precipitates of liver were judged inadequate serum substitutes. While transformation success was as good in neonatal calf serum (95.5 ± 1.9%) as it was in horse serum (94.7 ± 4.0 %), horse serum was selected over neonatal calf serum because the latter is more expensive, in all cases, the use of sterile serum eliminated or markedly decreased bacterial growth common 88 Table 5-1. Per cent transformation of Anodonta imbecilis glochidia in media with various protein sources or sera for 6 trials with 2 plates counted per treatment. Serum or Protein Source Mean % (s.d.) Neonatal calf serum 95.5^ (1.87) Horse serum 94.7^ (3.98) Salmon liver 91.5^ (5.39) Trout liver 8 3.0^^ (13.83) Fish plasma 81.8*^ (7.47) Rabbit pancreas 67.5^ (20.17) ■^Treatments with the same letters were not significantly different from each other (p < 0.01). 89 in cultures with fish plasn^a. This was a major problem in earlier work (Isom and Hudson 1982). Results from the second group of experiments testing transformation success for Anodonta imbecilis cultures incubated either in a CO^ (5%) atmosphere or ambient air indicated that there was significantly (p Cu > Cd = Zn > Ni > Fe > Mn (Jones 1964) . This order was based on literature values and was explained on the basis of the solubility and reactivity of the metal ions. Another ordination related metal toxicity with the electron configurations of their outer electron orbitals (Kaiser 1980) . The most toxic group was comprised of Sn^"*", As^"*", Se"^"^ and Pb^"^ which have filled d and s orbitals, but unfilled p orbitals. The least toxic 124 group — Na"*", Be^"*", Ba^"*", Al^"*" and Cr^"^, had configurations like inert gases. However, responses to heavy metals vary considerably according to the organisms involved (Hellawell 1986) , therefore only trends in toxicity can been identified. Another area of interest in toxicity testing is the impact of metal mixtures on biota. While water guality criteria are established on the basis of single compound toxicity test (EPA 1979) , many metals enter aquatic systems as mixtures. For example, zinc and cadmium occur together both in uncontaminated waters (Lake 1979, NRC 1979) and in industrial effluents (Casarett and Doull 1975, Hemelraad et al . 1987). Nickel, cadmium and mercury may be discharged due to the manufacture of batteries (Occhiogrosso et al. 1979). Thus, there is a need for information on the impact of metal mixtures to aquatic biota. Although many studies have been performed to determine the toxicity of heavy metals to invertebrates, relatively few have used animals from flowing waters (Whitton and Say 1975) . It is particularly important to determine such effects because rivers and streams have been the recipients of much industrial waste over the years. The older literature attributed faunal declines in contaminated streams to the presence of various metals, but did not quantify the relationship (Carpenter 1924, Jones 1940 and 1958). 125 During the last 20-30 years, a marked decline in species diversity and density of freshwater mussels has been observed in many streams that receive mining and industrial effluents (Havlik and Marking 1987, Clarke 1970) . Although molluscs are among the most sensitive to heavy metal pollution (Wurtz 1962), few experimental data quantifying their susceptibility to metals are available. It has been noted in numerous field surveys that mussel species are declining (Rasmussen 1980) , but the fact that they often carry high body burdens of metals (Anderson 1977, Foster and Bates 1978, Jones and Walker 1979) suggested that metals were killing the mussels. In fact, almost no verifying data exist (Imlay 1971) even for adults, much less earlier life stages. With the placement of over 70 species of freshwater mussels on the threatened or endangered species list (USFWS 1989) , the establishment of acceptable exposure limits for these species has become a priority. In accordance with the need for basic experimental data, a series of acute toxicity (96-h) tests were performed to determine whether juvenile mussels were sensitive to metal pollution. Possibly they bioaccumulate significant concentrations of metals, but are unable to withstand the same levels in direct exposure. On the other hand, the loss of mussels from metal polluted rivers and streams may be caused by 126 habitat destruction, sedimentation, pesticides or other factors rather than metal toxicity. Six of the seven most toxic heavy metals were chosen for testing — mercury, zinc, nickel, cadmium, copper and chromium (Hellawell 1986) . Using methods developed earlier in this dissertation, juvenile mussels were exposed to each of these metals separately and in several mixtures to determine their 96-h LC50s (lethal concentration to 50% of the organisms) . Of the many environmental characteristics that can modify the toxicity of metals to biota, water hardness is among the most important (Sprague 1985) . Water hardness protects against the toxicity of heavy metals in two possible ways. First, metals become less soluble in hard water as they form complexes with carbonates. Second, water hardness, caused primarily by Ca^"*" and Mg^"*", may decrease membrane permeability and therefore uptake of metals from water (Everall et al. 1989). Therefore, the effect of water hardness on metal toxicity to mussels was also examined. Finally, separate experiments were performed to compare the sensitivity of mussels and the zooplankter Ceriodaphnia dubia to metal-contaminated sediments and an industrial effluent containing chromium. It was desirable to determine how similar their sensitivities were because zooplankton are commonly used in toxicity tests as surrogates for mussels (EPA, personal 127 communication) . Data from these experiments with single metals in soft and moderately hard water metal mixtures, and the contaminated sediments can be used both to determine whether ambient water levels of the metals could have caused the loss of mussel species and to help set more appropriate water quality criteria. They may also validate the use of zooplankton toxicity data in setting safe exposure limits for mussels. Materials and Methods Test Organisms One- to two-day old juveniles of the freshwater mussel Anodonta imbecilis were used as test organisms. Glochidia (larvae) of these mussels were cultured in vitro using one ml of horse serum and two ml of DME (Dulbecco's Modified Eagle's Medium) in a CO2 incubator. Details of the culture method are contained in Chapter 5 of this dissertation. The transformation process was observed periodically during culture until activity of the mussel's foot was visible through the shells. At that time (usually 6-10 days) , transformation was complete and cultures were transferred to soft reconstituted freshwater. Test Methodology Dissolved Metals. After 24 to 48 hours in water, juveniles were randomly distributed in test chambers which consisted of 15 X 60 mm pyrex Petri dishes. Ten 128 animals were placed in 15 ml of solution in each of two replicates per test concentration. Metals used in toxicity tests were: Cu * 5 H2O, ZnSO^ • 7H2O, NiSO^ • 6H2O, HgCl2 and CdCl2. A stock solution of each metal was made in deionized water and diluted to test concentrations with either soft or moderately hard reconstituted freshwater (Peltier and Weber 1985) . Five dilutions (in a 60% dilution series) plus a control (soft or moderately hard reconstituted freshwater) were used for each metal. Mussels were not fed during the tests, nor were test solutions renewed. The test endpoint, death, was determined based on absence of a visible heartbeat. The total number of survivors was recorded by replicate and concentration each day, and used to calculate a 96-h LC50. Metal Mixtures. Four mixtures of metals were evaluated for toxicity to Anodonta imbecilis juveniles. The combinations were chosen to include a pair of very toxic metals (Cd and Cu) , a pair comprised of a metal of low toxicity (Zn) and one of moderate toxicity (Ni) , a pair that consisted of two moderately toxic metals (Hg and Cr) , and one made of a very toxic metal (Cd) and a minimally toxic metal (Zn) . The level of toxicity assigned to each single metal was based on results of earlier tests with single metals. 129 The mean LC50 from single metal toxicity tests with A. iinbecilis was chosen as the highest concentration of each metal used in the mixture tests. Other concentrations were prepared by a series of 60% dilutions. Exposures were performed in soft reconstituted freshwater under conditions matching those described earlier for single metal tests. Rudiment Tests. Since no experimental data were available to determine whether mussels were more sensitive to dissolved or sediment-bound metals, two experiments tested the effect of sediment-sorbed cadmium and copper on Anodonta imbecilis juveniles. In each test, five grams of washed and dried (80° C) Miami River sediment (3% organic content) was put in a 50 ml glass vial with a teflon-lined lid. The metal solution was added (30 ml), the vial capped and then mixed overnight on a wrist-action shaker. Five dilutions of cadmium or copper in soft reconstituted freshwater and a control were prepared in this manner for each test. Shaking the solutions in the presence of sediments was done to load the sediments with the metal while removing them from the aqueous phase. After being shaken overnight, the sediments and solutions were transferred to 50 ml beakers and allowed to settle for at least 24 h prior to the introduction of juvenile mussels into the chambers. Mussels were not fed nor was the water aerated during the 96-h tests. 130 Five neonatal (<24 h old) Ceriodaphnia dubia were added to each chamber as reference organisms. These animals are very sensitive to metals and because they are pelagic organisms of a different taxonomic group, their sensitivity relative to that of A. imbecilis should enhance the usefulness of test results. If C. dubia has a sensitivity to metals at least as great as that of the mussels, it would be appropriate to use the zooplankter as a surrogate for mussels in future tests of metal toxicity. Since there are already numerous test organisms in use, it would be preferable not to add another unless it is necessary. The number of survivors of both groups were tallied each day and behavior was noted. Metal Effluent. An effluent from an airplane maintenance company (Flying Colors) provided by the City of Gainesville, Florida, was also tested for toxicity to juvenile mussels. This particular effluent was thought to be contaminated with metals based on tests with Microtox and B-galactosidase (C. Maziji, G. Bitton, and B. Koopman, unpublished data) . Sample analyses later verified the presence of 6,430 mg/L of chromium but found no other contaminants. A 96-h toxicity test was simultaneously performed with mussels and C. dubia. Ten mussels and five C. dubia neonates were placed in replicate 30 ml plastic test chambers containing 15 ml of effluent. Five dilutions (3%-0.4%) of the effluent 131 determined from a screening test and a control consisting of moderately hard reconstituted freshwater were used. Test Chemicals The metals used in these tests were prepared from reagent grade salts dissolved in soft or moderately hard reconstituted freshwater. They included CUSO4 * SHjO, CdCl2, ZnSO^ • 7H2O, K2Cr207, HgCl2 and NiS04 * 6H2O. Stock solutions were diluted in a 60% series and added to test chambers. Metal concentrations were measured as total metal using a Perkin-Elmer atomic absorption spectrophotometer Model 5000 with single element lamps, following EPA guidelines (U.S.E.P.A. 1983). Since determinations of mercury concentrations required the use of equipment not available at this time, mercury concentrations are given as nominal rather than measured. All single metal tests in each type of water were performed at least three times, while metal mixture toxicity was determined twice. Data Analysis Survival data were analyzed by several methods. A set of EPA (Peltier and Weber 1985) computer programs calculated the LC50. These programs, known as the TOX- DAT Multimethod, calculate the LC50 using moving average angle, probit and binomial methods. LC50s were then analyzed by ANOVA and Duncan's multiple range test to 132 determine if there were differences in toxicity among metals and metal mixtures. Mixture toxicity was calculated based on the concentration of individual metals. These values were then compared to the LC50 for the same metal in single toxicant tests to determine if synergistic, antagonistic or additive toxicity were evident. Calculations of the additive index followed the system of Marking and Dawson (1975) . The sum toxic action (S) is calculated as follows: Am + Bm Ai Bi For S - 1.0, Additive Index =1 -1.0; S For S - 1.0, Additive Index = S(-l) + 1. Am = LC50 of metal A in the mixture. Ai = LC50 of the same metal A alone. Bm = LC50 of metal B in the mixture. Bi = LC50 of metal B alone. Index values (S) of zero indicate a simple additive effect of the two metals compared to their toxicity when tested individually. Index values greater than zero result from mixtures that have synergistic toxicity. Those mixtures whose index values are less than zero are said to contain antagonistic toxicants (Marking and Dawson 1975) . All statistical analyses except calculation of LC50s utilized the SAS statistical package (SAS 1986) available 133 at the Northeast Regional Data Center, University of Florida, Gainesville. Results Dissolved Metals In general, 96-h LC50s for the mussels were lower than those measured at 48-h, and metal toxicity was reduced in moderately hard water compared to soft water (Tables 7-1 and 7-2) . These findings are in agreement with those from studies using other aquatic organisms (Alabaster and Lloyd 1982, Petrocelli 1985). In soft water, the order of metal toxicities to A. imbecilis at 48-h was: Cd > Cu > Hg > Ni > Cr > Zn. The 48-h LC50 for Cd^"^ (0.057 mg/L) was approximately three times lower than the value for Cu^"*" (0.171 mg/LO, four times lower than Hg^"^ (0,216 mg/L) and 5-6 times lower than the LC50s for Ni^^, Cr^"^ and Zn^^ (Table 7-2). The greatest increases in toxicity between 48-h and 96-h were seen for copper and chromium. While the other metals exhibited a 1.3-3 fold decrease in LC50s over that time period, the toxicities of chromium and copper increased approximately six and 7.5 times, respectively (Table 7- 2). Because there are no published studies of metal toxicity to juvenile mussels, direct comparisons with my results cannot be made. However, the trend toward the 134 Table 7-1. Comparative toxicities to Anocionta imbecilis juveniles of single metals at 48-h and 96-h in soft and moderately hard water. Toxicities for waters within the same time and metal category with the same letters were not significantly different based on ANOVA and Duncan's multiple range test (p - 0.05). Metal Cd Cr Cu Hg Ni Zn Soft water. ^ Moderately hard water, Type of Water 48-h 96-h MH^2 MH^ MH*^ mh'^ mh'^ S^ S^ MH^ MH^ MH^ S^ MH^ S^ 135 Table 7-2. Mean (s.d.) 48-h and 96-h LC50 values for juvenile Anodonta imbecilis mussels exposed to six metals in soft reconstituted freshwater. The number of test replicates is indicated by N. Metal Cd +2 Cr + 6 Cu +2 Hg + 2 Ni + 2 Zn +2 Time (h) N Mean LC50 (s.d.) (ma/L) 48 3 0.057 '0.006) 96 3 0.009 '0.003) 48 3 0.295 ;0.060) 96 3 0.039 '0.034) 48 3 0.171 '0.087) 96 3 0.086 '0.031) 48 3 0.216 '0.086) 96 3 0.147 '0.035) 48 5 0.240 ( 0.093) 96 5 0.190 ( 0.097) 48 3 0.355 ( 0.108) 96 3 0.268 ( 0.095) 136 relative greater toxicities of mercury, cadmium and copper versus those of nickel, zinc and chromium is in accordance with the literature on other invertebrate organisms. In tests with Lymnaea acummata, Khangarot et al. (1982) determined the toxicity order to be Hg > Cu > Cd > Ni > Cr > Zn. Anderson (1950) exposed D. magna to several metal solutions. Their relative toxicities to D. magna arranged in decreasing order were Hg > Cu > Cd > Zn > Ni. For Tubifex tubifex, the order was Hg > Cd > Cu > Cr > Zn > Ni (Brkovic-Popovic and Popovic 1977). In a comparison of heavy metal toxicity to the freshwater snail Riomphalaria glabrata, Ravera (1977) found copper to be more toxic than cadmium, and cadmium was more toxic than chromium. Gupta et al. (1981) exposed the mollusk Viviparus benaalensis to five metals and found their relative toxicities in the following order: Cu > Zn > Cr > Cd > Ni. Variability in the response of different taxa to individual metals may be accounted for by physiological differences in the organisms. Many metals are needed in trace amounts as cofactors or as components of specific enzymes. The extent to which an animal is affected by the presence of high levels of metals may depend on the importance of specific enzymes and transport systems in its metabolic processes. cadmium, because of its role as an antagonist to the uptake of calcium by induction of metallothionein 137 (Hammond and Beliles 1980) , causes skeletal deformities and nervous disorders (Alabaster and Lloyd 1982). It may also cause ion imbalances and interrupt energy production (Hiltibran 1971, Larsson et al . 1976). Copper is a necessary component of many enzymes. However, copper at high concentrations can cause precipitation of mucus on fish gills and to branchial cell damage (Ellis 1937, Baker 1969), Tissue levels of copper rise upon exposure to sublethal and sublethal concentrations (Calamari and Marchetti 1973, Kariya et al. 1967) resulting in liver and kidney damage (Leland and Kuwabara 1985) . Little else is apparently known of the harmful mode of action of copper on fish (Alabaster and Lloyd 1982) . Zinc is ubiquitous in the natural environment and an essential trace element for normal cell differentiation (Leland and Kuwabara 1985) . It is part of a number of metalloenzymes and serves as a cofactor for many other enzymes. Zinc levels in cells can affect carbohydrate, fat and protein metabolism as well as other metabolic processes. Ion imbalances, skeletal deformities and nervous system disorders have been noted in fish exposed to high concentrations of zinc (Bengtsson 1974a, Bengtsson 1974b, Lewis and Lewis 1971) . The primary site of action of mercury on cells is the sulfhydryl groups on surface membrane proteins (Luckey and Venugopal 1977) . As almost all enzymes 138 depend on the normal orientation of their sulfhydryl groups for proper conformation and function, the potential impact of mercury on cell processes is clearly evident (Leland and Kuwabara 1985) . Neither nickel nor chromium are very toxic to most animals (Hellawell 1986) . No functional action of nickel or chromium has been described (Hammond and Beliles 1980) . The toxicities of all six metals except mercury were significantly lower (p - 0.05) in moderately hard water than they were in soft water (Table 7-1) . Once again, cadmium was the most toxic to A. imbecilis with a 48-h LC50 of 0.137 mg/L and a 96-h LC50 of 0.107 mg/L Cd"*"^ (Table 7-3). Mercury was the second most toxic metal in moderately hard water. At 48-h, the LC50 was 0.223 mg/L Hg"*"^, and its toxicity increased only slightly at 96-h to 0.171 mg/L. These values for mercury are not significantly different (p<0.05) from comparable measures in soft water. The toxicity of copper to A. imbecilis in moderately hard water was almost half that observed in soft water. Except for the 96-h LC50 for chromium, Ni and Cr were also half as toxic in moderately hard water as in soft water. Finally, zinc toxicity increased from 0.588 mg/L to 0.438 mg/L at 48-h and 96-h respectively, in moderately hard water. This was an increase of 1.37 times. 139 Table 7-3. Mean (s.d.) 48-h and 96-h LC50 values for juvenile Anodonta imbecilis mussels exposed to six metals in moderately hard reconstituted freshwater. The number of test replicates is indicated by N. Metal Cd + 2 Cr + 6 Cu + 2 Hg + 2 Ni + 2 Zn + 2 Time (h) N Mean LC50 (s.d. 3 (n>q/ 0.137 'D 48 '0.034) 96 3 0.107 '0.128) 48 3 1.187 '0.313) 96 3 0.618 '0.168) 48 3 0. 388 0.036) 96 3 0.199 < 0.006) 48 3 0.223 < 0.061) 96 3 0.171 < 0.038) 48 3 0.471 ( 0.035) 96 3 0.252 ( 0.050) 48 2 0.588 ( 0.035) 96 2 0.438 ( 0.132) 140 The impact of water hardness on metal toxicity has been noted in studies with other aquatic species. Chromium toxicity to the worm Tubifex tubifex was reduced from 1.5 mg/L at 48-h in soft water to 4 . 8 mg/L Cr^"^ in hard water (Brkovic-Popovic and Popovic 1977) . There was a 2-20 fold decrease in chromium toxicity to fathead minnows and bluegill as water hardness was increased 1800% (Pickering and Henderson 1966) . Similar increases in LC50 values were noted for lead toxicity to rainbow trout (Davies et al. 1976), the effects of zinc on stickleback (Jones 1938) and rainbow trout (Lloyd 1960) , copper toxicity to rainbow trout and the carp, Cyprinus carpio. (Tabata 1969) , and cadmium's lethality to goldfish (Alabaster and LLoyd 1982) and rainbow trout (Brown 1968) . Metal solubility and therefore bioavailability are decreased in hard water (Sprague 1985) . In summary, at 48-h, three metal toxicity groups were identified in soft water. Zinc was the least toxic (p - 0.05). Cu, Cr, Ni and Hg formed a moderately toxic set. Cadmium was the most toxic. The same trend was evident at 96-h, but the moderately toxic group — Hg, Ni, Cu, Cr — changed order. Cadmium remained the most toxic metal in soft water after 4-day exposures and zinc was the least toxic (Tables 7-4 and 7-5) . In moderately hard water, chromium was the least toxic at both time periods while cadmium was the most toxic. With longer 141 Table 7-4. Anodonta imbecilis toxicity data for metals dissolved in soft reconstituted freshwater by ANOVA and Duncan's multiple range test. The test organism was Anodonta imbecilis. Metals connected by the same line are not significantly different (p— 0.05). Metals were increasingly toxic going from left to right. SOFT WATER 48-h LC50S Zn Cr Ni Hg Cu Cd 96-h LC50S Zn Hg Ni Cu Cr Cd 142 Table 7-5. Results f.f 1^.1,1^^1 ifSiJaST^ toxicity data for metals dissoiv ^^^^^^'s multiple -SiJ^ln^^i^^oiirroin,'?r°o. lift to rl,.t. MODERATELY_HARD_WATER Cr Zn 48-h LC50S Nl Cu Hg Cd 96-h LC50S cr Zn Ni CU Hg Cd 143 exposure, the toxicities of individual metals to Anodonta imbecilis in moderately hard water became more similar. At 96-h, the toxicities of Cr and Zn were significantly different from all other metals and each other. However, Ni, Hg, Cu and Cd were not significantly different from each other. Metal Mixtures The toxicity of a particular metal can increase (synergism) , decrease (antagonism) or remain unchanged (additive) when combined with another metal (Marking 1985) . Which of these responses will occur depends on whether the presence of one facilitates the uptake of the other metal or whether or not they compete for the same transport sites on the membrane (Sprague 1985) . In the present study, there was a trend toward a greater toxicity of metals in combinations containing Ni, Zn, Hg or Cu than for single metal exposures (Tables 7-6 and 7-7) . However, the increases in toxicity were not significant (p - 0.05) based on ANOVA and Duncan's multiple range test. In contrast, cadmium had a higher 96-h LC50 in mixtures with Zn (0.029 mg/L) and Cu (0.012 mg/L) than it did alone (0.009 mg/L) and the LC50 for chromium alone, 0.039 mg/L, increased to 0.148 mg/L in combination with Hg. Of the four mixtures tested, the combination of cadmium and copper produced the greatest toxicity to A. 144 Table 7-6. Comparisons of single metal LC50s at 48-hours to the same metal in combination with another for Anodonta imbecilis. Values represent the mean (s.d.) LC50 for two tests. 48-h LC50 fmq/L) Additive^ Mixture Individual In Combination Index Ni 0.240(0.093) 0.128(0.043) -0.14 and Zn 0.355(0.108) 0.217(0.055) Cd 0.057(0.006) 0.050(0.14) -0.58 and Zn 0.355(0.108) 0.249(0.071) Hg 0.216(0.086) 0.094(0.014) -0.43 and Cr 0.295(0.060) 0.170(0.054) Cd 0.057(0.006) 0.037(0.020) -1.36 and Cu 0.171(0.087) 0.062(0.035) '^ Calculation of Additive Index followed Marking and Dawson (1975) . 145 Table 7-7. Comparisons of single metal LC50s at 96-hours to the same metal in combination with another for Anodonta imbecilis. Values represent the mean (s.d.) LC50 for two tests. 96-h LC50 (mq/L^ Additive^ Mixture Individual In Combination Index Ni 0.190(0.097) 0.088(0.032) -0.07 and Zn 0.268(0.095) 0.162(0.051) Cd 0.009(0.003) 0.029(0.029)* -2.76 and Zn 0.268(0.095) 0.145(0.025) Hg 0.147(0.035) 0.088(0.012) -1.32 and Cr 0.039(0.034) 0.148(0.064) Cd 0.009(0.003) 0.012(0.001) -0.58 and Cu 0.086(0.031) 0.021(0.003) Significantly different from LC50 of individual metal (p — 0.05). All other combinations and individual LC50S were not significantly different. ^Calculation of Additive Index followed Marking and Dawson (1975) . 146 imbecilis. The 48-h LC50 for copper individually was 0.171 mg/L while in combination with cadmium, copper had an LC50 of 0.062 mg/L (Table 7-6). Likewise, the individual 48-h LC50 for cadmium was 0.057 mg/L, but in the mixture with copper it decreased to 0.037 mg/L. Marking and Dawson (1975) developed an index to determine whether chemicals in mixtures exerted an antagonistic, synergistic or additive effect on each other. Index values statistically indistinguishable from zero are representative of additive effects. Synergistic effects are identified by index values greater than zero and antagonism of chemicals is indicated by values less than zero. If the calculated index range contains zero, it is said to be indistinguishable from zero. The additive index (Marking and Dawson 1975) for Cu and Cd after two days' exposure of A, imbecilis was -1.36 with a range of -1,27 to +1.66 (Table 7-6). Therefore, Cu and Cd exert an additive toxicity effect on each other. At 96-h, copper and cadmium had generally lower LC50S individually and in the mixture than at 48-h (Table 7-7) . However, cadmium was slightly less toxic at 96-h in combination with copper. Because of this, copper and cadmium were determined to be slightly antagonistic to each other based on the Marking and Dawson (1975) index. The only other antagonistic effect was measured between Hg and Cr at 96-h. No synergistic effects were seen in any of the four mixtures. 147 Similar evaluations of metal fixture toxicity by other investigators have produced various results. Attar and Maly (1982) determined that Cd and Zn were antagonistic to each other in toxicity tests with D. „^. Thompson et al, (19B0) found that the toxicity of Z„-CU mixtures to blue,ills was additive. Lloyd <1961) found the combined effect of Cu and Zn on rainbow trout was additive at low concentrations and synergistic at higher levels. The presence of high levels of zinc inhibited the uptaKe of cadmium in adult M^dsnta S^^^^ (Hemelraad et al. 1987) and no lethality was noted. goHimpnt- Tests NO toxicity to either A. imbg£ills or C. dafela was detected in sediment tests wxth copper and cadmium. Based on earlier results with aqueous exposures, the nominal agueous concentration range of 1,000 ug/L <200 .g/Kg sediment, to 130 ug/L (26 mg/Kg sediment) would have Rilled both species without the presence of sediments. However, in the presence of sediments all mussels and zooplanKtcn survived for the duration of the 96-h tests. Changes in mussel behavior were noted. Hussels did not burrow into the substrate in chambers with metals, in contrast to observations of their behavior in control chambers, and offspring were seen developing in c. d^bU in lower metal concentrations « eoo ug/L) , while in containers with higher metal levels 148 no reproduction was visible. Zooplankton molts were visible on the bottom of test chambers containing <600 ug/L metal and food was seen in their guts. The non-lethal effect of the metals on both organisms was probably due to sediment binding of metals. There was virtually no metal measured in the aqueous phase of any test chamber. However, based on measurements of acid-digested samples, sediment metal concentrations were lower than expected. It is possible that some metal was lost by sorption to glass during preparation of the sediments prior to initiation of the tests. Effluent Toxicity The "Flying Colors" effluent was very toxic to both A. imbecilis and C. dubia. Their 48-h LC50s were 0.95% effluent and 0.57% effluent, respectively (Table 7-8). At 96-h, the effluent toxicity to both animals increased. A concentration of less than 0.6% was lethal to half the A. imbecilis juveniles (LC50=0 . 58%) , while 0.43% was lethal to half of the C. dubia neonates. Analysis of this effluent by the City of Gainesville showed that it contained 6.4 3 mg/L Cr"^^. The LC50 of 0.6% for mussels is equal to 0.039 mg/L, the same as the 96-h LC50 in soft reconstituted water but much lower than 149 Table 7-8 . The toxicity of "Flying Colors" effluent to Ceriodaphnia dubia. Anodonta imbecilis and Microtox. LC50 values are in per cent whole effluent. LC50 Organism Time (h) % effluent 95% C.I Ceriodaphnia dubia 48 0.57 0.10-0.95 96 0.43 0-0.67 Anodonta imbecilis 48 0.95 0.74-1,18 96 0.58 0-0.040 Microtox* 0.25 0.07 N/A From C. Maziji, unpublished data, 150 the value for moderately hard water at 96-h (0.618 mg/L) . However, metals are usually less toxic in hard water than in soft water as evidenced by data from this dissertation as well as by literature data. Discussion The freshwater mussel A. imbecilis is as sensitive to dissolved metal pollution as are zooplankton and may be more sensitive than some insects (Table 7-9) . These data show that while mussels can accumulate high levels of metals from their environment and live for some time, they may also be adversely affected by much lower concentrations. Before using a surrogate animal to determine safe exposure levels for another, it is prudent to verify that they have comparable sensitivities. Having done so in the laboratory, it appears that zooplankton can be used to identify waters that may be toxic to mussels. However, the habitats of these two organisms is so different that they may not be exposed to the same level of toxicant in nature. Certainly, laboratory findings using controlled conditions may produce conservative estimates of toxicity. No accounting of substrate, water flow, interactions with other organisms or the effect of various food types was made. These factors may have substantial impact on the determination of LC50s (Sprague 151 Table 7-9. Comparative toxicities of selected metals in soft water to several invertebrates and fish. Organism Metal Water Hardness (mg/L) Time LC50 Ref. (h) (mg/L) A. imbecilis Cd Daphnia Cd Chironomus Cd Bluegill Cd 39 45 25 20 48 48 48 96 0.057 0.065 8.05 1.96 3 5 2 A. imbecilis Cr Daphnia Cr Chironomus Cr Bluegill Cr A. imbecilis Cu Daphnia Cu Chironomus Cu Bluegill Cu 39 25 36 39 45 25 44 48 48 48 24 48 48 48 96 0.295 1.800 11.80 0.280 0.171 0.065 0.327 0.884 6 4 4 3 5 1 A. imbecilis Hg Daphnia Hg Chironomus Hg Fathead minnow Hg A. imbecilis Ni Daphnia Ni Chironomus Ni Bluegill Ni 45 45 25 45 39 45 25 25 48 48 48 96 48 48 48 48 0.216 0.005 0.029 0.168 0.240 0.510 0.327 69.50 3 5 1 3 5 1 A. imbecilis Zn Daphnia Zn Chironomus Zn Bluegill Zn 39 45 25 20 48 48 48 96 0.355 0.100 8.20 5.38 3 5 2 * References: (1) Mayer and Ellersieck 1986. 1966. (3) Biesinger and Christensen 1972. (4) Heath 1979. (5) Khangarot and Ray 1989. (6) Ray 1987. All others from this study. (2) Mount Smith and Khangarot and 152 1985) . However, the acute toxicity test is accepted as giving the most precise results and is still the most widely used method for determining relative sensitivity to pollutants. Water hardness has a major effect on metal toxicity to mussels. This has been demonstrated for many other organisms and is related to metal chelation and to physiological responses of the organisms (Sprague 1985) . The affect of water hardness on the toxicity of metals to mussels is important to know because many of the streams with depauperate mussel fauna have received acid mine drainage or industrial effluents over the years. They also have low water hardness (Jones 1940 and 1958, Wurtz 1962, Dieffenbach and Ryck 1976), While measured concentrations of metals in these streams are lacking for the most part, there has been enough circumstantial evidence to support the assertion that metals were responsible for the faunal decline. Comparison of LC50 values for metals and ambient water quality guidelines established by EPA (USEPA 1976) indicates that current standards may be adequate to protect mussel species. In most cases, the maximum limit allowed to be released during a 30-day period (Table 7- 10) is lower than the mussel 96-h LC50 in moderately hard water. The only exception was seen in the case of nickel. However, since this value was based on total 153 Table 7-10. Comparisons between EPA^ water quality moderately hard freshwater) . Metal EPA Guidel 24-h Max. ines (ug/L) Max. Limit 48-h LC50 for A. imbP-nilis fua/L) Cd 0.038 4.6 137 Cr 0.290 21 1, 190 Cu 5.6 32 388 Hg 0.2 4.1 223 Ni 130 2500 471 Zn 47 450 588 154 metal concentration in water with a hardness of 150 mg/L CaC03 , and my data were generated at a somewhat lower hardness (39 mg/L CaC03), the values are not likely to be different in terms of biological impact. Standards are not set for waters with low alkalinity. Since higher water hardness decreases metal toxicity (Sprague 1985) perhaps receiving stream hardness should be considered in determining safe levels of metals in effluents. Determination of the impact on mussels of in. situ chronic exposures (several months) to low metal concentrations is critically needed. Metals are sometimes believed to contribute to the decline of mussel species in rivers meeting water quality standards. This assumption is supported by information from both the Tennessee Valley Authority on metal concentrations in rivers judged to be good mussel habitat based on species diversity (Jenkinson and Heuer 1986) and from long-term studies on the Clinch River, Virginia, which is one of the best remaining mussel habitats (American Electric Power Service Corp. 1989) . However, water quality standards do not take into account the potential for bioaccumulation or sublethal responses. Long-term exposure could affect fecundity or behavior, e.g. filtering capacity or ability to burrow, that could lead to species decline. The impact of low concentrations of metal mixtures is not considered in the development of water quality 155 criteria (U.S.E.P.A. 1976). From this research, however, metals can be more toxic to mussels at lower concentrations in combination than they may be singly. This conclusion is in agreement with the literature on other aquatic organisms. Since sources of metal pollution rarely produce pure metal wastes, it seems reasonable to assess the impact of mixtures in setting effluent concentration limits. Comparisons of mussel sensitivities to metals with those of other invertebrate species, particularly zooplankton, suggest that setting water quality standards based on tests with the latter will adequately protect mussel fauna. Similarly, the responses of A. imbecilis and C. dubia, a standard effluent toxicity test organism, indicate that the latter animal may serve as an acceptable surrogate for mussels in metal effluent toxicity tests. The two organisms appear to be equally sensitive to chromium waste. However, analyses of the toxicity of other metal effluents should be performed to determine how broadly the responses of these two organisms overlap. CHAPTER 8 THE TOXICITY OF SEVERAL PESTICIDES, ORGANIC COMPOUNDS AND A WASTEWATER EFFLUENT TO THE FRESHWATER MUSSEL, Anodonta imbecilis, THE ZOOPLANKTER, Ceriodaphnia dubia AND THE FATHEAD MINNOW, Pimephales promelas Introduction Pesticides (insecticides, herbicides, molluscicides, piscicides and nematicides) have been used to eliminate or control pests since at least Grecian times (Edwards 1973, Ruzicka 1973). Pliny the Elder advocated the use of arsenic to control insects in A.D. 70, Other inorganic compounds were used for centuries, and although many were persistent in soils resulting in crop damage (Edwards 1973), they were not as refractory or globally distributed as the modern organic pesticides. Development and use of organic chemicals to control pests increased as a result of the growth of the petrochemical industry in the 1940s (Nimmo 1985) . New pesticides were effective in reducing crop losses caused by insects, competitive weeds and soil microbes, and the incidences of malaria, typhus, dysentery and other diseases (Edwards 1973) . However, because of their biocidal capacity, they also posed a serious threat to non-target organisms (Hellawell 1986) . Therefore, it is surprising that little thought was given to the impact of 156 157 the widespread use of such pesticides might have until the 1960s (Carson 1962, Rudd 1964, Butler and Springer 1963) . In the past 20 years, vast amounts of data have accumulated showing the effects of pesticides on non- target organisms, particularly fish and zooplankton (Mayer and Ellersieck 1986, Johnson and Finley 1980) . Due to differences in chemical structure, there are wide ranges in the sensitivities of non-target organisms to pesticides. However, some generalizations can be made. Organochlorine insecticides (e.g., DDT, lindane, toxaphene, chlordane) are the most persistent and most toxic to fish of all organic pesticides. They are very insoluble in water (1 ug/L-7 mg/L) , tending to partition into lipids and organic substrates (Nimmo 1985) . Thus, they accumulate in biota and the ecosystem. While the mode of action of organochlorine insecticides has not been clearly established (Ware 1978) , they cause spontaneous nerve impulses that eventually lead to convulsions and death. Organophosphate insecticides (e.g., malathion, diazinon) which were developed after the organochlorines, tend to be more water soluble (24-2500 mg/L) , less persistent and less toxic to fish (Nimmo 1985) . The organophosphate insecticides prevent the breakdown of acetylcholine by acetylcholinesterase. In so doing, they 158 interfere with the normal passage of nerve impulses and cause paralysis (Ware 1978) . The newer carbamate insecticides (e.g., carbaryl, carbofuran) are water soluble (4 0-7 00 mg/L) highly unstable and toxic to fish in only very high concentrations (Nimmo 1985) . Like the organophosphates, carbamate insecticides are cholinesterase inhibitors (Ware 1978) . Pyrethroid insecticides such as Karate (ICI Americas) are derivatives of pyrethrum, a natural botanical insecticide (Hellawell 1986) . They inhibit normal sodium and potassium conductance in the neuron that eventually blocks the passage of nerve impulses. Paralysis of muscles follows (Murphy 1980) . They are highly toxic to fish and invertebrates, but are rapidly degraded by photolysis and therefore do not persist (Ware 1978, Johnson and Finley 1980), Pentachlorophenol (PCP) has been used as a wood preservative (fungicide), and as a molluscicide in some parts of the world where snails serve as vectors of human parasites (Cheng 1974, Ware 1978, Hellawell 1986). It is the second most heavily used pesticide in the United States (U.S.E.P.A. 1986a). Being highly chlorinated, PCP is very toxic to both plants and animals. Its mechanism of action is via a combination of plasmolysis, protein precipitation and uncoupling of oxidative phosphorylation (Ware 1978) . 159 Modern herbicides are also less persistent than chlorinated insecticides and less toxic to non-target aquatic biota than were the defoliants widely used in 1960s (2,4-D and 2,4,5-T) (Edwards 1973). The triazines, e.g. atrazine, are used extensively in agriculture and silviculture to control broadleaf and grassy weeds (Ware 1978, Weed Science Society of America 1979). Triazines are strong inhibitors of photosynthesis (Ware 1978, Weed Science Society of America 1979) . Endothall herbicides, e.g. Hydrothol-191 and Aquathol-K, are among the most effective for use in the control of the aquatic weeds Hvdrilla verticillata , valisneria americana and Myriophvllum spicatum in canals, lakes and streams (Dumas 1976) . Endothalls have half- lives of 3-10 days depending on the formulation and are biodegraded by microorganisms (Rienert and Rodgers 1987). The inorganic endothall herbicides, such as Aquathol and Aquathol-K, are less toxic to non-target biota but are not as effective as the organic endothall, Hydrothol-191 (Reinert and Rodgers 1987) . Organic compounds other than pesticides are released into the aquatic environment. Included among these are sodium dodecyl sulfate (SDS) , a bacteriocidal surfactant, and ethylediamine tetraacetate (EDTA) a chelating agent used in shampoos and detergents, surfactants disrupt cell membranes (Ware 1978), while EDTA removes cations from solution. A number of organic 160 solvents are used in industry, and as diluents of hydrophobic pesticides for use in aquatic toxicity tests. Two of the most common are acetone and methanol . Even though more stringent effluent water quality criteria were instituted in amendments to the Clean Water Act of 1977 (National Pollution Discharge Elimination System, NPDES) it is necessary to monitor the impacts of effluents on aquatic biota on a continual basis. The Environmental Protection Agency (EPA) requires testing for such impacts on aquatic organisms as part of the registration and re-registration processes for pesticides (Zucker 1985a and 1985b), as part of the pre- manufacturing notification for toxic chemicals under the Toxic Substances Control Act, and for re-permitting of many municipal wastewater facilities (NPDES) . Receiving waters for wastewater and industrial effluents are usually streams and rivers. Likewise, water associated with agriculture is often flowing in canals and ditches. However, standard toxicity test batteries do not include representatives of the lotic fauna. The most common test organisms are fish and zooplankton which primarily inhabit lakes (Johnson and Finley 1980, Mayer and Ellersieck 1986, Peltier and Weber 1985, Buikema et al. 1982). Thus, little is known about the sensitivity of most benthic invertebrates. A wastewater entering the aquatic environment affects hundreds or thousands of species. Therefore, 161 using just a few test species, i.e., fish and zooplankton, may lead to underestimation of the impact on the ecosystem. Patrick et al. (1968) and others have shown that macroinvertebrates and algae are often more sensitive to toxicants than are fish. Additionally, algae and macroinvertebrates are organisms on which fish depend for food (Buikema et al. 1982). Fish may be indirectly affected by a decline in the density or biomass of such organisms, regardless of their direct response to the toxicant. There is concern over the status of one group of stream organisms in particular, the unionid mussels. With the recent designation of over 70 species of unionid mussels as endangered or threatened (USFWS 1989), it has become necessary to assess the impact of pesticides, herbicides and other organic pollutants on their survival. Significantly, more than 10 species of freshwater mussels found in Florida are candidates for inclusion on the threatened or endangered species list. Federal law (Endangered Species Act) mandates that pesticide and herbicide use must be limited to levels that are not detrimental to designated species in watersheds containing endangered or threatened organisms. It is impossible to set valid standards or limits without appropriate data. Finally, the use of indigenous species for toxicity testing is ideal when possible because the organisms are acclimated to ambient conditions and. 162 therefore, provide a good indication of biotic responses in their locale (Buikema et al. 1982). Therefore, the EPA was interested in testing the toxicity to freshwater mussels of several pesticides (atrazine, carbaryl and Karate) undergoing evaluation for registration or re-registration. Determining the toxicity of other selected organic pollutants to mussels was essential in assessing the adequacy of current water quality standards and the development of more protective water quality criteria if necessary. It was also important to assess the comparability of zooplankton and mussel sensitivities because the EPA Office of Pesticide Programs is advocating the use of Daphnia magna as a surrogate for freshwater mussels in toxicity tests. The goals of my research in this area were to: (1) determine the toxicity of several pesticides, organic compounds and an organic effluent to juvenile Anodonta mussels, and (2) compare their sensitivities with common test organisms such as D. magna, Ceriodaphnia dubia and Pimephales promelas, the fathead minnow. Materials and Methods Test Organisms A. imbecilis glochidia were cultured jji vitro using methods described earlier (Chapter 3) . After their transformation, juveniles were put in soft reconstituted freshwater (Peltier and Weber 1985) and used for tests 153 usually within two days. Daphnia magna were obtained from a local source using EPA approved culture methods. Ceriodaphnia dubia were cultured in this laboratory. Pimephales promelas (fathead minnow) larvae used in the effluent toxicity test were obtained from EPA-Newtown, OH, via overnight mail and used immediately. Test Conditions Aqueous Exposures. Toxicity tests with pesticides and pure compounds were performed for 48-h using methods developed earlier in this dissertation. All tests were performed in an environmental chamber with 16 hours of light and eight hours of darkness, at a temperature of 22° + 1° C. Five test concentrations were used, plus a control which was soft reconstituted freshwater. Two replicates, each containing 10 juvenile mussels, were used per concentration in either 200 ml crystallizing dishes (Karate, carbaryl and atrazine) or 15 X 60 mm glass Petri dishes with lids. All tests were static except for those using carbaryl and Karate which were known to decompose rapidly. Solutions of these two pesticides were renewed at 24-h. Details of the test protocol were given in Chapter IV. Hydrothol-191, an endothall derivative (Pennwalt Corp., Philadelphia, PA.), was dissolved directly in soft reconstituted freshwater to make a stock of 530 mg/L. Because lindane is only slightly soluble in water (10 164 mg/L) , a stock solution was made in methanol. Small aliquots of the stock were added to the test chambers directly. Na * PCP has low solubility in water, as well. It was dissolved in 0.01 N NaOH and pH was adjusted to 7.0. Stocks of the remaining compounds--SDS , methanol, acetone and EDTA— were prepared by dissolving the reagents directly in soft reconstituted water. Dilutions used in definitive tests were made as appropriate based on range-finding tests. Test solutions were made by 60% dilution of the stocks with soft reconstituted freshwater. All toxicant concentrations given are nominal except for those of toxaphene and chlordane. Karate, atr^^ine and carbarvl ■ Tests with Daphnia magna neonates (< 24 h) were performed separately but concurrently with mussel tests for Karate, carbaryl and atrazine. The EPA was particularly interested in comparisons between Daphnia magna and Anodonta imbecilis sensitivities to two of the pesticides, i.e. carbaryl and atrazine, because use of these pesticides is already being limited in watersheds with endangered mussels to levels deemed safe by zooplankton tests. However, there were no data to support the assumption that D. magna and unionid mussels have comparable sensitivities to carbaryl or atrazine. Additionally, the toxicity of Karate (ICI Americas), a pyrethroid insecticide, was assessed because it is currently being tested for registration. 165 Since carbaryl and atrazine have such low solubilities in water (40 mg/L and 33 mg/L at 30° C, respectively) , a saturated solution of each was prepared as follows: 100 mg of the pesticide was added to 1 L of soft reconstituted water and stirred overnight; the supernatant was filtered and small aliquots of this stock were added directly to the test chambers. For each 48-h toxicity test, ten D. magna neonates were randomly placed in each chamber which consisted of a 200 ml crystallizing dish containing 100 ml of toxicant solution. Two replicates were prepared for each of five dilutions and a control containing soft reconstituted freshwater (Peltier and Weber 1985) . The D, magna were fed at 24-h during the tests. Toxaphene and Chlordane Tests. Chlorinated pesticides are known to adsorb to soils or sediments where they can remain for many years (Ware 1978, Menzer and Nelson 1980) . While both of these pesticides have been recently de-registered in the United States for most uses (52 C.F.R., 51 C.F.R), they persist in the sediments of many bodies of water. Little was known about the toxicity of sediment-sorbed pesticides to mussels, or whether such infaunal molluscs are differentially susceptible to sediment-bound or aqueous concentrations. Therefore, I made a comparison of the toxicity of toxaphene and chlordane in both water and laboratory- spiked sediments. 166 TWO sets of chambers were prepared for each of these insecticides. One series was prepared to determine toxicity in the usual way, dissolved in water. In addition, the diminution of toxicity as a result of sediment adsorption was assessed using a series of test chambers that contained both water and sediments. Thirty ml of soft reconstituted freshwater were put into each 50 ml glass vial with or without sediment (5 g dried, 3% organic content) . An appropriate volume of stock pesticide in acetone was injected directly into each vial using a micro-syringe, then the vial was capped with a teflon-lined top. The vials were mixed over night on a wrist-action shaker, after which their contents were emptied into 50 ml beakers. Both sets of test chambers (with and without sediments) were left for 24 h prior to addition of the mussels, to allow sediments to settle and contents to equilibrate. Ten juvenile Anodonta imbecilis and five Ceriodaphnia dubia neonates were added to each of two chambers at each test concentration both with and without sediment. Toxaphene concentrations ranged from 1,829 ug/L to 0 ug/L. Chlordane was used at concentrations of 905 ug/L to 0 ug/L. Water samples from chambers used in chlordane and toxaphene tests were analyzed by gas chromatography to determine aqueous concentrations. Samples were extracted with three 10 ml aliquots of methylene chloride and later transferred to iso-octane. The extracts were blown down 167 to 1 ml with nitrogen and stored in crimpseal vials until analyzed on a Varian 3700 gas chromatograph. A 30 m DB-5 column with a 0.53 mm diameter and a 1 um coating was used with an initial temperature of 150° C ramped to 250° C at 5° C per minute. The injector temperature was 22° C and the detector was set at 300° C. The recovery rate for both pesticides was 28%. Effluent Toxicity Test. An industrial effluent sample was obtained from the Buckman Street Wastewater Treatment Facility, Jacksonville, FL for use in assessing the relative sensitivity of juvenile A. imbecilis mussels versus those of standard effluent test organisms. The 7- d effluent toxicity tests were performed using moderately hard reconstituted freshwater (Horning and Weber 1986) as diluent and control water. Forty A. imbecilis juveniles were exposed to each effluent concentration, 20 in each of two replicate chambers. Ten Ceriodaphnia dubia neonates and 2 0 Pimephales promelas larvae were exposed to each test concentration, the former in individual 30 ml plastic containers to permit monitoring of reproduction, the latter in two groups of 10 at each dilution in 1 L pyrex beakers. Toxicity was assessed based on protocols established in the Ceriodaphnia dubia survival and reproduction test, and the fathead minnow survival and growth test (Horning and Weber 1986) . Survival of test organisms was recorded 168 daily until termination at seven days. Zooplankton reproduction and larval fathead minnow growth were used as indicators of sublethal affects. Water was changed daily in all test chambers (Horning and Weber 1985) . Data Analysis Survival data were analyzed by several methods. A set of EPA (Peltier and Weber 1985) computer programs calculated the LC50s. These programs, known as the TOX- DAT Multimethod, calculate the LC50 using moving average angle, probit and binomial methods. Results of these analyses (LC50s) were then used to determine differences in toxicity among the chemical with ANOVA and Duncan's multiple range test. All statistical analyses except LC50S were performed using the SAS statistical package (SAS 1986) at the Northeast Regional Data Center, University of Florida, Gainesville. Results Aqueous Exposures Of the 12 organic compounds tested for toxicity to A. imbecilis, PCP was the most toxic, while methanol was the least toxic (Table 8-1) . Forty-eight hour LC50s for A. imbecilis exposed to acetone and methanol were 37.02 and 36.3 mg/L , respectively (Tables 8-1 and 8-2). A solvent must be non-toxic at concentrations ^10 ml/L to be used in toxicity tests. If a pesticide is not soluble 169 Table 8-1. Summary of acute toxicity test results for juvenile Anodonta imbecilis with twelve pesticides and organic compounds. N is the number of experiments performed with each toxicant. LC50 values with the same letters are not significantly different from each other (p^O.05) . 48-h Chemical N LC5Q (mg/L) Methanol 5 37.02 (4.7)^ Carbaryl 1 36.3^^ Acetone 2 33.83 (11.31)^^ Atrazine 1 33^^ SDS 3 19.04 (4.19)^ Lindane 3 > 10 ^ Hydrothol 3 4.85 (2.29)^ EDTA 3 1.35 (0.35)^ Karate 1 > l*'^ PCP 5 0.61 (0.26)'^ Chlordane 2 N/C"*" Toxaphene 2 N/C Solubility in water, 170 Table 8-2. Comparisons between the acute toxicities of several organic compounds for A. imbecilis, D. magna and L. macrochirus . LC50 (mg/L) Chemical A. imbecilis^ D. magna^ L. macrochirus^ Acetone 33.83 .0039^ Methanol 37.02 ll®*? 29.40^^ PCP 0.610 0.33^ 0.240^ SDS 19.04 10. 3^^ °- 48-h LC50. ^ 96-h LC50. '-' T.OT.7-i c ai-iH Ta7qV-,q>- 1 Q Q C d Lewis and Weber 1985.^ "^ Macek and McAllister 1970, Ceriodaphnia dubia. "5poirier et al. 1986. Ceriodaphnia dubia. -^ Salmo gairdneri 171 in 10 ml/L of solvent, it is considered to be insoluble in water and therefore not a threat to aquatic life. With that in mind, it is safe to use either of these solvents in pesticide tests with mussels because neither is lethal at the maximum allowable concentration. Forty- eight hour LC50 values for D. magna are 0.0039 mg/L acetone (Macek and McAllister 1970) and 11 mg/L methanol (Poirier et al. 1986). The sensitivity of S. gairdneri (rainbow trout) to methanol was 29.40 mg/L, intermediate between the values for the two invertebrates (Mayer and Ellersieck 1986) . SDS is often used as a reference toxicant in tests with zooplankton to verify that a particular set of test organisms is healthy in comparison with accepted norms. The 48-h LC50s for A. imbecilis exposed to SDS was 19.04 mg/L (Table 8-2). There are no established benchmark values for SDS toxicity to mussels. However, mussels were less sensitive to the surfactant than D. magna based on published values (Lewis and Weber 1985) (Table 8-2). A. imbecilis was relatively insensitive to all of the herbicides and insecticides tested (Table 8-3) . Lindane, an organochlorine insecticide, was not toxic to mussels at concentrations as high as its solubility limit in water, lOmg/L. Under such circumstances, a pesticide is said to be non-toxic to aquatic biota. Lindane is toxic to most other aquatic biota at low concentrations (Table 8-3) . Based on literature values, lindane is 172 Table 8-3. Comparative 48-h LC50s of five pesticides for three aquatic species. LC50 fua/L) Chemical A, imbecil is^ D. maqna'^ L. macrochirus^ Carbaryl 36,300 56f 15,800^ Lindane > 10,000* 485® 77f Karate > 1,000* 27 N/A Atrazine 33,000 9,800 42,000^ Hydrothol -p 4,280 360^ 940^ Mayer and Ellersieck 1986. ^ Pennwalt Corp. *} Johnson and Finley 1980. ® Henderson et al. 1959. ^ 96-h LC50. ^ 48-h LC50. * Solubility in water. '^48-h LC50. 173 toxic to both D. magna and L. macrochirus at <500 ug/L. However, Bluzat and Seuge (1979) calculated the 48-h LC50 for Lymnaea stagnalis to be 7,300 ug/L. Molluscs do not appear to be susceptible to lindane at concentrations that are normally found in water. The toxicity of the aquatic herbicide Hydrothol-191, to mussels was also very low. At 48-h, the mussel LC50 was 4,280 ug/L. In comparison, I determined the 48-h LC50 for rpriodaohnia dubia to be 190 ug/L and the 96-h value for fathead minnow larvae to be 468 ug/L in earlier tests (Chapter 4). Literature values for other aquatic organisms are also much lower than those of the mussels (Pennwalt Corp. 1980, Johnson and Finley 1980). PCP was acutely toxic to juvenile A. imbecilis mussels (Table 8-2). It was the only pesticide to which the mussels responded at a level similar to other species. The 48-h LC50 was 610 ug/L for mussels, 330 ug/L for D. magna (Lewis and Weber 1985) and 240 ug/L for bluegill (Macek and McAllister 1970). PCP is known to be toxic to virtually all biota, including molluscs (Ware 1978) and has been used as a molluscicide for years. The 48-h LC50S for two snail species, Lymnaea stagnalis and Gillia altilis, were found to be 240 ug/L and 810 ug/L PCP (U.S.E.P.A. 1986a). Therefore, it is not surprising that PCP was toxic to A. imbecilis. 17 4 ..,v,.-» car'-'-Y'' ""'^ Atrazlne Karate, while not toxic to mussels at its solubility li.it in water (1 r,g/L, ICI Americas), was lethal to halt of the D. maana at a concentration of 27 ug/L (48-h LC50) (Table 8-3) . No data are available yet on the sensitivity of other aquatic biota to Karate since it is still undergoing pre-registration testing. The LC50 for the carbamate insecticide, carbaryl, , T o i\ This value is much was 36,300 ug/L at 48-h (Table 8-3). This va „f= Kft nrr/T. for D. magna and higher than literature values of 56 ug/L tor _ _^i_ 15 800 ug/L for L. macrochirus (Mayer and Ellersieclc 1936). A. imbecllls was also less sensitive to carbaryl than the snail Lymnaea stMnalis (Bluzat and Seuge 1979). carbamates are typically very effective insecticides, but are non-toxic to mammals and non-insect arthropods. They are also transitory in the environment (Hellawell 1986, Mount and Oehme 1931) . D. m^ tested concurrently with A. imbecilis had a 48-h LC50 of 1.9 mg/L carbaryl. The herbicide atrazine was virtually non-toxic to juvenile A. imbeciUs mussels, having an LC5C of 33,000 ug/L (Table 3-3). Comparable values are 9.3 ug/L for D. ^^ neonates tested in this laboratory and 42,000 ug/L for bluegill sunfish (Mayer and Ellersieck 1986) . T,^vaphene "«^ rhiordane Tests Determinations of the toxicities of toxaphene and chlordane to A. imbecilis indicated that mussels were 175 tolerant of these insecticides at concentrations several orders of magnitude higher than were C. dubia. D. magna or L. macrochirus (Table 8-4) . Acute toxicity values for most aquatic organisms range from 2-40 ug/L for toxaphene and from 3-115 ug/L for chlordane (Johnson and Finley 1980). However, neither toxaphene (up to 1.83 mg/L) nor chlordane (ug to 0.90 mg/L) was toxic to A. imbecilis at 48-h in chambers without sediment. After four days' exposure, half of the mussels were killed by 0.74 + 0.07 mg/L toxaphene and 0.88 + 0.05 mg/L chlordane. There were no mussel deaths due to toxaphene or chlordane exposure in test chambers containing sediment (Table 8- 5). In those chambers, aqueous concentrations were markedly lower than in their counterparts without sediment. Ceriodaphnia dubia neonates were considerably more effected by both toxaphene and chlordane than were juvenile mussels (Table 8-4) . After 48-h, there were no survivors in test vessels without sediments. In contrast, no C. dubia died during the first two days in test chambers that contained sediments. By 96-h, all zooplankton had died in the toxaphene + sediment tests, while the LC50 in the chlordane + sediment chamber was 0.450 mg/L. Mussels were 1-2 orders of magnitude less sensitive than are fish or zooplankton. 175 Table 8-4. Acute toxicities of toxaphene and chlordane in soft water to A. imbecilis, D. magna and L. macrochirus. Organism LC50 (mg/L) Toxaphene Chlordane Anodonta imbecilis Ceriodaphnia dubia^ Daphnia magna '^ Lepomis macrochirus 0.74 ± 0. 07 0 88 + 0.05 N/C"^ N/C+ 0.010 0.029 0.018 0.092 XT LC50 at 48-h. " LC50 at 96-h. ^ Mayer and Ellersieck 1986. *D. pulex. ± N/C= not calculable because all died before 48-h. 177 Table 8-5. Measured concentrations of toxaphene and chlordane in replicate test vessels with and without sediments. Aqueous Concentration (mg/L) No sediment Sediment Toxaphene 0 340 824 910 1009 1829 0 0 466 331 625 559 Chlordane 0 191 350 ± 97.2 754 ± 41,2 864 ± 45.2 905.5 + 75.7 0 0 49.58 114.5 + 109.8 0 205 + 111.7 178 Effluent Toxicity Test The Buckman Street Wastewater Treatment Facility- effluent, known to contain several organic compounds including diazinon (Koopman et al. 1989, Dutton 1988), was less toxic to mussels than to either C. dubia or P. promelas (Table 8-6) . The tested sample, which was not analyzed for specific chemical contents, contained a volatile organic compound based on its smell. Juvenile mussels were 4-5 times less sensitive than were the zooplankton, and even less sensitive than that compared to fathead minnows which died at the lowest effluent concentration (6%) in 24-h. Ninety-six hour LC50s for A. imbecilis and C. dubia were 35.35% and 7.08%, respectively, while at 7-d the LC50s had decreased to 16.24% for the mussels and 4.97% for C. dubia. Reproduction levels in the controls were too low to determine subchronic effects on C. dubia. Discussion In contrast to the results of toxicity tests with metal pollutants, A. imbecilis was found to be generally less sensitive to organic pollutants than are standard toxicity organisms such as D. magna, Ceriodaphnia dubia, the fathead minnow and bluegill sunfish. The reasons for the apparent tolerance of mussels to pesticides, 179 Table 8-6 . Comparative toxicity of an effluent from the Buckman Street Wastewater Treatment Facility, Jacksonville, Florida to A. imbecilis. C. dubia and Pimephales promelas. Wastewater Organism 96-h LC50 7-d LC50 A. imbecilis 35.35 16.24 C. dubia 7.08 4.97 P. promelas N/C* N/C N/C=not calculable; all fathead minnows died in 24-h even in only 6% effluent. 180 herbicides and effluents, all having different chemical structures, characteristics and mechanisms is unknown. Pesticides and herbicides are generally used to eliminate specific organisms. In this role, their efficacy on target organisms is maximized and in recent times, their impacts on non-target organisms is minimized. Evidently, the physiology of the mussel A. imbecilis is sufficiently different from that of targeted plants and animals that they are not susceptible to chemicals that interfere with various processes in targeted biota. The only compound to which A. imbecilis showed sensitivity was PCP, a known molluscicide. It appears that the mode of action of this particular chemical is general enough that it can kill a broad spectrum of living organisms, including molluscs (Ware 1978) . It is particularly noteworthy that A. imbecilis was tolerant of extraordinarily high concentrations of the organochlorine pesticides lindane, toxaphene and chlordane. Organochlorines, in general, interfere with ion balance in the neuron via inhibition of Mg^"'"-Ca^"^ and Na -K ATPases. Lindane also perturbs cell division and causes proliferation of lysosomes (Ramade 1987) . As a result, such compounds are highly toxic to birds, mammals, fish and zooplankton (Ramade 1987, Johnson and Finley 1980) . It is interesting that both the mussels I 181 tested and the gastropod mollusc Lymnaea staqnalis were unaffected by lindane (Bluzat and Seuge 1979) . Acute exposures of mussels to high concentrations of sediment-sorbed toxaphene and chlordane were not lethal. This is of interest because while the use of both of these pesticides has recently been discontinued (C.F.R. V. 51 and v. 52) , they have been applied extensively to various agricultural crops over the last 30 years. In fact, toxaphene was the most heavily used pesticide in the 1960s and 1970s, replacing DDT for many uses after 1971 (U.S.E.P.A. 1986b). As a result, they remain in the sediments to which they have an affinity. However, this does not appear to be a threat to mussels. Since D. magna is currently used by EPA as a surrogate for mussels in deriving safety limits for pesticide use and that species was shown to be more sensitive to pollutants than was A. imbecilis. it appears that mussels are being adeguately protected. However, it is impossible to determine from these data what the effects of chronic exposures to pesticides or other organic compounds might be. Further testing is necessary to determine whether long-term exposure of mussels to pesticides or other organic compounds is responsible for the loss of mussels from rivers and streams where they were once plentiful. In recent years, there has been a move toward the use of microcosms and mesocosms to evaluate the impact of 182 pollutants on systems more complex than single-species exposure chambers (Cairns 1985, Giesy 1985, Taub 1973) . Because multi-species systems better mimic natural ecosystems, they are useful in measuring some of the inter-related responses of species. In a limited sense, such an approach was used during several phases of this dissertation. The response of Ceriodaphnia dubia and Anodonta imbecilis to organic compounds and metals was evaluated simultaneously. Since fish eat zooplankton, and freshwater mussels use fish as hosts for their larvae, there is a unique species interdependence. A more complex exposure system permitting longterm studies with all three species would have provided more concrete answers to questions about the real impact of pollutants on the survival of mussels, fish and zooplankton. However, several scenarios can be imagined. In cases where host fish species died because of their sensitivty to pesticides or metals, mussel larvae could not transform into free-living juveniles that eventually grow into adults. Over time, mussels would decline and disappear even though they were not directly eliminated by pollutants. If zooplankton were more sensitive to toxic substances than were fish, there would be a reduced food resource for fish fry potentially reducing growth and survival of young fish. The density of individual fish species might be lowered leading to changes in 183 competition and predation interactions among fish. Since this might change the availability of host fish for mussel larvae, the impact of pollutants on zooplankton could also affect mussel survival. Finally, some species of freshwater fish rely heavily on mollusks as food, e.g. redear sunfish. To the extent that fish consume mollusks, they may be negatively impacted by the loss of mussels from the food chain. In such cases, fish are indirectly affected by the pollutants that kill mussels. Aquatic toxicology is beginning to pass from its infancy into a discipline that can be increasingly adept at assuming an ecosystem-level perspective. We have many techniques to measure the affects of toxicants on single species of organisms in the laboratory. The current move toward development of laboratory micro- and mesocosm test protocols will make controlled tests more representative of natural ecosystems. CHAPTER 9 CONCLUSIONS The overall purpose of this study was to evaluate the sensitivity of the freshwater mussel Anodonta imbecilis relative to that of the typical toxicity test organisms, e.g. Pimephales promelas, Ceriodaphnia dubia and Daphnia magna. While macroinvertebrate animals comprise a major component of the fauna of flowing waters few but the insects have been used as test organisms. Currently many unionid mussels are listed as endangered or threatened species. However, no method to assess the impact of various toxicants on the survival of unionid mussesl has been available. In the process of determining whether fish or zooplankton are good indicators of mussel sensitivity to pollutants, several other goals were accomplished. These included: determining the toxicity of the aguatic herbicide Hydrothol-191 to the fathead minnow and Ceriodaphnia dubia relative to temperature at application, the simplification of in vitro culture technigues for Anodonta imbecilis^ and the development of a test method to assess the acute toxicity of toxicants to Anodonta imbecilis. 184 185 T.e conclusions fro. this study were as (oliows: , Hydrotnol-i91 is .ore toxic to PiM^Elmi^ ^^^-=^' +.v,,^i-iQi in controlling that may be as effective as Hydrothol 191 „ost undesirable .acrophytes. At levels one-t„ent.eth ,0.031 .g/L, the concentration allowed for field ..plication ,1-3 .,/M, Hydrothol significantly decreased the growth Of larval fathead minnows. With a half-l.fe ., XO days, normal use of this herbicide could have a ► ™ fish arowth and development. The measurable impact on fish grow 48.h LC50 for anadanta imb^sili^ «as 4.s5 m,/L. The effect of water temperature on the toxicity of „,drothol-191 to fathead minnows was determined to be impaired by Hydrothol concentrations below those that impaired growth at 25° C 2 Hydrothol-191 was found to be highly toxic to e^^i^^^Stoia iibia based both on survival and reproductive impairment, .cute toxicity was measured at 0.«0 m,/.. While survival was significantly lower after seven days in concentrations as low as 0.190 mg/L. f r dnbia was affected at Reproductive capacity of C. dubia ^ m ucf/L Water temperature Hydrothol concentrations of 15 ug/L. • . n^ a factor in determining the toxicity was not as important a factor 186 of Hydrothol-191 to C. dubia as it was for the fathead minnow. 3. The in vitro culture of Anodonta imbecilis was considerably simplified by the substitution of horse serum for fish plasma, and commercially available culture media for the idiosyncratic medium used previously. Fish plasma is difficult to obtain except in cases where an aquaculture facility is located nearby, while horse serum can be purchased from several commercial sources, substitution of horse serum also reduced the incidence of bacterial and fungal contamination of the glochidia cultures. Commercial culture media were also found to be adequate as nutrient sources for Anodonta imbecilis. in vitro propagation of freshwater mussels is an advantage for those interested in replenishing excised populations of endangered species because it obviates the necessity of finding a suitable host fish species, many of which have not been identified. If a population of endangered mussels were located, they could be cultured in the laboratory to the juvenile stage and then returned to suitable natural habitats. Being able to culture mussels in vitro is also an advantage for those who are interested in using them as a toxicity test organism. Since metamorphosis of the glochidia can be followed under a microscope, tests can be scheduled for a time when adequate numbers of juveniles will be available. 187 4. A simple, inexpensive and fast toxicity test protocol was developed for A. imbecilis. The procedure includes many of the same materials and methods used in methodologies already in widespread use. Thus, others interested in using freshwater mussels as test organisms can do so with ease. 5. Juvenile mussels were found to be sensitive to metal pollution. In comparison to data for zooplankton from toxicity tests performed both in this study and in others, A. imbecilis was at least as sensitive as Ceriodaphnia dubia and Daphnia magna to Cr^"*", Cd^"^, Ni^"*" + 2 and Cu . Mussels were generally more sensitive to metal toxicants than were fish and Chironomus, another benthic invertebrate. A. imbecilis was about as sensitive to a wastewater effluent containing Cr^"*" as was C. dubia, but less sensitive than were fathead minnow larvae. 6. Anodonta imbecilis was not as sensitive to eleven of twelve organic compounds and pesticides as were fish, Daphnia magna, and other common test species. Mussels were tolerant of very high concentrations of toxaphene and chlordane, as well as lindane and atrazine. Their 48-h LC50 for Hydrothol-191 was eight times higher than was the same value for Ceriodaphnia dubia. Juvenile mussels were found to be as sensitive to PCP as were both fish and zooplankton. PCP is a potent fungicide and molluscicide that is toxic to most organisms. In general, relative to widely used pelagic organisms, the freshwater mussel Anodonta imbecilis was found to be sensitive to metals but not so to organic compounds. While the species used is not endangered or threatened, it was chosen to represent the unionid mussels in toxicity tests because it has a broad distribution, can be propagated in the laboratory and has a relatively long reproductive period. There was a significant need to determine the sensitivity of mussels to metal and pesticide pollution. Further testing should be performed to expand the database on mussels. Based on this dissertation research, it appears that using D. magna as a surrogate for mussels in toxicity tests is acceptable for organic compounds. Since mussels appear to be more susceptible to metal pollution than are D. magna or C. dubia, zooplankton are not good substitutes for mussels in such tests. 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American Society for Testing and Materials. Philadelphia, PA pp. 2-10. Zucker, E. 1985a. Acute toxicity test for freshwater invertebrates. United States Environmental Protection Agency, Office of Pesticide Programs, Washington, D.C. EPA-540/9-85-005 . 211 Zucker, E. 1985b. Acute toxicity test for freshwater fish. United States Environmental Protection Agency, Office of Pesticide Programs, Washington, D.C. EPA-540/9-85-006. BIOGRAPHICAL SKETCH Anne E. Keller was born January 6, 1952, in Quantico, Virginia. During her first 17 years, she and her military family travelled throughout the United States and Europe. Once she entered Lake Forest College, Illinois, family trips were replaced by self-created adventures. After graduation from Lake Forest College, Anne attended the University of South Florida where she received a master's degree in zoology in 1976 and was certified to teach. From 1977 to 1982, Anne taught high school biology and chemistry in several Florida schools. Then, she returned to graduate school to pursue a Ph.D. in environmental engineering sciences at the University of Florida. In 1984, she received her M.S. degree having performed research on the response of Florida freshwater fish to changes in lake acidity. She is currently completing her Ph.D. in the speciality area of environmental toxicology. 212 I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of Doctos— of Philoj I'hdmas L. CrismaxL,,->eTrairman Professor of-'EInvironmental Engineering Sciences I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of Doctoj; of Phi).