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Pesticide effects on freshwater zooplankton: an ecological perspective
Environmental Pollution 112 (2001) 1±10
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Review
Pesticide eects on freshwater zooplankton: an ecological perspective T. Hanazato Suwa Hydrobiological Station, Shinshu University, 5-2-4 Kogandori, Suwa 392-0027, Japan Received 1 February 2000; accepted 11 March 2000
``Capsule'': Eects of pesticides on freshwater zooplankton are considered from an ecological perspective. Abstract The eects of pesticides on zooplankton are reviewed and their ecological signi®cance is discussed. Toxicity is shown to vary depending on animal species, genotype, life stage, and size at birth. Natural stresses such as food shortage, oxygen depletion and odors of potential predators can also aect toxicity. Populations in the growth phase are vulnerable to pesticides but have the potential to recover rapidly from the damage. Pesticides may aect the population dynamics by controlling individual survival and reproduction, and by altering the sex ratio. Furthermore, toxic chemicals may control predation risk by changing swimming behavior and body morphology, and this in turn in¯uences the population dynamics. Many zooplankton display morphological and behavioral responses to predators when exposed to their odor-producing chemicals. However, pesticides induce a maladaptive response to predator odor, and this poses an ecological risk. The following patterns are recognized as eects of pesticides at the community and ecosystem levels: (1) induction of dominance by small species; (2) an increase of species richness and diversity; and (3) elongation of the food chain and reduction of energy transfer eciency from primary producers to top predators. # 2001 Elsevier Science Ltd. All rights reserved. Keywords: Pesticides; Zooplankton; Toxicity tests; Ecological interpretation
Contents Introduction .................................................................................................................................................................. 2 Individual-level responses.............................................................................................................................................. 2 Population-level responses............................................................................................................................................. 4 Biological interactions ................................................................................................................................................... 5 Community- and ecosystem-level responses .................................................................................................................. 6 Concluding remarks ...................................................................................................................................................... 7 References ..................................................................................................................................................................... 8
E-mail address: [email protected] (T. Hanazato). 0269-7491/01/$ - see front matter # 2001 Elsevier Science Ltd. All rights reserved. PII: S0269-7491(00)00110-X
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1. Introduction Surface water bodies are contaminated with many anthropogenic toxic chemicals that can aect their natural communities. It is necessary to assess the eects of these chemicals in order to conserve aquatic ecosystems. Among the anthropogenic chemicals, pesticides may cause the most serious problems because they are designed speci®cally to kill organisms (both the noxious target organisms and other non-target ones) and they are released into the natural environment intentionally. It has been widely documented that pesticide concentrations in the natural environment are often high enough to kill certain organisms (Hatakeyama et al., 1991, 1994) and aect the structure and function of natural communities (Helgen et al., 1988; Hatakeyama et al., 1990). Pesticides exert their impacts at multiple levels Ð including molecules, tissues, organs, individuals, populations and communities Ð and a variety of ecotoxicological tests have been designed to assess these eects (Cairns and Niederlehner, 1995). However, this assessment is hindered by the fact that natural ecosystems are diverse and the eects are complicated. Zooplankton are frequently used in ecotoxicological tests because they are one of the groups most sensitive to toxic chemicals and they occupy a central position in the lentic (standing water) food chain. The responses of zooplankton to toxicity tests are considered to be informative of relative impacts on the ecosystem as a whole. In the present paper I review the results of pesticide toxicity tests conducted with zooplankton at dierent levels of organization, ranging from individuals to communities, and discuss the eects of chemicals from an ecological perspective. 2. Individual-level responses In order to evaluate pesticide toxicity, standardized acute and chronic toxicity tests have been intensively conducted using cladocerans, in particular the species Daphnia (OECD, 1981; ASTM, 1994). In these tests, the direct toxic eects on animals are quanti®ed and standard test statistics, such as LC50, are determined. These tests indicate clearly that pesticides exert species-speci®c eects on zooplankton. This has been documented both in standard laboratory tests and in larger-scale mesocosms (small arti®cial ponds, or enclosures set up in lakes and ponds), where changes in zooplankton population density have been followed after a chemical application. For example, Hanazato (1991a) applied the carbamate insecticide carbaryl to outdoor experimental ponds in Japan and found that the large-sized cladoceran Daphnia was most sensitive to the chemical, followed in order by the medium-sized cladocerans Moina and Diaphanosoma, the small Bosmina, and ®nally roti-
fers, which show the lowest sensitivity. This same tendency for larger species to have the greatest sensitivity to carbaryl was observed in a lake in Ohio, USA (Havens, 1994a), and the same response was observed in Japan for the organophosphorus insecticide fenthion (Hanazato and Kasai, 1995). Gliwicz and Sieniawska (1986) have also shown in a laboratory study that there is a positive relationship between body size and sensitivity to lindane in cladocerans. This was considered to be due to the fact that cladocerans are most sensitive to chemicals just after molting, and that large-sized animals molt more times than smaller ones prior to reproduction. Size-related responses have been documented for a wide range of stressors, including acidi®cation, eutrophication and global warming (Geller and MuÈller, 1981; Richman and Dodson, 1983; Havens and Hanazato, 1993; Moore and Folt, 1993; Stemberger and Lazorchak, 1994). Rotifers are the zooplankton that are least sensitive to lake acidi®cation as well as pesticides, and Havens and Hanazato (1993) explained this by the fact that rotifers have high species diversity and thus are likely to contain both tolerant and sensitive taxa. The same explanation may apply to the tolerance of rotifers to pesticides. However, there are exceptions to the sizebased rule. Large cladoceran Simocepharus is less sensitive to the organochlorine insecticide endosulfan than the smaller Ceriodaphnia (Barry and Logan, 1998). Small Bosmina is more sensitive to DDT and arsenate than larger Daphnia (May-Passino and Novak, 1984), and small species are more sensitive to the pyrethroid insecticide fenvalerate than large ones among three species of Cladocera, Daphnia magna, D. galeata mendotae and Ceriodaphnia lacustris (Day and Kaushik, 1987). Sensitivity to pesticides also varies within species depending on genotype. Baird et al. (1990) determined LC50 values of eight clones of D. magna in acute toxicity tests and their LOEC (lowest eective concentration) in chronic toxicity tests for cadmium and 3,4-dichloroaniline (DCA), and found a large dierence in sensitivity among the clones. The LC50 ranged over three orders of magnitude for cadmium and one order for DCA, and the range of LOEC was within one order of magnitude for cadmium (0.6±6 mg/l) and two orders for DCA (25± 50 mg/l). The developmental stage of an animal also can aect its tolerance to pesticide exposure. Day and Kaushik (1987) performed acute toxicity tests and documented that juveniles of D. magna were more sensitive to fenvalerate than were adults. Hanazato (1991b) exposed D. ambigua to carbaryl for a short time (10 h) at dierent life stages ranging from egg to the third juvenile instar (neonate), and demonstrated that the ®rst instar was most sensitive to the chemical. Similarly, Breukelman (1932) found that the ®rst instar of D. magna was more sensitive to mercuric chloride than individuals at other
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developmental stages. This suggests that ®rst instar Daphnia should be used in standardized acute and chronic toxicity tests (OECD, 1981; ASTM, 1994). Enserink et al. (1990) demonstrated that smaller neonates are more sensitive than larger ones to cadmium, so that sensitivities to pesticides may also dier even among neonates depending on their body size. Neonatal size is aected by various factors, some of which can be readily controlled under laboratory conditions. These include food conditions, oxygen concentration, maternal clutch number, maternal body size, pesticide concentration and odor of potential predators. It has been suggested that size increases with increasing clutch number (Glazier, 1992; Lampert, 1993; Hanazato and Dodson, 1995). However, Glazier (1992), Lampert (1993) and Hanazato and Dodson (1995) have also pointed out that neonatal size depends on maternal body size more than clutch number. On the other hand, mother daphnids produce larger neonates under foodlimited conditions (Gliwicz and Guisande, 1992; Glazier, 1992) whereas they produce smaller neonates in water with low oxygen concentrations or with high insecticide concentrations (Hanazato and Dodson, 1995). Furthermore, neonatal size is controlled by the maternal response to odor released from invertebrate (e.g. the phantom midge Chaoborus, the backswimmer Notonecta) and vertebrate (®sh) predators (Dodson and Havel, 1988; Dodson, 1989; Stibor, 1992; Machacek, 1993). Aquatic organisms in the natural environment often are exposed to multiple pesticides and toxic chemicals simultaneously. The chemicals have complex eects on the organisms, sometimes being antagonistic or additive, and at other times synergistic (Lichtenstein et al., 1973; Cairns, 1983; Hoegland et al., 1993; Wijk et al., 1994). Therefore, the eect of a pesticide on a zooplankton species may be modi®ed by other chemicals contaminating the water body. In addition, natural stressors such as predation, food shortage, oxygen de®ciency, and high temperature can aect pesticide toxicity. Folt et al. (1999) tested the eects of sodium dodecyl sulfate, low food availability and high temperature on Daphnia and found that a combination of these stressors was more harmful than either one alone. Hanazato and Dodson (1995) demonstrated that predator odor, low oxygen concentration and an insecticide can reduce the population ®tness of Daphnia synergistically. The odor of predators induces morphological and behavioral changes in Daphnia as defense mechanisms, but it also reduces the population growth rate and tolerance to environmental stress (Havel and Dodson, 1987; Black and Dodson, 1990; Hanazato and Dodson, 1992; Hanazato, 1998a). This is a cost associated with the defense. These results indicate that natural stressors alter the sensitivity of zooplankton to pesticides, and that zooplankton species in the natural
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environment may be more sensitive to pesticides than the same species cultured under favorable conditions in the laboratory. Thus standard tests may underestimate toxicity. A good example of zooplankton being exposed to both anthropogenic and natural stressors is the diel vertical migration (DVM) of Daphnia in a pesticidecontaminated lake. The animals migrate downward into dark waters in the morning to avoid ®sh predation and upwards in the evening to obtain algae, which are a primary source of their food (Lampert, 1989; Fig. 1). Fish odor is a strong cue for inducing the DVM of Daphnia (Lampert and Loose, 1992), even when they must migrate downward into water that is low in oxygen and lacking in food resources. During DVM, Daphnia experience stress from oxygen and food de®ciency, and they become more sensitive to pesticides than expected from laboratory toxicity tests. Pesticides also can aect the feeding rate of zooplankton. Day and Kaushik (1987) measured the ®ltering rate (volume of water cleared of food per unit time) of the cladoceran Ceriodaphnia lacustris exposed to fenvalerate for 24 h. They found that the chemical concentration at which the ®ltering rate was reduced signi®cantly was 20-fold lower than the 48-h EC50 obtained in an acute toxicity test. This indicates that feeding behavior is quickly aected by pesticides (<24 h) even at sub-lethal concentrations. Lower feeding rates may result in reduced rates of growth and reproduction of the animals, which may be detected in chronic toxicity tests. Another behavior that may be aected by pesticide exposure is swimming. When this occurs it is a lifethreatening situation for zooplankton, which swim to obtain food, maintain their position in the water
Fig. 1. A scheme showing Daphnia performing diel vertical migration (DVM) when exposed to anthropogenic (pesticides) and natural stressors in lakes.
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column and avoid predators. Dodson et al. (1995) quanti®ed the behavioral response of Daphnia to carbaryl exposure by analyzing digitized three-dimensional video records of free-swimming animals exposed to the chemical. The Daphnia showed two kinds of swimming behavior in response to pesticide exposure: (1) `spinning' (extreme and continuous escape behavior) in response to acutely toxic levels of carbaryl, and (2) `irritation' (an increase in escape-like behavior) in response to sublethal levels. Pesticide exposure changes the swimming speed of various zooplankton species, sometimes increasing it (Preston et al., 1999) but at other times causing a decrease (Taylor et al., 1995). Such behavioral changes may be used as endpoints for assessing chemical toxicity (Goodrich and Lech, 1990; Di Delupis and Rotondo, 1998).
3. Population-level responses Eects of pesticides on zooplankton populations can be predicted from individual-level responses to pesticide exposure. As mentioned earlier, the neonate is the life stage most sensitive to pesticides, suggesting that among other things, populations composed of a large proportion of neonates are most sensitive to pesticides, and that a population vulnerable to pesticides is likely in a growing phase (i.e. producing neonates intensively). It was noted that small neonates are more sensitive to chemicals than large ones, and that neonate size is aected by various environmental factors. This suggests that a Daphnia population exposed to a low oxygen concentration and predator odor, which induce production of small neonates, will become more sensitive to pesticides, while a population under food limitation stress, which has larger neonates, will be less sensitive. It has also been demonstrated that pesticides induce the production of smaller neonates in Daphnia (Hanazato and Dodson, 1995), suggesting that a Daphnia population exposed to pesticides will become even more sensitive to chemicals. In the standardized chronic toxicity test using Daphnia proposed by the OECD (1981) and ASTM (1994), reproduction (number of ospring produced) of the test animals is analyzed. Analysis of reproduction may be useful for assessing the eects of chemicals on population growth. However, Hanazato (1998b) has cautioned that the number of ospring per clutch is not aected directly by toxic chemicals. Rather, it is a function of maternal body size, which in turn is governed by growth rate during the juvenile stage. Thus, a reduction of reproductive output by chemicals in a Daphnia results indirectly from the perturbation of juvenile growth (Fig. 2). This is an important distinction that should be understood by persons carrying out the tests.
Cladocerans usually display asexual reproduction, allowing their populations to increase rapidly. Sometimes sexual reproduction and formation of resting eggs (diapause stage) occurs, and this can be triggered by unfavorable environmental conditions. Sexual reproduction is an important process in population dynamics, and is initiated by production of males and haploid resting eggs. Dodson and co-workers have found that anthropogenic chemicals, including nonylphenol, dieldrin and atrazine, aect the production of Daphnia males and alter the sex ratio (Shurin and Dodson, 1997; Dodson et al., 1999a, b). This means that the chemicals control the Daphnia population dynamics not only by aecting individual survivorship and fecundity but also by determining the sex ratio, which in¯uences the success of resting egg production. Interestingly, the Daphnia sex ratio is 1±2 orders of magnitude more sensitive to atrazine than survivorship or fecundity (Dodson et al., 1999b). Dodson and Hanazato (1995) found that the yearly maximum percentage of males in populations of three Daphnia species in Lake Mendota, USA, decreased drastically from 1885 to 1975. They hypothesized that anthropogenic chemicals, which began to be widely used in the 1940s, altered the Daphnia sex ratio in the lake through their hormone-like eects. Eects of toxicants on the population dynamics of zooplankton have rarely been analyzed in the laboratory (van der Hoeven, 1990), but they have been observed in experimental ponds and enclosures (Hurlbert et al., 1972; Papst and Boyer, 1980; Kaushik et al., 1985; Day et al., 1987; Hanazato and Yasuno, 1987; Yasuno et al., 1988; Hanazato, 1991a). Hanazato and Yasuno (1987) analyzed summer population dynamics (changes in density, growth rate, birth and death rates) of the cladoceran Moina micrura in an experimental pond treated with the insecticide carbaryl, which has a high dissipation rate in water (Hanazato
Fig. 2. A schematic drawing showing the relationship among neonatal size (size at birth), adult size and clutch size in zooplankton. A large neonate grows to a large adult, which in turn produces many ospring. A small neonate matures at a small size and produces fewer eggs. Toxic chemicals such as pesticides reduce the growth rate of neonates and cause the neonates to mature at a smaller size with subsequently fewer eggs (redrawn from Hanazato, 1998b).
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and Yasuno, 1990a). The population at ®rst seemed to be eliminated by the chemical, but it redeveloped 2 days after the application and showed rapid growth with a high birth rate. This phenomenon was linked with improved food resource conditions, due to the local extinction of less-tolerant competitors. Similar increases in of population size after pesticide application has often been observed in some zooplankton species (Hurlbert et al., 1972; Papst and Boyer, 1980; Kaushik et al., 1985; Day et al., 1987; Yasuno et al., 1988; Hanazato and Yasuno, 1990b; Hanazato, 1991a). The eects of pesticides on natural populations may dier depending on their stage of development. In a pond experiment conducted by Hanazato and Yasuno (1990a), the rotifer Keratella valga was able to recover from losses due to carbaryl application when the population was in a growing phase. On the other hand, recovery did not occur when the same treatment was applied during a period of population decline. Hanazato and Yasuno (1990a) considered that the population had some intrinsic factors controlling its cycle, and that these factors might have prevented a response to the sudden change in environmental conditions caused by application of the chemical, such as the increased available food density after disappearance of competitors. 4. Biological interactions Biological interactions such as competition and predation can play a major role in regulating the population dynamics of zooplankton. Pesticides also aect these interactions and cause secondary eects on the structure and function of the biological community. Thus, analysis of the eects of pesticides on biological interactions is an important issue in ecotoxicological studies. It was mentioned earlier that larger animals tend to be more sensitive to pesticides than smaller ones. Interestingly, the same tendency can be applied to competitive superiority (Hanazato, 1998c). Namely, large animals such as Daphnia are vulnerable to pesticide contamination but they are superior in competition with other zooplankters, while small animals such as Bosmina and rotifers are tolerant to the chemicals but inferior in terms of competition. In the case of competition between Daphnia and rotifers, however, not only exploitative competition but also interference competition (where large herbivores damage and/or kill smaller ones due their feeding activities) is an important mechanism (Burns and Gilbert, 1986; Gilbert, 1988; MacIsaac and Gilbert, 1991). There is evidence that very large daphnids also can eat small rotifers. As a result, relatively low pesticide concentrations, which damage only large taxa (Daphnia), may aect the population dynamics of small zooplankton species
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indirectly through competitive relationships. In fact, it is often observed in zooplankton communities that rotifers bloom after destruction of the Daphnia population by pesticides (Hurlbert et al., 1972; Papst and Boyer, 1980; Kaushik et al., 1985; Day et al., 1987; Yasuno et al., 1988). Zooplankters change their swimming behavior when exposed to pesticides, as mentioned earlier. Such behavioral changes may alter predator±prey relationships. Dodson et al. (1995) observed that the `spinning' behavior of Daphnia induced by carbaryl increased the probability of Daphnia being eaten by ®sh. This suggests that even sub-lethal concentrations of pesticides contribute to Daphnia population crashes in the presence of predators (Dodson et al., 1995). Preston et al. (1999) demonstrated that the toxicant pentachlorophenol increases swimming speed in the rotifer Brachionus calyci¯orus, thus increasing the encounter rate of the prey animals with their predators, and consequently reducing prey survivorship. Similar indirect eects of pesticides, which increase risk for prey organisms, have been reported for grass shrimp (Farr, 1977) and juvenile guppy (Brown et al., 1985). In contrast, the insecticide lindane seems to reduce predation risk in D. magna by suppressing the animals' swimming activity, which then reduces their encounter rate with the predator Hydra (Taylor et al., 1995). This same eect, of course, might be detrimental if it resulted in a lowered amount of food intake or reduced ability to maintain a vertical position in the water column of a lake. In any event, it can be concluded that, even at sub-lethal concentrations, pesticides may aect the survivorship of zooplankton in the presence of predators by controlling prey behavior. The fact that animals exposed to pesticides have a higher predation risk provides another important consideration in ecotoxicology. That is, there is an increase in the rate of pesticide uptake by predators that feed on pesticide-exposed animals, thus promoting biomagni®cation in the aquatic food web. Daphnia and other cladocerans can also develop morphological responses to predators. Daphnia develops protuberant structures on its carapace such as neckteeth, high helmets and long tailspines when exposed to chemicals released from predators such as the larva of the phantom midge Chaoborus, the backswimmer Notonecta, the predacious cladoceran Leptodora, and ®sh (Grant and Bayly, 1981; Krueger and Dodson, 1981; Hebert and Grewe, 1985; Dodson, 1989; Tollrian, 1990, 1994; Fig. 3). These structures reduce the vulnerability of Daphnia to the predators by making handling and ingestion more dicult (Havel and Dodson, 1984; Mort, 1986; Parejko, 1990). Hanazato (1991c, 1992) and Hanazato and Dodson (1993) found that the same morphological changes can be induced by organophosphorus and carbamate insecticides in several species of cyclomorphic Daphnia, if the animals are
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exposed to high concentrations of the chemicals during the ®nal embryo to the ®rst instar stages. Furthermore, Hanazato (1995) and Hanazato and Dodson (1992, 1995) observed that sub-lethal concentrations of the chemicals aect the morphology of Daphnia when the animals are exposed to both insecticides and predator (Chaoborus) odor simultaneously. Barry (1998) documented the induction of a protuberant structure (crest) by the insecticide endosulfan in Daphnia longicephala. Induction occurred even when the insecticide was applied alone (without the predator odor) and at a very low concentration (more than three orders of magnitude lower than the EC50). Because these morphologies are anti-predator defenses, such eects of pesticides must alter predator±prey relationships. The induction of protuberant structures in Daphnia is a positive eect of natural predator odor, which decreases the prey's mortality due to predation. However, there is a negative eect as well, since animals with anti-predator devices have lower clutch sizes and increased maturation time. In Daphnia this reduces population growth rate (Hanazato and Dodson, 1992), probably due to the energy costs associated with the morphological response. Hanazato and Dodson (1992, 1995) demonstrated that carbaryl reduces the growth rate and productivity of Daphnia more in the presence
Fig. 3. Daphnia species that change their morphology in response to predator odor [a: Daphnia pulex; b: D. galeata mendotae; c: D. retrocurva; d: D. ambigua; e: D. carinata; in each pair, the left picture shows the typical morph (grown without the odor) and that on the right indicates the morph bearing protuberant structures (grown with the odor)] and predators releasing the odor [f: phantom midge larva (Chaoborus: 2±10 mm in body length; g: backswimmer (Notonectidae: ca. 10 mm); h: predacious Cladocera (Leptodora: 2±10 mm)].
of predator odor, and found synergism in the combined eect of the anthropogenic and natural chemicals. These observations suggest that pesticides might also disturb natural organic chemical communication in aquatic communities (Hanazato, 1999). There is increasing evidence that aquatic organisms communicate with one another using organic chemicals (Larsson and Dodson, 1993). This is a largely unstudied but potentially critical area of pesticide impacts. 5. Community- and ecosystem-level responses Community-level responses to pesticides have often been assessed in enclosures and experimental ponds to which pesticides are experimentally applied (Kennedy et al., 1995), and in natural ponds that are occasionally contaminated with pesticides at relatively high concentrations (Helgen et al., 1988). The most frequently reported response is a change in community structure from dominance by Daphnia to dominance by small zooplankters such as rotifers and Bosmina (Hurlbert et al., 1972; Papst and Boyer, 1980; Kaushik et al., 1985; Day et al., 1987; Yasuno et al., 1988; Hanazato and Yasuno, 1990b; Hanazato, 1991a). Therefore, it is suggested that pesticides induce dominance of small zooplankton species, and consequently they reduce the mean body size of individuals in the community (Havens and Hanazato, 1993; Hanazato, 1998c). Because Daphnia depresses the populations of most small zooplankton taxa through competition, a Daphniadominant community has low species richness. However, if the Daphnia population has been destroyed by pesticides, small zooplankton species increase in abundance, and the species richness of the community becomes higher (Hanazato, 1994). The increased diversity due to pesticide contamination (Hanazato, 1998c) diers from the response of the zooplankton to other stressors such as heavy metals and acidi®cation (a kind of chemical stress), which reduce diversity (Sprules, 1975; Confer and Kaaret, 1983; Schindler, 1990; Havens, 1991, 1994b; Locke, 1992; Locke and Sprules, 1994). Hanazato (1998c) explained that heavy metals and acidi®cation exert continuous stress (direct toxic eects), so that the species richness and diversity are maintained at low levels. On the other hand, pesticides have a relatively short life span in the environment and zooplankton communities are exposed to the chemical stress for a shorter period of time. Therefore, the process of community recovery from destruction by pesticides is usually observed for populations of diverse small taxa, and the community diversity increases during the process (Hanazato, 1998c). Hanazato (1998c) therefore concluded that short-term or intermittent perturbation by chemical stress (in general) would increase the diversity of natural communities. Hanazato (1997)
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Fig. 4. The main pathways of carbon and energy ¯ow from algae through zooplankton to ®sh in lake ecosystems contaminated and uncontaminated with pesticides (redrawn from Hanazato, 1998c).
has also demonstrated that repeated application of an insecticide at a moderately toxic level to a rotifer community maintains the species richness at a high level. Probably the chemical suppress the population growth of sensitive species moderately and allows an increase of species that are tolerant to the chemical but inferior in terms of competition. This is another case of pesticides increasing the biodiversity of a plankton community, and is consistent with the intermediate disturbance hypothesis of Connell (1978), and Menge and Sutherland (1987). Changes in zooplankton community structure induced by pesticides may also aect the functioning of lake ecosystems (Hanazato, 1998c). In an uncontaminated lake containing abundant Daphnia, a large proportion of phytoplankton primary production is transferred to top predators (®sh and birds) through the algae! zooplankton!®sh pathway. In a lake that is impacted by insecticides, where there are many small zooplankton and few Daphnia, a smaller proportion of primary production is transferred upwards, due to longer food chains that include invertebrate predators (Fig. 4). Because energy loss occurs during transfer from one trophic level to another, eciency of energy transfer from primary producers to top predator is lower in a pesticide-stressed ecosystem than in an ecosystem uncontaminated by chemicals. Hanazato (1998c) suggested that reduced energy transfer eciency is a common eect of various anthropogenic stresses such as contamination with insecticides, herbicides and heavy metals, acidi®cation, eutrophication and global warming.
6. Concluding remarks In standardized toxicity tests for zooplankton, survival and reproduction are the attributes most often observed. However, analysis of these factors alone is not enough for assessing the full ecological impacts of toxicants (Forbes and Calow, 1999). Analysis of toxicant eects on other life history characteristics such as ospring size, morphology and behavior is necessary to evaluate the eects on populations, communities and ecosystems. In the natural environment, aquatic organisms are exposed to multiple pesticides simultaneously, and therefore multiple eects must be assessed (Cairns, 1983). Furthermore, the organisms are exposed to a combination of pesticides and natural stressors such as high and low temperature, food shortage, oxygen depletion and predator odor, whose eects may be worse in combination than alone. Few studies have addressed how a combination of stressors aects aquatic organisms (Folt et al., 1999) and pursuit of such studies should be encouraged. Ecosystems are comprised of a huge number of species that have complex interactions with one another. Biological interactions play an important role in maintaining ecosystems, and pesticides or other toxicants can in¯uence them. Therefore, eects of toxicants on biological interactions need to be studied in ecotoxicology investigations. Up to now, however, such studies have been few. The presence of biological interactions mediated by a chemical cue released from organisms (e.g. development
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of protuberant structure in Daphnia in response to odors from potential predators) was discovered in the 1980s (Larsson and Dodson, 1993). In the 1990s, it was demonstrated for the ®rst time that pesticides disturb the natural organic chemical communication between predator and prey (Hanazato, 1999). Hanazato (1999) then considered that the eect of pesticides was similar to that of environmental endocrine disrupters Ð anthropogenic chemicals that disturb chemical communication between tissues or organs in the bodies of animals. These ®ndings suggested the presence of a new ecological risk posed by pesticides. There are many common patterns in the structural and functional responses of ecosystems to pesticides and other anthropogenic stressors such as acidi®cation, heavy metal pollution and climatic warming, as originally hypothesized by Odum (1985). Identi®cation of these common patterns is helpful for predicting the ecological risks of environmental problems created by human activities, and is thus an important research topic in both ecotoxicology and fundamental ecology. Accordingly, more emphasis needs to be placed on ecological analysis in future ecotoxicology research.
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Hatakeyama, S., Shiraishi, H., Kobayashi, N., 1990. Eects of aerial spraying of insecticides on nontarget mafrobenthos in a mountain stream. Ecotoxicology and Environmental Safety 19, 254±270. Hatakayama, S., Shiraishi, H., Sugaya, Y., 1991. Monitoring of the overall pesticide toxicity of river water to aquatic organisms using a freshwater shrimp, Paratya compressa improvisa. Chemosphere 22, 229±235. Hatakayama, S., Fukushima, S., Kasai, F., Shiraishi, H., 1994. Assessment of herbivide eects on algal production in the Kokai River (Japan) using model stream and Selenastrum bioassay. Ecotoxicology 3, 143±156. Havel, J.E., Dodson, S.I., 1984. Chaoborus predation on typical and spined morphs of Daphnia pulex: behavioral observation. Limnology and Oceanography 29, 487±494. Havel, J.E., Dodson, S.I., 1987. Reproductive costs of Chaoborus induced polymorphism in Daphnia pulex. Hydrobiologia 150, 273± 281. Havens, K.E., 1991. Crustacean zooplankton food web structure in lakes of varying acidity. Canadian Journal of Fisheries and Aquatic Sciences 48, 1846±1852. Havens, K.E., 1994a. An experimental comparison of the eects of two chemical stressors on a freshwater zooplankton assemblage. Environmental Pollution 84, 245±251. Havens, K.E., 1994b. Experimental perturbation of a freshwater plankton community: a test of hypotheses regarding the eects of stress. Oikos 69, 147±153. Havens, K.E., Hanazato, T., 1993. Zooplankton community responses to chemical stressors: a comparison of results from acidi®cation and pesticide contamination research. Environmental Pollution 82, 277± 288. Hebert, P.D.N., Grewe, D.M., 1985. Chaoborus-induced shifts in the morphology of Daphnia ambigua. limnology and Oceanography 30, 1291±1297. Helgen, J.C., Larson, N.J., Anderson, R.L., 1988. Response of zooplankton and Chaoborus to temephos in a natural pond and in the laboratory. Archives of Environmental Contamination and Toxicology 17, 459±471. Hoegland, K.D., Drenner, R.W., Smith, J.D., Cross, D.R., 1993. Freshwater community response to mixtures of agricultural pesticides: Eects of atrazine and bifenthrin. Environmental Toxicology and Chemistry 12, 627±637. van der Hoeven, N., 1990. Eect of 3,4-dechloroaniline and metavanadate on Daphnia populations. Ecotoxicology and Environmental Safety 20, 53±70. Hurlbert, S.H., Mulla, M.S., Wilson, H.R., 1972. Eects of an organophosphorus insecticide on the phytoplankton, zooplankton, and insect populations of freshwater ponds. Ecological Monographs 42, 269±299. Kaushik, N.K., Stephenson, G.L., Solomon, K.R., Day, K.E., 1985. Impact of permethrin on zooplankton communities in limnocorals. Canadian Journal of Fisheries and Aquatic Sciences 42, 77±85. Kennedy, J.H., Johnson, Z.B., Wise, P.D., Johnson, P.C., 1995. Model aquatic ecosystems in ecotoxicological research: consideration of design, implementation, and analysis. In: Homan, D.J., Battner, B.A., Burton Jr., G.A., Cairns Jr., J. (Eds.), Handbook and Ecotoxicology. CRC Press, FL, pp. 117±162. Krueger, D.A., Dodson, S.I., 1981. Embryological induction and predation ecology in Daphnia pulex. Limnology and Oceanography 26, 219±223. Lampert, W., 1989. The adaptive signi®cance of diel vertical migration of zooplankton. Functional Ecology 3, 21±27. Lampert, W., 1993. Phenotypic plasticity of the size at ®rst reproduction in Daphnia: the importance of maternal size. Ecology 74, 1455± 1466. Lampert, W., Loose, C.J., 1992. Plankton towers: bridging the gap between laboratory and ®eld experiments. Archiv fur Hydrobiolgie 126, 53±66.
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Larsson, P., Dodson, S.I., 1993. Chemical communication in planktonic animals. Archiv fur Hydrobiologie 129, 129±155. Lichtenstein, E.P., Liang, T.T., Anderegg, B.N., 1973. Synergism of insecticides by herbicides. Science 181, 847±849. Locke, A., 1992. Factors in¯uencing community structure along stress gradients: zooplankton responses to acidi®cation. Ecology 73, 903±909. Locke, A., Sprules, W.G., 1994. Eects of lake acidi®cation and recovery on the stability of zooplankton food webs. Ecology 75, 498±506. Machacak, J., 1993. Comparison of the response of Daphnia galeata and Daphnia obtusa to ®sh produced chemical substance. Limnology and Oceanography 38, 1544±1550. MacIsaac, H.J., Gilbert, J.J., 1991. Discrimination between exploitative and interference competition between Cladocera and Keratella cochlearis. Ecology 72, 924±937. May-Passino, D.R., Novak, A.J., 1984. Toxicity of arsenate and DDT to the cladoceran Bosmina longirostris. Bulletin of Environmental Contamination and Toxicology 33, 325±329. Menge, B.A., Sutherland, J.P., 1987. Community regulation: variation in disturbance, competition, and predation in relation to environmental stress and recruitment. American Naturalist 130, 730±757. Moore, M.A., Folt, C.L., 1993. Zooplankton body size and community structure: eects of thermal and toxicant stress. Trends in Ecology and Evolution 8, 178±183. Mort, M.A., 1986. Chaoborus predation and the function of phenotypic variation in Daphnia. Hydrobiologia 133, 39±44. Odum, E.P., 1985. Trends expected in stressed ecosystems. BioScience 35, 419±422. OECD, 1981. Daphnia sp. 14 day reproduction test (including an acute immobilization test). (Guidelines for the testing of chemicals No. 202). OECD, Paris Papst, M.H., Boyer, M.G., 1980. Eects of two organophosphorus insecticides on the chlorophyll a and pheopigment concentration s of standing ponds. Hydrobiologia 69, 245±250. Parejko, K., 1990. Mochlonyx predation on typical and spined juvenile Daphnia. Freshwater Biology 25, 211±217.
Preston, B.L., Cecchine, G., Snell, T.W., 1999. Eects of pentachlorophenol on predator avoidance behavior of the rotifer Brachionus calyci¯orus. Aquatic Toxicology 44, 201±212. Richman, S., Dodson, S.I., 1983. The eect of food quality on feeding and respiration by Daphnia and Diaptomus. Limnology and Oceanography 28, 948±956. Schindler, D.W., 1990. Experimental perturbations of whole lakes as tests of hypotheses concerning ecosystem structure and function. Oikos 57, 25±41. Shurin, J.B., Dodson, S.I., 1997. Sublethal toxic eects of cyanobacteria and nonylphenol on environmental sex determination and development in Daphnia. Environmental Toxicology and Chemistry 16, 1269±1276. Sprules, W.G., 1975. Midsummer crustacean zooplankton communities in acid-stressed lakes. Journal of the Fisheries Research Board of Canada 32, 389±395. Stemberger, R.S., Lazorchak, J.M., 1994. Zooplankton assemblage responses to disturbance gradients. Canadian Journal of Fisheries and Aquatic Sciences 51, 2435±2447. Stibor, H., 1992. Predator induced life-history shifts in a freshwater cladoceran. Oecologia 92, 162±165. Taylor, E.J., Morrison, J.E., Blockwell, S.J., Tarr, A., Pascoe, D., 1995. Eects of lindane on the predator-prey interaction between Hydra oligactis Pallas and Daphnia magna Strauss. Archives of Environmental Contamination and Toxicology 29, 291±296. Tollrian, R., 1990. Predator-induced helmet formation in Daphnia cucullata (Sars). Archiv fur Hydrobiologie 119, 191±196. Tollrian, R., 1994. Fish-kairomone induced morphological changes in Daphnia lumholzi (Sars). Archiv fur Hydrobiologie 130, 69±75. van Wijk, R.J., Postma, J.F., van Houwelingen, H., 1994. Joint toxicity of ethyleneamines to algae, daphnids and ®sh. Environmental Toxicology and Chemistry 13, 167±171. Yasuno, M., Hanazato, T., Iwakuma, T., Takamura, K., Ueno, R., Takamura, N., 1988. Eects of permethrin on phytoplankton and zooplankton in an enclosure ecosystem in a pond. Hydrobiologia 159, 247±258.
www.elsevier.com/locate/envpol
Review
Pesticide eects on freshwater zooplankton: an ecological perspective T. Hanazato Suwa Hydrobiological Station, Shinshu University, 5-2-4 Kogandori, Suwa 392-0027, Japan Received 1 February 2000; accepted 11 March 2000
``Capsule'': Eects of pesticides on freshwater zooplankton are considered from an ecological perspective. Abstract The eects of pesticides on zooplankton are reviewed and their ecological signi®cance is discussed. Toxicity is shown to vary depending on animal species, genotype, life stage, and size at birth. Natural stresses such as food shortage, oxygen depletion and odors of potential predators can also aect toxicity. Populations in the growth phase are vulnerable to pesticides but have the potential to recover rapidly from the damage. Pesticides may aect the population dynamics by controlling individual survival and reproduction, and by altering the sex ratio. Furthermore, toxic chemicals may control predation risk by changing swimming behavior and body morphology, and this in turn in¯uences the population dynamics. Many zooplankton display morphological and behavioral responses to predators when exposed to their odor-producing chemicals. However, pesticides induce a maladaptive response to predator odor, and this poses an ecological risk. The following patterns are recognized as eects of pesticides at the community and ecosystem levels: (1) induction of dominance by small species; (2) an increase of species richness and diversity; and (3) elongation of the food chain and reduction of energy transfer eciency from primary producers to top predators. # 2001 Elsevier Science Ltd. All rights reserved. Keywords: Pesticides; Zooplankton; Toxicity tests; Ecological interpretation
Contents Introduction .................................................................................................................................................................. 2 Individual-level responses.............................................................................................................................................. 2 Population-level responses............................................................................................................................................. 4 Biological interactions ................................................................................................................................................... 5 Community- and ecosystem-level responses .................................................................................................................. 6 Concluding remarks ...................................................................................................................................................... 7 References ..................................................................................................................................................................... 8
E-mail address: [email protected] (T. Hanazato). 0269-7491/01/$ - see front matter # 2001 Elsevier Science Ltd. All rights reserved. PII: S0269-7491(00)00110-X
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T. Hanazato / Environmental Pollution 112 (2001) 1±10
1. Introduction Surface water bodies are contaminated with many anthropogenic toxic chemicals that can aect their natural communities. It is necessary to assess the eects of these chemicals in order to conserve aquatic ecosystems. Among the anthropogenic chemicals, pesticides may cause the most serious problems because they are designed speci®cally to kill organisms (both the noxious target organisms and other non-target ones) and they are released into the natural environment intentionally. It has been widely documented that pesticide concentrations in the natural environment are often high enough to kill certain organisms (Hatakeyama et al., 1991, 1994) and aect the structure and function of natural communities (Helgen et al., 1988; Hatakeyama et al., 1990). Pesticides exert their impacts at multiple levels Ð including molecules, tissues, organs, individuals, populations and communities Ð and a variety of ecotoxicological tests have been designed to assess these eects (Cairns and Niederlehner, 1995). However, this assessment is hindered by the fact that natural ecosystems are diverse and the eects are complicated. Zooplankton are frequently used in ecotoxicological tests because they are one of the groups most sensitive to toxic chemicals and they occupy a central position in the lentic (standing water) food chain. The responses of zooplankton to toxicity tests are considered to be informative of relative impacts on the ecosystem as a whole. In the present paper I review the results of pesticide toxicity tests conducted with zooplankton at dierent levels of organization, ranging from individuals to communities, and discuss the eects of chemicals from an ecological perspective. 2. Individual-level responses In order to evaluate pesticide toxicity, standardized acute and chronic toxicity tests have been intensively conducted using cladocerans, in particular the species Daphnia (OECD, 1981; ASTM, 1994). In these tests, the direct toxic eects on animals are quanti®ed and standard test statistics, such as LC50, are determined. These tests indicate clearly that pesticides exert species-speci®c eects on zooplankton. This has been documented both in standard laboratory tests and in larger-scale mesocosms (small arti®cial ponds, or enclosures set up in lakes and ponds), where changes in zooplankton population density have been followed after a chemical application. For example, Hanazato (1991a) applied the carbamate insecticide carbaryl to outdoor experimental ponds in Japan and found that the large-sized cladoceran Daphnia was most sensitive to the chemical, followed in order by the medium-sized cladocerans Moina and Diaphanosoma, the small Bosmina, and ®nally roti-
fers, which show the lowest sensitivity. This same tendency for larger species to have the greatest sensitivity to carbaryl was observed in a lake in Ohio, USA (Havens, 1994a), and the same response was observed in Japan for the organophosphorus insecticide fenthion (Hanazato and Kasai, 1995). Gliwicz and Sieniawska (1986) have also shown in a laboratory study that there is a positive relationship between body size and sensitivity to lindane in cladocerans. This was considered to be due to the fact that cladocerans are most sensitive to chemicals just after molting, and that large-sized animals molt more times than smaller ones prior to reproduction. Size-related responses have been documented for a wide range of stressors, including acidi®cation, eutrophication and global warming (Geller and MuÈller, 1981; Richman and Dodson, 1983; Havens and Hanazato, 1993; Moore and Folt, 1993; Stemberger and Lazorchak, 1994). Rotifers are the zooplankton that are least sensitive to lake acidi®cation as well as pesticides, and Havens and Hanazato (1993) explained this by the fact that rotifers have high species diversity and thus are likely to contain both tolerant and sensitive taxa. The same explanation may apply to the tolerance of rotifers to pesticides. However, there are exceptions to the sizebased rule. Large cladoceran Simocepharus is less sensitive to the organochlorine insecticide endosulfan than the smaller Ceriodaphnia (Barry and Logan, 1998). Small Bosmina is more sensitive to DDT and arsenate than larger Daphnia (May-Passino and Novak, 1984), and small species are more sensitive to the pyrethroid insecticide fenvalerate than large ones among three species of Cladocera, Daphnia magna, D. galeata mendotae and Ceriodaphnia lacustris (Day and Kaushik, 1987). Sensitivity to pesticides also varies within species depending on genotype. Baird et al. (1990) determined LC50 values of eight clones of D. magna in acute toxicity tests and their LOEC (lowest eective concentration) in chronic toxicity tests for cadmium and 3,4-dichloroaniline (DCA), and found a large dierence in sensitivity among the clones. The LC50 ranged over three orders of magnitude for cadmium and one order for DCA, and the range of LOEC was within one order of magnitude for cadmium (0.6±6 mg/l) and two orders for DCA (25± 50 mg/l). The developmental stage of an animal also can aect its tolerance to pesticide exposure. Day and Kaushik (1987) performed acute toxicity tests and documented that juveniles of D. magna were more sensitive to fenvalerate than were adults. Hanazato (1991b) exposed D. ambigua to carbaryl for a short time (10 h) at dierent life stages ranging from egg to the third juvenile instar (neonate), and demonstrated that the ®rst instar was most sensitive to the chemical. Similarly, Breukelman (1932) found that the ®rst instar of D. magna was more sensitive to mercuric chloride than individuals at other
T. Hanazato / Environmental Pollution 112 (2001) 1±10
developmental stages. This suggests that ®rst instar Daphnia should be used in standardized acute and chronic toxicity tests (OECD, 1981; ASTM, 1994). Enserink et al. (1990) demonstrated that smaller neonates are more sensitive than larger ones to cadmium, so that sensitivities to pesticides may also dier even among neonates depending on their body size. Neonatal size is aected by various factors, some of which can be readily controlled under laboratory conditions. These include food conditions, oxygen concentration, maternal clutch number, maternal body size, pesticide concentration and odor of potential predators. It has been suggested that size increases with increasing clutch number (Glazier, 1992; Lampert, 1993; Hanazato and Dodson, 1995). However, Glazier (1992), Lampert (1993) and Hanazato and Dodson (1995) have also pointed out that neonatal size depends on maternal body size more than clutch number. On the other hand, mother daphnids produce larger neonates under foodlimited conditions (Gliwicz and Guisande, 1992; Glazier, 1992) whereas they produce smaller neonates in water with low oxygen concentrations or with high insecticide concentrations (Hanazato and Dodson, 1995). Furthermore, neonatal size is controlled by the maternal response to odor released from invertebrate (e.g. the phantom midge Chaoborus, the backswimmer Notonecta) and vertebrate (®sh) predators (Dodson and Havel, 1988; Dodson, 1989; Stibor, 1992; Machacek, 1993). Aquatic organisms in the natural environment often are exposed to multiple pesticides and toxic chemicals simultaneously. The chemicals have complex eects on the organisms, sometimes being antagonistic or additive, and at other times synergistic (Lichtenstein et al., 1973; Cairns, 1983; Hoegland et al., 1993; Wijk et al., 1994). Therefore, the eect of a pesticide on a zooplankton species may be modi®ed by other chemicals contaminating the water body. In addition, natural stressors such as predation, food shortage, oxygen de®ciency, and high temperature can aect pesticide toxicity. Folt et al. (1999) tested the eects of sodium dodecyl sulfate, low food availability and high temperature on Daphnia and found that a combination of these stressors was more harmful than either one alone. Hanazato and Dodson (1995) demonstrated that predator odor, low oxygen concentration and an insecticide can reduce the population ®tness of Daphnia synergistically. The odor of predators induces morphological and behavioral changes in Daphnia as defense mechanisms, but it also reduces the population growth rate and tolerance to environmental stress (Havel and Dodson, 1987; Black and Dodson, 1990; Hanazato and Dodson, 1992; Hanazato, 1998a). This is a cost associated with the defense. These results indicate that natural stressors alter the sensitivity of zooplankton to pesticides, and that zooplankton species in the natural
3
environment may be more sensitive to pesticides than the same species cultured under favorable conditions in the laboratory. Thus standard tests may underestimate toxicity. A good example of zooplankton being exposed to both anthropogenic and natural stressors is the diel vertical migration (DVM) of Daphnia in a pesticidecontaminated lake. The animals migrate downward into dark waters in the morning to avoid ®sh predation and upwards in the evening to obtain algae, which are a primary source of their food (Lampert, 1989; Fig. 1). Fish odor is a strong cue for inducing the DVM of Daphnia (Lampert and Loose, 1992), even when they must migrate downward into water that is low in oxygen and lacking in food resources. During DVM, Daphnia experience stress from oxygen and food de®ciency, and they become more sensitive to pesticides than expected from laboratory toxicity tests. Pesticides also can aect the feeding rate of zooplankton. Day and Kaushik (1987) measured the ®ltering rate (volume of water cleared of food per unit time) of the cladoceran Ceriodaphnia lacustris exposed to fenvalerate for 24 h. They found that the chemical concentration at which the ®ltering rate was reduced signi®cantly was 20-fold lower than the 48-h EC50 obtained in an acute toxicity test. This indicates that feeding behavior is quickly aected by pesticides (<24 h) even at sub-lethal concentrations. Lower feeding rates may result in reduced rates of growth and reproduction of the animals, which may be detected in chronic toxicity tests. Another behavior that may be aected by pesticide exposure is swimming. When this occurs it is a lifethreatening situation for zooplankton, which swim to obtain food, maintain their position in the water
Fig. 1. A scheme showing Daphnia performing diel vertical migration (DVM) when exposed to anthropogenic (pesticides) and natural stressors in lakes.
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column and avoid predators. Dodson et al. (1995) quanti®ed the behavioral response of Daphnia to carbaryl exposure by analyzing digitized three-dimensional video records of free-swimming animals exposed to the chemical. The Daphnia showed two kinds of swimming behavior in response to pesticide exposure: (1) `spinning' (extreme and continuous escape behavior) in response to acutely toxic levels of carbaryl, and (2) `irritation' (an increase in escape-like behavior) in response to sublethal levels. Pesticide exposure changes the swimming speed of various zooplankton species, sometimes increasing it (Preston et al., 1999) but at other times causing a decrease (Taylor et al., 1995). Such behavioral changes may be used as endpoints for assessing chemical toxicity (Goodrich and Lech, 1990; Di Delupis and Rotondo, 1998).
3. Population-level responses Eects of pesticides on zooplankton populations can be predicted from individual-level responses to pesticide exposure. As mentioned earlier, the neonate is the life stage most sensitive to pesticides, suggesting that among other things, populations composed of a large proportion of neonates are most sensitive to pesticides, and that a population vulnerable to pesticides is likely in a growing phase (i.e. producing neonates intensively). It was noted that small neonates are more sensitive to chemicals than large ones, and that neonate size is aected by various environmental factors. This suggests that a Daphnia population exposed to a low oxygen concentration and predator odor, which induce production of small neonates, will become more sensitive to pesticides, while a population under food limitation stress, which has larger neonates, will be less sensitive. It has also been demonstrated that pesticides induce the production of smaller neonates in Daphnia (Hanazato and Dodson, 1995), suggesting that a Daphnia population exposed to pesticides will become even more sensitive to chemicals. In the standardized chronic toxicity test using Daphnia proposed by the OECD (1981) and ASTM (1994), reproduction (number of ospring produced) of the test animals is analyzed. Analysis of reproduction may be useful for assessing the eects of chemicals on population growth. However, Hanazato (1998b) has cautioned that the number of ospring per clutch is not aected directly by toxic chemicals. Rather, it is a function of maternal body size, which in turn is governed by growth rate during the juvenile stage. Thus, a reduction of reproductive output by chemicals in a Daphnia results indirectly from the perturbation of juvenile growth (Fig. 2). This is an important distinction that should be understood by persons carrying out the tests.
Cladocerans usually display asexual reproduction, allowing their populations to increase rapidly. Sometimes sexual reproduction and formation of resting eggs (diapause stage) occurs, and this can be triggered by unfavorable environmental conditions. Sexual reproduction is an important process in population dynamics, and is initiated by production of males and haploid resting eggs. Dodson and co-workers have found that anthropogenic chemicals, including nonylphenol, dieldrin and atrazine, aect the production of Daphnia males and alter the sex ratio (Shurin and Dodson, 1997; Dodson et al., 1999a, b). This means that the chemicals control the Daphnia population dynamics not only by aecting individual survivorship and fecundity but also by determining the sex ratio, which in¯uences the success of resting egg production. Interestingly, the Daphnia sex ratio is 1±2 orders of magnitude more sensitive to atrazine than survivorship or fecundity (Dodson et al., 1999b). Dodson and Hanazato (1995) found that the yearly maximum percentage of males in populations of three Daphnia species in Lake Mendota, USA, decreased drastically from 1885 to 1975. They hypothesized that anthropogenic chemicals, which began to be widely used in the 1940s, altered the Daphnia sex ratio in the lake through their hormone-like eects. Eects of toxicants on the population dynamics of zooplankton have rarely been analyzed in the laboratory (van der Hoeven, 1990), but they have been observed in experimental ponds and enclosures (Hurlbert et al., 1972; Papst and Boyer, 1980; Kaushik et al., 1985; Day et al., 1987; Hanazato and Yasuno, 1987; Yasuno et al., 1988; Hanazato, 1991a). Hanazato and Yasuno (1987) analyzed summer population dynamics (changes in density, growth rate, birth and death rates) of the cladoceran Moina micrura in an experimental pond treated with the insecticide carbaryl, which has a high dissipation rate in water (Hanazato
Fig. 2. A schematic drawing showing the relationship among neonatal size (size at birth), adult size and clutch size in zooplankton. A large neonate grows to a large adult, which in turn produces many ospring. A small neonate matures at a small size and produces fewer eggs. Toxic chemicals such as pesticides reduce the growth rate of neonates and cause the neonates to mature at a smaller size with subsequently fewer eggs (redrawn from Hanazato, 1998b).
T. Hanazato / Environmental Pollution 112 (2001) 1±10
and Yasuno, 1990a). The population at ®rst seemed to be eliminated by the chemical, but it redeveloped 2 days after the application and showed rapid growth with a high birth rate. This phenomenon was linked with improved food resource conditions, due to the local extinction of less-tolerant competitors. Similar increases in of population size after pesticide application has often been observed in some zooplankton species (Hurlbert et al., 1972; Papst and Boyer, 1980; Kaushik et al., 1985; Day et al., 1987; Yasuno et al., 1988; Hanazato and Yasuno, 1990b; Hanazato, 1991a). The eects of pesticides on natural populations may dier depending on their stage of development. In a pond experiment conducted by Hanazato and Yasuno (1990a), the rotifer Keratella valga was able to recover from losses due to carbaryl application when the population was in a growing phase. On the other hand, recovery did not occur when the same treatment was applied during a period of population decline. Hanazato and Yasuno (1990a) considered that the population had some intrinsic factors controlling its cycle, and that these factors might have prevented a response to the sudden change in environmental conditions caused by application of the chemical, such as the increased available food density after disappearance of competitors. 4. Biological interactions Biological interactions such as competition and predation can play a major role in regulating the population dynamics of zooplankton. Pesticides also aect these interactions and cause secondary eects on the structure and function of the biological community. Thus, analysis of the eects of pesticides on biological interactions is an important issue in ecotoxicological studies. It was mentioned earlier that larger animals tend to be more sensitive to pesticides than smaller ones. Interestingly, the same tendency can be applied to competitive superiority (Hanazato, 1998c). Namely, large animals such as Daphnia are vulnerable to pesticide contamination but they are superior in competition with other zooplankters, while small animals such as Bosmina and rotifers are tolerant to the chemicals but inferior in terms of competition. In the case of competition between Daphnia and rotifers, however, not only exploitative competition but also interference competition (where large herbivores damage and/or kill smaller ones due their feeding activities) is an important mechanism (Burns and Gilbert, 1986; Gilbert, 1988; MacIsaac and Gilbert, 1991). There is evidence that very large daphnids also can eat small rotifers. As a result, relatively low pesticide concentrations, which damage only large taxa (Daphnia), may aect the population dynamics of small zooplankton species
5
indirectly through competitive relationships. In fact, it is often observed in zooplankton communities that rotifers bloom after destruction of the Daphnia population by pesticides (Hurlbert et al., 1972; Papst and Boyer, 1980; Kaushik et al., 1985; Day et al., 1987; Yasuno et al., 1988). Zooplankters change their swimming behavior when exposed to pesticides, as mentioned earlier. Such behavioral changes may alter predator±prey relationships. Dodson et al. (1995) observed that the `spinning' behavior of Daphnia induced by carbaryl increased the probability of Daphnia being eaten by ®sh. This suggests that even sub-lethal concentrations of pesticides contribute to Daphnia population crashes in the presence of predators (Dodson et al., 1995). Preston et al. (1999) demonstrated that the toxicant pentachlorophenol increases swimming speed in the rotifer Brachionus calyci¯orus, thus increasing the encounter rate of the prey animals with their predators, and consequently reducing prey survivorship. Similar indirect eects of pesticides, which increase risk for prey organisms, have been reported for grass shrimp (Farr, 1977) and juvenile guppy (Brown et al., 1985). In contrast, the insecticide lindane seems to reduce predation risk in D. magna by suppressing the animals' swimming activity, which then reduces their encounter rate with the predator Hydra (Taylor et al., 1995). This same eect, of course, might be detrimental if it resulted in a lowered amount of food intake or reduced ability to maintain a vertical position in the water column of a lake. In any event, it can be concluded that, even at sub-lethal concentrations, pesticides may aect the survivorship of zooplankton in the presence of predators by controlling prey behavior. The fact that animals exposed to pesticides have a higher predation risk provides another important consideration in ecotoxicology. That is, there is an increase in the rate of pesticide uptake by predators that feed on pesticide-exposed animals, thus promoting biomagni®cation in the aquatic food web. Daphnia and other cladocerans can also develop morphological responses to predators. Daphnia develops protuberant structures on its carapace such as neckteeth, high helmets and long tailspines when exposed to chemicals released from predators such as the larva of the phantom midge Chaoborus, the backswimmer Notonecta, the predacious cladoceran Leptodora, and ®sh (Grant and Bayly, 1981; Krueger and Dodson, 1981; Hebert and Grewe, 1985; Dodson, 1989; Tollrian, 1990, 1994; Fig. 3). These structures reduce the vulnerability of Daphnia to the predators by making handling and ingestion more dicult (Havel and Dodson, 1984; Mort, 1986; Parejko, 1990). Hanazato (1991c, 1992) and Hanazato and Dodson (1993) found that the same morphological changes can be induced by organophosphorus and carbamate insecticides in several species of cyclomorphic Daphnia, if the animals are
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exposed to high concentrations of the chemicals during the ®nal embryo to the ®rst instar stages. Furthermore, Hanazato (1995) and Hanazato and Dodson (1992, 1995) observed that sub-lethal concentrations of the chemicals aect the morphology of Daphnia when the animals are exposed to both insecticides and predator (Chaoborus) odor simultaneously. Barry (1998) documented the induction of a protuberant structure (crest) by the insecticide endosulfan in Daphnia longicephala. Induction occurred even when the insecticide was applied alone (without the predator odor) and at a very low concentration (more than three orders of magnitude lower than the EC50). Because these morphologies are anti-predator defenses, such eects of pesticides must alter predator±prey relationships. The induction of protuberant structures in Daphnia is a positive eect of natural predator odor, which decreases the prey's mortality due to predation. However, there is a negative eect as well, since animals with anti-predator devices have lower clutch sizes and increased maturation time. In Daphnia this reduces population growth rate (Hanazato and Dodson, 1992), probably due to the energy costs associated with the morphological response. Hanazato and Dodson (1992, 1995) demonstrated that carbaryl reduces the growth rate and productivity of Daphnia more in the presence
Fig. 3. Daphnia species that change their morphology in response to predator odor [a: Daphnia pulex; b: D. galeata mendotae; c: D. retrocurva; d: D. ambigua; e: D. carinata; in each pair, the left picture shows the typical morph (grown without the odor) and that on the right indicates the morph bearing protuberant structures (grown with the odor)] and predators releasing the odor [f: phantom midge larva (Chaoborus: 2±10 mm in body length; g: backswimmer (Notonectidae: ca. 10 mm); h: predacious Cladocera (Leptodora: 2±10 mm)].
of predator odor, and found synergism in the combined eect of the anthropogenic and natural chemicals. These observations suggest that pesticides might also disturb natural organic chemical communication in aquatic communities (Hanazato, 1999). There is increasing evidence that aquatic organisms communicate with one another using organic chemicals (Larsson and Dodson, 1993). This is a largely unstudied but potentially critical area of pesticide impacts. 5. Community- and ecosystem-level responses Community-level responses to pesticides have often been assessed in enclosures and experimental ponds to which pesticides are experimentally applied (Kennedy et al., 1995), and in natural ponds that are occasionally contaminated with pesticides at relatively high concentrations (Helgen et al., 1988). The most frequently reported response is a change in community structure from dominance by Daphnia to dominance by small zooplankters such as rotifers and Bosmina (Hurlbert et al., 1972; Papst and Boyer, 1980; Kaushik et al., 1985; Day et al., 1987; Yasuno et al., 1988; Hanazato and Yasuno, 1990b; Hanazato, 1991a). Therefore, it is suggested that pesticides induce dominance of small zooplankton species, and consequently they reduce the mean body size of individuals in the community (Havens and Hanazato, 1993; Hanazato, 1998c). Because Daphnia depresses the populations of most small zooplankton taxa through competition, a Daphniadominant community has low species richness. However, if the Daphnia population has been destroyed by pesticides, small zooplankton species increase in abundance, and the species richness of the community becomes higher (Hanazato, 1994). The increased diversity due to pesticide contamination (Hanazato, 1998c) diers from the response of the zooplankton to other stressors such as heavy metals and acidi®cation (a kind of chemical stress), which reduce diversity (Sprules, 1975; Confer and Kaaret, 1983; Schindler, 1990; Havens, 1991, 1994b; Locke, 1992; Locke and Sprules, 1994). Hanazato (1998c) explained that heavy metals and acidi®cation exert continuous stress (direct toxic eects), so that the species richness and diversity are maintained at low levels. On the other hand, pesticides have a relatively short life span in the environment and zooplankton communities are exposed to the chemical stress for a shorter period of time. Therefore, the process of community recovery from destruction by pesticides is usually observed for populations of diverse small taxa, and the community diversity increases during the process (Hanazato, 1998c). Hanazato (1998c) therefore concluded that short-term or intermittent perturbation by chemical stress (in general) would increase the diversity of natural communities. Hanazato (1997)
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Fig. 4. The main pathways of carbon and energy ¯ow from algae through zooplankton to ®sh in lake ecosystems contaminated and uncontaminated with pesticides (redrawn from Hanazato, 1998c).
has also demonstrated that repeated application of an insecticide at a moderately toxic level to a rotifer community maintains the species richness at a high level. Probably the chemical suppress the population growth of sensitive species moderately and allows an increase of species that are tolerant to the chemical but inferior in terms of competition. This is another case of pesticides increasing the biodiversity of a plankton community, and is consistent with the intermediate disturbance hypothesis of Connell (1978), and Menge and Sutherland (1987). Changes in zooplankton community structure induced by pesticides may also aect the functioning of lake ecosystems (Hanazato, 1998c). In an uncontaminated lake containing abundant Daphnia, a large proportion of phytoplankton primary production is transferred to top predators (®sh and birds) through the algae! zooplankton!®sh pathway. In a lake that is impacted by insecticides, where there are many small zooplankton and few Daphnia, a smaller proportion of primary production is transferred upwards, due to longer food chains that include invertebrate predators (Fig. 4). Because energy loss occurs during transfer from one trophic level to another, eciency of energy transfer from primary producers to top predator is lower in a pesticide-stressed ecosystem than in an ecosystem uncontaminated by chemicals. Hanazato (1998c) suggested that reduced energy transfer eciency is a common eect of various anthropogenic stresses such as contamination with insecticides, herbicides and heavy metals, acidi®cation, eutrophication and global warming.
6. Concluding remarks In standardized toxicity tests for zooplankton, survival and reproduction are the attributes most often observed. However, analysis of these factors alone is not enough for assessing the full ecological impacts of toxicants (Forbes and Calow, 1999). Analysis of toxicant eects on other life history characteristics such as ospring size, morphology and behavior is necessary to evaluate the eects on populations, communities and ecosystems. In the natural environment, aquatic organisms are exposed to multiple pesticides simultaneously, and therefore multiple eects must be assessed (Cairns, 1983). Furthermore, the organisms are exposed to a combination of pesticides and natural stressors such as high and low temperature, food shortage, oxygen depletion and predator odor, whose eects may be worse in combination than alone. Few studies have addressed how a combination of stressors aects aquatic organisms (Folt et al., 1999) and pursuit of such studies should be encouraged. Ecosystems are comprised of a huge number of species that have complex interactions with one another. Biological interactions play an important role in maintaining ecosystems, and pesticides or other toxicants can in¯uence them. Therefore, eects of toxicants on biological interactions need to be studied in ecotoxicology investigations. Up to now, however, such studies have been few. The presence of biological interactions mediated by a chemical cue released from organisms (e.g. development
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of protuberant structure in Daphnia in response to odors from potential predators) was discovered in the 1980s (Larsson and Dodson, 1993). In the 1990s, it was demonstrated for the ®rst time that pesticides disturb the natural organic chemical communication between predator and prey (Hanazato, 1999). Hanazato (1999) then considered that the eect of pesticides was similar to that of environmental endocrine disrupters Ð anthropogenic chemicals that disturb chemical communication between tissues or organs in the bodies of animals. These ®ndings suggested the presence of a new ecological risk posed by pesticides. There are many common patterns in the structural and functional responses of ecosystems to pesticides and other anthropogenic stressors such as acidi®cation, heavy metal pollution and climatic warming, as originally hypothesized by Odum (1985). Identi®cation of these common patterns is helpful for predicting the ecological risks of environmental problems created by human activities, and is thus an important research topic in both ecotoxicology and fundamental ecology. Accordingly, more emphasis needs to be placed on ecological analysis in future ecotoxicology research.
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