Download Pesticide effects on freshwater zooplankton: an ecological perspective

Environmental Pollution 112 (2001) 1±10
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Review
Pesticide e€ects 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'': E€ects of pesticides on freshwater zooplankton are considered from an ecological perspective.
Abstract
The e€ects 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 a€ect toxicity. Populations in the growth phase are vulnerable to pesticides but have the
potential to recover rapidly from the damage. Pesticides may a€ect 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 e€ects 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 eciency 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
2
T. Hanazato / Environmental Pollution 112 (2001) 1±10
1. Introduction
Surface water bodies are contaminated with many
anthropogenic toxic chemicals that can a€ect their natural communities. It is necessary to assess the e€ects 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 a€ect 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
e€ects (Cairns and Niederlehner, 1995). However, this
assessment is hindered by the fact that natural ecosystems are diverse and the e€ects 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 di€erent
levels of organization, ranging from individuals to
communities, and discuss the e€ects 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 e€ects on animals are quanti®ed and standard test statistics, such as LC50, are determined. These
tests indicate clearly that pesticides exert species-speci®c
e€ects 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 e€ective concentration) in
chronic toxicity tests for cadmium and 3,4-dichloroaniline (DCA), and found a large di€erence 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 a€ect
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 di€erent
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 di€er even
among neonates depending on their body size. Neonatal
size is a€ected 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 e€ects 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 e€ect 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 a€ect pesticide
toxicity. Folt et al. (1999) tested the e€ects 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 a€ect 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 a€ected 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 a€ected 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.
4
T. Hanazato / Environmental Pollution 112 (2001) 1±10
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
E€ects 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 a€ected 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 o€spring produced) of the test
animals is analyzed. Analysis of reproduction may be
useful for assessing the e€ects of chemicals on population growth. However, Hanazato (1998b) has cautioned
that the number of o€spring per clutch is not a€ected
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, a€ect 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
a€ecting 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 e€ects.
E€ects 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 o€spring. 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 e€ects of pesticides on natural populations may
di€er 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 a€ect
these interactions and cause secondary e€ects on the
structure and function of the biological community.
Thus, analysis of the e€ects 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 a€ect 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 e€ects 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 e€ect, 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 a€ect 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 dicult (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
6
T. Hanazato / Environmental Pollution 112 (2001) 1±10
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 a€ect 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 e€ects of
pesticides must alter predator±prey relationships.
The induction of protuberant structures in Daphnia is
a positive e€ect of natural predator odor, which
decreases the prey's mortality due to predation. However, there is a negative e€ect 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
e€ect 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)
di€ers 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
e€ects), 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)
T. Hanazato / Environmental Pollution 112 (2001) 1±10
7
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 a€ect 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, eciency 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 eciency is a common e€ect 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
e€ects on other life history characteristics such as o€spring size, morphology and behavior is necessary to
evaluate the e€ects on populations, communities and
ecosystems.
In the natural environment, aquatic organisms are
exposed to multiple pesticides simultaneously, and
therefore multiple e€ects 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 e€ects may be
worse in combination than alone. Few studies have
addressed how a combination of stressors a€ects 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, e€ects 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
8
T. Hanazato / Environmental Pollution 112 (2001) 1±10
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 e€ect 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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