Ecological Applications, 18(7), 2008, pp. 1728–1742
? 2008 by the Ecological Society of America
AN UNFORESEEN CHAIN OF EVENTS: LETHAL EFFECTS
OF PESTICIDES ON FROGS AT SUBLETHAL CONCENTRATIONS
RICK A. RELYEA1AND NICOLE DIECKS
Department of Biological Sciences, University of Pittsburgh, Pittsburgh, Pennsylvania 15260 USA
using laboratory experiments and a range of pesticide concentrations that are held constant
for short periods of time (1–4 days). From these experiments, one can estimate the
concentration that causes no effect on survival. However, organisms in nature frequently
experience multiple applications of pesticides over time rather than a single constant
concentration. In addition, organisms are embedded in ecological communities that can
propagate indirect effects through a food web. Using outdoor mesocosms, we examined how
low concentrations (10–250 lg/L) of a globally common insecticide (malathion) applied at
various amounts, times, and frequencies affected aquatic communities containing zooplank-
ton, phytoplankton, periphyton, and larval amphibians (reared at two densities) for 79 days.
All application regimes caused a decline in zooplankton, which initiated a trophic cascade in
which there was a bloom in phytoplankton and, in several treatments, a subsequent decline in
the competing periphyton. The reduced periphyton had little effect on wood frogs (Rana
sylvatica), which have a short time to metamorphosis. However, leopard frogs (Rana pipiens)
have a longer time to metamorphosis, and they experienced large reductions in growth and
development, which led to subsequent mortality as the environment dried. Hence, malathion
(which rapidly breaks down) did not directly kill amphibians, but initiated a trophic cascade
that indirectly resulted in substantial amphibian mortality. Importantly, repeated applications
of the lowest concentration ( a ‘‘press treatment’’ consisting of seven weekly applications of 10
lg/L) caused larger impacts on many of the response variables than single ‘‘pulse’’ applications
that were 25 times as great in concentration. These results are not only important because
malathion is the most commonly applied insecticide and is found in wetlands, but also because
the mechanism underlying the trophic cascade is common to a wide range of insecticides,
offering the possibility of general predictions for the way in which many insecticides impact
aquatic communities and the populations of larval amphibians.
The field of toxicology has traditionally assessed the risk of contaminants by
pipiens); pesticide pulse; wood frog (Rana sylvatica).
acetylcholine esterase inhibitor; amphibian decline; ecotoxicology; leopard frog (Rana
Understanding and predicting the impacts of anthro-
pogenic chemicals on nontarget organisms is a challeng-
ing proposition for ecologists and toxicologists.
Traditionally, toxicologists have taken a tiered approach
that begins with short-term (1–4 days), single-species
laboratory tests to determine which constant concentra-
tions of a contaminant cause 50% lethality (LC50), and
which concentrations cause no observable effect
(NOEC). If these studies indicate sufficient risk relative
to expected exposure concentrations, subsequent tests
include longer-term, single-species experiments (e.g., life
cycle tests) in the laboratory (EPA 1998). In this way,
the tiered approach is designed to incorporate greater
One limitation of this approach is that these
laboratory tests tell us how constant concentrations of
a contaminant affect a species. The situation contrasts
with real-world pesticide applications in aquatic sys-
tems, in which pesticides can not only vary in
concentration, but also in the timing of the application,
the rate of breakdown, and the frequency of repeated
applications (McConnell et al. 1998, Viant et al. 2006).
This is important because the sensitivity to contami-
nants may vary over ontogeny, and multiple applica-
tions cause a species to be exposed to a greater amount
of pesticide over ontogeny than a single application
(e.g., if the contaminant bioaccumulates or if the effects
are cumulative). Although chronic, whole-life-cycle
studies expose animals throughout ontogeny, fewer
studies have examined the importance of specific
application timing over ontogeny (Forget et al. 1998,
Howe et al. 1998, Bridges 2000), and even fewer have
directly compared the impact of single vs. multiple
applications on aquatic communities (but see Hanazato
and Yasuno  for effects on zooplankton and
Boone et al.  for effects on tadpoles). In the latter
two cases, both studies have found that multiple
applications of the insecticide carbaryl had effects
Manuscript received 4 March 2008; accepted 19 March 2008.
Corresponding Editor: D. S. Schimel.
similar to single applications (although tadpole devel-
opment was more advanced with multiple applications
to tadpoles at high densities). However, the concentra-
tions used in these experiments (500 and 3500 lg/L,
respectively) were relatively high relative to each taxon’s
LC50 (the concentration expected to kill 50% of a
population), and a single application was generally
capable of causing the maximal effect on the taxa.
Therefore, additional applications could cause no
further impacts. However, under lower and more
commonly observed concentrations (e.g., ;1–10% of
LC50 values), one might expect to observe substantial
differences between single and multiple applications.
Hence, there is a strong need to investigate how a range
of application times, amounts, and frequencies affect
A second limitation with the tiered approach is that
single-species tests exclude interspecific interactions
among taxa that live with the focal organism, while
short-term tests (i.e., 1–4 days) preclude observations of
trophic cascades that could indirectly affect the focal
organism but require more time to appear. A number of
studies have demonstrated that density-mediated indi-
rect effects initiated by contaminants might be quite
common. In general, most of these studies have
documented relatively simple indirect effects in which a
contaminant causes a sensitive taxon to decline in
abundance (predators, herbivores, competitors), which
allows an insensitive, interacting taxon to increase in
abundance (reviewed in deNoyelles et al. 1994, Brock et
al. 2000a, b, Fleeger et al. 2003, Relyea and Hoverman
2006). This type of indirect effect appears to be common
and is observable in a matter of days or weeks. However,
indirect effects have the potential to be more complex
and to take considerably longer (i.e., months) to appear.
If we combine the two issues of single vs. multiple
applications and the importance of interspecific interac-
tions in a community-oriented approach, we arrive at a
number of interesting insights. For example, there is a
long history in ecological studies of examining commu-
nity impacts either as ‘‘pulse’’ experiments (i.e., a single
disturbance) or ‘‘press’’ experiments (i.e., multiple
disturbances [Bender et al. 1984, Paine et al. 1998,
Clements and Newman 2002]). A number of community
toxicology studies have identified important indirect
effects by conducting pulse experiments, typically
consisting of a pesticide addition at the beginning of
the experiment (Havens 1994, Boone et al. 2004, Mills
and Semlitsch 2004, Relyea 2005). Additional studies
have focused on a community’s ability to rebound from
pulse treatments and found that the resiliency depends
on pesticide breakdown rate, organism generation time,
dispersal rates, and the regional species pool (Wallace et
al. 1996, Spawn et al. 1997, Woin 1998, Brock et al.
2000a, b). In the case of press experiments, we might
predict that multiple applications of a pesticide could
hold the community in a constant state of disturbance,
thereby preventing resilience. In this case, we might
observe much lower concentrations of a pesticide having
larger impacts on the community than single pulses of
much greater magnitude. Surprisingly, few studies have
directly compared the importance of single vs. multiple
pesticide applications on aquatic communities (Hanaza-
to and Yasuno 1990, Boone et al. 2001).
Indirect effects of pesticides are not simply of interest
to basic ecology, but also can have important applica-
tions for conservation. For example, amphibians are a
group of serious conservation concern due to ongoing
global population declines (Alford and Richards 1999,
Stuart et al. 2004), and some of these declines have been
correlated with insecticide use (Davidson et al. 2001,
2002, Davidson 2004). However, observed insecticide
concentrations in natural wetlands are frequently lower
than concentrations known to have direct lethal effects
on amphibians (McConnell et al. 1998, LeNoir et al.
1999, Sparling et al. 2001). Hence, there is a disconnect
between correlative regional patterns of declines with
pesticides and the most common explanation for the
mechanism(s) underlying these declines (i.e., direct
toxicity). In this study, we addressed whether low
concentrations of an insecticide (at variable concentra-
tions, times, and frequencies) can initiate a chain of
indirect effects that ultimately affects amphibians.
We addressed this question using tadpoles living at
low and high densities in communities containing
zooplankton, phytoplankton, and periphyton. In this
community, zooplankton feed primarily on planktonic
algae (i.e., phytoplankton), whereas tadpoles feed
primarily on attached algae (i.e., periphyton; Fig. 1A).
Using this community, we can make a priori predictions
about the impacts of insecticides at low concentrations
(;1–10% of LC50 values for tadpoles). At these
concentrations, one would expect no direct lethal effects
on the amphibians but substantial direct lethal effects on
the zooplankton, because aquatic invertebrates are
highly sensitive to insecticides (deNoyelles et al. 1994,
Brock et al. 2000b, Fleeger et al. 2003). If much of the
zooplankton assemblage is eliminated by the insecticide,
there should be a subsequent increase in phytoplankton
(providing that phytoplankton is limited by herbivory
and not by nutrients). The increase in phytoplankton
should then reduce the amount of light reaching the
periphyton on the bottom of the pond and thereby cause
a reduction in periphyton. Because tadpoles rely on
periphyton as their food source, this trophic cascade
should cause a reduction in growth and development,
thereby causing a change in the age and size at
metamorphosis (Fig. 1B). In short, low concentrations
of an insecticide (,LC50) that rapidly degrade should
negatively impact amphibians via a cascade of indirect
effects that occurs long after the pesticide has left the
system. Moreover, this negative effect on the tadpoles
should be most pronounced under four conditions: (1)
when tadpoles are resource limited, (2) when a species of
tadpole has an inherently long larval period that would
subject it to a longer exposure of the cascade, (3) when
October 20081729PESTICIDE-INDUCED TROPHIC CASCADES
multiple applications of the pesticide continue to
reinforce the trophic cascade and prevent the zooplank-
ton community from recovering, and (4) when the
pesticide is applied early rather than late in the
amphibian’s larval period.
Of the large number of insecticides currently applied
around the world, we chose to work with the insecticide
malathion because it is the most common insecticide
used in the United States, and can be found in natural
water bodies (Kiely et al. 2004). Malathion is a broad-
spectrum insecticide that kills a range of invertebrates
including the mosquitoes that serve as vectors for West
Nile virus and malaria (Gratz and Jany 1994, Walker
2000). Annual applications of malathion amount to 10–
14 million kg of active ingredient applied to nearly 106
ha of cropland as well as home, garden, government,
and industrial use (Kiely et al. 2004; National Pesticide
Use Database, available online).2In general, malathion
is highly toxic to most aquatic invertebrates (including
zooplankton) and moderately toxic to larval amphibians
(USEPA Ecotox Database, available online).3
Expected concentrations for malathion in water range
from 0 to 1600 lg/L, and wetland surveys have found up
to 600 lg/L (California Department of Fish and Game
1982, USDA 1997, McConnell et al. 1998, LeNoir et al.
1999, Relyea 2004). Evaluating typical concentrations in
standing water is difficult because major aquatic surveys
have only examined either streams and rivers or well
water. For example, stream surveys indicate that
a low concentration of a pesticide that is only directly toxic to zooplankton. A time line is also provided for (C) when the pesticide
additions occurred and (D) when each response variable was measured. D.O. is dissolved oxygen. Malathion was the only pesticide
used. ‘‘Light meter’’ indicates when water transparency was measured.
Hypothesized indirect effects in a simple aquatic food web that contains either (A) no pesticide or (B) contamination by
RICK A. RELYEA AND NICOLE DIECKS1730
Vol. 18, No. 7
Alford, R. A., and S. J. Richards. 1999. Global amphibian
declines: a problem in applied ecology. Annual Review of
Ecology and Systematics 30:133–165.
Altwegg, R., and H. U. Reyer. 2003. Patterns of natural
selection on size at metamorphosis in water frogs. Evolution
Arar, E. J., and G. B. Collins. 1997. In vitro determination of
chlorophyll a and pheophytin a in marine and freshwater
algae by fluorescence. Method 445.0. National Exposure
Research Laboratory, USEPA, Cincinnati, Ohio, USA.
Barry, M. J., and D. C. Logan. 1998. The use of temporary
pond microcosms for aquatic toxicity testing: direct and
indirect effects of endosulfan on community structure.
Aquatic Toxicology 41:101–124.
Bender, E. E., T. J. Case, and M. E. Gilpin. 1984. Perturbation
experiments in community ecology: theory and practice.
Berven, K. A. 1981. Mate choice in wood frog, Rana sylvatica.
Berven, K. A. 1982. The genetic basis of altitudinal variation in
wood frog Rana sylvatica. I. An experimental analysis of life
history traits. Evolution 36:962–983.
Berven, K. A., and D. E. Gill. 1983. Interpreting geographic
variation in life-history traits. American Zoologist 23:85–97.
Boone, M. D., C. M. Bridges, J. F. Fairchild, and E. E. Little.
2005. Multiple sublethal chemicals negatively affect tadpoles
of the green frog, Rana clamitans. Environmental Toxicology
and Chemistry 24:1267–1272.
Boone, M. D., C. M. Bridges, and B. B. Rothermel. 2001.
Growth and development of larval green frogs (Rana
clamitans) exposed to multiple doses of an insecticide.
Boone, M. D., and C. M. Bridges-Britton. 2006. Examining
multiple sublethal contaminants on the gray treefrog (Hyla
versicolor): effects of an insecticide, herbicide, and fertilizer.
Environmental Toxicology and Chemistry 25:3261–3265.
Boone, M. D., R. D. Semlitsch, J. F. Fairchild, and B. B.
Rothermel. 2004. Effects of an insecticide on amphibians in
large-scale experimental ponds. Ecological Applications 14:
Bridges, C. M. 2000. Long-term effects of pesticide exposure at
various life stages of southern leopard frog (Rana sphenoce-
phala). Archives of Environmental Contamination and
Bridges, C. M., and M. D. Boone. 2003. The interactive effects
of UV-B and insecticide exposure on tadpole survival, growth
and development. Biological Conservation 113:49–54.
Brock, T. C. M., J. Lahr, and P. J. van den Brink. 2000a.
Ecological risks of pesticides in freshwater systems. Part 1:
Herbicides. Alterra Report 088, Wageningen, The Nether-
Brock, T. C. M., R. P. A. van Wijngarden, and G. van Geest.
2000b. Ecological risks of pesticides in freshwater systems.
Part 2: Insecticides. Alterra Report 089, Wageningen, The
California Department of Fish and Game. 1982. Monitored
aquatic incidents during broadscale aerial application over
San Francisco Bay area, 1981. Administrative Report 82-2.
Sacramento, California, USA.
Chase, J. M., and T. M. Knight. 2003. Drought-induced
mosquito outbreaks in wetlands. Ecology Letters 6:1017–
Clements, W. H., and M. C. Newman. 2002. Community
ecotoxicology. Wiley, Chichester, UK.
Davidson, C. 2004. Declining downwind: amphibian popula-
tion declines in California and historical pesticide use.
Ecological Applications 14:1892–1902.
Davidson, C., H. B. Shafer, and M. R. Jennings. 2001. Declines
of the California red-legged frog: Climate, UV-B, habitat,
and pesticides hypotheses. Ecological Applications 11:464–
Davidson, C., H. B. Shafer, and M. R. Jennings. 2002. Spatial
tests of the pesticide drift, habitat destruction, UV-B, and
climate-change hypotheses for California amphibian declines.
Conservation Biology 16:1588–1601.
deNoyelles, F., S. L. Dewey, D. G. Huggins, and W. D. Kettle.
1994. Aquatic mesocosms in ecological effects testing:
detecting direct and indirect effects of pesticides. Pages
577–603 in R. L. Graney, J. H. Kennedy, and J. H. Rodgers,
Jr., editors. Aquatic mesocosm studies in ecological risk
assessment. Lewis, Boca Raton, Florida, USA.
Denver, R. J., N. Mirhadi, and M. Phillips. 1998. Adaptive
plasticity in amphibian metamorphosis: response of Scaphio-
pus hammondii tadpoles to habitat desiccation. Ecology 19:
EPA. 1998. Guidelines for ecological risk assessment. U.S.
Environmental Protection Agency, Washington, D.C., USA.
Fairchild, J. F., T. W. LaPoint, J. L. Zajicek, M. K. Nelson,
F. J. Dwyer, and P. A. Lovely. 1992. Population-,
community-, and ecosystem-level responses of aquatic
mesocosms to pulsed doses of a pyrethroid insecticide.
Environmental Toxicology and Chemistry 11:115–129.
Fleeger, J. W., K. R. Carman, and R. M. Nisbet. 2003. Indirect
effects of contaminants in aquatic ecosystems. Science of the
Total Environment 317:207–233.
Forget, J., J. F. Pavillon, M. R. Menasria, and G. Bocquene.
1998. Mortality and LC50 values for several stages of the
marine copepod Tigriopus brevicornis (Muller) exposed to the
metals arsenic and cadmium and the pesticides atrazine,
carbofuran, dichlorvos, and malathion. Ecotoxicology and
Environmental Safety 40:239–244.
Gerhardt, H. C. 1994. The evolution of vocalization in frogs
and toads. Annual Review of Ecology and Systematics 25:
Giddings, J. M., R. C. Biever, M. F. Annunziato, and A. J.
Hosmer. 1996. Effects of diazinon on large outdoor pond
microcosms. Environmental Toxicology and Chemistry 15:
Gosner, K. L. 1960. A simplified table for staging anuran
embryos and larvae with notes on identification. Herpeto-
Gratz, N. G., and W. C. Jany. 1994. What role for insecticides
in vector control programs? American Journal of Tropical
Medicine and Hygiene 50:11–20.
Guerrant, G. O., L. E. Fetzer, Jr., and J. W. Miles. 1970.
Pesticide residues in Hale County, Texas, before and after
ultra-low volume aerial application of malathion. Pesticide
Monitoring Journal 4:14–20.
Hanazato, T., and M. Yasuno. 1987. Effects of a carbamate
insecticide, carbaryl, on summer phyto- and zooplankton
communities in ponds. Environmental Pollution 48:145–159.
Hanazato, T., and M. Yasuno. 1989. Effects of carbaryl on
spring zooplankton communities in ponds. Environmental
Hanazato, T., and M. Yasuno. 1990. Influence of time of
application of an insecticide on recovery patterns of a
zooplankton community in experimental ponds. Archives of
Environmental Contamination and Toxicology 19:77–83.
Havens, K. E. 1994. An experimental comparison of effects of
two chemical stressors on a freshwater zooplankton assem-
blage. Environmental Pollution 84:245–251.
Havens, K. E. 1995. Insecticide (carbaryl, 1-napthyl-n-methyl-
carbamate) effects on a freshwater plankton community:
zooplankton size, biomass, and algal abundance. Water, Air,
and Soil Pollution 84:1–10.
October 2008 1741PESTICIDE-INDUCED TROPHIC CASCADES
Howard, R. D. 1980. Mating behaviour and mating success in
wood frogs, Rana sylvatica. Animal Behaviour 28:705–716.
Howe, G. E., R. Gillis, and R. C. Mowbray. 1998. Effect of
chemical synergy and larval stage on the toxicity of atrazine
and alachlor to amphibian larvae. Environmental Toxicology
and Chemistry 17:519–525.
Jurgensen, T. A., and K. D. Hoagland. 1990. Effects of short-
term pulses of atrazine on attached algal communities in a
small stream. Archives of Environmental Contamination and
Kiely, T., D. Donaldson, and A. Grube. 2004. Pesticide
industry sales and usage: 2000 and 2001 market estimates.
USEPA, Washington, D.C., USA.
Kosinski, R. J., and M. G. Merkle. 1984. The effects of four
terrestrial herbicides on the productivity of artificial stream
algal communities. Journal of Environmental Quality 13:75–
LeNoir, J. S., L. L. McConnell, G. M. Fellers, T. M. Cahill,
and J. N. Seiber. 1999. Summertime transport of current-use
pesticides from California’s Central Valley to the Sierra
Nevada Mountain Range, USA. Environmental Toxicology
and Chemistry 18:2715–2722.
Lozano, S. J., S. L. O’Halloran, K. W. Sargent, and J. C.
Brazner. 1992. Effects of esfenvalerate on aquatic organisms
in littoral enclosures. Environmental Toxicology and Chem-
McConnell, L. L., J. S. LeNoir, S. Datta, and J. N. Seiber.
1998. Wet deposition of current-use pesticides in the Sierra
Nevada mountain range, California, USA. Environmental
Toxicology and Chemistry 17:1908–1916.
Mills, N. E., and R. D. Semlitsch. 2004. Competition and
predation mediate the indirect effects of an insecticide on
southern leopard frogs. Ecological Applications 14:1041–
Mitsch, W. J., and J. G. Gosselink. 1986. Wetlands. Van
Nostrand Reinhold, New York, New York, USA.
Niemi, G. J., P. DeVore, N. Detenbeck, D. Taylor, A. Lima, J.
Pastor, J. D. Yount, and R. J. Naiman. 1990. Overview of
case studies on recovery of aquatic systems from disturbance.
Environmental Management 84:571–587.
Odenkirchen, E., and S. P. Wente. 2007. Risks of malathion use
to federally listed California red-legged frog (Rana aurora
draytonii). USEPA Environmental Fate and Effects Division,
Washington, D.C., USA.
Paine, R. T., M. J. Tegner, and E. A. Johnson. 1998.
Compounded perturbations yield ecological surprises. Eco-
Pusey, B. J., A. H. Arthington, and J. MacClean. 1994. The
effects of a pulsed application of chlorpyrifos on macroin-
vertebrate communities in an outdoor artificial stream
system. Ecotoxicology and Environmental Safety 27:221–
Rand, G. M., J. R. Clark, and C. M. Holmes. 2000. Use of
outdoor freshwater pond microcosms: II. Responses of biota
to pyridaben. Environmental Toxicology and Chemistry 19:
Relyea, R. A. 2004. Synergistic impacts of malathion and
predatory stress on six species of North American tadpoles.
Environmental Toxicology and Chemistry 23:1080–1084.
Relyea, R. A. 2005. The impact of insecticides and herbicides
on the biodiversity and productivity of aquatic communities.
Ecological Applications 15:618–627.
Relyea, R. A., and J. R. Auld. 2004. Having the guts to
compete: how intestinal plasticity explains costs of inducible
defenses. Ecology Letters 7:869–875.
Relyea, R. A., and J. R. Auld. 2005. Predator- and competitor-
induced plasticity: how changes in foraging morphology
affect phenotypic trade-offs. Ecology 86:1723–1729.
Relyea, R. A., and J. T. Hoverman. 2006. Assessing the ecology
in ecotoxicology: a review and synthesis in freshwater
systems. Ecology Letters 9:1157–1171.
Relyea, R. A., N. M. Schoeppner, and J. T. Hoverman. 2005.
Pesticides and amphibians: the importance of community
context. Ecological Applications 15:1125–1134.
Rohr, J. R., and P. W. Crumrine. 2005. Effects of an herbicide
and an insecticide on pond community structure and
processes. Ecological Applications 15:1135–1147.
Sanders, H. O., and O. B. Cope. 1966. Toxicities of several
pesticides to two species of cladocerans. Transactions of the
American Fisheries Society 95:165–169.
Schiesari, L. 2006. Pond canopy cover: a resource gradient for
anuran larvae. Freshwater Biology 51:412–423.
Semlitsch, R. D. 1987. Density-dependent growth and fecundity
in the paedomorphic salamander Ambystoma talpoideum.
Semlitsch, R. D., D. C. Scott, and J. H. K. Pechmann. 1988.
Time and size at metamorphosis related to adult fitness in
Ambystoma talpoideum. Ecology 69:184–192.
Smith, D. C. 1987. Adult recruitment in chorus frogs: effects of
size and date at metamorphosis. Ecology 68:344–350.
Sparling, D. W., G. M. Fellers, and L. L. McConnell. 2001.
Pesticides and amphibian population declines in California,
USA. Environmental Toxicology and Chemistry 20:1591–
Spawn, R. L., K. D. Hoagland, and B. D. Siegfried. 1997.
Effects of alachlor on an algal community from a midwestern
agricultural stream. Environmental Toxicology and Chemis-
Stuart, S. N., J. S. Chanson, N. A. Cox, B. E. Young, A. S. L.
Rodrigues, D. L. Fischman, and R. W. Waller. 2004. Status
and trends of amphibian declines and extinctions worldwide.
USDA (U.S. Department of Agriculture). 1997. Environmental
monitoring report. Cooperative medfly project Florida.
Spray operations Hillsborough area. United States Depart-
ment of Agriculture Report, Washington, D.C., USA.
USGS (U.S. Geological Survey). 2000. Organophosphate
pesticide occurrence and distribution in surface and ground
water of the United States, 1992–1997. Report 00-187,
United States Geological Survey, Washington, D.C., USA.
Viant, M. R., C. A. Pincetich, and R. S. T. Eerderna. 2006.
Metabolic effects of dinoseb, diazinon and esfenvalerate in
eyed eggs and alevins of Chinook salmon (Oncorhynchus
tshawytscha) determined by H-1 NMR metabolomics.
Aquatic Toxicology 77:359–371.
Walker, K. 2000. Cost-comparison of DDT and alternative
insecticides for malaria control. Medical and Veterinary
Wallace, J. B., J. W. Grubaugh, and M. R. Whiles. 1996. Biotic
indices and stream ecosystem processes: results from an
experimental study. Ecological Applications 6:140–151.
Wang, T. 1991. Assimilation of malathion in the Indian River
estuary, Florida. Bulletin of Environmental Contamination
and Toxicology 47:238–243.
Weis, J. S., G. Smith, T. Zhou, C. Santiago-Bass, and P. Weis.
2001. Effects of contaminants on behavior: biochemical
mechanisms and ecological consequences. BioScience 51:209–
Woin, P. 1998. Short- and long-term effects of the pyrethroid
insecticide fenvalerate on an invertebrate pond community.
Ecotoxicology and Environmental Safety 41:137–156.
Wong, C. K., K. H. Chu, and F. F. Shum. 1995. Acute and
chronic toxicity of malathion to the fresh-water cladoceran
Moina macrocarpa. Water, Air, and Soil Pollution 84:399–
RICK A. RELYEA AND NICOLE DIECKS1742
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