mosquitofish dominate amphibian and invertebrate community development in experimental wetlands

Mosquitofish dominate amphibian and invertebrate

community development in experimental wetlands

Christopher D. Shulse1*, Raymond D. Semlitsch2 and Kathleen M. Trauth3

1Missouri Department of Transportation, PO Box 270, Jefferson City, MO 65102, USA; 2Division of Biological

Sciences, University of Missouri, Columbia MO 65211, USA; and 3Department of Civil and Environmental

Engineering, University of Missouri, Columbia, MO 65211, USA

Summary

1. Restored and constructed habitats can play important conservation roles. Predators help

shape communities in these habitats through complex interactions with prey, other predators

and biotic and abiotic characteristics of the environment. However, introduced predators can

have dramatic effects that may be difficult to predict.

2. Using regression models, we compared influences of introduced invasive western mosqui-

tofish Gambusia affinis to those of two naturally colonizing predators (crayfish and dragon-

flies), and vegetation, on three anuran species in experimentally constructed wetlands. Using

analyses of covariance, we also examined influences of mosquitofish and vegetation on aqua-

tic invertebrate communities.

3. We found that mosquitofish reduced abundances of grey treefrogs Hyla versicolor and

H. chrysoscelis and boreal chorus frog Pseudacris maculata, but had no significant influence

on green frog Lithobates clamitans. Mosquitofish also reduced invertebrate abundance, but

their effect on richness was less clear. Vegetation cover did not significantly increase most

anuran or invertebrate abundances. However, vegetation increased invertebrate richness.

After fish removal, invertebrate abundance increased. Fish removal may have facilitated cho-

rus frog re-colonization into wetlands with low abundance of invertebrate predators.

4. Our results indicate that mosquitofish are detrimental to wetland communities, and we

recommend that managers avoid stocking mosquitofish. We also encourage temporary or dra-

inable wetlands to prevent mosquitofish persistence if colonization occurs. Implementing these

recommendations will improve the conservation potential of restored wetlands.

Key-words: aquatic communities, introduced predators, rotenone, wetland restoration

Introduction

A major challenge for restoration ecologists involves pre-

dicting pathways of ecological succession in the presence

of multiple biotic and abiotic conditions. Predators play

key roles in shaping natural communities through interac-

tions with prey and other predators (Van Buskirk 1988;

Griffen 2006). These interactions are often complex,

thereby making discernment of mechanisms generating

natural community patterns and structure difficult

(DeWitt & Langerhans 2003). Most prey species are con-

sumed by multiple predators, but prey responses to differ-

ent predators are not the same. The reaction by prey to

one predator may make it more or less vulnerable to

another, depending on the nature of interactions between

the two predators (Sih, Englund & Wooster 1998). These

interactions are important during restoration because as

succession proceeds, food webs develop based upon con-

ditions present at a site, some of which can be manipu-

lated by the restoration ecologist. For example, wetland

hydroperiod plays a major role in shaping wetland com-

munities (Pechmann et al. 1989) so designing wetlands

with temporary or permanent hydroperiods will have a

direct impact on community composition (Pechmann

et al. 2001).

Introduced predators can dramatically alter community

development, particularly if the predator is invasive and

prey do not possess adaptive traits to reduce mortality

(Nystr€om et al. 2001). Introduced fish have been impli-

cated in aquatic community disruptions. Eastern Gambu-

sia holbrookii (Girard 1859) and western G. affinis*Correspondence author. E-mail: [email protected]

© 2013 The Authors. Journal of Applied Ecology © 2013 British Ecological Society

Journal of Applied Ecology 2013, 50, 1244–1256 doi: 10.1111/1365-2664.12126

mosquitofish are small poeciliids native to the south-

eastern United States, but introduced throughout the

world because of their purported effectiveness at control-

ling mosquitoes (Pyke 2008). Mosquitofish readily con-

sume invertebrates, small fish and amphibian eggs and

larvae (Pyke & White 2000; Richard 2002), and they can

alter the composition of the aquatic invertebrate commu-

nity (Hurlbert, Zedler & Fairbanks 1972). Mosquitofish

are the most widespread fish in the world (Pyke 2008), and

the IUCN lists them among the 100 worst invasive species

(Lowe, Browne & Boudjelas 2000). Introductions have

been associated with amphibian declines in California,

Australia and China (Lawler et al. 1999; Pyke & White

2000; Karraker, Arrigoni & Dudgeon 2010), and negative

effects have been recorded in experiments using eggs and

larvae of amphibian species within their native range

(Grubb 1972; Baber & Babbitt 2004; Stanback 2010).

Dragonfly naiads and crayfish are top invertebrate

predators in many wetlands. Dragonflies are carnivorous

and consume other aquatic invertebrates and small fish

(Merrill & Johnson 1984; Van Buskirk 1988). They are

also efficient consumers of larval amphibians (Smith 1983;

Semlitsch & Gibbons 1988). Crayfish are highly omnivo-

rous and consume detritus, vegetation, invertebrates,

carrion, fish eggs and young, and amphibian eggs and

larvae (Momot 1995; Dorn & Wojdak 2004). Although

dragonflies are generalist predators (Wallace et al. 1987),

their trophic impact is likely to be narrower than crayfish,

which can directly impact multiple trophic levels (Dorn &

Wojdak 2004). Introduced crayfish can disrupt aquatic

communities and have been implicated in amphibian

declines (Gamradt & Kats 1996; Axelsson et al. 1997).

However, the results from other studies suggest that cray-

fish are inefficient predators of larval amphibians (Fauth

1990; Lefcort 1996). Nevertheless, crayfish can destroy

vegetation (Axelsson et al. 1997), thus lowering habitat

complexity and potentially contributing to reduced

amphibian abundance.

We compared the influences of introduced western mos-

quitofish to two native predators (crayfish and dragon-

flies) on three amphibian species in experimental

constructed wetlands. Grey treefrogs Hyla versicolor/chry-

soscelis complex, boreal chorus frogs Pseudacris maculata

and green frogs Lithobates clamitans were selected because

each species employs different mechanisms to cope with

predation (Smith 1983; Van Buskirk 2003). Grey treefrogs

and boreal chorus frogs are palatable to fish, but green

frogs are not (Kats, Petranka & Sih 1988). Furthermore,

chorus frogs prefer temporary wetlands, whereas grey

treefrogs will reproduce in both temporary and permanent

water, and green frogs require relatively permanent water

(Kats, Petranka & Sih 1988). We predicted mosquitofish

would have a greater negative impact on hylids than on

green frogs, and mosquitofish effects would be greater

than those of crayfish and dragonflies. We also examined

the influence of mosquitofish on aquatic invertebrates and

whether vegetation attenuates fish impacts. We hypothesized

mosquitofish would lower invertebrate abundance and

richness and vegetation would attenuate predation

because habitat complexity can provide refuge for prey

(Sass et al. 2006; Hartel et al. 2007).

This research is part of a larger study that aims to

improve the conservation potential of restored and con-

structed wetlands (see Shulse 2011 and Shulse et al. 2012).

The current study examines the roles of predators and

vegetation in determining amphibian and invertebrate

communities following wetland construction. We focus on

mosquitofish because of their widespread use and the per-

ception that they are benign to native wildlife (Pyke

2008). Our goals were to investigate whether mosquitofish

influence wetland communities differently than native pre-

dators and to present wetland management recommenda-

tions based on replicated experimentation in the field.

Materials and methods

During October and November 2006, we constructed replicate

wetland arrays at three upland grassland habitats in north-eastern

Missouri, USA (Fig. 1), managed by the Missouri Department of

Conservation (MDC). Six wetlands (23 m diameter, 0�76 m maxi-

mum depth) were constructed at each location (n = 18). A com-

plete description of wetland designs, placement and surrounding

landscapes is given in Shulse et al. (2012). A goal of another

study at these wetlands was to examine the influences of within-

wetland slope, mosquitofish and vegetation on amphibian meta-

morph production and species richness (Shulse et al. 2012).

Therefore, we randomly assigned one of the six combinations of

slope, mosquitofish and vegetation to each wetland (Table 1).

Planted wetlands received 50 cordgrass Spartina pectinata divi-

sions spaced evenly apart and radiating from the centre. Non-

surviving plants were replaced during autumn 2007. All other

vegetation was allowed to develop naturally.

In March 2007, we captured mosquitofish in a Missouri

Department of Transportation compensatory mitigation wetland

in Audrain County, Missouri, and released them into the three

selected wetlands at each MDC location at a rate of 3089 fish

ha�1, which is slightly higher than the rate of 2471 fish ha�1

(1000 fish acre�1) recommended by Duryea et al. (1996). This

resulted in a founding population of 125 adult mosquitofish per

stocked wetland. Fish were re-stocked where samples indicated

low populations in spring 2008. Reconnaissance sampling in early

spring 2009 revealed healthy mosquitofish populations in all

stocked wetlands so no further re-stocking occurred. MDC per-

sonnel removed mosquitofish from stocked wetlands at one loca-

tion (Redman) on 17 September 2009 using the piscicide

rotenone (chemical restoration). Rotenone was applied to stocked

wetlands at another site (Sears) on 10 March 2010. Rotenone was

applied at label rates. Dead mosquitofish were observed in all

wetlands. However, during the second sampling period in 2010,

mosquitofish were captured in one treated wetland (Sears 1).

Therefore, it was assumed that this wetland contained survivors

so we considered it fish-stocked for 2010 analyses. The stocked

wetlands at the third location (White) were not treated and

reconnaissance sampling in early spring 2010 indicated healthy

populations.

Amphibians, mosquitofish and invertebrates were sampled

three times within each season using aquatic funnel traps and dip

© 2013 The Authors. Journal of Applied Ecology © 2013 British Ecological Society, Journal of Applied Ecology, 50, 1244–1256

Mosquitofish dominate wetland communities 1245

nets. Aquatic funnel traps were deployed for 48 h in 2007 and

2008 and overnight in 2009 and 2010, using two kinds of com-

mercially available minnow traps: collapsible nylon mesh traps

(3-mm mesh; 38 9 26 9 26 cm; 6 cm openings) or galvanized

steel wire traps (6-mm mesh; 42 cm long; 2�5 cm openings). Two

traps of each were used per wetland and placements were stag-

gered so that traps of the same model were directly across from

one another at each cardinal direction. Pair direction assignment

was random. One dip net (3-mm nylon mesh) sweep was con-

ducted from the water’s edge at each cardinal direction and

sweeps were ~1�5 m long with the net pressed to the substrate

and pulled towards the sampler. During the second 2007 sam-

pling period, a zooplankton canvas D-net with 500-micron mesh

bottom was added to the protocol to capture very small organ-

isms. Sweeps of approximately 1�5 m occurred at each ordinal

direction using the D-net. This resulted in four dip net sweeps

and 4 D-net sweeps, spaced evenly apart, for each wetland during

each sampling period after 2007–1. Data from all methods were

combined to calculate amphibian, mosquitofish and invertebrate

abundances and invertebrate taxa richness at each wetland during

each sampling period. All organisms were released unharmed at

point of capture after recording. We were unable to distinguish

between eastern grey treefrogs and Cope’s grey treefrogs in the

field so grey treefrogs are considered as the Hyla versicolor/chry-

soscelis complex.

Within-wetland vegetation was measured using four 1-m² quad-

rats spaced at cardinal directions around the perimeter of each

wetland. Quadrats were placed at the edge of each wetland to

assess vegetation cover within 1 m of the shore and at 3 m from

the shore. The percentages of open water, emergent, floating and

submerged vegetation were visually estimated within each quad-

rat. The three categories of vegetation were combined and aver-

aged for all quadrats over all sampling periods within a season at

each wetland to calculate an average measure of vegetation cover

for the season. Percentage vegetation cover was transformed to

the arcsine square root of the proportion for analyses. Develop-

ment of natural vegetation occurred faster in some non-planted

wetlands than in planted wetlands. Therefore, we used vegetation

cover as a continuous covariate within our analyses as opposed

to a treatment factor (below).

We analysed each year separately to look for overall patterns

in abundance or taxa richness. For regressions and ANCOVAs, a

single wetland was used as the unit of replication. All statistical

analyses were performed using SPSS version 16�0 (2007 SPSS

Chicago, IL, USA). To explain relationships between abundances

of amphibians, predators and vegetation cover, we developed

regression models with negative binomial distributions and log-

link functions using the generalized linear model option in SPSS.

We used abundances of mosquitofish, crayfish and dragonfly

naiads, along with vegetation cover, as independent variables. We

conducted Spearman’s rank correlation tests between independent

variables to avoid including two variables strongly correlated

with one another (r ≥ 0�70) in models. Dragonfly abundance and

vegetation cover were highly correlated in 2008 and 2009

(Table 2); therefore, in models for these years, we focused our

analyses on predators and excluded vegetation cover. Each

regression model contained grey treefrog, boreal chorus frog or

green frog abundances as dependent variables, and either all four

independent variables or the three predator variables (2008 and

2009). Only crayfish, dragonflies and vegetation cover were

included in the model for grey treefrogs in 2010 because no grey

treefrogs were captured in wetlands containing fish in that year.

To test the hypothesis that mosquitofish reduce invertebrate

abundances, we used the cumulative number of invertebrates

Fig. 1. Locations of study sites within

north-eastern Missouri, USA.

Table 1. Wetland treatment combinations. One wetland of each

treatment combination was constructed at each of the three study

sites

Treatment

combination Slope Mosquitofish

Prairie

cordgrass

1 4:1 Stocked Not planted

2 4:1 Not stocked Not planted

3 15:1 Stocked Not planted

4 15:1 Stocked Planted

5 15:1 Not stocked Not planted

6 15:1 Not stocked Planted


1246 C. D. Shulse, R. D. Semlitsch & K. M. Trauth

(log10-transformed) captured during all sampling periods each

year at each wetland as dependent variables in generalized linear

models with mosquitofish as a factor and vegetation cover as a

covariate. We excluded crayfish, snails, bivalves and daphniids

from ANCOVAs. Data for snails and bivalves were not collected

consistently, and daphniids were challenging to quantify at very

high numbers. Crayfish grow large enough to escape fish preda-

tion (Stein 1977), and even the smallest crayfish we observed were

too large for gape-limited mosquitofish to consume.

To test whether mosquitofish reduce invertebrate richness, we

used the cumulative number of invertebrate taxa captured during

all sampling periods per year at each wetland as dependent vari-

ables in generalized linear models. Mosquitofish presence was

included as a factor and vegetation cover as a covariate. Inverte-

brate richness values included daphniids but excluded crayfish,

snails and bivalves. We attempted to identify each invertebrate to

family, but we were unable to identify some to this level in the

field (See Table S1, Supporting Information). To achieve normal

distribution, invertebrate taxa richness values were log10-

transformed for 2007. Because comparisons of taxa richness

among different assemblages should account for differences in

sampling effort and abundance (Gotelli & Colwell 2001), we plot-

ted rarefied richness curves for each year using EstimateS version

8.2 (Colwell 2005). We included daphniids in rarefaction analyses,

but we capped the number of daphnia at 100 per sample due to

the aforementioned quantification problems. For rarefaction, we

defined sample as the total individuals captured by all methods

during a single sampling period at each wetland. Finally, we

Table 2. Spearman’s correlation matrix for independent variables

in amphibian generalized linear regression models

N = 18 Mosquitofish Crayfish Dragonfly

Vegetation

2007 �0�13 �0�20 0�142008 �0�37 �0�18 0�712009 �0�34 �0�56 0�782010 �0�44 �0�61 0�68

Mosquitofish

2007 0�16 �0�642008 0�16 �0�632009 0�11 �0�502010 0�05 �0�44

Crayfish

2007 �0�342008 �0�392009 �0�562010 �0�63

Table 3. Parameter estimates for independent variables in amphibian abundance generalized linear regression models. Significant param-

eters and their corresponding statistics are in boldface

Species Year Parameter b SE Wald v2

95% CI

Sig.Lower Upper

Grey treefrog

complex

2007 Mosquitofish �0�02 0�01 4�38 �0�05 �0�001 0�04Crayfish �0�03 0�01 7�59 �0�05 �0�009 0�006Dragonfly 0�11 0�08 1�87 �0�05 0�28 0�17Vegetation 19�81 5�79 11�73 8�47 31�16 0�001

2008 Mosquitofish �0�06 0�02 11�68 �0�09 �0�03 0�001Crayfish �0�02 0�01 2�18 �0�05 0�007 0�14Dragonfly �0�11 0�05 5�57 �0�21 �0�02 0�02

2009 Mosquitofish �0�03 0�01 8�29 �0�05 �0�009 0�004Crayfish �0�60 0�27 4�97 �1�13 �0�07 0�03Dragonfly �0�05 0�03 2�82 �0�11 0�009 0�09

2010* Crayfish �0�01 0�01 0�98 �0�04 0�01 0�32Dragonfly 0�003 0�02 0�03 �0�03 0�04 0�87Vegetation 2�77 1�84 2�27 �0�84 6�38 0�13

Boreal chorus

frog

2008 Mosquitofish �0�02 0�006 10�89 �0�03 �0�009 0�001Crayfish �0�04 0�03 1�61 �0�11 0�02 0�21Dragonfly 0�01 0�07 0�02 �0�013 0�15 0�88

2010 Mosquitofish �0�04 0�01 6�87 �0�07 �0�01 0�009Crayfish 0�001 0�02 0�001 �0�04 0�04 0�97Dragonfly �0�12 0�05 6�58 �0�21 �0�03 0�01Vegetation 1�46 2�54 0�33 �3�53 6�44 0�57

Green frog 2008 Mosquitofish �0�003 0�01 0�06 �0�02 0�02 0�80Crayfish �0�05 0�02 4�91 �0�10 �0�006 0�03Dragonfly �0�12 0�06 4�35 �0�24 �0�008 0�04

2009 Mosquitofish 0�003 0�003 0�96 �0�003 0�009 0�33Crayfish �0�10 0�03 8�34 �0�17 �0�03 0�004Dragonfly �0�02 �0�02 0�54 �0�06 0�03 0�46

2010 Mosquitofish 0�008 0�006 2�10 �0�003 0�02 0�15Crayfish �0�07 0�02 13�05 �0�11 �0�03 <0�001Dragonfly 0�007 0�01 0�27 �0�02 0�03 0�60Vegetation �0�13 1�54 0�007 �3�15 2�88 0�93

*No grey treefrogs were captured in wetlands containing mosquitofish in 2010.



performed Wilcoxon signed rank tests to evaluate invertebrate

abundance and taxa richness of chemically restored wetlands

before and after treatment.

Results

Regression analyses revealed negative associations

between grey treefrogs and mosquitofish during 2007,

2008 and 2009 (all P < 0�05; Table 3). Treefrogs were pos-

itively associated with vegetation in 2007. No treefrog lar-

vae were captured in wetlands containing mosquitofish

during 2010. Treefrog abundance was also negatively

associated with crayfish abundance in 2007 and 2009 and

larval dragonflies in 2008. Treefrogs were most abundant

during 2007 (Fig. 2), but they were captured in only 39%

of the wetlands (Fig. 3). During subsequent years, they

were captured in roughly half of the wetlands but their

abundance dropped and remained at relatively low levels.

Boreal chorus frog tadpoles were never captured in

large numbers (i.e. >15) in wetlands containing mosquito-

fish. We did not perform a regression analysis for chorus

frogs in 2007 because their larvae were captured in only

three wetlands. No fish, crayfish or dragonflies were cap-

tured in these wetlands during the first two sampling peri-

ods when chorus frogs were breeding. Chorus frogs were

negatively associated with mosquitofish in 2008

(P = 0�001) and 2010 (P = 0�009), and they were captured

in 67% of wetlands in 2008 and 44% in 2010. Only two

larval chorus frogs were captured in 2009 and they

occurred in a fish-free wetland. Chorus frogs were also

negatively associated with dragonflies in 2010 (P = 0�01).The peak abundance for chorus frogs occurred during

2008 (Fig. 2). Although their larvae were nearly absent in

2009, both their abundance and occurrence increased

sharply in 2010 (Figs 2 and 3).

Green frogs did not occur in enough numbers to per-

form regression analyses in 2007, but their abundance was

consistently negatively associated with crayfish during

2008, 2009 and 2010 (all P < 0�05). Green frogs were also

negatively associated with dragonflies in 2008 (P = 0�04).There were no statistically significant relationships

between green frog abundance and mosquitofish. Green

frog abundance and occurrence increased over the course

of the study and peaked during the first sampling period

of 2010 (Figs 2 and 3).

Analyses using ANCOVAs revealed that invertebrate abun-

dance was significantly reduced in the fish-stocked wetlands

during all four sampling years (2007: F1,15 = 13�25,P = 0�002; 2008: F1,15 = 21�07, P < 0�001; 2009: F1,15 = 55�15,P < 0�001; 2010: F1,15 = 15�60, P = 0�001). Mean invertebrate

abundance was consistently higher in fish-free wetlands

throughout the duration of our study (Fig. 4). ANCOVAs also

indicated that mosquitofish significantly reduced invertebrate

taxa richness during the first 3 years (2007: F1,15 = 6�9,P = 0�02; 2008: F1,15 = 19�1, P = 0�001; 2009: F1,15 = 14�61,P = 0�002), but not in 2010 (P = 0�56). The vegetation cover

covariate had no significant effects on invertebrate abundance

during any year, but it did significantly increase taxa richness

Sample period2010-32010-22010-12009-32009-22009-12008-32008-22008-12007-32007-22007-1

Mea

n an

uran

abu

ndan

ce

30

25

20

15

10

5

0

Error bars: +/– 1 SE

Green frogGray treefrogBoreal chorus frog

Fig. 2. Mean abundance trends for anu-

rans over all sampling periods. The mean

abundance of chorus frogs was 102 in

2008. The scale of the Y-axis has been

capped at 30 to more clearly illustrate

trends. Fish stocked: N = 9 for 2007–2009and N = 4 for 2010. Fish unstocked:

N = 9 for 2007–2009 and N = 14 for 2010.



in all years except 2007 before natural vegetation cover had

developed (2007: P = 0�48; 2008: F1,15 = 10�04, P = 0�006;2009: F1,15 = 12�35, P = 0�003; 2010: F1,15 = 5�05, P = 0�04).Average invertebrate taxa richness generally increased in all

wetlands throughout the duration of the study, but most

fish-free wetlands were consistently richer during all sampling

periods (Fig. 5).

The rarefied richness curves illustrate the higher individ-

ual abundances in fish-free wetlands, but they also reveal

that in 2007 and 2009, taxa richness reached levels in fish-

stocked wetlands nearly as high as in those without fish,

even though fewer individuals were captured (Fig. 6a).

Because the rarefaction curves for fish treatments fail to

approach an asymptote, and vastly different numbers of

individuals were captured in the two treatments, we also

plotted rarefied richness based on samples (Fig. 6b). The

sample-based curves suggest that taxa richness was some-

what higher within fish-free wetlands, but during 2007,

taxa richness was similar for the two treatments across

samples and in 2010, taxa richness was nearly equal at

samples below 10.

A Wilcoxon signed rank test revealed a statistically sig-

nificant increase in invertebrate abundance (excluding

crayfish, daphniids, snails and bivalves) following rote-

none applications to fish-stocked wetlands, N = 12,

Z = �2�20, P = 0�03, with a large effect size (r = 0�64).The median invertebrate abundance in fish-stocked wet-

lands was 50�5 in 2009 prior to treatment. In 2010, after

treatment, the median increased to 248�5. Invertebrate

abundances were low in the three untreated fish-stocked

wetlands in 2009 and 2010 (2009: mean = 34�0,range = 22–49; 2010: mean = 28�3, range = 17–35).

There was also a statistically significant increase in

invertebrate taxa richness following rotenone application

Fig. 3. Occurrence trends for anurans over all four study years at

all wetlands. Green frogs became more common as wetlands

aged, grey treefrog occurrence was relatively stable, and chorus

frog occurrence was variable, possibly due to predator population

fluctuations.

Sample period2010-32010-22010-12009-32009-22009-12008-32008-22008-12007-32007-22007-1

Mea

n in

vert

ebra

te a

bund

ance

120

100

80

60

40

20

0

Error bars: +/–1 SE

FishNo fish

Fish treatment

Fig. 4. Mean invertebrate abundance

trends over all sampling periods. Wetlands

with fish consistently contained fewer

invertebrates. Fish stocked: N = 9 for

2007–2009 and N = 4 for 2010. Fish

unstocked: N = 9 for 2007–2009 and

N = 14 for 2010.



to fish-stocked wetlands, N = 12, Z = �2�03, P = 0�04,with effect size r = 0�59. The median richness in fish-

stocked wetlands was 7�5 in 2009 prior to treatment, and

in 2010, after treatment, the median increased to 11�5.However, invertebrate taxa richness also increased in the

three untreated fish-stocked wetlands from 2009 to 2010

(2009: mean = 8�7, range = 7–10; 2010: mean 11�3,range = 9–13).

Discussion

AMPHIB IAN COMMUNITY DEVELOPMENT

Our results illustrate the dramatic role aquatic predators

play in wetland community development. While natural

predators altered amphibian communities over time,

introduced mosquitofish impeded community development

from the outset. Chorus frogs and grey treefrogs appeared

to be particularly sensitive to mosquitofish and our results

may reflect avoidance by breeding adults, predation, tro-

phic effects or a combination thereof. Nevertheless, meta-

morph production data recorded at the same wetlands

during 2007 and 2008 using terrestrial pitfall traps and

drift fences reinforce our results (Shulse et al. 2012).

Although some models in this study revealed negative

influences on hylids from invertebrate predators, none

were as consistent as those observed for mosquitofish.

Our wetlands contained water during all four study

years. As a result, wetlands without mosquitofish

developed high populations of invertebrate predators and

by 2010, almost all contained ranid larvae. Heightened

competition from ranids (Faragher & Jaeger 1998; Boone,

Semlitsch & Mosby 2008) and susceptibility to inverte-

brate predators (Skelly 1995; Smith & Van Buskirk 1995)

may explain why larval chorus frogs were nearly absent

by 2009. Some hylids can detect fish and invertebrate

predators in wetlands (Resetarits & Wilbur 1989, 1991;

Binkley & Resetarits 2008). During diurnal early spring

reconnaissance trips in 2007 and 2008, chorus frogs called

selectively from fish-free wetlands, but in 2009, chorusing

had nearly ceased in all wetlands. Instead, frogs chorused

from nearby ephemeral swales and ditches (Shulse, per-

sonal observation). Chorus frogs prefer fish-free wetlands

with vegetation (Shulse et al. 2010, 2012), but dragonflies

Sample period2010-32010-22010-12009-32009-22009-12008-32008-22008-12007-32007-22007-1

Mea

n in

vert

ebra

te ta

xa ra

ichn

ess

12

10

8

6

4

2

0


FishNo fish

Fish treatment

Fig. 5. Mean invertebrate taxa richness over all sampling periods. Although this figure illustrates a clear difference in richness between

the two treatments, the rarefied richness curves based on individuals suggest that differences are due to a sampling effect – there were

simply fewer individuals of most taxa to capture in the wetlands containing fish. Fish stocked: N = 9 for 2007–2009 and N = 4 for 2010.

Fish unstocked: N = 9 for 2007–2009 and N = 14 for 2010.



1,2501,0007505002500

Taxa

30

25

20

15

10

5

0

FishNo fish 2007

1,2001,0008006004002000

30

25

20

15

10

5

0

FishNo fish 2008

3,0002,5002,0001,5001,0005000

Taxa

30

25

20

15

10

5

0

FishNo fish

Individuals

2009

4,0003,0002,0001,0000

30

25

20

15

10

5

0

FishNo fish

Individuals

2010

Taxa

30

25

20

15

10

5

0

27252321191715131197531

FishNo fish 2007

27252321191715131197531

30

25

20

15

10

5

0

FishNo fish 2008

Sample27252321191715131197531

Taxa

30

25

20

15

10

5

0

FishNo Fish 2009 30

25

20

15

10

5

Sample4136312621161161

FishNo fish 2010

(a)

(b)

Fig. 6. Rarefied taxa richness curves for

2007–2010 based on individual abundance

(a) and samples (b). The sample-based

curves suggest richness was higher in fish-

free wetlands, but individual-based curves

reveal many taxa were present in both

treatments, albeit at lower abundances in

wetlands with fish. Fish stocked: N = 27

for 2007–2009 and N = 12 for 2010. Fish

unstocked: N = 27 for 2007–2009 and

N = 42 for 2010.



also appear to prefer similar habitat (Table 2). Dragonfly

populations increased as wetlands aged (Fig. 7c), as did

vegetation cover (See Fig. S1, Supporting Information).

By 2009, chorus frogs may have avoided the wetlands,

even those with high cover, in favour of nearby ephem-

eral, low-cover aquatic habitat containing few predators

and little competition. During 2010, chorusing frogs

returned in limited numbers, but mostly to chemically

restored wetlands (Shulse, personal observation). Our cho-

rusing observations were validated by capture results. In

2010, the highest abundances of chorus frogs were cap-

tured in wetlands that were either 1) chemically restored

prior to the breeding season or 2) had relatively low pred-

atory insect populations, illustrating that these anurans

detect both fish and invertebrate predators. Other studies

have shown that fish removal leads to increased breeding

of fish-sensitive anurans (Br€onmark & Edenhamn 1994;

Vredenburg 2004). Early colonizing amphibians may have

Sample period2010-32010-22010-12009-32009-22009-12008-32008-22008-12007-32007-22007-1

Mea

n m

osqu

itofis

h ab

unda

nce

50

40

30

20

10

0

Error bars: +/–1 SE

Mea

n cr

ayfis

h ab

unda

nce

30

20

10

0

Sample period2010-32010-22010-12009-32009-22009-12008-32008-22008-12007-32007-22007-1


FishNo fish

Fish treatment

(a)

(b)

Fig. 7. Mean abundances of mosquitofish

(a), crayfish (b) and dragonflies (c) over all

sampling periods at all wetlands. Total

mosquitofish plummeted after rotenone

treatments in late 2009 and early 2010, but

numbers increased during summer 2010 at

the stocked wetlands. Crayfish numbers

were similar in both treatments, suggesting

that mosquitofish have little impact on

crayfish populations, and dragonfly abun-

dances were consistently lower in fish-

stocked wetlands. Dragonflies exhibited a

pattern similar to that of most other inver-

tebrate taxa sampled.



a hierarchy of breeding habitat preferences with fish

avoidance as the strongest filter, followed by aquatic

invertebrate predators. Based on the results of Shulse

et al. (2012), this hierarchy can be extended below inverte-

brate predators to include vegetation cover followed by

within-wetland slope. Our observations suggest that the

primary habitat trade-off for breeding chorus frogs is

between exposure to predators and subjection to breeding

site stochasticity.

Grey treefrogs were also less abundant in fish-stocked

wetlands, but they were able to persist in fish-free

wetlands throughout the duration of the study. Their high

numbers during 2007 indicate that they, like chorus

frogs, are early colonizers that prefer wetlands with low

predator levels. Adult female grey treefrogs will minimize

predation risk to their eggs and larvae by avoiding wet-

lands containing fish (Binkley & Resetarits 2008). How-

ever, larvae will often develop bright red pigment on their

tails and altered body shape in the presence of high popu-

lations of aquatic invertebrate predators (McCollum &

Leimberger 1997). This ‘dragonfly morph’ appears less

susceptible to invertebrates than the typical morph

(McCollum & Van Buskirk 1996) and may indirectly con-

tribute to their ability to continue to breed in permanent

wetlands. We often observed ‘dragonfly morph’ grey tree-

frog larvae in our wetlands with varying shades and

amounts of red pigment. Because these anurans are mid-

spring to early summer breeders whose larvae emerge dur-

ing mid- to late summer, highly ephemeral wetlands that

become dry by mid-summer may reduce or eliminate

recruitment. However, breeding later in more permanent

wetlands may expose larvae to the highest seasonal levels

of dragonflies.

The increasing abundance of green frogs over the dura-

tion of the study, like invertebrates, probably reflects

hydroperiod. Green frogs overwinter as larvae and there-

fore require permanent or semi-permanent wetlands.

Green frogs were negatively associated with crayfish,

although it is not clear whether this reflects mortality or

avoidance. Anderson & Brown (2009) observed reduced

hatching of green frogs in the presence of crayfish, even

when the crayfish had no direct access to the eggs. Inter-

estingly, many sparsely vegetated wetlands contained high

populations of crayfish. The negative correlations between

crayfish and vegetation (Table 2) illustrate the effects that

these shredders have upon aquatic vegetation. These

effects, in addition to predation, may explain the negative

associations between crayfish and anurans. Green frog

abundances were never significantly negatively associated

with mosquitofish, perhaps reflective of their ability to

persist with fish. However, they did not appear to be facil-

itated by fish as has been demonstrated for bullfrog tad-

poles (Werner & McPeek 1994; Adams, Pearl & Bury

2003).

While our study did not reveal strong associations

between most anurans and vegetation, grey treefrogs were

strongly positively associated in 2007, perhaps reflecting

the sparse vegetation present at the time. Treefrogs may

have used the planted cordgrass for chorusing or cover.

We also found strong positive associations between total

amphibian metamorph production and vegetation cover

during 2008 at the same wetlands (Shulse et al. 2012).

Mea

n dr

agon

fly a

bund

ance

30

20

10

0

Sample period2010-32010-22010-12009-32009-22009-12008-32008-22008-12007-32007-22007-1


FishNo fish

Fish treatment(c)

Fig. 7. Continued.



Chorus frog metamorph production during 2008 was also

positively associated with vegetation cover, but a model

that combined mosquitofish abundance with vegetation

cover best explained chorus frog abundance that year

(Shulse et al. 2012). Habitat complexity may increase in

importance, surpassing other features such as within-

wetland slope, as wetlands age (Shulse et al. 2012). Stud-

ies have shown that high vegetation cover is important

for tadpole survival (Babbitt & Tanner 1998; Baber &

Babbitt 2004), but predators that easily penetrate dense

vegetation may have lowered the importance of cover in

our models. For some anurans, predator population

crashes (and manipulations that directly alter predator

levels) may be more important modifiers of reproductive

success than habitat alterations.

INVERTEBRATE COMMUNITY DEVELOPMENT

Our results for invertebrates are concordant with those of

previous studies that have demonstrated that mosquitofish

are injurious to aquatic invertebrates (i.e. Hurlbert, Zedler

& Fairbanks 1972; Jassby et al. 1977a,b; Lawler et al.

1999). Mosquitofish may alter aquatic communities by

selectively feeding on large zooplankton reducing pressure

on smaller zooplankton species, phytoplankton and bacte-

ria (Jassby et al. 1977a,b). Studies suggest mosquitofish

are primarily zooplanktivores (Garcia-Berthou 1999;

Reynolds 2009) and may prefer zooplankton to larval

amphibians (Reynolds 2009). We captured a total of 5

daphniids in Sears 1 during sampling period 2010–1 after

chemical restoration which nearly eliminated mosquito-

fish. None were captured during subsequent sampling

periods when the fish population recovered, or at any

other fish-stocked wetland during the study. After autumn

rotenone treatments in 2009 at other wetlands, daphniid

captures increased during the following spring to the high-

est levels recorded (C. D. Shulse & R. D. Semlitsch,

unpublished data). Because daphnia are preyed upon by

other invertebrates, high abundances shortly after fish

removal may reflect ideal conditions before predator pop-

ulations recover.

Our rarefied richness curves (Fig. 6) and invertebrate

taxa captured (Table S1, Supporting information) illus-

trate that most taxa present in fish-free wetlands were also

present in those with fish, although many at compara-

tively very low numbers. This suggests that our ANCOVA

results illustrating lower richness in fish-stocked wetlands,

along with differences in mean richness between the two

treatments (Fig. 5), may be partly explained by mosquito-

fish reducing richness simply through reducing abundance

(i.e. a sampling effect). Culicids, chironomids, gerrids,

amphipods and hydrachnids were also very rare in

fish-stocked wetlands. Unlike daphnia, these were also rel-

atively uncommon in fish-free wetlands. Mosquitofish

likely prefer daphnia, but they appear to prey indiscrimi-

nately on most aquatic invertebrates once daphnia popu-

lations are depleted.

Stewart & Downing (2008) found macroinvertebrate

richness and abundance increased along with vegetation

in constructed wetlands. While our results also indicate

that invertebrate richness is bolstered by vegetation, we

found no evidence that vegetation increased overall

invertebrate abundance. Reynolds (2009) found that high

levels of both aquatic invertebrates and vegetation cover

reduced mosquitofish predation on anuran larvae. How-

ever, our results illustrate that aquatic invertebrate

abundance is severely reduced by mosquitofish, suggest-

ing vegetation cover provided insufficient refuge for

invertebrates.

MANAGEMENT IMPLICATIONS

Complex interactions between predators, wetland hydro-

period and successional processes shape wetland commu-

nities, but mosquitofish break down natural processes,

alter populations of other predators and grazers and

impact multiple trophic levels. Our results suggest that

outside their native range and ecosystems, mosquitofish

reduce the ecological value and conservation potential of

wetlands, particularly those restored or created as com-

pensatory mitigation for the destruction of natural wet-

lands. Some invertebrates and amphibians responded

positively to both fish removal and low invertebrate pred-

ator populations. These conditions are likely to be similar

to those that occur when a wetland re-fills after drying.

However, rotenone can have negative consequences for

amphibians (Fontenot, Noblet & Platt 1994) so caution is

warranted.

Building wetlands of varying sizes and depths creates

hydroperiod diversity across the landscape similar to

natural conditions (Semlitsch 2002; Petranka et al. 2007;

Shoo et al. 2011), and constructing drainable ponds will

allow managers to control hydroperiods. These non-

toxic approaches will ensure that mosquitofish and

other aquatic predators are occasionally eliminated in

some pools, thereby increasing the ecological value of

restored and constructed wetland complexes. Further-

more, native predators should be encouraged to colo-

nize through management regimes that mimic pre-

settlement conditions and natural successional processes.

However, non-native predators can have devastating

impacts and they should not be stocked or allowed to

persist.

Our experimental approach at replicated environments

goes beyond traditional laboratory or observational inves-

tigations to provide a unique comparison between the

impacts of an invasive introduced predator to those of

naturally colonizing predators. We encourage researchers,

ecologists and managers to work collaboratively to incor-

porate experimentation into restoration projects. Doing so

will yield valuable information that will improve the eco-

logical value of restoration projects, reduce the threat of

invasive species and increase the conservation potential of

restored habitats.



Acknowledgements

We thank the Missouri Department of Transportation, Missouri Depart-

ment of Conservation, University of Missouri Division of Biological Sci-

ences and University of Missouri Department of Civil and Environmental

Engineering. Thanks to S. Becker, K. Kettenbach, L. Rehard, D. Kuschel,

A. Robertson, D. Lund, A. Leary, B. McMurray and G. Schmitz for assis-

tance in the field. Thanks to K. Smith and an anonymous reviewer for

insightful comments on an earlier version of this manuscript. This project

was funded by a United States Environmental Protection Agency Region

VII Grant CD-98769101-0, a Missouri Department of Conservation Wild-

life Diversity Fund Grant and a Missouri Department of Transportation

Research and Development Grant (RI 07-005). Organisms were captured

under Missouri Department of Conservation Wildlife Collector’s Permits

13438, 13769, 14120 and 14533 and University of Missouri Animal Care

and Use Protocol 4189.

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Received 26 October 2012; accepted 23 May 2013

Handling Editor: Marc Cadotte

Supporting Information

Additional Supporting Information may be found in the online version

of this article.

Fig. S1. Vegetation cover at each wetland over the course of the

study.

Table S1. Invertebrate taxa captured at wetlands based on mos-

quitofish treatment.



mosquitofish dominate amphibian and invertebrate community development in experimental wetlands

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