EFFECTS OF INTRODUCED PEACOCK ON NATIVE LARGEMOUTH SALMOIDES IN SOUTHEAST

By

JEFFREY E. HILL

A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY

UNIVERSITY OF FLORIDA

2003 ACKNOWLEDGMENTS

Many people provided tremendous assistance in the completion of this project.

All deserve my acknowledgment. I apologize beforehand to anyone inadvertently omitted.

The primary acknowledgment goes to my wife, Susan, for her unfailing support

throughout my entire graduate career. I thank her for sacrificing in order for me to fulfill

our shared goal. 1 especially thank her for putting up with my obsession for .

I am gratefial for the constant support of my family—^my father, mother, and sister

(Baker, Jacqueline, and Kim). They shared the dream of my doctorate—I am thankfial to have fulfilled our collective aspiration.

I greatly appreciate the guidance and support of my doctoral committee—Drs.

Charles E. Cichra (Chair), Carter R. Gilbert, William J. Lindberg, Leo G. Nico, and Craig

W. Osenberg. It was a great pleasure to work for Dr. Cichra—the experience I received in extension, research, and teaching, along with strong mentorship and his fiiendship, were instrumental in my professional development. Carter Gilbert has been a tremendous

influence and is one of my real "fish heroes". I especially thank Carter for staying involved in my graduate education after his retirement—that meant a lot to me. Leo

Nico's remarkable field experience with nonindigenous and south Florida systems was invaluable. The many discussions (and some disagreements) over nonindigenous

fishes with Leo helped shape my scientific philosophy. I appreciate the support and

ii encouragement that Bill Lindberg always gave. Our discussions inside and outside of the classroom about ecology and the philosophy of science (e.g., tautology!) helped open my

eyes to the wider world of science. I thank Craig Osenberg for agreeing to help out on

the committee of someone studying a crazy system like the south Florida canals. I especially appreciate his assistance with preparing for my qualifying exam and his

extraordinary input during the revision process. I have learned a great deal from each committee member and thank them for being a part of my graduate education.

A special acknowledgment is due to Jeff Sowards for all of the hard work in the field. Thirty-six hour "days" can be rough, but he never complained. Jeff was instrumental in the field work, in the lab, and as an "ideas man" to keep the project

doable. I could not have done it without his help. Also, Sharon Fitz-Coy was a tremendous help in, among other things, collecting prey for experiments and identifying insects in fish stomach contents. She was always willing to help.

Special thanks are due to Paul Shafland (Florida Fish and Wildlife Conservation

Commission; FWC) for a great deal of assistance, facilitation, literature, and advice during this project. Paul's largest contributions, however, came during many discussions of scientific philosophy regarding nonindigenous fish.

Craig Watson and Roy Yanong (Tropical Aquaculture Laboratory) provided exceptional support throughout the process. They made the tribulations more bearable.

Debbie Pouder facilitated all work done at the Sam Mitchell Aquaculture

Demonstration Farm (SMADF) in Blountstown, Florida. I appreciate her help and that of the staff, especially Randall Kent.

iii Ken Portier of the Department of Statistics at the University of Florida helped

with the design and analysis of experiments in Chapter 5.

I appreciate the hospitality of Don and Judy Bemecker who provided a great place

to stay and many delicious meals during field work.

Many others are due thanks for their contributions. At the University of Florida,

Department of Fisheries and Aquatic Sciences, Mike Allen, Mary Cichra, Doug CoUe,

Ruth Francis-Floyd, Tom Glancy, Scott Graves, Robert Leonard, Allen Riggs, Beth

Sargent, Troy Thompson, Larry Tolbert, and Dan VanGenechten helped with various

aspects of the project. Kelly Jacoby drew the maps. Susan Morgan greatly facilitated the

graduation process. At the Florida Caribbean Science Center, U.S. Geological Survey,

Gary Hill, Howard Jelks, and Bill Stranghoener assisted with facilities and equipment.

Frank Morello, Jerry Krummrich, and Fred Cross (FWC) issued scientific collecting

permits; Ted Hoehn (FWC) kindly provided access and instruction to the FWC fish

database; Eric Nagid (FWC) helped answer questions about use and content of the

database.

I gratefully acknowledge the financial and facilities assistance of the Department

of Fisheries and Aquatic Sciences, the Sam Mitchell Aquaculture Demonstration Farm,

and the Tropical Aquaculture Laboratory of the University of Florida. Financial

assistance was further provided by a Dean's Fellowship for Graduate Research from the

Institute of Food and Agricultural Sciences, University of Florida, the Florida

Department of Agriculture and Consumer Services, by Ryan Kelley Memorial

Scholarships from the International Women's Association, and the Roger

Rottmann Memorial Scholarship from the Florida Chapter, American Fisheries Society.

iv TABLE OF CONTENTS gage

ACKNOWLEDGMENTS ii

ABSTRACT vii

CHAPTER

1 INTRODUCTION 1

2 STUDY SYSTEM 11

Study 11 Canal System 14 Study Sites 18

3 A QUALITATIVE RISK ASSESSMENT OF PEACOCK CICHLIDS CICHLA OCELLARIS AND SOME IMPLICATIONS FOR THE METHODOLOGY 24

Introduction 24 Background Information 27 Qualitative Risk Assessment 36 Discussion 43

4 ONTOGENETIC DIETARY SHIFTS AND DIETARY OVERLAP OF A NONINDIGENOUS FISH AND AN ECOMORPHOLOGICALLY SIMILAR NATIVE FISH 49

Introduction 49 Methods 51 Results 56 Discussion 68

5 EXPERIMENTAL PREY HANDLING AND SELECTIVITY OF TWO MORPHOLOGICALLY SIMILAR PREDATORY FISHES- NATIVE AND INTRODUCED PEACOCK CICHLIDS 79

V Introduction 79 Methods 82 Results 90 Discussion 99

6 THE INTRODUCTION OF PEACOCK CICHLIDS INTO SOUTHEAST FLORIDA: POPULATION-LEVEL CONSEQUENCES FOR LARGEMOUTH BASS 109

Introduction 1 09 Methods 112 Results 114 Discussion 119

7 SUMMARY 127

REFERENCES 133

BIOGRAPHICAL SKETCH 154

vi Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy

EFFECTS OF INTRODUCED PEACOCK CICHLIDS CICHLA OCELLARIS ON NATIVE LARGEMOUTH BASS MICROPTERUS SALMOIDES IN SOUTHEAST FLORIDA

By

Jeffrey E. Hill

December 2003

Chair: Charles E. Cichra Major Department: Fisheries and Aquatic Sciences

The predatory peacock Cichla ocellaris , native to South America, was

introduced into southeast Florida to enhance and to provide

biological control over nonindigenous fishes. This action was controversial and there is

concern over the effects of the introduction on native species, especially the

ecomorphologically similar largemouth bass Micropterus salmoides . I conducted a

formal qualitative risk assessment to evaluate the risks associated with the peacock

cichlid introducfion and to test predictions of the algorithms. The esfimate of risk for peacock cichlids was Low. The evaluation elucidated shortcomings in the methodology

(e.g., emphasis on a single output) and I made recommendations for its improvement.

Part of the concern over the peacock cichlid introduction is their similarity to largemouth

bass. I tested and found support for the hypothesis that morphological similarity equals

similarity in prey handling ability and prey selection. I documented substantial overlap in

vii trophic morphology and dietary ontogeny between juveniles and found similar prey handling ability and prey selection by adults in tank experiments. These results suggest that largemouth bass and peacock cichlids have similar effects on the prey base.

However, largemouth bass had exclusive use of highly defensive prey (i.e., crayfish).

Largemouth bass populations might be expected to have declined, given evidence of

direct negative effects of peacock cichlids. I investigated this question by comparing largemouth bass populations in southeast Florida canals with populations in lakes and

streams (statewide and in south Florida). Additionally, I compared canals with peacock cichlids to canals without peacock cichlids. The data did not support a hypothesis of declines in largemouth bass abundance due to peacock cichlids. These results question common assumptions concerning the effects of nonindigenous species— 1) that direct effects of predation and competition are the predominant effects and 2) the narrow view that invaded communities are organismic (i.e., Clementsian)—and challenge the uncritical acceptance of the premise that nonindigenous fishes will invariably cause ecological harm. My results suggest that other factors (e.g., indirect effects) may be important and that invasion biology should develop a pluralistic theory that incorporates

Gleasonian concepts (i.e., individualistic communities).

viii CHAPTER 1 INTRODUCTION

Introduced species are considered threats to natural ecosystems and biodiversity

worldwide (Lodge 1993). For example, Miller et al. (1989) reported that nonindigenous species were a factor in the extinction of 27 North American fishes (68% of extinctions).

Similarly, Lassuy (1995) stated that nonindigenous species were factors in listing 48 imperiled freshwater fishes under the Endangered Species Act in the USA (70% of listed

fishes). Although ecological damage due to nonindigenous species is difficult to esfimate, economic costs in the billions of dollars are cited for aquafic and terrestrial

systems in the USA (OTA 1993). Indeed, the overwhelming consensus is that species introductions may result in deleterious effects (e.g., lowered abundance of native species) within the receiving system (Magnuson 1976; Courtenay 1993, 1995; Minns and Cooley

2000).

Nonindigenous fishes have a long history in the USA and have entered this

country through several intentional and unintentional pathways (Fuller et al. 1999).

Many fishes have high economic and social value and have been purposefially stocked to

provide fisheries benefits. For example, centrarchids (e.g., Lepomis and Micropterus ).

percids (e.g., Sander [formerly Sfizostedion] ). salmonids (e.g., Oncorhynchus and

Salmo ), and other fishes have been transferred to many regions to enhance recreational fisheries (Fuller et al. 1999; Heidinger 1999). Purposeftil introducfions for recreafional fishing are perhaps the most studied cases of interaction between native and

1 2

nonindigenous fishes. Nevertheless, thorough empirical evaluations are few (but see

Huckins et al. 2000). Such research is costly and may be controversial (e.g., Churchill et

al. 2002). Introduced salmonids have been intensely scrutinized, yet there is wide variation in their reported effects on native salmonids (Fausch 1988). In a review of the

effects of salmonid introductions, Fausch (1988) found that some authors documented

little or no effect of introduced salmonids, yet other authors concluded that such fishes

constitute a serious threat to native species (see also Fuller et al. 1999; Crawford 2001).

Similarly, data for non-salmonid introductions are often mixed and inconclusive (Taylor

et al. 1984; Shafland 1996a; Courtenay 1997; Fuller et al. 1999). Regardless of the potential or realized benefits, the possible dangers have led to many calls for a cessation

of purposeful fish introductions (Magnuson 1976; Courtenay and Robins 1973, 1989;

Courtenay 1995; Minns and Cooley 2000). These papers point out the numerous

theoretical negative impacts of fish introductions and that some introduced fishes have

proven to be notable and costly pests (e.g., the unintenfionally introduced sea lamprey

Petromyzon marinus in the Great Lakes; Mills et al. 1994).

Given increasing transfers of fishes outside their natural ranges (Nico and Fuller

1 999), there has been significant scientific and public interest in predicting the effects of species invasions on native aquatic faunas (Lodge 1993; OTA 1993; Moyle and Light

1996a; Kolar and Lodge 2002). Indeed, inadequate predicfive ability in this field has

frustrated scientists and resource managers. This situafion is partly due to the

surprisingly few empirical evaluafions of the effects of nonindigenous fishes (Taylor et

al. 1984; Moyle et al. 1986; Fuller et al. 1999). In fact, much of the literature on harmfiil effects is anecdotal. Furthermore, a lack of a truly pluralistic ecological theory of bio- ;

3

invasion may be a factor hampering the advance of this field (Schoener 1987; Ross 1991 see also Lodge 1993; Williamson 1996; Shrader-Frechette 2001; Sakai et al. 2001).

General ecological theory provides a fi-amework for a variety of predictions of the effects of species introductions. For example, there is a rich literature concerning the effects of species introductions and removals in temperate lakes on lake productivity and

food web structure (e.g., Carpenter et al. 1985; Carpenter et al. 1987; Mittelbach et al.

1995) . Some researchers have used ecological theory to predict and subsequently

document negative effects of species introductions (e.g., Huckins et al. 2000); others have predicted and then demonstrated effects that were positive for aquatic resources (e.g.,

Carpenter et al. 1985; Carpenter et al. 1987). Therefore, ecological theory is not biased toward an invariable prediction of negative effects.

Despite this rich body of ecological theory, many scientists and resource managers interested in the ecological effects of nonindigenous species assume that invading species have only net negative effects on native species (see reviews in Stauffer

1984; Taylor et al. 1984; Moyle et al. 1986; Li and Moyle 1999; but see Williamson

1996) . The current emphasis on negative effects of introduced species are founded on two principles: 1) the organismic concept of equilibrial, stable, and competitively- structured communities (Clements 1936) and 2) that direct effects of (via competition and predation) are far more important than direct or indirect positive effects that might negate or overwhelm the direct negative interactions. The expectation that competition will play a key role was influenced by Cause's articulation of the

competitive exclusion principle (Cause 1 934; Hardin 1 960), the Hutchinsonian concept of niche (Hutchinson 1957), Elton's (1958) work on species invasions, and MacArthur 4

and colleague's papers on competition and island biogeography (e.g., MacArthur and

Levins 1967; MacArthur and Wilson 1967). Such theory (e.g., island biogeography)

predicts that any species addition must lead to a species extinction (MacArthur and

Wilson 1967). Under this view, the observed increase in invasions (e.g., Nico and Fuller

1 999) is certain to lead to increased rates of extinction (or extirpation) of native species

(and previously established nonindigenous species)—a prediction leading to considerable

concern for conservation biologists and resource managers over a loss of biodiversity.

Furthermore, the native species most at risk of extinction should be those species that are

most similar to invading species with respect to resource use (Hardin 1960; MacArthur

and Levins 1967; Diamond 1975; but see Resetarits 1995). Because similarity in

resource use is often associated with similarity in morphology (Wainwright 1988, 1996;

Wainwright and Richard 1995), species of similar morphology are expected to be the

most sensitive to invading species. Moreover, proponents of this view of species

invasions aggressively challenge any notion of "empty niches" on semantic and

theoretical grounds (e.g., Herbold and Moyle 1986; Courtenay 1995), practically

precluding the possibility that species introductions can take place without any

deleterious effects for the receiving system.

Nonindigenous fishes may adversely affect native species directly or indirectly

via competition, predation, habitat alteration, reproductive inhibition, and shared

pathogens (Taylor et al. 1984; Moyle et al. 1986). Indeed, many conclusions of impacts

of nonindigenous fishes assume that these mechanisms of negative interaction will

directly translate into reductions in the density of native species. If native fish populations are low or are perceived to have declined coincidentally with the introduction of a nonindigenous fish, then evidence for the existence of a putative mechanism is often

assumed to be proof that the mechanism is a contributing factor. This apparent cause-

and-effect scenario may be correct for some cases of species introductions. Nevertheless,

the existence of a negative interaction between the two species may not be important to

the native species at the population level if, for example, the total strength of the

interaction (per capita effect x population size) is low, if other factors (e.g., abiotic

events) are strong, or if density-dependence at later life stages leads to compensatory

responses (e.g., Vonesh and De la Cruz 2002). As a result, data are required that directly

link the interactions between introduced and native species and the observed effects on

population density.

Despite these views that invasions should lead to reductions in abundance or

extinctions of native species, there is growing evidence that new species can be added to

native communities with little noticeable effect (Moyle and Light 1 996a; Williamson

1996; Gido and Brown 1999). Indeed, perhaps as many as 80-90% of all established

nonindigenous species fall into this category (Williamson 1996). For example, Moyle

and Light (1996a) noted several examples of invading fishes integrating into Ireshwater

and marine communities without obvious changes in native species.

A partial explanation for the failure of many scientists and resource managers to

expect observed cases of integration of invading species into communities is a conscious or tacit adherence to the Clementsian view of organismic communities (Clements 1936).

On the other hand, if communities are considered loosely-structured assemblages of species that result from probabilistic processes, that are transient in time and space, and are dependent on the characteristics of individual species (individualistic concept; Gleason 1926), then integration of invaders into local assemblages is a possibility.

Clearly, there is a great deal of evidence that plant and associations have changed

over time (e.g., Graham et al. 1996), implying that communities may change in species composition and are not fixed in structure. Exchanges of species during numerous faunal

interchanges throughout the Neogene period (i.e., last 25 million years) have often led to enrichment of faunas rather than species replacements, suggesting that many communities are not saturated with species (Vermeij 1991). Moreover, Rosenzweig

(2001) argued on theoretical grounds that species introductions often permanently increase local biodiversity. Likewise, Gido and Brown (1999) reviewed fish community data for 125 drainages in North America and concluded that fish introductions generally increased species richness and that this pattern suggests that North American fish communities are not saturated with species. Mcintosh (1995) reviewed the controversy associated with the competing paradigms of Clements and Gleason and noted that,

although debate is ongoing, many animal ecologists had joined with plant ecologists in viewing communities as transient, individualistic assemblages of species.

Although it is a mistake to assume that biotic factors are not important in

individualistic communities, strong community structuring by biotic factors is a defining

characteristic of the organismic concept of community (Mcintosh 1 995). The emphasis of the organismic concept is apparent in much of the literature on nonindigenous fishes.

For example, even though abiotic factors are frequently considered when evaluating risks of establishment of nonindigenous fishes (Baltz and Moyle 1993; Moyle and Light

1996a, b; RAM Committee 1998), most studies have considered biotic interacfions to be

preeminent once a nonindigenous species becomes established (Taylor et al. 1984; 7

Courtenay 1993; but see Moyle and Light 1996b). However, in his review of fish

introductions and stream fish assemblage structure, Ross (1991) pointed out that the

establishment of a nonindigenous species with no apparent effect on the native

assemblage suggests strong abiotic effects within the community.

Another important reason for the general prediction of harm to native species

following a species introduction is the prominent emphasis on the direct negative effects

of competition and predation. Both mechanisms can have strong negative effects on

species abundance (Garman and Neilson 1982; McComas and Drenner 1982; Crowder

and Binkowski 1983; Kaufinan 1992; Mills et al. 1994). However, indirect effects can be

both facilitative and powerfiil. For example, Lawlor (1979) reported that 30-40% of

interactions in eight bird communities were positive. Miller (1994) estimated the

magnitude of direct and indirect interactions of five weedy plant species. He found that

although direct negative effects were large, indirect effects were generally positive,

relatively strong, and reduced the influence of competition. Additionally, Stone and

Roberts (1991) argued that a high proportion (i.e., 20-40%) of species interactions must

be advantageous within the context of a community and that "hypercompetitive

communities" (i.e., when every species suffers ft-om the presence of every other species) must be rare. In some cases, positive interactions between putative competitors can

overwhelm direct negafive effects and result in increases in populafion abundance (i.e., facilitation) (e.g.. Miller 1994).

The exotic peacock cichlid Cichla ocellaris (Cichlidae) was intenfionally introduced into southeast Florida canals in 1984-1987 to provide biological control of other exotic fishes and to create a novel sport fishery (Shafland 1995). This relatively 8

large, predatory species is ecomorphologically similar to the native largemouth bass

Micropterus salmoides () (Norton and Brainerd 1993) that already occurred

in these habitats. In their presentation of empirically-derived "rules" of biological

invasions, Moyle and Light (1996a) concluded that the piscivore trophic group is one of the most likely to become successful invaders and to subsequently alter the fish

assemblages of receiving systems. The ecomorphological similarity between these two

species is also of concern given the potential for resource competition. Indeed, there has been reported substantial overlap in size-specific gape width (Hill 1998), diet (Shafland

1995; 1999b), and habitat use (Lilyestrom and Churchill 1996) between largemouth bass

and peacock cichlids. Additionally, there has been speculation and concern regarding possible deleterious effects of peacock cichlids on native fishes in Florida (Courtenay and

Robins 1989; OTA 1993; Courtenay 1994; Cox 1999; see also Lachner et al. 1970).

Indeed, Cox (1999) reviewed species introductions into North America and and listed the peacock cichlid as an "invasive exotic" species in Florida, a designation

requiring actual or probable harm resulting from its introduction.

To evaluate the actual effect of peacock cichlids on native largemouth bass, I first considered the potential direct and indirect effects of the introduction on largemouth bass.

Using these theoretical effects and additional biological information, I conducted a

qualitative risk assessment (RAM Committee 1 998) to provide a framework to formalize the predictions of peacock cichlid effects and to evaluate this methodology using peacock

cichlids as a test case (Chapter 3).

Given the natural history of this system, predation by peacock cichlids on small largemouth bass and competition for food between similar size classes of peacock 9

cichlids and largemouth bass were the most likely interactions; in the absence of mitigating factors, both would lead to a decline in largemouth bass population density in response to an introduction of peacock cichlids. Other potential mechanisms commonly cited in the literature such as spawning site competition, habitat alteration, hybridization,

and introduction of disease (Taylor et al. 1984; Li and Moyle 1999) were considered but

discarded as unlikely. Although both species are substrate spawners, there is relatively

little temporal overlap in spawning activity (Shafland 1999b). Unlike common carp

Cyprinus carpio , grass carp Ctenopharyngodon idella , or some species

( Oreochromis spp., Tilapia spp.) (Taylor et al. 1984; Fuller et al. 1999), peacock cichlids are not known to alter habitat by destroying aquatic macrophytes, by disturbing sediments, or by any other mechanism. The possibility of hybridization is extremely remote because peacock cichlids are phylogenetically distinct from largemouth bass (or

any other native Florida fish) at the familial level. Additionally, samples of peacock cichlids were screened for known pathogens by personnel of the US Fish and Wildlife

Service prior to release (Shafland 1 995). On the other hand, a diet study conducted in

Florida confirmed the use of largemouth bass as prey by peacock cichlids in Florida

(Shafland 1999b). However, the potential for food competition had received little

investigation (but see Shafland 1995, 1999b; Hill 1998). In the present study, I assessed food competition indirectly, largely through overlap in food resource use. Some work on trophic overlap had been conducted with subadult and adult largemouth bass and peacock cichlids; Hill (1998) had previously investigated their similarity in gape size and Shafland

(1 999b) had compared diets of both species in a southeast Florida canal. I therefore determined overlap in trophic morphology (i.e., morphological ability to handle prey) and ;

10

dietary overlap from stomach contents using small juveniles (i.e., up to about 160 mm

total length). Additionally, I investigated patterns of prey selectivity of subadult and

adult largemouth bass and peacock cichlids in two experimental settings. Although

overlap in resource use is not a direct proxy for competition (Colwell and Futuyma 1971

Sale 1979; Schoener 1982), shared use of a resource is generally considered to be a

prerequisite for exploitation competition (Moermond 1979; Sale 1979; but see

Rosenzweig 1981 for situations where competition is an important interaction but overlap

in resource use is essentially zero; Chapter 4). Additionally, resource overlap is often

used to justify conclusions of interspecific competition in evaluations of species

introductions.

Another objective for the present study was to test the hypothesis that peacock

cichlids have had large negative effects on largemouth bass populations in southeast

Florida. Based on the expectation that predation and food competition lead to reductions

in largemouth bass density, I predicted that south Florida canals would have reduced

abundance of largemouth bass relative to other types of Florida systems (i.e., lakes and

streams) and that canals with peacock cichlids would have lower abundance of

largemouth bass than canals lacking peacock cichlids. Therefore, I used population data

for largemouth bass from south Florida canals and other Florida systems to test these hypotheses. An additional objective was to determine if the present study supported the

prevailing view among many invasion biologists based on Clementsian (i.e., organisimic)

communities or a more pluralistic alternative that incorporates Gleasonian (i.e., individualistic) communities. CHAPTER 2 STUDY SYSTEM

Study Species

The largemouth bass Micropterus salmoides (Centrarchidae) is native to much of

eastern North America, including southern Florida where it is considered a distinct subspecies, the Florida largemouth bass Micropterus salmoides floridanus (Lee 1980).

The biology and ecology of largemouth bass are well studied (e.g., Heidinger 1975;

Carlander 1 977). Largemouth bass is a habitat generalist, with small individuals occurring in or near cover and larger individuals occupying a wide range of cover types

and open water habitats (Keast et al. 1978; Mesing and Wicker 1986; Armett et al. 1996;

Annett 1998; pers. obs.). It is predatory, with a general ontogenetic dietary progression from to insects and small crustaceans to decapod crustaceans and fish (Chew

1974). Adults are considered primarily piscivorous (Chew 1974; Carlander 1977;

Cailteux et al. 1996). Nests are constructed by males in relatively shallow water, often near cover. Largemouth bass in southern Florida in winter to early spring (peak in

March in canals) over a relatively short time period (Shafland 1 999b). This results in a more or less distinct yearly cohort that appears in spring and declines in numbers through the year (pers. obs.).

Members of the Cichla (Cichlidae) are relatively large, predatory fishes that

superficially resemble the North American black basses (i.e., Micropterus ). Cichla are native to most major Atiantic drainages of tropical South America. In English, this genus

11 12

is known as peacock cichlids or . Cichla comprise a group of five described and perhaps six or more undescribed species (Machado-Allison 1971; Kullander 1986;

Kullander and Nijssen 1989). To date, the introduced form established in Florida has

been designated as the peacock cichHd C. ocellaris , a species considered native to northeastern South America, although there remains some uncertainty about the alpha of the genus (see Kullander and Nijssen 1989).

During 1984-1987, fmgerling peacock cichlids were intentionally introduced into freshwater canals in Broward and Miami-Dade counties of southeast Florida by the

Florida Game and Fresh Water Fish Commission—currently Florida Fish and Wildlife

Conservation Commission; FWC (Shafland 1 995). The introduction of this piscivore was planned to provide a biological control on the large populations of prey fish, primarily of exotic origin, to correct an unbalanced fishery (sensu Swingle 1950) and to supplement recreational fisheries that were principally based on largemouth bass (Shafland 1995).

Peacock cichlids spawn in nests constructed in relatively shallow water

throughout the warmer months (Zaret 1 980a). Therefore, unlike largemouth bass, cohorts of age-0 peacock cichlids arise throughout much of the year, although they appear later than the single largemouth bass cohort. Peacock cichlids are habitat generalists and occur

syntopically with largemouth bass at all life stages (pers. obs.).

Largemouth bass and peacock cichlids are not naturally sympatric. Nevertheless, these species co-occur in Hawaii and Puerto Rico due to the introducfion of both species and in Florida due to the introducfion of peacock cichlids. Both species are important sport fish in all three regions. Devick (1980) reported that peacock cichlids in Hawaiian reservoirs complemented largemouth bass as a sport fish due to differing spawning 13

seasons. Water withdrawals from the reservoirs often led to loss of nests or fry of

substrate spawning fishes (e.g., largemouth bass and peacock cichlids). The extended

breeding period of peacock cichlids relative to largemouth bass often allowed its successftil reproduction and subsequent recruitment into the sport fishery, in effect carrying the sport fishery over periodic years of largemouth bass recruitment failure

(Devick 1972; Devick 1980). Lilyestrom and Churchill (1996) found both species to

coexist in Puerto Rican reservoirs with little evidence of competition based on slight differences in habitat use and diet and major differences in diel activity periods and spawning seasons. Shafland (1995, 1999a, b, c) noted overlap in diet but concluded that peacock cichlids had not harmed largemouth bass populations in Florida canals. Also of note is the fact that two or more species of peacock cichlids co-occur in many South

American river systems (Jepsen et al. 1997; Winemiller et al. 1997; L. G. Nico, USGS, pers. comm.).

Largemouth bass and peacock cichlids are ecomorphologically similar based on gape width, general morphology, feeding mechanics, and diet (Norton and Brainerd

1993; Hill 1998). Given this similarity in ecology, a few experimental studies have explored possible competitive interactions between largemouth bass and peacock cichlids. Swingle (1967) concluded that peacock cichlids were less efficient predators of

tilapia (probably Tilapia zillii but listed as T. melanopleura) and fathead minnows Pimephales promelas than largemouth bass under a similar stocking rate and conditions in an pond. Lilyestrom et al. (1994) stocked age-0 largemouth bass

(mean TL = 48.8 ± 1 .0 mm) and peacock cichlids = 1 1 (mean TL 8. ± 1 . 1 mm) into experimental ponds in Puerto Rico and found that largemouth bass reduced growth and 14

survival of peacock cichlids but that largemouth bass grew faster and survived better with

peacock cichlids. As pointed out by the authors, the results might have been strongly

influenced by the disparity in initial sizes of the two species, although these were chosen

to reflect the size relationship early in the peacock cichlid spawning season.

Although peacock cichlids have been introduced into several regions, there is little evidence of deleterious effects except in Lake Gatun, Panama (Zaret and Paine

1973) and in Venezuelan reservoirs (cited in Winemiller et al. 1997). The accidental introduction of peacock cichlids into Lake Gatun, Panama, and the subsequent environmental harm, is one of the most cited examples of the danger of fish introductions. Native fishes in the massive reservoir were reportedly devastated by this novel predator and local extirpations occurred (Zaret and Paine 1973). The predation effects apparently cascaded throughout the food web and resulted in significant changes in zooplankton, insect, and bird populafions (Zaret and Paine 1973). Less frequently cited is the report of Welcomme (1988) menfioning the rebound of nafive fish populations in subsequent years.

Canal System

Southeast Florida is drained by a vast network of artificial canals. Much of the canal network was constructed during the early to mid 20"' Century (mainly fi-om 1913 to

1969), primarily to lower the local water table to create agricultural land, to facilitate

urban development, and to protect these lands from flooding (VanArman et al. 1984).

Flood control is largely achieved by redirecting water fi-om the -Lake

Okeechobee-Greater system generally south and east to the coast.

Historically, most of the flow of this system was to the south and southwest. Although 15

some reaches are channelized streams, most of the canal system is entirely new, permanent, deep-water habitat—a freshwater habitat that is naturally scarce in southern Florida (Loftus and Kushlan 1987). Many canals have substantial connection with the surficial aquifer, especially in eastern and southern portions of the region. The

canal system is operated and maintained by the South Florida Water Management District

(SFWMD) which controls flow and water routing with a combination of pump stations and water control gate structures.

Although canals are found in much of southern Florida, peacock cichlids are largely confined to systems along the southeast coast, particularly the eastern portions of

Broward and Miami-Dade counties (about 530 km of canal; Shafland and Stanford 1999).

This species is also found in some southern Palm Beach County canals (total range of about 650 km of canals; Shafland and Stanford 1999). In Florida, peacock cichlids generally inhabit canals that flow through suburban and urban areas, including the densely-populated Ft. Lauderdale and Miami metropolitan areas. Some are found in rural

canals (i.e., pass through agricultural lands) in western portions of the peacock cichlid

range. Peacock cichlid distribution is limited by cool winter temperatures (Shafland

1995; i.e., lower lethal temperature of 15° C; Swingle 1967; Guest et al. 1980). Its range

expands slightly north (i.e., into Palm Beach County) and west during years with mild winters.

Canal morphometry is characterized as box-cut, with vertical walls cut deeply

(often > 3-5 m) into the substrate. In much of the region, the substrate is porous coral rock. A narrow shelf (usually 1-2 m wide), often with submersed vegetation, occurs along the banks and may be the only littoral zone for the canal. Smaller, lateral canals 16

are found in many systems. These are often shallower and have more stable aquatic macrophyte communities (pers. obs.).

Water chemistry can vary widely both temporally and spatially throughout the canal system (Shafland 1999a; pers. obs.). Intermittent releases of water from Lake

Okeechobee or the Water Conservation Areas south of the lake greatly increase flow and alter water physico-chemical parameters (e.g., alkalinity, color, dissolved oxygen, pH,

temperature, and turbidity). These changes can occur rapidly (i.e., within minutes or hours), especially in larger canals that are used as main conduits for de-watering. In particular, water releases can stress fishes by inducing hypoxic conditions. Tidal influences are minimal due to salinity control gates near the coast. These structures only open to release fresh water, although euryhaline fishes may swim up the current and enter the freshwater canal system.

The primary purpose of the canals is drainage and therefore SFWMD maintains an extensive program to remove aquatic macrophytes and overhanging terrestrial vegetation. Many canals support dense stands of aquatic macrophytes, often exotic

invasives (e.g., hydrilla Hydrilla verticillata ), that inhibit flow. The SFWMD uses

biological (i.e., grass carp Ctenopharvngodon idella ), chemical, and mechanical control methods, but mechanical harvesting is the most common. Mechanical harvesting occurs throughout the year and may be frequent in main canals during warmer months.

Coverage of aquatic macrophytes exceeding 90% (percent areal coverage, PAC) can be reduced to < 5% in a single day by mechanical harvesters (pers. obs.). Such drastic changes in habitat structural complexity likely have dramatic effects on invertebrate and fish populations, including largemouth bass and peacock cichlids. Aquatic macrophytes 17

serve as important habitat and as a refuge from predation for many invertebrates and small fishes (Savino and Stein 1982; Werner et al. 1983; Gilinsky 1984; Chick and

Mclvor 1997; Jacobsen and Perrow 1998; Burks et al. 2001). Individual fishes, including age-0 largemouth bass and peacock cichlids, are also removed directly within the harvested vegetation (L. G. Nico, USGS, pers. comm.; pers. obs.). Electrofishing catch- per-unit-effort (CPUE) for age-0 individuals is significantly lower immediately after harvesting activities (unpubl. data).

The peacock cichlid is one of about 25 exotic fish species established in Florida

(Fuller et al. 1999; Hill 2002), with about 18 of these established in southeast Florida

canals (Loftus and Kushlan 1987; Fuller et al. 1999; Loftus 2000; Hill 2002). In the

canals, nonindigenous fishes, particularly cichlids, may dominate the fish assemblage in

numbers and biomass (Courtenay et al. 1974; Shafland 1999a). The native fish fauna is

relatively depauperate, with about 43 freshwater and 1 3 euryhaline species (Loftus and

Kushlan 1987; Loftus 2000; Table 1). The native freshwater species are affiliated mostly with North American temperate families and are relatively recent invaders due to the

short geologic time that southern Florida has been exposed above sea-level (Gilbert

1987).

The general perception in the literature is that native fishes in south Florida urban canals are subjected to numerous stressors (e.g., canal morphometry, variable and often harsh water physico-chemistry conditions, frequent aquatic macrophyte manipulation, and presumably strong biotic interactions with exotic fishes) that result in reduced population abundances (Loftus and Kushlan 1987; Courtenay 1997; Annett 1998). For example, several authors have suggested that cichlids exert strong negative effects on 18

native fishes (Buntz and Manooch 1968; Hogg 1976; Noble and Germany 1986;

Courtenay and Robins 1989; Courtenay 1997). On the other hand, freshwater canals that traverse marsh areas serve as important refligia for large-bodied fishes (native and nonindigenous) during low water periods and generally enhance populations of such fishes (Loftus and Kushlan 1987); these canals are found mostly west of the distribution of peacock cichlids.

Study Sites

Most field work for this dissertafion was conducted in Cutler Drain Canal (C-

1 OOC) and Canal (C-2) in the Biscayne Bay Drainage, Miami-Dade

County, Florida (Figures 1 and 2). Collections ofjuvenile largemouth bass and peacock

cichlids were made from the S-1 19 water control structure (25° 38.58' N, 80° 20.31 ' W)

to about 1 .6 km upstream of the structure in Cutler Drain Canal and from the Don Shula

Expressway bridge (downstream of the boat ramp on Snapper Creek Drive [25° 42.00' N,

80° 21 .25' W]) upstream up to, and including, the two lateral canals at the crossing of

SW 99"* Avenue in Snapper Creek Canal. These two canals were chosen because they are within the center of the range of peacock cichlids in southeast Florida, are relatively large canals, are accessible by boat, are in reasonably safe locations for personnel and equipment, and have different characteristic water chemistry and clarity.

Cutler Drain Canal (Figure 2A) in the Howard area is the smaller of the two canals and lacks large lateral canals in the study area. Under most conditions, water

clarity is relatively high, flow is low, and groundwater is the major water source. Cutler

Drain Canal is used as a main flow conduit only during infrequent high water events. I 19

I N Miami-Dade County A

Creek Canal Cutler Drain

Everglades National Park

25 km

Figure 1 . Location of select canals in Miami-Dade County, Florida. Snapper Creek Canal and Cutler Drain Canal were study canals. Inset map shows Florida. Map redrawn from Shafland et al. 2001. observed a substantial change in aquatic macrophyte species composition during the

study period. In 1999, hydrilla Hydrilla verticillata , fanwort Cabomba caroliniana, and 20

Figure 2. (A) Cutler Drain Canal (C-IOOC) and (B) Snapper Creek Canal (C-2), Miami- Dade County, Florida. Maps redrawn from Hidalgo 1997. 21

bacopa Bacopa sp. were the main species, with hydrilla being dominant. By 2001 , Asian hygrophila Hygrophila polysperma became the dominant macrophyte, with bacopa also being relatively common. Concurrently, numerous large grass carp were found in the

area, presumably stocked by the SFWMD for aquatic macrophyte control. Hydrilla is a highly selected food of grass carp (Shireman and Smith 1983), a possible factor in the

species replacement. Also, Cutler Drain Canal had low overall macrophyte density

during 1999 due to macrophyte removal. Qualitatively, abundance of both native and

exotic fish was relatively high (see Table 1 for a list of species).

Snapper Creek Canal (Figure 2B) is a major canal conduit in southeast Florida.

The study area encompassed a large, mainstem canal and two large lateral canals. Water

flow is frequently high in the main canal and water characteristics are most influenced by

surface water. Snapper Creek Canal receives most of its water fi-om the Tamiami Canal

system, a significant drainage from the Water Conservation Areas and, ultimately, Lake

Okeechobee from the north. The water is highly colored and occasionally turbid. Water clarity is usually greater in the lateral canals. Aquatic macrophytes are generally abundant in Snapper Creek Canal, especially in the laterals. Hydrilla dominated the main

canal and cabomba in the laterals. Significant amounts of muskgrass Chara sp., pond weed Potamogeton sp., and vallisneria Vallisneria americana also occurred. Snapper

Creek Canal has one of the most diverse fish faunas of the canal system (Table 1) and fish abundance was often high, especially in the lateral canals. 22

Table 1.- Fish species collected from two freshwater canals in southeast Florida in 1997- "*" species. "?" indicates a tentative identification. 2001 . A indicates a nonindigenous A

Cutler Drain Canal had 3 1 species (21 native and 1 0 nonindigenous). Snapper Creek

Canal had 40 species (22 native and 1 8 nonindigenous).

Scientific Name Common Name Occurrence

Cutler Drain Snapper Creek

Amiidae Amia calva Bowfin X

Anguillidae Anguilla rostrata American eel

Atherinidae Labidesthes sicculus Brook silverside X

Catostomidae Erimyzon sucetta Lake chubsucker

Centrarchidae gloriosus X X Lepomis gulosus Warmoutii X X Lepomis macrochirus Bluegill X X Lepomis microlophus Redear sunfish X X Lepomis punctatus Spotted sunfish X X Micropterus salmoides Largemouth bass X X

Centropomidae Centropomus undecimalis Common snook

Cichlidae ocellatus * X X Cichla ocellaris* Peacock cichlid X X bimaculatum* Black acara X X Cichlasoma citrinellum * Midas cichlid X X Cichlasoma managuense* Jaguar guapote X Cichlasoma urophthalmus * Mayan cichlid X Geophagus surinamensis * Redstriped eartheater X Hemichromis letoumeauxi * Jewel cichlid X X Heros severus * Green severum X * Oreochromis aureus Blue tilapia X * Oreochromis mossambicus Mozambique tilapia X? * Tilapia mariae X

Clariidae Clarias batrachus* Walking X

Clupeidae Dorosoma cepedianum Gizzard shad X 23

Table 1 . Continued.

Scientific Name Common Name Occurrence

Cutler Drain Snapper Creek

Cyprinidae Ctenopharyngodon idella * Grass carp X X Cvprinus carpio * Common carp (koi) X Notemigonus crysoleucas Golden shiner X X

Cyprinodontidae goodei Bluefin killifish X

Elopidae Elops saurus Ladyfish

Fundulidae Fundulus chrysotus Golden topminnow X

Ictaluridae Ameiurus natalis Yellow bullhead X X Ameiurus nebulosus Brown bullhead X X

Lepisosteidae Lepisosteus platyrhinchus Florida gar

Loricariidae Ancistrus sp.* X Hypostomus sp.* X Pterygoplichthys multiradiatus* X

Mugilidae Mugil cephalus Striped mullet

Percidae Etheostoma fiisiforme Swamp darter

Poeciliidae Gambusia holbrooki Eastern mosquitofish X X Heterandria formosa Least killifish X Poecilia latipinna Sailfin molly X X CHAPTER 3 A QUALITATIVE RISK ASSESSMENT OF PEACOCK CICHLIDS CICHLA OCELLARIS AND SOME IMPLICATIONS FOR THE METHODOLOGY

Introduction

Understanding and predicting the effects of introduced species are of interest for

conservation biology and community ecology. Moreover, resource managers and

regulatory authorities need such information. Unfortunately, prediction has proven

elusive because of uncertainty due to the complexity of interacting factors in ecological

systems (Baltz 1991 ; Ross 1991 ; FAO 1996). Furthermore, the outcome of some species

introductions has been both unpredicted and undesired (i.e., Frankenstein Effect; Moyle

etal. 1986).

There have been many methods used to predict the effect of species introductions.

Quantitative modeling methods (e.g., Kolar and Lodge 2002) can be used, but these are

data-intensive, require ecosystem-, taxon-, and stage-specific information, and have received little testing. As an alternative to quantitative models, qualitative methods also exist to estimate the effects of nonindigenous species (e.g., loop analysis; Levins 1974

[cited in Li et al. 2000]; Li and Moyle 1981; Li et al. 2000). Less formal methods include soliciting the opinions of experts who have extensive knowledge of the species and systems involved. Major drawbacks of this method are subjectivity and non-repeatability

(Ambrose et al. 1996). A slightly better approach might involve a more formal method to integrate expert opinions into a risk assessment. Indeed, the Generic Nonindigenous

24 25

Aquatic Organism Risk Analysis Review Process (i.e., RAM methodology) has been developed by the federal Aquatic Nuisance Species Task Force as a method to predict risks associated with newly established species, species proposed for intentional introduction, and individual introduction pathways (RAM Committee 1998). This methodology uses information on species life history and ecology, characteristics of its native range and the region of potential introduction, and vectors of introduction to assign qualitative probabilities that the focal species will colonize open waters, successfully reproduce and become established, and adversely affect native species or economic activity (RAM Committee 1998).

The RAM methodology has been recently used by the U.S. Fish and Wildlife

Service to synthesize information and to guide decision-making regarding proposals to

list certain fishes as Injurious Wildlife Species (e.g., black carp Mylopharyngodon

piceus ; Nico et al. 2001 ). There have been only three formal applications of the RAM methodology for fish and all have been assessments for specific organisms or groups of organisms (i.e., not for introduction pathways)—black carp (Nico et al. 2001), snakeheads (Family Channidae) (W. R. Courtenay, Jr. and J. D. Williams, USGS, in prep.), and Asian swamp eel Monopterus albus (L. G. Nico, USGS, in prep.). Another risk assessment using this methodology is planned to include bighead carp

Hypopthalmichthys nobilis and silver carp H. molitrix (J. D. Williams, USGS, pers. comm.). Of these risk assessments, only the black carp risk assessment was conducted prior to the occurrence of reproducing populations of the focal species (or one or more members of the family for snakeheads) in open waters within the USA. Furthermore, black carp was considered the test case for the use of the RAM methodology, partly 26

because there was still time to apply risk management strategies if warranted (Nico et al.

2001).

The black carp risk assessment tested the RAM methodology as a vehicle for risk

assessment (i.e., would the method work as envisioned and produce an estimate of risk

and certainty?). The black carp risk assessment demonstrated that the methodology could

achieve these goals. On the other hand, there has been no testing of the predictions of the

risk algorithm relative to any test case. In order to test these predictions, a number of

species or groups that are already established should be chosen to represent a variety of

nonindigenous species scenarios (e.g., various taxa, geographic distributions, and socio-

economic importance). Risk assessments could then be conducted and the results of the

risk assessments (i.e., risk predictions) could be compared to the observed outcomes (up until that time). This process could provide insight into biases or grossly incorrect predictions based on the risk algorithm and suggest improvements in the process. Indeed, changes in the risk assessment methodology were envisioned by the authors who acknowledged the existence of uncertainty in the risk assessment process (RAM

Committee 1998).

My objective is to begin this process of testing the RAM methodology using the peacock cichlid Cichla ocellaris (Cichlidae) as a test case. The emphasis of this assessment will be Florida, the only region of the continental USA with established

Cichla (see Cichla introductions and Limiting factors on Cichla distribution, below).

This species was chosen because 1 ) it has been introduced into the USA (i.e., Florida) as a sport fish and for biological control (i.e., has socio-economic importance), 2) it is predatory and ecomorphologically similar to native species (i.e., is of potential ecological 27

concern), 3) unlike the previously assessed fishes, it is limited to a rather narrow potential geographic range in the USA, 4) the species is not a new introduction (i.e., was

introduced into Florida in 1984), and 5) it has been the subject of fisheries research in

Florida and elsewhere. I will provide a brief overview of the biology, ecology, and history of introductions of Cichla to provide a background for the risk assessment

process. I will then describe the RAM methodology and assign estimates of risk and

certainty to the various rating elements. Lastly, I will discuss the predictions of the risk

assessment algorithm relative to the data on observed effects of Cichla in Florida and

suggest some modifications to the methodology.

Background Information

Cichla Taxonomy and Life History

The genus Cichla is composed of five valid nominal species (i.e., C. intermedia

Machado-Allison, C. monoculus Spix, C. ocellaris Schneider, C. orinocensis Humboldt, and C. temensis Humboldt) and perhaps another six undescribed species (KuUander and

Nijssen 1989). There has been considerable confusion over the alpha taxonomy of Cichla

(Lowe-McConnell 1969; Kullander 1986; Kullander and Nijssen 1989; Shafland 1995); the name C. ocellaris (i.e., peacock cichlid) has been applied to most Cichla that have been introduced outside of South America.

The peacock cichlid is a relatively large (up to about 650 mm total length [TL] and 5.4 kg) species native to Atlantic drainages of northeast South America (Kullander and Nijssen 1989). Cichla superficially resemble the North American black bass of the genus Micropterus (Norton and Brainerd 1993). 28

Peacock cichlids are predatory and adults are largely piscivorous (Lowe-

McConnell 1969; Zaret 1980a; Shafland 1995, 1999b; Jepsen et al. 1997; Winemiller et

al. 1997). Juveniles progress through ontogenetic dietary stages of feeding on zooplankton, insects, and decapod crustaceans prior to feeding on fishes (Zaret 1980a;

Chapter 4).

Peacock cichlids form pairs during spawning. Although Cichla probably only

spawn once per year in South American rivers (Jepsen et al. 1999), introduced populations in lakes, reservoirs, and canals spawn throughout the warmer months (Lowe-

McConnell 1969; Zaret 1980a; Jepsen et al. 1999; Shafland 1999b). Both parents guard a

nest (cleaned substrate of wood, rock, or sand) where up to 1 0,000 eggs are laid (Zaret

1980a). The parents further protect the free-swimming fry for up to eight to 10 weeks

(Zaret 1980). Fry occupy open water areas until they disperse from the guarded schools

into littoral vegetation (Schroder and Zaret 1979; Zaret 1980a). Sexual maturity often is attained within one year at lengths of about 240-290 mm TL (Fontenele 1950; Shafland

1995). Subaduhs and adults are habitat generalists (pers. obs.).

Cichla Introductions

Cichla are important food and sport fish and have been introduced into several tropical and subtropical regions (Zaret 1980a; Welcomme 1981, 1988; Lever 1996; Fuller et al. 1999). Cichla have been introduced into experimental ponds in Alabama (Swingle

1967), Florida (Ogilvie 1966), and Georgia (Ogilvie 1966) and into power plant cooling

reservoirs in (Shafland 1995; Fuller et al. 1999). These introductions did not result in established populations. Cichla have been intentionally stocked by fisheries agencies 29

and are established in open waters of Florida, Hawaii, and Puerto Rico (Erdman 1984;

Shafland 1995; Fuller et al. 1999).

In Hawaii and Puerto Rico, there were no native freshwater fish faunas to interact

with introduced Cichla . Peacock cichlids were introduced as sport fish into reservoirs

along with other nonindigenous species such as centrarchids ( Lepomis spp. and

largemouth bass Micropterus salmoides ) and cichlids (Tilapia and Oreochromis spp.)

(Erdman 1984; Devick 1980; Lilyestrom et al. 1994). Although the dynamics of other species were not reported, peacock cichlids and largemouth bass were not considered to

significantly compete in Hawaii (Devick 1980) or Puerto Rico (Lilyestrom et al. 1994;

Lilyestrom and Churchill 1996).

Peacock cichlids were intentionally introduced into freshwater canals in southeast

Florida in 1984-1987 (Shafland 1995). The primary reasons for introducing peacock

cichlids into Florida were 1 ) unbalanced prey-to-predator biomass ratios (sensu Swingle

1950) resulting from the high abundance of illegally introduced cichlids, particularly spotted filapia Tilapia mariae, and 2) to supplement recreational fisheries that were based

on largemouth bass (Shafland 1 995). Studies evaluating the effects of the peacock cichlid introducfion have not found evidence of negative effects on nafive fish

assemblages, including largemouth bass (Shafland 1999a, c; Chapter 6), or on fishery

values (Shafland and Stanford 1999). The Florida peacock cichlid fishery is esfimated at

US$ 8 million annually (Shafland and Stanford 1999).

On the other hand, the introducfion of Cichla into Lake Gatun, Panama, is one of the most frequently cited examples of the danger of fish introducfions. Cichla accidentally introduced into the Chagres River expanded downstream into the Panama 30

Canal reservoir in the 1960s (Zaret and Paine 1973). Once in the reservoir, Cichla populations expanded and this novel predator apparently decimated populations of small native fishes (Zaret and Paine 1973). The introduction also correlated with changes in zooplankton, insect, and bird populations (Zaret and Paine 1973). However, native fish populations may have rebounded since their initial reduction (Welcomme 1988).

Limiting Factors on Cichla Distribution

Cichla are fi-eshwater fishes with an upper lethal salinity of about 1 8 ppt

(Shafland and Hilton 1986). This factor would limit the ability of peacock cichlids to disperse via river mouths, estuaries, and salt marshes. Additionally, Cichla are true tropical fishes and cannot survive cold temperatures. The lower lethal temperature for

Cichla in laboratory studies was about 15° C (Swingle 1967; Guest et al. 1980). Indeed, low winter temperatures resulted in the elimination of reproducing populations of

purposefially introduced Cichla in some Texas power plant cooling reservoirs (Fuller et al. 1999). Experimental stocks of Cichla also have succumbed to cool temperatures in

Florida (Ogilvie 1966; Shafland 1995; pers. obs.). Salinity and temperature limit the distribution of Cichla in Florida to fi-eshwater canals and borrow lakes in the southeast portion of the state (Shafland 1995; see also Shafland and Pestrak 1982). Moreover, winter temperatures would make the establishment of Cichla in other portions of the continental USA unlikely in the absence of reliable thermal refuges. The possible limiting effects of water chemistry parameters (e.g., pH, hardness) have not been

investigated. However, C. ocellaris appears to be adaptable (i.e., survives and reproduces) to a wide range of water conditions (pers. obs.). Additionally, unlike some established nonindigenous fishes in Florida (e.g., walking catfish Clarias batrachus), 31

Cichla are not capable of breathing air or crossing land barriers. However, because they

are highly-favored sport fish, humans are likely to illegally transport Cichla to new waterbodies.

Ecological Interactions

Theoretically, peacock cichlids could have several ecological interactions with other species based on the dynamics of predation and competition. These interactions could have negative, positive, or mixed consequences for native species.

Cichla are predatory and can consume small native fishes, including juvenile largemouth bass (Hill 1998). Indeed, in Florida, fish comprise 92% by number of the diet

of peacock cichlids (Shafland 1 999b). The effect of predation by Cichla could be strong—this was implicated in the observed changes in native fish abundance in Lake

Gatun, Panama (Zaret and Paine 1973). Furthermore, there are many other examples of significant changes in native fish abundance following the introduction of piscivorous

fish (e.g., Kaufman 1992; Mills et al. 1994; see also Taylor et al. 1984). In addition to the direct effect of predation on native fishes, increased predation risk due to Cichla could force vulnerable prey (including juvenile largemouth bass) to refuge and potentially compete with each other for food within the limited refuge area (Mittelbach

1984, 1988; Osenberg et al. 1994). These direct and indirect effects of predation could reduce the abundance of largemouth bass and other native fishes.

Nevertheless, predation by peacock cichlids on abundant nonindigenous cichlids

(e.g., spotted tilapia), should have benefits for native fishes. Indeed, prey fishes can compete with early life stages of piscivores (e.g., largemouth bass) (Osenberg et al. 1 994;

Olson et al. 1995) and many prey fishes also are nest or larval fish predators (Bain and 32

Helfrich 1983; Popiel et al. 1996; Trexler et al. 2000). As a result, the introduction of a nonindigenous fish species that reduces the density of competitors or egg predators might

actually lead to increased population sizes of native fish. Interestingly, peacock cichlids were introduced, in part, to reduce numbers of small nonindigenous fishes (Shafland

1995). Thus this possible positive effect of peacock cichlids was an implicit part of the original management goal. Some correlative data support this outcome within the Florida canal system (Shafland 1995, 1999a). For example, the introduction of peacock cichlids into Canal was associated with a decline in the mean prey-to-predator biomass ratio fi-om 21.1 (1983-1988) to 10.0 (1988-1991) and a 36% decline in the biomass and 55% decline in numerical abundance of spotted tilapia (the most abundant exofic species) fi-om 1983-1988 compared to 1988-1993 (Shafland 1999a). Although

suggesfive, there are no data to help evaluate if these trends (i.e., decreased prey-to- predator biomass ratios and reduced abundance of nonindigenous spotted tilapia) were caused by peacock cichlids.

The ontogenetic dietary shifts of peacock cichlids could lead to competition with various life stages of native fishes for zooplankton, insects, decapod crustaceans, or fish

(Chapter 4). This mechanism could lead to reductions in native fish abundance following

the introduction of Cichla . For example, peacock cichlids and largemouth bass overlap in

diet (Shafland 1999b; Chapter 4; L. G. Nico, USGS, and J. E. Hill unpubl. data), which might indicate the potential for competition mediated through reductions in prey fish abundance, shifts in community composition, or changes in prey habitat use or behavior

(Soluk and Collins 1988; He and Kitchell 1990; Hambright et al. 1991; Soluk 1993; 33

Persson et al. 1 996). Nevertheless, the indirect effects of peacock cichhd predation on

native fishes have not been investigated.

Peacock cichhds also might facilitate largemouth bass (or other native predators)

by altering the behavior of their prey. For example, largemouth bass and peacock

cichlids have different hunting strategies and activity periods. Peacock cichlids often

rove in small schools and rely on speed rather than ambush as a hunting strategy (Erdman

1969; Devick 1972; Shafland 1995). Moreover, peacock cichlids are diurnal predators

with different primary activity periods than largemouth bass (Lowe-McConnell 1969;

Zaret 1980a; Lilyestrom and Churchill 1996). As a result of these differences,

individuals of prey species may be presented with conflicting behavioral choices

regarding activity patterns (e.g., foraging and refiiging), perhaps enhancing the hunting

success of one or both predators (Soluk and Collins 1988; Soluk 1993). This possibility

has not been studied for this system.

Peacock cichlids might enhance populations of native predators by serving as

prey themselves. For example, largemouth bass prey on juvenile peacock cichlids

(Shafland 1999b). If largemouth bass dynamics are limited by food for the larger,

piscivorous life stages, then the introduction of a new food source (i.e., peacock cichlids)

could benefit largemouth bass. Although Florida canals often have high prey biomass,

perhaps suggesting that food limitation is not very important, much of this potential food

is morphologically unavailable to largemouth bass. Gape size (i.e., throat width) limits

the maximum size of prey for largemouth bass (Lawrence 1 958; Hambright 1 99 1 ) and

many canal prey species rapidly grow out of vulnerable sizes (Shafland et al. 1985; Hill

1998; L. G. Nico, USGS, unpubl. data). By breeding throughout the warm months 34

(Shafland 1 999b), peacock cichlids provide numerous, appropriately-sized offspring for

largemouth bass predation throughout the year (Chapter 4; pers. obs.). Small peacock

cichlids are relatively abundant (Shafland 1999a, c; Chapter 4), are vulnerable (i.e., small, elongate, and with weak spines; Hill 1998) to a wide range of sizes of largemouth bass, and comprise 9% (by number) of prey consumed by largemouth bass in Tamiami Canal

(Shafland 1999b).

Spawning site competition via territorial aggression has been postulated as a

significant negative centrarchid-cichlid interaction (Hogg 1976; Taylor et al. 1984; Noble

and Germany 1986). However, evidence of this mechanism is largely anecdotal or correlative and no quantitative investigation has demonstrated spawning site competition for these fishes. Centrarchids would be the most likely native group to experience spawning site competition from peacock cichlids (or other cichlids) due to general

similarities in reproductive behavior (Heidinger 1975; Zaret 1980a; Annett et al. 1996).

However, in a study of the effect of blue tilapia Oreochromis aureus on largemouth bass

reproduction in experimental ponds, non-nesting (i.e., all female) blue tilapia had a similar effect on largemouth bass reproduction as did all male or mixed sex treatments

(i.e., nest-building and territorial) (Shafland and Pestrak 1983). This result suggests that spawning site competition is not the mechanism responsible for observed reductions of largemouth bass recruitment that occurred in some lakes with introduced blue tilapia (see

Taylor et al. 1984; Germany and Noble 1986). Additionally, there is relatively small overlap in spawning times of largemouth bass and peacock cichlids (Shafland 1999b; pers. obs.) and little agonistic behavior has been observed between the two species (P. L.

Shafland. FWC, pers. comm.; Chapter 5; pers. obs.). However, there is temporal, and 35

perhaps spatial, overlap in spawning activity between peacock cichlids and other

centrarchids (e.g., sunfish Lepomis spp.) (pers. obs.).

Of course, negative and positive effects of peacock cichlids could both exist.

Moreover, these effects may not be spatially or temporally consistent. Depending upon the direction and intensity of these effects, the outcome for native fishes could be mixed

(e.g., stage or context dependent), or even cancel. For example, peacock cichlids could enhance juvenile largemouth bass growth by reducing numbers of competing fishes while simultaneously decreasing adult largemouth bass growth by depleting numbers of

available prey. This type of effect is documented for bluegill Lepomis macrochirus and

largemouth bass in temperate lakes (Olson et al. 1995). Nevertheless, additional research

is required to document mixed effects in this system.

Fishing Effects

The influence of peacock cichlids on sport fishing could have mixed effects for the other major canal sport fish—largemouth bass. Combined recreational fishing values for largemouth bass and peacock cichlids (the most commonly targeted and valuable sport fish) in southeast Florida canals exceed US$ 13 million annually (Shafland and

Stanford 1999). Fishing tactics for both species are similar, although some important differences occur (Hidalgo 1997). These differences (e.g., fishing time of day and lure choice) are exploited by anglers specifically targeting one or the other species (Hidalgo

1 997). Given the high angler effort and increasing popularity of peacock cichlids in southeast Florida canals (Shafland and Stanford 1999), exploitation of largemouth bass could be increasing. This could result fi-om by-catch of anglers targeting peacock cichlids and by the promotion of canal fishing resulting from the attention that peacock 36

cichlids have generated. On the other hand, the introduction of peacock cichHds also

might reduce exploitation mortality of largemouth bass. Because of the differences in

fishing for the two species, peacock cichlid fishing may actually redirect angler effort

away from largemouth bass and therefore reduce largemouth bass total mortality. This

interesting question has not been investigated.

Qualitative Risk Assessment

The potential of nonindigenous species to be introduced, become established, and

become invasive can be qualitatively modeled given information concerning introduction

pathways and the biology and ecology of the organism in question (RAM Committee

1998). There are two main sections of the risk model: 1) Probability of Establishment

and 2) Consequences of Establishment. An overall Organism Risk Potential (ORP) is the

final output of the model (i.e., low, medium, or high) and is calculated based on the risk

estimates of both sections. Within these sections of the risk model there are seven rating

elements, each assigned a qualitative estimate of risk (i.e., low, medium, and high) and an

estimate of certainty (i.e., very certain, reasonably certain, moderately certain, reasonably

uncertain, and very uncertain). Four of the seven elements address the Probability of

Establishment and three elements attempt to predict the Consequences of Establishment.

Rating Elements of Risk Model—Probability of Establishment

1) Estimate probability of the exotic organism being on, with, or in the

pathway. The peacock cichlid is established in southeast Florida (Shafland 1995) and is thus in the Pathway. Various members of the genus Cichla are occasionally imported in the aquarium industry (pers. obs.). State and commonwealth 37

agencies have imported or exchanged peacock cichHds for fishery introductions

(Maciolek 1984; Shafland 1995; Fuller et al. 1999). High - Very Certain

2) Estimate probability of the organism surviving in transit. The peacock cichlid has survived transit on numerous occasions. High - Very Certain

3) Estimate probability of the organism successfully colonizing and maintaining a population where introduced. The peacock cichlid has been

reproducing in southeast Florida since the mid-1980s and is considered established in the

state (Shafland 1 995). Cichla have been intentionally introduced and established

reproducing populations in Hawaii and Puerto Rico in reservoirs (Fuller et al. 1999).

Nevertheless, peacock cichlids have been unsuccessful at establishing persistent populations in Texas power plant cooling reservoirs and possibly in southern Florida in

the 1960s due to cool temperatures (Fuller et al. 1999). A congener, the speckled

peacock cichlid C. temensis , was intentionally introduced alongside C. ocellaris yet failed to become established in southeast Florida (Shafland 1995). Because of thermal constraints, it is unlikely that Cichla could become established in the continental USA outside of southeastern Florida (see Limiting factors on Cichla distribution, above). High

- Very Certain

4) Estimate probability of the organism to spread beyond the colonized area.

The initial introduction area of the peacock cichlid nearly encompasses its present range in Florida. Peacock cichlids have successfully colonized adjacent canals and have been illegally or naturally transplanted into artificial lakes and ponds in southeast Florida.

5° Limifing factors for peacock cichlids include a lower lethal temperature of about 1 C

(Swingle 1967; Guest et al. 1980) and an upper salinity tolerance of about 18 ppt (Guest 38

et al. 1980; Shafland 1995). Relatively warm winters allow peacock cichlids to

temporarily colonize canals slightly north and west of the targeted area of introduction

(Shafland 1995; Loftus 2000; L. G. Nico, USGS, pers. comm.). Nevertheless, given the

inability of peacock cichlids to survive cool winter temperatures or salt water, there is

little possibility for significant expansion outside of the present range in southeast

Florida. Additionally, due to cool winter temperatures, it is unlikely that Cichla could become established in other areas of the continental USA. Low - Very Certain

Rating Elements of Risic Model—Consequences of Establishment

5) Estimate economic impact if established. The peacock cichlid could negatively affect the largemouth bass recreational fishery in southeast Florida via predation and competition for food (see Ecological interactions, above). Moreover, peacock cichlid fishing may have negafive effects on largemouth bass populafions (see

Fishing effects, above). Also, peacock cichlids theoretically could enhance largemouth bass populations via ecological and fisheries mechanisms (see Ecological interacfions and

Fishing effects, above). However, studies investigating the effects of peacock cichlids on largemouth bass have not reported significant reductions in largemouth bass abundance

(Shafland 1999a, c; see also Chapter 6). The largemouth bass fishery in the area occupied by peacock cichlids has been estimated at US$5 million annually (see Shafland and Stanford 1999).

Major objectives for introducing peacock cichlids into Florida were to create a novel sport fishery and to supplement recreational fisheries for largemouth bass

(Shafland 1995). These objectives have been met and the recreational fishery has been estimated at US$8 million annually (combined fishery of largemouth bass and peacock 39

cichlids of US$13 million) (Shafland and Stanford 1999). Therefore, the overall economic effect has been positive for the region of introduction. The economic risk is estimated to be low due to the sustained values of largemouth where peacock cichlids have been present since 1984-1987 (Shafland and Stanford 1999) and the relatively small potential effect on statewide fishery values (i.e., the small geographic area occupied—portions of three south Florida counties) regardless of any possible negative impacts on the largemouth bass fishery. Additionally, the similarity of

largemouth bass and peacock cichlid fishing (i.e., similar equipment and techniques;

Hidalgo 1997) suggests that largemouth bass anglers, guides, and tackle shops could adjust to any losses to the largemouth bass fishery by switching to peacock cichlids. Low

- Very Certain

6) Estimate environmental impact if established. There is a potential for peacock cichlids to adversely affect native fishes by predation and competition for food

(see Ecological interactions, above). Predatory fishes can exert top-down effects that may alter trophic relations and energy flow in aquatic systems (Zaret and Paine 1973;

Carpenter et al. 1987; Northcote 1988; Mittelbach et al. 1995). Nevertheless, it is also possible that peacock cichlids may positively affect native fishes by reducing numbers of other exotic fishes and by serving as prey themselves. Current data indicate the operation

of mechanisms by which peacock cichlids negatively effect native species (e.g., predation), but also suggest that peacock cichlids have not significantly reduced native

fish populations in southeast Florida (Shafland 1995, 1999a, c; Chapter 6). Furthermore, peacock cichlids are confined to artificial water bodies within a small geographic area that lacks highly specialized or endemic nafive fishes. Based on exisfing information, the 40

risk of environmental harm is relatively low, but risks may increase if conditions within

the system significantly change. One possible change that could increase the risk of

environmental harm from peacock cichlids is climatic change. An increase in mean low

temperatures in southern Florida might lead to an expansion of the range of peacock

cichlids. Nevertheless, peacock cichlids are less tolerant of cool temperatures than are

many established tropical fishes in Florida (Shafland and Pestrak 1982; Shafland 1995)

and experience winterkills even within the present thermal refuge of the canal system, (P.

L. Shafland, FWC, pers. comm.), a region with mean monthly low air temperatures in

January (Florida's coldest month) of > 1 8.3° C (Shafland and Pestrak 1 982). Low-

Moderately Certain

7) Estimate impact from social and/or political influences. Peacock cichlids

have had fairly broad approval and support from within the angling community before

and since the introduction in the mid-1980s. The recreational fishery is popular and

supports important socio-economic activity (Shafland and Stanford 1999). A few

anglers, environmental groups, and some scientists, particularly ecologists, express

concern over the peacock cichlid. Anglers against the introduction are concerned with

possible effects on the largemouth bass fishery. Environmental groups and many

ecologists are against the introduction of any non-native species, their concerns generally

based on the theoretical or potential effects of nonindigenous species on native species

(e.g., Courtenay and Robins 1989; Minns and Cooley 2000). An additional concem for

environmental groups and governmental agencies is the possibility of peacock cichlids invading the nearby Everglades National Park. Given the widespread acceptance and 41

support of peacock cichlids in south Florida, the overall effect is estimated to be positive.

Low - Moderately Certain

Organism Risk Potential (ORP)

The Organism Risk Potential (ORP) has two components: 1) the Probability of

Establishment and 2) the Consequences of Establishment. The Probability of

Establishment is the lowest ranking amongst the first four ratings elements of the risk

model (i.e., within pathway, entry potential, colonization potential, and spread potential).

Given the fact that peacock cichlids are established in Florida, intuitively this probability

should be rated as "High". Nevertheless, because the potential for spread is rated as

"Low", the overall Probability of Establishment from the model is rated as "Low" (Figure

3). The fourth rating element (i.e., spread potential) was estimated as "Low" because the

area of introduction (i.e., the area of canals where peacock cichlids were stocked), the

intended target area of the introduction, and the potential maximum range of colonization

for peacock cichlids in the continental USA (i.e., Florida) are very similar (i.e., about 650

km of canals; Shafland and Stanford 1 999; see Limiting factors on Cichla distribution,

above).

The Consequences of Establishment rating is defined as the highest of the probabilities assigned to the economic and environmental impacts. Perceived impact

(i.e., social and polifical) does not influence the assessment of Consequences of

Establishment unless the economic and environmental risk estimates are both "Low".

All three elements were estimated to be "Low" for the potential for negative consequences. Therefore, the Consequences of Establishment are estimated to be "Low"

(Figure 3). 42

Probability Organism Entry Colonization Spread of within Potential Potential Potential = Low Establishment Pathway (High) (High) (Low) (High)

Consequences of Economic Environmental Perceived = Low Establishment (Low) (Low) (Low)

Probability of Consequences of ORP Risk Establishment Establishment = Low (Low) (Low)

Figure 3. Organism Risk Potential (ORP) for peacock cichlid Cichla ocellaris within the continental USA based on the RAM methodology (RAM Committee 1998), given a

"Low" Probability of Establishment. The Probability of Establishment is the lowest

rating among the four rating elements (i.e., within pathway, entry potential, colonization

potential, and spread potential). The Consequences of Establishment is the highest rating

between economic and environmental ratings. The ORP is the average of Probability of

Establishment and Consequences of Establishment ratings. An ORP of "Low" is defined as "acceptable risk - organism(s) of little concern (does not justify mitigation)" (RAM Committee 1998).

The ORP rating is the average of the two risk potentials assigned for the

Probability of Establishment and the Consequences of the Establishment, with the

average rounded up (i.e., a conservative estimation). Therefore the ORP for peacock cichlids introduced into the continental USA is "Low" (i.e., "Low" Probability of

Establishment and Consequences of Establishment) (Figure 3). Conversely, had the risk of spread rating element been disregarded, then the Probability of Establishment would

have been assigned a rating of "High" and the ORP would have been "Medium" (i.e., —

43

average of "High" Probability of Establishment and "Low" Probability of

Consequences). An ORP of "Medium" is considered unacceptable risk and would

require mitigation (RAM Committee 1998).

Discussion

Risk assessments attempt to predict effects of nonindigenous species and are used

as a basis for economic, environmental, and policy decisions. As a result, it is important

to evaluate the methods used to conduct risk assessments to determine if risk assessments

yield accurate predictions and thus are useful tools.

Although the peacock cichlid is well-established in southeast Florida (Shafland

1995), the RAM method algorithm resulted in an estimate of the Probability of

Establishment as "Low". This somewhat nonsensical outcome of the model (i.e., rating

the probability that an established species would become established as low) was due to

the similarity in the geographic area of the initial introduction and the small potential

maximum range in the continental USA (i.e., there is little chance of range expansion).

Previous risk assessments using this methodology have been conducted on fishes with

large potential ranges in the USA (i.e., black carp, Asian swamp eel, and snakeheads).

Therefore, the situation with peacock cichlids may represent a relatively unusual case

one not adequately anticipated in the methodology. Perhaps any apparent conflict over the assignment of a "Low" risk of establishment is semantic. The Probability of

Establishment label could be modified to Probability of Establishment and Spread to avoid confusion in such cases (see also Kolar and Lodge 2002). Regardless of the label, the essence of this risk element (i.e., Probability of Establishment) is the potential for an organism to access the region of interest, and then survive, reproduce, and spread once 44

there. Additionally, this issue suggests that each of rating elements should be considered

separately as well as in combination, and that any method of risk assessment requires

some level of flexibility in interpretation, factors acknowledged by the authors of the

methodology (RAM Committee 1998).

The qualitative risk assessment methodology combines estimates of risk for two

very different processes into a single estimate of the risks posed by a nonindigenous

species (i.e., establishment and consequences). The assignment of the average risk rating

between these two processes as the overall risk potential that this organism represents

(i.e., ORP) can be misleading. An organism may have a low probability of establishment

but a high potential for adverse effects and have the same Medium ORP rating as a

species with a high potential for establishment but a very low probability of negative

consequences. A Medium ORP is defined as an unacceptable risk, justifying mitigation

(RAM Committee 1998). A logical argument would be to place more emphasis on the

potentially harmfial species even though there is a relatively low chance of becoming

established because of the greater danger posed by a successful invasion. Indeed, it

might be a waste of governmental resources and an undue economic burden to attempt to

exclude species that have low potential for environmental or economic harm regardless of

their perceived ability to colonize. It may be more appropriate to consider the two processes independently and do away with or otherwise de-emphasize the ORP because

the ORP rating seems to lose information rather than simply condense it.

Two alternative ORP estimates for peacock cichlids (i.e., "Low" [Figure 1] and

"Medium") were presented in this risk assessment. The estimate of "Low" was based on a strict interpretation of the RAM method algorithm. The estimate of "Medium" risk 45

logically could be argued based on the fact that peacock cichlids are established in the

continental USA (i.e., should have a "High" Probability of Establishment). An ORP result of "Medium" risk would place the peacock cichlid into the category of

unacceptable risk (RAM Committee 1 998). However, it seems illogical to require mitigation against a socio-economically important sport fish when there is no evidence

that it has harmed native fish species in the nearly two decades since it has been

intentionally introduced (Shafland 1995, 1999a,c; Shafland et al. 2001; Chapter 6).

The RAM algorithm can only return a "Medium" or "High" ORP if a

nonindigenous species is established or has a medium or higher probability of becoming

established (RAM Committee 1 998). This presupposes that if established, any nonindigenous organism will have an unacceptable risk of negative impacts. This inherent bias prejudices a supposedly objective process to lead to exclusion or restriction of any species that may become established. If this strict precautionary principle (FAO

1 995) is the foundation for the RAM methodology, then the entire assessment of risk hinges on the estimate of the ability of the organism to become established. The section concerning the possible consequences to the environment or economy become of no importance, except perhaps in defining regulatory, eradication, or control priorities. For example, the upcoming risk assessment of bighead carp and silver carp must conclude

that these species represent an unacceptable risk and require mitigation (i.e., the ORP must be either "Medium" or "High") due to the fact that both already are established in the Mississippi River (and could spread to other regions), regardless of the assessor's assignment of risk associated with the presence of these species. Therefore, the outcome of the risk assessment for these species is largely predetermined. 46

Regardless of any inherent bias, the predictive abihty of any risk assessment tool

can only be as good as the information put into the model. Although it provides a

meaningful way of approaching the problem and attempts to interject much-needed

objectivity into the field, the method still leaves substantial room for personal

interpretation and subjectivity (i.e., "uncertainty of the assessor"; RAM Committee

1998). For example, everyone will not agree with my estimates of risk and certainty for

peacock cichlids or other species. Some disagreements will be over interpretation of

data, others over philosophy. In particular, some ecologists would rate the environmental

risk of peacock cichlids as high due to demonstrated predation and the potential for

competition affecting native species regardless of existing data. Furthermore, data are

lacking for many non-fish components of southeast Florida's aquatic systems, increasing

uncertainty. Although empirical rules of bio-invasion have been proposed that seek to

provide a framework for prediction based on past invasions (Moyle and Light 1996a),

their universality seems doubtful and the performance of many nonindigenous species

has been unpredicted (Moyle et al. 1986). Indeed, in any risk assessment process, a

major impediment is uncertainty associated with the organisms (RAM Committee 1 998).

In the case of peacock cichlids, in contrast to other fishes assessed by this methodology, there are fairiy substantial data on introduced populations within the region of concern

(i.e., continental USA) (Shafland 1995, 1999a, b, c; Hill 1998; Shafland and Stanford

1999; Shafland et al. 2001; Chapter 4, 5, and 6), thereby increasing the estimates of certainty (RAM Committee 1998).

In the RAM methodology guide, it is argued that the actual ratings assigned to individual elements are not as important as the transparent presentation of information 47

upon which the ratings are based (RAM Committee 1 998). AUhough this is an important

point, the fact that the algorithm produces a single term, the ORP, as a guide to decision

making (i.e., acceptability of risk and requirement of mitigation) places a distinct

emphasis on the ORP as the end product of the risk assessment. This factor, coupled with

the previously mentioned bias relative to any species capable of becoming established

(i.e., a de facto assumption that all established nonindigenous species represent an

unacceptable risk), somewhat limit the utility of the process. Indeed, it could be argued

that this process merely adds a pseudo-quantitative shroud to the subjective method of

expert opinion. On the other hand, the process does provide a vehicle for gathering

information on the focal species and forces a qualitative evaluation of various categories

of risk.

In conclusion, it is clear that the peacock cichlid test case pointed out problems

associated with the RAM methodology. Established species with small potential ranges

yield a nonsensical result for the Probability of Establishment. This could be largely

addressed by changing the label to Probability of Establishment and Spread. The

combining of probabilities of a species becoming established with subsequent

consequences of the establishment into an ORP output places too much emphasis on a

single value and loses important information in the process. This could be addressed by

reducing the prominence of the ORP (or eliminating it altogether) and retaining the

Probability of Establishment and Spread along with the Consequences of Establishment as the primary outputs of the RAM model. Regulators and managers should have clear objectives in mind prior to commissioning a risk assessment and should determine if the threat of successful introduction alone is the primary concern or if negative effects of the 48

nonindigenous species are more (or equally) important. Additionally, the built-in

assumption of the RAM model that all nonindigenous species capable of establishing persistent populations represent an unacceptable risk should be acknowledged. The

RAM guide states that a risk assessment cannot be used to determine an acceptable level

of risk, yet the ORP ratings are defined explicitly in terms of acceptable and unacceptable risk (RAM Committee 1998). The level of acceptable risk is not a question for the

assessor, but one for society, and definitions should be changed to remove conclusions of

acceptability. Also, conclusions of the need for mitigation are not needed, although

suggestions for mitigation (or eradication or control) may be appropriate within the risk

assessment. In essence, the definitions of "Low", "Medium", and "High" risk should be

self-explanatory and interpreted within the context of the stakeholder environment.

Lastly, additional test cases should be assessed to determine if other biases occur in the methodology. CHAPTER 4 ONTOGENETIC DIETARY SHIFTS AND DIETARY OVERLAP OF A NONINDIGENOUS FISH AND AN ECOMORPHOLOGICALLY SIMILAR NATIVE FISH

Introduction

Changes in resource use during ontogeny have considerable influence on the dynamics of size-structured populations (Werner and Gilliam 1984). For piscivorous fishes, ontogenetic diet shifts can be dramatic. For example, the largemouth bass

Micropterus salmoides typically progresses through relatively discrete stages of feeding on zooplankton, insects, and decapod crustaceans before becoming largely piscivorous

(Keast and Eadie 1985; Olson 1996; but see Schramm and Maceina 1986; Cailteux et al.

1996; Shafland 1999b; Huskey and Tumigan 2001 for studies reporting continued and substantial consumption of decapod crustaceans in addition to fish). These dietary shifts are size-related and dependent upon the relative size of the predator's gape and the prey's body depth or girth (Hambright 1991). For gape-limited species, growth-related

increases in gape size allow for capture of larger, more energetically profitable prey (e.g., inifiation of piscivory; Pasch 1975; Ludsin and DeVries 1997). Indeed, individual

growth is enhanced by switching to a fish diet (Pasch 1975; Timmons et al. 1980; Olson

1996). Moreover, increased growth also leads to increased survival of age-0 largemouth bass because mortality is strongly size-dependent (Gutreuter and Anderson 1985; Ludsin and DeVries 1997).

49 50

Trophic morphology, including gape width, is considered to be an indicator of a fish's ability to handle prey (Wainwright and Richard 1995; Wainwright 1996). Indeed,

numerous studies have investigated gape width and its relationship to prey use in largemouth bass (e.g., Lawrence 1958; Hambright 1991; Hambright et al. 1991). Given this link between trophic morphology and prey use, species that are morphologically similar might be expected to exhibit similar ontogenetic dietary shifts and have

considerable overlap in diet. An ecological implication of this is that if prey are limiting and are depleted by predators, then competition will occur between species that

significantly overlap in diet.

The peacock cichlid Cichla ocellaris (Cichlidae) was intentionally introduced into the native range of largemouth bass (Centrarchidae) in southeast Florida in 1984-1987

(Shafland 1995). Norton and Brainerd (1993) described these species as ecomorphologically similar based on morphology, diet, and feeding mechanics.

Moreover, Hill (1 998) found broad similarity in length-specific gape width between these species. Although the ontogenetic dietary patterns of largemouth bass diet are well known (Keast and Eadie 1985; Olson 1996), little information is published on the diet of juvenile Cichla (Lowe-McConnell 1969; Zaret 1980a; Lilyestrom and Churchill 1996;

Shafland 1999b). Nevertheless, available information and the similarity in morphology of these species imply that largemouth bass and peacock cichlids may have similar diets through ontogeny. Moreover, preliminary observations of canals in southeast Florida suggested that juvenile peacock cichlids were abundant and used similar habitats as juvenile largemouth bass (pers. obs.). 51

Studies documenting ontogenetic dietary stages or dietary overlap of naturally co- occurring fishes are common (e.g., Nico and Taphom 1988; Winemiller 1989, 1991;

Murie 1995; Scott and Angermeier 1998). Additionally, much research has investigated shared resource use of introduced and native fishes, including species that are

morphologically similar (e.g., Mathur 1977; Parrish and Margraf 1990; Matthews et al.

1992; Huckins 1997; Huckins et al. 2000; Declerck et al. 2002; Sutton and Ney 2002;

Stoffels and Humphries 2003). However, studies are rare that compare dietary ontogeny and estimate the overlap in diet of a nonindigenous fish with an ecomorphologically similar, but phylogenetically distant, native fish.

This study investigated the hypothesis that an introduced fish will have an ontogenetic dietary progression similar to that of a native species of similar morphology, regardless of their phylogenetic relations. Additionally, this study documented the temporal co-occurrence ofjuvenile largemouth bass and peacock cichlids of similar size and ability to handle prey, and estimated the extent to which the two species overlapped in prey use in two southeast Florida canals.

Methods

Collections for this study were made in Cutler Drain Canal (C-IOOC) and Snapper

Creek Canal (C-2), Biscayne Bay Drainage, Miami-Dade County, Florida. Both canals occur in urban and suburban areas of the Greater Miami Metropolitan Area. Cutler Drain

Canal averages about 25 m wide and 3^ m deep. It had highly variable aquatic macrophyte species composition and abundance due to biological and mechanical weed control methods used by the South Florida Water Management District (SFWMD). In sampled reaches, aquatic macrophyte coverage ranged from about 2% percent areal 52

coverage (PAC) to > 75% PAC. Asian hygrophila Hygrophila polysperma was the dominant submersed macrophyte. Water clarity was generally > 3 m of Secchi depth.

Collections were made from the S-1 19 water control structure (25° 38.58' N, 80° 20.31'

W) upstream for about 1 .6 km. Snapper Creek Canal is a large, important conduit for water drainage. Collections were made from the Don Shula Expressway bridge

(downstream of the boat ramp on Snapper Creek Drive [25° 42.00' N, 80° 21.25' W]), upstream along the main canal and in two adjacent lateral canals accessed near the

crossing of SW 99 Avenue—a total of about 1 .8 km. In sampled reaches, aquatic macrophyte abundance ranged from < 5% PAC to > 50% PAC in the main canal and from < 20%) PAC to near 100%) PAC in the laterals. The main canal had highly variable aquatic macrophyte abundance, with hydrilla Hydrilla verticillata being the dominant species. The lateral canals had relatively little mechanical harvesting of weeds by the

SFWMD, and therefore had relatively stable aquatic macrophyte abundance. Hydrilla and fanwort Cabomba caroliniana were the dominant species in the lateral canals.

Muskgrass Chara sp., a macroalgae, also was common in the southern lateral canal.

Snapper Creek Canal had highly colored and occasionally turbid water (Secchi depth usually 1-2 m).

Four collections were made in Snapper Creek Canal in 1 999 (July, September,

October, and November) and two collections were made in both canals in 2001 (August and October). Juvenile largemouth bass and peacock cichlids were collected by daytime boat electrofishing. An attempt was made to collect all individuals < 150 mm TL until target numbers were reached, 25 (in 1999) or 50 (in 2001). Fish were euthanized and 53

preserved in formalin. Formalin was injected into the abdominal cavity of specimens

collected in 2001.

In the laboratory, specimens were transferred to ethanol. All fish were measured

for maximum total length (TL), and a representative sample covering the collected size

range of largemouth bass and peacock cichlids was measured for gape width (GW).

Gape width measurements were made as external mouth width with the mouth closed to

approximate intercleithral distance (Lawrence 1958; Hambright 1991; Hill 1998).

Wainwright (1996) determined that intercleithral distance is the limiting dimension for

swallowing prey in centrarchids. External mouth width has an approximate 1:1

relationship to intercleithral distance in largemouth bass (Lawrence 1958). Additionally,

Hill (1998) evaluated this method for peacock cichlids and found that peacock cichlid

prey maxed out at about 96.5% of predicted gape size (maximum sizes ranged from 88-

104% with a SD of 5.6%).

Visceral masses of all fish were removed, stomachs were opened with forceps,

and prey items were removed, identified, and counted with the aid of a dissecting

microscope. Only items occurring in the stomach portion of gastro-intestinal tract were

examined and used for diet analyses. All prey measurements and predator GW

measurements were made with dial calipers. Predator total-length measurements were

made with a measuring board. Fish prey were measured for maximum standard length

(SL). Decapod crustaceans in the fish stomachs were found with abdomens folded

underneath the cephalothorax and were therefore measured for folded body length (FBL;

i.e., dp of rostrum to outside of bend in abdomen). Maximum body depth (BD) was measured for each prey if the specimen was intact. Decapod crustacean BD was 54

measured with the abdomen folded underneath the cephalothorax. An estimate of the volume of prey in the stomachs was obtained by water displacement using either a graduated cylinder or a pipette. A qualitative estimate of stomach fullness also was

recorded. The values ranged from 1 to 4, with 1 being empty, 2 having some food but less than 25% of the estimated stomach capacity, 3 having 25-75% of stomach capacity filled, and 4 having > 75% of estimated stomach capacity.

Juvenile largemouth bass and peacock cichlids were grouped into 10-mm TL size classes. Estimates of length-specific gape width were used to construct histograms of gape size distribution for each sampling date to determine the extent of overlap in morphological ability ofjuvenile largemouth bass and peacock cichlids to handle prey.

Estimates of gape width were obtained by regressing gape width measurements on total length (PROC GLM; SAS Institute, Gary, ). Analysis of covariance

(ANCOVA) was used to test for differences in slopes and intercepts of developed regressions for largemouth bass and peacock cichlids (PROC GLM; SAS Institute, Gary,

North Carolina). All statistical analyses were conducted at a significance level of a =

0.05 unless otherwise noted.

The stomach contents ofjuvenile largemouth bass and peacock cichlids were

described by frequency of occurrence, numerical abundance, and volume for all prey, as well as prey size for fish and decapod crustacean prey. A predator was defined as a piscivore when fish prey were found in at least 60% of predator stomachs that contained

any food (Bettoli et al. 1992; Gailteux et al. 1996). Prey sizes (i.e., prey BD and prey length offish and decapod crustacean prey) were regressed on predator total length and the resulting relations were compared with ANCOVA (PROG GLM; SAS Institute, Gary, 55

North Carolina). Also, relative prey size was estimated in two ways—prey body depth

divided by predator gape width, and prey length (SL for fishes and FBL for decapod

crustaceans) divided by predator total length, multiplied by 100 to obtain percentages.

Estimates of stomach fullness were compared between species within sampling

dates by the Wilcoxon rank sum test (PROC NPARIWAY; SAS Institute, Gary, North

Carolina) and within species and among sampling dates by the Kruskal-Wallace test

(PROC NPARIWAY; SAS Institute, Gary, North Carolina). A significant Kruskal-

Wallace test was followed by Dunn's non-parametric multiple comparison test

(Hollander and Wolfe 1973). Due to the conservative nature of Dunn's test, a significance level of a = 0.10 was used to distinguish mean rank differences among sampling dates (Hollander and Wolfe 1973).

Dietary overlap values between predators within sampling date were estimated using Schoener's Index (Schoener 1970):

Cxy= 1 -0.5(X \Pxi-Pyi\),

where C;^^ is the index value, p^i is the proportion of food / in the stomach contents of

species x, and p^, is the proportion of food i in the stomach contents of species y. This

index ranges from 0 (no overlap) to 1 (complete overlap). 1 defined proportions using either numerical abundances or prey volumes. Although subjective, values of dietary overlap > 0.6 have been considered as significant dietary overlap (Zaret and Rand 1971;

Mathur 1 977). Schoener's index of dietary overlap has been recommended for use when resource availability data are lacking (Hurlbert 1978; Wallace 1981). 56

Results

A total of 548 fish were examined for stomach contents (247 largemouth bass and

301 peacock cichlids) (Figure 4). Of these specimens, 216 largemouth bass and 200

peacock cichlids had food in their stomachs. The results for stomach contents include

only those individuals with food in the stomach.

Largemouth Bass

Peacock Cichlid

^ <^ \^

Total Length Group (mm)

Figure 4. Length-frequency for juvenile largemouth bass (N = 247) and peacock cichlids (N = 301) electrofished from Snapper Creek Canal (1999 and 2001) and Cutler Drain

Canal (2001 ), Miami-Dade, County, Florida.

Prey items were sorted into eight prey categories—zooplankton, amphipods,

insects, decapod crustaceans, fish, larval fish/eggs, anurans, and organic material. The

zooplankton in the stomachs consisted mostly of calanoid copepods and cladocerans.

The insects were mostly mayfly larvae (Ephemeroptera), chironomids (Diptera), and water boatmen (Corixidae; Hemiptera). However, odonate larvae (Odonata), water 57

striders (Gerridae; Hemiptera), and some terrestrial insects were present. A few

terrestrial insects were found in largemouth bass stomachs, but most were aquatic insects.

No terrestrial insects were found in peacock cichlid stomachs. All decapod crustaceans

found in peacock cichlid stomachs and most in largemouth bass stomachs were grass

shrimp Palaemonetes sp. Crayfish Procambarus sp. also were found in largemouth bass

stomachs. Swamp darter Etheo stoma fusiforme and, to a lesser extent, bluefm killifish

Lucania goodei , were the dominant fish prey; poeciliids (i.e., Gambusia holbrooki and

Poecilia latipinna), bullhead catfish Ameiurus sp., loricariid catfish, cichlids (including

peacock cichlids), and centrarchids also were consumed. Larval fish in the stomachs

were yolk-sac fry and were not identified to species. Organic materials included well-

digested , detritus, and plant material.

Ontogenetic Dietary Shifts

For largemouth bass, insects, and to a lesser extent, zooplankton and amphipods

dominated frequency of occurrence for small individuals (i.e., < 70 mm TL) (Figure 5A).

Zooplankton became less important and dropped out of the diet after the 80 mm TL size

class. The longest largemouth bass with zooplankton was 86 mm TL. Amphipods

declined in importance and were not found in fish larger than 130 mm TL. Insects

remained frequent in the diet over the range of largemouth bass length examined in this

study (< 1 70 mm TL), but were replaced by fish as the most frequently occurring prey

category in specimens of 120 mm TL and larger. Decapod crustaceans occurred in only

13.7% of largemouth bass < 90 mm TL, but were commonly consumed (in 40.2% of

stomachs) by largemouth bass > 90 mm TL. The smallest largemouth bass that consumed a fish was 67 mm TL. An increasing fraction of largemouth bass consumed 58

Peacock Cichlid Total Length Group (mm)

Zooplankton —o— Amphipod - A - Insect — — Decapod » Rsh

Figure 5. Frequency of occurrence (i.e., percent of predators with prey category in stomach contents) of main prey categories in juvenile (A) largemouth bass (N = 216) and (B) peacock cichlids (N = 200) electrofished from Snapper Creek Canal (1999 and 2001) and Cutler Drain Canal (2001), Miami-Dade County, Florida. The horizontal line at

60 % corresponds to the frequency of fish in the diet recommended by Bettoli et al. (1992) as the minimum frequency for categorizing a predator as piscivorous. 59

fish as largemouth bass size increased. Nevertheless, the frequency of occurrence offish

did not exceed 60% (i.e., piscivory as defined by Bettoli et al. 1992) until the 120 mm TL group.

Peacock cichlids exhibited a similar pattern of diet change, although there were important differences, particularly regarding the extent of piscivory (Figure 5B). The

smallest size class of peacock cichlids (i.e., 30-50 mm TL group) ate zooplankton at a high frequency. Zooplankton was frequent or moderately frequent in the diets of peacock cichlids up to the 100 mm TL group. The largest peacock cichlid with zooplankton in its stomach was 105 mm TL. Insects were eaten by a high frequency of peacock cichlids in the 60-80 mm TL groups, but rapidly declined in percent occurrence and dropped out of

the diet by the 1 1 0 mm TL group. Decapod crustaceans were moderately frequent in the stomachs of peacock cichlids > 70 mm TL, and of high frequency in fish > 140 mm TL.

The smallest peacock cichlid with a fish in its stomach was 54 mm TL. Fish were found in a low to moderate percent of peacock cichlid stomachs in the 30-80 mm TL groups, but rapidly increased in frequency with increasing peacock cichlid size. Piscivory (sensu

Bettoli et al. 1992) was attained by peacock cichlids > 100 mm TL.

By number, diets of small largemouth bass (i.e., < 70 mm TL) were dominated by zooplankton (Figure 6A). Additionally, zooplankton made up about half of the number of prey for largemouth bass in the 70 mm TL group. Amphipods made up only a small percentage by number of the prey of largemouth bass. Insect numbers were important in the 70 mm TL group and dominated the 80-1 30 mm TL groups. Moderate numbers of insects also were eaten by the largest size classes of largemouth bass. Decapod crustaceans were of relatively minor importance by number for largemouth bass > 90 mm 60

TL. Fish were of low frequency by number in fish > 90 mm TL except for the 140+ mm

TL groups where fish were dominant.

A 100%^

80%' I E 3 60%- z SS5 >.

•4-1 40%- c i 20%' 0) Q. 0%.

Largemouth Bass Total Length Group (mm)

Peacock Cichlid Total Length Group (mm)

Zooplankton SAmphipod m Insect Q Decapod Fish

Figure 6. Percent contribution to diet by number of main prey categories in juvenile (A) largemouth bass (N = 216) and (B) peacock cichlids (N = 200) electrofished from Snapper Creek Canal (1999 and 2001) and Cutler Drain Canal (2001), Miami-Dade County, Florida. 61

Zooplankton was an important prey category by number for peacock cichlids up

through the 1 00 mm TL size class (Figure 6B). Amphipods were relatively unimportant

by number for peacock cichlids. Insects contributed most to prey numbers for peacock

cichlids in the 80 mm TL group, but were relatively important by number for fish of 60-

110 mm TL. Decapod crustaceans were of low to moderate numbers in peacock cichlids

of 70-140 mm TL, but were about 35% of the prey numbers for fish in the 140+ mm TL

group. Fish were important by number for peacock cichlids beginning at the 1 00 mm TL

group and were dominant for peacock cichlids > 1 1 0 mm TL.

Zooplankton and amphipods contributed relatively little to the volume of stomach

contents of largemouth bass in this study, except for the 30-50 mm TL group (Figure

7A). Insects made up over half of the volume of this size class (i.e., 30-50 mm TL

group) and generally made up at least moderate amounts of the prey volume for

largemouth bass in the 60-100 mm TL groups and again in the 130 mm TL group. The

decapod crustacean category was an important constituent of prey volume for largemouth

bass in the 70-130 mm TL groups. Fish prey made up a small to moderate amount of the

prey volume for largemouth bass in the 60-100 mm TL groups but became dominant in

all larger size classes.

Zooplankton made up a substantial percent of the prey volume only for peacock

cichlids 30-50 mm TL (Figure 7B). Insects contributed a relatively small amount to prey

volume for peacock cichlids in the 30-50, 70, and 80 mm TL groups, and made up the

largest amount of prey volume in the 60 mm TL group. Decapod crustaceans were dominant by volume in the 70 mm TL group but were of moderate to low volume in 62

larger peacock cichlids. Fish were important volumetrically for peacock cichlids 30-80 mm TL, and were dominant for larger peacock cichlids.

Largemouth Bass Total Length Group (mm)

Peacock Cichlid Total Length Group (mm)

Zooplankton 0 Other SAmphipod Ulnsect 0 Decapod Fish

Figure 7. Volume contribution to diet of prey categories in juvenile (A) largemouth bass (N = 216) and (B) peacock cichlids (N - 200) electrofished from Snapper Creek Canal (1999 and 2001) and Cutler Drain Canal (2001), Miami-Dade County, Florida. The "Other" category includes larval fish/fish eggs, organic materials, and anurans. ,

63

The Scaling of Gape Width with Fish Length

The relations between gape width and total length for both predators were:

- = - Largemouth bass GW 0. 11 3 TL 2.2 1

r^ = 0.965, N = 87, TL range = 58-175 mm; and

Peacock cichlid - GW = 0.0997 TL - 1 .570,

^ = 0.968, N = 1 10, TL range = 39-165 mm (Figure 8). The ANCOVA revealed that the regression slopes were different (p < 0.0001), with largemouth bass having a slightly larger length-specific gape width. Nevertheless, there was substantial overlap of gape

measurements, especially for fish below 1 00 mm TL (Figure 8). The distribution of gape sizes for both predators also overlapped considerably during the sampling period (Figure

Largemouth bass Peacock cichlid

14- E E 12"

6"

0 0 50 100 150 200 Total Length (mm)

Figure 8. Length-specific gape width measurements from juvenile largemouth bass (N = 87) and peacock cichlids (N = 110) electrofished in southeast Florida canals. 64

July 1999 - Snapper Creek Canal August 2001 - Snapper Creek Canal

LMB = 50 = 18 LMB LMB PC = 50 PC = 13 PC

a-

L li . 11 1 ml 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 9 10 11 12 13 14 15 16 17

September 1999- Snapper Creek Canal October 2001 - Snapper Creek Canal

LMB = 41 LMB = 19 PC = 50 PC = 30 >• >. U 40 U AO c I 30 20 P 20

4UU iUJL nfl. IImiJi 2 3 4 5 6 7 S g 10 11 12 13 14 15 16 17 2 3 4 5 6 7 8 0 10 11 12 13 14 15 16 17

October 1999- Snapper Creek Canal August 2001 - Cutler Drain Canal

LMB = 8 LMB = SO = PC 31 PC = 47 50 >< U 40 U 40 C « 30 CT 0) 20

n n i . i.hl.l 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17

November 1999- Snapper Creek Canal October 2001 - Cutler Drain Canal

LMB = 50 ^ 60 PC = 50 ^50 >> U 40 C » 30 D" Q) 20

, , ,1,1 5 B 7 8 9 10 11 12 13 14 15 16 17 5 6 7 8 9 10 11 12 13 14 15 16 17 Gape Width Group (mm) Gape Width Group (mm)

Figure 9. Gape width frequency for juvenile largemouth bass (LMB) and peacock cichlids (PC) electro fished from Snapper Creek Canal (1999 and 2001) and Cutler Drain Canal (2001), Miami-Dade County, Florida. Sample sizes are shown for each sampling event. 65

Prey Size and Stomach Fullness

Prey size increased with predator size for fish and decapod crustacean prey

(largemouth bass, r = 0.426, p = 0.0015; peacock cichlid, r = 0.490, p < 0.0001) (Figure

10). Moreover, there was relatively little difference in prey size between largemouth bass

and peacock cichlids on either a length-specific or gape-specific basis. Indeed,

ANCOVA detected no differences in either the slope (p = 0.608) or intercept (p = 0.165)

values of the prey length-predator length regressions. The largest estimates of prey

length relative to predator length were 34.4% for largemouth bass and 35.4% for peacock

cichlids. Prey body depth also was significantly correlated with largemouth bass total

length (r = 0.437, p = 0.0024) and peacock cichlid total length (r = 0.271, p = 0.0037).

The ANCOVA revealed no significant difference in slope between species (p = 0.177),

but did detect a significant difference in intercept values (p = 0.0013), with largemouth

bass having a higher value. Highest estimates of prey body depth relative to predator

gape size were 101.4% for largemouth bass and 95.0% for peacock cichlids.

Stomach fullness estimates (Table 2) for the July, September, and October 1999

Snapper Creek Canal samples were not significantly different between predator species.

However, juvenile largemouth bass had fuller stomachs than did peacock cichlids in the

November 1999 sample (Wilcoxon; p = 0.0163). In 2001, largemouth bass stomachs

were significantly fiiller than peacock cichlid stomachs in all samples (Wilcoxon; p <

0.0001 for all four sampling date by canal combinafions). There were no significant

differences in stomach fullness for largemouth bass among sampling events (Kruskal-

Wallace, p = 0.0541). In contrast, stomach fullness did vary significantly for peacock cichlids (K-W, p < 0.0001). Dunn's test demonstrated that peacock cichlid stomachs 66

50

A Largemouth bass 40 Peacock cichlid a A 'a CD A >=- 30 i I' 5 20 a I

A ft A 10 A

I— I— I— — — — I I I I —I 20 40 60 80 100 120 140 160 180 Predator Total Length (mm)

B

itiiiiii 0 20 40 60 80 100 120 140 160 180 Predator Total Length (mm)

Figure 10. (A) Prey length (mm) of decapod crustacean (folded body length) and fish (standard length) prey eaten by juvenile largemouth bass and peacock cichlids electrofished from Snapper Creek Canal (1999 and 2001) and Cutler Drain Canal (2001), Miami-Dade County, Florida. Results include 53 prey fi-om largemouth bass and 123 prey from peacock cichlids. (B) Body depth (mm) of decapod crustacean and fish prey eaten by juvenile largemouth bass and peacock cichlids electrofished fi-om Snapper Creek Canal (1999 and 2001) and Cutler Drain Canal (2001), Miami-Dade County, Florida. Results include 53 prey from largemouth bass and 1 1 3 prey fi-om peacock cichlids. The diagonal lines indicate estimated maximum gape for both predators based on the length- specific gape width regressions developed in the present study. 67

Table 2. Percent (%) ofjuvenile largemouth bass and peacock cichlid stomachs within

each stomach fullness category (i.e., 1 = empty, 2 = < 25% of estimated stomach capacity, 3 = 25-75% of stomach capacity, and 4 = > 75% of stomach capacity) for each sampling event. Specimens were electrofished from Snapper Creek Canal (SC) and Cutler Drain Canal (CD), Miami-Dade County, Florida.

Largemouth Bass Peacock Cichlid Stomach Fullness Stomach Fullness

Sampling Event Canal 1 2 3 4 1 2 3 4 July 1999 SC 33.3 11.1 44.4 11.1 23.1 15.4 38.5 23.1 September 1999 SC 20.0 5.0 45.0 30.0 6.7 13.3 50.0 30.0 October 1999 SC 12.5 50.0 12.5 25.0 16.1 29.0 25.8 29.0 November 1999 SC 22.2 0 22.2 55.6 24.1 41.4 27.6 6.9 August 2001 SC 8.0 26.0 32.0 34.0 33.3 41.2 17.7 7.8 October 2001 SC 12.2 14.6 51.2 22.0 29.4 56.9 9.8 3.9 August 2001 CD 12.2 38.8 34.7 14.3 50.0 37.5 6.2 6.3 October 2001 CD 6.1 49.0 28.6 16.3 56.0 40.0 4.0 0

were significantly fuller in the 1999 samples than in either canal in 2001, except that the mean rank for Snapper Creek Canal in August 2001 was not separable from the mean rank for Snapper Creek Canal in November 1999. Peacock cichlids in the October 2001

Cutler Drain Canal sample had the smallest mean rank stomach fullness of all the peacock cichlid samples.

Diet Overlap

Diets ofjuvenile largemouth bass and peacock cichlids overlapped on all dates

(Table 3). Of the eight samples collected, four samples had a Schoener's Index value >

0.60 for prey numerical abundance. In particular, largemouth bass and peacock cichlids overlapped considerably in volume. Five of eight overlap values exceeded 0.60 and the other three values exceeded 0.50. 68

Table 3. Schoener's Index values for dietary overlap by number and volume between juvenile largemouth bass and peacock cichlids electrofished from two southeast Florida "*" = canals (Miami-Dade County). A indicates large overlap (i.e., Cxy > 0.60). CD Cutler Drain Canal; SC = Snapper Creek Canal.

Schoener's Index Values Sampling Event Canal Numerical Occurrence Volume July 1999 SC 0.102 0.957 September 1999 SC 0.658 * 0.836 October 1999 SC 0.499 0.522 November 1999 SC 0.012 0.873 August 2001 SC 0.920 * 0.576 October 2001 SC 0.551 0.654 ' August 2001 CD 0.932 * 0.586 October 2001 CD 0.895 * 0.785

Discussion

The results of the present study broadly agree with earlier investigations of diets

ofjuvenile largemouth bass in Florida (Chew 1974; Cailteux et al. 1996; Huskey and

Tumigan 2001), although there were some important differences. Because prior studies

ofjuvenile largemouth bass diets were conducted in lakes or impoundments in central or

north-central Florida (versus southeast Florida canals), differences in stomach contents

could be the result of a suite of factors, including habitat type, latitude, or the presence of peacock cichlids.

Zooplankton was eaten by larger juvenile largemouth bass in the southeast Florida canals than was documented for most other Florida systems. Consumption of zooplankton in the present study occurred in largemouth bass in the 30-80 mm TL groups. Chew (1974) reported zooplankton from the stomachs of largemouth bass in the

3 1-75 mm TL range from two Florida lakes, but in the 76-296 mm TL range only in

Lake Weir. Neither Cailteux et al. (1996) nor Huskey and Tumigan (2001) found 69

zooplankton in fish of the size ranges reported in the present study (i.e., 58 mm TL and

larger).

The present study documented high consumption of insects by juvenile

largemouth bass in the canals relative to the previous studies. Insects were important

prey in , but not in Lake Hollinsworth (Chew 1974). Likewise, insects were

relatively important prey for juvenile largemouth bass in vegetated and unvegetated

north-central Florida lakes (Cailteux et al. 1996), but were a comparatively small part of

the diet offish in east-central Florida lakes and impoundments (Huskey and Tumigan

2001).

Consumption of decapod crustaceans varied widely in the studies. The data for

canal largemouth bass showed that decapod crustaceans were major prey for fish of 90

mm TL and larger. Chew (1 974) found that decapod crustaceans were important in the diet ofjuvenile largemouth bass in Lake Weir, but were absent fi-om stomach contents

fi-om Lake Hollinsworth. Cailteux et al. (1996) found that decapod crustacean were important prey in vegetated lakes but were of only minor importance in unvegetated lakes. Huskey and Tumigan (2001) reported substantial use of decapod crustaceans by largemouth bass of the sizes included in the present study.

Although nearly all size classes of largemouth bass in the present study included individuals with fish prey in their stomachs, only largemouth bass > 120 mm TL surpassed 60% frequency of occurrence, the designation of a "true" piscivore as defined by Bettoli et al. (1992; see also Cailteux et al. 1996). The study canals generally had dense beds of aquatic macrophytes along the shoreline and often throughout the deeper water column as well. Cailteux et al. (1996) found that largemouth bass in unvegetated 70

lakes became ftilly piscivorous (i.e., > 60% frequency of occurrence of fish in diet) by 60

mm TL, but did not become fiilly piscivorous until > 120 mm TL in vegetated lakes.

Similar results were documented by Bettoli et al. (1992) for heavily-vegetated Lake

Conroe, Texas, where largemouth bass did not become fijlly piscivorous until > 140 mm

TL. Predatory fishes are less efficient at capturing fish prey in structurally complex

habitats (Savino and Stein 1982; Hayse and Wissing 1996). Moreover, largemouth bass

has been found to increase diet breadth (i.e., feeding on invertebrates and fish) in

structurally complex habitats (Anderson 1984). Therefore, the late switch ofjuvenile

largemouth bass to piscivory in the southeast Florida canals may be related to the

abundant submersed vegetation.

Very little informafion has been published on the diet ofjuvenile peacock

cichlids. Lowe-McConnell (1969) reported palaemonid shrimp in 80-130 mm TL (?;

type of length measurement is uncertain but assumed to be TL) peacock cichlids (N = 4)

in Guyana. The reference is unclear, but fish also may have been eaten by peacock

cichlids < 1 60 mm TL. Additionally, the author found zooplankton and filamentous

algae in 40 mm TL (?) peacock cichlids. Again, the reference is unclear, but the sample

size of peacock cichlids < 160 mm TL was not likely to have exceeded 15 or 20

individuals. In his investigation of an introduced population occurring in Lake Gatun,

Panama, Zaret (1980a) examined stomach contents of seven Cichla specimens (either C. ocellaris or a congener) ranging in size fi-om about 55 to 90 mm SL (or about 70 to 1 12 mm TL using the conversion equation of Hill [1998]). The stomachs contained terrestrial insects, palaemonid shrimp, and fish. In Puerto Rico, introduced peacock cichlids < 1 50 mm TL ate fish and aquatic insects (Lilyestrom and Churchill 1996). In the only 71

published study from Florida, Shafland (1999b) documented fish of several species, but no non-fish prey, from peacock cichlids 100-149 mm TL in Tamiami Canal. The present

study was in general agreement with previous studies, but documented consumption of a much wider variety of prey.

The present study documented considerable overlap in gape width between juvenile largemouth bass and peacock cichlids of similar sizes. This finding agrees with

Hill (1998), who reported broad similarity in gape size of these species over a wide range

of lengths (i.e., largemouth bass \ mm TL; peacock cichlids 50-430 mm TL).

Given the extent of temporal and spatial overlap in total length distributions found in this

study, juvenile largemouth bass and peacock cichlids of similar size and morphological

prey-handling ability co-occurred in both canals. Morphological factors, coupled with

syntopy, implied that juvenile largemouth bass and peacock cichlids had a high potential

for undergoing similar dietary shifts and therefore sharing common food resources.

As suggested by morphology, the ontogenetic prey-use pattern of largemouth bass

and peacock cichlids was similar (i.e., zooplankton, followed by insects, then decapod

crustaceans and fish). The main dietary differences were in a slightly broader range of

prey eaten by largemouth bass (e.g., inclusion of crayfish in addition to grass shrimp in

the decapod crustacean category and inclusion of anurans in the diet) and in the timing

(i.e., total length) of shifts in prey proportions. For example, peacock cichlids consumed

zooplankton at larger sizes than did largemouth bass (i.e., through the 1 00 mm TL group

versus the 80 mm TL group for largemouth bass). Furthermore, amphipods and insects 1

72

100

Largemouth Bass Peacock Cichlid

Figure 1 1 . Frequency of occurrence (% of predator stomachs with the prey type) for

largemouth bass (N = 1 13; 100-349 mm TL) and peacock cichlids (N = 156; 100-348 mm TL) sampled from Tamiami Canal, Miami-Dade County, Florida. Data from Shafland 1999b.

E] Insect B Decapod

Fish

Largemouth Bass Peacock Cichlid

Figure 12. Percent (%) by number of prey for largemouth bass (N = 1 13; 100-349 mm TL) and peacock cichlids (N = 156; 100-348 mm TL) sampled from Tamiami Canal, Miami-Dade County, Florida. Data from Shafland 1999b.

were generally more important for largemouth bass than peacock cichlids, although certain peacock cichlid size classes included these in high numbers and frequency.

Stomach contents data from another southeast Florida canal also show a similar pattern of 73

consumption of decapod crustaceans and fish by largemouth bass and peacock cichlids of

100 to about 350 mm TL (Shafland 1999b; Figures 1 1 and 12).

Initiation of piscivory is an important ontogenetic event for predatory fishes

(Aggus and Elliott 1975; Pasch 1975; Ludsin and DeVries 1997). Fish were included in

the peacock cichlid diet at a generally higher frequency, number, and volume than in the

largemouth bass diet at all predator sizes and peacock cichlids may have started to eat

fish at a shorter total length than did largemouth bass. As a result, peacock cichlids

became piscivorous (sensu Bettoli et al. 1992) before largemouth bass (i.e., by the 100 mm TL group versus the 120 mm TL group). These trends were even more pronounced based on numeric and volumetric proportions of fish in the diets ofjuvenile predators

(e.g., for predator size classes > 100 mm TL, largemouth bass averaged [mean ± 1 SE]

56.0 ± 1 1 .9% fish by volume versus 88.1 ± 1.7% for peacock cichlids).

The percentage of empty stomachs was high for peacock cichlids and much lower for largemouth bass. This translated into significantly higher stomach fiallness estimates for largemouth bass on several sampling dates. These data agree with the results of small and large tank studies, which documented lower feeding rates for peacock cichlids than

for largemouth bass (Chapter 5). Relatively low daily rations for a congener, Cichla

monoculus . also have been reported (Rabelo and Araujo-Lima 2002). Surprisingly, however, juvenile peacock cichlids may grow faster than juvenile largemouth bass.

There was substantial overlap in the total lengths ofjuveniles throughout the sampling period. Because largemouth bass spawn earlier than peacock cichlids (Shafland 1999b), the overlap in length frequencies in the present study suggests faster growth of age-0 peacock cichlids. Unfortunately, there are no studies available that directly measured 74

growth rates ofjuvenile largemouth bass and peacock cichlids. If lower stomach fullness

are due to lower feeding rates, then these results suggest that peacock cichlids are more

efficient than largemouth bass (i.e., that despite lower consumption rates they are able to

grow faster). Because a shift to a fish diet has been linked to increases in growth rates in

piscivorous fishes (Pasch 1975; Timmons et al. 1980; Olson 1996), early and substantial

consumption of fish prey by juvenile peacock cichlids may have resulted in their apparent

faster growth rates relative to largemouth bass.

Not only were the stomach contents ofjuvenile largemouth bass and peacock

cichlids similar in taxonomic identity, they also were similar in relative prey size. For

both species, increasing prey size was weakly correlated with increased predator length.

The only statistically significant difference in prey size between the species being

marginally larger (i.e., deeper-bodied) prey in largemouth bass. This result may be due to

the fact that a largemouth bass has a slightly larger gape width than a peacock cichlid of

the same total length.

Competition for food can be an important interaction for juvenile largemouth bass. For example, sunfishes (i.e., Lepomis spp.) and juvenile largemouth bass have been

shown to compete for food in the littoral zones of temperate lakes, possibly limiting

largemouth bass recruitment (Osenberg et al. 1994; Olson et al. 1995). Likewise, growth

and survival of largemouth bass juveniles is density-dependent (Olson et al. 1995;

Garvey et al. 2000). The presumed mechanism is intraspecific food competition. The results of these studies suggest that food may be limiting for juvenile largemouth bass under some circumstances and that the presence of fishes with similar food requirements and habitat use may be detrimental to largemouth bass abundance. 75

Species with similar ecological requirements which consume resources from the

same habitat should have higher competition coefficients than a species pair which feeds

in different habitats or consumes different resources within a habitat (Rosenzweig 1981).

Juvenile largemouth bass and peacock cichlids overlapped considerably in diet and such

large overlap index estimates have been considered significant (Zaret and Rand 1971;

Mathur 1977). Nevertheless, diet overlap indices by themselves do not necessarily

demonstrate competition (Colwell and Futuyma 1971; Hulbert 1978). For example, high

overlap can occur without competition when resources are abundant. Rock-dwelling

cichlids (i.e., "mbuna"; Cichlidae) in Lake Malawi, , exhibit extreme trophic

differentiation based on differences in tooth structure, mouth morphology (e.g., size and

orientation), and pharyngeal jaw apparatus (Fryer and lies 1972). Normally, most of

these species are herbivorous, grazing on aufwuchs (i.e., bio-cover of algae and bacteria)

in specific microhabitats (Ribbink 1991). Nevertheless, during times of high zooplankton

abundance, mbuna will abandon their trophic specialties and feed on superabundant

zooplankton (McKaye and Marsh 1983; Ribbink et al. 1983). Dietary overlap is very high during these periods, but competition is not. On the other hand, Rosenzweig (1981) modeled two-species competition and concluded that a competition coefficient of zero

(implying little or no resource overlap) was possible, even between two strong competitors, due to competitor-induced habitat shifts. Indeed, low overlap in diet could be the consequence of competition. Diets of native lake trout Salvelinus namaycush significantly differ fi-om diets of introduced Micropterus dolomieu and

rock bass Ambloplites rupestris in Canadian lakes (Vander Zanden et al. 1999). In these lakes, smallmouth bass and rock bass eat littoral prey fish and lake trout feed primarily on 76

zooplankton and include fish prey as only a small part of the diet. However, in similar

lakes without introduced smallmouth bass or rock bass, lake trout diets contain a

considerably higher percentage of littoral prey fish. These results imply a competition-

induced diet shift.

Indeed, high dietary overlap values have been used both to refute and to

demonstrate the existence of interspecific competition in fishes (Colwell and Futuyma

1971; Hulbert 1978). In some studies, high dietary overlap between two fishes is

interpreted as evidence that food resources are not limiting (i.e., therefore competition is not occurring) or to explain observed interspecific habitat or activity period differences

(e.g., Zaret and Rand 1971; Keast 1978, 1985; Keast et al. 1978; George and Hadley

1979; Winemiller 1989; Hirst and DeVries 1994). Much of this research is based on morphologically similar species, feeding guilds, or entire species assemblages that are native to the study area, apparently have coevolved, and presumably have reacted to past competition through character displacement or resource partitioning. In other studies,

particularly investigations of nonindigenous fishes, dietary overlap is used as evidence of

food competition or of the high potential for food competition to occur (e.g., Mathur

1977; Parrish and Margraf 1990; Zale and Gregory 1990; Scoppettone 1993; Ogle et al.

1995;Declercket al. 2002).

Dietary data trom southeast Florida canals show high overlap, and previous studies in temperate lakes demonstrate that largemouth bass are food-limited.

Unfortunately, there are no good tests of food limitation in southeast Florida canals, although there are studies of the abundances of small fishes upon which largemouth bass and peacock cichlids feed. During the past two decades (i.e., since the introduction of 77

peacock cichlids in 1 984) qualitative and quantitative surveys of fishes in the canals where largemouth bass and peacock cichlids co-occur have found high numbers of small fishes, both native and nonindigenous (boat electrofishing surveys: L. G. Nico, USGS,

unpubl. data; pers. obs.; blocknet sampling: Shafland et al. 1985; Shafland 1999a;

Shafland et al. 2001). These high densities suggest that small fish availability does not limit predators such as largemouth bass and peacock cichlids. However, dynamics of invertebrate prey, which may influence early growth patterns of largemouth bass and peacock cichlids, have not been investigated in these canals. Finally, if food was limiting, then largemouth bass population densities should have declined following the introduction of peacock cichlids, a potential competitor. Nevertheless, available data do not suggest that largemouth bass populations have declined during the 20-year period following the introduction of peacock cichlids (Shafland 1995, 1999a, c; Shafland and

Stanford 1999; Shafland et al. 2001; Chapter 6).

Fishes that are piscivorous as adults typically undergo marked changes in diet related to increases in body length and mouth size (Keast 1985, Keast and Eadie 1985;

Olson 1996; Mittlebach and Persson 1998; present study). Knowledge of the extent to

which ecomorphological similarity is predictive of the timing (i.e., predator size) and prey progression of such shifts can be important in evaluating the risks associated with fish introductions, whether planned or illegal. The results of the present study suggest that ecomorphological similarity to a native fish is predictive of dietary ontogeny and dietary overlap for nonindigenous species. This implies that such information can be used for initial assessments of risk prior to fish introductions in the absence of diet data for early life history stages of native or introduced populations of the nonindigenous 78

species. Studies of nonindigenous fishes typically concentrate on adult stages. However,

careful attention also must be paid to early life stages and the possible interactions of

such stages with native fishes for those species that differ in habitat and food use through

ontogeny (Courtenay 1995). Given the ecomorphological convergence of many cichlids

and centrarchids (Norton and Brainerd 1993), using similarity to predict food resource

use may be particularly valuable when cichlids invade systems where centrarchids are

native (e.g., Florida) or when centrarchids invade systems in Africa, Central America, or

South America. CHAPTER 5 EXPERIMENTAL PREY HANDLING AND SELECTIVITY OF TWO MORPHOLOGICALLY SIMILAR PREDATORY FISHES- NATIVE LARGEMOUTH BASS AND INTRODUCED PEACOCK CICHLIDS

Introduction

The morphology of predatory fishes is known to have a profound effect on their

prey use. For example, gape limitation is an important constraint for many predatory

fishes. The size of the mouth, pharyngeal, or intercleithral opening (i.e., gape size) places limits on maximum prey size and influences the energetic return from submaximal prey

(Lawrence 1958; Werner 1974, 1977, 1979; Zaret 1980b; Wainwright 1988; Wainwright and Richard 1995). Furthermore, patterns of prey selection frequently scale with patterns of gape size (Wainwright and Richard 1995; Mittelbach and Persson 1998). The constraint of gape limitation often results in an ontogenetic dietary progression characterized by the initial consumption of zooplankton, then of intermediate-sized invertebrates, and finally by fish (Keast 1985; Wainwright and Richard 1995).

Given this relationship of gape size to prey use, a reasonable prediction is that an introduced species with gape size (and other features of morphology) similar to a native species will also exhibit similarifies in prey handling ability and prey selection.

Additionally, such similarities suggest that the nonindigenous species will exert comparable pressures on prey species. These predictions are of general concern following a fish introduction because they imply the potential for food competition

79 80

between the invader and any ecomorphologically similar native species. Indeed,

competition is the most often cited mechanism whereby introduced fishes negatively

effect native fishes (Moyle et al. 1986).

The native largemouth bass (Centrarchidae) is an economically important predatory fish in southern Florida (Shafland and Stanford 1999). The peacock cichhd

Cichla ocellaris (Cichlidae), a species native to South America, was legally introduced into freshwater canals in southeast Florida in 1984-1987 by the Florida Game and Fresh

Water Fish Commission (currently Florida Fish and Wildlife Conservation Commission)

(Shafland 1995). The predatory peacock cichlid was stocked to reduce the abundance of

other nonindigenous fishes, particularly spotted tilapia Tilapia mariae , and to augment recreational fisheries that primarily targeted largemouth bass (Shafland 1995).

Largemouth bass and peacock cichlids are ecomorphologically similar based on appearance, body size, gape size, feeding mechanics, and diet (Norton and Brainerd

1993; Hill 1998; Chapter 4). Dietary overlap ofjuveniles in southeast Florida canals is generally large (Chapter 4) and adults of both species are regarded as piscivorous

(Carlander 1977; Shafland 1995, 1999b). Additionally, experimental studies of largemouth bass prey selection have demonstrated that fish are often selected at a higher

rate than non-fish prey (e.g., Lewis et al. 1961). There was considerable overlap in diets of subadult and adult largemouth bass and peacock cichlids in Tamiami Canal (Shafland

1999b). Nevertheless, largemouth bass consumed substantial amounts of non-fish prey

(54 % by number; primarily decapod crustaceans, but also insects) in addition to fish whereas co-occurring peacock cichlids were almost exclusively piscivorous (Shafland

1999b). 81

These results suggest that ecomorphological similarity can predict coarse patterns

but is not precise enough to predict patterns of a finer scale. Alternately, some other

mechanism (e.g., differential encounter rates mediated by habitat, predator behavior, or competition) may be operating to modify the prey use patterns of one or both species.

Prey selection by predators results from the interaction of differential encounter rates with prey, probabilities that a predator will attack a prey item if encountered, and probabilities that an attack results in consumption (O'Brien 1979; Osenberg and

Mittelbach 1989). Encounter rate is fundamental to the process because it sets the initial conditions from which the remaining predation components operate (Osenberg and

Mittelbach 1 989). For example, rarely encountered prey may represent a small portion of a predator's diet even if the product of attack and success probabilities approaches unity

(i.e., 1.0). Encounter rates, especially under field conditions, are difficult to estimate

(Osenberg and Mittelbach 1989). Moreover, structurally complex habitat can greatly reduce encounter rates and feeding efficiency of predators (Savino and Stein 1982;

Anderson 1984; Hayse and Wissing 1996). Compefifion between predators also can reduce encounter rates with prey by prey depletion, by modifying prey behavior in response to predation risk, or by direct interference with predator feeding behavior

(Hobson 1979; Mittelbach 1984; Fausch et al. 1997; Turner et al. 1999). Alternately, species-specific hunting behavior of predators may influence the rate of encounter with different prey types (Cooper et al. 1985; Magnhagen 1986; Grant and Noakes 1987).

However, by creating situations where encounter rates are similar between two predators

(e.g., by making encounters almost certain), attack and success probabilities can be estimated and comparisons can be made. 82

I conducted a series of tank experiments with largemouth bass and peacock

cichlids to test predictions concerning prey use by similar nonindigenous and native

fishes (i.e., similar morphology equals similar prey handling ability, prey selection, and

thus potential effects on prey base). I used estimates of prey handling times to test for differences in the ability of both predators to handle prey. For example, prey handling and the resulting information on prey profitability are important for predictions of prey

selection based on optimal diet models (Emlen 1966; MacArthur and Pianka 1966). This experiment also provided information on the probability of attack and successful consumption of prey for both predators within the experimental system. This information was used to formulate and test predictions of prey selection for largemouth bass and peacock cichlids given choices among prey types.

Communities often have multiple species of predators (Sih et al. 1998) and a number of studies have investigated the equivalence of multiple predators in their effect on prey assemblages (Harris 1995; Morin 1995; Kurzava and Morin 1998; Baber and

Babbitt 2003). The introduction of peacock cichlids into southeast Florida canals represents the addition of a predator to a predator assemblage formerly dominated by a

single species, the largemouth bass (Chapter 2). Therefore, I tested assemblages made up of varying densities and percent composition of largemouth bass and peacock cichlids for predator assemblage effects on prey selection and prey mortality.

Methods

Experiment 1 - Prey Handling

I conducted a laboratory experiment to test the relative ability of largemouth bass and peacock cichlids to capture and consume fish and non-fish prey. This experiment 83

was conducted in the University of Florida Demonstration Aquaculture Recirculating

System (see Hill and Cichra in prep, for a complete description). This indoor system has

1.1 -m diameter by 0.6-m high, circular fiberglass tanks connected to biofilters. Tank interiors were light blue and had single airstones and 50-mm central standpipes. Water temperature was maintained between 25.5° and 27° C and the light:dark cycle was 13 h:ll h. Eight largemouth bass were electrofished from Lake Alice, Alachua County,

Florida and eight peacock cichlids were electrofished from Cutler Drain Canal, Miami-

Dade County, Florida. The largemouth bass averaged 322 mm total length (TL) (255-

408 mm TL) and the peacock cichlids averaged 340 mm TL (251^07 mm TL). Prey were collected from ponds at the Department of Fisheries and Aquatic Sciences,

University of Florida, Gainesville, Florida.

Prey handling times, attack probability, and probability of success given an attack were estimated for single prey items presented to individual predators. Prey used in the experiment were of three representative prey types— 1) a soft-rayed fish (i.e., sailfin

molly Poecilia latipinna), 2) a spiny-finned fish (i.e., bluegill Lepomis macrochirus ), and

3) a decapod crustacean (i.e., crayfish [a mix of Procambarus fallax and P.

paeninsulanus]) . Each prey type is present in southeast Florida. Prey sizes were chosen

to encompass a range of about 20-100% of predator gape size (i.e., prey body depth

divided by predator gape width and multiplied by 1 00) based on relafions for bluegill, largemouth bass, and peacock cichlid found in Hill (1998) and for crayfish and sailfin molly developed in the present study. Length-specific body depth relations for crayfish and sailfin molly were developed from fresh specimens (PROC GLM; SAS Institute,

Cary, North Carolina): 84

Crayfish: BD = 0.339 TL - 2.841 ; = 0.90

Sailfin Molly: BD = 0.3 1 8 TL - 3.890; = 0.98 where BD = maximum body depth. Handling time was measured as the total time from

initial prey contact to the cessation of all feeding movements of mouth and opercula, including time spent recapturing released prey. In the small tanks, initial searching and pursuit times were essentially zero. Prey mass was estimated using length-mass regressions developed from fresh specimens (PROC GLM; SAS Institute, Gary, North

Carolina).

Bluegill: LogioWT = 2.790 LogioTL - 4.525; r^ = 0.95

Crayfish: LogioWT = 3.498 LogioTL - 5.472; ^ = 0.97

Sailfin Molly: LogioWT = 3.388 LogioTL - 5.436; r^ = 0.99

Energy values for prey were taken from the literature— 1 160 cal/g for centrarchids

(Miranda and Muncy 1989), 1 100 cal/g for other fish (Pope et al. 2001), and 750 cal/g for

crayfish (Pope et al. 2001) for wet weights.

I incorporated effects of prey size and predator size by dividing prey body depth

by predator gape (and multiplied by 100), which I called relative prey size, and then categorized relative prey size into 10% size classes. Two-way analysis of variance

(ANOVA) was used to detect differences in mean prey profitability across relative prey size classes for sailfin molly and bluegill prey using predator species and relative prey size class as main effects (PROC GLM; SAS Institute, Gary, North Carolina). Non- significant interacfion terms (i.e., predator x size class) were dropped from the model

(bluegill interacfion F5^54 = 1 .2 = 0.3 1 sailfin = = 1 , p 9; molly interacfion Fs.ss 1 .55, p

0.212). One-way ANOVA was used for crayfish prey because peacock cichlids did not 85

eat crayfish. ANOVAs were followed by the least squares means multiple comparison procedure. ANOVA (followed by Tukey's HSD multiple comparison procedure) was also used to test for the effect of prey size and type on the probability of successful consumption of prey for both predators given that an attack had taken place (PROC

GLM; SAS Institute, Gary, North Carolina). Two-sample t-tests were conducted to investigate differences between predators in the probability of successfully consuming each prey type (PROC TTEST; SAS Institute, Cary, North Carolina). Probabilities were

arcsine-square root transformed prior to analysis. All tests were conducted at a Type I error rate (a) of 0.05.

Experiment 2 - Prey Selection by Individual Predators

Prey selection patterns were determined empirically in the same laboratory tanks

that were used in Experiment 1 . A single predator per tank was tested and eight largemouth bass and eight peacock cichlids were tested. Eight largemouth bass were electrofished from Lake Alice, Alachua County, Florida and eight peacock cichlids were electrofished from Cutler Drain Canal, Miami-Dade County, Florida. The largemouth bass averaged 322 mm total length (TL) (255^08 mm TL) and the peacock cichlids averaged 340 mm TL (251^07 mm TL). Prey were collected from ponds at the

Department of Fisheries and Aquatic Sciences, University of Florida, Gainesville,

Florida. Predators were fasted the day prior to each experiment. Predators not being

tested were fed crayfish, eastern mosquitofish Gambusia holbrooki , sailfin mollies, bluegills, and redear sunfish Lepomis microlophus . Prey species selectivity was tested using three representative prey types—a soft-rayed prey (i.e., sailfin molly), a spiny- finned prey (i.e., bluegill), and a decapod crustacean (i.e., crayfish). Five of each prey 86

type were presented simultaneously to each predator (i.e., total of 1 5 prey per trial). Prey body depths were 40-60% of predator gape size. Remaining prey were retrieved and

counted after one day (i.e., about 24 h).

Prey selectivities for largemouth bass and peacock cichlids feeding on fish and crayfish prey were calculated using the Manly-Chesson index (Manly 1974; Chesson

1978, 1983), accounting for prey depletion during the experiment (i.e., Case 2; Chesson

1983):

ln((n,o-r,o)/«,.o) ^, = 1^...^^

7=1

where m is the number of prey categories, r/o is the number of prey type i in the diet and

n,o is the number of prey / in the environment. Values of a range from 0 to 1 , with 0

being avoidance, 1 being complete preference, and \lm being random selection. One- way ANOVA (followed by the Tukey HSD multiple comparison procedure) was used to

test the hypothesis of random prey selection for each predator (i.e., all prey categories equally selected) (PROC GLM; SAS Institute, Cary, North Carolina). Two-sample t-tests

were used to determine if prey selectivity (i.e., a) differed between largemouth bass and peacock cichlids for sailfin molly and bluegill prey (PROC TTEST; SAS Institute, Cary,

North Carolina). The Wilcoxon signed rank test was used to test for differences in mean numbers of prey eaten per predator between largemouth bass and peacock cichlids

(PROC NPARl WAY; SAS Institute, Cary, North Carolina). A significance level of a =

0.05 was used for all statistical analyses.

Data on probability of attack and probability of successfiil consumption from

Experiment 1 were used to predict prey selection for largemouth bass and peacock 87

cichlids in Experiment 2, assuming that all available prey were encountered (given the small size of the tanks), and subject to the constraint that the total number of prey eaten was equal to that observed. Chi-square tests were used to test the hypothesis that the predicted diet was not different from the observed diet for each predator species (PROC

FREQ; SAS Institute, Gary, North Carolina).

To determine if decapod crustaceans are acceptable prey for subadult and adult

peacock cichlids given the lack of consumption of crayfish in the experiments, I stocked

1 0 crayfish into tanks with individual peacock cichlids and offered no alternative prey for five days. This experiment was conducted at the Florida-Caribbean Science Center, U. S.

Geological Survey, Gainesville, Florida in indoor, 1 .2-m diameter by 0.6-m high, circular fiberglass tanks. Tank interior color was light blue and the interior was bare except for a

single airstone and a central 1 0-cm standpipe. There was a constant, slow flow of well water through each tank and water temperature was maintained at about 26.7° C by the use of two 300-w heaters. Peacock cichlids were collected by electrofishing from Cutler

Drain Canal and Snapper Creek Canal in Miami-Dade County, Florida and crayfish were trapped from ponds at the Department of Fisheries and Aquatic Sciences, University of

Florida, Gainesville, Florida. A total of five peacock cichlids ranging from 196 to 355 mm TL were tested. Crayfish body depths were smaller than predator gape width (Hill

1998).

Experiment 3 - Prey Selection by Predator Assemblages

I conducted an experiment at the Sam Mitchell Aquaculture Demonstration Farm

(SMADF), University of Florida, Blountstown, Florida, to test for differences in the effect of predafion by assemblages of largemouth bass and peacock cichlids on prey 88

assemblages consisting of three representative prey types. The experimental tank system

consisted of 6-m diameter by 1 .2-m high circular fiberglass tanks with a central 10-cm, screened standpipe, several 15-cm airstones around the tank perimeter, and a constant flow of pond water from a single, 0.4-ha, aerated pond. Tank interior color was off- white. Water in each tank was maintained at a depth of about 45 cm and had a temperature of 26 - 29° C during the experiment. Morning dissolved oxygen readings in

the tanks exceeded 8.0 mg/L. Water clarity (i.e., vertical secchi depth) in the pond exceeded pond depth (> 2.4 m) and the mean value in the tanks for color was 8.82 Pt-Co

(SD = 1.29) and for turbidity was 0.32 NTU (SD = 0.18). The tanks were outdoors under a 20% shade cloth covering.

Largemouth bass and peacock cichlids were used as predators and golden shiners

Notemigonus crysoleucas , redear sunfish, and crayfish were used as prey. Largemouth bass were angled from ponds at SMADF and the Department of Fisheries and Aquatic

Sciences (FAS), University of Florida, Gainesville, Florida. Peacock cichlids were electrofished from Cutler Drain Canal, Miami-Dade County, Florida. The largemouth bass averaged 340 mm TL (299-376 mm TL) and the peacock cichlids averaged 324 mm

TL (290-365 mm TL). Golden shiners and redear sunfish were purchased from bait and

fingerling suppliers. Crayfish were trapped from ponds at FAS. It is unlikely that any prey had experience with large predatory fish. Prey species were chosen to represent distinct prey types— 1) an elongate, soft-rayed fish (i.e., golden shiner), 2) laterally-

compressed, spiny-finned fish (i.e., redear sunfish), and 3) a decapod crustacean (i.e., crayfish). These species are present in southeast Florida canals. Prey sizes used were smaller than predator gape size according to length-specific gape width relations from 89

Hill (1998). Prey were chosen to be similar in body depth (mean BD = 16-18.5 mm), the prey dimension of most importance to gape-limited predators (Lawrence 1958;

Hambright 1991). Mean prey length and weight were compared by one-way ANOVA followed by the Tukey HSD multiple comparison test (PROC GLM; SAS Institute, Gary,

North Carolina). Golden shiners were longer than crayfish and redear sunfish (ANOVA,

F2,78 = 69.31, p < 0.0001) and heavier than redear sunfish (ANOVA, F2,7g = 38.03, p <

0.0001). Crayfish and redear sunfish were of similar length, but crayfish were heavier than golden shiners and redear sunfish (ANOVA, F2,78 = 38.03, p < 0.0001).

Nine predator treatments were tested with an equal initial food resource level of

50 each of golden shiners, redear sunfish, and crayfish. Predator density was four, six, or

eight predators (i.e., low, medium and high predator density). Within each density level, the species composition of predators was 100% largemouth bass, 100% peacock cichlid,

or 50% of each species, giving nine treatment combinafions (i.e., 3 predator densities x 3 species combinations = 9). A tenth treatment, zero predators, was used to estimate losses not due to predation, but was not included in the analyses because no losses were observed in this treatment. Treatments were randomly assigned to tanks and each

treatment was run three times. Trials were run for three days. Tanks were inspected at least twice daily for prey mortality. Missing prey were not replaced. At the end of the trial, the tanks were drained and prey were removed and counted. Predator densities, initial prey numbers, and trial duration were chosen to provide a range of resource availability fi-om abundant (i.e., the most selected prey would not be eradicated) to scarce

(i.e., highly selected prey will be eradicated and alternative prey should be exploited). 90

Prey selectivity values for golden shiners, redear sunfish, and crayfish were estimated for each predator density x combination treatment using the Manly-Chesson

arcsine-square root- index (i.e., a , Case 2, with depletion) (Chesson 1983). Values were transformed prior to analysis. Two-way ANOVA, followed by the least squares means multiple comparison procedure was used to test for the effects of predator assemblage

characteristics (i.e., predator density, combination, and density x combination interaction) on mean a values (PROC GLM; SAS Institute, Gary, North Carolina). Non-significant interactions were excluded fi^om the model and ANOVA was repeated with only the main effects.

In addition to examining prey selectivity (which is a measure of the relative

mortality imposed on prey types), I also compared the absolute mortality imposed on

prey. I used two-way ANOVA to test the effect of predator density and combination on

the logio-transformed odds of a predator eating a prey during the trial versus leaving it

alive at its end, followed by the least squares means multiple comparison procedure

(PROC GLM; SAS Institute, Gary, North Carolina). Non-significant interactions (i.e., predator density x predator combination) were excluded fi'om the model and ANOVA was repeated with only the main effects. One-way ANOVA was used to test for differences in mean per capita consumption of crayfish by largemouth bass (PROC GLM;

SAS Institute, Gary, North Carolina). Type I error rate (a) was set at 0.05.

Results

Experiment 1 - Prey Handling

In general, handling times increased with prey size but varied between largemouth bass and peacock cichlids and among prey types; however, there was considerable 91

variation in these relationships leading to pronounced variation in prey profitability.

Sailfin mollies were more energetically favorable prey for peacock cichlids than for largemouth bass (ANOVA, Fi,58= 16.48, p = 0.0001; Figure 13A). Nevertheless, sailfin mollies were of high value relative to other prey for largemouth bass (Figures 13A and

13B). Addifionally, the relafive size of sailfin mollies did not have a significant effect on their profitability (ANOVA, F4,58 = 0.43, p = 0.785). Contrary to sailfin molly prey, profitability of bluegill prey did not depend on predator species (ANOVA, Fi,59= 2.33, p

= 0.1322), but was influenced by relafive prey size (ANOVA, F5,59= 8.14, p < 0.0001;

Figure 13B). The smallest size classes of bluegill (i.e., 40 and 50%) were more profitable than were larger bluegills. Only largemouth bass ate crayfish during this experiment.

The relative size of crayfish prey did not have a significant effect on crayfish profitability for largemouth bass (ANOVA, F3,43 = 2.50, p = 0.072; Figure 13B); however, crayfish in the larger than 70 % of gape size were offered to largemouth bass but were not eaten.

Largemouth bass and peacock cichlids always attacked fish prey (i.e., sailfin

mollies and bluegills) placed into the experimental tanks (Table 4). Largemouth bass usually attacked crayfish, but peacock cichlids never attacked crayfish. Addifionally, successful consumption of attacked prey varied among prey types. Largemouth bass were more successfial at consuming fish than crayfish (ANOVA, F2,i78 = 9.33, p =

0.0001) and peacock cichlids were more successful with sailfin mollies than with bluegills (ANOVA, F|,48 = 7.68, p = 0.008). Prey size did not influence predator success

(given that an attack occurred) for largemouth bass eating sailfin mollies (ANOVA, F5,42

= 0.95, p = 0.460) or bluegills (ANOVA, F5,54 = 0.35, p = 0.877) or for peacock cichlids 92

Figure 13. Estimated mean profitability (± 1 SE) to largemouth bass (LMB) and peacock cichlids (PC) of sailfm molly (SFM) (A), bluegill (BG) (B), crayfish (CF) (B) versus relative prey size. BD = body depth; GW = gape width. 93

eating sailfin mollies (all attacked sailfin mollies were successfiilly consumed) or bluegills (ANOVA, F5,23 = 0.42, p 0.832). However, largemouth bass had lower success after an attack for the largest size class of crayfish eaten (i.e., 60%) compared to the smallest crayfish size class (i.e., 30%) (ANOVA, F3,69 = 2.94, p = 0.039). Comparing the predators to each other, peacock cichlids were more successfial in consuming sailfin

mollies than were largemouth bass (t-test, to.os, 47 = -2.34, p = 0.024), but there was no

significant difference between predators for bluegills (t-test, to.05, gv = 0.09, p = 0.928).

Because only largemouth bass ate crayfish, it may be inferred that largemouth bass were more successful in their consumption of crayfish prey.

Table 4. The probability of attack [Prob(A)], probability of successful consumption given an attack [Prob(SC)], and the product of those probabilities [Prob(AxSC)] for individual largemouth bass (N = 8) and peacock cichlids (N = 8) sequentially offered three prey types. No Prob(SC) was estimated for peacock cichlids feeding on crayfish because there were no attacks (i.e., ND = no data).

Largemouth Bass Peacock Cichlid

Prob(A) Prob(SC) Prob(AxSC) Prob(A) Prob(SC) Prob(AxSC)

Sailfin 1.0 0.896 0.896 1.0 1.0 1.0 Molly

Bluegill 1.0 0.738 0.738 1.0 0.704 0.704

Crayfish 0.884 0.548 0.484 0.0 ND 0.0 94

Experiment 2 - Prey Selection by Individual Predators

Prey selection by both largemouth bass (ANOVA, ¥2,21 = 32.61, p < 0.0001) and

peacock cichlids (ANOVA, ¥2,21 = 22.10, p < 0.0001) in the small tanks was non-random.

Prey fish species were selected at a high frequency by predators relative to crayfish

(Figure 14 and Table 5). Bluegills and sailfin mollies were not selected at a significantly different rate by largemouth bass; however, peacock cichlids had higher selectivity index values for sailfin mollies than for bluegills. T-tests revealed no difference between

selectivity estimates for bluegill (to.05, i4 = 1 -57, p = 0.138), but peacock cichlids had higher mean selectivity for sailfin molly prey than did largemouth bass (to.05,14 = -2.24, p

= 0.042). Largemouth bass ate significantly more prey per individual than did peacock

cichlids (Wilcoxon, z = 1 .981, p = 0.048).

1^ Largemouth Bass

0.8 • A Peacock Cichlid

.e 0.6 4

u i CO

0.2

Sailfin Molly Bluegill Crayfish

Figure 14. Observed mean prey selectivity values (a) (± 1 SE) for individual largemouth bass (N = 8) and peacock cichlids (N = 8) preying on sailfin mollies, bluegills, and crayfish in small tanks. The horizontal dashed line indicates random prey selecfion (i.e., a = \/m = 0.333, where m = number of prey categories). =

95

Prey selection by largemouth bass differed from the predicted pattern based on

= = 0.026; attack and success probabilities estimated in Experiment 1 {"i 7.28, df 2, p

Table 5). In particular, crayfish were underrepresented in the observed diet and both fish prey were overrepresented. The diet change was in the direction of the more profitable

prey (i.e., fish > crayfish), suggesting that attack probabilities decreased for crayfish in the presence of more profitable prey (i.e., fish). On the other hand, prey selecfion by

peacock cichlids did not significantly deviate from the predicted values (x^ = 0.245, df l,p = 0.621).

Table 5. Predicted and observed prey selection for largemouth bass (N = 8) and peacock cichlids (N = 8) simultaneously presented with three prey types. Predicted diets were derived from estimates of the product of the probability of attack and the probability of successfial consumption given an attack obtained in Experiment 1 (Table 3). % = percent of prey; N = number of prey (based on observed consumpfion of 57 prey by largemouth bass and 35 prey by peacock cichlids).

Largemouth Bass Peacock Cichlid

Predicted Observed Predicted Observed Diet Diet Diet Diet

% N % N N % N

Sailfin 42 24 51 29 59 21 66 23 Molly

Bluegill 35 20 44 25 41 14 34 12

Crayfish 23 13 0 96

No crayfish were eaten by peacock cichlids in the prey handhng or prey

selectivity trials. Unlike largemouth bass, peacock cichlids also did not eat crayfish

during regular feedings (i.e., non-experimental), but readily ate offered fish prey. Indeed, crayfish were not acceptable prey for most of the peacock cichlids, even in the absence of alternative prey. Of the 50 crayfish offered to peacock cichlids that had no alternative

prey for five days, only one (i.e., 2%) was eaten. All five experimental peacock cichlids subsequently ate several fish prey on the day after the experiment concluded.

Experiment 3 - Prey Selection by Predator Assemblages

There was no significant effect of predator assemblage on prey selectivity in the

present study (Figure 1 5). Predator density had no significant effect on prey selectivity

= (golden shiner = 1 .2 = 0.3 1 redear sunfish ANOVA, = 0.97, ANOVA, ¥2,22 1 , p 6; ¥2,22 p

0.394; crayfish ANOVA, ¥2^2 = 0.74, p = 0.488). Moreover, ANOVA failed to detect a significant effect of assemblage species composition on prey selectivity for golden

shiners {¥2,22 = 1-97, p = 0.163), redear sunfish {¥2,22 = 2.43, p = 0.1 1 1), or crayfish (¥2,22

= 3.06, p = 0.067).

In contrast, predator density and species composition did effect the absolute

mortality rates of prey (Figure 1 6). Largemouth bass, peacock cichlids, and the mixed

combination had similar odds of eating redear sunfish (ANOVA, ¥2,22 = 1 -90, p = 0. 1 74),

= but not golden shiners (ANOVA, ¥2,22 = 4.80, p = 0.019) or crayfish (ANOVA, ¥2^2

3.70, p = 0.041 ). In both cases, largemouth bass had greater odds of eating these prey than did peacock cichlids. The 50% predator combinations were intermediate between, and not statistically different from, the single species combinations. 97

Golden Shiner A Redear Sunfish Crayfish 0.8 • •

0.6 •f '.5 5 O 0) 0) 0.4- (0

0.2

Predator Treatment

Figure 15. Mean selectivity values (i.e., a) (± 1 SE) for golden shiners, redear sunfish, and crayfish by assemblages of largemouth bass (LMB) and peacock cichlid (PC)

predators varying in predator density (i.e., four, six, and eight predators) and combination

(i.e., 100 % PC, 50 % PC: 50 % LMB, and 100 % LMB).

Increasing density of predators increased the odds of mortality of golden shiners

(ANOVA, ¥222= 3.71, p = 0.041) and redear sunfish (ANOVA, F2,22= 9.20, p = 0.001).

Indeed, at high predator densities, fish prey were largely depleted (i.e., usually < 5% of

original numbers remaining). However, predator density had little effect on crayfish

mortality (ANOVA, F2,22 = 1 -22, p = 0.3 14). Because peacock cichlids ate few or no

crayfish during the experiment, an analysis was conducted using only data for treatments

containing largemouth bass. The analysis revealed that largemouth bass density did not have a significant effect on per capita consumption of crayfish (ANOVA, F4J3 = 0.35, p =

0.842), implying that individual largemouth bass exerted a relatively constant effect on

crayfish mortality regardless of density (Figure 17). 98

100 A ^ $ • 2 80 A 60 • A

\ 40. Q.

20 • « 4 * ^

N

Figure 1 6. Mean predation mortality (%) (± 1 SE) for golden shiners, redear sunfish, and crayfish by assemblages of largemouth bass (LMB) and peacock cichlid (PC) predators varying in predator density (i.e., four, six, and eight predators) and combination (i.e., 100 % PC, 50 % PC: 50 % LMB, and 100 % LMB).

a5 (3

Lai3 4 6 Number of Largemouth Bass

Figure 1 7. Mean per capita consumption (± 1 SE) of crayfish by largemouth bass in a 3- day tank experiment. Golden shiners and redear sunfish prey were also consumed during the experiment. 99

Discussion

One of the implicit and seldom tested assumptions of ecological studies is the premise that morphological similarity between species leads to similarity in food resource use (Harris 1995; Wainwright and Richard 1995) and therefore increased likelihood that two species will compete. This assumption is often invoked in many discussions of

introduced species (i.e., introduced and native species that are similar in morphology are likely to compete for resources [e.g., Matthews et al. 1992; Huckins et al.2000]).

Overall, the results of this study support the hypotheses that morphologically similar species have similar general patterns of prey handling and prey selection.

Similar prey handling ability of largemouth bass and peacock cichlids was most evident with bluegills, a species representative of spiny-finned, laterally-compressed fish prey. Southeast Florida canals have several abundant centrarchid (e.g., redear sunfish

and spotted sunfish Lepomis punctatus ) and cichlid (e.g., jewel cichlid Hemichromis letoumeauxi and spotted tilapia) prey species that are similar to bluegill in morphology

(Hill 1998; Shafland 1999a) and that are commonly consumed by largemouth bass and

peacock cichlids (Shafland 1999b; L. G. Nico, USGS, and J. E. Hill, unpubl. data). This category includes spotted tilapia, an abundant nonindigenous fish that was the primary target of biological control efforts that were part of the management plan that resulted in the peacock cichlid introduction (Shafland 1995).

Although there was support for the assumption of similar prey use by similar predators, potentially important differences were evident between largemouth bass and peacock cichlids in prey profitability and in the probability of successfully consuming

(i.e., probability of attack x probability of success given an attack) encountered sailfin 100

mollies and crayfish. Peacock cichlids were better than largemouth bass at preying on

sailfin mollies (i.e., they incurred lower handling times and had higher capture success).

However, sailfin mollies also were the most profitable prey for largemouth bass and the significance of this advantage for peacock cichlids is unknown. In contrast, peacock cichlids rarely (or never) consumed crayfish, providing largemouth bass with exclusive use of this resource. The ability of largemouth bass to exploit non-fish prey could be an important refuge against potential competifion with introduced peacock cichlids.

Hypothefically, the lack of crayfish in peacock cichlid stomachs sampled from

canals (e.g., Shafland 1999b) could be due to low (or zero) encounter rates between crayfish and peacock cichlids. For example, peacock cichlids are diurnal predators (e.g.,

Lilyestrom and Churchill 1996) and crayfish are crepuscular and nocturnal (Pennak

1989), thus limiting potential encounters. However, my results show that even when

crayfish are present and encountered, peacock cichlids will not consume them. It is possible that crayfish are extremely difficult for peacock cichlids to subdue and eat or that peacock cichlids do not normally recognize crayfish as potential prey. These hypotheses are not distinguishable given the data gathered in this study. However, peacock cichlids (mostiy juveniles) in Florida regularly consume another decapod

crustacean, the grass shrimp Palaemonetes sp. (Shafland 1999b; Chapter 4; L. G. Nico,

USGS, and J. E. Hill, unpubl. data). Nevertheless, grass shrimp, in contrast to crayfish, lack strong chelae, possibly making them more vulnerable to peacock cichlids.

Additionally, studies of various Cichla species outside of Florida have documented consumption of decapod crustacean prey (again mostly Palaemonetes spp.; Lowe-

McConnell 1969; Zaret 1980a; Jepsen et al. 1997), although adult Cichla are strongly 101

piscivorous (Lowe-McConnell 1969; Lilyestrom and Churchill 1996; Jepsen et al. 1997;

Winemiller et al. 1997).

Morphologically similar species are often expected to have similar effects on the

between prey base (e.g., Hambright et al. 1 991 ; Harris 1 995). However, a disagreement predicted diets based on experimental data (present study) and published diets of

largemouth bass and peacock cichlids in southeast Florida canals (Shafland 1 999b) implies that this assumption may be too simplistic (i.e., other factors may influence prey use and result in differing observed diets in similar species). Fish prey were more profitable and readily captured and consumed by largemouth bass than were crayfish.

Additionally, the present study documented avoidance of crayfish (i.e., acp « l/wj) when largemouth bass were presented with alternative prey (i.e., fish). Conversely, Shafland

(1999b) found that largemouth bass in a southeast Florida canal ate 45% (by number) decapod crustaceans, mostly fi-eshwater prawns Macrobrachium sp., but also crayfish.

The inconsistency between experimental and field data may be due to the relationship of

encounter rates between largemouth bass and their potential fish and non-fish prey (i.e., encounter rates with fish are low enough that largemouth bass include non-fish prey in

their diet). Encounter rates for fish versus non-fish prey could be modified by competition, predator hunting behavior, prey anti-predator behavior, or habitat structural

complexity. Additional research is needed to test these alternative (but not mutually exclusive) hypotheses.

The heavy use of non-fish prey by largemouth bass in the canals raises concern over the possible effects of food competition with peacock cichlids, the other prominent

predator in the system. Indeed, a shift in diet (i.e., induced resource partitioning) is a 102

prediction of competition theory (Sale 1979). Both species regularly consume fish

(Shafland 1995, 1999b; Chapter 4; L. G. Nico, USGS, and J. E. Hill, unpubl. data) and

could potentially deplete fish prey to the point of forcing a shift in diet. However, the

high abundance of prey fish (numerical and biomass) in many southeast Florida canals

(Shafland et al. 1985; Shafland 1995, 1999a; L. G. Nico, USGS, unpubl. data; pers. obs.)

suggests that fish prey depletion is not likely. Additionally, due to the interconnecfivity

of the canal system, large-bodied fishes such as largemouth bass could move to areas

with higher prey abundance if fish prey become locally depleted. The presence of

peacock cichlids also might reduce the encounter rate of largemouth bass with fish prey

through interference competition. Nevertheless, Experiment 3 documented largemouth

bass effectively preying upon fish (and selecting fish at a much higher rate than crayfish)

in the presence of peacock cichlids, suggesting that interference is not the causal

mechanism. Importantly, population data also suggest that competition with peacock

cichlids has not led to significant reductions in largemouth bass abundance in Florida

canals (Shafland 1999a, c; Shafland et al. 2001 ; Chapter 6).

Characterisfics of the predator, prey, and habitat influence prey use and might be

responsible for the heavy reliance of largemouth bass on crayfish prey. For example,

predator and prey behavior can modify encounter rates (Cooper et al. 1985; Magnhagen

1986; Grant and Noakes 1987; Winemiller and Taylor 1987). In canals, largemouth bass

hunting behavior may increase encounters with crayfish by hunting where crayfish occur

(i.e., in structurally complex habitats) and when they are active (i.e., low-light

conditions). Moreover, prey fish may take refuge fi-om predators in structurally complex

habitats. Habitat structural complexity (e.g., aquatic macrophytes) reduces hunting 103

efficiency of predators and can lead to reduced encounters with fish relative to non-fish prey (Savino and Stein 1982; Anderson 1984; Hayse and Wissing 1996). Several studies have documented significant consumption of decapod crustaceans by largemouth bass in

vegetated (i.e., structurally complex) systems in Florida (Chew 1974; Schramm and

Maceina 1986; Cailteux et al. 1996; Huskey and Tumigan 2001; Wheeler and Allen

2003). Cailteux et al. (1996) compared the diets of largemouth bass in vegetated and non-vegetated lakes in central Florida and found that diets were dominated by decapod crustaceans in vegetated lakes but by fish in non-vegetated lakes. They suggested that the structural complexity of aquatic macrophytes was responsible for the diet difference.

The southeast Florida canals often have extensive and dense beds of aquafic macrophytes

(Chapter 2), potentially an important factor in the observed pattern of significant use of non-fish prey by largemouth bass. The effect of macrophytes on largemouth bass diets in southeast Florida canals could be determined by comparing largemouth bass stomach contents fi-om areas with macrophytes with those from areas without macrophytes (i.e.,

harvested areas; Chapter 2).

The present study generally agrees with previous studies that have found differences in prey vulnerability that are correlated to prey defenses (e.g., Webb 1986;

Hoyle and Keast 1 987). The three prey types used in the experiments were chosen to present the predators with prey of differing morphology. Crayfish possess formidable morphological defenses—large chelae and a strong exoskeleton. Bluegills are relafively deep-bodied and possess spines in the dorsal and anal fins. However, sailfin mollies have few morphological attributes that could be considered defensive against subadult and adult largemouth bass or peacock cichlids. Such defenses do not provide absolute 104

predation refuges for prey (except when body depth exceeds mouth size for gape-hmited

predators [e.g., Lawrence 1958; Hambright et al. 1991]), but often increase handling times and the probability of escape (Gillen et al. 1981; Webb 1986; Wahl and Stein 1988;

Nilsson and Bronmark 2000). Gape-limited predators typically swallow fish prey headfirst and decapod crustacean prey tail-first to avoid interference fi-om spines or claws

(e.g., Lawrence 1 958). This often requires the predator to reorient prey prior to swallowing. To accomplish this, the predator may forcibly release the prey and then rapidly suck the item back into its mouth. Furthermore, prey with defensive morphology

may struggle ft-ee from the predator's mouth. In either case, released prey may escape on

its own (e.g., reftige in cover) or other predators may interfere and steal the prey item

(i.e., kleptoparasifism) or facilitate its escape (e.g., Nilsson and Bronmark 2000).

Therefore, attacking large and defensive prey has a potential energetic penalty of increased handling time and potential loss of prey due to escape and therefore defensive prey may be selected by predators at a lower rate than less defensive prey.

For largemouth bass, crayfish were less likely than fish to be attacked and more likely to escape if attacked. Largemouth bass also selected crayfish at a low rate relative to fish prey. However, the spines and deep body of bluegills did not confer an advantage over the relatively defenseless sailfm molly if largemouth bass were predators. The lack of difference differences in selection between sailfin mollies and bluegill for largemouth bass may be due to the relatively small size of bluegill prey in the experiment. Larger bluegills, with their more developed spines and deeper bodies, were less profitable prey

than were small bluegills (Experiment 1 ) and thus bluegill size may have limited the effecdveness of their morphological defenses. However, bluegills were less profitable 105

prey than were sailfin mollies for peacock cichlids, possibly due to the morphological

defenses of bluegills. This translated into higher selection values for sailfin mollies in

Experiment 2. This pattern, including a general refusal by peacock cichlids to eat highly

defensive crayfish, suggests that prey morphological defenses may be relatively more

effective against peacock cichlids than largemouth bass.

In the present study, largemouth bass ate more prey per capita than did peacock

cichlids. This is consistent with lower stomach fullness estimates for juvenile peacock

cichlids collected in Florida canals (Chapter 4). Similarly, relatively low daily rations

have been reported for a peacock cichlid congener, (Rabelo and

Araujo-Lima 2002). In the present study, predator length was not a factor because

individuals were chosen to obtain similar mean and extreme lengths. Predator mass is a

potential factor, but largemouth bass are only slightly heavier for their length than are

peacock cichlids (unpubl. data). It is possible that largemouth bass are under less stress

in the confines of an experimental tank than are peacock cichlids and therefore feed better

(see Schreck et al. 1997). Peacock cichlids in tanks are easily disturbed whereas

largemouth bass quickly become accustomed to the presence of people near their tanks

(pers. obs.). Additionally, peacock cichlids seem to be prone to disease (e.g.,

Ichthyophthirius multifiliis and opportunisfic bacteria) and shedding of fin epithelium

when first collected from canals and placed into a tank (pers. obs.). The outbreak of

disease in fishes is often stress-related (Wedemeyer 1 997) and confinement stress

correlates with epithelial loss in fishes (e.g., striped bass Morone saxatilis ; Noga et al.

1998). However, the peacock cichlids used in the experiments responded well to

appropriate prophylactic and veterinary treatments and began feeding consistently within 106

one or two days. Another possibility is that largemouth bass may have eaten more prey

simply because they were better at capturing it. However, this hypothesis is not

supported by the estimates of the probability of attack and success obtained in

Experiment 1 which showed that peacock cichlids were equally or more successful in

prey capture (Table 4). Conversely, peacock cichlids may have lower energy

expenditures than largemouth bass, maintaining body mass and growth with fewer

resources than an equivalent largemouth bass. This would likely translate into the ability

of peacock cichlids to sustain positive population growth at lower resource levels than

would be possible for largemouth bass. Given this case, peacock cichlids would be

favored during any resource crunch period. Nevertheless, additional data (e.g.,

bioenergetics modeling, growth studies) are needed to test this hypothesis.

A relative lack of effects of predator assemblage characteristics on prey selection

suggested similar effects of both predators on experimental prey assemblages. Similarity

of predators has been found in several experimental studies. For example, salamanders

(Notophthalmus viridescens and Ambystoma opacum ) (Morin 1995) and a salamander

(Notophthalmus viridescens) and a fish (banded sunfish Enneacanthus obesus) (Kurzava

and Morin 1998) had similar predatory effects on tadpole assemblages. Baber and

Babbitt (2003) also found fiinctional similarity among three fish species (eastern

mosquitofish Gambusia holbrooki , flagfish Jordanella floridae, and golden topminnow

Fundulus chrysotus ) preying on tadpoles; nevertheless, they found a significant

difference between the effect of these predators and the morphologically and behaviorally

different walking catfish Clarias batrachus . On the other hand, largemouth bass and peacock cichlids, although morphologically similar, expressed some important 107

differences in their effects on prey. The previously mentioned greater individual prey consumption of largemouth bass resulted in a greater effect of largemouth bass on the odds of golden shiner and crayfish death. As found in Experiment 2, the main difference between the predators was in the consumption of crayfish.

In conclusion, the present study provides some support for the assumptions that predators with similar morphology will have similar ability to handle prey and that this ability will lead to similar patterns of prey selection. Therefore, using trophic groupings based on morphology may be a valid methodology for some ecological studies (Briand and Cohen 1984; Harris 1995; but see Paine 1992; Yodzis and Winemiller 1999 for problems with this approach). Moreover, similarity in morphology with native species could be predictive of the diets of nonindigenous species and therefore the potential for dietary overlap and food competition. However, the predictions of experimental data

(i.e., that largemouth bass and peacock cichlids would have similar, fish-dominated diets) were not congruent with published field data on largemouth bass diets in canals (i.e., comparable proportions of fish and non-fish prey) and implied that the expectation that morphological similarity will yield similar effects on prey in the field may be too simplistic for complex systems. Thus, the incorporation of ecological interactions, species-specific behavior, and habitat influences may be required for better predictive ability. For example, some altemafive hypotheses were presented in this discussion that

may explain the differences in observed diets (i.e., in canals) in the context of prey encounter rates, namely competition, predator or prey behavior, and habitat

characteristics. Of these altemafives, the presence of aquatic macrophytes (i.e., habitat 108

structural complexity) was perhaps the simplest explanation, although the present study did not specifically attempt to refute one or more of these hypotheses.

Additionally, the present study suggested that prey morphological defenses (e.g., spines, chelae, and hard exoskeleton) may be relatively more effective against peacock cichlids than largemouth bass (in terms of influencing prey selection) but that peacock cichlids may be more efficient users of food resources than are largemouth bass.

Although peacock cichlids might be compefifively superior to largemouth bass if peacock cichlids are indeed more efficient predators; the ability of largemouth bass to utilize non- fish prey (along with apparently high prey abundance) would represent a competition refiige fi-om peacock cichlids. Indeed, population data imply that competition has not had a significant negative effect on largemouth bass abundance (Shafland 1999a, c; Shafland

etal. 2001; Chapter 6). CHAPTER 6 THE INTRODUCTION OF PEACOCK CICHLIDS INTO SOUTHEAST FLORIDA: POPULATION-LEVEL CONSEQUENCES FOR LARGEMOUTH BASS

Introduction

The introduction of predatory fishes outside their natural ranges often causes substantial concern for biologists and fishery managers. Nonindigenous predatory fishes are considered likely to cause negative changes in receiving systems due to predation or

competition effects on native species (Taylor et al. 1984; Moyle et al. 1986; Moyle and

Light 1996a). Recent introductions of predatory snakeheads (Channidae) and Asian swamp eel Monopterus albus into North America, as well as the Esox lucius eradication efforts in , have captured the attention of the media and public. Introductions of predatory Nile perch Lates niloticus into Lake Victoria, Africa

(e.g., Hughes 1986; Kauftnan 1992; Goldschmidt et al. 1993), and peacock cichlids

Cichla sp. into Lake Gatun, Panama (Zaret and Paine 1973), are frequently cited to demonstrate the danger that predatory fishes pose when introduced into new regions.

The predatory South American peacock cichlid Cichla ocellaris was intentionally introduced into freshwater canals of southeast Florida during 1984-1987 by the Florida

Game and Fresh Water Fish Commission (currently Florida Fish and Wildlife

Conservation Commission [FWC]) (Shafland 1995). The purposes of the introduction were to control excessive prey fish biomass, primarily resulting from the abundance of

the exotic spotted tilapia Tilapia mariae , and to supplement urban recreational fisheries

109 110

(Shafland 1995). The largemouth bass Micropterus salmoides , an ecomorphologically

similar native species (Norton and Brainerd 1993), was already present in these canals.

Both species currently support valuable recreational fisheries in southeast Florida

(Shafland and Stanford 1999).

There is continuing speculation that peacock cichlids negatively affect largemouth bass populations (Courtenay and Robins 1989; OTA 1993; Cox 1999; see also Lachner et

al. 1970). Indeed, there is evidence that peacock cichlids exert some negative pressures

on largemouth bass. Peacock cichlids consume small largemouth bass (Shafland 1999b;

unpublished data). For example, Shafland (1999b) documented that largemouth bass

occurred in 10% of 156 peacock cichlid stomachs from Tamiami Canal and made up 6%

by number of prey. Peacock cichlids and largemouth bass also have similar feeding

mechanics (Norton and Brainerd 1993), length-specific gape size (Hill 1998; Chapter 4),

and dietary patterns (Lilyestrom and Churchill 1996; Shafland 1999b; Chapter 4, 5).

Therefore, peacock cichlids might negatively effect largemouth bass populations via

resource competition. Conversely, existing data suggest that peacock cichlids have not

affected largemouth bass populations in southeast Florida (Shafland 1995, 1999a, b, c;

Shafland and Stanford 1999; Shafland et al. 2001). Additionally, there is no evidence for

deleterious effects of peacock cichlids on largemouth bass populations in other regions

where peacock cichlids and largemouth bass have been introduced (i.e., Hawaii and

Puerto Rico) (Devick 1980; Lilyestrom et al. 1994; Lilyestrom and Churchill 1996).

Courtenay (1997) has argued that the absence of long-term, quantitative data prior

to fish introductions explains the substantial lack of documentation of negative effects of

introduced fishes, especially in Florida. There are no data for southeast Florida canals Ill

covering the period prior to the initial exotic fish invasions (i.e., 1950s and 1960s; Hill

2002), and relatively few data for the period before peacock cichlids were introduced.

Therefore, it is difficult to compare largemouth bass populations before and after the peacock cichlid introduction, in sites with and without peacock cichlids, a powerful approach that could have been used to conduct a more conclusive assessment (e.g.,

Huckins et al. 2000). However, an analysis using the pre-introduction data on largemouth bass densities does not support the hypothesis of negative effects of peacock cichlids on largemouth bass populations in southeast Florida (Shafland 1999a).

Underlying questions about effects of peacock cichlids on largemouth bass is the assumption that south Florida urban canals are favorable habitat for largemouth bass and native fishes in general. If these canals represent marginal habitat (e.g., supporting only sink populations; Armett 1998) then potential effects of peacock cichlids may be relatively unimportant for the regional dynamics of largemouth bass. Based on canal morphometry, variable water physico-chemical parameters, frequent changes in aquatic macrophyte abundance due to control efforts, and presumably strong negative biotic interactions with numerous species of abundant exotic fishes, several authors have argued that urban canals are poor habitat for native fishes (Hogg 1976; Loftus and Kushlan 1987;

Courtenay 1997; Annett 1998.). Annett (1998) suggested that the canals represent population sinks for largemouth bass, given a lack of observed spawning within the canals. Conversely, the perception of marginal largemouth bass populations was not supported by my own qualitative observations consisting of hundreds of hours of field

work (e.g., electrofishing) and observations in the canal system. For example, it is clear that largemouth bass spawn and successfully recruit in southeast Florida canals (Chapter 112

of 4; pers. obs.). Indeed, I was surprised by the range of sizes and the abundance largemouth bass in these canals.

I used available data (i.e., FWC database; FWC 2001) to test the hypotheses that

in 1) largemouth bass population densities in the canals were similar to densities found

other Florida systems (i.e., Florida lakes and streams) and 2) largemouth bass population

densities in canals have been negatively effected by the introduction of peacock cichlids.

Shafland (1999a) provided data on Black Creek Canal before, during, and after the

introduction of peacock cichlids (see also Shafland et al. 1985; Shafland et al. 2001). The

present statistical analysis is the first to compare canals with statewide data for other

types of systems (i.e., lakes and streams) and to compare canals containing peacock

cichlids with canals that lack peacock cichlids. If canal populations of largemouth bass

are similar to other systems in Florida, then this result would question the common

assumption that urban canals in south Florida are poor largemouth bass habitat.

Moreover, a significant reduction in abundance of largemouth bass due to peacock

cichlids would be unlikely.

Methods

I used the FWC database (FWC 2001) which provides data from statewide fish

populafion sampling conducted by FWC personnel from 1 956 to 2000 (mostly in the

1980s and 1990s). The FWC database contains blocknet data and electrofishing data.

The blocknet data set contained 790 observations (i.e., one or more blocknets [using

either rotenone or detonator cord] at a site) for 71 water bodies and the electrofishing data

set contained 934 observadons (i.e., one or more boat electrofishing transects at a site) for

300 Florida water bodies. A type I error rate of a = 0.05 was used for all analyses. I used 113

tests that did a type II error rate of (3 0.20 to estimate detectable effect sizes (i.e., A) for not yield significant differences (Ott 1993).

The present study focused on harvestable largemouth bass (i.e., > 254 mm total length [TL]) because electrofishing is not the best method for estimating the abundance

of age-0 fish (i.e., age-0 largemouth bass may not be fully recruited to the gear)

(Reynolds 1996) and harvestable fish are of immediate fishery importance. Indeed, if

harvestable-size largemouth bass abundance is maintained over time, recruitment is

confirmed in the system. Most comparisons in the present study used electrofishing

catch-per-unit-effort (CPUE) data. This method is appropriate for monitoring population

abundance of harvestable-size fish (Reyonlds 1996) and is a better method than is the use

of blocknets (Tate et al. 2003).

I combined all blocknet or electrofishing observations to create a mean value if a

water body had more than one observation; regional and statewide means were obtained

from these water body means (PROC MEANS; SAS Institute, Gary, North Carolina). I

used estimates of numerical abundance (fish/ha) and standing crop (kg/ha) from blocknets as well as numerical (fish/min) and biomass (g fish/min) electrofishing CPUE

for the analysis comparing canals to other Florida systems. Differences among statewide

blocknet or electrofishing CPUE data for water body types (i.e., Florida canals, lakes, and streams) were determined by one-way analysis of variance (ANOVA), using the general linear models procedure (PROC GLM) followed by the Tukey HSD multiple comparison procedure in SAS (SAS Institute, Cary, North Carolina).

I used electrofishing CPUE (i.e., numerical and biomass) to compare largemouth bass populations in south Florida canals with south Florida lakes and to compare canals 114

with and without peacock cichlids. I considered water bodies to be located in south

Florida if they were in or south of Charlotte, Glades, and Martin counties and Lake

Okeechobee. Differences in electrofishing CPUE data for south Florida canals versus south Florida lakes and for canals with and canals without peacock cichlids were tested by two-sample t-tests following a test for equality of variance (PROC TTEST; SAS

Institute, Gary, North Garolina).

Results

Based on electrofishing data, the density and biomass of harvestable largemouth bass varied among different types of water bodies sampled throughout Florida, with

GPUE approximately one and a half to two times higher in lakes than in canals and

streams (Figures 18 and 19). Harvestable largemouth bass density and biomass based on

blocknet data showed a slightly different pattern, with largemouth bass being

approximately 1 0-fold more abundant in canals and lakes compared to streams (Figures

20 and 21).

These comparisons based on statewide samples potentially confound differences

among types of water bodies with latitudinal effects (i.e., canals are more common in

southern Florida). I therefore compared south Florida canals with south Florida lakes (no

south Florida streams were included in the database), but found no significant differences

at the a priori a = 0.05 level (Figures 22 and 23). The analysis of electrofishing GPUE

data also failed to detect a significant difference between canals with and canals without peacock cichlids (Figures 24 and 25). .

115

0.6*

Florida Canals Florida Lakes Florida Streams

Figure 18. Mean (± SE) electrofishing CPUE (fish/min) for harvestable largemouth bass

(> 254 mm TL) in Florida canals (N = 43), lakes (N = 1 14), and streams (N = 21). Mean CPUE for Florida lakes was significantly greater than Florida canals and streams (ANOVA, F2,i75 = 8.90, p = 0.0002). Data were obtained from FWC 2001

Figure 19. Mean (± SE) electrofishing CPUE (biomass; g/min) for harvestable largemouth bass (> 254 mm TL) in Florida canals (N = 43), lakes (N = 114), and streams (N = 21). Mean CPUE for Florida lakes was significantly greater than Florida canals and streams (ANOVA, F2,i75= 6.52, p = 0.0019). Data were obtained from FWC 2001. 116

Florida Canals Florida Lakes Florida Streams

Figure 20. Mean (± SE) number (fish/ha) of harvestable largemouth bass (> 254 mm TL)

in Florida canals (N = 1 8), lakes (N = 39), and streams (N = 14) based on blocknet data. Mean CPUE for Florida canals and lakes were significantly greater than Florida streams (ANOVA, F2,68 = 5.57, p = 0.0057). Data were obtained from FWC 2001.

Florida Canals Florida Lakes Florida Streams

Figure 21. Mean (± SB) biomass (kg/ha) of harvestable largemouth bass (> 254 mm TL) in Florida canals (N = 1 8), lakes (N = 49), and streams (N = 14) based on blocknet data. Mean CPUE for Florida canals and lakes were significantly greater than Florida streams (ANOVA, F2,68 = 4.79, p = 0.01 13). Data were obtained fi-om FWC 2001. .

117

Figure 22. Mean (± SE) electrofishing CPUE (fish/min) for harvestable largemouth bass

= = 1 ). (> 254 mm TL) in south Florida canals (N 39) and lakes (N 1 Means were not significantly different (t-test, to.o5,4g= -1-83, p = 0.0734; A = 0.038 fish/min). Data were obtained from FWC 2001.

Figure 23. Mean (± SE) electrofishing CPUE (biomass; g/min) for harvestable largemouth bass (> 254 mm TL) in south Florida canals (N = 39) and lakes (N = 11).

Means were not significantly different (t-test, to.o5,48 = -0.99, p = 0.3262; A = 24.6 g/min). Data were obtained fi-om FWC 2001 . .

118

0.4 «

Canals with Peacock Cichlids Canals without Peacock Cichllds

Figure 24. Mean (± SE) electrofishing CPUE (fish/min) for harvestable largemouth bass (> 254 mm TL) in southeast Florida canals with (N = 27) and without (N = 13) peacock = cichlids. Means were not significantly different (t-test, to.os, 38 = 1 -33, p = 0. 1 92; A

0. 1 69 fish/min). Data were obtained from FWC 200 1

250 «

Canals with Peacock Cichllds Canals without Peacock Cichllds

Figure 25. Mean (± SE) electrofishing CPUE (biomass; g/min) (± 1 SE) for harvestable largemouth bass (> 254 mm TL) in southeast Florida canals with (N = 27) and without (N = = 13) peacock cichlids. Means were not significantly different (t-test, to.o5,38 = 0.57, p 0.575; A = 128.5 g/min). Data were obtained from FWC 2001 119

Discussion

The similarity of largemouth bass populations in Florida canals to other types of aquatic systems in Florida on both a statewide and regional (i.e., south Florida) basis is somewhat surprising given the general impression from the literature that urban canals are poor habitat for native fishes (e.g., Hogg 1976; Loftus and Kushlan 1987; Courtenay

1997; Annett 1998). In contrast, these results imply that canals are not poor habitats

overall for largemouth bass. In particular, there is little evidence that canal populations of largemouth bass are sink populations (Annett 1998). Indeed, largemouth bass successftilly spawn and recruit within the canal system, based on the presence of nesting fish, free-swimming fiy, and small juveniles (Chapter 4; pers. obs.).

Shafland and colleagues have presented additional data that support the hypothesis that canals in southeast Florida are not marginal habitats. Mean electrofishing

CPUE (fish/min) of harvestable largemouth bass in canals with peacock cichlids are

comparable to the mean statewide value used by FWC (Figure 26; see Shafland et al.

2001). Indeed, electrofishing data for several canals showed higher CPUE estimates for

harvestable largemouth bass (i.e., fish/min) than the statewide average (Shafland et al.

2001). Recreational sport fishing for largemouth bass in southeast Florida, where peacock cichlids also occur, has an estimated annual economic value of US$5 million

(Shafland and Stanford 1 999). Fishing tournaments in the canals yield catch rates that may exceed tournaments held in famous fishing waters (e.g., )

(Shafland et al. 2001). Recreational angler creel data collected for two large canals in

Miami-Dade County rank them both in the top 10 statewide for angler effort directed at largemouth bass and in angler success. Black Creek Canal was number one (182 120

hr/ha/yr) and Tamiami Canal was number five (48 hr/ha/yr) in angler effort directed at largemouth bass (Shafland and Stanford 1999). Tamiami Canal had the highest angler

success rate for largemouth bass (0.41 fish/hr) and Black Creek Canal was tenth (0.30

fish/hr) (Shafland and Stanford 1999). Clearly, persistent largemouth bass populations

exist in the canals regardless of negative abiotic or biotic influences.

1986-1989 1986-1999 1997-1999 Statewide Average

Figure 26. Mean electrofishing CPUE (fish/min) for harvestable largemouth bass in south Florida canals that have peacock cichlids. The data for 1986-1989 include 23 canals; the data for 1986-1999 include 34 canals; and the data for 1997-1999 include 32

canals. The data and the value for statewide average were obtained from Shafland et al. 2001.

Although electrofishing and blocknet data showed different trends in the

statewide habitat comparison, canal populations of largemouth bass were not

distinguishable from another Florida system (i.e., either lakes or streams) in both cases.

Tate et al. (2003) noted differences between electrofishing and blocknet data for

largemouth bass populations in two vegetated lakes in north Florida. It is therefore not

surprising to have somewhat conflicting trends in the larger and more heterogeneous 121

FWC database that encompasses the entire state. Also, it is possible that the physical nature of the canal systems in Florida, being intermediate between lakes and streams,

may be reflected in this inconsistent outcome. In contrast to the statewide electrofishing

data, no significant differences were found between south Florida canals and lakes. Both

tests (i.e., fish/min and g/min) could detect differences of about 10% with reasonable

power (i.e., 1 - P = 0.80) (Figures 22 and 23). These results suggest that regional

influences may be important for largemouth bass and that south Florida canals are not

marginal habitat.

Introductions of piscivores, especially species similar to native fishes, are

generally predicted to result in negative effects for the native species (Taylor et al. 1984;

Moyle et al. 1986; Moyle and Light 1996a). Hypothetically, peacock cichlids might

negatively influence an ecomorphologically similar native species such as the largemouth bass by predation or food competition. However, the results of the present study did not

support the hypothesis that the introduction of peacock cichlids has caused major

reductions in largemouth bass abundance.

Largemouth bass are successftil in south Florida canals despite the presence of

many nonindigenous species, including peacock cichlids. It is plausible that the density

of largemouth bass would be even greater if not for the possible deleterious effects of peacock cichlids. However, the available data provide no support for this hypothesis; largemouth bass abundance and biomass was not significantly different in canals with

and without peacock cichlids. However, an assumption of this comparison is that these

canals differ only in the presence or absence of peacock cichlids. This assumption is not

valid. The geographic distribution of these two types of canals is adjacent, but non- 122

overlapping. Although still within the same south Florida counties, canals without peacock cichlids are north and west of those with peacock cichlids. Therefore, geographic effects might confound peacock cichlid effects. In their fisheries survey of

south Florida canals, Shafland et al. (2001) reported a number of differences between the two types. Peacock cichlids are found in canals that are often more urban in character

(but there is wide variation), have a more box-cut morphometry, are deeper, have

narrower littoral zones, and have less shoreline vegetation than the canals lacking peacock cichlids. Canals in the area occupied by peacock cichlids also have greater abundance and variety of nonindigenous fishes. Interestingly, many of these factors

suggest that canals lacking peacock cichlids should support higher densities of largemouth bass than canals with peacock cichlids, independent of the actual presence or

effects of peacock cichlids. Indeed, Shafland et al. (2001) commented that the canals lacking peacock cichlids were better habitat for sunfish Lepomis spp. However, the analysis failed to distinguish between mean electro fishing CPUE of harvestable

largemouth bass for these systems. On the other hand, given reasonably high power (i.e.,

1 - P = 0.80), the detectable effect sizes of the analyses were > 50% difference in

abundance (i.e., fish/min) (Figure 24) and > 70% difference in biomass (i.e., g/min)

(Figure 25), rather large differences. Therefore, further investigation may be warranted concerning the relationship of largemouth bass populations in the two canal types.

The results of the present study suggest that peacock cichlids have not had major negative effects on largemouth bass populations. However, the present study does not provide data investigating possible changes in the abundance of largemouth bass in the canals since the peacock cichlid introduction. Appropriate data were lacking in the FWC 123

database to conduct such an analysis (FWC 2001). However, Shafland et al. (2001) summarized largemouth bass population data from blocknets for the pre-introduction

period (i.e., 1980-1984) and a transitional period when peacock cichlids had been introduced but were not yet considered established (i.e., 1984-1987), as well as data for canals with peacock cichlids and canals lacking peacock cichlids for the period of 1980-

1997 (i.e., covering four years of pre-introduction, three years of transitional period, and ten years of post-introduction data) (Figure 27). These data include both harvestable and sub-harvestable largemouth bass. Although error estimates were not presented, these data do not support a hypothesis of reduction in largemouth bass abundance in response to effects of peacock cichlids.

Figure 27. Mean biomass (kg/ha) of largemouth bass in south Florida canals based on blocknet data. The pre-introduction period includes data from 1980-1984 (N = 22 canals). The transitional period includes data from 1984-1987 (N = 22 canals), the time period when peacock cichlids were introduced but not yet established. Canals with peacock cichlids includes data from 1980-1997 (N = 28 canals). Canals lacking peacock cichlids includes data from 1980-1997 (N = 1 1 canals). Data obtained from Shafland et al. 2001. 124

Despite anecdotal reports and theoretical predictions suggesting negative effects of peacock cichlids in Florida (Courtenay and Robins 1989; OTA 1993; Cox 1999; see

also Lachner et al. 1970), no published data support the hypothesis that peacock cichlids have significantly reduced largemouth bass populations. For example, a major governmental report on nonindigenous species cited a non-peer reviewed environmental

magazine (i.e., Florida Environments; McClelland 1992) as the source for a conclusion that peacock cichlids were "slowly reducing populations of indigenous bass and bream"

(p. 263; OTA 1993). The cited article (i.e., McClelland 1992) presented no data, only a speculative statement not attributed to any source. Similarly, Cox (1999) listed the peacock cichlid in Florida as an "invasive exotic", a designation that strongly implies that

the peacock cichlid is a harmful species. However, only a brief mention of peacock cichlid predation on native fishes and invertebrates was given as the justification (Cox

1999).

The lack of detectable effect may be due to many factors. The interaction

strength of these mechanisms (i.e., predation and competition), and thus the magnitude of the effect, may be low. Additionally, abiotic factors or other biotic interactions, including intraspecific interactions, may be preeminent. Such dynamics may keep the abundance of predatory fishes below the level where competitive interactions between them become important. For example, Zom and Seelbach (1995) concluded that abiotic or biotic processes, especially prior abiotic events that reduce recruitment, were responsible for the lack of a positive linear relationship between predatory fish biomass and available fish habitat in warm water streams. Furthermore, a nonindigenous fish 125

could integrate into the existing fish assemblage under non-equilibrium conditions, if the

system is not saturated with species, or if there are underexploited resources.

The common knowledge, often repeated in the literature, that south Florida canals are uniformly poor habitat for native fishes and that nonindigenous fishes therein have

significantly reduced native species abundance was not supported by the data for

largemouth bass. Other recent studies also have quesfioned the percepfion of across-the- board reduced native fish populations in canals, as well as the extent of negative effects that nonindigenous fishes exert in south Florida. Recent fisheries studies conducted by

the FWC in south Florida canals document the persistence and abundance of native

fishes, including largemouth bass, bluegill Lepomis macrochirus , and redear sunfish

Lepomis microlophus (Shafland 1999a, c; Shafland and Stanford 1999; Shafland et al.

2001). Indeed, Shafland (1996) argued that few data document negative effects of

nonindigenous fishes in south Florida and urged a scientific evaluation of data to replace

the anecdotal reports that form the usual basis of discussion on this topic. Moreover, in

an analysis of the most comprehensive and long-term dataset for south Florida, covering

the greater Everglades system and adjacent canals, Trexler et al. (2000) concluded that

nonindigenous fishes had not significantly affected native fishes, especially in the less

disturbed habitats.

Clearly, the interactions of native and nonindigenous fishes in this region require

a great deal of additional research. One priority is the documentation of the influence of nonindigenous fishes on the aquatic systems of south Florida (Shafland 1996; Courtenay

1997). Of particular interest, however, is the fact that nearly all of the available

information on the effects of nonindigenous fishes in south Florida is contrary to the 126

current paradigm of invasion biology. Much of the ecological theory adopted by the field of invasion biology (see Vermeij 1996; Williamson 1996), particularly with respect to

freshwater systems, is based on island biogeography and tightly-structured, stable

communities (sensu Clements 1936) (Stauffer 1984; Li and Moyle 1999). The

overwhelming consensus based on this theoretical background is that species

introductions lead to significant negative consequences for native species and ecosystems

It (Chapter 1 ). Finding expected results merely reinforces existing dogma. is with the

study of systems that yield unexpected and contrary results that this field will advance

and produce a mature, pluralistic, and ultimately more useful theory. CHAPTER 7 SUMMARY

Florida has numerous established and reproducing nonindigenous fishes (Fuller et

al. 1999; Hill 2002). Such nonindigenous fishes represent an important challenge to fishery managers and environmental regulators in Florida and throughout the USA

(Magnuson 1976; Hill 2002). Clearly, some introduced fishes in the USA have become

invasive; yet others have had relatively little detectable effect on their receiving system

(Moyle and Light 1996a).

The intentional introduction of the peacock cichlid Cichla ocellaris into the native range of the ecomorphologically similar largemouth bass Micropterus salmoides in southeast Florida was made to increase recreational fishing opportunities and to exert biological control over other nonindigenous fishes (Shafland 1995). Nevertheless, purposeful introductions are controversial and concern has been expressed over the possible negative effects of peacock cichlids in southeast Florida (Courtenay and Robins

1989; OTA 1993; Courtenay 1994; Cox 1999). Indeed, introduced piscivores are considered likely to cause deleterious effects in their receiving systems (Moyle and Light

1996a). Moreover, morphologically similar species should use similar resources and thus be more prone to compete with one another.

Concern over the potential effects of nonindigenous fishes has prompted increasing regulation, including the listing or proposed listing of additional fish species as

injurious wildlife by the United States Fish and Wildlife Service (USFWS) (e.g.,

127 128

al. snakeheads Channa spp. and black carp Mvlopharyngodon piceus ; Nico et 2001).

Risk assessments are used to provide information and assist in decision-making regarding possible regulation of nonindigenous species. The Generic Nonindigenous Aquatic

Organism Risk Analysis Review Process (i.e., RAM methodology; RAM Committee

1998) is being used by the USFWS as their risk assessment method. The black carp risk assessment was a test case for the application of the methodology (Nico et al. 2001); however, there have been no previous tests of the predictions of the RAM method algorithms. My application of the RAM method using Cichla as a test case elucidated

shortcomings of the method (Chapter 3). In particular, I pointed out that the precautionary principle (FAO 1995) is implicit in the RAM algorithm (i.e., the presupposition that any nonindigenous species represents an unacceptable risk if

established or if establishment is of at least medium probability). Given this approach, the risk assessment outcome depends solely on the estimate of the Probability of

Establishment (i.e., disregards the risks estimated under the Consequences of

Establishment). Thus, any species of medium or greater Probability of Establishment has

at least a medium Overall Risk Potential (ORP) (i.e., unacceptable risk requiring mitigation; RAM Committee 1998). Furthermore, the RAM methodology confounds the

role of the risk assessment (i.e., to estimate risk) with that of society (i.e., to determine

the acceptable level of risk). Additionally, 1 recommended that the ORP be deemphasized in favor of retaining the Probability of Establishment (and Spread) and the

Consequences of Establishment as the primary model outputs.

Part of the concern over the introduction of peacock cichlids into southeast

Florida stems from the hypothesis that similarity in morphology leads to similarity in 129

resource use (and therefore potential competition). 1 examined this issue for juvenile largemouth bass and peacock cichlids by describing their gape width relations under the

assumption that trophic morphology equates to prey handling ability (Wainwright and

Richard 1995), by comparing ontogenetic shifts in diet, and by estimating the degree of

overlap in prey use between the two species (Chapter 4). Furthermore, I tested this

hypothesis (i.e., similar morphology equals similar resource use) by documenting prey handling ability and prey selectivity of subadult and adult largemouth bass and peacock

cichlids under experimental conditions (Chapter 5). I found considerable evidence for

similarity in prey use by largemouth bass and peacock cichlids and thus support for the

hypothesis of similar morphology equals similar resource use. Juveniles had similar gape

width relations and ontogenetic dietary shifts, and overlapped considerably in diet.

Larger individuals had similar prey handling ability and prey selection patterns.

Additionally, the results suggested that largemouth bass and peacock cichlids have

similar effects on the prey base. Nevertheless, I documented several important

differences. In particular, the results suggested that largemouth bass are more capable of

consuming highly defensive prey (e.g., crayfish) and that peacock cichlids are better at

feeding on relatively defenseless prey (e.g., sailfm molly). Indeed, although largemouth

bass generally avoided crayfish prey if alternative, more profitable prey (i.e., fish) were

present, largemouth bass had exclusive use of crayfish as a food resource. Furthermore,

largemouth bass consistently consumed more prey than did peacock cichlids of equal

size.

Given the overall similarity in prey handling ability and prey selection under

experimental conditions (Chapter 5), the differences in diets between largemouth bass 130

and peacock cichlids in Tamiami Canal were surprising (i.e., non-fish prey making up about five times more [as a percentage by number] of the largemouth bass diet compared to the peacock cichlid diet; Shafland 1999b). This difference is suggesfive of compefifion-induced resource partifioning (Sale 1979). Nevertheless, given the apparently high abundance of fish prey in southeast Florida canals, a simpler explanation might be differential encounter rates between prey types for largemouth bass due to

habitat structural complexity (i.e., aquatic macrophytes) leading to a broadening of diet

(Anderson 1984; Cailteux et al. 1996). This question would be a potentially fhiitful area for additional research.

Although some authors have suggested that peacock cichlids have had negative effects on largemouth bass and other native fishes (Courtenay and Robins 1989; OTA

1993; Cox 1999; see also Lachner et al. 1970), no population data have been presented that suggest any such effects. A partial explanation might be the relative lack of pre- introduction data, a common concern in evaluations of the effects of fish introductions

(Courtenay 1997). Shafland (1999a) presented pre- and post-introduction data for Black

Creek Canal and concluded that peacock cichlids had not significantly reduced any native

fish population. I used the Florida Fish and Wildlife Conservation Commission database

(FWC 2001) to further investigate this question by comparing populations of largemouth

bass in canals with populations in Florida lakes and streams (Chapter 6). Because of

potential latitudinal effects, I also compared canals to south Florida lakes. Additionally, I compared largemouth bass populations in canals that contain peacock cichlids to canals that lack peacock cichlids. My results suggest that canals are not marginal largemouth 131

bass habitat and do not support a hypothesis of a reduction in largemouth bass abundance due to peacock cichlids.

Although the direct effects of predation and food competition by introduced peacock cichlids hypothetically should have negative effects on largemouth bass populations, existing data do not suggest reductions in largemouth bass populations

(Shafland 1999a; Shafland et al. 2001; Chapter 6). Observations suggest that food resources (at least fish prey) may not be limiting for these predators. Additionally, indirect mechanisms, perhaps mediated through prey or human exploitation, may have facilitative effects that are as yet unknown (see Stone and Roberts 1991). Indeed, the

potential for positive indirect effects of species introductions is an intriguing area for fiiture research. Such effects may well be responsible for many instances when

nonindigenous species have little apparent effect on receiving communities.

It can be argued that the predominant theoretical view of invasion biology is based upon the Clementsian concept of organismic communities (Clements 1936), namely that communities are stable, highly-structured, biotically-controlled, and

persistent in time and space (Chapter 1). Under such a concept, species introductions are significant perturbations with effects rippling throughout entire communities. Of

particular concern is the prediction of increased rates of extinction concurrent with

increased rates of invasion (MacArthur and Wilson 1 967). On the other hand, many species introductions seem to occur with little or no detectable negative effect on other

species (Williamson 1 996). Such integration is extremely difficult to accommodate

under the prevailing view within invasion biology (see Stauffer 1 984; Vermij 1996;

Williamson 1996; Li and Moyle 1999). Most explanatory arguments attack the perceived 132

flaws of the studies that fail to document negative effects, point out a lack of pre- introduction data, or claim that sufficient time has not passed for the negative effects to

surface. Indeed, a common prediction is that a species that is not a problem today will become a problem eventually. Nevertheless, the results of this dissertation call into question the uncritical acceptance of the presupposition that nonindigenous fishes will invariably cause ecological harm.

Instead of adapting invasion biology in such a way as to accommodate contrary results, perhaps invasion biology should be expanded to include these results into a more pluralistic and comprehensive whole. For example, integration of invading species into receiving systems becomes a theoretical possibility if communities can be viewed as loosely-structured, primarily abiotically-controlled, stochastic, and transient in time and

space (i.e., individualistic concept; Gleason 1926). Given the considerable evidence that many North American fish communities are not saturated with species (Gido and Brown

1999), that species introductions frequently result in permanent increases in species richness (Vermeij 1991; Gido and Brown 1999; Rosenzweig 2001), and that community

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Jeffrey E. Hill was bom in 1 966 in Florence, Alabama, to Baker E. and Jacqueline

C. Hill. He became interested in science and the outdoors at an early age—his father

took him hunting, fishing, and catching minnows in the creek; his mother and siblings

read dinosaur books to him; Myrtle Alexander, a local biology teacher, taught him about

reptiles and mollusks. By high school, Jeffrey knew he wanted to pursue a career in

or fish ecology. He maintains a lifelong love of aquarium keeping, fishing, hunting, and the wonder of the outdoors.

Jeffrey graduated from Bradshaw High School, Florence, Alabama, listed in the

"Mark of Excellence" (top 5% of graduating class) and delivered the Farewell Speech for the Class of 1985. He attended the University of Florida from 1985 to 1988 in the

Undergraduate Honors Program as a National Merit Scholar. Jeffrey married Susan

Bemecker in 1989 and completed his Bachelor of Science degree at the University of

North Alabama in 1 991 with a major in biology and a minor in geography. He has worked in Florida in the private sector in mariculture research and owned and operated a tropical fish farm. Jeffrey earned a Master of Science degree in 1 998 in the Department of Fisheries and Aquatic Sciences, University of Florida. His thesis was titled "Estimate of gape limitation on forage size for the peacock cichlid, Cichla ocellaris . an exotic fish established in Florida." He entered the doctoral program in the Department of Fisheries and Aquafic Sciences in 1998 as a Dean's Fellow. Jeffrey's graduate awards include the

154 155

Roger Rottman Memorial Scholarship from the Florida Chapter of the American

Fisheries Society and the Department of Fisheries and Aquatic Sciences Graduate

Student of the Year Award. After graduation, Jeffrey will serve as associate research

faculty at the Tropical Aquaculture Laboratory of the University of Florida in Ruskin,

Florida. He plans to continue his career in research, teaching, and extension. I certify that I have read this study and that in my opinion it conforms to

acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of Doctor of Philosophy.

Charles E. Cichra, Chair Professor of Fisheries and Aquatic Sciences

I certify that I have read this study and that in my opinion it conforms to

acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of Doctor of Philosophy.

William J. Lindb^^ Associate Professor of Fisheries anc Aquatic Sciences

I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is flilly adequate, in scope and quality, as a dissertation for the degree of Doctor of Philosophy.

Carter R. Gilbert / Professor Emeritus of Zoology

I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of Doctor of Philosophy.

Leo (\. Nico ^ ^Research Biologist, U.S. Geological Survey I certify that I have read this study and that in my opinion it conforms to acceptable standards of scholarly presentation and is fully adequate, in scope and quality, as a dissertation for the degree of Doctor of Philosophy.

Craig W. Osenberg Professor of Zoology

This dissertation was submitted to the Graduate Faculty of the College of Agricultural and Life Sciences and to the Graduate School and was accepted as partial fulfillment of the requirements for the degree of Doctor of Philosophy.

December 2003

Dean, Graduate School