The effects of predatory fish on amphibian species richness and distribution

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S0006-3207(96)00

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Biological Conservation 79 (1997) 123 131 Copyright © 1996 Elsevier Science Limited Printed in Great Britain. All rights reserved 0006-3207/97 $17.00 + .00

ELSEVIER

THE EFFECTS OF P R E D A T O R Y FISH O N A M P H I B I A N SPECIES RICHNESS AND DISTRIBUTION Stephen J. Hecnar* & Robert T. M'Closkey Department of Biological Sciences, University of Windsor, Windsor, N9B 3P4, Ontario, Canada (Accepted 19 June 1996)

of water bodies increases, the importance of desiccation decreases and predation becomes the dominant mortality factor (Heyer et al., 1975; Wilbur, 1984). In permanent water bodies, fish are likely the most important predators (Petranka et al., 1987; Kats et al., 1988). Efficient fish predators can search out and easily capture amphibian larvae and are considered the only aquatic predators capable of complete elimination of some species' tadpoles (Heyer et al., 1975). Fish predators can reduce the abundance of amphibians (Macan, 1966; Sih et al., 1992), eliminate subpopulations or cause local extinctions (Burger, 1950; Petranka, 1983; Sexton & Phillips, 1986), and can alter distribution patterns (Macan, 1966; Petranka, 1983; Bradford et al., 1993; Brrnmark & Edenhamn, 1994). Many ecologists assume that predation is a major selective force restricting some amphibian species to temporary pond habitats or resulting in the evolution of adaptations that permit co-existence with predators in permanent water bodies. Much of the amphibian predation research has concerned anti-predator defenses in larvae such as chemical repellents (Voris & Bacon, 1966; Kruse & Francis, 1977; Kats et al., 1988), reduced mobility (Woodward, 1983), shifts in activity patterns (Taylor, 1983; Petranka et al., 1987), or avoidance of predators by use of chemical cues (Petranka et aL, 1987). Research also suggests that adults may alter habitat selection by not calling or ovipositing in water bodies that contain predatory fish (Resetarits & Wilbur, 1989, 1991; Kats & Sih, 1992; Hopey & Petranka, 1994). Lack of defensive adaptations is considered to be an important reason why temporary pond amphibian species cannot coexist with predatory fish in permanent water habitats (Kats et al., 1988). Predation is a complex interaction and not all fish can be considered to be important predators on amphibians (Petranka, 1983; Hayes & Jennings, 1986; Sexton & Phillips, 1986; Br6nmark & Edenhamn, 1994). Some species by virtue of their diet (e.g. planktivores, herbivores) or habitat use (deep water) would not be important predators. Of > 180 fish species found in Canada, amphibians were documented as diet items of only 13 species (Scott & Crossman, 1973). Fish are also

Abstract

Amphibian communities at 178 ponds across southwestern Ontario, Canada, were studied to determine if presence of predatory fish was related to altered amphibian species richness or distribution on a geographic scale. Ponds are an important amphibian habitat in the study area and many have been stocked with fish. Surveys conducted over three years were used to construct amphibian species lists for individual ponds. Species richness and presence/ absence were compared among ponds classified by the type offish present. Amphibian species richness was significantly lower at ponds having predatory fish present than at non-predatory, or fish-free, ponds. Not all amphibian species were negatively affected by the presence of predatory fish: those having either large bodies or clutch size co-occurred with predatory fish more frequently than those with small bodies or clutch size. Introduction of predatory fish by humans has likely resulted in altered amphibian species assemblages and reduced community diversity on a geographic scale. Copyright © 1996 Elsevier Science Limited Keywords: amphibian, fish, predation, species richness, distribution.

INTRODUCTION The role of predation in structuring populations and communities occupies a position of central interest in ecology. Predation may act to reduce abundance, alter distribution, or act as a major selective force on prey species (Begon et al., 1990). In communities, predation can act either to increase or decrease species diversity (Huston, 1994). In conservation biology, concern exists regarding detrimental effects that introduced exotic predators can have on native communities (Drake et al., 1989; Primack, 1993). Predation is also of interest and concern in amphibian ecology and conservation. Conceptual models of amphibian communities suggest that as the permanency *To whom correspondence should be addressed. 123

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S. J. Hecnar, R. T. M'Closkey

gape-limited (Zaret, 1980) and size-structured (Stein et al., 1988) predators, which suggests that some fish species may not be large enough to be efficient predators of amphibians. Risk of predation in amphibians is sizedependent, and larvae can outgrow their predators (Wilbur, 1980; Semlitsch & Gibbons, 1988). The extensive fish-amphibian predation literature suggests that centrarchids (sunfish, bass), eocids (pike), and salmonids (trout) are important predators, whereas fish such as cyprinids (minnows, carp) are not. Some studies have investigated patterns of amphibian distribution in relation to fish presence, but they have generally taken a single-species approach or have been limited in spatial extent (Petranka, 1983; Bradford, 1989; Bradford et al., 1993; Fellers & Drost, 1993; Br6nmark & Edenhamn, 1994). Few studies have examined either the response of natural amphibian communities to predatory fish (e.g. Macan, 1966; Sexton & Phillips, 1986), or investigated patterns on a geographic scale. In recent years much attention has focused on the question of global amphibian decline and predation (primarily exotic fish introductions) has been suggested as one of several possible causal factors (Barinaga, 1990; Pechmann et al., 1991; Wake & Morowitz, 1991; Blaustein et al., 1994a; Pechmann & Wilbur, 1994). Exotic fish species have been widely introduced on a global scale (Courtenay & Stauffer, 1984) and have been linked to extinctions of amphibian populations (Hayes & Jennings, 1986; Bradford, 1989; Bradford et al., 1993; Fellers & Drost, 1993). In this study we investigated amphibian species richness and distribution with respect to the presence of predatory fish in ponds in southwestern Ontario, Canada. We took a community-based approach by using repeated surveys to construct local species lists across a large geographic area which has been highly impacted by humans and subjected to numerous fish introductions. Questions of interest include: are predatory fish capable of altering community structure (species richness) or distribution of individual species on a geographic scale? Based on the extensive amphibian predation literature, we predicted that fish predation can reduce amphibian species richness, but that individual species would be differentially affected. Because of the importance of size and chemical defenses in predator-prey relations, amphibian species having largebodied adults or tadpoles, or unpalatable species should be better able to coexist with predatory fish.

METHODS Study area and species We conducted the study in southwestern Ontario, Canada (42,962 km 2) where we surveyed 178 ponds between Point Pelee (42° 10' N, 82° 30' W) and Tobermory (45 ° 12' N, 81 ° 40' W). Southwestern

Ontario lies within the Eastern Deciduous forest, but its landscapes have been highly modified by humans for agriculture. Loss of original wetlands in portions of the study area is >90% (Snell, 1987). Today, ponds constructed primarily for agricultural use are an important amphibian habitat in southwestern Ontario. These ponds range from semi-permanent to permanent and have often been stocked with game fish by landowners and government authorities for angling. Essentially the same pool of pond-dwelling amphibians occurs across the study area, but the patterns of local species richness and occurrence differed among regions within southwestern Ontario (Hecnar, 1996). Among regions, mean local amphibian species richness (s-diversity) ranged from 1.95 ± 0.131 to 3.93 + 0.222 species/pond and the frequency of occurrence for individual species ranged from 1 to 80% of ponds (Hecnar, 1996). Pond selection and surveys We located ponds using topographic maps, aerial photographs, information provided by landowners, and by chance. We added ponds to our survey list if permission for access was obtained. From late March to late July in 1992, 1993, and 1994 we surveyed 178 ponds. The ponds we used averaged 6422+2594.3 m 2 in area and 1.9+0.083 m maximum depth. Although a few new ponds were added to the list in the second and third years, most ponds (67%) were surveyed in each of the three years. In each year we surveyed each pond on at least three dates and combined day and night visits. Searches involved from three to seven people searching from the pond perimeter out to ~ 10 m, and wading, canoeing, and dip netting through the pond. When woodlands were adjacent we extended searches to detect terrestrial stages of amphibians. We identified amphibians visually or by their calls. If any life stage of a species (adult, larvae, eggs) was encountered at a pond over the study period we considered that species to be present. Conversely, if we detected no signs of a species we considered it to be absent. We identified large fish visually and obtained information on stocking and recent catches from landowners and government authorities. We identified small fish visually and by dip netting. We did not electrofish or set nets for fish to minimize disturbance to the sites. With the intensive repeated surveys we were able to construct accurate amphibian species lists for individual ponds. Data analyses We classified each pond into one of three fish categories: fish-free, non-predatory fish, or predatory fish. Nonpredatory fish ponds were those having small species such as minnows, darters, sticklebacks, or primarily scavengers or herbivorous species such as catfish and carp. Predatory fish ponds were those having large predatory species such as sunfish, bass, trout, perch, and pike which are documented as preying on amphibians. We made the distinction between non-predatory and

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Fish predation on amphibians

predatory fish because not all fish can be considered important predators on amphibians. Although some small fish species may eat some amphibian eggs or small tadpoles, we assumed that they are not important predators. Ponds with predatory fish also frequently contained non-predatory fish. To compare amphibian species richness among fish categories we used one-way analysis of variance (ANOVA) on raw data. Bartlett's test indicated that variances among categories were homogeneous (p = 0.212) so data transformation was unnecessary. For a posteriori multiple comparisons we used Tukey's Highest Significant Difference (HSD) test. To determine if the presence or absence of individual amphibian species was dependent on fish category we used contingency table analyses and computed the G-statistic. We used William's correction rather than Yate's correction because the latter is overly conservative (Sokal & Rohlf, 1981). We initially planned to do 2 × 3 analyses, but because amphibian species richness did not differ between fish-free and non-predatory fish ponds (see Results) we pooled both into one nonpredatory category and conducted two × two analyses for a more powerful test. To investigate how size characteristics of amphibian species (adult, larval, clutch) related to fish predation we divided the amphibian species into predatory and nonpredatory fish pond groups. If a species occurred in proportionately more of the predatory ponds than the non-predatory ones (see Results, Fig. 2), we placed it in Table 1. Species of amphibians and predatory fish encountered in southwestern Ontario ponds from 1992 to 1994

Amphibians Green frog Northern leopard frog Wood frog Bullfrog Pickerel frog Mink frog Northern spring peeper Western chorus frog Eastern gray treefrog American toad Red-spotted newt Spotted salamander Fish Pumpkinseed Bluegill Green sunfish Largemouth bass Smallmouth bass Black crappie Yellow perch Rainbow trout Brown trout Speckled trout Northern pike Longnose gar Bowfin

Rana clamitans R. pipiens R. sylvatica R. catesbeiana R. palustris R. septentrionalis Pseudacris crucifer P. triseriata Hyla versicolor Bufo americanus Notophthalmus viridescens Arnbystoma maculatum Ambystoma jeffersonianum laterale complex Lepomis gibbosus L. macrochirus L. cyanellus Micropterus salrnoides M. dolomieui Pomoxis nigromaculatus Percaflavescens Onchorhynchus mykiss Salmo trutta Salvelinus fontinalis Esox lucius Lepisoteus osseus Amia calva

the predatory group, and vice versa. We used the basic occurrence patterns as the classification criterion because some of the two × two tests lacked sufficient power to detect effects (see Results, Table 2). The occurrence criterion makes biological sense and permitted us to include all 13 amphibian species in the size comparisons. We used midpoints of average size ranges reported in the literature for adults, larvae, and clutches of each species. We compared the size data between predatory and non-predatory groups using t-tests. Clutch size was square root transformed to homogenize variance, but transformation of adult and larval size was not required (Bartlett's test, p = 0 . 2 4 4 and 0.187 respectively). We used the BIOM-pc software package (Sokal & Rohlf, 1991) for G-tests and conducted all other analyses using SYSTAT (Wilkinson, 1990).

RESULTS We observed 13 amphibian and 13 predatory fish species at ponds in southwestern Ontario from 1992 to 1994 (Table l). Of the 178 ponds surveyed, we classified 75 as fish-free, 46 as non-predatory, and 57 as predatory fish ponds. We also observed non-predatory fish at 84.2% of predatory fish ponds. Centrarchid and salmonid fish species had the highest frequency of occurrence at predatory fish ponds. Percent occurrence was sunfish (65), bass (49), trout (20), pike (10), perch (8), and other fish (4). Mean local amphibian species richness (at-diversity) was 3-55 ± 0.133 species/pond (range 0-9), but was not correlated (Pearson product moment) with either pond area (r=0-14, p=0-270, n = 155) or water depth ( r = -0.10, p = 0-674, n-- 155). Amphibian species richness differed significantly among ponds according to fish category (ANOVA; F2,175--5-45, p=0.005; Fig. 1). Multiple comparisons (HSD) indicated that amphibian species richness was 4.5

t~ I,M

z

z

4.0

w ~

3.5

z in

"7 3.o II

2.5 FISHFREE

NONPRED PREDATOR

FISH C A T E G O R Y

Fig. 1. Amphibian species richness at ponds (mean ± 1 SE) among fish categories in southwestern Ontario.

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S. J. Hecnar, R. T. M ' C l o s k e y Table 2. Presence and absence of amphibian species at ponds among fish categories

Number of ponds where a species was present is followed in parentheses by the number of ponds where it was absent. Fish-free and non-predatory classes were pooled into one category for analysis. Species

Fish category Non-predatory Predatory

Fish-free R. clamitans R. pipiens R. sylvatica R. catesbeiana R. palustris R. septentrionalis B. americanus P. crucifer P. triseriata H. versicolor N. viridescens A. laterale A. maculatum

61(14) 53(22) 16(59) 2(73) 3(72) 0(75) 50(25) 38(37) 16(59) 16(59) 22(53) 2(73) 3(72)

41(5) 35(11) 8(38) 4(42) 1(45) 1(45) 33(13) 27(19) 1(45) 14(32) 16(30) 1(45) 0(46)

G

54(3) 32(25) 6(51) 4(53) 3(54) 1(56) 43(14) 14(43) 1(56) 5(52) 4(53) 0(57) 0(57)

4.330* 4.692* 2.515 0-282 0.346 --~ 0-883 13.745"** 8-048** 6-926** 14-720"** --~ _a

*p < 0.05, **p < 0.01, ***p < 0.001. "Sample size too small to detect effect. significantly lower in predatory fish ponds than in either non-predatory (p = 0.008) or fish-free (p = 0.018) ponds. Species richness did not differ between non-predatory and fish-free ponds ( p = 0.818). For individual amphibian species, patterns of cooccurrence with fish differed (Table 2, Fig. 2). Presence of 6 of i 3 amphibian species was significantly dependent on the presence of predatory fish (Table 2). The leopard frog, spring peeper, chorus frog, gray treefrog, and redspotted newt co-occurred with predatory fish less frequently than expected. Presence of the green frog was also dependent on presence of predatory fish but it cooccurred more frequently than expected. Presence of the woodfrog, bullfrog, pickerel frog and the American toad was independent of the presence of predatory fish (Table 2). Although both A m b y s t o m a species (A. laterale, and A. maculatum) did not co-occur with predatory fish a G-test was not conducted because insufficient power existed to detect a significant effect with the small

1°°11

[ ] NON-PRED

t~11

8oll I

I IIPRED FISH / .

=

60

.

~D 40

~

*

z

~ ~ . . , , ~0~o,.~ +~, ~ ~"

~.

s~ ~.

~. ~., ,.,-.,,~,~,. v ' x ~"

b"

Fig. 2. Proportion of predatory fish and non-predatory ponds occupied by amphibian species in southwestern Ontario. *Species whose presence is significantly dependent on presence of predatory fish (see Table 2). tA. jeffersonianum-laterale complex.

Table 3. Adult and larval body sizes (cm) and clutch sizes (no. of eggs) of amphibian species encountered in southwestern Ontario ponds

Adult frog and toad body size is snout-vent length. All larval sizes and adult salamander sizes are total length. Values are average range midpoints. Species R. clamitans R. pipiens R. sylvatica R, catesbeiana R. palustris R. septentrionalis B. americanus P. crucifer P. triseriata H. versicolor N. viridescens A. laterale A. maculatum

Adult -c

Size~ Larvae

Clutch

7-4 7.1 5.3 12.1 6.0 5-9 7-1 2-6 2.9 4.2 9-0 12.0

9.0~ 8.5L 4-5L I 1-5v-v6-7~w 8.8L 2.7-L 2.9-w 2.9-L 4.5 -L 3.83 3.73

3750-L 400w0 ~-8250~ 16000-~ 2500 wa 3000v 6000E 900wR 60y~ 175w0 ~g 76y2853

15.5

5.2L

150L

aSources: -C-Conant & Collins 1991; -L-Logier1952; MMinton 1972; VVogt 1981; WWalker 1946; WRWright 1914.

Fish predation on amphibians

sample size. However, when data for both Ambystomids were pooled, the test was significant (G=4.300, p < 0.05) with ambystomids co-occurring with fish less often than expected. Sample size was also too small for testing the mink frog R. septentrionalis. Size differences existed between species that occurred at proportionately more predatory than non-predatory ponds (Table 3). Clutch size was significantly different (t11=-3.20, p=0.008) with predatory pond species having larger clutches (6250+2510.0 eggs) than nonpredatory pond species (1184 ± 483.4 eggs). Larval size (total length) was significantly different (t1~=-2.3l, p=0.041) with predatory pond species having larger larvae (7.7 + 1.48 cm) than non-predatory pond species (4-5 ~0.64 cm). Adult size did not differ between predatory and non-predatory pond species (tll=--0.17, p=0-869). Because anurans have a different body plan than caudates, we repeated the adult size analysis for the 10 anuran species alone. Adult size was significantly different (t8 = 2.34, p=0-048) with predatory pond species having larger adult body size (7.8 • 1.15 cm) than non-predatory pond species (4.4 ± 0.82 cm). DISCUSSION Two major patterns are clear. First, amphibian species richness (a-diversity) was lower in ponds where predatory fish were present compared to non-predatory or fish-free ponds. Second, not all amphibian species appeared to be negatively affected by predatory fish. Species with large bodies and large clutch sizes cooccurred with predatory fish more frequently than small-bodied species. Body sizes of individuals determine their risk of predation and growth rate determines how long individuals remain at risk (Heyer et al., 1975; Wilbur, 1980, 1984). If suitable refuges are available in the littoral zones of ponds, tadpoles can quickly outgrow predators (Heyer et al., 1975). Because of the spatial scale of this study it was, of necessity, non-experimental. Correlation does not necessarily indicate cause; however, our results are both consistent with and complement the extensive fishamphibian literature which is based largely on smallscale studies. Although most other studies have taken a small-scale approach, it appears that the results of local interactions between predatory fish and amphibians can translate to patterns at regional or geographic scales. Over 160 species of predatory fish have been introduced in 120 countries worldwide (Welcomme, 1984), and indigenous predatory fish are often translocated within their ranges to habitats that previously lacked fish (Crossman, 1984; Page & Burr, 1991). The potential for impact of introduced fish on amphibian diversity is global in scope. Population or species effects As a group, the ranid frogs (excepting the leopard frog) did not appear to be negatively affected by the presence

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of fish predators. Ranid frogs have several characteristics that would lessen the impact of predation. First, ranids have relatively large clutch sizes, and both adults and larvae tend to be large-bodied. Second, the larvae of green frogs and bullfrogs are unpalatable to fish (Kats et al., 1988). However, other ranids such as the leopard frog and woodfrog have palatable larvae which may explain why these two ranids occurred in proportionately more non-predatory ponds (Table 2). Experimental evidence suggests that adult woodfrogs may be able to detect predatory fish presence and choose alternate breeding sites (Hopey & Petranka, 1994). The pickerel frog also commonly occurs with predatory fish (Kats et al., 1988), and the adult is known to have highly toxic skin secretions (Vogt, 1981). At two ponds which were highly productive stocked fish ponds, pickerel frogs were the numerically dominant amphibian species (S. J. Hecnar and R. T. M'Closkey, unpublished data). In contrast, the green frog occurred in predatory fish ponds more often than expected. However, it is unlikely that this indicates a beneficial relationship between the species. Neither green frogs nor fish are important items in each others, diets; however, co-occurrence can be explained by similarity in habitat requirements. The green frog is an aquatic species and both adults and larvae overwinter in ponds in the study area (Hecnar, personal observations). For both green frogs and fish, deep permanent waters are essential for survival. The same is likely the case for the bullfrog. In aquatic frogs (e.g. green frog, bullfrog) high fecundity and relatively small eggs are probably adaptations which act to lessen the risk of predation (Wilbur, 1984). All three species in the treefrog group (spring peeper, chorus frog, and gray treefrog) occurred less often than expected in predatory fish ponds. All three species have palatable larvae and the peeper and chorus frog do not increase refuge use in the presence of fish predators (Kats et al., 1988). Also, adults of both the peeper and chorus frog are small enough that consumption by larger fish is plausible. The size of adult gray treefrogs and the fact that they enter water only to amplex and oviposit, likely precludes any substantial effect of fish predation. Predation on small vulnerable eggs and larvae is far more likely. Occasionally, we have dip-netted gray treefrog larvae showing apparent predator damage to tail fins. Gray treefrog larvae have bright reddish orange tail fins which might function to deflect attacks of visually-oriented predators away from the body, similar to the black tail tip found in pond populations of Blanchard's cricket frog Acris crepitans blanchardi (Caldwell, 1982). Experimental evidence indicates that adult diploid gray treefrogs Hyla chrysoscelis altered their choice of calling sites and oviposition sites based on predators or competitors present (Resetarits & Wilbur, 1989, 1991). The American toad did not appear to be affected by the presence of predatory fish. The toad has toxic eggs, larvae, and adults (Voris & Bacon, 1966; Licht, 1968;

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S. J. Hecnar, R. T. M'Closkey

Kruse & Stone, 1984). Largemouth bass quickly learn to avoid feeding on toad tadpoles and select more palatable species such as hylids (Kruse & Stone, 1984). Our sample size for ambystomid salamanders was low, but as a group they never co-occurred with predatory fish. This disassociation of salamanders with fish has often been reported (Petranka, 1983; Semlitsch, 1988; but see Sprules, 1974; Figiel & Semlitsch, 1990). Predatory fish can alter growth, density, and activity of ambystomid larvae (Figiel & Semlitsch, 1990; Sih et al., 1992). Although adult ambystomids may have a refuge from predation in their relatively large size, noxious skin secretions, and aposematic coloration (Brodie, 1983), larvae are palatable (Kats et al., 1988) and fish are capable of eliminating larval recruitment (Ireland, 1989) or extirpating local populations (Burger, 1950; Sexton & Phillips, 1986). The effects of fish predation on the distribution of amphibians have been noted in other studies. Petranka (1983) studied the spatial segregation of the smallmouth salamander Ambystoma texanum and fish in 90 streams in the central United States. Eggs and larvae were largely restricted to first and second order streams where no fish occurred and sunfish were found to be capable of eliminating all the larvae from individual pools. Bradford et al. (1993) reported on the decline of the mountain yellow-legged frog R. muscosa in the Sierra Nevada mountains of California. Introduced fish were implicated as causing historic and recent population extinctions. From a survey of 312 lake sites they concluded that isolation of remaining frog populations by introduced fish increased the likelihood of local extinctions and prevented recolonization. A similar conclusion was reached by Fellers and Drost (1993) who documented the disappearance of the cascades frog Rana cascadae from 14 of 15 historic localities in California, although they also suggested other factors in addition to predation. Br6nmark and Edenhamn (1994) studied the distribution of the European treefrog Hyla arborea, a species which has been declining in northwestern Europe in recent years. In a study of 25 ponds they concluded that fish presence was an important factor in reducing successful reproduction and limiting the occurrence of the treefrogs. Community effects Semlitsch (1988) reported on the allotopic distribution of mole salamander Ambystoma talpoideum and spotted salamander A. maculatum from 21 sites in South Carolina. Experiments produced evidence that predation on eggs and larvae excluded A. talpoideum from ponds with predatory fish, but A. maculatum co-occurred with predatory fish in five of the sites. Although sunfish readily ate A. talpoideum eggs and larvae of both species, the jelly egg mass envelope in A. maculatum prevented fish predation.

Bradford (1989) studied the distribution of the mountain yellow-legged frog and Pacific treefrog Pseudacris regilla with respect to introduced salmonid fish in 67 mountain lakes in the Sierra Nevada Mountains of California. The yellow-legged frog was distributed allopatrically with introduced fish in permanent waters, but the Pacific treefrog occurred only in ephemeral habitats that lacked fish. Few studies have viewed entire natural amphibian communities with respect to fish predation. Sexton and Phillips (1986) reported on a natural experiment in three ponds where river flooding introduced predatory fish into one pond, non-predatory fish into another, and the third pond remained unaffected. After fish invasion, species richness in the predatory fish pond was reduced from 11 to two species, but no effect was observed in the pond invaded by minnow species. In the agricultural landscapes of southwestern Ontario, the extent of natural wetlands has been drastically reduced in the last century (Snell, 1987) and many permanent ponds have been constructed for agricultural or recreational use (Hecnar, 1996). The historical shift in the nature of habitats and the introduction of predatory fish coupled with differential susceptibility to predation among amphibian species have likely resulted in an altered composition of species in less diverse communities. Role of refuges and habitat structure Researchers often refer to amphibians as being temporary or permanent water species. However few species show perfect fidelity to either habitat (Kats et al., 1988; this study). Species that are normally restricted to ephemeral sites because of fish predation face desiccation during egg and larval development as the major mortality factor (Heyer et al., 1975; Wilbur, 1984). By residing in permanent waters that lack fish predators, populations of so-called temporary pond species would reduce their risk of extinction in drought years. This may be one reason that temporary species often occur in permanent water habitats. Considering the complexity of predation and habitat heterogeneity, it is not surprising that temporary pond species often co-occur with predatory fish. Habitat heterogeneity provided by shallow and heavily vegetated perimeters of permanent ponds may provide suitable nursery refuges for rapid growth of tadpoles, permitting them to outgrow their risk of predation (Heyer et al., 1975). Amphibian species richness was positively correlated (Pearson product moment) with the amount of emergent vegetation at predatory fish ponds (r=0.38, n=56, p=0.003; S.J. Hecnar and R.T. M'Closkey, unpublished data). The most diverse pond that we surveyed (amphibian species richness = 9) was a government-run fish pond stocked with large speckled and rainbow trout. Populations of spring peepers, gray treefrogs, leopard frogs, and woodfrogs used refuge habitats in the shallow littoral zone. Although common

Fish predation on amphibians

in surrounding ponds, red-spotted newts were never observed and were presumably excluded from the site.

Alternate explanations The impact o f introduced fish on amphibians may not be limited to direct predation. Introduced fish were implicated as carriers of a pathogenic fungus that resulted in near complete mortality of eggs in the western toad Bufo boreas populations in Oregon (Blaustein et al., 1994b). Indirect facilitation by sunfish preying on the invertebrate and salamander predators of bullfrog larvae can result in increased bullfrog density (Werner & McPeek, 1994). In larval amphibian communities, predators that selectively feed on competitively dominant species can promote the coexistence of subordinate competitor species (Morin, 1983). The red-spotted newt is a highly toxic species which produces tetrodotoxins in skin secretions (Brodie et al., 1974; Formanowicz & Brodie, 1982; Brodie, 1983). Notophthalmus adults also have countershading which aids in concealment and aposematic coloration to advertise its toxicity (Brodie, 1983; Duellman & Trueb, 1986). Therefore, the negative effect of fish on newt distribution may not be via predation. Newts and fish have broadly overlapping foraging niches, but newts may be less efficient in resource exploitation, resulting in competitive exclusion from some sites by sunfish (Bristow, 1991). In the field, we generally observed larger newt populations in ponds lacking predatory fish than in ponds where newts and fish were syntopic. In 1992, a landowner stocked trout in a pond having an abundant newt population which resulted in a local extinction in 1993. Nearby in another pond that was stocked with trout and bass, newts have continued to coexist with fish for many years. While introduced fish have undoubtedly played a role in amphibian decline at some localities in southwestern Ontario it would be naive to assume that it is the only factor. Massive clearance of woodlands has also occurred in the study area and many amphibian species require woodlands at some point in their life cycles. The amount of woodlands occurring in the vicinity of ponds is the most important geographic factor affecting amphibian diversity that we have detected (Hecnar & M'Closkey, 1996). Further study is required to determine the relative importance of habitat loss (deforestation) and habitat degradation (fish introduction) in affecting amphibian species richness in southwestern Ontario.

ACKNOWLEDGEMENTS D. Hecnar, T. Hecnar, J. Cotter, D. Chalcraft, R. Poulin, A. Plante, D. Peterson, J. Barten, and C. Watson helped with fieldwork. We thank the Ontario Ministry of Natural Resources and Parks Canada, the AusableBayfield, Maitland Valley, Essex Region, Saugeen Valley, and St Clair Region Conservation Authorities, and

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many private landowners for their continued co-operation and interest. S. and V. Hecnar provided field accommodations. Funding was provided through a Natural Sciences and Engineering Research Council of Canada grant to R.T.M. and a Parks Canada contract.

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