habitat heterogeneity and the functional signifiance of fish in river food webs

Habitat Heterogeneity and The Functional Signifiance of Fish in River Food WebsAuthor(s): Mary E. PowerReviewed work(s):Source: Ecology, Vol. 73, No. 5 (Oct., 1992), pp. 1675-1688Published by: Ecological Society of AmericaStable URL: http://www.jstor.org/stable/1940019 .Accessed: 16/04/2012 13:23

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IFcologt' 73(5), 1992, pp. 1675-1688 1992 by the Ecological Society of Amenica

HABITAT HETEROGENEITY AND THE FUNCTIONAL SIGNIFICANCE OF FISH IN RIVER FOOD WEBS'

MARY E. POWER Department of Integratvfe Biology, ULniversity of Califbrnia, Berkelea, Berkeley, California 94720 USA

. lbstract. Predation by fish (roach, II(esJerolelucas s)';n ;netricus, and steelhead Onco- uiYnchus ;nhvkiss) produced strong cascading effects on biota associated with boulder- bedrock substrates in pools of a northern California river, but not on gravel-dwelling biota. Enclosure-exclosure experiments in the South Fork Eel River of northern California (39044' N, 123039' W) showed that fish, by suppressing densities of damselfly nymphs and other small predators, released algivorous chironomids (Iseudochir'onomalus richardsoni) from predation. Chironomids in turn dramatically reduced algal standing crops. In contrast, fish had little effect on algae or invertebrates associated with gravel. Gravel-dwelling heptageniid mayflies were behaviorally inhibited from using tops of stones in fish enclosures, and stone surfaces had more chironomid tubes in fish enclosures than in fish exclosures. However, no effects on epilithic algae or densities of invertebrates comparable to those of biota on boulder-bedrock substrates were detected. These spatially varying predator effects in a river parallel results from marine benthic systems, where strong effects of large predators documented for rocky intertidal habitats and unvegetated soft bottoms are not conspicuous in seagrass beds.

Kev words: benthos; fish; food webs; omnivore: refuges; ri ver coitnnunities; strong interactions; stdbstrate heterogeneity: trophic cascades.

INTRODUCTION

The effects of predators on their prey depend not only on biological attributes of both, but also on the setting of their interaction. Predators that are functionally important in one environmental setting may play only a minor role in another. For example, the starfish f'isaster is a keystone species structuring intertidal communities along the Washington State coast, but may be "just another starfish" in Alaska (Paine 1980:670). Predator impacts may vary over space for a number of reasons. Fretwell (1 977, 1987) and Oksa- nen et al. (1 98 1) predict that the functional significance of predators should shift along gradients of environmental productivity. In habitats too unproductive to support permanent predator populations, transient predator individuals might occur, but would be inca- pable of limiting populations of their prey (but see Holt [1984. 1985] and T. Oksanen [1990] for discussions of spillover effects). Predator impacts also vary over space in relation to structural variation of the environment (Huffaker 1958, Smith 1972). The efficiency with which predators deplete their prey declines with increasing availability of prey refuges. Losses to predation can be offset by prey immigration, which varies with the de- gree of habitat isolation (Holt 1984, Cooper et al. 1990). All else being equal, the net impacts, or functional significance, of predators should be weakest in contin- uous habitats with refuges, and strongest in isolated

I Manuscript received 8 July 199 1; revised 22 October 199 1 accepted 28 October 199 1.

habitats that lack refuges (Huffaker 1958, Hastings 1977, Caswell 1978, Crowley 1981, Holt 1984).

Substrates in rivers and streams vary considerably in their patchiness and quality as habitats and refuges (Hart 1978, Minshall 1984, Huryn and Wallace 1987). In river reaches where sediment supply from the wa- tershed is high relative to the capacity of the river to transport bedload, fine-grained gravel sediments will cover much of the bed (Dietrich et al. 1989). Boulders and bedrock formations will emerge as habitat islands from periodically mobile gravel beds. Conversely, in rivers that are "sediment-starved" relative to their transport capacities, coarse bedrock and boulder substrates will predominate on the bed surface (Dietrich et al. 1989), and the isolated deposits of finer sediments may function as habitat islands for gravel-dwelling biota.

For a river reach of the first type, I compare previ- ously reported (Power 1990a) effects of fish on the biota of boulder-bedrock habitat islands with new results on fish effects on biota of the gravel bed. I used direct behavioral observations and diet data to ascertain that fish fed in both microhabitats, and an enclosure-exclosure experiment to assess their direct and indirect impacts on lower trophic levels.

STUDY SITE AND BIOTA

The study site is a 1-km reach of the South Fork of the Fel River (39044' N, 123039' W) in Mendocino County, California, USA. The study reach is surrounded by an old-growth conifer forest dominated by Doug- las fir (Isciiudotsuiga mnczn-iesil') and coastal redwood

1676 MARY E. POWER Ecology, Vol. 73, No. 5

(Seqiuo)ia selmperlwires). Nearly all precipitation falls between October and April (Anderson et al. 1987). Following winter floods. baseflow discharge drops from 10-25 m3/s to <1 m3/s during the summer, and the river slows to form a series of large lentic pools con- nected by short riffles. In mid-June 1989, an estimated 30)/4) of the wetted area in the 1-km study reach was lentic (flow <5 cm/s): by late July this lentic area was 59%/o. by late August, 60%. and by early September, 86' /%.

The bed of river pools was largely composed of gravel particles 2-64 mm in median diameter surrounding boulders >256 mm in median diameter and emergent bedrock formations. In the 1-km study reach, the estimated percentage (by area projected to the water surface) of boulders and bedrock in the wetted channel increased from 39% to 44% to 50% from 14 February to 20.July to 10 September, because lateral gravel bars dried as water levels dropped. During winter floods, gravel is transported around the more stationary boulders and bedrock. Biota on boulder and bedrock "islands" are less subject to abrasion than biota on smaller stones during episodes of bed movement (Hynes 1970, Sousa 1979, McAuliffe 1984, Power and Stewart 1987).

Because the active channel maintained by winter floods is wide (z30 m), during summer much of the river bed is sunlit. Turfs of filamentous green macroalgae (primarily Cladophora gloincrala L.) develop on boulders or bedrock outcrops, probably by vegetative regrowth from basal pads that survive winter scour. During most years. turfs of Cladophora attain lengths of several metres by June. and, in places, Cladophora canopy covers most of the river surface (Power 1990a, b. c). Gravel substrates are thinly coated with diatoms and detritus for most of the year. Short, attached fil- aments of green algae first appear on pebble and cobble substrate only in late summer or fall of most years, suggesting infrequent recruitment from zoospores (Power 1990c).

While some invertebrates occur on both gravel and turf-covered boulders, many taxa are specialized for one microhabitat or the other (Minshall 1984, Dudley et al. 1986). Heptageniid mayfly nymphs, common primary consumers in gravels, are morphologically specialized, with flattened bodies that enable the insects to wedge themselves into crevices and spaces between stones (Hynes 1970:1 23). Gravel sediments are also inhabited by tube-dwelling chironomids, stoneflies, naucorids (lhrtisuas inormH)on), and dragonfly (aesh- nid. gomphid, and libellulid) nymphs. (Cladophora turfs on boulders and bedrock harbor perching damselfly (lestid and coenagrionid) nymphs, hydrophilid beetle larvae (Inochrus sp.), and a guild of algal-weaving chi- -onom ids dominated by the species Pseudochironornus richardsoni (Power 1990a, h). Oligochaete worms associated with detritus in chironomid retreats are also common in algal turfs (Power 1990h). Caddis larvae (Lupidostonma sp., Gumiaga sp., Ilic/ops vchc sp.), meg-

alopteran larvae (Sialils sp.), and a variety of mayflies (baetids. siphlonurids, ephemerellids) occur in both m icrohabitats.

The most common fish in the study reach in early summer are California roach, Ilesperolcucas stmmn l cii- cus) and juvenile steelhead (Oncorhynchus ;uvkiss [=LSaIho gairdneri]). In contrast to habitat-specific invertebrates described above, these fishes swim freely between gravel and turf-covered boulders. By midsum- mer, roach and the few overwintering stickleback (Gas- tcrosteis acIIlealu.S) produce large numbers of fry.

METHODS

Enclosures/exclosures

Food webs were allowed to develop in the presence or absence of large fish in 1 2 large (6 m2) pens. Pens were 3 m long x 2 m wide x 1.3 m high, with walls of plastic screen lined with black plastic shade cloth. Most stream invertebrates and fish fry could pass through the 3-mm mesh shade cloth walls. The shade cloth extended 60 cm below the bottom edge of the wall, and this skirt was anchored with gravel to pre- clude passage of large fish. Otherwise, the enclosed river bed was unmodified. Periodic hand cleaning and drift deflectors of 20 cm wide aluminum flashing set at the water surface ;1 m upstream from pens kept drifting detritus from clogging pen walls. Each pen was built around boulders or bedrock that supported Cla- dophora turfs. These boulders or bedrock islands were surrounded by gravel inside each pen, which was con- tiguous with gravel outside enclosures. During July when flow inside and outside pens in pools was too low to measure, rates of fine sediment outfall were similar inside pens (X + 1 SE = 0.78 ? 0.06 mg cm 2 d ', n = 104) and in the open river (1.05 + 0.18 mg-cm '-d ', i = 23) (l = 1.35, P > .1 from Welch's approximate t test for samples with unequal variance [Zar 1984:131]). Maximum water depths in pens ranged from 62 to 83 cm, overhead forest canopy measured on a spherical densiometer (Lemmon 1957) ranged from 3 1.8 to 69.5% open, and boulder or bedrock covered 27-58% of the enclosed area (projected to the water surface). (The remaining area was gravel.)

The 12 pens were distributed in groups of four over a 1-km reach. Within each group, two pens were selected randomly to be stocked with fish ("enclosures"): the other two were left unstocked enclosureses"). Stocked enclosures received 20 juvenile steelhead (28- 50 mm SL [standard length, Lagler et al. 1977:403]) and 40 roach (30-70 mm SL). Physical attributes of pens in both treatments (means ? l SE) were on average similar. Maximum depths, canopies (% open), and / boulder-bedrock were: 72 ? 1 cm, 54.4 ? 6.3'%, and 37 ? 3% in stocked enclosures, and 78 + 4 cm, 51.5 ? 3.8%, and 38 ? 5% in enclosures, respectively. Fish were size matched among enclosures. Stocked groups reflected proportions and size distributions of fishes

October 1992 SPATIALLY VARYING TROPHIC CONTROL BY FISH 1677

(427 roach, 186 juvenile steelhead, and 1 stickleback) captured by trapping and electroshocking the study reach in early June, when experiments began. Densities of fish in enclosures (10 individuals/) were within ranges observed (5-48 individuals/) near large turf- covered boulders in the open river, but were higher than overall fish densities (1-2 individuals) estimated from electroshocking surveys of larger river reaches encompassing both gravel and boulder-bedrock substrates (Brown and Moyle 1989; M. Power and J. Nielsen, unpublished data).

Sampling

Algal standing crops and condition were monitored nondestructively in each pen at 23-48 sites along five cross-stream transects, where these intercepted boulders or bedrock. At sites spaced at 10-cm intervals along each transect, I noted dominant and subdomi- nant taxa of macroscopically conspicuous algae, measured the visually estimated modal height of these algae, and their condition and density (% canopy cover) (see Power and Stewart [1987] for more details of methods). No macroscopically conspicuous algae were observed on gravel substrates in pools, inside or outside of enclosures, throughout the 6-wk experiment.

In addition, at the onset of experiments (5 June), 3 wk (22 June), and 5-6 wk (16 and 24 July) afterwards, I collected cores of algae and of sediments from pens and from nearby sites in the open river for samples of algae and associated biota. Organisms in gravel substrates were sampled with a 6.4 cm diameter metal corer that was pushed 5 cm into the bed. These samples were elutriated in the field and preserved in 70% eth- anol. Attached algal turfs and associated organisms from boulder and bedrock outcrops were collected with 9.0 cm diameter cores, fitted at the top end with a three-layered sleeve cut from nylon stockings. The sleeve (mesh size of -0.3 mm) retained organisms as the sample was lifted through the water column, until the sample could be blackflushed into plastic bags and preserved (see Power [1 990a] for details of methods).

Block effects were insignificant, and were dropped from the analysis. Algal biomass and the densities of the dominant invertebrate primary consumers and predators from enclosures, enclosures, and open sites were subjected to planned comparisons at the end of the experiment with one-way Kruskal-Wallis tests, followed by Dunn's nonparametric multiple comparison test (Zar 1984:200) when results were significant. For some of the predatory invertebrates, heteroscedasticity among samples could not be eliminated by data trans- formation. I applied the relatively conservative Krus- kal-Wallis test for all comparisons so that the assess- ment of treatment effects for different variables was not confounded by the use of tests with different power.

At the end of the experiment, gut contents of enclosed and free swimming roach and steelhead were examined. Individuals were dissected under 10 x mag-

nification in the laboratory. The presence or absence of diatoms, filamentous algae, and detritus in their guts was noted. Macroinvertebrates were identified and measured under dissecting microscopes (10-70 x). Published regressions of head capsule width to dry mass (Smock 1980) were used to estimate dry biomass of ingested invertebrates.

Cladophora biomass sampled from cores was measured as damp mass (after 50 spins in a salad spinner [dryer]), which was converted to dry mass according to the regression: dry mass = 0.09 (damp mass) + 0.02, r2= 0.98 (Power 1990b). Epilithic algae were sampled from cobble surfaces by aspirating areas within 1-cm2 templates. Cells were counted in a Palmer chamber, and cell counts were converted to dry mass estimates by assuming 0.420-0.904 mg ash-free dry mass per 106 cells (from estimates of Stevenson and Stoermer [ 1 982] for diatoms), and that diatoms had ash contents of 50% (Winberg 1971).

Behavioral observations

I observed fish to compare the feeding behavior of caged and free-swimming individuals, and to estimate the feeding effort (number of bites) that caged individuals allocated to turfs on boulders and bedrock, and to gravel substrates. Visibility was excellent, as the water was clear and < 1 m deep, the surface was calm, and I was able to stand within 1-2 m of the fish I observed, near outer walls of enclosures or on the river shore. Before initiating observations, I would wait quietly for 20 min. Within this time, fish disturbed by my approach resumed apparently normal activities.

To quantify feeding rates and microhabitat use by fish, I recorded the bites of a focal individual for timed periods of 1-9 min. During observations, I noted the species and estimated the length of the fish, and recorded the bites it allocated to turf-covered boulders, gravel sediments, enclosure walls, or the water surface. Most observations stopped after 5 min or when I lost sight of the focal individual. To reduce bias and de- pendence in observations, I chose focal individuals by a systematic spatial rule. I visually divided each enclosure or observed open area into six sections: left and right sides of upstream, center, and downstream thirds, and selected sequential focal individuals in nonadja- cent sections.

As a second check for enclosure effects on fish feeding behavior, I used the Morisita-Horn Index (Krebs 1989) to estimate overlap in microhabitat use by enclosed and free-swimming fish:

2 PijPik CH = 2 P., + 2>2

where pi is the proportion of resource i used by consumer j or k. CH ranges from 0 (no overlap) to 1.0 (complete overlap). Enclosed fish sometimes foraged on enclosure walls, which were not available as foraging


TABLE 1. Roach and juvenile steelhead feeding rates (bites/min) inside and outside enclosures, 12 June-14 July 1989.

Fish Inside Outside SL (cm) V 1 SE ti A 1 SE 11

Roach (lesperolucias sls tletricus) 3 .00 .00 4 .16 .06 7 4 .04 .02 1 1 .04 .01 8 5 .03 .01 40 .03 .02 8 6 .04 .01 64 .06 .02 14 7 .03 .01 48 .05 .03 8 8 .02 .01 17 .05 .05 "

Steelhead (Onchorhvxnchus invkiss) 3 .01 .00 8 .01 .01 6 4 .03 .01 1 2 .01 .01 7 5 .02 .01 1 7 .04 .02 5 6 .02 .01 1 9 .02 1

substrates for free-swimming fish. Feeding observations on enclosure walls were not included in computations of overlaps between enclosed and free-swim- m iing roach and steelhead.

Preference of roach and steelhead for boulder vs. gravel substrates was computed as Manly's alpha for undepleted resources (Chesson 1 978, Krebs 1989:396):

a,,1 lt'b, + I' J, _n,, tlie,

where a,, is the preference index for boulder-bedrock substrate. r;, is the proportion of bites on benthic substrates allocated to boulder-bedrock, n,, is the proportion of boulder-bedrock substrate in enclosures (based on area projected to the water surface), and rg, and ns, are the proportions of bites allocated to gravel substrates and proportions of substrates in enclosures, respectively (r,, + r,, 1; n,, + n,,= 1; Manly's alpha ranges from 0 [complete avoidance] to 1.0 [exclusive selection]).

On 14 July (week 6), 1 counted mayflies on tops of stones (28-64 mm estimated median diameter) in all enclosures. After approaching an enclosure for these counts, I would stand quietly for 20 min. Then I counted the mayflies on top of every other stone lying along a longitudinal transect from the downstream to the upstream end of an enclosure. Dark heptageniid mayflies were quite visible against the light-colored stones. After repeating these counts on each of three days, I collected stones from all enclosures for censuses of insects clinging to all surfaces, and for samples of epilithic diatoms grown on top surfaces. To choose stones, I stretched a metre tape along the length of each enclosure. and picked the stone under each 20-cm interval. Selected stones and associated biota were netted (0.3- mm mesh) and transferred to ziplock plastic bags at the water surface. After macroinvertebrates were picked off and chironomid tubes were counted, I aspirated

epilithic algae from 1-cm2 areas circumscribed by a template. These algal samples were preserved in vials in 2% formalin, and later counted under 400 x in the laboratory.

RESULTS

Fish feeding behavior

Feeding behavior of enclosed roach and steelhead resembled that of their free-swimming conspecifics in both rate and substrate choice. Size-specific per capita feeding rates, except for the smallest (3 cm SL) size class of roach, were similar for enclosed and free-swimming fish of both species (Table 1). In addition, roach and steelhead apportioned their feeding among gravels, boulder substrates, and the water surface similarly in enclosures and in the open river (Fig. 1): overlaps in the allocation of bites to these three microhabitats were C, = 0.98 for enclosed and free-swimming roach and C, = 0.95 for enclosed and free-swimming steelhead. Enclosed and free-swimming juvenile steelhead allocated more bites to the water surface than did roach (Fig. 1). Preferences (Manly's alpha) of enclosed fish for boulder-bedrock vs. gravel substrates were 0.68 and 0.77 for roach and steelhead, respectively, suggesting that enclosed individuals of both species pre-

1.0- Ci) Stee IRoach

0.8 -

0

H 0. outside

0

0D4 -ED

0

QL 0.0 Boulder-Turf Gravel Surface Wall

FG 1.0 P

ming roacha)ad Steelhead

and I bEo i en optdprtetet

0.8 b 0 El~~ inside H 0.6 - Moutside

0 0.4-

0

Boulder-Turf Gravel Surface Wall

FIG. 1. Proportion of bites allocated to turf-covered bedrock, gravel, and the water surface by enclosed and free-swimming roach (a) and steelhead (b). Histogram bars are averages and 1 SE Of Six means computed per treatment.


-U Roach Diet by Number Steelhead Diet by Number

>5 40 a 40 b 0 - El free-swimming (n =23) . 30 f t enclosed (n =25) -3 free-swimming (n 12)

enclosed (n =15) -I 20 - 2.0

1.0 - 1.0 -

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Z O) (n E E a) CZ CZ a-)

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175 -175-

FI. 150 - C El free-swimming and 150 El free-swimming i enclosed El enclosed E cz125 - 125

-F 100( 100

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FiG. 2. Numbers of invertebrates per individual free-swimming and enclosed roach (a) and steelhead (b), and estimated biomass of these invertebrates in roach (c) and steelhead (d) guts. "hept" =heptageniid mayflies; "mayf' =other mayflies; ..chiro" chironomids; "dipt" =other dipterans; "nauic" =naucorid bugs, "gerr" =gerrids; "drag" =dragonflies; "dams"= damsclflies: "cadd" =caddisf'ijes; "'elm" =etmid beetles; "hydro" =hydrophilid beetles; "plec" =stoneflies; and "siat" sialids; "unid" =unidentified. Data are averages with 1 SE.

erred boulder to gravel substrates for feeding. Pref- erences of free-swimming fish could not be assessed because the relative availability of the two substrate types within their home ranges was unknown.

Fish gut contents

Guts of roach and steelhead (both >30 mm SL) collected at the end of the experiment contained both gravel- and algae-dwelling insects (Fig. 2a-d). Fourteen categories (orders or families of insects and an unidentified category) were discriminated. No significant differences could be detected in the frequency of consumption of any of these categories by free-swimming vs. enclosed roach or steelhead (t tests on arcsine-square- root-transformed proportions, with a critical level of .004 calculated by the Bonferroni criterion). Mayflies were major components of the diets of both roach (20 and 26%, respectively, of the individuals counted in free and enclosed fish) and steelhead (51 and 38% of the individuals counted in free and enclosed fish, respectively). Damselfly larvae were, because of their large size, important components of the biomass ingested by free-swimming roach and free and enclosed

steelhead, although they were not numerous in fish guts (Figs. 2c, d). Chironomid larvae were infrequent in both roach and steelhead guts, and were insignificant components of their diets by biomass (Fig. 2a-d). Roach also consumed filamentous algae and diatoms in addition to arthropods, while steelhead were strictly car- nivorous (Power 1990a).

Impacts of fish on boulder-bedrock biota At the onset of the experiment, algal turfs covered

all sites on boulder-bedrock substrates sampled in both enclosures and exclosures (Power 1 990a). Subsequent- ly, algal standing crops declined in all pens and in the open river. Declines were greatest in enclosures with fish, and least in enclosures (Fig. 3). Five weeks after experiments began, algae on boulders in stocked and unstocked pens differed conspicuously (Power 1990a) and significantly (Table 2, Dunn's multiple comparison test indicated significant differences between enclosures and enclosures [P < .02]). In six out of six stocked enclosures, Cladophora turfs were reduced to low, pros- trate mats only 1-2 cm high, with a webbed, knotted appearance characteristic of infestation by algal-weav-


TABLE 2. Results of Kruskal-Wallis tests for differences among fish enclosures (n = 6), fish exclosures (n 6), and open sites (n = 3).

Kruskal-Wallis Results Variable Units Date H P graphed

Algal turfs Algae (damp mass/cm2) 16 July 8.23 .016 Fig. 3 Chironomids (individuals/cm2) 16 July 8.64 .013 Fig. 4a Odonates (individuals/cm2) 16 July 8.01 .018 Fig. 4c Megalopterans (individuals/cm2) 16 July 5.25 .072 Fig. 4d Total mayflies (individuals/cm2) 16 July 6.23 .044 Fig. 4e Large mayflies (individuals/cm2) 16 July 1.79 .409 Fig. 4f Large mayflies (individuals/cm2) 22 June 9.68* .008* Fig. 4f

Gravel Chironomids (individuals/cm2) 24 July 4.86 .088 Fig. 4g Plecopterans (individuals/cm2) 24 July 5.17 .075 Fig. 4h Megalopterans (individuals/cm2) 24 July 2.11 .348 Fig. 4j Total mayflies (individuals/cm2) 24 July 2.16 .339 Fig. 4k Large mayflies (individuals/cm2) 24 July 1.79 .409 Fig. 41

* From an unplanned, a posteriori comparison (Dunn's procedure) of large-mayfly densities at the experiment midpoint, 22 June. Large mayflies were the only sampled taxon to show significant differences on 22 June; all other significant differences occurred in planned comparisons of densities in samples taken at the end of the experiment.

ing midges (Power 1990a). In six out of six unstocked exclosures. (ladophora remained erect and became heavily overgrown with diatoms and cyanobacteria (Nostoc sp.). In exclosures, large amounts of Nostoc sloughed off of (ladophora and floated to the water

Attached Algae (mostly Cladophora)

5000

.- * fish j 4000 -a- - no fish

E 0---- open

0' 3000-

CZ) 2000----

E

5 Jun 22 June 16 July

FIG. 3. Changes in algal standing crops over the course of the experiment in enclosures, enclosures, and the open river. Points are averages + 1 SE of six means per cage treatment, six means from open water on 22 June and three means from open water on 5 June and 7 July (each mean is computed from three subsamples per cage or open river site).

surface or sank to the bed. Almost no Nostoc was produced in stocked enclosures (Power 1990a).

Midge densities (predominantly Pseudochironornus richardsoni) increased markedly in enclosures, exclosures, and in the open river from 5 to 22 June. Sub- sequently, densities of chironomid larvae decreased in all habitats, but by 16 July were 10 times higher in enclosures than enclosures (Fig. 4a, Table 2; Dunn's multiple comparison test indicated significant differences between enclosures and enclosures [P < .02]). Chironomid declines in fish enclosures were associated with loss of algal habitat. Densities of chironomids per unit biomass of algae remained nearly constant in enclosures from 22 June to 16 July, but dropped in fish exclosures and the open river (Fig. 4b).

Reductions of chironomid densities in enclosures and the open river coincided with increases in densities of fish fry (Power 1 990a) and predatory invertebrates from 22 June to 16 July. More odonates (Fig. 4c) (primarily lestid damselfly nymphs) and megalopteran larvae (Fig. 4d) were collected from cores in fish enclosures than in enclosures or the open river (Power 1990a). By 16 July, odonate densities were significantly different among treatments, and megalopteran densities were nearly so (Table 2). Fry of roach and stickleback, which by early July had appeared in the open river, recruited to enclosures but not to enclosures (Power 1990a).

Densities of large (head capsule width >0.5 mm)

FIG. 4. Densities of insect larvae in turf-covered boulder-bedrock and gravel substrates: (a) chironomids in turfs (individuals per unit area): (b) chironomids in turfs (individuals per unit algal biomass); (c) odonates (mostly lestid damselfly nymphs) in turf; (d) megalopterans (mostly Sialis sp.) in turf; (e) total ephemeropterans in turf: (Q large (head width >0.5 mm) ephemeropterans in turf; (g) chironomids in gravel; (h) plecopterans in gravel; (i) odonates in gravel; (j) megalopterans (mostly Sialis sp.) in gravel; (k) total ephemeropterans in gravel; (1) large (head width >0.5 mm) ephemeropterans in gravel. Bars are ? 1 SE. Note differences in scale on the y axes.


Turf Biota Gravel Biota 120000- 10000

a Chironomids 8000 9 Chironomids E 100000 -E 80

- 80000- 6000 - fish

S 60000- -- no fish

> 40000- */'/ > u Vo 4000 - -- - - open

2 0000 - >

C: 20000 X- , 2000

0 4 5 Jun 22 Jun 16 Jul 5 Jun 22 Jun 24 Jul

1000 -

60 b Chironomids h Pegaoptera 2) 5000 E 750 -

40-

3500 - 2 500 -

20 - >ag peeotr 00

> ~~~~~~~~~v250-

0 0 5 Jun 22 Jun 16 Jul 5 Jun 22 Jun 24 Jul

300 2500

C\1 C Odonaes i Odonates E 200ate E 2000

0i 1500 Ci) ~~~~~~~~~~~e-~

V001 -0 1000 >100 >

- ~~~~~~~~~ ~~5QQ -

0 - 0 -

5 Jun 22 Jun 16 Jul 5 Jun 22 Jun 24 Jul

d ~~~~~~~~~~~750- E 200 Meaopea1j Megaloptera -~~~~ 4 ~~~~E

500

5-0 100/25

/~~~~

5 Jun 22 Jun 16 Jul 0 5 Jun 22 Jun 24 Jul

25000- 3000-

C e Total Ephemeroptera k Total Ephemeroptera E 20000- I - ~~~~~~~2000-

15000 I

V 10000 V--

> 1000 --

c 5000

0 *ln0 2 u 5 Jn 22 Jun 16 Jul 5 Jun 22 Jun 2 u

15000- 2500- f I ~~~~~~CjLarge Ephemeroptera E fLarge Ephemeroptera E 2000

10000- 1c 50

-~~~~ -, / >~~~~~~~~C 1000-

*~~~~ 5000 / V~~~~~~> 00 > 5000 ~~~~~~~~- 50

0 ~~~~~~~~~~~~~00


TABLE 3. Estimated dry biomass (mg/cm2) of algae and arthropods sampled in gravel and turfs on boulders and bedrock.

Fish No fish Open

Boulders Gravel Boulders Gravel Boulders Gravel

5 June 1989 (onset of experiment) Primary producers

Algae 25.77 23.90 26.90 Primary (1P) consumers

Mavflies 0.211 0.200 0.498 0.090 0.738 0.112 Caddisflies 0.118 0.061 0.141 0.032 0.046 0.017 Midges 0.063 0.010 0.083 0.008 0.117 0.004 Total 0.392 0.271 0.722 0.130 0.901 0.113

Secondary (20) consumers Odonates 0.000 0.100 0.000 0.200 0.000 0.000 Plecopterans 0.002 0.125 0.015 0.002 0.009 0.021 Megalopterans 0.000 0.0004 0.0004 0.001 0.000 0.001 Total 0.002 0.225 0.015 0.203 0.009 0.022

Resource/consumer ratios Algae/ I cons. 73.39 0.74-1.48 29.08 1.54-3.08 29.86 1.77-3.54 1?/'? cons. 196.00 1.20 46.88 0.64 100.11 6.04

22 June 1989 (midway through experiments) Primary producers

Algae 12.02 11.20 14.91 Primary (1?) consumers

Mayflies 0.460 0.173 0.371 0.153 1.394 0.100 Caddisflies 0.046 0.013 0.010 0.023 0.065 0.052 Midges 0.340 0.010 0.420 0.010 0.420 0.004 Total 0.432 0.196 0.801 0.186 1.879 0.156

Secondary (2?) consumers Odonates 0.0051 0.055 0.051 0.064 0.008 0.073 Plecopterans 0.003 0.009 0.002 0.013 0.0004 0.019 Megalopterans 0.0002 0.011 0.000 0.009 0.003 0.010 Total 0.052 0.075 0.053 0.086 0.011 0.104

Resource/consumer ratios Algae/ IOcons. 27.82 1.02-2.04 14.00 1.08-2.15 7.94 1.28-2.56 10/2' cons. 8.31 2.61 15.71 2.16 170.81 1.50

16 July 1989 (end of experiment) Primary producers

Algae 3.03 0.20-0.40 17.59 0.20-0.40 8.83 0.20-0.40 Primasy (I1?) consumers

Mavflies 0.023 0.033 0.062 0.040 0.199 0.089 Amphipods 0.004 0.000 0.014 0.000 0.035 0.000 Caddisflies 0.020 0.001 0.049 0.003 0.093 0.030 Midges 0.096 0.005 0.010 0.009 0.033 0.003 Total 0.143 0.039 0.135 0.052 0.360 0.122

Secondary (20) consumers Odonates 0.072 0.200 0.597 0.000 0.000 0.100 Stoneflies 0.000 0.003 0.000 0.000 0.000 0.009 Alderflies 0.007 0.044 0.029 0.027 0.004 0.007 Total 0.079 0.247 0.626 0.027 0.004 0.116

Resource/consumer ratios Algae/l Icons. 21.18 5.13-10.26 130.30 3.85-7.69 24.53 1.64-3.28 11?20 cons. 1.81 0.16 0.22 1.93 90.00 1.052

mayflies increased from 5 to 22 June in fish enclosures (Fig. 4f) and were significantly higher than in enclosures on that date (Table 2, Dunn's multiple comparison test indicated significant differences between enclosures and enclosures [P < .01]). Subequently, mayflies decreased in all cages and the open river (Fig. 4e, f). Decreases of mayflies in turf-covered boulders

of enclosures coincided with the buildup of predatory invertebrates in these cages.

These results, and results from separate experiments showing that small predatory insects and fish fry can cause fourfold reductions in the numbers of midges colonizing C(ladophora (Power 1990a), indicate that roach and steelhead, by suppressing small predators,


1200

C\jE 1000 No fish E 80 l|Fish E 800-

- 600

O 400-

200-

0 T U (n, X E c 0) Ci) E * _ T o a z E co- P - E E c - E ' s 0 , oP~~ ~ oC) 0~~% G) ci C 01) Z>0oN~~c3-~,- -~

00 c-C0_N

CZ

< (I) c

FIG. 5. Taxonomic composition and abundance of algae in samples collected from I1-cm2 areas aspirated on three cobbles from each cage. Bars are means with 1 SE of averages from six enclosures and six exciosures. Generic names on the x axis are sequentially cyanobacteria (Nostoc to Schizothrix), chiorophytes (Aphanochacte to ,4nkistrodesmus), and diatoms (Epi- themnia to 4 mphipleura); "bgs" =blue-green algae (cyanophytes); "ig" =large ( >20 ,um diameter); "g" =green algae (chlorophytes); "sml" =Usmall (<20 gm diameter).

release algivorous chironomids, which in turn suppress algae. During the biologically active summer season, 96-98% of the plant biomass and 70-84% of the ar- thropod primary consumer biomass in pools of the South Fork Eel were associated with turfs on boulder- bedrock substrates (Table 3). Because pools comprise 60-80% of the wetted river area during summer, these food chain dynamics influence the river community at large.

Impacts of fish feeding on gravel hiota

In contrast to the strong, cascading effects of fish on the invertebrate and algal turfs of boulders and bedrock, few fish effects on the biota of gravel were apparent. Algal standing crops on gravel particles were scant, and their taxonomic compositions were similar in stocked and unstocked enclosures at the end of the experiment (Fig. 5). Of all taxa examined, only differences in Nostoc between treatments approached significance (P = .003 from a t test, NS at the critical level of.002 computed from the Bonferroni criterion for the 26 taxa compared). Nostoc on gravel in fish enclosures was probably sloughed from the copious Nostoc that overgrew Cladophora turfs on boulders in the absence of fish (Power 1 990a). Most of the "epilithic" diatom cells counted had a senescent or empty appearance, suggesting that they were also detrital outfall from healthier assemblages of the same taxa found growing epiphytically on Cladophora (J. Marks, personal com- m1utn1icatiOni).

Mayflies and predatory invertebrates showed no significant differences in abundance in gravel cores sampled from enclosures and enclosures throughout the experiment (Fig. 4h-1, Table 2). Similarly, densities of chironomid larvae in gravel cores also showed no sig-

nificant treatment effects (Fig. 4g), although there were significantly more chironomid tubes (occupied and empty) on the top surfaces of stones collected from enclosures than from enclosures (P = .02, Mann-Whit- ney U, Fig. 6). (These results are unchanged when tube counts are normalized for the surface area of stones.) Densities of chironomid tubes on stone surfaces may have been higher in enclosures because fish suppressed foraging by heptageniid mayflies. More mayflies were counted on top surfaces of stones in enclosures than in enclosures during three daytime censuses (P = .02, Mann-Whitney U, Fig. 7a). When entire stones were netted, however, the numbers of heptageniids and other large, mobile insects clinging to stone surfaces did not differ between enclosures and enclosures (Fig. 7b).

Chironomid Tubes on Stones

15 -

ci) C 0

10.

al) X5 oXXX

a 5

Fish No Fish FIG. 6. Numbers of chironomid tubes counted on stones

32-64 mm median diameter collected from enclosures and exciosures. Bars are means (with I SE) Of Six averages per treatment of counts from three stones per cage. Stones were collected systematically at evenly spaced, predetermined points along longitudinal transects.


1.2 a 0.8

0 ) 0 .5 0.8-

000

Fish No Fish

4.-

b El Stocked

0Unstocked

00

0 2

CZ

0 7-C

D O S S / S S w Sw -j ._~~~~1 CZ4 - a .

FIG. 7. (a) Number of mayflies (mostly heptagenjids) sighted on stones on the beds of enclosures and exclosures. Bars are means (with I SE) Of Six averages per treatment of counts on 3 2-34 stones per pen. Observations were made on every other stone >28 mm median diameter that lay along a longitudinal transect from the downstream to the upstream end of each pen. (b) Numbers of individuals clinging to cobbles (64-128 mm median diameter) netted from enclosures and exelosures. Bars show means with I SE for the taxa captured on netted cobbles: "Hept" =heptageniid mayflies-, "Siph" =siphlonurid mayflies-, "Baet" =baetid mayflies; "Ephem" =ephemerellid mayflies-, "Plec" =plecopterans, "Lepid"=

Lepidostoma (caddisfly)-, and "Coena" =coenagrionid dam-

7Trophic /cewl hiotnass in graw'c and tw.-fi on b)oudders in rit'er poolS

Dry biomass of attached algal turfs decreased in var- ious treatments from ;25 mg/cm2 at the onset of experiments on 5 June to 3-18 at the end of the experiment in md -July (Table 3). The bio a f epilithic algae on gravels in enclosures, exelosures, and open sites in pools appeared similar and very scant throughout the 6-wk experiment, but was only sampled at the endpoint in mid-July, when standing crops (dry masses) were estimated from cell counts to range from 0.17-0.36 mg/cm2 . Algal standing crops were, in mid- July, one to two orders of magnitude greater on boulder-bedrock substrates than on pool gravel.

Epibenthic primary consumers associated with algal turfs on boulders (mayflies, caddisflies, chironomids,

amphipods) had a collective biomass estimated at 0.4- 0.9 mg/cm2 at the onset of the experiment, and at 0. 1- 0.4 mg/cm2 at the end. The collective biomass of these algivore-detritivores in gravel was estimated at 0.1- 0.3 mg/Cm2 at the onset of experiments, and from 0.04- 0. 1 mg/cm2 at the end. Primary consumer biomass, averaged across all treatments, was 4 times higher on turf-covered boulders than in pool gravel in early June, 6.5 times higher in late June, and 3 times higher by mid-July (Table 3).

Epibenthic carnivores (odonates, megalopterans, and plecopterans) had a collective biomass estimated at 0.002-0.015 mg/cm2 in turfs on boulders in early June (Table 3). In gravel, this guild had a higher estimated biomass (0.02-0.23 mg/cm2), largely composed of large dragonfly nymphs, which did not occur in turfs. In contrast to primary consumer biomass, collective biomass of predatory invertebrates was estimated to be 17 times higher in gravel than on boulders in early June. By late June, predatory invertebrate biomass was only twice as high in gravels. By mid-July, this biomass distribution was reversed. Averaged across all treatments, predatory invertebrate biomass was twice as high on turf-covered boulders as in gravel, due to the increase ofthese predators within fish exclosures (Table 3).

Resource: consumer biomass ratios were generally higher for boulder-bedrock than for gravel-dwelling biota, suggesting that turf-dwelling consumers might be less food limited. In early June, ratios of algal biomass to that of primary consumers were 30-65 for turf- covered boulders, and 0.6-3.2 for gravel biota (assuming similar algal standing crops in gravels to those estimated for mid-July) (Table 3). By the end of the experiment, these ratios ranged from 2 1-130 for turf- dwelling organisms, and from 1.4-9.2 for biota associated with gravel. Over the course of the experiment, these resource: consumer ratios were 1-2 orders of magnitude higher in turfs than in gravels. In early June, ratios of primary consumers to epibenthic carnivores ranged from 46-196 for turf fauna, and from 1-5 for gravel fauna. By mid-July, these ratios ranged from 0.2-1.8 for turf fauna and 0.2-1.9 for gravel fauna (Table 3).

The proportion of the biota in river pools associated with gravel vs. tufts on boulders and bedrock was estimated by multiplying standing crops in open pool sites by the proportion of pool area occupied by these substrates. These computations suggest that in June, 98% of the algal, 84% of the invertebrate primary consumer, and 24% of the invertebrate primary carnivore biomass was associated with turfs on boulders. By mid- July, these estimates were 96, 70, and 3%, respectively.

DISCUSSION

Predatory effects of fish were strong on boulder-bedrock substrates and weak in gravel substrates in pools of the South Fork Eel. These results parallel results


from marine benthic systems, in which the effects of epibenthic predators documented for rocky intertidal and unvegetated soft-sediment habitats are not conspicuous in seagrass beds (Peterson 1979, Wilson 1991). Spatial variation in predator impacts can lead to spatial variation in the number of functionally significant (sen- su Frctwell 1977) trophic levels in benthic food webs, even without loss of species from top trophic levels. In boulder-bedrock microhabitats of the South Fork Eel, fish were functionally significant as fourth-level consumers in a food chain. By suppressing predatory invertebrates and fish fry, fish released algivorous chironomids. which in turn lowered standing crops of attached algae. As predicted by food chain dynamics theory (Hairston et al. 1960, Fretwell 1977, 1987, Ok- sanen et al. 1981), standing crops of algae on boulders and bedrock were lower in fish enclosures with four trophic levels than in fish enclosures with three trophic levels (Power 1990a).

Fish also fed on gravel substrates immediately adjacent to boulder-bedrock formations, but they had no detectable effects there on invertebrate or algal standing crops. Differences between boulder-bedrock and gravel with respect to isolation, cover, seasonal disturbance, productivity, and local prey characteristics may con- tribute to shifts in the functional significance of fish between the two microhabitats.

IIa!itat isolation and (age effects

Predator impacts on prey in patches, or cages, depend on rates of prey depletion relative to prey immigration rates (e.g., Holt 1984, T. Oksanen 1990, Cooper et al. 1990). In pools of the South Fork Eel, boulders and bedrock emerge as habitat islands from the gravel bed. In addition, turf-dwelling invertebrates in rivers are often weak swimmers (Hynes 1970), and, in lentic pools, would not be passively dispersed by flow. Therefore, immigration by turf-dwelling biota to builders in enclosures might be less likely to offset losses to fish predation than immigration by gravel-dwelling biota from adjacent gravel outside cages.

Features of cages that impede migration of stream insects, such as small-mesh walls or solid floors, can artificially exaggerate effects of enclosed predators (Frid and James 1988, Cooper et al. 1990, Wilson 1991). Cooper et al. (1990) found that studies using cages with 3-mm mesh were more likely to detect significant fish effects on invertebrates than studies with cages with 6-mm mesh. During the summer period of roach and stickleback recruitment in the South Fork Eel, 6-mm mesh walls would have permitted passage of fish that within a few weeks would have been capable of at- tacking larger insects (head capsule width >0.5 mm). This would also prevent detection of fish effects among treatments. In turfs on boulders, densities of large insect taxa such as damselfly nymphs increased in exclosures, indicating that these insects were not impeded by 3-mm mesh walls, either because they could cross

them, or because subgravel or aerial dispersal pathways (Williams and Hynes 1976) were available to them.

Cages also prevent fish from moving among sites to feed, and can increase their feeding impacts, even when the densities and per capita feeding rates of enclosed fish are within the range observed for free-ranging fish in similar microhabitats (Butler 1 989). Inferences can be drawn about the importance of these effects in the South Fork Eel from observations of insect and algal densities in enclosures and enclosures compared with densities in open sites. At the end of the experiment, algal standing crops sampled in open sites were intermediate between standing crops sampled in enclosures and enclosures (Fig. 3). The densities of chironomids were similar in open sites and enclosures, and significantly higher in enclosures; these differences were par- ticularly apparent when chironomid densities were normalized for algal standing crops, which better re- flects habitat available for these small organisms than planar area (Fig. 4a, b). In contrast, the densities of predatory odonates and megalopterans sampled in open sites were more similar to densities in fish enclosures than in enclosures (Fig. 4d, e). The second result sug- gests that invertebrate predators can be depressed to similar densities by fish predation at ambient and enclosed (and possibly elevated) levels. Why, then, were chironomid densities so low in open sites, given the low measured densities of invertebrate predators? A likely explanation is that additional small predators, in particular fish fry, which could not be sampled with cores, also preyed on midges in open habitats (Power 1 990a). In addition, densities of both midges and predatory invertebrates in the study reach at large were probably not accurately estimated by the limited number of samples collected from open sites in the vicinity of the pens. Larger scale transect surveys and visual reconnaissance have shown that during years of algal blooms, Cladophora in the main channel of the South Fork Eel becomes densely infested with tuft-weaving midges, and that these infestations precede the alga's decline (Power 1990a, h). If levels of fish predation were elevated in enclosures relative to open-channel habitats, their impacts on midges and algae may be accelerated. Nevertheless, these experimental results support the hypothesis that fish predation on small, intermediate predators brings about the eventual infestation of macroalgae by midges in the open river channel.

Ph Bsical refuges from predation A second explanation for the failure to detect similar

fish effects on gravel biota is that refuges provided by interstitial spaces in gravels greatly reduce foraging efficiency of fish (Coleman and Hynes 1970, Hynes 1970, Williams and Hynes 1974, Brusven and Rose 1981, Hemphill and Cooper 1984, Minshall 1984, Williams 1984, but see Reice 1983, Flecker and Allan 1984). Cover provided by real or simulated macrophytes also


reduced predation by fish on invertebrates in soft-bot- tomed littoral freshwater habitats (Crowder and Coo- per 1982, Savino and Stein 1982, Gilinsky 1984, Her- shey 1985) and marine habitats (Reise 1977, Peterson 1982, Virnstein et al. 1983, Robertson 1984, Sum- merson and Peterson 1984, Wilson 1991). In the South Fork Eel, however, invertebrates in macroalgae on boulders, except for algae-weaving midge larvae, were more depleted by fish than were invertebrates in gravel.

Differential prey vzlnerahilitv

Retreats constructed from macroalgae by tuft-weaving midges defend them from fish but not invertebrate predators. In field and laboratory feeding trials, we have found that silk-reinforced algal tufts afford chironomid larvae apparently complete protection against fish, but not against predatory invertebrates (lestids, aeshnids, and naucorids) (Power et al. 1992). Despite their heavy consumption of algivorous mayflies (Fig. 2), fish had negative indirect effects of algae on boulders because their predation on small predators released tuft-weaving midges, a key herbivore guild (Power 1990a).

Many temperate freshwater fishes, like roach and steelhead in the South Fork Eel, feed opportunistically at several trophic levels (Vadas 1990). Feeding impacts of potentially omnivorous fish are, however, con- strained by the size, morphology, or behavior of their prey (Drenner et al. 1978, Werner et al. 1983, Osenberg and Mittelbach 1989). In a southern California stream, trout suppressed large, mobile invertebrate predators, while more cryptic, gravel-dwelling insects showed ten- dencies to increase in the presence of trout (Hemphill and Cooper 1984, Cooper 1988). Several other studies in freshwater ponds have shown that fish predation on larger, more conspicuous, or more active taxa such as odonates (Crowder and Cooper 1982, Menzie 1978 cited in Gilinsky 1984) and predatory chironomids (Cuker 1981, Gilinsky 1984) indirectly increased populations of herbivorous chironomids.

Fish effects may not have been detected in gravel microhabitats in the South Fork Eel because prey there were vulnerable to both top and intermediate predators, so that invertebrates and fish fry in enclosures had impacts similar to those of fish in enclosures. This hypothesis is supported by the increase of large mayflies in enclosures from 5 June to 22 June, followed by their subsequent decline in enclosures in mid-July after odonate and megalopteran densities had built up (Fig. 4f).

Disturbance, productivity, and the energy I1aSC of/gravel vs. boulder wnicrohabitats

(ladophora, the dominant macroalga in northern California rivers, is scoured from all substrates by winter floods. Spring regrowth appears to be vegetative from basal pads, as has been documented for Great

Lakes Cladophora recovering from ice scour (Blum 1982). Cladophora in the South Fork Eel proliferates in the spring on boulder and bedrock substrates, but not on more severely scoured gravels (Power 1990a, c). As Cladophora proliferates, so do highly nutritious diatoms that grow epiphytically on the macroalga. In contrast, the darker, stagnant, silty gravel beds of pools support low standing crops of algae and invertebrates (Fig. 4, Table 3), suggesting that gravel beds of pools are more seasonally disturbed and less productive than larger substrates.

Diatoms sampled from the gravel bed appeared senescent or empty, and may have been derived in large part from epiphytes (with which they overlapped tax- onomically) sloughed from turf-covered boulders (J. Marks, personal comnmnunication). Other indirect evi- dence that diatoms were not actively growing on pool gravels came from their lack of response to mayfly activity. Although behavioral depression of mayflies by fish may have promoted the persistence of chironomid tubes on top surfaces of cobbles in fish exclosures, there was no corresponding effect on epilithic diatoms, which were equally abundant in enclosures and exclosures. Also, despite their diurnal avoidance of stone surfaces, mayflies in enclosure gravel were as abundant as mayflies in exclosures, suggesting that they could feed on detritus that infiltrated gravel interstices. If primary consumers fed primarily on infiltrating detritus, they could remain within gravel interstices where they would be less available to fish. Low standing crops of prey, coupled with the lower per capita availability of prey near or in interstitial refuges, may have depressed foraging profitability of gravel substrates for fish below the threshold required for fish to impose significant impacts, as predicted by Fretwell (1977, 1987) and Oksanen et al. (1981).

Changes in trophic control oi'er sinall spatial scales

This study and others (Power 1984, 1987, Power et al. 1989, Schlosser and Ebel 1989, Peckarsky et al. 1990) demonstrate abrupt change over small spatial scales (centimetres) in the functional importance of predators in rivers. Prey characteristics and microhabitat isolation, cover, disturbance, and productivity remain to be investigated as factors that separately or in combination alter control by predators of lower trophic levels in the Eel River. Shifts in impacts of predators over small spatial scales provide opportunities to study the constraints placed on predators by their habitats. We need to know more about these constraints if we are to predict where and when trophic interactions will be strong (Menge and Sutherland 1976, Paine 1980, Hansson 1989, Menge and Olson 1990).

ACKNOWLEDGMENTS I thank S. Kupferberg, J. Marks, J. Nielsen, and B. Rainey

for help in the field and the laboratory, P. Kotanen, W. Sousa,


and an anonymous reviewer for statistical advice, and B. Diet- rich, T. Dudley, P. Kotanen, L. Oksanen, W. Sousa, and T. Wootton for valuable comments on the paper.

LITERATURE CITED

Anderson, S., K. L. Markham, L. F. Trujilow, W. F. Shelton, and D. A. Grillo. 1987. Water resources data for Cali- fornia, Water Year 1985. 2. Pacific Slope basins from Ar- royo Grande to Oregon State line except Central Valley. United States Geological Survey, Sacramento, California, USA.

Blum, J. L. 1982. Colonization and growth of attached algae at the Lake Michigan water line. Journal of Great Lakes Research 8:10-15.

Brown, L. R., and P. B. Moyle. 1989. Eel River Survey: third year studies. Report to the California Department of Fish and Game. University of California, Davis, California, USA.

Brusven, M. A., and S. T. Rose. 198 1. Influence of substrate composition and suspended sediment on insect predation by the torrent sculpin, C'oitus rhotheus. Canadian Journal of Fisheries and Aquatic Sciences 38:1444-1448.

Butler, M. J. 1989. Community responses to variable predation: field studies with sunfish and freshwater macroinvertebrates. Ecological Monographs 59:311-328.

Caswell, H. 1978. Predator-mediated coexistence: a nonequilibrium model. American Naturalist 112:127-154.

Chesson, J. 1978. Measuring preference in selective predation. Ecology 61:44-49.

Coleman, M. J., and H. B. N. Hynes. 1970. The vertical distribution of the invertebrate fauna in the bed of a stream. Limnology and Oceanography 15:31-40.

Cooper. S. D. 1988. Responses of aquatic insects and tad- poles to trout. International Vereinigung fir theoretische und angewandte Limnologie. Verhandlungen 23:1698-1703.

Cooper, S. D., S. J. Walde, and B. L. Peckarsky. 1990. Prey exchange rates and the impact of predators on prey populations. Ecology 71:1503-15 14.

Crowder, L. B., and W. E. Cooper. 1982. Habitat structural complexity and the interaction between bluegills and their prey. Ecology 63:1802-1813.

Crowleyv P. H. 1981. Dispersal and the stability of predator- prey interactions. American Naturalist 118:673-701.

Cuker, B. E. 198 1. Control of epilithic community structure in an arctic lake by vertebrate predation and invertebrate grazing. Dissertation. North Carolina State University, Ra- leigh, North Carolina, USA.

Dietrich, W. E., J. W. Kirchner, H. Ikeda, and F. Iseya. 1989. Sediment supply and the development of the coarse surface layer in gravel-bedded rivers. Nature 340:215-217.

Drenner, R. W., J. R. Strickler, and W. J. O'Brien. 1978. Capture probability: the role of zooplankter escape behavior in the selective feeding of planktivorous fish. Journal of the Fisheries Research Board of Canada 35:1370-1373.

Dudley. T. L., S. D. Cooper, and N. Hemphill. 1986. Effects of macroalgae on a stream invertebrate community. Journal of the North American Benthological Society 5:93-106.

Flecker, A. S., and J. D. Allan. 1984. The importance of predation, substrate and spatial refugia in determining lotic insect distributions. Oecologia (Berlin) 64:306-3 13.

Fretwell. S. D. 1977. The regulation of plant communities by food chains exploiting them. Perspectives in Biology and Medicine 20:169-185.

1987. Food chain dynamics: the central theory of ecology? Oikos 50:291-301.

Frid, C. L. J., and R. James. 1988. The role of epibenthic predators in structuring the marine invertebrate commu- nitv of a British coastal salt marsh. Netherlands Journal of Sea Research 22:307-3 14.

Gilinsky, E. 1984. The role of fish predation and spatial heterogeneity in determining benthic community structure. Ecology 65:455-468.

Hairston, N. G., F. E. Smith, and L. B. Slobodkin. 1960. Community structure, population control, and competition. American Naturalist 94:421-425.

Hansson, L. 1989. Predation in heterogeneous landscapes: how to evaluate total impact? Oikos 54:117-119.

Hart, D. D. 1978. Diversity in stream insects: regulation by rock size and microspatial complexity. Vereinigung fur thcoretische und angewandte Limnologie, Verhandlungen 20:1376-1381.

Hastings, A. 1977. Spatial heterogeneity and the stability of predator-prey systems. Theoretical Population Biology 12:37-48.

Hemphill, N., and S. D. Cooper. 1984. Differences in the community structure of stream pools containing or lacking trout. Vereinigung fur theoretische und angewandte Lim- nologie, Verhandlungen 22:1858-1861.

Hershey, A. E. 1985. Effects of predatory sculpin on the chironomid communities in an arctic lake. Ecology 66:113 1- 1138.

Holt, R. D. 1984. Spatial heterogeneity, indirect interactions, and the coexistence of prey species. American Nat- uralist 124:377-406.

1985. Population dynamics of two-patch environ- ments: some anomalous consequences of an optimal habitat distribution. Theoretical Population Biology 28:18 1-208.

Huffaker, C. B. 1958. Experimental studies on predation: dispersion factors and predator-prey oscillations. Hilgardia 27:343-383.

Huryn, A. D., and J. B. Wallace. 1987. Local geomorpho- logy as a determinant of macrofaunal production in a mountain stream. Ecology 68:1932-1942.

Hynes, H. B. N. 1970. The ecology of running waters. Uni- versity of Toronto Press, Toronto, Ontario, Canada.

Krebs, C. 1989. Ecological methodology. Harper and Row, New York, New York, USA.

Lagler, K. F., J. E. Bardach, R. R. Miller, and D. R. M. Passino. 1977. Ichthyology. Wiley, New York, New York, USA.

Lemmon, P. E. 1957. A new instrument for measuring forest overstory canopy. Journal of Forestry 55:667-668.

McAuliffe, J. R. 1984. Competition for space, disturbance, and the structure of a benthic stream community. Ecology 65:894-908.

Menge, B. A., and A. M. Olson. 1990. Role of scale and environmental factors in regulation of community structure. Trends in Ecology and Evolution 5:52-57.

Menge, B. A., and J. P. Sutherland. 1976. Species diversity gradients: synthesis of the roles of predation, competition, and temporal heterogeneity. American Naturalist 110:351- 369.

Minshall, G. W. 1984. Aquatic insect-substratum relationships. Pages 358-400 in V. H. Resh and D. M. Rosenberg, editors. The ecology of aquatic insects. Plenum, New York, New York, USA.

Oksanen, L., S.D. Fretwell,J. Arruda, and P. Niemela. 1981. Exploitation ecosystems in gradients of primary productivity. American Naturalist 118:240-261.

Oksanen, T. 1990. Exploitation ecosystems in heterogeneous habitat complexes. Evolutionary Ecology 4:220-234.

Osenberg, C. W., and G. G. Mittelbach. 1989. Effects of body size on the predator-prey interaction between pump- kinseed sunfish and gastropods. Ecological Monographs 59: 405-432.

Paine, R. T. 1980. Food webs: linkage, interaction strength and community infrastructure. Journal of Animal Ecology 49:667-685.

Peckarsky, B. L., S. C. Horn, and B. Statzner. 1990. Stonefly


predation along a hydraulic gradient: a field test of the harsh-benign hypothesis. Freshwater Biology 24:18 1-191.

Peterson, C. H. 1979. Predation, competitive exclusion, and diversity in the soft-sediment benthic communities of es- tuaries and lagoons. Pages 233-264 in R. J. Livingston, editor. Ecological processes in coastal and marine systems. Plenum, New York, New York, USA.

1982. Clam predation by whelks (Busycon spp.): experimental tests of the importance of prey size, prey density, and seagrass covfr. Marine Biology 66:159-170.

Pimm, S. L. 1982. Food webs. Chapman and Hall, London, England.

Power, M. E. 1984. Depth distributions of armored catfish: predator-induced resource avoidance? Ecology 65:523-528.

1987. Predator avoidance by grazing fishes in temperate and tropical streams: importance of stream depth and prey size. Pages 333-352 in W. C. Kerfoot and A. Sih, editors. Predation: direct and indirect impacts on aquatic communities. University Press of New Hampshire, Han- over, New Hampshire, USA.

1990a. Effects of fish in river food webs. Science 250:411-415.

1990b. Benthic turfs versus floating mats of algae in river food webs. Oikos 58:67-79.

* 1990c. Seasonal and hydrologic controls of algal blooms in Northern California rivers. Technical comple- tion report. Water Resources Center, University of Cali- fornia at Riverside, Riverside, California, USA.

Power, M. E., T. L. Dudley, and S. D. Cooper. 1989. Grazing catfish, fishing birds, and attached algae in a Panamanian stream. Environmental Biology of Fishes 26:285-294.

Power, M. E., J. C. Marks, and M. S. Parker. 1992. Variation in the vulnerability of prey to different predators: community-level consequences. Ecology 73, in press.

Power, M. E., and A. J. Stewart. 1987. Disturbance and recovery of an algal assemblage following flooding in an Oklahoma stream. American Midland Naturalist 117:333- 345.

Reice, S. R. 1983. Predation and substratum: factors in lotic community structure. Pages 325-345 in S. Bartell and T. Fontaine, editors. Dynamics of lotic ecosystems. Ann Ar- bor Science, Ann Arbor, Michigan, USA.

Reise, K. 1977. Predator exclusion experiments in an intertidal mud flat. Helgolhnder wissenshaftliche Meeresun- tersuchungen 30:263-27 1.

Robertson, A. I. 1984. Trophic interaction between the fish fauna and macrobenthos of an eelgrass community in West- ern Port Victoria. Aquatic Botany 18:135-153.

Savino, J. F., and R. A. Stein. 1982. Predator-prey inter-

actions between largemouth bass and bluegills as influenced by simulated submersed vegetation. Transactions of the American Fisheries Society 111:255-266.

Schlosser, I. J., and K. K. Ebel. 1989. Effects of flow regime and cyprinid predation on a headwater stream. Ecological Monographs 59:41-57.

Smith, F. E. 1972. Spatial heterogeneity, stability, and diversity in ecosystems. Transactions of the Connecticut Academy of Arts and Sciences 44:309-335.

Smock, L. A. 1980. Relationships between body size and biomass of aquatic insects. Freshwater Biology 10:375-383.

Sousa, W. P. 1979. Disturbance in marine intertidal boulder fields: the nonequilibrium maintenance of species diversity. Ecology 60:1225-1239.

Stevenson, R. J., and E. F. Stoermer. 1982. Seasonal abundance patterns of diatoms on Cladophora in Lake Huron. Journal of Great Lakes Research 8:169-183.

Summerson, H. C., and C. H. Peterson. 1984. Role of predation in organizing benthic communities of a temperate- zone seagrass bed. Marine Ecology Progress Series 15:63- 77.

Vadas, R. L., Jr. 1990. The importance of omnivory and predator regulation of prey in freshwater fish assemblages of North America. Environmental Biology of Fishes 27: 285-302.

Virnstein, R. W., P. S. Mikkelsen, K. D. Cairns, and M. A. Capone. 1983. Seagrass beds versus sand bottoms: the trophic importance of their associated benthic invertebrates. Florida Scientist 46:363-38 1.

Werner, E. E., J. F. Gilliam, D. J. Hall, and G. G. Mittelbach. 1983. An experimental test of the effect of predation risk on habitat use in fish. Ecology 61:233-242.

Williams, D. D. 1984. The hyporheic zone as a habitat for aquatic insects and associated arthropods. Pages 430-455 in V. H. Resh and D. M. Rosenberg, editors. The ecology of aquatic insects. Plenum, New York, New York, USA.

Williams, D. D., and H. B. N. Hynes. 1974. The occurrence of benthos deep in the substratum of a stream. Freshwater Biology 4:233-256.

Williams, D. D., and H. B. N. Hynes. 1976. The recoloniza- tion mechanisms of stream benthos. Oikos 27:265-272.

Wilson, W. H. 1991. Competition and predation in marine soft-sediment communities. Annual Review of Ecology and Systematics 21:221-241.

Winberg, G. G. 1971. Symbols, units and conversion factors in studies of fresh water productivity. International Bio- logical Program Cable Printing Service, London, England.

Zar, J. 1984. Biostatistical analysis. Prentice-Hall, Engle- wood Cliffs, New Jersey, USA.

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