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Trophic interactions among detritus, benthic midges, and predatory fish in a freshwater marsh.


New concepts have been proposed to explain the relative influences of resource limitation and predation on food web dynamics. Early models (Hairston et al. 1960, Menge and Sutherland 1976, Wiens 1977) suggested that competition for resources or predation can influence a particular trophic level, but that one of these mechanisms will be dominant. Argument over the relative importance of each mechanism was largely resolved by acknowledging that resources and predation may each be important at different times or places, or to different species within overall food webs (Sih et al. 1985, Schoenly and Cohen 1991, Hunter and Price 1992, Belovsky and Joern 1995, Polls et al. 1996). A new twist to the top-down vs. bottom-up discussion has been developed by Rosemond et al. (1993) and Osenberg and Mittelbach (1996). They demonstrated that not only can both resource limitation and predation affect a trophic level, but those mechanisms can exert significant influences simultaneously. Belovsky and Joern (1995) argue that while many mechanisms may operate in concert, only one will exert regulation. My experiments in a western New York marsh further develop those themes by simultaneously evaluating the influences of plant litter resources (bottom-up) and fish predation (top-down) on benthic midge distribution and abundance.

While multi-level trophic interactions have been investigated extensively in a diversity of aquatic habitats, including lakes (Carpenter et al. 1987, Bronmark et al. 1992, Diehl 1995, Osenberg and Mittelbach 1996), ponds (Hall et al. 1970, Hambright et al. 1986, Warren 1989, Diehl 1992), rivers (Power et al. 1996), streams (Rosemond et al. 1993), and tree holes (Kitching 1987), food web interactions in marshes are poorly understood. Macroinvertebrates hold a central position in marsh food webs, yet little is known about what factors limit their populations (Batzer and Wissinger 1996). Several characteristics of marshes are likely key influences on the trophic dynamics of their macroinvertebrate faunas. In these habitats, detritus is an important basal energy source for food webs (Odum and Heald 1972, Rasmussen 1985, Murkin 1989, Neill and Cornwell 1992, Bunn and Boon 1993). Along benthic substrates in particular, the consumer community is typically dominated by detritivores such as midge larvae (Batzer and Wissinger 1996). In addition to providing food, plant litter in marshes may serve as valuable structural habitat for an assortment of invertebrates (Dvorak 1987, Campeau et al. 1994). Although plant litter is clearly important to marsh invertebrates, the large volumes occurring in marshes might suggest that this resource is seldom limiting. However, this hypothesis remains largely unrested. Rasmussen (1985) found that supplementing micro-detrital foods increased growth rates but not densities of marsh midges. Neckles et al. (1990) experimentally removed plant litter from a seasonal marsh, but found that consumer abundance was not affected. In this study I tested whether plant litter supplies limited marsh midge abundance, and furthermore determined whether litter influenced midges via a trophic link or a refuge phenomenon.

Top-down interactions of predators have been thoroughly studied in many food webs (see Kerfoot and Sih 1987, Carpenter et al. 1987, Cappuccino and Price 1995, Polis and Winemiller 1996). However, experiments addressing interactions of aquatic predators and consumers are lacking from marshes (Batzer and Wissinset 1996). Several factors might influence the functional significance of predation to marsh invertebrates. Strong (1992) proposed that top-down effects such as trophic cascades might be most prevalent in simple, species-poor habitats. This description applies to the bottom substrates of most freshwater marshes, where the benthic macroinvertebrate and fish communities typically consist of only a few taxa tolerant of anoxic conditions (Batzer and Resh 1992). Studies in the lake littoral (Wellborn and Robinson 1991, Osenberg and Mittelbach 1996) suggest that invertebrate communities commonly coexisting with fish can become relatively resistant to their predation. However, because marshes periodically dry and fish are temporarily eliminated, marsh invertebrates may be relatively inexperienced with fish predation and thus particularly susceptible. Additionally in North American marshes, the exotic fish Cyprinus carpio occurs frequently. The historically limited duration of interaction between native invertebrates and exotic carp might intensify their trophic interactions (Ebenhard 1988). In contrast, other features of marshes might lessen the strength of predator-prey interaction. Marshes can be particularly harsh environments for fish because oxygen levels plummet during hot periods of summer, and winter kill is common (in northern climates). Unlike the littoral regions of lakes, marsh fish cannot always migrate to deep water refugia during periods of environmental stress. The poor survival rate of marsh fishes may reduce their importance as top predators in marsh food webs in comparison to other fish-bearing lentic habitats (e.g., Bronmark 1994, Diehl 1995). In this study I manipulate both primary and secondary predatory fish numbers, and examine numeric responses of marsh midge populations.

Factors other than trophic interaction can influence food web structure and function. Winemiller's (1996) work in floodplain lakes suggests that temporal and spatial variation in food web structure can be affected by the integrated influences of habitat disturbance, ecological succession, and life history patterns, as well as predator-prey interaction. In floodplains and marshes, life history patterns might be of particular relevance. Fish and insect communities there are typically dominated by juveniles. Thus, the specific reproductive and colonization patterns of both consumers and predators may strongly influence food web dynamics in wetlands. In this study I examine temporal variations in the composition of the marsh community, and their ramifications for trophic dynamics.


Cayuga Marsh is located in western New York's Iroquois National Wildlife Refuge (INWR) (43 [degrees] N, 53 [degrees] w). In 1994, [similar to] 60% of its 30-ha surface was covered by emergent plants, primarily cattails (Typha spp.) and to a lesser extent bur-reed (Sparganium americanum). These plant stands were interspersed with numerous open-water pools. Submersed beds of Ceratophyllum demersum and algal metaphyton and floating patches of Lemna sp. were scattered throughout the site.

I conducted experiments in a 100 x 30 m open-water pool located near the center of Cayuga Marsh. At the high water mark in April 1994, depths in this pool ranged from 0.9 to 1.0 m. With the exception of a few man-made channels, this pool was among the deeper portions of Cayuga Marsh. By summer's end, water depths had receded to [less than] 0.6 m, largely from surface evaporation and plant transpiration. The study pool's substrate consisted of a relatively firm, clay-based mud. Although dense stands of emergent cattail and bur-reed surrounded the open water, few emergent plants grew in the pool itself. Moderate densities of submersed C. demersum occurred throughout the water column of the pool.

In 1991, 3 yr prior to this study, Cayuga Marsh had been intentionally drawn-down by INWR personnel, and kept dry over that summer. The site was reflooded the following spring of 1992, using water diverted from an adjacent marsh. An assortment of fish species characteristic of marsh environments were introduced with the flood waters. During the 1994 study season, I sampled the benthic fish community in the Cayuga Marsh study pool, using standard 6-mm wire-mesh Gee's minnow traps, with their funnel openings enlarged to 4 cm. I placed unbaited traps along benthic substrates of the study pool, and then retrieved them after 24 h. During May and early June 1994 I collected only adult brook stickleback (Culaea inconstans) and an occasional adult brown bullhead (Ictalurus nebulosus) [ILLUSTRATION FOR FIGURE 1 OMITTED]. In mid-June I captured large numbers of young-of-the-year carp (Cyprinus carpio) and bullhead. Following this initial peak, the numbers of juvenile carp and bullhead steadily declined over the rest of the summer. In late summer, I detected a smaller peak of young-of-the-year pumpkinseed sunfish (Lepomis gibbosus) [ILLUSTRATION FOR FIGURE 1 OMITTED].

Core sampling of mud substrates in the study pool over the 1994 season indicated that Chironomus plurnosus and Chironomus tentans midge larvae, and to a lesser extent Glyptotendipes barbipes larvae, accounted for [greater than] 95% of the overall numbers of benthic macroinvertebrates. Muscilium sp. fingernail clams and unidentified oligochaete worms were the only other benthic invertebrates commonly collected. In May, the midge larvae that occurred were mature fourth instars that had hatched the previous summer. These midges had already begun emerging in May, and most of that cohort had eclosed into adults by late June. The new 1994 cohort of midges first appeared in samples during June [ILLUSTRATION FOR FIGURE 2 OMITTED].

I placed bottle-traps (Riley and Bookbout 1990) that collect free-swimming invertebrates along benthic substrates of the pool for 24 h. However, they captured few nektonic invertebrates. When these samplers were placed higher up in the water column, they collected numerous invertebrates (Ephemeroptera, Odonata, Hemiptera, and Amphipoda). Thus, the lack of nektonic invertebrates collected along benthic substrates was not simply due to trap inefficiency, but suggests minimal use of the benthic substrates by nektonic species.

This sampling indicated that the animal community along the benthic substrates of this marsh pool was primarily comprised of midge larvae, an assortment of small insectivorous fish (mostly juvenile carp and brown bullhead, and to a lesser extent adult stickleback), and small numbers of adult brown bullheads. The coexistence of midges and this fish assemblage provided opportunity for them to interact trophically. Both juvenile and adults stages of common carp (Zur and Sarig 1980, Chapman and Fernando 1994) and brown bullhead (Ivlev 1961, Sinnott and Ringlet 1987, Kline and Wood 1996) are known to consume invertebrates readily, especially chironomid midge larvae. The degree of midge feeding by these fish species can be particularly pronounced in wetlands (Chapman and Fernando 1994) or the littoral zones of lakes (Kline and Wood 1996). While adult brown bullheads readily consume midge larvae, they also consume smaller fish, including conspecifics (Kline and Wood 1996). Thus, adult brown bullheads in Cayuga Marsh had the potential directly to influence midges through their feeding, or indirectly to influence midges by consuming other predators, such as small insectivorous fish. Detritivorous midge larvae (Coffman and Ferrington 1984) could also be influenced by bottom-up interactions with detritus. In Cayuga Marsh the primary source of detritus was litter from cattails, which can serve as both food and habitat for midges (see review of Batzer and Wissinger 1996).



Mesocosm treatments

I manipulated the trophic structure in the Cayuga Marsh study pool using 42 1.5 x 1.5 m square enclosure/exclosure mesocosms. Each consisted of a 12 mm diameter plastic-pipe frame with one of three wall types: (1) 12-mm mesh galvanized hardware cloth (24 replicates), (2) 3-mm mesh hardware cloth (12 replicates), and (3) open walls (i.e., corner pipes only; 6 replicates). Calibration studies indicated that the 12mm mesh restricted movement of coarse detritus and large fish, but not small fish, and the 3-mm mesh was a barrier to all fish except very small newly hatched larvae.

I erected the 42 mesocosms in the study pool during late April and early May 1994. I arranged them in two parallel rows, each with 21 mesocosms. Rows were separated by 5 m of open water, and within each row I maintained a 2-m buffer of open water between adjacent mesocosms. Each of the two rows was partitioned into three 7-mesocosm blocks. Each location block had 4 mesocosms with 12-mm mesh walls, 2 mesocosms with 3-mm mesh walls, and 1 mesocosm with no walls, arranged in a random order. Once erected, the sides of mesh-walled mesocosms (1.2 m high) penetrated - 10 cm into the benthic mud substrate. During May when water depths were at a maximum, walls extended 10-20 cm above the water surface.

By varying the mesh size of the walls and adding fish or litter to selected mesocosms, I was able to produce habitats of different trophic structure. Within the template of 42 mesocosms, I developed 7 treatments each with 6 replicates (one treatment/location block). The assemblage of large fish, small fish, and/or litter detritus differed among all treatments. Comparative detail on each of the 7 treatments is presented in Table 1, and hereafter I refer to the 7 treatments by the following names: (1) Ambient Controls, (2) Large-Fish Enclosures, (3) Large-Fish Exclosures or Caged Ambient Detritus, (4) Small-Fish Exclosures, (5) Small-Fish Enclosures, (6) Cattail-Litter Supplements, and (7) Nonnutritive-Litter Supplements.

For Large-Fish Enclosures, adult bullheads (the most abundant large fish) were collected in nearby portions of Cayuga Marsh during May, and were immediately transferred to mesocosms. Similarly, for Small-Fish Enclosures, juvenile carp (the most abundant small fish) were collected from Cayuga Marsh and transferred to mesocosms, but not until July, after they became abundant from the spawn [ILLUSTRATION FOR FIGURE 1 OMITTED]. For both fish introductions, the specific numbers of fish added (3 large bullheads or 40 small carp) were chosen so that densities in mesocosms mimicked ambient densities. For the Small-Fish Exclosures, calibration studies indicated that exclosure reduced densities of small insectivorous fish by only 90% (some larval fish penetrated the mesocosm walls soon after hatching and became trapped inside). For Cattail-Litter Supplements and Nonnutritive-Litter Supplements, litter was added in May to mimic the natural period when dead emergent plants fall into the water. The volume of litter introduced was within the range of natural inputs into similar sized, non-open-water areas of Cayuga Marsh, and it resulted in a loose layer of material covering the entire benthic substrate of the mesocosms. Preliminary aquarium studies indicated that the fiberglass tape used to mimic cattail litter in Nonnutritive-Litter Supplements remained intact even after soaking for many months. Algal periphyton grew on the fiber glass tape, and sensitive aquatic microcrustaceans persisted for many months in aquaria containing tape, indicating that the material had no significant toxic side effects.


During 9-13 May, after all mesocosms were in place but before manipulations, I collected a baseline sample of the benthic midge populations in each mesocosm, using a 4 cm diameter corer. Within a 1 x 1 m plot in the center of each mesocosm, I collected six core samples at randomly selected locations. Because I assumed that the mesh walls themselves could influence adjacent habitat, I did not sample within a 25-cm buffer zone along each mesocosm's interior perimeter. I sampled nektonic invertebrate communities along benthic substrates by placing invertebrate activity traps (two traps/mesocosm) on the bottom-center of each mesocosm for 24 h (1-L plastic bottles with an inverted funnel opening on one end [Riley and Bookhout 1990]). A baseline fish sample was collected from each of the 42 mesocosms, using modified minnow-traps (described above) placed near the bottom-center of each habitat (two traps/mesocosm). Previous sampling in Cayuga Marsh and elsewhere indicated that these samplers readily captured fish as large as 20-cm standard length. These traps did not collect large and very small fish, and fish behavior probably affected capture efficiency. However, they were useful for this study because samplers could be placed to collect selectively those fish occurring along benthic substrates, and sampling procedures minimally disrupted the habitats within mesocosms.

After the May baseline samplings, I initiated all manipulations, except for additions of small carp, which were added later in July. Subsequently I sampled midges, nektonic invertebrates, and large and small fish in each of the 42 mesocosms, as described above, on a 5-wk rotating schedule (13-17 June, 18-22 July, and 22-26 August 1994). However for these three samples, I increased my core sampling effort from 6 to 8 subsamples/mesocosm (pooled sample area = 0.01 [m.sup.2] mesocosm). Except for these scheduled samplings, I left the interiors of mesocosms undisturbed during interim periods.

Data on the gut contents of fish are not provided. During the May to August experimental period, all fish collected in traps were returned to mesocosms to minimize disturbance. At the experiment's end, few fish remained because of habitat-wide declines in fish densities [ILLUSTRATION FOR FIGURE 1 OMITTED].

Tests of mesocosm cage artifacts

When artificial structures are placed into natural aquatic environments, cage artifacts may confound results of the desired contrasts (Peckarsky and Penton 1990). In this study I was particularly concerned that midges and/or small fish might congregate in or avoid mesocosms of a particular design, regardless of other manipulations. I looked for such artifacts in this study by comparing midge and small fish numbers between mesocosm types with similar overall biotic structures but different construction. I carried out four tests:

1) In May when midge densities should have been similar in open, coarse-wall, and fine-wall mesocosms, and numbers of small fish should have been equal in open and coarse-wall cells, I compared benthic-midge and small-fish densities (which then were very rare) among the three mesocosm types to determine if their densities had been significantly influenced while mesocosms were being erected.

2) In June, July, and August I compared open Ambient Controls and coarse-wall Large-Fish Enclosures (Table 1), which both contained ambient levels of detritus, midges, and large and small fish throughout the summer.

3) In June (prior to the yearly spawn) I compared fine-wall Small-Fish Exclosures and coarse-wall Large-Fish Exclosures (Table 1), both of which contained no large fish and few if any small fish at that time.

4) In August I compared fine-wall Small-Fish Enclosures and coarse-wall Large-Fish Exclosures (Table 1), which then both supported near ambient densities of small fish and no large fish.

Experiments on trophic interaction

Comparisons among the seven mesocosm treatments (Table 1) allowed me to test several hypotheses about trophic interactions of benthic midge populations in the Cayuga Marsh study pool using a series of orthogonal comparisons of midge and small-fish densities. For predator experiments I used both exclosure and enclosure approaches to reduce possible biases introduced when a single technique is used (Walde and Davies 1984).

Hypothesis 1: Large fish significantly influence midge densities or reduce small fish numbers. - Ambient Controls + Large-Fish Enclosures vs. Large-Fish Exclosures.

Hypothesis 2: Small insectivorous fish significantly reduce midge densities. - Test 1. Ambient Controls + Large-Fish Enclosures + Large-Fish Exclosures vs. Small-Fish Exclosures. Test 2. Small-Fish Exclosures vs. Small-Fish Enclosures.

Hypothesis 3: Availability of cattail litter limits midge densities. - Ambient Controls + Caged Ambient Detritus vs. Cattail-Litter Supplements.

Hypothesis 4: Midge response to cattail litter results from increased habitat heterogeneity. - Test 1. Ambient Controls + Caged Ambient Detritus vs. Nonnutritive-Litter Supplements. Test 2. Cattail-Litter Supplements vs. Nonnutritive-Litter Supplements (for this test, the hypothesis would be confirmed if midge densities do not differ).

Testing was conducted in a sequential manner, and the order was predetermined. For fish-predator tests, I first tested for cage effects, then large-fish effects, and lastly small-fish effects. For litter resource tests, I first tested for cage effects, then natural-litter effects, and lastly artificial-litter effects. If initial tests in a series suggest negligible differences, those treatments were combined as replicates for subsequent tests to increase sample sizes. For example, since cages' effects were minor (see section below, Cage artifacts of mesocosms), Ambient Controls were considered as replicates of Large-Fish Enclosures to test Hypothesis 1 and as replicates of Caged Ambient Detritus to test Hypotheses 3 and 4. When tests of Hypothesis 1 indicated that large fish were not affecting midge or small-fish densities, I included the three treatments used for those tests as replicates of ambient small-fish densities for testing Hypothesis 2.

Statistical analyses

Statistical tests were conducted using Statview SuperANOVA Bundle computer software (Abacus Concepts, SuperANOVA 1989). To contrast densities of midges or small fish between or among treatments, I pooled all subunits of core (6 or 8) or trap (2) samples that were collected within individual mesocosms into single sample units, which then served as the replicates. I considered C. plumosus, C. tentans, and G. barbipes midges as one trophospecies (Pimm et al. 1991), because relative numbers of each were consistent among all mesocosms, and there were no species-specific responses to manipulations of detritus or fish predation. Within-mesocosm midge distributions were similar in all treatments (coefficients of variation averaged 1.55). I similarly pooled all species of small fish (juvenile carp and bullhead, and adult stickleback), because all species collected could potentially feed on midge larvae, and there were no species-specific responses by fish to treatments (only in the Small-Fish Enclosures did the small-fish population consist exclusively of carp). I did not contrast numbers of nektonic invertebrates among treatments because so few were captured along benthic substrates.

To evaluate whether mesocosm design alone influenced midge or small-fish distributions, I contrasted May baseline samples of midges and small fish among open, coarse-wall, and fine-wall mesocosms using two-way ANOVA (mesocosm type x location block, with full interaction; two-tailed). Scheffe tests were used to determine specific differences among the three mesocosm types. I did not use randomized-block ANOVAs because of possible interactions between treatment and location block (Abacus Concepts, SuperANOVA 1989). Six location blocks existed, with each of the two rows of mesocosms containing three blocks. For other tests for cage artifacts involving paired contrasts, I used either two-way ANOVA or paired t tests for those analyses (two-tailed).

Analyses of treatment effects on midge and small-fish abundances in June, July, and August involved paired-orthogonal comparisons. I used two-way ANOVA (treatment x location) when replication existed within location blocks. Paired t tests were used when each treatment had only a single replicate per block. For those experiments that were sampled in June, July, and August, I considered each month independently in my statistical testing, rather than as repeated measures, because seasonal life-history patterns of both midges and fish caused different biotic communities to develop during each sampling period. Furthermore, I was specifically interested in how the outcomes of ecological interactions changed seasonally. To account for any procedure-wise error that was introduced by repeating analyses three times, I reduced the test alpha level from P [less than] 0.05 to 0.017 for each of the monthly contrasts per experiment. One-tailed tests were used when examining impacts of large fish on small fish or of small fish on midges, and also when determining whether additions of natural or artificial litter benefited midges. However, I used two-tailed tests when addressing impacts of large fish on midges (these fish could potentially impact midge populations negatively or positively). If Bartlett's test for homogeneity of variance (Snedecor and Cochran 1967) indicated that treatments had unequal variances (test results not presented), I applied a [log.sub.10] (x + 1) transformation (Allan 1984) to those sample data prior to analyses to normalize data and homogenize variance.


Cage artifacts of mesocosms

Some minor cage artifacts were detected in this study. In the May baseline samples, midge densities varied among the three mesocosm types ([F.sub.2,24] = 4.70, P = 0.019). Sheffe post hoc tests indicated that coarse-and fine-wall mesocosms supported similar numbers of midges (P = 0.99), but both of these treatments supported fewer midges than the mesocosms with no walls (both P [less than] 0.04). As mentioned previously, small fish were rare in May, and none were collected in baseline samplings. Differences in midge densities were likely due to increased disturbances associated with erecting mesocosms with walls over those without. However, this disturbance only affected the cohort midges that had overwintered from the previous 1993 season. Any impacts on midges resulting from erecting mesocosms subsequently disappeared when those midges emerged in June. However because of this initial cage effect, I restrict my ecological interpretations to the new cohort of midges that began first appearing in June (1994 cohort, [ILLUSTRATION FOR FIGURE 2 OMITTED]).

After May, I no longer detected any significant cage effects associated with coarse mesh. In June, July, and August, the Ambient Controls and coarse-wall Large-Fish Enclosures supported very similar midge and small-fish densities [ILLUSTRATION FOR FIGURE 3 OMITTED]. Since both of these treatments should represent near-ambient conditions, the similarities in their midge and small-fish numbers verified that realistic conditions were being maintained in coarse-wall mesocosms from June through August.

Minor cage effects might have persisted somewhat longer in fine-wall mesocosms. In June, I collected more than twice as many new-cohort midges in the coarse-wall Large-Fish Exclosures than in the fine-wall Small-Fish Exclosures, although this difference was not significant ([F.sub.1,6] = 4.41, P = 0.080) (negligible numbers of small fish occurred in both treatments at that time). Subsequent testing in August, after small fish became abundant [ILLUSTRATION FOR FIGURE 1 OMITTED], found no cage effects on midges inside fine-wall vs. coarse-wall mesocosms. The fine-wall Small-Fish Enclosures and the coarse-wall Large-Fish Exclosures supported very similar densities of midges (t = 0.83, df = 5, P = 0.44). At that time, these two mesocosm types also had statistically similar densities of small fish (t = 2.06, df = 5, P = 0.09), and should have had similar ambient levels of detritus. Overall the mesocosms appeared to be reasonable arenas for testing ecological hypotheses.

Hypotheses on trophic interaction

Midge and/or small-fish numbers varied significantly among the six location blocks for many of the below analyses, but in no instance did significant statistical interactions exist between an experimental treatment and the location of mesocosms.

Hypothesis 1: Large fish significantly influence midge densities or reduce small-fish numbers. - In June, July, and August, the 12 habitats with large fish (Ambient Controls + Large-Fish Enclosures) had midge densities very similar to the Large-Fish Exclosures ([ILLUSTRATION FOR FIGURE 4A OMITTED]; June: [F.sub.1,6] = 0.01, P = 0.93, July: [F.sub.1,6] = 0.04, P = 0.85, August: [F.sub.1,6] = 0.03, P = 0.86). Few small fish were present in mesocosms until after the early summer spawn, and the presence or absence of large fish did not affect post-spawn small-fish densities in July ([F.sub.1,6] = 0.13, P = 0.73) or August ([F.sub.1,6] = 2.17, P = 0.19) [ILLUSTRATION FOR FIGURE 4B OMITTED]. Hypothesis 1 was not supported.

Hypothesis 2: Small insectivorous fish significantly reduce midge densities. - In July, after small fish first became abundant from the spawn, the Small-Fish Exclosures supported far fewer small fish ([ILLUSTRATION FOR FIGURE 5B OMITTED], [F.sub.1,18] = 14.72, P [less than] 0.001) than the treatment supporting ambient densities of small fish (Ambient Controls + Large-Fish Enclosures + Large-Fish Exclosures). By August, ambient densities of small fish had declined, and now only marginal differences in their density were detected between treatments ([F.sub.1,12] = 2.49, P = 0.07). Conversely, midge densities were significantly higher in the Small-Fish Exclosures ([ILLUSTRATION FOR FIGURE 5A OMITTED], [F.sub.1,18] = 8.26, P = 0.005). By August, the differences in midge densities between those two treatments had become even more pronounced ([ILLUSTRATION FOR FIGURE 5A OMITTED], [F.sub.1,12] = 12.56, P = 0.002). If I had achieved a complete exclosure of small fish, these differences in midge numbers might have been even larger.

In the post-spawn experiment in which small carp were enclosed from July to August, the Small-Fish Enclosures supported significantly fewer midges by the experiment's end in August than did the Small-Fish Exclosures ([ILLUSTRATION FOR FIGURE 6 OMITTED]; t = 3.05, df = 5, P = 0.014). Evidence from both the former fish-exclosure experiment and the latter fish-enclosure experiment support Hypothesis 2.

Hypothesis 3: Availability of cattail litter limits midge densities.--In June, habitats with ambient levels of detritus (Ambient Controls + Caged Ambient Detritus) and the Cattail-Litter Supplements supported similar numbers of midges ([ILLUSTRATION FOR FIGURE 7A OMITTED]; [F.sub.1,6] = 1.52, P = 0.26). In July, midge densities in these two treatments began to diverge, but differences were not yet significant ([F.sub.1,6] = 3.72, P = 0.051). However by August, significantly more midges occurred where cattail litter had been supplemented ([ILLUSTRATION FOR FIGURE 7A OMITTED]; [F.sub.1,6] = 11.08, P = 0.008). Small-fish numbers were similar between habitats with ambient and supplemented litter in both July (P = 0.41) and August (P = 0.32). Data in August support Hypothesis 3.

Hypothesis 4: Midge response to cattail litter results from increased habitat heterogeneity. - In June and August, midge densities were very similar in the Nonnutritive-Litter Supplements and habitats with ambient levels of detritus ([ILLUSTRATION FOR FIGURE 7B OMITTED]; June: [F.sub.1,6] = 1.33, P = 0.29, August: [F.sub.1,6] = 0.02, P = 0.90). In July, somewhat fewer midge larvae were collected in the presence of artificial cattail ([F.sub.1,6] = 9.89, P = 0.020), although after accounting for procedure-wise error this difference was not significant (P [greater than] 0.017). Small-fish numbers were similar between habitats with ambient and supplemented artificial litter in both July (P = 0.38) and August (P = 0.47). Results show that enhancing structure alone did not benefit midges, and during July the artificial litter might even have been somewhat detrimental to midges (for unknown reasons). In July and August, supplements of natural detritus that enhanced levels of both food and structure provided marginal but nonsignificant benefits for midges over supplements of artificial detritus that enhanced structure only (July: t = 2.12, df = 5, P = 0.044, August: t = 1.67, df = 5, P = 0.078). Hypothesis 4 was not supported.


The dynamic nature of food web interaction is now recognized (Schoenly and Cohen 1991, Polls et al. 1996). In Cayuga Marsh, considerable temporal variation existed in the structure and function of the benthic food web, with trophic interactions of midges changing from June to August. In June, neither predation nor resources limited benthic midge populations. In July, predation but not resource limitation became an important influence. In August, both predation and resources were limiting midge numbers.

Top-down interactions of predators

Manipulations of top-predators have been shown to affect multiple lower trophic levels in some cases (Carpenter et al. 1987, Power 1990, Bronmark 1994) but not others (Hambright et al. 1986). Strong (1992) suggests that trophic cascades might occur most often in low-diversity habitats, although Power et al. (1996) indicate cascades can also occur in speciose habitats. Additionally, Power (1992) suggests that cascading effects of fish predation may be most pronounced in structurally simple habitats. However, results in Cayuga Marsh do not provide support for either of these contentions. While the benthic substrate of Cayuga Marsh supported few species and was structurally simple, direct feeding by large fish such as brown bullhead did not significantly affect abundances of organisms in lower trophic levels. However, large fish (brown bullhead as well as carp) still provided an important, indirect impact on the food web by serving as the brood stock for the mid-summer pulse of small fish.

Small insectivorous fish strongly influenced benthic midge populations in Cayuga Marsh. By summer's end, the benthic substrates of the study pool supported only low densities of midges, except where small fish had been excluded. In some aquatic systems, including the marsh-like lake littoral, fish predation often results in resistant species becoming prevalent (Vanni 1987, Wellborn and Robinson 1991, Osenberg and Mittelbach 1996). However in Cayuga Marsh, benthic substrates in late summer supported few invertebrates of any kind. The short mid-summer burst of predator activity had relatively long-term ramifications for midges. They were unable to mount a late-summer response to declining small-fish numbers because the univoltine midges had finished breeding for that year. Thus, the impacts of mid-summer predation would persist until the new cohort of midges arrived the next year. The life-history patterns of both midges and fish influenced the structure and function of this wetland food web (see also Winemiller 1996).

Bottom-up interactions of resources

While detritus is a valuable source of energy for most food webs (Vanni and de Ruiter 1996), remarkably few empirical studies suggest that consumer populations are limited by detritus availability. Richardson (1991) found that supplementing detritus into an oligotrophic montane stream increased detritivore numbers. Dodson and Hildrew (1992) also found that supplementing detritus in more eutrophic streams increased consumer numbers, but the impact was minimal in streams that naturally retained large amounts of detritus. David et al. (1991) found that excluding litter from oak forest soils reduced invertebrate numbers, but supplements produced no response. They concluded that under normal conditions the soil fauna was not litter limited. Similarly, a review by Batzer and Wissinger (1996) suggests that detritus availability in wetlands seldom limits the abundance of invertebrate consumers. Apparently consumer limitation by detritus is not a widespread phenomenon.

However, this study showed that supplies of litter detritus limited marsh detritivore abundance, and most evidence points toward a nutritional rather than structural basis for midge response. Experiments with artificial litter did not indicate that the litter's physical shape benefited midges. Furthermore, midges' response to natural detritus was restricted to the late summer period when litter had largely lost its macro-structure through decomposition. While detritus from emergent plant litter is an important food resource for marsh midges (Neill and Cornwell 1992, Bunn and Boon 1993), new litter probably must be conditioned microbially before it becomes palatable (Smock and Stoneburner 1980, Smock and Harlowe 1983). The importance of litter conditioning has been long established for consumer-detritus interaction in streams (Cummins et al. 1973). The progressively stronger response of Cayuga Marsh midges to litter over the summer may have resulted from changing nutritional quality. During the early summer, cattail supplements may have increased only heterogeneity and not food resources. Campeau et al. (1994) compared midge response to natural vs. artificial litter in a Canadian marsh, and similarly concluded that a late-summer response was nutritionally based.

Concurrent limitation by predators and resources

Theory often assumes that competition for resources, a density-dependent phenomenon, becomes inconsequential when consumer populations are suppressed by predation (Hairston et al. 1960, Wiens 1977). However, Rosemond et al. (1993) and Osenberg and Mittelbach (1996) maintain that predation and resource limitation are not mutually exclusive processes, but operate in concert on populations. Rosemond et al. (1993) found that algae in streams were being simultaneously influenced by nutrients and grazing snails. Their evidence for dual bottom-up and top-down trophic limitation is especially convincing because the resources were dissolved nutrients that should not affect grazer efficiency. As discussed above, for detritivores it is more difficult to separate nutritional effects of litter from possible refuge effects. However, it seems clear that even though Cayuga Marsh midges had been severely impacted by predation, they could still respond to increased food levels even though they were probably not experiencing pronounced competition for resources. As seen in small-fish exclusions, ambient levels of food could have supported considerably higher numbers of midges.

Even when both top-down and bottom-up interactions influence a population, one factor may take precedence over the other. Osenberg and Mittlebach (1996) maintain that both predators and resources were affecting populations of lacustrine snails, but resource limitation was the more important factor. Conversely, in terms of trophic interaction in Cayuga Marsh, predation was the dominant influence on midge populations (at least over the period of study). Belovsky and Joern (1995) suggest that while many factors can concurrently be influencing a population, in terms of population regulation a single process will always have primacy. The specific process may vary across temporal and spatial scales, however. Their experiments indicate that grassland-grasshopper populations were being regulated by resource limitation in Montana and by avian predators in Nebraska. They suggest that ecologists should be working toward defining the environmental and biotic conditions that determine which mechanisms come to dominate populations rather than searching for broad generalizations.


I thank Bobbi Peckarsky and Alex Flecker for help in developing and evaluating this study. John Grandy and Christopher Pusateri provided valuable assistance with its laboratory aspects. Funding was provided by a post-doctoral research fellowship to me from the National Science Foundation (DEB-9303248). Portions of this work were conducted while I was employed by the Department of Biology at Canisius College, Buffalo, New York.


Abacus Concepts. SuperANOVA. 1989. Abacus Concepts Inc., Berkeley, California, USA.

Allan, J. D. 1984. Hypothesis testing in ecological studies of aquatic insects. Pages 484-507 in V. H. Resh, and D. M. Rosenberg, editors. The ecology of aquatic insects. Praeger, New York, New York, USA.

Batzer, D. P., and V. H. Resh. 1992. Macroinvertebrates of a California seasonal wetland and responses to experimental habitat manipulation. Wetlands 12:1-7.

Batzer, D. P., and S. A. Wissinger. 1996. Ecology of insect communities in nontidal wetlands. Annual Review of Entomology 41:75-100.

Belovsky, G. E., and A. Joern. 1995. The dominance of different regulating factors for rangeland grasshoppers. Pages 359-386 in N. Cappuccino and P. W. Price, editors. Population dynamics: new approaches and synthesis. Academic Press, San Diego, California, USA.

Bronmark, C. 1994. Effects of tench and perch on interactions in a freshwater, benthic food chain. Ecology 75:1818-1828.

Bronmark, C., S. P. Klosiewski, and R. A. Stein. 1992. Indirect effects of predation in a freshwater, benthic food chain. Ecology 73:1662-1674.

Bunn, S. E., and P. I. Boon. 1993. What sources of organic carbon drive food webs in billabongs? A study based on stable isotope analysis. Oecologia 96:85-94.

Campeau, S., H. R. Murkin, and R. D. Titman. 1994. Relative importance of algae and emergent plant litter to freshwater marsh invertebrates. Canadian Journal of Fisheries and Aquatic Science 51:681-692.

Cappuccino, N., and P W. Price, editors. 1995. Population dynamics: new approaches and synthesis. Academic Press, San Diego, California, USA.

Carpenter, S. R., J. F. Kitchell, J. R. Hodgson, P. A. Cochran, J. J. Elser, M. M. Elser, D. M. Lodge, D. Kretchmer, X. He, and C. N. von Ende. 1987. Regulation of lake primary productivity by food web structure. Ecology 68:1863-1876.

Chapman, G., and C. H. Fernando. 1994. The diets and related aspects of feeding of Nile tilapia (Oreochromis niloticus L.) and common carp (Cyprinus carpio L.) in lowland rice fields in northeastern Thailand. Aquaculture 123:281-307.

Coffman, W. P., and L. C. Ferrington, Jr. 1984. Chironomidae. Pages 551-562 in R. W. Merritt, and K. W. Cummins, editors. An introduction to the aquatic insects of North America. Second edition. Kendall/Hunt, Dubuque, Iowa, USA.

Cummins. K. W., R. C. Peterson, F. O. Howard, J. C. Wuycheck, and V. I. Holt. 1973. The utilization of leaf litter by stream detritivores. Ecology 54:336-45.

David, J. F., J. F. Ponge, P. Arpin, and G. Vanier. 1991. Reactions of the macrofauna of a forest mull to experimental perturbations of litter supply. Oikos 61:316-326.

Diehl, S. 1992. Fish predation and benthic community structure: the role of omnivory and habitat complexity. Ecology 73:1646-1661.

-----. 1995. Direct and indirect effects of omnivory in a littoral take community. Ecology 76:1727-1740.

Dodson, M., and A. G. Hildrew. 1992. A test of resource limitation among shredding detritivores in low order streams in southern England. Journal of Animal Ecology 61:69-77.

Dvorak, J. 1987. Production-ecological relationships between aquatic vascular plants and invertebrates in shallow waters and wetlands - a review. Archiv fur Hydrobiologie Beiheft Ergebnisse der Limnologie 27:181-184.

Ebenhard, T. 1988. Introduced birds and mammals and their ecological effects. Swedish Wildlife Research 13:1-107.

Hairston, N. G., F. E. Smith, and L. B. Slobodkin. 1960. Community structure, population control and competition. American Naturalist 44:421-425.

Hall, D. J., W. E. Cooper, and E. E. Werner. 1970. An experimental approach to the production dynamics and structure of freshwater animal communities. Limnology and Oceanography 15:839-928.

Hambright, K. D., R. J. Trebatoski, R. W. Drenner, and D. Kettle. 1986. Experimental study of the impacts of bluegill (Lepomis macrochirus) and largemouth bass (Micropterus salmoides) on pond community structure. Canadian Journal of Fisheries and Aquatic Science 43:1171-1176.

Hunter, M. D., and P. W. Price. 1992. Playing chutes and ladders: heterogeneity and relative forces of bottom-up and top-down forces in natural communities. Ecology 73:724732.

Ivlev, V. S. 1961. Experimental ecology of the feeding of fishes. Yale University Press, New Haven, Connecticut, USA.

Kerfoot, W. C., and A. Sih, editors. 1987. Predation: direct and indirect impacts on aquatic communities. University Press of New England, Hanover, New Hampshire, USA.

Kitching, R. L. 1987. Spatial and temporal variation in food webs in water-filled treeholes. Oikos 48:280-288.

Kline, J. L., and B. M. Wood. 1996. Food habits and diet selectivity of the brown bullhead. Journal of Freshwater Ecology 11:145-151.

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.

Murkin, H. R. 1989. The basis for food chains in prairie wetlands. Pages 316-338 in A. G. van der Valk, editor. Northern prairie wetlands. Iowa State University Press, Ames, Iowa, USA.

Neckles, H. A., H. R. Murkin, and J. A. Cooper. 1990. Influences of seasonal flooding on macroinvertebrate abundance in wetland habitats. Freshwater Biology 23:311-322.

Neill, C., and J. C. Cornwell. 1992. Stable carbon, nitrogen, and sulfur isotopes in a prairie marsh food web. Wetlands 12:217-224.

Odum, W. E., and E. J. Heald. 1972. Trophic analyses of an estuarine mangrove community. Bulletin of Marine Science 22:671-738.

Osenberg, C. W., and G. G. Mittelbach. 1996. The relative importance of resource limitation and predator limitation in food chains. Pages 134-148 in G. A. Polis and K. O. Winemiller, editors. Food webs: integration of patterns and dynamics. Chapman and Hall, New York, New York, USA.

Peckarsky, B. L., and M. A. Penton. 1990. Effects of enclosures on stream microhabitat and invertebrate structure. Journal of the North American Benthological Society 9:249-261.

Pimm, S. L., J. H. Lawton, and J. E. Cohen. 1991. Food web patterns and their consequences. Nature 350:669-674.

Polis, G. A., R. D. Holt, B. A. Menge, and K. O. Winemiller. 1996. Time, space, and life history: influences on food webs. Pages 435-460 in G. A. Polis and K. O. Winemiller, editors. Food webs: integration of patterns and dynamics. Chapman and Hall, New York, New York, USA.

Polis, G. A., and K. O. Winemiller. editors. 1996. Food webs: integration of patterns and dynamics. Chapman and Hall, New York, New York, USA.

Power, M. E. 1990. Effects of fish in river food webs. Science 250:411-414.

-----. 1992. Habitat heterogeneity and the functional significance of fish in river food webs. Ecology 73:1675-1688.

Power, M. E., M. S. Parker, and J. T Wootton. 1996. Disturbance and food chain length in rivers. Pages 286-297 in G. A. Polis and K. O. Winemiller, editors. Food webs: integration of patterns and dynamics. Chapman and Hall, New York, New York, USA.

Rasmussen, J. B. 1985. Effects of density and microdetritus enrichment on the growth of chironomid larvae in a small pond. Canadian Journal of Fisheries and Aquatic Science 42: 1418-1422.

Richardson, J. S. 1991. Seasonal food limitation of detritivores in a montane stream: an experimental test. Ecology 73:873-887.

Riley, T. Z., and T. A. Bookhout. 1990. Response of aquatic macroinvertebrates to early-spring drawdown in nodding smartweed marshes. Wetlands 10:173-185.

Rosemond, A. D., P. J. Mulholland, and J. W. Elwood. 1993. Top-down and bottom-up control of stream periphyton: effects of nutrients and herbivores. Ecology 74:1264-1280.

Schoenly, K., and J. E. Cohen. 1991. Temporal variation in food web structure: 16 empirical cases. Ecological Monographs 61:267-298.

Sih, A., P. Crowley, M. McPeek, J. Petranka, and K. Strohmeier. 1985. Predation, competition, and prey communities: a review of field experiments. Annual Review of Ecology and Systematics 16:269-311.

Sinnott, T. J., and N. H. Ringler. 1987. Population biology of the brown bullhead (Ictalurus nebulosus Lesueur). Journal of Freshwater Ecology 4:225-234.

Smock, L., and K. L. Harlowe. 1983. Utilization and processing of freshwater wetland macrophytes by the detritivore Asselus forbesi. Ecology 64:1556-1565.

Smock, L. A., and D. L. Stoneburner. 1980. The response of macroinvertebrates to aquatic macrophyte decomposition. Oikos 35:397-403.

Snedecor, G. W., and W. G. Cochran. 1967. Statistical methods. Iowa State University Press, Ames, Iowa, USA.

Strong, D. R. 1992. Are trophic cascades all wet? Differentiation and donor-control in speciose ecosystems. Ecology 73:747-754.

Vanni, M. J. 1987. Effects of food availability and fish predation on a zooplankton community. Ecological Monographs 57:61-88.

Vanni, M. J., and P. C. de Ruiter. 1996. Detritus and nutrients in food webs. Pages 25-29 in G. A. Polls and K. O. Winemiller, editors. Food webs: integration of patterns and dynamics. Chapman and Hall, New York, New York, USA.

Walde, S. J., and R. W. Davies. 1984. Responses of lotic prey populations to an invertebrate predator: evolution of in situ enclosure/exclosure experiments. Ecology 65:1206-1213.

Warren, P. H. 1989. Spatial and temporal variation in the structure of a freshwater food web. Oikos 55:299-311.

Wellborn, G. A., and J. V. Robinson. 1991. The influence of fish predation in an experienced prey community. Canadian Journal of Zoology 69:2515-2522.

Wiens, J. A. 1977. On competition and variable environments. American Scientist 95:590-597.

Winemiller, K. O. 1996. Factors driving temporal and spatial variation in aquatic floodplain food webs. Pages 298-312 in G. A. Polis and K. O. Winemiller, editors. Food webs: integration of patterns and dynamics. Chapman and Hall, New York, New York, USA.

Zur, O., and S. Sarig, 1980. Observations on the feeding of common carp (Cyprinus carpio) on chironomid larvae. Bamidgeh 31:25-26.
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