WETLANDS, Vol. 26, No. 2, June 2006, pp. 430–437 䉷 2006, The Society of Wetland Scientists

SPECIES-SPECIFIC RESPONSES OF AQUATIC MACROPHYTES TO FISH EXCLUSION IN A PRAIRIE MARSH: A MANIPULATIVE EXPERIMENT Vincent D. Evelsizer1 and Andrew M. Turner Department of Biology Clarion University Clarion, Pennsylvania, USA 16214 1 Present address: Iowa Department of Natural Resources 109 Trowbridge Hall University of Iowa Iowa City, Iowa, USA 52242

Abstract: An exclosure experiment was carried out at two sites in Delta Marsh, Manitoba, Canada to investigate the role of fish in limiting the growth of submersed macrophytes. The experiment consisted of three treatments: (1) fine-mesh exclosures designed to exclude both planktivorous fish and carp, (2) coarsemesh exclosures designed to exclude adult carp but admit smaller fish, and (3) reference plots marked with corner stakes but without sides. Treatments were established in mid-May, and macrophyte biomass was sampled from within the exclosures in late August to assess treatment effects. Exclosure treatments had strong effects on the macroalgae Chara, with biomass 11.9-fold greater in full-exclosure plots than in reference plots, and 3-fold greater in carp exclosures than in reference plots. Exclosure treatments had no effect on above-ground or below-ground biomass of Stuckenia pectinata, the most widespread and abundant macrophyte in Delta Marsh. Thus, fish appear to limit macrophyte growth in Delta Marsh, but the effect of fish exclusion was dependent on species composition of the macrophyte assemblage. Key Words: carp

macrophytes, biomanipulation, turbidity, wetland restoration, Chara, Stuckenia, Delta Marsh,

INTRODUCTION

tal approaches are necessary to identify the underlying mechanisms. One potential determinant of macrophyte abundance in wetlands is the abundance of planktivorous and benthivorous fish. A number of studies have shown that high densities of zooplanktivorous fish are capable of depressing zooplankton standing crop, which in turn reduces phytoplankton grazing by zooplankton and leads to high phytoplankton standing crops and thus low water clarity (Turner and Mittelbach 1992, Carpenter and Kitchell 1993, Schriver et al. 1995, Jeppesen et al. 1997). Low water clarity usually limits the growth and species diversity of submersed aquatic plants (e.g., Chambers and Kalff 1985, Duarte et al. 1986, Kantrud 1990). Benthivorous fish, such as common carp (Cyprinus carpio Linneaus) can contribute to turbidity by resuspending sediments as they forage (Meijer et al. 1990, Breukelaar et al. 1994). Carp may also limit the growth of aquatic macrophytes in a direct manner by uprooting plants during feeding or spawning activities (Anderson 1950, Tryon 1954, Atton 1959, Robel 1961, King and Hunt 1967). Thus, a key to the restoration of submersed macrophytes in marsh-

Shallow lakes and marshes tend to exist in one of two alternative conditions: a clear water state characterized by high water transparency and abundant submersed macrophytes, and a turbid state characterized by low water clarity and few submersed macrophytes (Timms and Moss 1984, Jeppesen et al. 1990, Scheffer 1998, Bayley and Prather 2003). Nutrient loading, food web structure, herbivory, and wave exposure are among the factors that influence lake state. Many shallow lakes and marshes in North America and Europe have recently experienced increases in turbidity and decreases in the abundance of submersed macrophytes (Chow-Fraser 1998, Scheffer 1998). Often, the decline of macrophytes is associated with human-induced disturbances such as the stabilization of water levels, invasion of planktivorous and/or benthivorous fish, or increased nutrient loading (Scheffer et al. 1993, Bouffard and Hanson 1997). However, because these disturbances often disrupt several mechanisms simultaneously, the specific factors most responsible for macrophyte decline are usually unknown, and experimen430

Evelsizer & Turner, FISH AND SUBMERSED MACROPHYTES es may be the limitation of planktivorous and benthivorous fish (e.g., biomanipulation; Shapiro and Wright 1984, Benndorf 1987, Jeppesen et al. 1990). European investigators have repeatedly tested these ideas and have accumulated substantial evidence documenting the important roles of fish in influencing submersed macrophytes in shallow lakes and marshes (e.g., Scheffer et al. 1993, Jeppesen et al. 1997), but the role of fish in North American wetlands has not been as widely evaluated (but see Hanson and Butler 1994, Zimmer et al. 2001, 2002). Here, we evaluate the effect of fish exclusion on submersed macrophyte biomass in East Delta Marsh, Manitoba, Canada. Delta Marsh is one of North America’s most prominent wetlands and has been the site of important research in wetland ecology (e.g., Murkin et al. 2000) but, like many wetlands, has experienced increased turbidity, a decrease in the abundance of submersed macrophytes, and a reduction in waterfowl use (deGeus 1987, Batt 2000). We experimentally excluded benthivorous fish and planktivorous fish from plots within the marsh and monitored the response of submersed macrophytes. Our goal was to test potential strategies for remediation at small scales to improve methods for subsequent ecosystem level manipulations.

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(Goldsborough 1995). Water pH ranges from 8.2 to 9.0, and total alkalinity averages 338 mg/l CaCO3, largely as bicarbonate (Anderson and Jones 1976). Delta Marsh historically had greater water clarity and supported dense beds of submersed macrophytes (Hinks and Fryer 1936, Walker 1959, 1965). Large numbers of waterfowl used the marsh as a migratory stop-over and breeding site (Hochbaum 1944, 1955). However, the common carp invaded Lake Manitoba and the marsh in 1948 (McCrimmon 1968), and Fairford Dam was built at the lake’s outlet in 1961, which stabilized the water level of Delta Marsh. Since that time, water clarity in the marsh has decreased, the abundance of submersed macrophytes has been reduced, and waterfowl use of the marsh has declined (Batt 2000). Fish, including common carp and fathead minnows (Pimephales promelas Rafinesque), are now extraordinarily abundant in the marsh (LaPointe 1986, Kiers and Hann 1995, Batt 2000, Evelsizer 2001). We conducted the exclosure experiment at two sites, Division Bay and 22-Bay (Figure 1). These sites were chosen based on historical data showing that S. pectinata was present at these sites in 1973–74 (Anderson and Jones 1976) and 1997 (T. Arnold and D. Wrubleski, personal communication). Exclosure Study

METHODS Study Sites Delta Marsh, located in south central Manitoba, Canada (50⬚11⬘ N, 98⬚19⬘ W), is a large (⬃22,000 ha) lacustrine marsh bordering the southern shoreline of Lake Manitoba (Figure 1). Nearby upland areas south of the marsh are intensively farmed. East Delta Marsh, the site of our studies, is a shallow (⬍ 2.5 m depth) network of bays, channels, and ponds bordered by hybrid cattail Typha X glauca Godr. and common reed (Phragmites australis (Cav.) Trin ex. Steud.) (deGeus 1987, Shay et al. 1999). Sago pondweed (Stuckenia pectinata (L.) Borner) is the most abundant submersed macrophyte in the marsh (Anderson and Low 1976, Anderson 1978). Summer water clarity is generally low, with vertical extinction coefficient values (kd, photosynthetic active radiation) of 3.2–5.0 m⫺1 and turbidity values of 12–25 NTU (Evelsizer 2001). The marsh is moderately brackish (classification of Stewart and Kantrud 1972), with specific conductivity values that generally range between 1000 and 3000 ␮S/cm and total dissolved solids concentrations ranging from 519 to 3230 mg/l (Goldsborough 1995, Evelsizer 2001). The waters are nutrient-rich, with water-column nitrogen: phosphorus ratios generally ⬎16 and total phosphorus concentrations generally ⬎ 50 ␮g/l

The experiment consisted of three treatments: (1) a fine-mesh exclosure designed to exclude all fish (hereafter, full exclosure), (2) a coarse mesh exclosure designed to exclude adult carp but admit smaller fish (hereafter, carp exclosure), and (3) reference plots marked with corner stakes but without sides. Study plots were 3 ⫻ 3 m square and were placed in open water areas of known macrophyte beds and spaced at least 4 m apart. Full exclosures were built with sides of 0.25-mm nylon mesh attached to a frame of welded wire fence and supported by corner posts. Carp exclosures were built with 5 ⫻ 10 cm welded steel mesh fencing. In addition to excluding large carp, the mesh fencing may have reduced access of other large animals (e.g., turtles and muskrats) to the plots, but we never sighted either of these species during our regular visits to the exclosures. For both treatments, exclosure sides extended at least 15 cm into the sediments, and exclosures were covered with poultry-wire mesh fencing. Reference plots were marked at all four corners with 2.1 m steel fence posts. Each of the two study sites received seven full exclosures, seven carp exclosures, and 14 reference plots. Treatments were randomly assigned to plots, and plots were arranged parallel to shore and placed so as to standardize water depth (mean depth within cages ⫽ 76 cm in Division Bay, 56 cm in 22-Bay) and maintain

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Figure 1.

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Map of east Delta Marsh, Manitoba, Canada showing location of study sites in 22-Bay and Division Bay.

uniform sediment characteristics. Exclosures were constructed on shore and transported to each site with an airboat to minimize disturbance to the study site. Placement of exclosures was completed on May 25, less than one month after ice-out and before significant above-ground macrophyte growth had begun. Two minnow traps were placed within each full-exclosure plot to capture any fish that invaded as eggs or larvae, and all exclosures were inspected at least twice a week to ensure that large fish had not jumped into or otherwise invaded the plots.

Peak standing crop of S. pectinata foliage occurs by mid-August to early September (Anderson and Low 1976). Therefore, we sampled vegetation biomass in each exclosure between 15 and 22 August. Four randomly selected locations within the interior of each exclosure (⬎ 0.5 m from edge in order to minimize any edge effects) were sampled for estimation of above-ground and below-ground biomass. At each location, a 50 ⫻ 50 ⫻ 65 cm sheet metal quadrant was pushed into the sediments. All foliage was removed from within the quadrant for estimation of above-

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Figure 2. Treatment effects on Chara biomass in 22-Bay. Bars denote mean plus one standard error, n⫽14 enclosures in reference treatment and 7 enclosures for carp and full-exclosure treatments.

ground biomass. Below-ground biomass was estimated by removing all sediment within the quadrant to a depth of 15 cm. The sediment was rinsed and sieved with a 1.6-mm screen to retain roots and tubers. Roots were harvested to a depth of 15 cm to maintain consistency with the earlier studies of Anderson and Low (1976) and Anderson (1978). Vegetation was air-dried for at least 10 d, then sorted by species, oven dried (60⬚ C for 24 h) to a constant mass, and weighed. We evaluated treatment effects on water clarity by measuring light extinction and turbidity on four occasions during the summer. The vertical attenuation of photosynthetic active radiation (PAR) was estimated using a Li-Cor LI-192 underwater quantum sensor to record light intensity at the surface and at a depth of 20 cm within each exclosure. The extinction coefficient kd was estimated as the instantaneous rate of light attenuation (Wetzel and Likens 2000). We sampled turbidity using a 5.5-cm-diameter clear tube, which extended through the entire water column. Water was allowed to mix inside the tube before filling 500-ml sample bottles, which were then read by a turbidimeter. We randomly selected 7 of the 14 reference plots for water clarity measurements (n ⫽ 21 measurements per site). Water clarity measurements were averaged across the summer for each plot, and plot means were the unit of analysis. The effects of exclosure treatment and site on macrophyte biomass variables (above-ground biomass, below-ground biomass, root/shoot allocation) and water clarity (turbidity, extinction coefficient) were analyzed with two-way ANOVA, with plot means as the unit of observation. One abundant macrophyte, Chara sp., was present at just one site, so the analysis of treatment effects on Chara was a simple one-way ANOVA. We also tested whether Chara suppressed S. pectinata at this site by using Chara biomass as a covariate in the

analysis of exclosure effects on S. pectinata biomass. In cases where ANOVA found significant exclosure effects, Tukey’s post-hoc test was used to make pairwise comparisons among exclosure treatments. The design is unbalanced, with more reference plots than full exclosure or carp exclosure plots, which can make ANOVA less robust to violations of assumptions, particularly homogeneity of variances. Data were analyzed with the General Linear Models routine of SPSS 12.0, using type III sums of squares. Type III sums of squares are invariant with respect to cell frequencies and, thus, are useful for unbalanced designs. In addition, the treatment effects presented here are either highly significant or far from significant, so the outcome is unlikely to be affected by the unbalanced design. Data were log-transformed when necessary in order to promote homoscedasticity. RESULTS Two species of submersed macrophytes were abundant in the study plots. Chara sp., a green macroalgae (Chlorophyta: Charophyceae), composed 63% of the above-ground biomass in 22-Bay (reference plot mean) but comprised less than 2% of macrophyte biomass in Division Bay. Stuckenia pectinata comprised most of the remaining macrophyte biomass at the two sites. Myriophyllum sibiricum (Komarov), Utricularia macrorhiza (LeConte), and Vallisneria americana (Michx.) were present in a few plots but comprised less than 1% of macrophyte biomass in all treatments. Because these species were not present in most plots, statistical analysis of treatment effects on their abundances was not possible, and they are not considered further. There was a strong effect of exclosure treatment on Chara biomass in 22-Bay (Figure 2; F2,25 ⫽ 29.2; P ⬍

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Table 1. Biomass of Stuckenia pectinata and water quality parameters within exclosure plots in 22-Bay and Division Bay. There is a significant effect of site on above- and below-ground biomass of S. pectinata. All other treatment and site effects are not significant. Standard errors of means are shown in parentheses. For vegetation variables, n ⫽ 14 for reference pots, 7 for carp and full exclosure plots. For water transparency variables, n ⫽ 7 for all three treatments. 22-Bay

Above-ground biomass (g/m2) Below-ground biomass (g/m2) Root/shoot ratio Extinction coefficient (m⫺1) Turbidity (NTU)

Division Bay

Reference

Carp Exclosure

Full Exclosure

Reference

Carp Exclosure

Full Exclosure

19.4 (4.5)

18.7 (5.3)

6.8 (1.1)

12.5 (3.1)

2.2 (0.8)

9.4 (3.6)

3.6 (0.5) 4.9 (0.7)

3.3 (0.6) 5.7 (1.1)

2.3 (0.3) 2.9 (0.3)

2.4 (0.5) 4.9 (0.6)

1.1 (0.4) 2.5 (0.5)

2.0 (0.7) 4.0 (0.8)

4.3 (0.1) 12.6 (0.2)

4.2 (0.1) 12.8 (0.3)

4.4 (0.1) 12.8 (0.3)

4.4 (0.1) 12.7 (0.4)

4.5 (0.1) 12.2 (0.4)

4.5 (0.1) 12.5 (0.3)

0.001). Full-exclusion plots had the greatest Chara biomass, followed by the carp exclosure and reference plots (Figure 2). Full-exclusion treatment means were significantly different from both carp exclosure and reference means (Tukey’s Test, P ⬍ 0.001), but carp exclosure means were not significantly different from reference means (P ⫽ 0.38). The relative abundance of Chara was also greatest in the full-exclusion treatment, where it comprised 96.4% of the total plant biomass, and lowest in reference plots, where it accounted for 62.6 % of the plant biomass. Exclosure effects on relative abundance were significant (F2,25 ⫽ 3.6; P ⫽ 0.04), with full-exclosure means different from reference means (P ⫽ 0.03), but other pairwise comparisons were not significant (P ⬎ 0.10). Stuckenia pectinata biomass was relatively low, with above-ground biomass in reference plots averaging just 15.9 g m⫺2 and below-ground biomass 3.0 g m⫺2. Exclosure treatments had no effect on aboveground biomass of S. pectinata, below-ground biomass, or shoot/root allocation (P ⬎ 0.10; Table 1). There was a significant effect of study site on S. pectinata biomass, with average above-ground biomass in reference plots 1.76-fold higher in 22-Bay than in Division Bay (F1,50 ⫽ 4.17; P ⫽0.05), and below-ground biomass 1.66-fold higher in 22-Bay than in Division Bay (F1,50 ⫽ 7.33; P ⬍ 0.01; Table 1). There was no site effect on root/shoot allocation (P ⬎ 0.10) and no nonadditive effects of exclosure treatment and site on any of the variables (P ⬎ 0.10 for all interaction terms). An analysis of co-variance (ANCOVA) testing for the joint effects of Chara biomass and exclosure treatments on above-ground biomass of S. pectinata showed that the biomass of S. pectinata was not related to Chara biomass (ANCOVA; F1,24 ⫽0.181; P ⫽0.20). Averaged across the summer, there were no effects

of exclosure treatment on light extinction or turbidity (P ⬎ 0.10 for all treatment effects; Table 1). There was no significant site effect on light extinction, although 22-Bay had slightly higher water clarity (F1,36 ⫽ 3.72; P ⫽0.06) and no site effect on turbidity (Table 1). DISCUSSION Marshes worldwide have been altered by human activities from a clear-water state with abundant submersed vegetation to a turbid state with little or no vegetation (Jeppesen et al. 1998b, Scheffer 1998). In order to restore these systems, it is essential to identify the mechanisms responsible for the state transition. Field experiments can be an effective means of identifying such mechanisms (Dunham and Beaupre 1998). Here, we showed that a local reduction in fish abundance can induce increased plant growth, but this response depends on the initial state of the system. At a site with Chara present, fish exclusion had a strong positive effect on overall abundance of submerged macrophytes, but at a site dominated by S. pectinata, fish exclusion had no effect on macrophyte biomass. The failure of S. pectinata growth rates to respond to either of the fish exclosure treatments was contrary to our expectations, but a review of the literature suggests that this result is consistent with other studies of submersed macrophytes in wetlands. Studies conducted in wetlands dominated by S. pectinata have found that the abundance of this species is often unrelated to water clarity (Bales et al. 1993, Jeppesen et al. 1994, Moss 1994, Jeppesen et al. 1998a). Stuckenia pectinata is relatively tolerant of high turbidity because of its growth habit of forming a canopy of foliage near the water’s surface (Kantrud 1990). We found that the relative abundance of S. pectinata in 22-Bay decreased in fish exclosure plots. Similarly, Anderson (1950) and

Evelsizer & Turner, FISH AND SUBMERSED MACROPHYTES King and Hunt (1967) found that fish removal will allow other species to re-establish themselves in the plant community and eventually lead to a lower abundance of S. pectinata. These other species presumably compete with S. pectinata and reduce its dominance in the plant assemblage. If S. pectinata is less sensitive to turbidity than other taxa, the results of experimental studies examining the effects of improved water clarity on the overall abundance of submersed plants will depend on the availability of propagules for these other, less tolerant taxa. Exclusion of all fish in 22-Bay resulted in a 11.9fold increase in Chara standing crop, but carp exclusion resulted in just a 3-fold increase. This difference suggests that small, planktivorous fish play an important role in regulating water clarity and growth of submerged macrophytes. Planktivore exclusion can have a positive effect on macrophytes via a trophic cascade that results in higher zooplankton abundance, lower phytoplankton abundance, and improved water clarity (Turner and Mittelbach 1992, Carpenter and Kitchell 1993, Schriver et al. 1995, Jeppesen et al. 1997). An important role of planktivorous fish is certainly plausible, as small fish were very abundant in the marsh (Kiers and Hann 1995). Overnight sets of Beamish minnow traps (Evelsizer 2001) showed that fathead minnows (a planktivore) were very abundant at both study sites (catches ⬎ 800 per set). Spottail shiners (Notropis hudsonius Clinton) and Iowa darters (Etheostoma exile Girard) were moderately abundant (catches ⬎ 30 per set), and other zooplanktivorous fish including brook stickleback (Culaea inconstans) and yellow perch (Perca flavescens Mitchell) were also present. Despite the circumstantial evidence implicating small fish and trophic cascades in the suppression of submerged macrophytes, our measurements of water clarity (turbidity and light attenuation) failed to detect consistent treatment effects. This is not entirely surprising, as the mesh walls of the exclosures permitted some water exchange with the surrounding marsh. It is possible that any shifts in water clarity were transitory and largely missed by our snapshot samples, or that they were small in magnitude but important to the plants. It is also possible that some factor other than light availability limited plant growth, but water clarity in 22-Bay was greater than in Division Bay, which is consistent with the greater plant biomass and presence of Chara in 22-Bay. Thus, the results of the exclosure study support the hypothesis that fish contribute to the suppression of vegetation in Delta Marsh, but the mechanism of suppression remains unclear. Large-scale manipulations of Delta Marsh are being considered, and our work was designed to provide data regarding the potential success of such manipulations.

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However, it is important to consider the extent to which our results can be extrapolated to large-scale manipulations (Dunham and Beaupre 1998). Our exclosures permitted water exchange with the surrounding marsh, so treatment effects on water clarity may have been diluted. In addition, propagule limitation may slow the invasion of new species (van der Valk and Davis 1978) and thus inhibit the response of the macrophyte community to short-term improvements in water clarity. These particular processes would tend to make restoration of macrophytes more difficult at small scales than at larger scales, and we expect that exclosures like ours may underestimate the potential for recovery of submerged macrophytes. Anderson (1978, see also Anderson and Low 1976) also studied factors limiting the growth of S. pectinata in Delta Marsh. Their study focused on spatial variation in growth and found that sediment characteristics, exposure to wave action (fetch), and mean water depth were the most important factors in determining the biomass of S. pectinata at a particular site. These data also provided a valuable benchmark that, when contrasted with more recent surveys, show that production of S. pectinata in Delta Marsh has decreased by approximately 40% over the past 30 years (Wrubleski 1998, Evelsizer 2001). The factors identified by Anderson (1978) as responsible for driving patterns of spatial variation within Delta Marsh fail to account for the recent decline of submersed macrophytes, as there is no evidence that sediment characteristics, wave action, and water depth changed appreciably between 1970 and 2000. However, there is evidence that the abundance of planktivorous and benthivorous fish in Delta Marsh has increased (Wrubleski 1998), probably because of stabilized water levels associated with completion of Fairford Dam (Batt 2000). The experimental results presented here show that fish exclusion can increase the biomass of submersed macrophytes and induce a shift in species composition. They further suggest that it is not the direct action of carp uprooting plants that is responsible for macrophyte suppression, but rather some other yet unidentified mechanism associated with the presence of fish. Taken together with earlier studies of Delta Marsh, it seems likely that recent changes in the fish dynamics of Delta Marsh may have played a key role in the decline of submersed macrophytes and that the success of restoration efforts may well hinge on the ability to reduce fish abundance in the marsh. ACKNOWLEDGMENTS We thank the Delta Waterfowl Foundation for providing the financial and logistical support necessary to conduct this work. Todd Arnold was particularly in-

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strumental in the design and execution of this project. Luke Naylor, Chris and John DeRuyke, Faye Babineau, Andrea Evelsizer, Andrea LaShomb, Matt Chouinard, and others generously volunteered their time when extra help was needed. The manuscript benefited from the comments of Sharon Montgomery, Charles Williams, and two anonymous reviewers. LITERATURE CITED Anderson, J. M. 1950. Some aquatic vegetation changes following fish removal. Journal of Wildlife Management 14:206–209. Anderson, M. G. 1978. Distribution and production of sago pondweed (Potamogeton pectinatus L.) on a northern prairie marsh. Ecology 59:154–160. Anderson, M. G. and R. E. Jones. 1976. Submersed aquatic vascular plants of east Delta Marsh. Manitoba Department of Renewable Resources and Transportation Services, Winnipeg, MB, Canada. Anderson, M. G. and J. B. Low. 1976. Use of sago pondweed by waterfowl on the Delta Marsh, Manitoba. Journal of Wildlife Management 40:233–242. Atton, F. M. 1959. The invasion of Manitoba and Saskatchewan by carp. Transactions of the American Fisheries Society 88:203–205. Bales, M., B. Moss, G. Phillips, K. Irvine, and J. Stansfield. 1993. The changing ecosystem of a shallow brackish lake Hickling Broad Norfolk UK II. Long-term trends in water chemistry and ecology and their implications for restoration of the lake. Freshwater Biology 29:141–165. Batt, B. D. J. 2000. The Delta Marsh. p. 17–33. In H. R. Murkin, A. G. van der Valk, and W. R. Clark (eds.) Prairie Wetland Ecology. Iowa State University Press, Ames, IA, USA. Bayley, S. E. and C. M. Prather. 2003. Do wetland lakes exhibit alternative stable states? Submersed aquatic vegetation and chlorophyll in western boreal shallow lakes. Limnology and Oceanography 48:2335–2345. Benndorf, J. 1987. Food web manipulation without nutrient control: a useful strategy in lake restoration? Schweizerische Zeitschrift fu¨r Hydrologie 49:237–248. Bouffard, S. H. and M. A. Hanson. 1997. Fish in waterfowl marshes: waterfowl manager’s perspectives. Wildlife Society Bulletin 25: 146–157. Breukelaar, A. W., E. H. R. R. Lammens, and J. G. P. Klein Breteler. 1994. Effects of benthivorous bream (Abramis brama) and carp (Cyprinus carpio) on sediment resuspension and concentrations of nutrients and chlorophyll-a. Freshwater Biology 32:113–121. Carpenter, S. R. and J. F. Kitchell. 1993. The trophic cascade in lakes. Cambridge University Press, Cambridge, UK. Chambers, P. A. and J. Kalff. 1985. Depth distribution and biomass of submersed macrophyte communities in relation to Secchi depth. Canadian Journal of Fisheries and Aquatic Sciences 42:701–709. Chow-Frazer, P. 1998. A conceptual ecological model to aid restoration of Cootes Paradise Marsh, a degraded coastal wetland of Lake Ontario, Canada. Wetlands Ecology and Management 6:43– 57. deGeus, P. M. J. 1987. Vegetation changes in the Delta Marsh, Manitoba between 1948–80. M.Sc. Thesis. University of Manitoba, Winnipeg, MB, Canada. Duarte, C. M., J. Kalff, and R. H. Peters. 1986. Patterns in biomass and cover of aquatic macrophytes. Canadian Journal of Fisheries and Aquatic Sciences 43:1900–1908. Dunham, A. E. and S. J. Beaupre. 1998. Ecological experiments: scale, phenomenology, mechanism, and the illusion of generality. p. 27–49. In W. J. Resetarits Jr. and J. Bernardo (eds.) Experimental Ecology. Oxford University Press, New York, NY, USA. Evelsizer, V. D. 2001. An investigation of which factors limit sago pondweed growth in Delta Marsh, Manitoba. M.S. Thesis. Clarion University, Clarion, PA, USA. Goldsborough, L. G. 1995. Weather and water quality data summary (1994), University Field Station (Delta Marsh). University of

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