WETLANDS, Vol. 25, No. 3, September 2005, pp. 758–763 q 2005, The Society of Wetland Scientists
MACROINVERTEBRATE RESPONSE TO CATTAIL MANAGEMENT AT CHEYENNE BOTTOMS, KANSAS, USA Richard M. Kostecke1,3, Loren M. Smith1, and Helen M. Hands2 1 Wildlife and Fisheries Management Institute Department of Range, Wildlife, and Fisheries Management Texas Tech University, Lubbock, Texas, USA 79409 2 Cheyenne Bottoms Wildlife Area Kansas Department of Wildlife and Parks 56 NE 40 Road Great Bend, Kansas, USA 67530
3 Present address: The Nature Conservancy P.O. Box 5190 Fort Hood, Texas, USA 76544 E-mail:
[email protected]
Abstract: Cheyenne Bottoms, Kansas, USA has been designated by the Ramsar convention as a Wetland of International Importance. However, since that 1988 designation, cattail (Typha spp.) has become the dominant plant within the basin, and migratory bird use has decreased. We examined the effects of different cattail-management treatments (burned, disked, and grazed by 5 and 20 head of cattle) on macroinvertebrates used as food resources by migratory birds. We found few differences in diversity, biomass, or density of macroinvertebrates among treatments. When differences existed, diversity, biomass, and density were greater within the control or more heavily vegetated treatments (e.g., burned) than within less vegetated treatments (e.g., disked). Macroinvertebrate densities, particularly Chironomidae, ranged from 154 to 681/m2; however, they were up to seven times lower than historic densities and well below the 5000/m2 that has been suggested for supporting large numbers (0.5 million) of migratory waterbirds. Thus, Cheyenne Bottoms’ capacity to support migratory waterbirds may currently be reduced due to low macroinvertebrate densities in areas where cattail has invaded, as well as in areas where cattail has been managed. Research and management should be targeted at restoring the hydrology and dependent biotic communities that support migratory birds. Key Words: disking, grazing, migratory birds, prescribed burning, Typha, wetland macroinvertebrates, waterbirds, wetland management
INTRODUCTION
etative cover-to-water ratios has been investigated (Kaminski and Prince 1981, Murkin et al. 1982, Batzer and Resh 1992), there have been few comparisons of macroinvertebrate response to different types of vegetation manipulation (de Szalay and Resh 1997, Gray et al. 1999). In 1988, the Cheyenne Bottoms basin in central Kansas, USA was designated as a ‘‘Wetland of International Importance’’ by the Ramsar Convention on Wetlands and as a site of hemispheric importance by the Western Hemisphere Shorebird Reserve Network. Historically, the importance of Cheyenne Bottoms to migratory birds has been in part due to its extent of open wetland habitat and productive macroinvertebrate populations (Brooks and Kuhn 1987, Griffith and Welker 1987, Helmers 1991). However, hydrologic al-
Diversity, biomass, and density of wetland macroinvertebrates is regulated, in part, by plant community composition, physiognomic characteristics, and vegetation structure (Macan 1961, Krull 1970, Voigts 1976, Nelson and Kadlec 1984, Olson et al. 1995). Generally, vegetated wetlands harbor more macroinvertebrates than non-vegetated wetlands. Traditionally, wetland managers have manipulated wetland hydrology and vegetation to benefit migratory birds (see Smith et al. 1989). More recently, such manipulations have been used to manage macroinvertebrates as migratory bird food resources (e.g., de Szalay et al. 1996, de Szalay and Resh 1997, Anderson and Smith 2000). Although macroinvertebrate response to varying veg758
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Table 1. Cattail-management treatments and monitoring efforts by year at Cheyenne Bottoms, Kansas, USA. Year Treatment
1999
2000
Burned Burned and Grazed (5 head) Burned and Grazed (20 head)
Burned and monitored three 4-ha plots Burned, grazed and monitored three 4-h plots Burned, grazed, and monitored three 4-ha plots
Burned and Grazed (20 head)
Burned, un-grazed, and no plots established
Burned and Disked Burned and Disked
Burned, disked, and monitored three 4-ha plots Burned, un-disked, and no plots established
Control
Unburned and monitored three 4-ha plots
Monitored same plots Grazed and monitored same plots Un-grazed (inadequate forage) and monitored same plots Three 4-ha plots established, grazed, and monitored Unmonitored as plots not flooded due to drought Three 4-ha plots established, disked, and monitored Unmonitored as plots not flooded due to drought
teration and subsequent cattail (Typha spp.) encroachment began in the 1970s and has resulted in monotypic cattail stands that exclude migratory birds. Average counts of waterfowl and shorebirds in the 1980s were up to twice as large as counts in the 1990s (Kraft 1997; International Shorebird Survey, Manoment Center for Conservation Science, Manoment, Massachusetts, USA, unpublished data), when cattail covered 17 to 90% of each pool (Von Loh and Oliver 1999). Currently, a variety of techniques (e.g., prescribed burning and mechanical disturbance) are used to counter cattail encroachment at Cheyenne Bottoms (Kostecke et al. 2004). We evaluated the effects of cattail management treatments on macroinvertebrate assemblages to understand better their potential influences on food resources for migratory birds at Cheyenne Bottoms. Specifically, we examined the effects of prescribed burning, cattle grazing, and disking treatments on macroinvertebrate familial diversity (i.e., richness and Shannon’s diversity index), biomass, and density. METHODS Cattail Management Cattail coverage in Pool 3 (870 ha) at Cheyenne Bottoms Wildlife Area (CBWA) was approximately 82% in 1998 (Von Loh and Oliver 1999). In 1999, cattail was reduced in different parts of Pool 3 by prescribed burning, prescribed burning followed by cattle grazing (two stocking rates of 5 and 20 head per 11 ha), and prescribed burning followed by deep (15 cm) disking (Kostecke et al. 2004). We randomly established three 4-ha replicates of each treatment (Table 1). Flooding of Pool 3 occurred in mid-August 1999, and mean water depth was 0.62 m. After drawdown during spring 2000, Pool 3 was re-flooded in early September 2000. However, due to drought conditions, not all of Pool 3 could be flooded in 2000, including
all of the areas that contained control and disked replicates, thus affecting our monitoring scheme (Table 1). Because no untreated areas remained in Pool 3, we could not reestablish control replicates elsewhere in the pool. However, additional disking had occurred within Pool 3 during July and August 2000, which allowed us to establish three new, 4-ha disked replicates. Additionally, insufficient forage remained in replicates grazed by 20 head of cattle during 1999; therefore, new areas for these cattle were established for summer 2000 (C. D. Lee, Kansas State University, Manhattan, KS, USA, personal communication). However, in 2000, we continued to monitor replicates grazed by 20 head of cattle during 1999, even though they were not grazed after summer 1999. Mean water depth was 0.41 m during fall 2000. Macroinvertebrate Sampling We sampled macroinvertebrates on a weekly basis for 13 weeks starting on 14 August 1999 and ending on 11 November along twenty 200-m-long transects in each replicate. The first transect within each replicate was 10 m from the edge of the replicate, and spacing between transects was 10 m. In fall 2000, we sampled macroinvertebrates for nine weeks starting on 10 September and ending on 10 November. Termination of sampling each fall coincided with the departure of birds in both years and with freeze-up of the marsh in 2000. During each week, we sampled 10 randomlyselected transects out of the 20 transects established in each replicate; taking two 5-cm-diameter water-column samples (Swanson 1978), one 50 3 50-cm quadrat of clipped vegetation (DeCoster and Persoone 1970), and one 5-cm-diameter (10-cm deep) benthic core sample (Swanson 1983) from a random location along each of transect. We processed samples following Anderson and Smith (1996). We processed water-column samples on
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Table 2. Mean (SE) macroinvertebrate family richness, Shannon diversity index, density (no./m2), and biomass (mg/m2) in cattailmanagement treatments at Cheyenne Bottoms, Kansas during fall 1999. Treatment Variable
Control
Burned
Disked
Grazed, 20 head during 1999
Grazed, 5 head
Richness Shannon’s H Total biomass Total density Non-Chironomidae biomass Non-Chironomidae density Chironomidae biomass Chironomidae density
15.06a (1.18) 0.42a (0.26) 577.78a (225.46) 681 (126) 412.01a (134.19) 53a (9) 165.78a (38.92) 628a (129)
12.05ab (3.22) 0.79a (0.35) 322.33ab (169.75) 431 (183) 87.0b (38.13) 36a (3) 235.34a (122.55) 395a (184)
9.39bc (1.01) 0.46a (0.18) 303.00ab (158.99) 372 (100) 162.81b (50.69) 37a (9) 140.28a (59.45) 335a (89)
7.39c (0.57) 0.57a (0.24) 86.50b (35.81) 153 (70) 44.89b (10.95) 25a (5) 41.67a (27.38) 128a (73)
10.28abc (0.96) 0.78a (0.20) 247.61b (140.37) 215 (88) 181.88b (53.64) 54a (7) 65.72a (44.72) 161.36a (78)
Notes: For a variable, means followed by the same letter were not different (P . 0.05).
site. We placed benthic and epiphytic samples in plastic bags, which were labeled and refrigerated until we could process them. We processed all samples within 10 days. We used a 600-mm sieve to retain macroinvertebrates (Huener and Kadlec 1992). After sorting, we preserved samples in 70% ethyl alcohol. We used Merritt and Cummins (1996) and Pennak (1989) for identification of aquatic insects and other aquatic invertebrates, respectively. We used Borror et al. (1992) for identification of terrestrial arthropods. We identified invertebrates to family or, in a few instances, to order. We counted all macroinvertebrates within samples by family so that we could calculate densities per m2. We estimated biomass (mg dry mass/ m2) of each family within samples by drying macroinvertebrates at 558 C to a constant mass and weighing them on an analytical balance (60.001 g). We also calculated familial diversity (richness and Shannon’s diversity index; Magurran 1988) of each sample. Statistical Analyses We used multivariate analysis of variance (MANOVA) to test for the affects of treatments and the biweekly period 3 treatment interaction on eight response variables: 1) macroinvertebrate familial richness, 2) Shannon’s diversity index, 3) total macroinvertebrate biomass, 4) total macroinvertebrate density, 5) total non-Chironomidae biomass, 6) total non-Chironomidae density, 7) Chironomidae biomass, and 8) Chironomidae density. Because of differences in hydrology and location of replicates between years, we conducted separate analyses for 1999 and 2000. Although we sampled macroinvertebrates on a weekly basis, we averaged data across two-week periods to reduce the effects of the lack of statistical independence because our response variables are highly correlated (Milliken and Johnson 1992, Anderson and Smith 2000). For the same reason, we did not use re-
peated measures analysis of variance, but instead used MANOVA, which is much more robust towards problems of statistical independence (Milliken and Johnson 1992). We conducted MANOVAs so that each twoweek period was a separate variable in the same multivariate analysis. We also used a log (x 1 1) transformation to meet parametric assumptions of normality and to make variances more homogeneous (Zar 1996). However, in our results, we present non-transformed means and standard errors. Chironomidae was the only family we chose to analyze because chironomids dominated the macroinvertebrate biomass and density in our samples (Kostecke 2002) and are known to be a key food resource for waterbirds (e.g., Helmers 1991). Within each treatment, chironomids accounted for 63 6 8% of the total macroinvertebrate biomass and 90 6 2% of the total macroinvertrbate density during 1999 and 2000. We conducted all tests with SAS (SAS Institute 1999) using Wilk’s l as the test statistic and an alpha level of 0.05. Bonferroni t tests were used to determine differences between treatment means. If an interaction effect was detected, we used a 1-way ANOVA to examine treatment effects within biweekly time periods. RESULTS Family richness (F4, 10 5 6.58, P 5 0.01), but not Shannon’s diversity index (F4, 10 5 1.15, P 5 0.39), differed among treatments in 1999 (Table 2). The control had the greatest richness, whereas the grazed (20 head) treatment had the lowest richness. Non-chironomidae density (F4, 10 5 1.90, P 5 0.19), Chironomidae biomass (F4, 10 5 1.79, P 5 0.21), and Chironomidae density (F4, 10 5 2.04, P 5 0.16) did not differ among treatments (Table 2). Total macroinvertebrate biomass was greater in the control than in the grazed treatments (F4, 10 5 4.10, P 5 0.03). We detected a biweekly period X treatment interaction for total macroinvertebrate
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Table 3. Mean (SE) macroinvertebrate family richness, Shannon’s diversity index, density (no./m2), and biomass (mg/m2) in cattailmanagement treatments at Cheyenne Bottoms, Kansas during fall 2000. Treatment Variable
Burned
Disked
Grazed, 20 head during 1999
Grazed, 20 head during 2000
Grazed, 5 head
Richness Shannon’s H Total biomass Total density Non-Chironomidae biomass Non-Chironomidae density Chironomidae biomass Chironomidae density
23.50a (3.06) 0.80a (0.28) 664.00a (393.47) 320a (125) 303.87a (237.51) 32a (3) 360.25a (148.86) 288a (124)
11.00bc (3.37) 0.45a (0.22) 504.67a (167.07) 320a (91) 38.08a (28.38) 21a (5) 466.58a (169.86) 298a (89)
15.25bc (2.27) 0.27a (0.06) 701.76a (244.79) 505a (180) 54.80a (3.96) 26a (2) 647.17a (236.25) 479a (176)
6.50c (1.35) 0.24a (0.11) 426.92a (179.71) 259a (79) 11.95a (5.48) 9b (4) 415.00a (172.63) 250a (75)
18.58ab (2.25) 0.75a (0.43) 653.25a (182.52) 287a (78) 315.76a (154.06) 26a (6) 337.50a (119.18) 261a (80)
Notes: For a variable, means followed by the same letter were not different (P . 0.05).
density (F20, 21 5 3.38, P , 0.01). Total macroinvertebrate density differed among treatments during biweekly periods 4 and 5 (F4, 10 $ 6.03, P # 0.01). Total macroinvertebrate density in the control (mean 6 SE 5 794 6 75) was at least 7-fold greater than in the disked (142 6 24) and grazed (20 head) (114 6 37) treatments during period 4. During period 5, total macroinvertebrate density in the control (1417 6 209) was 11 times greater than in the grazed (20 head) treatment (126 6 93). Non-chironomidae biomass was greatest in the control (F4, 10 5 5.04, P 5 0.02) (Table 2). As in 1999, richness (F4, 10 5 10.72, P , 0.01), but not diversity (F4, 10 5 1.59, P 5 0.25), differed among treatments in 2000 (Table 3). In the absence of a control, the burned treatment had greatest richness; whereas the plots grazed by 20 head of cattle during 2000 had lowest richness, similar to 1999. Total macroinvertebrate biomass (F4, 10 5 0.32, P 5 0.86) and density (F4, 10 5 1.17, P 5 0.38), non-Chironomidae biomass (F4, 10 5 1.70, P 5 0.23), and Chironomidae biomass (F4, 10 5 0.57, P 5 0.69) and density (F4, 10 5 1.12, P 5 0.40) did not differ among treatments (Table 3). Non-chironomidae density was lowest in plots grazed by 20 head of cattle during 2000 (F4, 10 5 5.98, P 5 0.01) (Table 3). DISCUSSION Few differences in macroinvertebrate assemblages existed among the burned, disked, and grazed treatments at Cheyenne Bottoms. Although dependent on the intensity of the disturbance, macroinvertebrate assemblages are often similar pre- and post-vegetation manipulation (Batzer and Resh 1992, de Szalay et al. 1996, de Szalay and Resh 1997, Gray et al. 1999). Few treatment differences could exist for a variety of reasons. Because chironomids dominated the macroinvertebrate fauna within our samples, macroinvertebrate assemblages were highly uneven, and thus, di-
versity indices were uniformly low across treatments. Relative to biomass and density, reduced vegetative substrates available for colonization immediately following treatment might have been ameliorated by colonization or re-growth of vegetation (Murkin et al. 1982, Frid et al. 1999). Further, energy may not have been limiting to macroinvertebrates because detritus was present in all replicates. Although detritus levels may have differed among treatments, such differences might not affect macroinvertebrate production (Nelson et al. 1990, Murkin et al. 1992). Indeed, the amount of detritus needed to support high macroinvertebrate production may be relatively low (Neckles et al. 1990). Aquatic macroinvertebrates also tend to be spatially aggregated due to the patchiness of suitable habitat within a wetland (Downing 1991), which can result in large variance in density and biomass. Lack of treatment effects could also be masked by the effect of drying and reflooding the pool on an annual basis. Time since reflooding may not have been sufficient to allow for the development of treatment differences (Anderson and Smith 2000). Finally, differences among treatments may only have been detectable below the family level. When differences did exist, the more heavily vegetated treatments (e.g., control and burned; Kostecke et al. 2004) tended to have greater macroinvertebrate family richness, biomass, and density than disked and grazed treatments. Several taxa were found exclusively in the control and burned treatment (Kostecke 2002), thus increasing family richness. The presence of these taxa is likely related to the more complex habitat structure within the control and burned treatment, which had greater vegetative biomass, density, and diversity (Kostecke et al. 2004). Differences in biomass and density are also related to habitat structure. For example, taxa with greater biomass (e.g., Hydrophilidae) were found in greater densities within the burned treat-
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ment and, particularly, the control (Kostecke 2002). More complex habitat structure can accommodate a greater diversity of macroinvertebrates, as well as greater biomass and density (Krull 1970, Voights 1976, Nelson and Kadlec 1984, Olson et al. 1995). However, although greater macroinvertebrate diversity, biomass, and density may be found in more heavily vegetated areas (e.g., areas invaded by cattail), many migratory birds avoid such dense vegetation (Weller and Fredrickson 1974, VanRees-Siewert and Dinsmore 1999). Relatively high densities of important prey taxa, such as chironomids (Griffith and Welker 1987, Helmers 1991) are typically viewed as being an essential habitat element for some wetland birds (Davis and Smith 1998, Loesch et al. 1999). Chironomid densities within cattail-management treatments at CBWA were similar to densities found within other recently sampled habitats at CBWA (H. M. Hands, unpublished data), as well as other prairie wetlands (Davis and Smith 1998). However, our chironomid densities were 6 to 7 times lower than those reported at CBWA by Helmers (1991). This difference in chironomid density is possibly related to differences in environmental conditions (e.g., hydroperiod and sedimentation), as well as differences in study designs (e.g., whether bird exclosures were used and whether disturbed areas were sampled). Current chironomid densities at CBWA were well below 5000/m2, the density that may be needed to support large numbers (i.e, 0.5 million) of migratory birds (Davis and Smith 1998, Loesch et al. 1999). However, no management treatment in our study offered a particular advantage to migratory waterbirds in terms of increasing chironomid biomass and density. However, cattail control may still benefit waterbirds by removing dense vegetation that precludes their use of an area (Kostecke 2002). Additionally, our results suggest that wetland managers at CBWA should investigate hydrologic and biotic interactions other than cattail invasion as a cause for the reduction in macroinvertebrates. Restoration of natural hydrologic fluctuations and sediment removal may be needed before CBWA again has the capacity to support the large numbers of migratory waterbirds that historically used the basin. ACKNOWLEDGMENTS We thank the manager, K. Grover, and staff of Cheyenne Bottoms Wildlife Area for implementation of cattail-management treatments. K. Schneweis assisted with invertebrate processing. K. Grover, S. Phillips, M. Wallace, G. Wilde, and two anonymous reviewers commented on the manuscript. B. Harrington (Manomet Center for Conservation Sciences) gra-
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