WETLANDS, Vol. 28, No. 3, September 2008, pp. 874–881 ’ 2008, The Society of Wetland Scientists
NOTE INVERTEBRATE COMMUNITY VARIATION IN SEASONAL FOREST WETLANDS: IMPLICATIONS FOR SAMPLING AND ANALYSES Anthony T. Miller1,6, Mark A. Hanson1,2, James O. Church1,3, Brian Palik4, Shane E. Bowe5,7, and Malcolm G. Butler1 1 Department of Biological Sciences North Dakota State University Fargo, North Dakota, USA 58102 2
Minnesota Department of Natural Resources Wetland Wildlife Populations and Research Group 102 23rd St. NE Bemidji, Minnesota, USA 56601 E-mail:
[email protected] 3
Department of Ecology, Evolution, and Organismal Biology Iowa State University Ames, Iowa, USA 50011 4
Northern Research Station USDA Forest Service, Forestry Sciences Lab 1831 Hwy. 169 E Grand Rapids, Minnesota, USA 55744 5
Department of Biology Bemidji State University Bemidji, Minnesota, USA 56601 6
Present Address: Third Rock Consultants, LLC 2514 Regency Road, Suite 104 Lexington, Kentucky, USA 40503 7
Present Address: Red Lake Department of Natural Resources 15761 High School Drive Red Lake, Minnesota, USA 56671 Abstract: Using data sets from two separate studies, we assessed within-year variation in aquatic invertebrate communities in 31 seasonally flooded (seasonal) wetlands in aspen (Populus spp.) – dominated forests in north central Minnesota. Principal components analysis (PCA) indicated that, in each case, three axes explained . 55% of variance in aquatic invertebrates, with the first axis strongly correlated with sampling date. Indicator species tests showed that this variation along axis 1 was largely due to shifts in abundance of crustaceans, Diptera and other insects, leeches, and other taxa. Temporal shifts in aquatic invertebrate community structure pose a major obstacle for ecological studies of aquatic invertebrates in seasonal forest wetlands and should receive more attention from investigators planning research in these and perhaps other wetland habitats. Key Words: aquatic invertebrates, indicator species analysis, principal components analysis, seasonal forest wetlands, temporal variability
874
Miller et al., VARIATION IN WETLANDS: IMPLICATIONS FOR SAMPLING INTRODUCTION It is widely believed that invertebrate communities have potential to reflect ecological disturbance in aquatic habitats, including wetlands (Adamus 1996, Rosenberg and Resh 1996). Ecologists have long held that invertebrate communities in ponds and seasonally flooded wetlands were constrained by timing and duration of flooding, desiccation, temperature, chemical gradients, predation (Wiggins et al. 1980, Wellborn et al. 1996, Williams 1996, Brooks 2000), and by interactions between these abiotic and biotic processes (Schneider and Frost 1996) resulting in considerable regional variation in community composition across North America (Batzer et al. 2005). Given this coupling to environmental gradients, use of community-level analyses of wetland invertebrates has potential to reflect ecological changes following disturbance (Gernes and Helgen 2002, Burton et al. 1999). However, extreme temporal variation within invertebrate communities influences statistical inferences, and may be especially important when resulting data are related to environmental gradients or compared to features of wetlands or adjacent uplands. Small, seasonally flooded wetlands (Stewart and Kantrud 1971) are common throughout some forested regions of North America (Brooks et al. 1998, Palik et al. 2001, Palik et al. 2003) and are increasingly valued for their functional roles in maintaining natural patterns of biodiversity, hydrology, and landscape integrity both within and beyond wetland boundaries (Oertli 1993, Biggs et al. 1994, Niering 1997, Colburn 2004). Aquatic invertebrates are often the most abundant macrofauna in these sites (Brooks 2000) and have long been considered key elements of wetland food webs (Murkin 1989). More broadly, aquatic invertebrate communities integrate complex abiotic and biotic features of wetland environments; thus these assemblages are thought to reflect wetland characteristics and to indicate changes in functional relationships within and across wetland boundaries (Adamus 1996, Gernes and Helgen 2002). However, aquatic invertebrate communities are not always useful indicators of wetland condition or disturbance because community associations with environmental gradients are sometimes weak (Tangen et al. 2003). Furthermore, Palik et al. (2001) and Batzer et al. (2004) reported only weak relationships between invertebrate community structure and environmental characteristics such as water chemistry, nutrients, vegetation, or even hydroperiods of seasonal forest wetlands and Williams (1996) questions whether environmental gradients actually constrain inverte-
875
brates in temporary waters. Although wetland invertebrate populations are known to show high seasonal variability, little research has specifically assessed magnitude of temporal variability or its implications for sampling, analysis, or data interpretation. This is an important oversight if natural chronological variability masks other major sources of variance in invertebrate community structure, thus limiting usefulness of wetland invertebrates as ecological indicators. To illustrate the magnitude and consequences of within-year variation in aquatic invertebrate communities in wetland sites, we assessed invertebrate community composition and measured temporal patterns of variability in seasonal wetlands in a heavily forested region of north central Minnesota (USA). We used indirect gradient analyses, indicator species tests, and analysis of variance to portray intra-year patterns in aquatic invertebrates and to illustrate the extent to which sample collection date influenced a simple community analysis. METHODS Study Area We selected aquatic invertebrate data sets gathered during a single year from 31 seasonal forest wetlands. All study sites were located in north central Minnesota (USA) within a region classified as the Minnesota Drift and Lake Plain Section (Keys et al. 1995), an area consisting of ground and stagnation moraines, lake and outwash plains, and supporting dense mixtures of deciduous and coniferous forests (Almendinger and Hanson 1998). This landscape is overlain by deep, but variable, glacial till with extensive areas of lakes and wetlands. Data sets were derived from wetlands that were also sites of other concurrent studies, a first comprising 24 seasonal wetlands near Bemidji, MN, and a second based on seven similar sites near Remer, MN. Invertebrate samples were gathered in various years during 1999–2005, but data reported here were collected in 2000 (Remer sites) and 2005 (Bemidji sites). These specific data were selected for the present study because, during these years, we gathered samples from multiple sampling periods, allowing better analysis of temporal variation. Study site selection was constrained by several criteria, which differed slightly between the two broader research efforts. We used wetland surface area, emergent vegetation (often presence of black ash (Fraxinus nigra Marsh.) to indicate seasonal
876
WETLANDS, Volume 28, No. 3, 2008
inundation), and regional experience to identify sites with seasonal hydroperiods (typically 60–90 days). Bemidji sites were imbedded in aspen stands of various age-since-harvest. These wetlands ranged from 0.2–0.5 ha in surface area and had variable crown closure (5%–100%, estimated using spherical densiometers, Lemmon 1956). Emergent hydrophytes were common in Bemidji sites and included Carex utriculata Boott, Carex lacustris Willd., Sium suave Walt., Equisetum spp., and others. In contrast, Remer sites were all located immediately adjacent to uplands dominated by mature aspen (Populus spp.) where timber harvest had not occurred during the past 70–90 years. Most Remer sites were # 0.2 ha in surface area and were characterized by at least 75% peak crown closure. These sites supported little if any non-woody emergent hydrophytic vegetation, although Carex spp. occurred sporadically and were very abundant in some locations. In general, Bemidji wetlands were somewhat larger and exhibited longer hydroperiods compared to Remer sites. Land ownership of study areas was variable, but sites occurred within private, corporate, county, and state areas. Invertebrate Sampling Samples of aquatic invertebrates were collected from Remer sites at two-week intervals from May 1 to June 19, 2000; Bemidji sites were sampled approximately monthly from May 5 to July 11, 2005. Sampling began two-three weeks after iceout and continued until wetlands initially dried, or until early July. This resulted in four sets of samples from Remer sites and three sets from Bemidji sites. We used surface-associated activity traps (SATs) (Hanson et al. 2000) to sample aquatic and semi-aquatic invertebrates in all sites. SATs are especially useful in shallow wetlands because they gather representative samples, capture higher diversity of organisms than do conventional activity traps, and simultaneously collect animals both in the water column and associated with the water surface (Corixidae, Gerridae, Gyrinidae, etc.; Hanson et al. 2000). Invertebrates were gathered using one SAT placed along each of five randomly chosen transects in each study wetland. When emergent hydrophytes were present along transects, SATs were set at the deepest margin of the macrophyte zone; otherwise, traps were placed in open water midway along transects, equidistant between the wetland center and edge. Water depths at trap sites varied, but were rarely . 0.5 m, and typically decreased during the year. SATs were deployed for 24 hr by attachment to
Figure 1. Axis 1 scores (depicting seasonal variation in invertebrates) from seasonal ponds near A) Remer, MN, during 2000 and B) Bemidji MN, during 2005. Box plots depict median values (central horizontal line), along with the 10th, 25th, 75th, and 90th percentiles and outliers beyond 10 and 90 percentiles (indicated by points).
PVC frames fastened in sediments. Trap contents were condensed in the field by passage through a 0.4 mm-mesh sieve and preserved in 70% ethanol. Invertebrates were sorted, identified to the lowest feasible taxonomic level (typically family for insects and genus for crustaceans), and enumerated in the lab using stereomicroscopes. In some cases, we formed aggregate taxa to achieve better agreement between numbers of study sites (time/ site combinations) and invertebrate taxa (responses). Numbers of invertebrates collected in the five SATs in each wetland were combined, producing a single index to abundance value for invertebrates on each sampling date. Due to differences in design of parent studies, sampling frequencies, and
Miller et al., VARIATION IN WETLANDS: IMPLICATIONS FOR SAMPLING
877
Table 1. Eigenvalues and cumulative percent variance extracted from the first three ordination axes fit to invertebrate scores from wetland sites near Remer, MN. Axes are considered significant when observed eigenvalues exceed brokenstick. Axis
Eigenvalue
Percent Variance Explained
Cumulative Percent Variance Explained
Broken-stick Eigenvalue
1 2 3
327.80 235.47 168.47
25.80 18.53 13.26
25.80 44.33 57.59
181.13 136.07 112.54
potential wetland characteristics, data from Bemidji and Remer sites were analyzed separately. Unfortunately, we were unable to sample all wetland sites within a 24 hr period. Thus, samples from Remer sites were gathered over periods of three-four days during 2000; Bemidji sites were sampled during periods of 11–12 days during 2005. Statistical Analyses We used indirect gradient analysis (principal components analysis, PCA) to illustrate temporal variation in wetland invertebrate communities (ter Braak 1995, ter Braak and Smilauer 1998). PCA is sometimes unsuitable for community analyses (McCune and Grace 2002), but for several reasons we believe it was appropriate in this case. First, prior to PCA, all invertebrate data were natural-log (n+1) transformed to increase homogeneity of variance. Second, examination of resulting PCA plots revealed no distortion (‘‘arch’’) in ordination space. Finally, preliminary detrended correspondence analysis showed that axis lengths were short (, 2.0 standard deviations), indicating that linear models (such as PCA) were appropriate. We performed PCA using PC-ORD 4 (McCune and Mefford 1999). To identify variance contributed by time, we examined biplot overlays using Julian dates corresponding to the first date of each sampling period. To assess significance of patterns observed in PCA, we compared axis 1 site scores for each wetland among sampling (time) periods using repeated measures analysis of variance (rANOVA, SAS Proc Mixed)
followed by mean separation tests (LS Means) (Littell et al. 1996). To identify invertebrate taxa associated with specific sampling periods, we performed indicator species analyses using PC-ORD (McCune and Mefford 1999). As with PCA, indicator species values and comparisons were based on total catches of invertebrates in five SATs on each sampling date, and analyses were conducted separately using data from Bemidji and Remer sites. Monte Carlo tests derived from 5,000 permutations were subsequently applied to assess significance of maximum indicator species values (P , 0.05) (McCune and Mefford 1999). RESULTS AND DISCUSSION PCA identified three significant axes that cumulatively explained 57.6% of the variance in invertebrate site scores from Remer wetlands (Table 1) and four significant axes that cumulatively explained 65.1% of the variance in scores from Bemidji sites (Table 2). In both data sets, site scores shifted dramatically along axis 1, reflecting obvious transition in composition of aquatic invertebrates in our study wetlands from May 1 to June 19 (Remer sites, Figure 1a), and May 5 to July 11 (Bemidji sites, Figure 1b). Axis 1 explained 25.8% and 29.3% of variance in invertebrate community site scores in Remer and Bemidji wetlands, respectively. In both data sets, site scores were highly correlated with sampling (Julian) dates (r 5 0.79, Remer sites; r 5 0.87, Bemidji sites), indicating that these axes
Table 2. Eigenvalues and cumulative percent variance extracted from the first four ordination axes fit to invertebrate scores from wetland sites near Bemidji, MN. Axes are considered significant when observed eigenvalues exceed brokenstick. Axis
Eigenvalue
Percent Variance Explained
Cumulative Percent Variance Explained
Broken-stick Eigenvalue
1 2 3 4
700.23 355.05 287.32 214.61
29.29 14.85 12.02 8.98
29.29 44.14 56.16 65.14
430.02 310.50 250.73 210.89
878
WETLANDS, Volume 28, No. 3, 2008
Table 3. Results of indicator species tests assessing distribution of selected invertebrate taxa among sampling dates during 2000 (Remer sites). Maximum values reflect percent perfect agreement and X indicates sampling periods in which highest values were observed (McCune and Mefford 1999). Significance was inferred at P . 0.05. Taxa with significant P values showed association with indicated sampling period (May 1 5 1, May 15 5 2, June 5 5 3, June 19 5 4). P-values were derived from Monte-Carlo simulations using 5,000 permutations. Sampling Period Taxon Hirudinea Crustacea
Hydracarina Insecta
Mollusca
1
2
3
4
x Eubranchipus Conchostraca Daphnia Simocephalus Chydorus Scapholeberis Cyclopoida Calanoida Harpacticoida Ergasilus Ostracoda Collembola Odonata Hemiptera Trichoptera Haliplidae Dytiscidae Gyrinidae Hydrophilidae Unidentified Coleoptera Dixidae Culicidae Chaoboridae Unidentified Diptera Gastropoda
derived largely from seasonal chronology during these respective 50- and 62-day periods. Comparisons (rANOVA) among our sampling periods also indicated that invertebrate site scores varied through time in both Remer (F3,24 5 13.2; P , 0.01) and Bemidji data sets (F2,53.6 5 160.2; P , 0.01). LS means comparisons indicated that axis 1 site scores from Remer wetlands differed at intervals greater than approximately 15 days (P , 0.01) and Bemidji site scores differed among all three periods (P , 0.01) when samples were gathered at approximately 30-day intervals. Results from indicator species analysis help explain temporal patterns observed in our study wetlands. During initial sampling of Remer wetlands (May 1), Diptera (Culicidae, Chaoboridae, and Dixidae), Eubranchipus sp., Trichoptera, and Hydracarina were collected in a higher proportion of sites than was expected by chance, and these organisms dominated invertebrate communities
x x x x x x x x x x x x x x x x x x x x x x x x x x
Maximum Value Within Period
P-Value
78 57 31 56 47 30 61 44 15 45 36 34 49 36 28 44 69 44 38 18 32 52 77 98 77 33 57
, 0.01 , 0.01 0.39 0.20 0.32 0.84 , 0.01 0.11 0.81 0.13 0.38 0.69 , 0.01 0.79 0.42 0.18 , 0.01 0.13 0.07 0.97 0.28 , 0.01 , 0.01 , 0.01 , 0.01 0.65 0.23
(period 1, Table 3). Only one taxon, a zooplankter (Scapholeberis sp.), was associated with samples gathered during our May 15 effort (period 2). Hirudinea and our aggregate category for unidentified Coleoptera (other) were significant indicator taxa during our third and fourth sampling periods (periods 3 and 4, respectively; Table 3). Indicator species tests also identified taxa associated with temporal patterns in data from Bemidji sites. Here again, during our early sampling (period 1, May 5), Culicidae, Eubranchipus, and Trichoptera were collected in a higher proportion of sites than was expected by chance, along with Copepoda, Ostracoda, and Hydracarina (Table 4). Taxa associated with our second sampling of Bemidji sites included a group of other Diptera (various feeding groups) and Hirudinea (indicator values for Cladocera approached our chosen alpha, P 5 0.055). Late season (period 3, July 11) taxa from Bemidji sites included Chaoboridae, Odonata, pre-
Miller et al., VARIATION IN WETLANDS: IMPLICATIONS FOR SAMPLING
879
Table 4. Results of indicator species tests assessing distribution of selected invertebrate taxa among sampling dates during 2005 (Bemidji sites). Maximum values reflect percent perfect agreement among periods and X indicates sampling periods in which highest values were observed (McCune and Mefford 1999). Significance was inferred at P . 0.05. Taxa with significant P values showed association with indicated sampling period (May 5 5 1, June 6 5 2, July 11 5 3). P-values are derived from Monte-Carlo simulations using 5,000 permutations. Sampling Period Taxon Hirudinea Crustacea
Hydracarina Insecta
Mollusca
1
2
3
x Eubranchipus Conchostraca Cladocera Copepoda Ostracoda Odonata Predacious Hemiptera1 Other Hemiptera-various feeding groups2 Trichoptera Predacious Coleoptera3 Other Coleoptera-various feeding groups4 Culicidae Chironomidae Chaoboridae Other Diptera-predacious5 Other Diptera-various feeding groups6 Sphaeriidae
x x x x x x x x x x x x x x x x x x
MaximumValue Within Period
P-Value
36 75 36 37 41 38 40 43 54 44 21 36 34 69 37 55 33 38 42
, 0.02 , 0.01 0.37 0.06 , 0.01 , 0.01 , 0.01 , 0.01 , 0.01 , 0.01 0.05 0.18 0.71 , 0.01 0.12 , 0.01 0.15 0.05 , 0.01
1
Includes Belostomatidae, Gerridae, Hydrometridae, Nepidae, Notonectidae, Pleidae, Veliidae. 2 Predominantly Corixidae; occasionally other Hempitera. 3 Includes Carabidae, Dytiscidae, Gyrinidae, Hydrophilidae. 4 Includes Curculionidae, Haliplidae, Scirtidae, Hydraenidae. 5 Includes Ceratopogonidae, Sciomyzidae, and Tabanidae. 6 Includes Dixidae, Stratiomyidae, Tipulidae, and unidentified taxa.
daceous (Belostomatidae, Notonectidae, Veliidae) and other Hempitera (various feeding groups, Corixidae and others), and sphaeriid clams (Sphaeriidae) (Table 4). Wetland invertebrate communities are comprised of organisms characterized by diverse life history strategies in response to environmental gradients such as duration of ponding, desiccation, habitat isolation, biological interactions, and other site-level features of wetland habitats (Wiggins et al. 1980, Schneider and Frost 1996, Williams 1996, Hanson et al. 2005). As described by Wiggins et al. (1980), community composition changes dramatically throughout the year in temporary and seasonal wetlands as over-wintering detritivores and herbivores (e.g., Anostraca, Cladocera, Copepoda) are supplemented and sometimes replaced by nonwintering migrants (e.g., Coleoptera, Hemiptera). Because these communities exhibit extreme spatial and temporal variability, they present significant sampling and analysis challenges for investigators. Many useful devices and techniques have been developed specifically for wetland invertebrate sam-
pling (Murkin et al. 1983, Ross and Murkin 1989, Brinkman and Duffy 1996, Turner and Trexler 1997, Hanson et al. 2000, and many others). However, we see less evidence of concern for proper analysis and interpretation of temporal variability, even though it is an obvious and major source of invertebrate community variability in wetland habitats. Obtaining reliable samples of aquatic invertebrate populations is critical when presence/absence or quantitative abundance data is to be related to environmental gradients or ecosystem disturbance (Adamus 1996, Rosenberg and Resh 1996, Gernes and Helgen 2002). We believe that, in seasonal forest wetlands and elsewhere, investigators will benefit from sampling and analysis strategies that better control for, or even target, temporal dynamics in invertebrate populations. Our examples illustrated that invertebrate samples collected at intervals greater than approximately 15 days should be expected to exhibit statistical differences, even when gathered from the same un-manipulated wetland sites. These findings are to be expected and may be conservative because our community analyses were based on crude
880
WETLANDS, Volume 28, No. 3, 2008
aggregate taxa. Without allowance for within-year temporal variability, collection, analysis, and interpretation of invertebrate data sets often will provide little insight because within-year dynamics are almost certain to obscure informative patterns. Simple strategies for improving interpretation of invertebrate community data might include: 1) simultaneous sampling of invertebrates from multiple wetlands (same day or week), 2) sampling on similar dates each year during multi-year studies, or 3) coordinating efforts so as to better coincide with site-specific phenologies of key indicator organisms. We concur with previous authors who have suggested that better understanding of temporal change in invertebrate populations is needed for future invertebrate-based biomonitoring efforts (Rosenberg and Resh 1996) and for modeling community dynamics (Williams 1996). This need seems even greater for community-level analyses in temporary habitats and using combinations of animals with highly variable life histories. ACKNOWLEDGMENTS Funding for this project was provided by the Northern Research Station, USDA Forest Service, Minnesota Department of Natural Resources Division of Fish and Wildlife, and the North Dakota Water Resources Research Institute through a graduate research fellowship to A.T. Miller. We thank Potlach Corporation and Environmental Services Departments in Beltrami and Cass County, MN, for project coordination and access to wetlands used in this research. We also thank Brian Herwig, Jeffrey Lawrence, and three anonymous referees for helpful reviews of this manuscript. LITERATURE CITED Adamus, P. R. 1996. Bioindicators for assessing ecological integrity of prairie wetlands. Environmental Protection Agency. E.P.A. National Health and Environmental Effects Research Laboratory, Western Ecology Division, Corvallis, OR, USA. EPA/600/R-96/082. Almendinger, J. C. and D. S. Hanson. 1998. Ecological Land Classification Handbook for the Northern MN Drift and Lake Plains and the Chippewa National Forest. Minnesota Department of Natural Resources, Saint Paul, Minnesota, USA. Batzer, D. P., S. E. Dietz-Brantley, B. E. Taylor, and A. E. DeBiase. 2005. Evaluating regional differences in macroinvertebrate communities from forested depressional wetlands across eastern and central North America. Journal of the North American Benthological Society 24:403–14. Batzer, D. P., B. J. Palik, and R. Buech. 2004. Relationships between environmental characteristics and macroinvertebrate communities in seasonal woodland ponds of Minnesota. Journal of the North American Benthological Society 23:50–68. Biggs, J., A. Corfield, D. Walker, M. Whitfield, and P. Williams. 1994. New approaches to the management of ponds. British Wildlife 5:273–87.
Brinkman, M. A. and W. G. Duffy. 1996. Evaluation of four wetland aquatic invertebrate samplers and four sample sorting devices. Journal of Freshwater Ecology 11:193–200. Brooks, R. T. 2000. Annual and seasonal variation and the effects of hydroperiod on benthic macroinvertebrates of seasonal forest (‘‘vernal’’) ponds in central Massachusetts, USA. Wetlands 20:707–15. Brooks, R. T., J. Stone, and P. Lyons. 1998. An inventory of seasonal forest ponds on the Quabbin Reservoir watershed Massachusetts. Northeastern Naturalist 5:219–30. Burton, T. M., D. G. Uzarski, J. P. Gathman, J. A. Genet, B. E. Keas, and C. A. Stricker. 1999. Development of a preliminary invertebrate index of biotic integrity for Lake Huron. Wetlands 19:869–82. Colburn, E. A. 2004. Vernals pools: Natural History and Conservation. McDonald and Woodward Publishing Co., Blacksburg, VA, USA. Gernes, M. C. and J. C. Helgen. 2002. Indexes of biological integrity (IBI) for large depressional wetlands in Minnesota. Minnesota Pollution Control Agency, Saint Paul, MN, USA. Hanson, M. A., C. C. Roy, N. H. Euliss, Jr., K. D. Zimmer, M. R. Riggs, and M. G. Butler. 2000. A surface-associated activity trap for capturing water-surface and aquatic invertebrates in wetlands. Wetlands 20:205–12. Hanson, M. A., K. D. Zimmer, M. G. Butler, B. A. Tangen, B. R. Herwig, and N. H. Euliss, Jr. 2005. Biotic interactions as determinants of ecosystem structure in prairie wetlands: an example using fish. Wetlands 25:764–75. Keys, J., Jr., C. Carpenter, S. Hooks, F. Koenig, W. H. McNab, W. Russell, and M. L. Smith. 1995. Ecological units of the Eastern United States - first approximation (map and booklet of map unit tables). U.S. Forest Service, Atlanta, GA, USA. Lemmon, R. E. 1956. A spherical densiometer for estimating forest overstory density. Forest Science 2:314–20. Littell, R. C., G. A. Milliken, W. W. Stroup, and R. D. Wolfinger. 1996. SAS System for Mixed Models. SAS Institute Inc., Cary, NC, USA. McCune, B. and J. B. Grace. 2002. Analysis of Ecological Communities. MjM Design Software, Gleneden Beach, CA, USA. McCune, B. and M. J. Mefford. 1999. PC-ORD. Multivariate Analysis of Ecological Data, Version 4. MjM Software Design, Gleneden Beach, OR, USA. Murkin, H. R. 1989. The basis for food chains in prairie wetlands. p. 316–38. In A. G. van der Valk (ed.) Northern Prairie Wetlands. Iowa State University Press, Ames, IA, USA. Murkin, H. R., P. G. Abbott, and J. A. Kadlec. 1983. A comparison of activity traps and sweep nets for sampling nektonic invertebrates in wetlands. Freshwater Invertebrate Biology 2:99–106. Niering, W. A. 1997. Wetlands. National Audubon Society Nature Guides. Alfred A. Knopf, Inc., Chanticleer Press, Inc., NY, USA. Oertli, B. 1993. Leaf litter processing and energy flow through macroinvertebrates in a woodland pond (Switzerland). Oecologia 96:466–77. Palik, B., D. P. Batzer, R. Buech, D. Nichols, K. Cease, L. Egeland, and D. E. Streblow. 2001. Seasonal pond characteristics across a chronosequence of adjacent forest ages in northern Minnesota, USA. Wetlands 21:532–42. Palik, B. J., R. Buech, and L. Egeland. 2003. Using an ecological land hierarchy to predict seasonal-wetland abundance in upland forests. Ecological Applications 13:1153–63. Rosenberg, D. M. and V. H. Resh. 1996. Use of aquatic insects in biomonitoring, p. 87–97. In R. W. Merritt and K. W. Cummins (eds.) Aquatic Insects of North America, third edition. Kendal/ Hunt Publishing Company, Dubuque, IA, USA. Ross, L. C. M. and H. R. Murkin. 1989. Invertebrates. p. 35–38. In E. J. Murkin and H. R. Murkin (eds.) Marsh Ecology Research Program Long-Term Monitoring Procedures Manual. Delta Waterfowl and Wetlands Research Station Technical Bulletin 2.
Miller et al., VARIATION IN WETLANDS: IMPLICATIONS FOR SAMPLING Schneider, D. W. and T. M. Frost. 1996. Habitat duration and community structure in temporary ponds. Journal of the North American Benthological Society 15:64–86. Stewart, R. E. and H. A. Kantrud. 1971. Classification of natural ponds and lakes in the glaciated prairie region. US Fish and Wildlife Service, Washington, DC, USA. Resources Publication 92. Tangen, B. A., M. G. Butler, and M. J. Ell. 2003. Weak correspondence between macroinvertebrate assemblages and land use in Prairie Pothole Region wetlands, USA. Wetlands 23:104–15. ter Braak, C. J. 1995. Ordination. p. 91–173. In R. H. Jongman, C. J. ter Braak, and O. F. van Tongeran (eds.) Data Analysis in Community and Landscape Ecology. Cambridge University Press, Cambridge, UK. ter Braak, C. J. and P. Smilauer. 1998. CANOCO Reference Manual and User’s Guide to Canoco for Windows: Software for Canonical Community Ordination (version 4). Microcomputer Power, Ithaca, NY, USA.
881
Turner, A. M. and J. C. Trexler. 1997. Sampling aquatic invertebrates from marshes: evaluating the options. Journal of the North American Benthological Society 16:694–709. Wellborn, G. A., D. K. Skelly, and E. E. Werner. 1996. Mechanisms creating community structure across a freshwater habitat gradient. Annual Review of Ecology and Systematics 27:337–63. Wiggins, G. B., R. J. Mackay, and I. M. Smith. 1980. Evolutionary and ecological strategies of animals in annual temporary pools. Archiv fu ¨ r Hydrobiologie Supplement 58:97–207. Williams, D. D. 1996. Environmental constraints in temporary fresh waters and their consequences for the insect fauna. Journal of the North American Benthological Society 15:634–50.
Manuscript received 6 April 2007; accepted 21 March 2008.
