Abstract Effects of Wildfire on Water Quality and Benthic Macroinvertebrate Communities of a Chihuahuan Desert Spring System

ABSTRACT

EFFECTS OF WILDFIRE ON WATER QUALITY AND BENTHIC MACROINVERTEBRATE COMMUNITIES OF A CHIHUAHUAN DESERT SPRING SYSTEM

by Tara Jo Haan

Wildfire disturbances affect resource availability and alter community composition in arid environments. Traditionally, fire effects on arid-land aquatic ecosystems are under-studied compared to terrestrial ecosystems. Chihuahuan Desert spring systems offer a unique opportunity to study such effects on macroinvertebrate community resistance and resilience. I took advantage of a rare opportunity to employ a BACI design to observe changes in water quality and macroinvertebrate communities to wildfire in a spring system on Bitter Lake National Wildlife Refuge, New Mexico. The results suggest significant water quality and species-specific response to wildfire. I observed an increase in an endangered snail, Juturnia kosteri, but there were no significant community-based changes. These results suggest that arid-land aquatic communities can be resistant to abiotic/biotic changes caused by wildland fire. With climate change predicted to increase the frequency and intensity of arid-land fires, aquatic communities may be more vulnerable to severe events in the future.

EFFECTS OF WILDFIRE ON WATER QUALITY AND BENTHIC MACROINVERTEBRATE COMMUNITIES OF A CHIHUAHUAN DESERT SPRING SYSTEM

A Thesis

Submitted to the Faculty of Miami University in partial fulfillment of the requirements for the degree of Master of Science Department of Zoology

by Tara Jo Haan Miami University Oxford, OH 2012

Advisor ______Dr. David J. Berg

Reader ______Dr. Craig Williamson

Reader ______Dr. Ann L. Rypstra

TABLE OF CONTENTS

List of tables iii

List of figures iv

List of appendices v

Acknowledgements vi

1. Introduction 1

2. Study area and Sandhill wildfire 5

3. Objective 7

4. Methods 9

4.1 Sample collection 9

4.2 Water quality analysis 9

4.3 Macroinvertebrate community analysis 10

5. Results 12

5.1 Water quality 12

5.2 Macroinvertebrate communities 13

6. Discussion 18

7. Conclusions, implications, and future directions 28

8. Bibliography 31

LIST OF TABLES

Table 1: Water depth and velocity at Bitter Creek and Sago Spring

Table 2: Summary statistics for water quality variables

Table 3: Results of ARIMA analyses on water quality variables

Table 4: Average within-season Sorensen’s Similarity Index at Bitter Creek and Sago Spring

Table 5: Average among-season Sorensen’s Similarity Index at Bitter Creek and Sago Spring

iii

LIST OF FIGURES

Figure 1: The Chihuahuan Desert of North America

Figure 2. Bitter Creek and Sago Spring study areas within Bitter Lake National Wildlife Refuge

Figure 3. Map of Sandhill fire perimeter

Figure 4. Seasonal time series of temperature at Bitter Creek

Figure 5. Seasonal time series of dissolved oxygen at Bitter Creek

Figure 6. Seasonal time series of pH at Bitter Creek

Figure 7. Seasonal time series of salinity at Bitter Creek

Figure 8. Seasonal time series of specific conductance at Bitter Creek

Figure 9. Seasonal time series of total dissolved solids at Bitter Creek

Figure 10. Taxa richness and Simpson’s Index of Diversity at Bitter Creek

Figure 11. Taxa richness and Simpson’s Index of Diversity at Sago Spring

Figure 12. Total macroinvertebrate density at Bitter Creek and Sago Spring

Figure 13. Density of three endangered macroinvertebrates at Bitter Creek

Figure 14. Density of three endangered macroinvertebrates at Sago Spring

Figure 15. Density of all functional feeding groups at Bitter Creek and Sago Spring

Figure 16. Density of functional feeding groups with scrapers removed at Bitter Creek and Sago

Spring

Figure 17. Density and richness of insects at Bitter Creek and Sago Spring

Figure 18. Density and richness of detritivores at Bitter Creek and Sago Spring

Figure 19. Sorensen’s Similarity Index between Bitter Creek and Sago Spring

LIST OF APPENDICES

Appendix 1. Seasonal effects of hydrochemical variables.

Appendix 2. Raw macroinvertebrate data from Bitter Creek pre- and post-fire

Appendix 3. Raw macroinvertebrate data from Sago Spring pre- and post-fire

Appendix 4. α-, β-, and γ-diversity at Bitter Creek and Sago Spring

ACKNOWLEDGEMENTS

I would like to extend my sincere gratitude to my advisor, Dr. Dave Berg, for his constant patience and guidance throughout my time at Miami University. I would also like to thank my committee members, Dr. Ann Rypstra and Dr. Craig Williamson, for their suggestions, comments, and support. In addition, I would like to thank Dr. Jon Patten for his help and expertise in time series analysis and Mike Hughes of the Miami Statistical Consulting Center for assistance in ANCOVA models. Special thanks to Brian Lang of the New Mexico Department of Game and Fish for the collection of samples and his help throughout the course of my master’s research, as well as his timely responses to my many emails. Numerous undergraduate students aided in the sorting and identifications of macroinvertebrates, so I thank Keara Stanislawczyk, John Abeln, and Robert Firor for their assistance in that process. Funding for this research was provided by the National Science Foundation grant DEB-0717064 and the New Mexico Department of Game and Fish. Finally, I would like to thank my family for their endless support and encouragement throughout my education.

1. INTRODUCTION

Disturbances are discrete events that disrupt ecosystem structure, affect availability of resources and/or substratum, and change the physical landscape of an environment (Resh et al., 1988, Gresswell, 1999, Minshall, 2003). However, disturbances are also important in maintaining biodiversity of habitats and evidence suggests that moderate disturbance intensities and frequencies often allow for maximum species richness (Hobbs and Huenneke, 1992). Disturbances may be natural, such as droughts or floods, or they may be human-caused, such as the clearing of forests or the introduction of exotic species. Wildfires are an agent of disturbance which have been shown to have powerful effects in numerous ecosystems. For many years, wildfire effects in terrestrial environments were much more studied than in aquatic ecosystems (Minshall et al., 1989, Gresswell, 1999, Minshall, 2003), and wildfire studies in arid-land aquatic environments are particularly lacking. In addition, much of the previous research focusing on fire in aquatic systems emphasized alterations in water yield and quality, rather than effects on biota (Gresswell, 1999). Wildfires are characterized by their intensity, magnitude, frequency, scale, and pattern, and are heavily influenced by an areas’ local climate, vegetation and fuels, geomorphology, hydrology, and other environmental attributes, making the specific effects of a wildfire event widely variable and difficult to predict (Dwire and Kauffman, 2003). The effects of a wildfire on a stream ecosystem can be broken down into three response periods: (1) short-term (<1 year) changes result directly from changes in water chemistry and food quality; (2) mid-term (1-10 years) and (3) long-term (>10 years) responses refer to changes resulting from the removal and later successional replacement of the terrestrial cover in a stream catchment (Minshall et al., 1997). The most powerful short-term effects likely result from increased channel erosion leading to increased overland runoff and elevated sediment loads in streams, especially following significant precipitation events. Mid- and long-term responses, including return to pre-fire conditions, closely correspond to vegetative regrowth (Minshall et al., 1995, Ryan et al., 2011). Ultimately, however, the overall responses heavily depend on both the size and intensity of the fire, as well as the physical, chemical, and biological traits of each site (Gresswell, 1999). In addition, there is generally a gradual decrease in harmful fire effects with increasing stream size (Minshall, 2003).

1 The consequences of wildfire on freshwater macroinvertebrates are due to both direct and indirect effects of fire. Direct effects include increased temperature, nutrients, charcoal, and ash. These are associated with the time from the fire to the first major precipitation and runoff event. Indirect effects of fire refer to the channel alteration, increased erosion, and sediment transport and deposition that commence with the first runoff after the fire (Minshall, 2003). The direct effects of fire on macroinvertebrates are often minor, but communities are less resistant and resilient to the indirect effects, specifically intense sedimentation after a fire (Vieira et al., 2004). Community resistance refers to the magnitude of change from the pre-disturbed state and depends on the ability of organisms to avoid mortality or displacement after the fire. Community resilience refers to the rate of recovery to the pre-disturbed state and depends on the recolonization abilities of species, as well as the occurrence of repeated post-fire disturbances such as flooding, which may reset the recovery trajectory (Vieira et al., 2004). Macroinvertebrates have often been considered good indicators of stream quality. While there is not sufficient evidence to consider any one particular macroinvertebrate species as a gauge of water quality, the community as a whole has long been used as an indicator of pollution (Goodnight, 1973). Many macroinvertebrate species have life cycles greater than a year, allowing for the integration of longer-term pollution effects, but they may also respond rapidly to habitat stressors and alterations in water chemistry (Klemm et al., 2001). Furthermore, macroinvertebrates are relatively sedentary, allowing the detection of localized stressors, they are easy to collect, and communities are often heterogeneous, which increases the chances that at least some taxa will respond to disturbances. For these reasons, they are frequently preferred over fish communities as water quality monitors (Berkman et al., 1986). However, the laboratory work needed to complete the identifications of invertebrate taxa can be extremely difficult and requires a significant time commitment. Isolated and small wetlands offer a unique opportunity for examining the effects of wildfire. Because colonization of macroinvertebrates is limited (flying insects with an aquatic larval stage are the most likely early colonizers), most taxa occurring shortly after the fire likely survived in situ (Munro et al., 2009). Reports of recovery times of macroinvertebrate communities following wildfire vary widely in the scientific literature. Earl and Blinn (2003) found that macroinvertebrate densities of the Gila River drainage of south-western New Mexico returned to pre-fire conditions within 1 year following wildfire, but Minshall et al. (1995) noted

2 that macroinvertebrate communities from burned streams in Yellowstone National Park did not fully recover during their 5 year monitoring period. Furthermore, Minshall (2003) suggested that variations in invertebrate communities may not become stable until 7-10 years after the fire, and may even continue to show within-year variations from unburned reference streams for greater than 15 years post-fire. The suite of unique characteristics of every site can make recovery rates extremely challenging to forecast. The recovery time of macroinvertebrate communities to wildfire will depend heavily on dispersal abilities of taxa to recolonize a disturbed site. Williams and Hynes (1974) identified four main pathways that stream benthic invertebrates recolonize disturbed habitats: aerial travel, downstream drift, upstream movements, and vertical transport from deep substrates. In permanent streams and rivers, downstream drift is typically the most important of these pathways, while intermittent streams are more often recolonized by aerial dispersers. A useful method for measuring the magnitude by which a stressor has changed an environment is the Before-After-Control-Impact (BACI) design, in which data exist before and after a major event, as well as from both a reference and impacted site. For the best recognition of ecosystem distress, it is necessary to compare variables with a normal state (Rapport et al., 1985). In a BACI study, the pre-disturbance data of the disturbed site and the pre- and post- disturbance data of the control site can be used to determine what constitutes a “normal” state of the environment and are used to define the impact of a stressor. Because natural disturbances are often not easily predictable, true BACI studies in ecology are not very common due to the lack of pre-disturbance data. There are only a small number of BACI studies relating to fire and aquatic macroinvertebrate communities, and those generally have to do with prescribed burns, which give a clear opportunity for pre-disturbance data collections. For example, Bêche et al. (2005) employed a BACI study design to assess the effects of a prescribed fire on a stream and riparian zone located in the Sierra Nevada mountain range of California. Their results suggest that the fire had either no or short-lasting effects on the stream macroinvertebrates and water chemistry, likely due to the fact that it was a low- to moderate-intensity burn and only a small portion (<20%) of the watershed was affected. These are characteristics typical of prescribed burns. Similarly, Arkle and Pilliod (2010) found no detectable changes in macroinvertebrate communities or stream conditions following a prescribed fire in Payette National Forest, Idaho. They compared their findings with similar data taken after a nearby natural wildfire and found

3 that the prescribed fire did not mimic the ecological effects of the wildfire, either in-stream or in the surrounding riparian zone. In one natural wildfire BACI study, Vieira et al. (2004) examined the effects of repeated flooding on stream insect community resistance and resilience in a New Mexico forest following wildfire. They found that insect community richness and density were reduced to near zero after the fire, with rapid recolonization by a few species which are generalist feeders and have strong larval-dispersing abilities. They also showed that repeated hydrologic disturbances greatly altered community composition and richness, and those differences persisted even 4 years after the fire when flooding had diminished. Finally, they noted that responses heavily depend on specific species traits, hydrologic conditions following fire, and barriers to recolonization of the disturbed site. Springs located within Bitter Lake National Wildlife Refuge, Chaves County, New Mexico are the focus of this project. The present study represents an unreplicated “natural” experiment and presented a unique and rare opportunity to employ a BACI approach to observe changes in abiotic and biotic conditions of isolated arid-land springs. I was able to examine the effects of a wildfire disturbance by using data from both an affected and reference site, as well as by using pre- and post-disturbance data.

2. STUDY AREA AND SANDHILL WILDFIRE

The area of the Chihuahuan Desert from southern New Mexico to near Mexico City was once covered by a shallow ocean which extended from the Gulf of Mexico. This ocean, the Permian Sea, receded during the Late Cretaceous period (~66 mya) and drying of the region led to the establishment of a desert landscape containing isolated aquatic environments, many of which have high numbers of endemic species (Holsinger, 1972). Bitter Lake National Wildlife Refuge is located near Roswell, NM (Chaves County) in the Pecos River watershed of the Chihuahuan Desert (Fig. 1) and consists of about 9,900 Ha containing an assortment of water habitats, such as seeps, free-flowing springs, sink holes, marshes, and the refuge’s namesake, Bitter Lake. The refuge was established in 1937 to provide protection for migratory birds, but it also plays an essential role in the conservation of species which use and/or inhabit the many aquatic oases surrounded by an otherwise harsh desert landscape. The two spring systems of Bitter Creek and Sago Spring, located within the refuge, are the focus of this project (Fig. 2). Bitter Creek is 1.8 km long from spring heads to its entry into Bitter Lake and has substrates which consist of deep aqueous organic silts and clays. Nearby, the smaller Sago Spring measures 0.3 km in length from spring head to Bitter Lake, nearly half of which is subterranean. The substrates of Sago Spring differ from those of Bitter Creek and consist of gypsum sinkholes and white sand. The overland distance between Bitter Creek and Sago Spring is less than 0.5 km and the river distance between the two sites is just over 2 km. Groundwater sources for Bitter Creek and Sago Spring seem to be distinct and independent (Lang, 1998). Several species of shrubs and grasses comprise the riparian vegetation of both systems. Around 1100 hours on March 5, 2000 the Sandhill wildfire was ignited about 0.5 km west of the perimeter of the refuge and spread quickly onto refuge land behind wind gusts in excess of 50 mph (Lang, 2005). A total of about 425 Ha of refuge land was consumed by the fire, which was low- to- moderate intensity and burned in a mosaic pattern throughout about 75% of the burn area (Gavin, 2000, Outcalt and Kennard, 2008). The fire impacted upland scrub, consisting primarily of fourwing salt bush, and alkali sacaton-salt grass lowlands, as well as the entire riparian corridor of Bitter Creek, which was most heavily affected in the middle reach near the flume and the upper reach near headwater sources (Fig. 3) (Lang, 2005). Probably the most

5 evident biological effect of the fire was the formation of dense, homogenous stands of the invasive common reed, Phragmites australis, along the entire reach of Bitter Creek, which was found in only sparse patches prior to the fire (Lang, 2005). Bitter Creek and the unaffected Sago Spring provide habitat to several federally endangered species, including three macroinvertebrate taxa: the amphipod Gammarus desperatus (Amphipoda: Gammaridae) and the snails Pyrgulopsis roswellensis and Juturnia kosteri (Gastropoda: Hydrobiidae).

3. OBJECTIVE

The purpose of this project was to examine the effects of the Sandhill wildfire on both water quality and the macroinvertebrate communities of Bitter Creek. Sago Spring differs from Bitter Creek in several ways, such as in size and substrate composition, but because it was not affected by the fire, it served as a reference site for this study. I looked at changes in water quality associated with the fire by examining pre- and post-fire datasets and took advantage of the rare opportunity to use a BACI design associated with a natural ecological disturbance to investigate macroinvertebrate community response to wildfire, because data exist from both before and after the wildfire, as well as from both an impacted and reference site. Hydrochemical variables and biological indices were used to assess effects of the disturbance event. Indices such as species density, richness, diversity, and similarity all contribute different important information to the overall evaluation of the makeup of a community (Novak and Bode, 1992) and were used in this study to assess the macroinvertebrate communities of the two sites. In addition, characteristics such as dispersal ability and functional feeding groups of taxa were examined through time to assess shifts associated with specific species traits and ecosystem roles, rather than taxonomy. Although the spatial and temporal scale of wildfire effects on this particular spring environment are not easy to forecast, several a priori predictions were made:

Prediction 1: Species richness and diversity in Bitter Creek will decrease following the fire and community composition will shift toward an increase in dominance of species which are disturbance-adapted by having generalist feeding strategies (gatherers, filterers), and/or rapid colonizing abilities (aerial insects).

Prediction 2: It is widely known that wildfire processes lead to increased water temperatures (Ice et al, 2004). Direct heating from the fire, increased light availability due to removal of riparian vegetation, and increased sediment loads due to increased erosion and runoff will lead to higher water temperatures at Bitter Creek. In addition, decreased dissolved oxygen concentrations at Bitter Creek will occur after the fire due to reduced solubility of DO with

7 increased temperature and increased respiration due to the decomposition of burned vegetation in the stream.

Prediction 3: The endangered snails J. kosteri and P. roswellensis are scrapers and the endangered amphipod G. desperatus is a shredder, which tend to be specialist feeding groups. In addition, none of these species have strong dispersal abilities. For these reasons, I expect that all of these endangered species will be negatively affected by the fire and will decrease in density.

Prediction 4: There will be a pulse of increased nutrients following the wildfire due to increased runoff, but due to a change in input of organic matter from unburned detritus to burned materials and ash, the quality of detritus as a food source will decrease. I expect that this will cause an initial decrease in detritivore species richness and density at Bitter Creek.

Prediction 5: The macroinvertebrate communities of Bitter Creek and Sago Spring contain the same dominant species (J. kosteri, P. roswellensis, and an oligochaete worm), but exhibit some major differences, such as many more taxa being found in Bitter Creek than in Sago Spring. Although pre-fire similarity between the two sites is not extremely high, I predict that the similarity between macroinvertebrate communities of Bitter Creek and Sago Spring will decrease following the fire.

Prediction 6: The water quality and macroinvertebrate communities of Bitter Creek may return to pre-fire conditions or establish a new equilibrium over time, but I do not expect that this will be seen in the short 3-year post-fire window of observation of this study.

4. METHODS

4.1 Sample collection From Fall 1995 to Summer 1998 and from Spring 2000 to Winter 2003, six water quality variables (temperature, dissolved oxygen, salinity, pH, specific conductance, and total dissolved solids) were measured hourly at Bitter Creek using a Hydrolab RecorderTM for a 24-30 day logging period each seasonal quarter (Fall [November], Winter [February], Spring [May], and Summer [August]) by New Mexico Department of Game and Fish (NMDGF). The winter 2000 logging ran from February 8, 2000 to March 8, 2000, spanning the period prior to, during, and shortly after the fire. Due to equipment failure, no data were available for the summer 1998 and winter 2003 logging periods. Benthic macroinvertebrates were collected by NMDGF, which conducted monthly quantitative sub-sampling at multiple sites along both Bitter Creek and Sago Spring during the same time frames in which water quality data were acquired (1995-1998, 2000-2003). During the majority of sampling periods, 3-4 replicates were taken from each of 6 sampling locations along Bitter Creek and from each of 3 sampling locations along Sago Spring. Sampling was done using benthic grabs consisting of a customized stainless steel basket (5.08w x 7.62l x 2.54h cm) lined with a fine screen (mesh: 0.69 mm; gauge: 0.58 mm) and a 6 mm groove welded on three sides. This sampling was accomplished by inverting the sampler into the substrate, grooved-side down, and then sliding a stainless steel top into the groove. Organic material and substrate were washed from the basket and benthos were preserved in 70% ethanol. In addition, water depth and velocity were measured at each sampling location using a metric staff rod and Marsh-McBirney® flow meter. All water quality data and macroinvertebrate samples were transferred to Miami University for analysis.

4.2 Water quality analysis Summary statistics for all water quality variables were calculated. Time series analysis of water quality data from all years was performed using autoregressive integrated moving average (ARIMA) models in PROC ARIMA of SAS® version 9.2. As water quality variables were measured hourly at Bitter Creek, the ARIMA time series analysis was necessary because

9 values may be affected by the season during which they were measured, as well as the hour of day. My analysis factored out these natural fluctuations so that the effect of the wildfire could be isolated.

4.3 Macroinvertebrate community analysis Field methods for macroinvertebrate collection did not become consistent until July 1996, so only samples collected on or after that date were used in this study. A randomly reduced number of replicates was chosen from each season of each year, when available, in an attempt to mimic the timing of the collection of water quality data. A dissecting microscope was used to sort the macroinvertebrates to lowest convenient taxa and then assign them to morphospecies. For each sample, relative abundance of taxa, taxon richness, density, and Simpson’s Index of Diversity were calculated. In addition, functional feeding groups were assigned according to classification by Merritt et al. (2008) in order to assess changes associated with trophic dynamics rather than taxonomy (Cummins et al., 2005, Tomanova et al., 2006) Sorensen’ Similarity Index was calculated to compare communities within and between seasons and sites pre- and post-fire. In order to examine the effect of the fire on taxa exhibiting certain traits or fulfilling certain ecosystem roles, I also calculated the density and richness of insects with aerial adult life stages and the density and richness of macroinvertebrates which are detritivores. Pearson correlation coefficients were obtained between water quality variables and multiple macroinvertebrate community indices in order to determine which hydrochemical variables had the strongest relationships with community composition. Diversity at different temporal scales was assessed by calculating α-, β-, and γ-diversity at both Bitter Creek and Sago Spring before and after the fire. For time series analysis, it is necessary to establish a sufficiently long time series with many data points. I was able to accomplish this with the water quality data because the automatic and hourly collections of the Hydrolab RecorderTM allowed for evenly spaced time periods and many data points. For the macroinvertebrate data of this study, this type of dataset was unfortunately not possible as I had limited months of data and broader monthly collections were conducted, rather than hourly sampling. For these reasons, there were difficulties in applying time series techniques to analyze the communities. Therefore, analyses of covariance (ANCOVA) were performed in PROC GLM of SAS® version 9.2 to formally test the effect of

10 the fire on the biotic indices. In this model with a BACI approach, the factors of interest were phase (pre- or post-fire) and site (control or burn), with time being used as the covariate (measured as month since the start of observed data). This method allowed for the comparison of linear trajectories of indices of interest before and after the fire and between sites. The difference in the trajectories before the fire versus after the fire constituted the formal test of the effect of the disturbance.

5. RESULTS

5.1 Water quality Water depth and velocity ranges for both Bitter Creek and Sago Spring are presented in Table 1. Water depth of samples taken before and after the fire at both sites was always around or below 0.5 m and velocity was always below 0.4 m/sec. Summary statistics for all water quality variables measured at Bitter Creek are presented in Table 2. Time series graphs of all six variables were created and are shown in Figs. 4-9. Not surprisingly, water temperatures were warmest in spring and summer both pre- and post-fire, with the highest values being observed during the spring. However, post-fire ranges of water temperature in spring and summer were narrower than were seen pre-fire (Fig. 4). Pre-fire dissolved oxygen concentrations reached their highest levels during spring and fall (greater than 20 mg/L), and were occasionally observed at levels of super-saturation (beyond recordable range). Post-fire dissolved oxygen concentrations were notably depressed during spring, summer, and fall, as compared to pre-fire ranges (Fig. 5). For the most part, the pH of Bitter Creek from all seasons and years was around neutral to slightly basic. The post-fire winter pH values were noticeably elevated compared to pre-fire observations (Fig. 6). During all seasons from all years, salinity ranged from about 3-10 ppt, considered within the mesohaline range of brackish water (Lang, 1998). As salinity, specific conductance, and total dissolved solids (TDS) are all related, it should not be unexpected that these three graphs exhibit many of the same shapes, properties, and patterns. Post-fire fall salinity, specific conductance, and TDS values were visibly higher than pre-fire ranges (Figs. 7- 9). The results of the ARIMA analyses describe (1) significant (p<0.05) seasonal and fire effects for water temperature, specific conductance, salinity, and pH, (2) a significant seasonal effect, but no fire effect for total dissolved solids, and (3) no seasonal effect, but a significant fire effect for dissolved oxygen. The direction and magnitude of fire effects on these variables are given in Table 3 and the seasonal effects are shown in Appendix 1. Water temperature (+1.922C) and pH (+0.248) increased significantly after the fire, while dissolved oxygen (-1.492 mg/L), salinity (-0.909 ppt), and specific conductance (-656.829 µS/cm) exhibited significant decreases.

5.2 Macroinvertebrate communities The time periods analyzed, number of replicates sorted, and raw macroinvertebrate counts from Bitter Creek and Sago Spring are given in Appendices 2 and 3, respectively. Results from both pre- and post-fire samples indicate that the endangered gastropods J. kosteri and P. roswellensis, as well as an oligochaete worm are numerically dominant species in both Bitter Creek and Sago Spring. Taxa richness and Simpson’s Index of Diversity for Bitter Creek and Sago Spring samples are shown in Figs. 10 and 11, respectively. Simpson’s Index of Diversity ranges from zero to one with zero representing a homogeneous community and one representing a completely heterogeneous community. Both before and after the fire, average taxon richness at Bitter Creek ranged between about 4 and 12, while taxon richness at Sago Spring tended to be lower, ranging between about 3 and 8 both pre- and post-fire. On average, both taxon richness and Simpson’s Index of Diversity values were lower at Bitter Creek post-fire compared to pre- fire, although richness reached its’ highest level in November 2000, 8 months following the Sandhill wildfire. Average pre-fire diversity at Bitter Creek ranged from about 0.2 to 0.8, but was observed in a narrower range post-fire, never reaching above 0.6. On average at Sago Spring, taxon richness was marginally higher during the post-fire time frame, while diversity was slightly lower. During some sampling periods, standard error bars are fairly large, suggesting that the variability of replicates within each sample period can be quite large. No clear seasonality in richness or diversity was observed at either site. Total macroinvertebrate density from Bitter Creek and Sago Spring is shown in Fig. 12 and the densities of the three endangered macroinvertebrates J. kosteri, P. roswellensis, and G. desperatus are displayed in Figs. 13 and 14. Total density at Bitter Creek, which was driven heavily by the abundant J. kosteri, averaged nearly 33,000 individuals/m2 pre-fire and more than doubled post-fire, averaging about 68,000 individuals/m2. Total density at Sago Spring was driven primarily by the abundant P. roswellensis. Total average pre-fire density at Sago Spring was a little over 27,000 individuals/m2 and average post-fire density was nearly 65,000 individuals/m2. At Bitter Creek, J. kosteri was the most abundant of the three endangered species pre- and post-fire. Densities of J. kosteri were higher post-fire, reaching over 100,000 individuals/m2 in summer and fall 2000. Pyrgulopsis roswellensis and G. desperatus, on the other hand, were found at lower densities after the fire, with G. desperatus being completely absent from all samples after summer 2001. At Sago Spring, P. roswellensis was the most

13 abundant of the endangered macroinvertebrates both before and after the fire, being found at higher densities post-fire and reaching values over 100,000 individuals/m2 during the summer of 2002. Juturnia kosteri was found at higher densities after the fire, peaking in spring 2000 and summer 2001 and G. desperatus, the least abundant of the three, was also found at higher densities post-fire compared to pre-fire. All macroinvertebrate taxa were categorized into one of five functional feeding groups: scraper, collector-gatherer, collector-filterer, shredder, or predator. Scrapers graze directly on periphyton attached to substrates, collector-gatherers gather fine particulate organic matter from the surface of substrates, collector-filterers feed on organic materials suspended in the water column, shredders tear and ingest pieces of coarse particulate organic matter, and predators consume living prey organisms (Cummins and Klug, 1979, Wilfrid Laurier University, 2000). Shredders and scrapers tend to be specialist species and are sensitive to disturbance, while generalist feeders, such as collector-gatherers and collector-filterers, tend to be more tolerant to fluctuations in food availability (Rawer-Jost et al, 2000). The densities of these feeding groups at Bitter Creek and Sago Spring are shown in Fig 15. At both sites, the scraper feeding group was by far the most dominant, driven heavily by J. kosteri at Bitter Creek and P. roswellensis at Sago Spring. At Bitter Creek, scrapers were found in higher densities following the fire, peaking at densities in excess of 100,000 individuals/m2 in summer and fall 2000. At Sago Spring, scrapers were also found in higher densities post-fire, and were seen at densities greater than 100,000 individuals/m2 during summer 2002. Because the scraper feeding group was found at such high densities at both sites, this group was removed in order to better visualize the changes in the other four feeding groups, which is shown in Fig. 16. At Bitter Creek, collector-gatherers and collector-filterers were both found at higher densities after the fire. Collector-gatherers consisted primarily of chironomid midges and oligochaetes. Shredders and predators were found at the lowest densities throughout all years at Bitter Creek, although predators increased slightly in density after the fire, while shredders were found at lower densities following the fire and were completely absent from 42% of post-fire samples. At Sago Spring, densities of collector- gatherers and predators decreased post-fire, while collector-filterers and shredders increased. At this site, chironomids and oligochaetes also accounted for the majority of the collector-gatherer feeding group.

Densities and richness of insect taxa with aerial adult life stages are shown in Fig. 17 for both Bitter Creek and Sago Spring. At both sites, chironomid midges were the most abundant insects before and after the fire, followed by alderflies. At Bitter Creek, dragonfly, damselfly, and caddisfly larvae constituted a small proportion of insects, while none of these were found at Sago Spring. Average insect density and richness were higher at Bitter Creek than at Sago Spring throughout all years of observation. Density and richness of insects at Bitter Creek were slightly lower after the fire, while both were slightly higher after the fire at Sago Spring. No clear pattern or seasonality in insect richness or density was observed at either site. Furthermore, there were no differences between “strong” flyers (dragonflies, damselflies), and intermediate-to- “weak” flyers (midges, alderflies, caddisflies). Density and richness of detritivorous macroinvertebrates from Bitter Creek and Sago Spring are represented in Fig. 18. The classification of detritivores consists of the collection of the shredder, collector-gatherer, and collector-filterer FFGs. Both detritivore density and richness tended to be higher overall at Bitter Creek than at Sago Spring, both before and after the fire. Detritivore density at Bitter Creek was slightly higher post-fire, reaching the highest level in fall 2000, while detritivore richness declined only slightly after the fire. At Sago Spring, both density and richness of detritivores increased marginally post-fire. No obvious seasonality in detritivore density or richness was detected at either site. Sorensen’s Similarity Indices were calculated in order to compare the macroinvertebrate communities within and between seasons at each site. Sorensen’s Similarity Index ranges from zero to one with a value of zero indicating that the two communities do not share any species and a value of one indicating that the two sites have the same composition. Average within-season similarity at both sites is presented in Table 4, giving values for pre-fire, post-fire, and comparing seasonal pre- vs. post-fire similarity. While some noticeable trends were observed, for example noticeably higher winter similarity at Bitter Creek post-fire than was seen pre-fire, no significant differences (p>0.05) were detected within either of the two sites. When comparing the within-season similarities of Bitter Creek to those of Sago Spring, pre- vs. post-fire similarity was higher at Sago Spring in all seasons. In addition to within-season similarity, Sorensen’s Similarity Indices were also calculated between seasons at both sites (Table 5). Summer and fall macroinvertebrate communities were most similar at Bitter Creek pre-fire and spring and summer communities were most similar post-fire. At Sago Spring, fall and winter communities

15 were most similar pre-fire and summer and fall were most similar post-fire. At both sites, only summer vs. winter similarity values were significantly different (p<0.05) pre-fire vs. post-fire, with communities being more similar post-fire in both cases. Furthermore, similarity between Bitter Creek and Sago Spring macroinvertebrate communities was calculated for each sampling period and these values are graphically displayed in Fig. 19. Before the fire, these similarity values ranged between 0.22 and 0.75 and ranged between 0.28 and 0.53 post-fire. The overall average similarity between Bitter Creek and Sago Spring macroinvertebrate communities pre-fire was 0.44, and the overall post-fire average was 0.39. Values of α-, β-, and γ-diversity for both sites are presented in Appendix 3. Although there are differing views in the scientific community as to whether multiplicative or additive partitioning of species diversity is more useful (Veech et al., 2002), here we present γ-diversity as the sum of α- and β-diversity because this approach allows for both α-diversity and β-diversity to be presented as averages, where α is average within-sample diversity and β is average diversity not found in a single sample. This method makes α and β easier to directly compare because β-diversity is not reduced to describing changes only along an environmental gradient (Veech et al., 2002). In this study, rather than spatial diversity, α-, β-, and γ-diversity refer to temporal diversity, where γ-diversity is a total average community representation over an entire pre-fire or post-fire time frame, α-diversity refers to the average diversity in a single sampling period, and β-diversity represents the average differentiation between time periods. The results of this study show that γ-diversity was always higher at Bitter Creek than at Sago Spring and β- diversity was high relative to α-diversity in all seasons, when looking at only insects or detritivores, as well as when looking at overall pre- and post-fire diversity. Average water temperature was the most significantly correlated to macroinvertebrate community indices of all water quality variables, with temperature being significantly negatively correlated with diversity (n = 6; r = -0.83; p<0.05) and significantly positively correlated with total macroinvertebrate density (n = 6; r = 0.94, p<0.01) and scraper density (r=0.93, p<0.01). While dissolved oxygen showed fairly strong positive correlations with taxon richness (n = 6; r = 0.50; p>0.1) and diversity (n = 6; r = 0.62; p>0.1), these correlations were not significant. No other water quality variable showed strong correlations to any community indices. Although some clear trends were seen in many of the examined indices, the results of the ANCOVA analyses suggest that the observed increase in density of the endangered J. kosteri

16 was the only significant (p<0.05) effect of the fire. None of the community, trophic, or other taxon-specific indices showed significant changes due to the wildfire.

6. DISCUSSION

My results show that the Sandhill wildfire had significant effects on five of the six measured hydrochemical variables and led to a significant increase in the density of the endangered snail J. kosteri at Bitter Creek. However, no significant community-based changes were observed. Furthermore, the invasive P. australis became the dominant vegetation following the fire at Bitter Creek, which was previously covered by grass and shrubs. Here, I discuss the specific measured variables and indices as they relate to each of the predictions presented in the introduction.

There was no significant effect of the fire on taxon richness or Simpson’s Index of Diversity at Bitter Creek. Changes in total macroinvertebrate density were also not significant, even though it more than doubled at both sites. These results suggest high resistance and resilience of these communities to this particular disturbance event. While the severity of the disturbance, the characteristics of the affected species, and the type of vegetation present greatly influence the recovery of macroinvertebrates (Beganyi and Batzer, 2011), studies of invertebrate communities in all types of ecosystems have shown high resistance and resilience to different forms of disturbance. A flash flood in Sycamore Creek, Arizona (USA) reduced algae and invertebrate standing crop by 98%, but both had recovered within 2-3 weeks, with the macroinvertebrate communities being recolonized largely by the immigration of aerial adult mayflies and dipterans (Fisher et al, 1982). In Port Phillip Bay, Australia, most species of scallops decreased in abundance by 20-30 % following dredging, but had returned to pre- dredging levels following their next recruitment, within six months of the dredging disturbance (Currie and Parry, 1996). Furthermore, total richness, evenness, and diversity of ground- dwelling invertebrates in the Canary Islands showed no differences between burned and unburned plots in a prescribed fire study (García-Domínguez et al, 2010). Although I

18 hypothesized lower taxon richness and diversity in this system, due to the short duration and low- to-moderate severity of the Sandhill fire, it is not surprising that the macroinvertebrate community as a whole exhibited high resistance and resilience to the disturbance. The collector-gatherer and collector-filterer FFGs are generalists and were predicted to be the most abundant after the fire, while scrapers and shredders are more likely to be negatively affected by disturbance, but no significant effect was observed in any single FFG of this study. The use of FFGs has been criticized by some because macroinvertebrates are often opportunistic and have flexible feeding strategies (Palmer and O’Keeffe, 1993, MacNeil et al, 1997), but the use of these trophic guilds in addition to analyses based on taxonomy can improve the overall understanding of community response to disturbance. Macroinvertebrate community structure can exhibit high variability over a short time period, and using finer taxonomic resolution or functional groups, in addition to more inclusive variables like community richness or diversity, increases the likelihood of detecting these changes across a shorter temporal period (Kosnicki and Sites, 2011). Properties such as sediment loads and temperature have been shown to affect FFGs in some environments (Friberg et al., 2009). For instance, increased sedimentation can cause collector-filterers to lose their attachment to the substrate and can clog filtering apparatuses, while collector-gatherers, which often burrow, may not be negatively affected by sediment deposition (Berkman et al., 1986). In addition, increased turbidity due to sedimentation may hinder the feeding of visual feeders, such as predators. However, Hawkins et al. (1982) found that feeding guilds were most commonly affected by variations in food quality, rather than by physical or chemical characteristics of a habitat. In the present study, scrapers were the dominant FFG at Bitter Creek and Sago Spring, both before and after the fire. At Bitter Creek, scrapers were followed by collector-gatherers and collector-filterers, with predators and shredders being found in low densities in all years. Nicola et al. (2010) found similar compositions of FFGs in headwater streams of eastern Spain, with scrapers being by far the most abundant group; collector-filterers, shredders, collector-gatherers, and predators composed a small proportion of overall abundance. The density of FFGs also varied widely among sampled localities. The biomass of the scraper FFGs was usually higher in burned streams of Idaho, USA as well (Minshall et al., 2001). Although some trends were observed in the plotted time series, such as scraper density peaking after the fire, there was no significant fire effect on any single FFG in this study, suggesting that the trophic relationships

19 between these groups and with food sources in the spring were not significantly altered by the fire. Insect larvae were of interest in this study because their aerial adult stage makes them likely candidates to colonize Bitter Creek shortly after the fire, but no significant effect was seen on insect richness or density in this study. Disturbances in isolated arid-land springs are likely to eliminate poor dispersers, while vagile taxa may travel from spring to spring and have populations connected by dispersal (Myers et al., 2001). Caddisflies, mayflies, and small dipterans are generally capable of flight across only several kilometers, buy may travel further in strong winds (Gray and Fisher, 1981) Adult Chironomidae have intermediate flying abilities in that they are not necessarily limited to a particular site, but lack the ability to travel far from their larval habitats (Delettre et al., 1992). Because Bitter Creek and Sago Spring are separated by a small distance, it is assumed that insect transport is common between the two sites and that insects would be the most likely taxa to recolonize the disturbed site (Gray and Fisher, 1981). In severely burned streams in Washington, USA following a 2003 wildfire, macroinvertebrate communities tended to be dominated by chironomid midges (Mellon et al., 2008). Chironomids also increased in density in burned Idaho, USA streams (Minshall et al., 2001). However, average insect density and richness at Bitter Creek, which consisted mostly of chironomids, were both slightly lower after the fire, and no significant effect was seen. There were also no major indications of seasonal patterns in insect richness or density. Because of the climate in American southwest deserts, reproduction of many organisms in stream habitats is continuous, and short life cycles are common, so abundances in any particular insect group cannot generally be explained by synchronous emergence (Grimm and Fisher, 1989). Additionally, there was no difference in abundances of strong versus weak flyers in Bitter Creek after the fire. The lack of clear patterns in insect movement in isolated spring environments is likely due to the small size and remoteness of macroinvertebrate colonist groups (Stanley et al., 1994). The first major precipitation and runoff events after the fire are often considered the most likely to result in further harmful effects on the biotic communities (Minshall, 2003) and rain in the wet season often occurs in the form of locally-intense thunderstorms resulting in flash floods, which can scour substrates and eliminate benthic taxa (Gray and Fisher, 1981). The months receiving the most precipitation in the year following the fire were June, August, and October 2000. These precipitation events do not appear to coincide with decreases in species richness at

Bitter Creek, but Simpson’s Index of Diversity did decline slightly throughout that entire time period. However, these values of diversity were no lower than those occurring in the years prior to the fire. Macroinvertebrate density, including the densities of specific species, showed varying patterns during the months of high precipitation and no individual functional feeding group showed distinct patterns during this time that are considerably different from pre-fire observations, except that the density of scrapers, driven heavily by the presence of J. kosteri, peaked in October 2000. These results suggest that the high precipitation events, which likely resulted in large amounts of runoff into Bitter Creek after the fire, did not seem to have significant negative effects on the biotic communities or to cause a set-back in their recovery trajectories.

Prediction 2: Direct heating from the fire, increased light availability due to removal of riparian vegetation, and increased sediment loads due to increased erosion and runoff will lead to higher water temperatures at Bitter Creek. In addition, decreased dissolved oxygen concentrations at Bitter Creek will occur after the fire due to reduced solubility of DO with increased temperature and increased respiration due to the decomposition of burned vegetation in the stream.

The ARIMA analysis confirmed that water temperature was significantly increased and DO concentrations were significantly decreased following the Sandhill fire. Additionally, pH, salinity, and specific conductance were also significantly altered. While sensitivity to acidity varies among invertebrate orders, decreased stream pH has been shown to lead to overall reduced macroinvertebrate abundance, both in experimental studies and field observations (Courtney and Clements, 1998). The average pH of Bitter Creek increased in this study, becoming slightly more basic, likely due to ash deposition and the basic properties of burned material (Dissmeyer, 2000). This is not expected to have had negative effects on the macroinvertebrate communities. While several species can handle periods of increased salinity through tolerance or avoidance, studies suggest that higher salinity generally has adverse effects on macroinvertebrate richness and diversity (Nielsen et al., 2003, Amaral and Costa, 1999). However, it has also been shown that desert aquatic taxa can tolerate wide fluctuations in salinity, particularly if organisms are hypothesized to have marine origins (Herbst and Bromley, 1984). Gammarus amphipods within

Bitter Lake National Wildlife Refuge, for example, have shown to exhibit medium- to high- salinity tolerance (Seidel et al., 2010). Although salinity values were significantly changed by the wildfire in the present study, the aquatic habitats within the Chihuahuan Desert likely originated from an ancient ocean and many of the macroinvertebrate taxa that spend their entire life cycle in water would be expected to have marine ancestries and exhibit wide salinity tolerances, and therefore should not have been negatively affected by the salinity changes observed. The observed increase in temperature and decrease in DO concentrations are likely to be the most biologically significant water quality alterations. The increase in water temperature likely resulted from increased light availability to the stream and the input of sediment in Bitter Creek after the fire, while the decrease in DO was likely due to solubility effects with increased temperature and increased decomposition and respiration ensuing from increased nutrients in the burned site. Furthermore, as floating algal mats were the main primary producers found in Bitter Creek following the fire, it is a possibility that more of the DO produced was lost to the air than dissolved into the aquatic habitat. Pearson correlation coefficients revealed that water temperature was significantly negatively related to diversity of the macroinvertebrates of this study and, although not significant, temperature was negatively related to taxon richness. Dissolved oxygen displayed fairly strong, but not significant, positive correlations with taxa richness and diversity. The observed increased temperatures and decreased DO in this study may contribute to the slight observed drop in taxon richness and diversity. Although macroinvertebrate richness and diversity have been shown to exhibit a positive linear relationship with maximum stream temperature in tropical and temperate regions (Jacobsen et al., 1997), the temperature effects in the current study have also been observed following a moderate to high intensity wildfire in Colorado in 2002 (Hall and Lombardozzi, 2008). Elevated water temperatures and decreased DO concentrations were observed one year following the fire and these changes did not appear to have a significant effect on any single taxon alone, but overall richness and density of macroinvertebrates were lower after the fire. The changes in temperature and DO may also affect an endangered macroinvertebrate species of Bitter Creek. The endangered amphipod G. desperatus depends on cool, well-oxygenated waters, so the increase in temperature and decrease in dissolved oxygen concentrations at Bitter Creek after the fire may help explain the absence of this amphipod from many post-fire samples.

Both Bitter Creek and Sago Spring are listed as critical habitat for J. kosteri, P. roswellensis, and G. desperatus under the U.S. Endangered Species Act. Critical habitat refers to areas that are essential to the conservation of a species and may require specialized management practices and protection (Department of the Interior, 2011). Consequently, the responses of these species to a disturbance event are of serious and immediate significance to management agencies. In this study, there was an observed significant increase in the density of the endangered J. kosteri following the Sandhill fire, and this species accounted for the majority of total macroinvertebrate density and scraper density. Scrapers tend to be recognized as specialists, so they might not be expected to do well in a disturbed habitat. However, this increase might be partially explained by changes in primary productivity at Bitter Creek and reproductive patterns of the species. Parthenogenesis is common in hydrobiid snails of North America, meaning that the presence of one female is generally sufficient to initiate a new population of clonal offspring. In addition, the frequent and rapid reproductive rate of hydrobiid snails can result in huge population densities, in excess of hundreds of thousands per square meter (Mackie and Claudi, 2010). Wildfires have been shown to lead to dramatic increases in primary production in streams due to the availability of light and nutrients (Betts and Jones Jr., 2009, Kelly et al., 2006), which can stimulate the growth of algae (Goldman and Carptenter, 1974, Grimm and Fisher, 1981). Prior to the Sandhill fire, algal mats were located sporadically at spring vents, but following the fire, algae formed dense mats throughout most reaches of Bitter Creek (Lang, 2005). In addition, high salinity values can have negative effects on freshwater algae, reducing growth rate and photosynthesis (Jackson et al., 1987), so the decrease in salinity at Bitter Creek after the fire may have also supported the success of algae at this site. As algae is the major food source of springsnails, this abundance of food may be responsible for the extreme increase in J. kosteri at Bitter Creek, along with the associated, although not significant, increase of scrapers and overall macroinvertebrate density after the fire. Similarly, in Iceland streams

23 experiencing varying temperatures and nutrient enrichments, as water temperatures increased, mollusk-dominated macroinvertebrate communities were present (Gislason, 2012). Interestingly, there was no significant effect on the endangered P. roswellensis, which is also a scraper and feeds on algae, although the average density did decrease slightly following the fire. Populations of another springsnail species of the genus Pyrgulopsis found in Arizona exhibit the highest correlation to structural properties of a stream, such as substrate and vegetation density, rather than non-structural characteristics, such as water chemistry or primary productivity (Malcom et al., 2005). Because P. roswellensis prefers cobble substrates, it is possible that increased sedimentation in Bitter Creek after the fire, combined with the already silty substrate, was not preferable for the sustainability of large populations of this springsnail. Even though the effects were not significant, the endangered amphipod G. desperatus did not seem to be able to withstand some of the effects of the fire. Although found at fairly low densities before the fire, G. desperatus was found at even lower densities after the fire, and was completely absent from nearly half of all post-fire samples. For gammarid amphipods, laboratory experiments and in situ tests in different rivers and seasons have shown a positive relationship between water temperature and feeding rate (Coulaud et al., 2011). Although water temperatures were increased post-fire in this study, G. desperatus fell to lower densities and were absent from many samples, suggesting factors other than temperature were affecting their densities. Gammarus desperatus is known to be acutely sensitive to polyaromatic hydrocarbons, which were found in large quantities in burned materials collected from Bitter Creek after the fire (Lang, 2005). Furthermore, G. desperatus has high DO requirements, so the decrease in DO and presence of hydrocarbon compounds at the burned site may have negatively affected this species and led to the absence of the amphipod from many post-fire samples. Although the decrease in density of this endangered species at Bitter Creek may seem alarming, the results were not significant, indicating that further monitoring of this species is needed.

Although there was likely an input of burned material into Bitter Creek after the fire, altering the detrital quality, there were no significant changes in detritivore richness or density in this study. Laboratory food quality experiments examining the growth response of varying macroinvertebrate species to burned and unburned detritus showed that, although trophic generalists may be common in burned streams, macroinvertebrate communities are not likely to shift toward burned organic matter as a major food source post-fire. Of the 11 taxa examined in the experiments, only one, a species of mayfly, was able to grow on burned material (Mihuc and Minshall, 1995). In my study, G. desperatus, which was absent from many post-fire samples, falls into the shredder FFG and feeds primarily on detritus. It is possible that the burned material found after the fire was not a quality food source for this species. However, detritivores overall did not exhibit significant changes due to the wildfire, meaning that they were either able to utilize the burned organic material or that, due to the low- to-moderate intensity of the fire, the detritus was not sufficiently burned to limit significant nutritional content. In addition, as vegetation regrew, detritus inputs may have become more similar to pre-fire conditions.

Macroinvertebrate communities often exhibit high variability over short time periods (Kosnicki and Sites, 2011) and in this study β-diversity was high in relation to α-diversity at both sites in all seasons, indicating that there is high species turnover between time periods in these systems. Pre- vs. post-fire community similarity was higher at Sago Spring than at Bitter Creek in all seasons and the overall average similarity between Bitter Creek and Sago Spring macroinvertebrate communities decreased slightly post-fire. Decreases in macroinvertebrate similarity between streams and stream sections have been shown to occur with large increases in stream temperatures (Lessard and Hayes, 2003), so my average values of similarity along with

25 the significant increase in temperature seen in this study may suggest that the wildfire led to the communities of Bitter Creek being less similar to those of Sago Spring than they had been before the fire occurred. Furthermore, among-season similarity continued to be lower in the burned site than in the reference site, which is common in environments affected by wildfire (Minshall et al., 2001) because macroinvertebrate communities in disturbed sites may show greater year-to-year variability than those from undisturbed sites (Richards and Minshall, 1992).

The macroinvertebrate communities of Bitter Creek were not significantly altered by the Sandhill wildfire, so no substantial amount of time was needed to return to pre-fire conditions or establish a new “normal” within the system. Long-term negative effects on macroinvertebrate communities of burned streams are often largely due to the loss of riparian vegetation and increased sedimentation due to runoff. In forested environments, this runoff may frequently contain large woody debris, which can severely alter the stream channel and have adverse effects on the biota of a stream (Minshall et al., 2001). In streams of central Idaho, USA, for example, macroinvertebrate communities in more severely burned catchments exhibited greater year to year variation, which was attributed to the influence of increased input of sediment and woody debris and the interaction of these inputs with varying flow velocities of the burned streams (Arkle et al., 2010). The absence of large woody debris in the shrub and grass dominated landscape of Bitter Lake National Wildlife Refuge may have contributed to less alteration of the stream channel and less stress on the macroinvertebrate communities. Possibly the largest biological change witnessed at Bitter Creek after the fire is the dominance of the invasive P. australis, whose long-term effects on the macroinvertebrate communities of the system remain unknown. Non-native plant species have been known to benefit from wildfires in multiple types of ecosystems. The presence of bare ground, increased nutrients, and increased light availability can all aid in the establishment of non-native plants in disturbed environments (Freeman et al., 2007). In addition, non-native plants often have high

26 growth rate, high seed output, advantageous dispersal modes, and reach large biomass, all of which may aid in the suppression of native plants (Park and Blossey, 2008, Holomuzki and Klarer, 2010). Phragmites australis has the capability of greatly diminishing light availability to the biota of the stream and altering the structure of the land-water interface. Salinity has been shown to exhibit a negative relationship with the expansion of P. australis stands in other environments (Chambers et al., 2003), so perhaps the decrease in salinity at Bitter Creek, in combination with other favorable factors for invasive fauna, allowed for the dense growth of the common reed following the Sandhill fire. The shading effect of this dominating plant at Bitter Creek may explain the narrower range of temperatures observed in the post-fire spring and summer time series. Although the specific effects of this invasion at Bitter Creek have not yet been examined, it is important to note that this major alteration in the riparian corridor, which likely resulted from the Sandhill fire, could have unforeseen effects on the biotic communities of this system. One current proposed method for controlling the density of this invasive species within Bitter Lake National Wildlife Refuge is by the application of herbicides. When stress in an environment is alleviated, macroinvertebrates are often able to quickly recover (Moreno et al., 2010). Additionally, low intensity fires may have negligible effects in impacted streams (Bayley et al., 1992, Townsend and Douglas, 2004), while high intensity burns can have significant community effects (Malison and Baxter, 2010). The investigation of wildfire effects in this study indicated no significant changes in any of the community-based indices that were examined and showed significant effects on only one species within the system, J. kosteri, which increased in density. The properties of the Sandhill fire closely mimicked those of prescribed fires, in that it was short in duration, lasting less than 2 days, and low- to-moderate intensity. Prescribed fires have been shown to have insignificant effects on biota in disturbed aquatic habitats in different types of ecosystems (Bêche et al., 2005, Arkle and Pilliod, 2010). The duration, severity, and location (grassland, rather than forest) of the Sandhill fire may have led to the minimal consequences on the macroinvertebrate communities.

7. CONCLUSIONS, IMPLICATIONS, AND FUTURE DIRECTIONS

While many desert vertebrate and invertebrate species depend heavily on access to wetlands for survival, these habitats make up only a tiny fraction of the overall ecosystem (Thompson et al., 2002). Desert springs are some of the most vulnerable ecosystems in the world and the organisms which inhabit these environments are threatened by frequent groundwater depletion, surface water diversion, habitat destruction, and the introduction of non- native invasive species (Kodric-Brown and Brown, 2007). Desert springs and their associated riparian habitats are rarely considered in the characterization and prioritization of critical habitats in need of conservation due to the remoteness of the sites, the necessity and cost of repeated visits, and the large variability of these systems (Thompson et al., 2002). The wetlands of the southwestern United States and the aquatic biota which inhabit these environments may experience frequent wildfire events (Earl and Blinn, 2003). Although it should not be unexpected that habitats and species have become disturbance-adapted in areas where fire disturbances have been historically common (Tucker et al., 2011), wildfire frequency is expected to change in many areas of the world and the effects of these changes are unclear. Alterations in climate events due to the effects of greenhouse gasses may increase the risk of wildfire and fire intensity, as well as lead to a lengthened wildfire season and an increase in total number of wildfires occurring annually (Running, 2006, Westerling et al., 2006). Furthermore, fluctuations in extreme weather events can also increase the danger of severe wildfire seasons (Liu et al., 2010). Because these delicate spring habitats are home to endemic and endangered macroinvertebrate species and little is known about how these disturbances may affect the biota of these habitats, examining effects of wildfire in these systems may be important in informing future land management and species conservation. The use of BACI design when assessing effects of natural disturbance is rare in ecological studies. More often, reference locations are chosen after the disturbance based on their proximity and physical similarity to the affected site, while pre-disturbance data comparisons are frequently lacking. Without the use of a BACI design, response variables may be less apparent or non-detectable (Price et al., 2011). The use of this approach in the current study, combined with the before- and after-disturbance analysis of water quality, allowed the detection of significant abiotic alterations at Bitter Creek after the fire, most notably an increase

28 in water temperature and a decrease in DO concentrations. There were also no apparent effects on the macroinvertebrate community as a whole, but the endangered species, J. kosteri, seemed to be enhanced by this disturbance and the endangered amphipod G. desperatus was absent from many post-fire samples. Although the effects on the amphipod species were not significant, these results still suggest that the species in this system which are the most sensitive to wildfire disturbance tend to be the endangered macroinvertebrates. The lack of community, trophic, and other taxon-specific effects is likely due to the fact that the wildfire was short, of low- to- moderate intensity, and took place in a grass and shrub dominated landscape, along with the fact that aquatic macroinvertebrates have shown high resistance and resilience to such events in many types of ecosystems.

Based on the results presented in this study, here I offer some thoughts and views on possible future directions for the management of the biological communities within Bitter Lake National Wildlife Refuge:

 As the primary response of this wildfire on macroinvertebrates was an increase in the endangered J. kosteri, this species may be a useful indicator species within Bitter Creek and could be used to monitor future changes in response to other disturbances, extreme weather events, or pollution.

 Densities of the endangered amphipod G. desperatus were lower after the fire and this species was absent from nearly half of all post-fire samples used in this study. Although these effects were not significant, based on the acute sensitivity of this species to pollution and the fact that Bitter Creek and Sago Spring are critical habitat for the conservation of this species, further monitoring of G. desperatus populations is needed to fully understand its’ vulnerability to stressors in this ecosystem.

 The dominance of the common reed P. australis after the fire may have profound effects on the biological communities of Bitter Creek, but these effects remain unknown and warrant further research.

 The application of herbicides to control P. australis should take into account the sensitivity of many macroinvertebrate species and the endangered status of several taxa in Bitter Creek. The lack of community or trophic-based effects in this study might suggest that heavily controlled, low-intensity prescribed burns may be a safer and more effective method for controlling this invasive species, as well as promoting large populations of J. kosteri.

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Table 1. Pre- and post-fire ranges of water depth and velocity from sample sites at Bitter Creek and Sago Spring. Pre-fire Post-fire Water Velocity Water Velocity Site Water depth (m) Water depth (m) (m/sec) (m/sec) Bitter Creek 0.09-0.17 0.00-0.15 0.12-0.35 0.00-0.27 Sago Spring 0.07-0.54 0.01-0.21 0.06-0.28 0.02-0.35

Table 2. Summary statistics, by season, for six hydrochemical variables pre- and post-fire at Bitter Creek, Bitter Lake National Wildlife Refuge, New Mexico (n=number of observations). Pre-fire Post-fire Variable Statistic Fall Winter Spring Summer Fall Winter Spring Summer Dissolved oxygen (mg/L) minimum 2.9 2.9 1.0 0.2 1.8 0.8 0.7 0.7 maximum 21.8 18.0 20.0 17.0 16.5 18.2 16.2 13.5 mean 9.1 9.6 7.9 5.1 7.5 8.5 3.7 2.9 n = 1973 1914 1942 1299 2152 1329 2252 2265

Temperature (°C) minimum 5.0 5.0 10.0 16.7 4.2 5.4 12.4 17.5 maximum 18.3 22.9 27.6 26.9 20.6 18.4 28.4 27.0 mean 10.9 10.5 18.1 21.3 10.7 11.0 19.4 21.7 n = 1973 1914 1942 1299 2152 1329 2252 2265

pH minimum 7.1 6.9 6.9 6.6 7.2 7.3 6.8 6.8 maximum 8.5 8.3 8.6 7.9 8.5 9.0 8.1 8.3 mean 7.6 7.6 7.5 7.1 7.7 8.4 7.2 7.4 n = 1973 1914 1942 1299 2152 1329 2252 2265

Salinity (ppt) minimum 4.7 4.6 3.4 3.5 5.9 5.5 4.4 2.9 maximum 6.0 6.5 7.5 7.5 9.5 6.7 6.6 7.2 mean 5.2 5.5 5.2 5.1 7.0 5.9 5.4 5.0 n = 1973 1914 1942 1299 1430 1329 2252 1567 Specific Conductance (µS/cm) minimum 1067 395 6214 6395 10154 1134 3888 5650 maximum 10581 11480 12991 13115 16223 11705 11578 12610 mean 9196 9728 9214 9005 11636 10431 9143 9176 n= 1973 1914 1942 1299 2152 1329 2252 2265

Total Dissolved Solids (g/L) minimum 4.2 N/A 4.0 4.1 5.8 N/A 2.5 3.6 maximum 5.8 N/A 5.7 6.5 10.4 N/A 6.4 8.1 mean 5.6 N/A 5.2 5.6 7.2 N/A 5.6 5.6 n= 672 N/A 674 649 2152 N/A 1509 2265

Table 3. Results of ARIMA analysis of six hydrochemical variables at Bitter Creek (burned site), Bitter Lake National Wildlife Refuge, Chaves County, New Mexico. Water Quality Variable

Total Water Dissolved Specific Salinity pH Dissolved Temp. Oxygen Conductance Solids

Significant Yes Yes Yes Yes Yes No effect of fire? (t = 5.87; (t = -4.7; (t = -7.95; (t = 3.73; (t = -3.2; (t = 0.63; p<0.0001) p<0.0001) p<0.0001) p<0.001) p<0.01) p>0.5)

Effect of Increase Decrease Decrease Increase Decrease N/A fire (+1.922C) (-1.492 mg/L) (-0.909 ppt) (+0.248) (-656.829 µS/cm)

Table 4. Average within-season Sorensen’s Similarity Index (± standard error) at Bitter Creek and Sago Spring prior to and following the March 2000 Sandhill fire, as well as pre- versus post-fire. Pre-fire Post-fire Pre- vs. Post-fire Bitter Creek Winter 0.382 (±0.058) 0.536 (±0.041) 0.363 (±0.043) Spring 0.564 (N/A) 0.593 (N/A) 0.494 (±0.045) Summer 0.538 (N/A) 0.575 (±0.004) 0.494 (±0.059) Fall 0.625 (N/A) 0.607 (±0.013) 0.600 (±0.029)

Sago Spring Winter 0.751 (±0.023) 0.518 (±0.057) 0.729 (±0.054) Spring 0.737 (N/A) 0.500 (N/A) 0.536 (±0.055) Summer 0.909 (N/A) 0.751 (±0.046) 0.667 (±0.061) Fall 0.545 (N/A) 0.646 (±0.047) 0.688 (±0.053)

Table 5. Among-season Sorensen’s Similarity Index at Bitter Creek and Sago Spring prior to (above diagonal) and following (below diagonal) the March 2000 Sandhill fire. Bitter Creek: Winter Spring Summer Fall Winter 0.502 0.419 0.491 Spring 0.559 0.554 0.577 Summer 0.527 0.582 0.603 Fall 0.502 0.519 0.534 Sago Spring: Winter Spring Summer Fall Winter 0.640 0.565 0.682 Spring 0.601 0.465 0.629 Summer 0.678 0.619 0.679 Fall 0.647 0.666 0.731

Fig. 1. The Chihuahuan Desert of North America with Roswell, NM indicated (modified from Chihuahuan Desert Research Institute, Fort Davis, TX).

Fig. 2. Bitter Creek and Sago Spring study areas located within Bitter Lake National Wildlife Refuge, Chaves County, New Mexico (B.) (modified from Lang, 2005).

Fig. 3. Map of Sandhill fire perimeter located within Bitter Lake National Wildlife Refuge, Chaves County, New Mexico (U.S. Fish and Wildlife Service). The yellow circle represents the approximate location of headwater sources of Bitter Creek and the orange circle represents the approximate location of the flume of Bitter Creek.

Fig. 4. Seasonal time series of temperature (°C) measured hourly at Bitter Creek (burned site) prior to and following the March 2000 Sandhill fire (dashed line).

Fig. 5. Seasonal time series of dissolved oxygen (mg/L) measured hourly at Bitter Creek (burned site) prior to and following the March 2000 Sandhill fire (dashed line).

Fig. 6. Seasonal time series of pH measured hourly at Bitter Creek (burned site) prior to and following the March 2000 Sandhill fire (dashed line).

Fig. 7. Seasonal time series of salinity (ppt) measured hourly at Bitter Creek (burned site) prior to and following the March 2000 Sandhill fire (dashed line).

Fig. 8. Seasonal time series of specific conductance (µS/cm) measured hourly at Bitter Creek (burned site) prior to and following the March 2000 Sandhill fire (dashed line).

Fig. 9. Seasonal time series of total dissolved solids (g/L) measured hourly at Bitter Creek (burned site) prior to and following the March 2000 Sandhill fire (dashed line).

Fig. 10. Taxon richness (a) and Simpson’s Index of Diversity (b) averages (±standard error) at Bitter Creek (burned site) prior to and following the March 2000 Sandhill fire (black dashed line).

Fig. 11. Taxon richness (a) and Simpson’s Index of Diversity (b) averages (±standard error) at Sago Spring (reference site) prior to and following the March 2000 Sandhill fire (black dashed line).

Fig. 12. Average total macroinvertebrate density (±standard error) at Bitter Creek (burned site) (a.) and Sago Spring (reference site) (b.) prior to and following the March 2000 Sandhill fire (black dashed line).

Fig. 13. Average densities (±standard error) of three endangered macroinvertebrates at Bitter Creek (burned site) prior to and following the March 2000 Sandhill fire (black dashed line).

Fig. 14. Average densities (±standard error) of three endangered macroinvertebrates at Sago Spring (reference site) prior to and following the March 2000 Sandhill fire (black dashed line).

Fig. 15. Average densities of macroinvertebrate functional feeding groups at Bitter Creek (burned site) (a.) and Sago Spring (reference site) (b.) prior to and following the March 2000 Sandhill fire (black dashed line).

Fig. 16. Average densities of macroinvertebrate functional feeding groups with scrapers removed at Bitter Creek (burned site) (a.) and Sago Spring (reference site) (b.) prior to and following the March 2000 Sandhill fire (black dashed line).

Fig. 17. Average densities (±standard error) of larval insects at Bitter Creek (burned site) (a.) and Sago Spring (reference site) (b.), as well as richness averages (±standard error) of emerging insects at Bitter Creek (c.) and Sago Spring (d.) prior to and following the March 2000 Sandhill fire (black dashed line).

Fig. 18. Average densities (±standard error) of detritivores at Bitter Creek (burned site) (a.) and Sago Spring (reference site) (b.), as well as richness averages (±standard error) of detritivores at Bitter Creek (c.) and Sago Spring (d.) prior to and following the March 2000 Sandhill fire (black dashed line).

Fig. 19. Sorensen’s Similarity Index between Bitter Creek (burned site) and Sago Spring (reference site) prior to and following the March 2000 Sandhill fire (black dashed line).

Appendix 1. Seasonal effects of hydrochemical variables from ARIMA analysis. Seasons are represented by 3 dummy variables, with summer as the base level. Thus, s1 represents the difference between summer and fall, s2 represents the difference between summer and winter, and s3 represents the difference between summer and spring.

Hydrochemical Variable Seasonal Total Water Dissolved Specific Dummy Salinity pH Dissolved Temp. Oxygen Conductance Variable Solids t = 0.17; t = 0.53; t = -3.37; t = 7.02; t = -2.50; t = -1.70; s1 p>0.5 p>0.5 p<0.001 p<0.0001 p<0.05 p>0.05 t = 50.08; t = 1.07; t = -2.89; t = -6.23; t = -4.21; t = -4.95; s2 p<0.0001 p>0.1 p<0.01 p<0.0001 p<0.0001 p<0.0001 t = 64.79; t = -0.66; t = -2.00; t = -4.39; t = -3.40; t = -4.28; s3 p<0.0001 p>0.5 p<0.05 p<0.0001 p<0.001 p<0.0001

Appendix 2. Number of replicates sorted, raw macroinvertebrate counts (totaled for all replicates at each sample date), and assigned functional feeding groups from pre- and post-fire samples at Bitter Creek.

Pre-fire July Sept. Nov. Feb. May Aug. Nov. Feb. Mar. May 1996 1996 1996 1997 1997 1997 1997 1998 1998 1998 Number of replicates 3 3 3 3 3 3 3 12 3 3 Taxa Raw Counts: FFG: Chironomidae 1 1 collector-gatherer Chironomidae 2 3 collector-gatherer Chironomidae 6 1 11 1 collector-gatherer Chironomidae 7 13 1 2 4 1 3 19 1 collector-gatherer Chironomidae 10 2 collector-gatherer Chironomidae 11 1 46 15 collector-gatherer Dictotendipes 12 5 10 52 6 10 collector-gatherer Podonominae1 1 6 5 86 collector-gatherer Tanypodinae 54 collector-gatherer Unk. Chironomidae 1 1 2 collector-gatherer Unk. Chironomidae 3 3 collector-gatherer Chironomid Pupa 2 3 collector-gatherer Oligochaete A 54 27 75 74 126 146 402 1 80 collector-gatherer Pyrgulopsis roswellensis 476 73 73 12 37 259 578 277 2 114 scraper Juturnia kosteri 283 179 431 27 136 935 117 157 34 359 scraper Assiminea pecos 6 1 1 scraper Physella virgata 2 1 1 2 4 7 10 4 scraper Pyrgulopsis pecosensis 5 1 scraper Unk. Mollusc B 1 scraper Gammarus desperatus 60 5 3 4 4 8 92 shredder Hyalella 3 2 1 1 shredder Ostracode A 3 2 collector-filterer (continued on next page)

Ostracode B 5 10 4 1 4 2 6 collector-filterer Ostracode C 44 110 72 45 25 45 74 19 32 collector-filterer Ostracode F 9 30 1 15 7 41 15 collector-filterer Ostracode H 1 collector-filterer Ostracode R 1 collector-filterer Ostracode S 29 collector-filterer Ostracode T 1 collector-filterer Ostracode W 1 collector-filterer Water mite Brown B 1 1 predator Girardia A 5 predator Girardia B 1 37 predator Hirunidea 1 predator Unk. Annelid 1 2 collector-gatherer Megaloptera 1 9 4 6 predator Palpomyia tibialis 1 1 predator Anisoptera 2 predator Odonata (Petalulidae) 1 predator Odonata (Libellulidae) 1 1 2 predator Odonata (Coenagriidae) 1 1 predator Unk. Coleoptera F 1 predator Thysanoptera 1 predator Unk. Insect C 1 predator Aphidae (aphid A) 2 collector-gatherer Lethocerus 1 predator Unknown Clam 2 3 collector-filterer

Post-fire Mar. April May Aug. Oct. Nov. Mar. June Aug. Dec. Mar. Sept. 2000 2000 2000 2000 2000 2000 2001 2001 2001 2001 2002 2002 Number of replicates 9 3 3 3 3 3 3 3 3 3 3 3 Taxa Raw Counts: FFG: Chironomidae 7 3 2 1 5 collector-gatherer Chironomidae 10 3 2 5 2 1 collector-gatherer Chironomidae 11 1 collector-gatherer Chironomidae 13 1 collector-gatherer Dictotendipes 16 5 10 2 1 collector-gatherer Podonominae1 1 7 20 collector-gatherer Unk. Chironomidae 1 1 collector-gatherer Unk. Chironomidae 3 8 8 collector-gatherer Chironomid Pupa 1 1 collector-gatherer Oligochaete A 175 43 61 96 375 118 80 120 60 36 37 62 collector-gatherer Pyrgulopsis roswellensis 11 36 4 60 59 96 246 26 scraper Juturnia kosteri 420 78 344 2148 2257 1496 752 623 1514 684 922 914 scraper Physella virgata 18 9 7 1 3 13 1 scraper Gammarus desperatus 31 1 8 5 shredder Hyalella 3 1 7 2 2 shredder Ostracode A 10 16 1 6 3 5 2 6 4 collector-filterer Ostracode B 13 1 5 1 3 collector-filterer Ostracode C 13 28 12 100 9 79 79 8 2 collector-filterer Ostracode F 25 4 28 5 26 48 10 17 39 7 39 1 collector-filterer Ostracode G 13 collector-filterer Ostracode H 4 19 collector-filterer Ostracode R 2 collector-filterer Ostracode T 8 16 1 2 117 11 collector-filterer

(continued on next page)

Ostracode W 59 3 1 31 6 56 1 3 collector-filterer Water mite Brown B 4 2 1 predator Water mite Pale A 1 1 predator Water mite Brown A 4 predator Nematoda 4 3 1 predator Megaloptera 6 1 5 18 3 22 4 1 9 1 22 8 predator Palpomyia tibialis 2 1 predator Stilobezzia 1 predator Trichoptera 1 2 predator Anisoptera 1 predator Unk. Anisoptera 1 predator Odonata (Petalulidae) 1 predator Odonata (Libellulidae) 1 1 predator Odonata (Coenagriidae) 1 predator Unk. Odonata larvae 1 predator Coleoptera (Hydrophilidae) 1 predator Unk. Coleoptera F 1 predator Stratiomyidae 2 predator Unk. Winged Insect B 1 predator Lycosidae A 1 predator Unk. Spider E 1 predator Aphidae (aphid A) 1 collector-gatherer

Appendix 3. Number of replicates sorted, raw macroinvertebrate counts (totaled for all replicates at each sample date), and assigned functional feeding groups from pre- and post-fire samples at Sago Spring.

Pre-fire July Sept. Nov. Feb. May Aug. Nov. Feb. Mar. May 1996 1996 1996 1997 1997 1997 1997 1998 1998 1998 Number of replicates 2 2 2 2 2 2 2 4 2 2 Taxa Raw Counts: FFG: Chironomidae 6 1 collector-gatherer Chironomidae 7 3 collector-gatherer Chironomidae 10 1 collector-gatherer Chironomidae 11 1 collector-gatherer Dictotendipes 5 1 collector-gatherer Podonominae1 8 collector-gatherer Oligochaete A 1 17 147 160 107 81 58 collector-gatherer Pyrgulopsis roswellensis 153 266 323 4 400 274 76 197 12 8 scraper Juturnia kosteri 24 4 166 87 22 171 117 170 37 scraper Gammarus desperatus 3 3 2 36 5 2 15 10 1 shredder Ostracode B 4 4 8 8 30 collector-filterer Ostracode C 7 2 33 23 38 6 collector-filterer Ostracode F 3 2 collector-filterer Ostracode H 3 collector-filterer Girardia A 4 9 2 12 15 predator Girardia B 8 46 5 1 74 58 predator Notonectidae 1 predator

Post-fire April May Aug. Oct. Nov. Mar. June Aug. Dec. Mar. Sept. 2000 2000 2000 2000 2000 2001 2001 2001 2001 2002 2002 Number of replicates 2 2 2 2 2 2 2 2 2 2 2 Taxa Raw Counts: FFG: Chironomidae 7 1 3 2 8 collector-gatherer Podonominae1 1 2 5 4 collector-gatherer Unk. Chironomidae 3 1 collector-gatherer Chironomid pupae 1 collector-gatherer Oligochaete A 7 5 15 1 0 27 8 41 46 44 16 collector-gatherer Pyrgulopsis roswellensis 2 435 485 860 704 410 268 91 926 765 1358 scraper Juturnia kosteri 463 17 12 13 45 26 298 13 70 149 scraper Gammarus desperatus 106 7 8 72 64 171 39 40 65 2 234 shredder Ostracode B 2 1 27 1 collector-filterer Ostracode C 72 22 1 30 95 3 collector-filterer Ostracode F 2 2 11 1 collector-filterer Ostracode T 11 4 collector-filterer Water Mite Brown B 1 10 predator Girardia A 14 3 7 2 1 1 predator Girardia B 19 9 7 1 9 4 11 5 predator Megaloptera 1 predator

Appendix 4. Seasonal, insect, detritivore, and overall α-, β-, and γ-diversity at Bitter Creek and Sago Spring prior to and following the March 2000 Sandhill fire.

alpha Beta Gamma Bitter Creek: Pre -fire 1.83 31.17 33 Winter Post-fire 1.40 19.60 21 Pre-fire 4.67 23.33 28 Spring Post-fire 2.22 17.78 20 Pre-fire 2.33 11.67 14 Summer Post-fire 2.67 21.33 24 Pre-fire 3.50 17.50 21 Fall Post-fire 3.33 26.67 30 Pre-fire 2.1 18.9 21 Insect Post-fire 1.92 21.08 23 Pre-fire 2.50 22.50 25 Detritivore Post-fire 1.67 18.33 20 Pre-fire 5.00 45.00 50 Overall Post-fire 3.92 43.08 47 Sago Spring: Pre -fire 1.38 9.63 11 Winter Post-fire 2.00 10.00 12 Pre-fire 3.00 9.00 12 Spring Post-fire 2.25 6.75 9 Pre-fire 1.00 5.00 6 Summer Post-fire 2.17 10.83 13 Pre-fire 2.00 6.00 8 Fall Post-fire 2.33 11.67 14 Pre-fire 0.70 6.30 7 Insect Post-fire 0.45 4.55 5 Pre-fire 1.20 10.80 12 Detritivore Post-fire 0.82 8.18 9 Pre-fire 1.70 15.30 17 Overall Post-fire 1.55 15.45 17