Science, Irving K. Barber Faculty of (Okanagan)
Biology, Department of (Okanagan)
DSpace
UBCO
Meineke, Jennifer K.
Master of Science - MSc
University of British Columbia
Grasslands around the world are becoming degraded due to human causes. In British Columbia, Canada, grasslands cover only ~1% of the landmass, yet 20% of species listed as “at-risk” of extinction in British Columbia occupy grasslands. Grassland restoration is therefore important to managing imperiled species. Biosolids (post-treatment sewage) are being used to increase forage for livestock grazing in some areas. These grasslands appear to support high densities of insects and small mammal prey, potentially enhancing populations of owls and raptors. To examine impacts of biosolids on populations of prey and predators, I studied small mammals, raptors, and owls at the OK Ranch, a cattle ranch near Clinton, British Columbia. I quantified differences in North American deermouse (Peromyscus maniculatus) populations on biosolids and control pastures and studied populations of Short-eared Owls (Asio flammeus), Long-eared Owls (Asio otus), and American Kestrels (Falco sparverius), comparing yearly abundance, habitat use, and diet within this biosolids-influenced ecosystem. Deermouse populations were smaller in pastures with biosolids, likely due to an increase in grass cover in these areas, thus indicating these pastures may be being restored. American Kestrels and Short-eared Owls used biosolids pastures over control pastures disproportionate to availability. American Kestrels ate prey that were most abundant each year. When present on the study area, Short-eared and Long-eared Owls primarily ate voles, however Short-eared Owls diversified their diets more in 2019, when vole populations were low. This research has expanded our knowledge about how biosolids can impact small mammal populations and raptor and owl habitat use. These results suggest biosolids could be used in an explicit conservation capacity to help restore wildlife populations on degraded grasslands.
EFFECTS OF BIOSOLIDS ON A GRASSLAND COMMUNITY OF RODENTS AND BIRDS OF PREY IN BRITISH COLUMBIA by Jennifer K. Meineke B.Sc., Unity College, 2017 A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER OF SCIENCE in THE COLLEGE OF GRADUATE STUDIES (Biology) THE UNIVERSITY OF BRITISH COLUMBIA (Okanagan) September 2020 © Jennifer K. Meineke, 2020 ii The following individuals certify that they have read, and recommend to the College of Graduate Studies for acceptance, a thesis/dissertation entitled: Effects of biosolids on a grassland community of rodents and birds of prey in British Columbia submitted by Jennifer K. Meineke in partial fulfillment of the requirements of the degree of Master of Science. Dr. Karen E. Hodges, Biology Department, Irving K. Barber School of Arts and Sciences Supervisor Frank Doyle, M.Sc. R.P. Bio., Wildlife Dynamics Consulting Supervisory Committee Member Dr. Adam Ford, Biology Department, Irving K. Barber School of Arts and Sciences Supervisory Committee Member Dr. Cole Burton, Department of Forest Resources Management, UBC Faculty of Forestry University Examiner iii Abstract Grasslands around the world are becoming degraded due to human causes. In British Columbia, Canada, grasslands cover only ~1% of the landmass, yet 20% of species listed as “at-risk” of extinction in British Columbia occupy grasslands. Grassland restoration is therefore important to managing imperiled species. Biosolids (post-treatment sewage) are being used to increase forage for livestock grazing in some areas. These grasslands appear to support high densities of insects and small mammal prey, potentially enhancing populations of owls and raptors. To examine impacts of biosolids on populations of prey and predators, I studied small mammals, raptors, and owls at the OK Ranch, a cattle ranch near Clinton, British Columbia. I quantified differences in North American deermouse (Peromyscus maniculatus) populations on biosolids and control pastures and studied populations of Short-eared Owls (Asio flammeus), Long-eared Owls (Asio otus), and American Kestrels (Falco sparverius), comparing yearly abundance, habitat use, and diet within this biosolids-influenced ecosystem. Deermouse populations were smaller in pastures with biosolids, likely due to an increase in grass cover in these areas, thus indicating these pastures may be being restored. American Kestrels and Short-eared Owls used biosolids pastures over control pastures disproportionate to availability. American Kestrels ate prey that were most abundant each year. When present on the study area, Short-eared and Long-eared Owls primarily ate voles, however Short-eared Owls diversified their diets more in 2019, when vole populations were low. This research has expanded our knowledge about how biosolids can impact small mammal populations and raptor and owl habitat use. These results suggest biosolids could be used in an explicit conservation capacity to help restore wildlife populations on degraded grasslands. iv Lay Summary Grasslands around the world are becoming degraded due to human causes. Biosolids (post treatment sewage) are an organic soil amendment that may be used to restore degraded areas. I studied small mammals, raptors, and owls on biosolids-amended grasslands to determine how populations respond to biosolids. North American deermice (Peromyscus maniculatus) avoided biosolids pastures, which had more grass cover than control sites. Deermice are associated with degraded areas, so smaller populations of deermice in biosolids pastures indicates that these areas are being restored. American Kestrels (Falco sparverius) and Short-eared Owls (Asio flammeus) used areas with biosolids more than control areas, which may have been associated with increased abundance of their preferred prey in biosolids areas. This research has expanded our knowledge about how biosolids can impact small mammal populations and raptor and owl habitat use. v Preface I was responsible for designing the study, managing and incorporating existing data, organizing the fieldwork, analyzing the data, and writing and editing the thesis. Dr. Karen Hodges and Frank Doyle guided me throughout the process and provided suggestions for methods and analysis and feedback on thesis drafts. Hodges and Doyle came up with the initial idea for the study. Doyle assisted with lab techniques for dissecting raptor and owl pellets and analyzing their contents. Dr. Adam Ford provided comments and suggestions on the initial thesis proposal, in addition to the final written thesis. Ford also assisted with aspects of my statistical analysis and modeling. Louiza Oukil, Brooklyn Maher, and Conor McCune did the 2019 fieldwork, which consisted of small mammal trapping, raptor and owl monitoring, and pellet collection. Briana All, Kirstie Lawson, Megan Buers, and Emma Gaudreault collected the raptor and owl data and pellets during 2016-2018. Gabriela Tisnado assisted with raptor pellet dissection. Both data chapters will be submitted to peer reviewed journals. The publication from Chapter 2 will be authored by myself, Hodges, Doyle, and Louiza Oukil. The publication from Chapter 3 will be authored by myself, Hodges, and Doyle. This research will align with other publications from Hodges and Doyle about wildlife response to biosolids. Small mammal capture and handling was approved by the University of British Columbia Animal Care Committee (Certificate #A18-0068) and the British Columbia Ministry of Forests, Lands, Natural Resource Operations, and Rural Development (Permit #WL19-472657). vi Table of Contents Abstract ......................................................................................................................................... iii Lay Summary ............................................................................................................................... iv Preface .............................................................................................................................................v Table of Contents ......................................................................................................................... vi List of Tables .............................................................................................................................. viii List of Figures .................................................................................................................................x Acknowledgements ..................................................................................................................... xii Dedication ................................................................................................................................... xiv Chapter 1 Introduction..................................................................................................................1 1.1 Land Application of Biosolids ...................................................................................... 2 1.2 Deermouse Habitat Features and Use as an Indicator Species ..................................... 6 1.3 Vole Habitat Features and Population Dynamics ......................................................... 7 1.4 Vole Predators ............................................................................................................... 9 1.5 Research Objectives .................................................................................................... 12 Chapter 2 Responses of vole and deermouse populations to biosolids amendments on grazed grasslands .........................................................................................................................14 2.1 Background ................................................................................................................. 14 2.2 Methods....................................................................................................................... 18 2.3 Results ......................................................................................................................... 22 2.4 Discussion ................................................................................................................... 24 Chapter 3 Abundance, habitat use, and diet in raptors and owls on biosolids-amended grasslands......................................................................................................................................35 3.1 Background ................................................................................................................. 35 vii 3.2 Methods....................................................................................................................... 42 3.3 Results ......................................................................................................................... 46 3.4 Discussion ................................................................................................................... 48 Chapter 4 Conclusion ..................................................................................................................62 Bibliography .................................................................................................................................67 Appendices ....................................................................................................................................90 Appendix A: Small mammal trapping and data analysis ..................................................... 90 A.1: Small mammal trapping sites ........................................................................................ 90 A.2: Small mammal data analysis ......................................................................................... 90 Appendix B: Raptor and Owl Prey and Diet Analysis ........................................................ 93 B.1: Raptor and Owl Prey ..................................................................................................... 93 B.2: Raptor and Owl Diet Analysis ...................................................................................... 95 viii List of Tables Table 2.1 Deermouse captures and density estimates. Values are n (individuals caught), Lincoln Peterson estimate (deermice/ha) ± 95% CI. “--" is used when density was not calculated because no individuals were captured. ....................................................................................................... 29 Table 2.2 Mass of deermice in relation to biosolids application. Values are x̄ ± 1 SD (n). Mass is in grams. ........................................................................................................................................ 29 Table 2.3 Sex ratio and percentages of reproductive male and female deermice per trapping session. N is the total number of individuals captured across all three sites for that month on the treatment. No animals were caught in May on any of the 2018 biosolids pastures. ..................... 30 Table 2.4 Effects of treatment and month on grass cover and bare ground within biosolids and control pastures in 2019. Wald Type II Chi-Square tests were used to draw inference from Generalized Linear Mixed Models. .............................................................................................. 30 Table 3.1 Numbers of nests, pellets, and observations of Short-eared Owls, Long-eared Owls, and American Kestrels from 2017-2019a at the OK Ranch. Birds were often sighted in groups, so we present detections per 100 km (number of times that species was seen, regardless of group size), and counts per 100 km (total number of birds observed). 2017 was a high vole year, while 2018 and 2019 were low vole years. Long-eared Owls were detected only in 2017. We drove 3503, 2481, and 4646 survey kilometers in 2017, 2018, and 2019, respectively. ........................ 58 ix Table 3.2 Observations of American Kestrels over three years in relation to biosolids applications. 2017 data are from Buers et al. (2019). Birds were often sighted in groups, so we present detections (number of times that species was seen, regardless of group size), and counts (total number of birds observed). Between 2016-2019, 24%, 35%, 50%, 52% of the land had applications of biosolids, respectively. Chi-square testing was used to test for American Kestrel selection of sites with biosolids. ................................................................................................... 59 Table 3.3 Diet of American Kestrels in 2017 and 2018. In 2017, we had 54 pellets with 101 prey as published in Buers et al. (2019); here, we present the relative frequency of occurrence for comparison to 2018. In 2018, we had 47 pellets with 563 prey. 2017 was a high vole year, while 2018 was a low vole year. Relative frequency of occurrence is the percentage of a given prey type out of all prey recorded. Two pellets in 2018 contained a small number of feathers. .......... 60 Table 3.4 Diet of the Short-eared Owl. In 2016, we had 2 pellets with 6 prey: 5 voles, and 1 Meloe spp. In 2017, we had 90 pellets with 194 prey (Ormrod et al. in review); here, we present the relative frequency of occurrence for comparison to 2018 and 2019. In 2018 we had 11 pellets with 34 prey, and in 2019 we had 24 pellets with 47 prey. 2017 was a high vole year, while 2018 and 2019 were low vole years. In 2019, the majority of pellets were collected from a single nest site. Relative frequency of occurrence is the percentage of a given prey type out of all prey recorded......................................................................................................................................... 61 x List of Figures Figure 2.1 Locations of nine trapping grids at the OK Ranch, located near Clinton, British Columbia. ...................................................................................................................................... 31 Figure 2.2 Deermouse densities on the OK Ranch from June-August 2019. May is not included because of low capture numbers across all sites. Each point with bars represents the Lincoln Peterson estimate for deermice/ha and the 95% confidence interval per site per month. The points are jittered on the x-axis to avoid overlap. ......................................................................... 32 Figure 2.3 Grass cover and bare ground at the OK Ranch in 2019. Values are mean  standard deviation ........................................................................................................................................ 33 Figure 2.4 Deermice poulations were negatively related to grass cover in July at the OK Ranch. Deermice/ha = -1.02 + 76.7 grass; F = 7.61; R2= 0.52; P = 0.03. ................................................. 34 Figure 3.1 Map of biosolids applications at the OK Ranch from 2014-2019. The white outline is the approximate border of the OK Ranch study area. ................................................................... 55 Figure 3.2 American Kestrel detections on biosolids pastures in relation to biosolids availability between 2016-2019. A) The number of American Kestrel detections from 2016-2019 were 607, 332, 99, and 215, respectively. B) Standardized selection ratios (Manly’s ) for American Kestrel detections on biosolids treatments from 2016-2019. Values above the line indicate selection for biosolids and values farther from the line represent stronger selection. 2017 data are xi from Buers et al. (2019). ............................................................................................................... 56 Figure 3.3 Prey eaten by Short-eared Owls, Long-eared Owls, and American Kestrels at the OK Ranch from 2017-2019. 2017 was a high vole year, while 2018 and 2019 were low vole years. Small mammal refers either to pellets that had fur but no diagnostic bones, or other small mammal species that comprised <3.0% of the diet. The number above each bar is the number of pellets analyzed. Long-eared Owls were detected on the Ranch only in 2017. 2017 data are from Buers et al. (2019) and Ormrod et al. (in review). ........................................................................ 57 xii Acknowledgements I would like to begin by acknowledging that the land on which this study occurred is the traditional and unceded territory of the Secwepemc First Nation. I would like to thank my supervisors, Dr. Karen Hodges and Frank Doyle for selecting me for this project and sticking with me throughout. I could not have finished this thesis or this degree without their unwavering support. I would like to extend a huge thank you to my field technicians, Louiza Oukil, Conor McCune, and Brooklyn Maher. Your dedication and commitment to this project meant the world to me and I deeply appreciate all your hard work. I would also like to thank my volunteer, Gabriela Tisnado for her time, as well as the care and attention to detail she committed to her work on my project. I would like to thank my committee member, Dr. Adam Ford, for the helpful comments he provided on my proposal and final thesis, and for helping with my more advanced statistical modeling. Thank you also to Rene-Carl Dionne and John Lavery at SYLVIS, I really appreciate your willingness to help every time I emailed asking for documents and spatial data. This project could not have been completed without the hospitality provided by OK Ranch owners Lawrence and Allen Joiner, thank you for giving us access to your land and allowing us to stay and study the OK Ranch’s unique environment. I would also like to thank Carol Banman, Kelly Stacey, and Bill Spoonmore. I did not get to know you all as well as I would have liked, but I appreciate everything you did for this project and the kindness you showed to my lab mates and field technicians. A big thank you to my friends and lab mates at UBCO, Dr. Kristen Mancuso, Briana All, Leah Rensel, Angie Kelly, Logan Volkmann, and Sultana Majid for your encouragement and for xiii your help on all things including organizing small mammal trapping, statistical modeling, UBCO logistics, and for being there whenever I needed advice or support. I would like to thank family. My Mom and Dad, who always believed I could do this and always supported me in whatever ways they could. My sisters, Kelly and Karalyn, for always being there for me and saying they were proud of me. My Gram and Grandpa, for their support and for always being excited to hear about my adventures in British Columbia. My fiancé, Geordi, for sticking with me through the ups and downs of the last couple years, for talking me through hard times, celebrating my achievements, and for being there for me whenever and however he could, even with over 4,000 km between us. Finally, I would just like to end by thanking everyone who helped me in any way, big or small. I could not have done this without so much love and support. xiv Dedication To my family, for always believing I could. 1 Chapter 1 Introduction Grasslands are one of the most important biomes on earth. Grassland ecosystems cover over 40% of the earth’s landmass, contain 17% of worldwide plant diversity, support 11% of the world’s Endemic Bird Areas (Curtin and Western 2008), and can be found on every continent except Antarctica (White et al. 2000, Keersmaecker et al. 2016). More humans also call grasslands home than any other major terrestrial biome (White et al. 2000, Curtin and Western 2008, Blair et al. 2014). Grasslands provide humans with goods and services, including food, forage for livestock, medicine, and recreation (White et al. 2000, Blair et al. 2014, Fuhlendorf et al. 2018). Grasslands also provide services that help to regulate the earth’s environment, including climate control, erosion protection, and cycling water and nutrients (Matson et al. 1997, White et al. 2000, Onandia et al. 2019). Despite their importance, grasslands are being lost and degraded around the world (Blair et al. 2014). In northwestern North America, more than 90% of semi-arid grasslands have been degraded by human causes (Sullivan and Sullivan 2018). The main causes of grassland degradation are overgrazing, urban encroachment, disrupted fire regimes, and expansion of agriculture (Fleischner 1994, Matson et al. 1997, Gayton 2004, Vukovich and Ritchison 2008, Krzic et al. 2013, Newman et al. 2014). Grassland degradation leads to problems such as reduced soil quality, erosion, low levels of biodiversity, and increased densities of invasive species (Matson et al. 1997, Ceballos et al. 2010, Lee et al. 2014, Lyseng et al. 2018). Climate change also negatively impacts grasslands by exacerbating invasive species issues (Lee et al. 2014), contributing to woody plant encroachment (Kulmatiski and Beard 2013), and by causing droughts (Keersmaecker et al. 2016) that can lead to uncharacteristic wildfires (Neary and Leonard 2020). 2 In addition to grasslands’ importance to humans, many wildlife species also rely on grasslands. Wildlife, including arthropods, birds, and mammals, live in grasslands and use these areas for refuge, foraging, migrating, breeding, and raising young (White et al. 2000, Samson and Knopf 1994, Nixon et al. 2016, Nagy et al. 2020). When grasslands are lost or degraded, grassland species can lose their primary habitat, which places these populations at risk. Populations of grassland birds in North America have decreased by 53% since the 1970s, largely due to the degradation of grasslands (Askins et al. 2007, NABCI 2019, Rosenberg et al. 2019). In British Columbia, Canada, grasslands cover approximately 1% of the province’s area, however over 20% of species listed as “at-risk” in British Columbia occupy grasslands (Gayton 2004, Newman et al. 2014; B.C. Conservation Data Centre 2020). These species include the Prairie Falcon (Falco mexicanus), Short-eared Owl (Asio flammeus), Long-billed Curlew (Numenius americanus), Swainson’s Hawk (Buteo swainsoni), and Western Harvest Mouse (Reithrodontomys megalotis). Studies from the United States have shown that conservation efforts, focused on restoring grasslands and preventing habitat loss for grassland species on private agriculture and rangelands, have had positive effects on declining populations of wildlife (NACBI 2014, 2017). In this thesis, I study a wildlife community within a grazed grassland that has been altered by biosolids, a type of organic fertilizer used to increase forage for cattle but with a poorly explored capacity for ecological restoration of degraded areas. 1.1 Land Application of Biosolids Grassland restoration typically starts with re-establishing soil properties and vegetation structure to a historical state (Newman et al. 2014). Biosolids are a soil amendment with the potential for restoring degraded grasslands. Biosolids are the semi-solid leftovers after wastewater has gone through a sewage treatment plant (Elkhatib and Moharem 2015). After 3 these semi-solids have been treated for pathogens, they can be used for land application as fertilizer (Lu et al. 2012). The use of biosolids in restoration ecology is important because it is taking what was once a waste product and using it to better the environment (Paschke et al. 2005). Before land application of biosolids became commonplace, sewage sludge was frequently disposed of in non-environmentally friendly ways, such as landfilling, ocean dumping, or incineration (Paschke et al. 2005, Haynes et al. 2009, Lu et al. 2012). Land application of biosolids is considered to be the best, most environmentally friendly option for the disposal of biosolids (Singh and Agrawal 2008, Haynes et al. 2009, Kominko et al. 2017). Biosolids are effective as a fertilizer because they alter soil composition by increasing organic matter, carbon content, soil aggregation, and water retention in the soil (Gardner et al. 2010, Borden and Black 2011, Kominko et al. 2017). Biosolids also work better than chemical fertilizers and more traditional manure treatments when restoring soil formation and vegetation structure at degraded sites (Gardner et al. 2010, Lu et al. 2012). Biosolids provide iron, chlorine, manganese, and molybdenum and other micronutrients essential for plant growth that are typically not found in chemical fertilizers (Lu et al. 2012, Kominko et al. 2017). In addition, unlike chemical fertilizers, where nutrients can be lost if not taken up quickly by plants, biosolids continue to slowly release and add nutrients to the soil for a number of years after application (Lu et al. 2012, Sharma et al. 2017). This slow release of nutrients is important for plants because it means nutrients will be more readily available when needed, instead of in one initial burst (Lu et al. 2012). Further, just one application of biosolids can have significant long-term effects on the ecosystem (Newman et al. 2014, Avery et al. 2018, 2019). In British Columbia, biosolids are used on rangelands to increase forage for cattle on historically overgrazed areas (Wallace et al. 2009, Larney and Angers 2012, Avery et al. 2018). 4 Areas in British Columbia that have been overgrazed have reduced soil fertility, lower soil aggregation, reduced water holding capacity, and increased erosion (Krzic et al. 2013). Biosolids help to restore soils and grasslands to less degraded states (Wallace et al. 2009, Ippolito et al. 2010, Lu et al. 2012, Avery et al. 2018). Despite the positive effects biosolids have when applied on land, there are potential negative effects that should be considered as well. The primary concern when applying biosolids to an environment is the potential for heavy metal accumulation in the soil (Larney and Angers 2012, Lu et al. 2012). Biosolids contain heavy metals because of contamination that can come from leaching of copper pipes, industrial wastewater, and runoff from city streets (Haynes et al. 2009, Sharma et al. 2017). Risks associated with heavy metals and biosolids can be mitigated with proper treatment before application (Haynes et al. 2009, Kominko et al. 2017, Paramashivam et al. 2017), by applying biosolids in adherence with regulations (Sánchez-Monedero et al. 2004, Singh and Agrawal 2008, Lu et al. 2012), and through proper monitoring that follows best management practices (Basta et al. 2016, Avery et al. 2018). Specific to grassland restoration, another concern for the application of biosolids is reduced plant biodiversity and the potential for invasion by non-native plant species at application sites (McFarland et al. 2010, Newman et al. 2014, Blumenthal et al. 2017, Avery et al. 2019). These risks are associated with invasive species presence at the site before application and sites with low annual precipitation (MacFarland et al. 2010, Newman et al. 2014). For example, when three similar sites within southern interior British Columbia were compared, only one of the sites could be recommended for biosolids application because of slight differences in precipitation and initial plant species composition (Newman et al. 2014). The most commonly cited solutions to the invasive species problem are mitigating potential issues before biosolids 5 application, regularly monitoring and evaluating the application site for potential invaders, and heavily seeding the application site with native grasses before applying biosolids to the site (Meyer et al. 2004, McFarland et al. 2010, Borden and Black 2011, Newman et al. 2014). Because biosolids change components of the soil and vegetation structure at application sites, wildlife that live in these areas could also be affected by biosolids applications. However, despite many studies on soil and vegetation (Larney and Angers 2012, Lu et al. 2012, Sharma et al. 2017) in relation to biosolids application, there have been fewer studies looking into the impacts biosolids have on wildlife (Washburn and Begier 2011). Early studies on biosolids’ impacts on small mammals focused on heavy metal accumulation (Hegstrom and West 1989, Cheng et al. 1996). More recently, Washburn and Begier (2011) studied wildlife at a grassland site in North Carolina where biosolids had been applied regularly for 21 years. They found omnivorous ground foraging birds were more likely to use biosolids applications than insectivorous birds that forage in the air. In British Columbia, recent research into biosolids’ impacts on wildlife occurred in my study area, the OK Ranch (Lawson 2018, Buers et al. 2019, Gaudreault et al. 2019, Ormrod et al. in review). Pastures with biosolids applications enhanced habitat for Dusky Grouse (Dendragapus obscurus; Lawson 2018) and supported high densities of grasshoppers (Orthoptera; Gaudreault et al. 2019), potentially enhancing populations of American Kestrels (Falco sparverius; Buers et al. 2019). These previous studies show that biosolids can have bottom-up effects on wildlife communities. In this thesis, I investigate how wildlife responds to biosolids application, focusing on the population ecology of small mammals, primarily North American deermice (Peromyscus maniculatus) and voles (Microtus spp.) at the OK Ranch. Small mammals are ideal focal species 6 for this study as they have short generation times and respond quickly to changes in the environment (Bueno et al. 2012), allowing me to closely track the effects of biosolids over time. Deermice and voles are specifically interesting because they prefer different habitats (Stone 2007, Mulligan et al. 2013), and do not interact competitively (Galindo and Krebs 1985), which allows me to explore potential contrasting effects of biosolids on their populations. Because small mammals are also important prey for predators, I also investigate how changing populations of small mammals impacts the relative abundance and diet of raptors and owls. I also examine if biosolids have an impact on raptor and owl habitat use. Thus, I will examine how biosolids are impacting wildlife at different trophic levels within grazed semi-arid grasslands in British Columbia. 1.2 Deermouse Habitat Features and Use as an Indicator Species Deermice are geographically widespread generalist mammals found across large portions of North America and throughout mainland British Columbia (Nagorson 2002, Stone et al. 2007). Deermice prefer areas with sparse litter, simple vegetation structure, and bare ground, thus making them a species associated with burned, grazed, or degraded landscapes (Kaufman et al. 1988, Getz and Hoffman 1999, Pearson et al. 2001, Stone et al. 2007, Hales 2011, Bueno et al. 2012, DeGolier et al. 2015). Deermice are omnivorous, primarily eating seeds and arthropods; their preferred sparse environment makes it easier for them to forage (McGee et al. 1982, Sieg et al. 1986, Kaufman et al. 1988, Pearson 1999). Deermice can also move more quickly and easily in areas with simpler vegetative structure because their locomotion, primarily hops, bounds, and tight turns, requires less energy on bare ground (DeGolier et al. 2015). Deermice also prefer rocky areas because, in areas of sparse cover, rocks can provide shelter from predators (Kaufman et al. 1988, Pearson et al. 2001). 7 Because deermice are associated with degraded landscapes, they have been used as an indicator species to determine the success of grassland restoration projects (Bueno et al. 2012, Mulligan et al. 2013). In the midwestern United States, decreasing populations of deermice, in favor of increasing populations of voles, indicated a successful restoration project (Mulligan et al. 2013). Unlike deermice, grassland voles have more specific landscape requirements that must be met for their populations to thrive. 1.3 Vole Habitat Features and Population Dynamics Grassland species of vole in British Columbia include the meadow vole (Microtus pennsylvanicus), long-tailed vole (M. longicaudus), tundra vole (M. oeconomus), and montane vole (M. montanus). These four species use grasslands for food, cover, and breeding (Batzli and Jung 1980, Reich 1981, Smolen and Keller 1987, Sera and Early 2003). In contrast to the deermouse, cover is especially important to voles as it provides protection from predators and helps with thermoregulation (Birney et al. 1976, Klaus 2003). Additionally, voles are herbivores that rely on healthy abundant vegetation for their main dietary intake (Larsson and Hansson 1977, Jonsson et al. 2002). Because they rely on heavy cover, some vole species, such as the meadow and montane vole, are defined as habitat specialists (Hallett et al. 1983, Bueno et al. 2012, Sullivan and Sullivan 2018). Many vole species exhibit population cycles in parts of North America (Taitt and Krebs 1985, Krebs 1996, Krebs et al. 2018). These cycles typically occur every three to four years in montane, meadow, and tundra voles (Krebs and Myers 1974, Whitney et al. 1976, Reich 1981, Pinter et al. 1986) and have been studied on a species by species basis (Tiatt and Krebs 1985, Pinter 1986, Getz et al. 2006), in addition to studies on synchronous cycles within multi-species Microtus communities (Boonstra et al. 1998, Krebs et al. 2002, Krebs et al. 2018). Long-tailed 8 voles tend to have yearly annual population fluctuations as opposed to multi-year cycles (Tiatt and Krebs 1985), however, some studies have found their populations cycle synchronously with local Microtus populations of other species (Krebs et al. 2018). The exact drivers behind population cycles are still widely debated (Radchuk et al. 2016, Krebs et al. 2018, Oli 2019). Some theories are that population cycles are caused by intrinsic factors, such as fluctuation in age at maturity, dispersal, or sociality (Oli and Dobson 1999, Radchuk et al. 2016). There are also theories that the cycles are driven by extrinsic factors such as predators or disease (Hanski et al. 1991, Radchuk et al. 2016). In certain regions, intrinsic and extrinsic factors work together to cause population cycles, adding to the overall complexity (Radchuk et al. 2016). Specific to grassland ecosystems, vegetative cover affects vole cycles because cover influences predation risk, food availability, interactions with conspecifics, and microhabitat conditions (Birney et al. 1976, Johnston and Anthony 2008). Cattle grazing is one example of grassland degradation that negatively affects vole populations (Bueno et al. 2012). When cattle inflict high grazing pressure on a rangeland landscape, they reduce the vegetative cover below the threshold needed for vole populations to reach their peak, effectively diminishing the population cycle (Birney et al. 1976, Bueno et al. 2012). Areas that have been heavily grazed are associated with low vole density, low survival rates, and reduced body mass (Peles and Barret 1996, Klaus 2003, Bueno et al. 2012). Reduced vole populations in grazed areas are problematic because voles have considerable impacts on the surrounding ecosystem. Some researchers consider voles “keystone” species (Huitu et al. 2013, Sullivan and Sullivan 2018) because they are important prey for many predators. 9 1.4 Vole Predators Many different species of mammals and birds eat voles as a primary part of their diet (Fitzgerald 1977, Hanski et al. 1991, Sullivan and Sullivan 2018). Raptors and owls (Sherrod 1978, Baker and Brooks 1981, Korpimaki 1985, Korpimaki and Norrdahl 1991, Poulin et al. 2001) are particularly known for being specialist vole consumers, while long-tailed (Mustela frenata) and short-tailed (Mustela erminea) weasels are known mammalian vole specialists (Fitzgerald 1977). When resources are high, such as when the vole cycle is at its peak, more dietary overlap between species is possible, however, when prey densities are low, such as at the low end of a vole cycle, predators must reduce dietary overlap and minimize competition through partitioning and specialization (Lack 1946, MacArthur and Levins 1967, Gerstell and Bednarz 1999). Within grasslands in British Columbia, three raptor and owl species, American Kestrels, Short-eared Owls, and Long-eared Owls (Asio otis), inhabit the same areas and eat voles (Booms et al. 2014, Buers et al. 2019, Marks et al. 2020). Despite these similarities, these three species have different life-history strategies and respond differently to changes in prey density in areas they inhabit. American Kestrels are small diurnal falcons that migrate to British Columbia in the summer for breeding (Smallwood and Bird 2020). American Kestrels begin breeding at one year of age and breed every year (Smallwood and Bird 2020). American Kestrels are secondary cavity nesters; they nest within natural tree cavities and cavities made by woodpeckers, notably Northern Flickers (Colaptes auratus; Cooke and Hannon 2011, Smallwood and Bird 2020). The number of American Kestrels nesting within grassland habitats can vary greatly from 1-25 pairs/km2 (Smallwood and Bird 2020). This number is limited by the number of structures and cavities available for nesting within an area (Hamerstrom et al. 1973, Newton 1994). American Kestrels have limited site fidelity, often choosing to breed in different boxes or cavities within 10 the same general area in consecutive years (Steenhof and Peterson 2009). Because parasites and predation risk can be greater in cavities that were used the previous year, the switching of nesting locations from year to year is thought to be a strategy to limit risk and increase the chance of raising young (Aitken et al. 2002). American Kestrels are dietary generalists that are considered residents after migration (Poulin et al. 2001). Generalists are species that eat whatever prey is abundant in their area (Korpimäki 1985, Hanski et al. 1991). If American Kestrels are primarily eating one prey type, and that prey type suddenly decreases, they will simply switch to eating a more abundant prey type. American Kestrels are known for their dietary plasticity and ability to switch prey types when prey densities change (Rojas and Stappung 2004, Santillán et al. 2009, Smallwood and Bird 2020). Differences in seasonal availability of prey is a common factor that causes American Kestrels to change their diet (Sarasola et al. 2003, Buers et al. 2019). American Kestrels eat large quantities of insects (Yáñez et al. 1980, Rojas and Stappung 2004), however, rather than being classified as insectivores, they are more accurately classified as “carnivorous insectivores” because a large amount of their dietary biomass is commonly made up of vertebrates (Jaksić et al. 1981, Jaksić and Delibes 1987, Sarasola et al. 2003, Santillán et al. 2009). Common prey types for American Kestrels are grasshoppers (Orthoptera), beetles (Coleoptera), and small mammals, notably voles (Sherrod 1978, Sarasola et al. 2003, Buers et al. 2019, Smallwood and Bird 2020). American Kestrels are not currently listed as a conservation concern in British Columbia (B.C. Conservation Data Centre 2020). However, when British Columbia last assessed the status of American Kestrels in 2015, they noted a “declining trend” (B.C. Conservation Data Centre 2020), which can also be seen across most of North America (Smallwood et al. 2009, Sauer et al. 11 2017). Habitat loss and degradation of breeding, migration, and wintering areas have been linked to declines in American Kestrel populations (Smallwood et al. 2009) Long-eared and Short-eared Owls are medium-sized owls that, like the American Kestrel, migrate to British Columbia to breed in the summer, however, there is also evidence that some populations of Long-eared Owls in southern British Columbia that are year-round inhabitants (Marks et al. 2020, Wiggens et al. 2020). Short-eared Owls are crepuscular to cathemeral (Wiggins et al. 2020) and nest on the ground in grasslands (Clark 1975), while Long-eared Owls are nocturnal (Marks et al. 2020) and nest in stick nests built by other species in forest stands adjacent to grasslands (Marks 1986). Nesting density for both species are variable and can be difficult to determine (Marks et al. 2020, Wiggens et al. 2020). Long-eared Owl nesting density depends on the availability of nesting structures (Marks et al. 1986). In Idaho, Long-eared Owl nest densities ranged from 0.28-1.55 pairs/km2 (Marks et al. 1986). Long-eared Owls are weakly territorial and only defend the area immediately around their nest (Marks et al. 2020). Short-eared Owls are more territorial than Long-eared Owls and can defend larger territories of up to ~80 ha (Wiggins et al. 2020). A study in Montana found Short-eared Owls nest in densities averaging one nest per 5.5 hectares (Wiggens et al. 2020). Long-eared and Short-eared Owls may not breed every year, depending on the availability of their preferred prey type (Houston 2005, Booms et al. 2014, Wiggins et al. 2020) Short-eared Owls and Long-eared Owls are dietary specialists (Korpimaki and Norrdahl 1991, Poulin et al. 2001). Dietary specialists primarily consume one type of prey (Korpimäki 1985, Poulin et al. 2001). Some species are nomadic specialists because when populations of their preferred prey are low, they leave the area and move to where populations of their preferred prey are higher (Korpimäki 1985, Hanski et al. 1991). In North America and Europe, the Short-12 eared Owl is a well-documented nomadic vole specialist (Korpimäki 1985, Korpimaki and Norrdahl 1991, Booms et al. 2014). The Long-eared Owl is a nomadic vole specialist in Europe (Korpimaki and Norrdahl 1991, Tulis et al. 2015), but their nomadism and overall population responses to vole cycles in North America are poorly understood, due to their secretive nocturnal nature and lack of mark-recapture and telemetry data (Houston 2005). Studies in Saskatchewan indicate nomadic populations in parts of Canada (Houston 2005). Before widespread grassland degradation in North America, Short-eared Owls could easily move to follow vole population peaks (Booms et al. 2014). However, grassland degradation has interrupted vole cycles and has led to lower vole populations on degraded sites (Sullivan and Sullivan 2010, Booms et al. 2014). Short-eared Owl populations have been declining in North America since 1966 (Booms et al. 2014) and they are currently listed as a species of special concern in British Columbia (B.C. Conservation Data Centre 2020). Population declines of the Short-eared Owl have been attributed to reduced vole populations due to grassland degradation (Ims et al. 2008, Booms et al. 2014). Long-eared Owls are listed as secure throughout Canada (B.C. Conservation Data Centre 2020), although their populations are also in decline (Kirk and Hyslop 1998, Houston 2005, Marks et al. 2020). Like the Short-eared Owl, population declines of Long-eared Owls have been attributed to habitat loss and grassland degradation (Marks et al. 2020). Therefore, the restoration of degraded grasslands is needed to protect wildlife at multiple trophic levels. 1.5 Research Objectives The purpose of this thesis is to examine the bottom-up impacts biosolids have on a grazed semi-arid grassland in British Columbia. Within this system, I focus on small mammals, raptors, and owls to examine the response of interacting trophic levels to biosolids. In Chapter 2, I 13 address small mammal associations with biosolids and investigate how biosolids are changing vegetative structures and small mammal communities. In Chapter 3, I evaluate the relative abundances, habitat selection, and diets of American Kestrels, Short-eared Owls, and Long-eared Owls over four years in relation to small mammal densities and the biosolids-amended environment. Finally, in Chapter 4, I summarize the findings and make suggestions for future research on biosolids and wildlife. Because this thesis was a collaborative effort that I could not have completed without assistance from my supervisors, field technicians, and lab-mates, I will use the term “we” when discussing the work that was done and conclusions that were drawn within the two data chapters. 14 Chapter 2 Responses of vole and deermouse populations to biosolids amendments on grazed grasslands 2.1 Background In western North America, over 90% of semi-arid grasslands have been degraded (Sullivan and Sullivan 2018) due to human causes including overgrazing (Fleischner 1994, Krzic et al. 2013), urban encroachment (Gayton 2004), disrupted fire regimes (Allen and Palmer 2011), and expansion of agriculture (Matson et al. 1997). Grassland degradation leads to reduced soil quality, erosion, diminished biodiversity, and increased risk of invasive species (Matson et al. 1997, Ceballos et al. 2010, Lee et al. 2014, Lyseng et al. 2018). Grassland degradation negatively affects wildlife species that rely on grasslands for breeding, foraging, and migrating (Samson and Knopf 1994, Nixon et al. 2016). In British Columbia, Canada, grasslands cover approximately 1% of the total landmass (Newman et al. 2014). However, despite this small proportion, 20% of species listed as “at-risk” by the province of British Columbia occupy grasslands (Newman et al. 2014). The disproportion between grassland area, and the number of threatened or endangered species that inhabit grasslands, demonstrates the need to maintain or restore critical features of grassland landscapes for at-risk species. The goal of most grassland restoration projects is to restore degraded areas to historic states (Brown et al. 2008). Biosolids are a type of organic fertilizer that can aid in this goal. Biosolids are the solid leftovers after wastewater has gone through a sewage treatment plant (Elkhatib and Moharem 2015). Once these solids have been treated for pathogens, they can be used for land application (Lu et al. 2012). Biosolids are used in agriculture, rangelands, forestry, and restoration ecology (Dutch and Wolstenholme 1994, Fuchsman et al. 2010, Wallace et al. 2016). In the semi-arid south-central region of British Columbia, biosolids are applied for the 15 restoration of degraded rangelands (Larney and Angers 2012, Newman et al. 2014, Wallace et al. 2016, Avery et al. 2018). Biosolids are typically used to increase forage for cattle (Avery et al. 2018) but can have positive effects on the wider ecosystem (Wallace et al 2016). As an organic soil amendment, biosolids increase or improve soil organic matter, microbial activity, water-holding capacity, soil aggregation stability, and carbon and nitrogen content (Gardner et al. 2012, Lu et al. 2012, Cogger et al. 2013, Wallace et al. 2016). Grass species that can be challenging to restore after stress from intense grazing, such as bluebunch wheatgrass (Pseudoroegneria spicata), can increase after only one biosolids application (Anderson 1991, Avery et al. 2019). Biosolids can release nutrients into the soil for at least four years through bacterial decomposition, causing delayed and multi-year effects (Lu et al. 2012, Newman et al. 2014, Avery et al. 2019). For example, biosolids had the biggest effect on Nevada bluegrass (Poa secunda juncifolia) and needle-and-thread grass (Hesperostipa comata) 2-3 years after application (Newman 2008). Perennial grass cover increased dramatically in biosolids applications 16 years after the initial application (Avery et al. 2019). However, studies looking into biosolids typically focus on alterations to soil and vegetation cover in the application areas. Few studies address the wildlife that inhabits areas where biosolids have been applied (Washburn and Begier 2011). Small mammals are integral parts of grassland ecosystems due to their roles in nutrient cycling (Bakker et al. 2004, Lemoine and Smith 2019) and as prey for many predator species (Sullivan and Sullivan 2018). Small mammals are also likely to be affected by biosolids applications. Biosolids increase vegetative structure and cover (Avery et al. 2019), which are critical landscape features for many small mammal species (Johnston and Anthony 2008). Small 16 mammals also have short gestation times and high fecundity rates, meaning they respond rapidly to changing conditions (Bueno et al. 2012, Ricketts and Sandercock 2016). The montane vole (Microtus montanus) and the North American deermouse (Peromyscus maniculatus) are two small mammal species found in British Columbia grasslands where biosolids are applied. Deermice and voles prefer different habitats (Stone 2007, Mulligan et al. 2013), and do not interact competitively (Galindo and Krebs 1985), which may lead to contrasting effects of biosolids on their populations. The montane vole is an herbivorous habitat specialist that relies on vegetative cover (Birney et al. 1976, Bueno et al. 2012, Sullivan and Sullivan 2018). Grasslands with reduced cover can lead to low reproductive rates, small population sizes, and reduced body mass (Peles and Barret 1996, Klaus 2003, Bueno et al. 2012).). In grazed rangelands, vegetative cover can be decreased to the point where the area is no longer suitable for voles (Bueno et al. 2012). Loss of vegetative cover can also reduce the natural population cycles that occur within montane vole populations (Birney et al. 1976, Bueno et al. 2012). These cycles typically repeat every 3-4 years (Myers and Krebs 1974, Pinter 1986), but when vegetative cover is reduced, montane vole populations can become diminished and unable to reach their natural peak (Birney et al. 1976, Bueno et al. 2012). A decrease in montane vole habitat in rangelands, leading to smaller montane vole populations, can negatively impact many predators, including raptors and owls (Hamerstrom 1979, Poulin et al. 2001, Sullivan and Sullivan 2018). Because biosolids increase vegetative cover (Avery et al. 2019), rangelands with biosolids applications could improve habitat conditions for montane voles and bolster their populations. Deermice are geographically widespread generalists that inhabit a variety of landscapes. In contrast to montane voles, they typically favor open habitats with low vegetative cover, sparse 17 ground cover, and high amounts of rock and bare ground (Kaufman et al. 1988, Pearson et al. 2001, Stone 2007). Deermice are omnivorous, primarily eating seeds and insects (Jameson 1952). Deermice prefer areas of low cover because foraging and locomotion are easier for them on bare ground (McGee 1982, Pearson et al. 2001, DeGolier et al. 2015). The ease of collecting food, and the increased ability to evade predators, in areas of low vegetative cover typically outweighs the potential benefits of there being more food and shelter in areas with higher cover (McGee 1982, Kaufman et al. 1988, Matlack et al. 2001). Deermice have been used as an indicator species to inform researchers about the success of restoration projects because deermice prefer habitat features that correspond with degraded areas (Mulligan et al. 2013). In grazed rangelands, areas without biosolids application should be more suitable for deermice. Once biosolids have been applied, the increase in vegetative cover is thus predicted to cause deermice to decrease. The study area, the OK Ranch, is a cattle ranch in British Columbia where the impacts of biosolids on wildlife began being studied in 2016 (Lawson 2018, Buers et al. 2019, Gaudreault et al. 2019, Ormrod et al. in review). This previous work found biosolids applications enhanced habitat for Dusky Grouse (Dendragapus obscurus, Lawson 2018), while American Kestrels (Falco sparverius) used areas with biosolids more than control sites (Buers et al. 2019). Biosolids applications had a delayed effect on grasshopper (Orthoptera) populations, with the largest grasshopper densities found on pastures that had had biosolids spread 2-3 years before (Gaudreault et al. 2019). The purpose of our study is to examine the impact biosolids have on deermice and montane voles in relation to key habitat features within grazed semi-arid grasslands in British Columbia. We will compare two different years of biosolids applications (2016 and 2018) and 18 control sites with no biosolids application. Because vegetation response to biosolids can be delayed, we expect that the 2016 biosolids pastures will have the highest cover, while the control pastures will have the lowest. We have three objectives: 1) to determine the population size of deermice and montane voles, 2) to assess the condition and reproductive status of rodents and, 3) to quantify grass cover and bare ground, and link these to vole and deermouse populations. By determining population size and assessing condition and reproductive status we can determine both population and individual level responses to biosolids applications. Because montane voles are herbivorous and thrive in areas of high cover, we hypothesize population size, body mass, and reproduction will be highest on the pastures where biosolids were applied in 2016 and lowest on the control pastures (i.e. 2016 > 2018 > Control). Because deermice prefer areas with low cover for ease of foraging and locomotion, we hypothesize population size, body mass, and reproduction will be highest on control pastures (i.e. Control > 2018 > 2016). 2.2 Methods This research took place on the privately-owned OK Ranch located near Clinton, British Columbia (51°33’N, 122°0’W). The upland study area where we focused our research is 45.3 km2. The British Columbia Ministry of Forests Lands and Natural Resource Operations categorizes this site’s Biogeoclimatic Ecosystem Classification as interior Douglas fir (Pseudotsuga menziesii) with a very dry climate in a mild grassland phase (Lloyd et al. 1990). These grasslands are dominated by needle-and-thread grass, and, historically, bluebunch wheatgrass, however cattle grazing has reduced bluebunch wheatgrass in the area (Newman 2008, Newman et al. 2014). The OK Ranch sits at an elevation of 1,100 m, the annual precipitation is 401 mm, and the average yearly temperature from 2014-2019 was 3.2º C, with January and July averages being -6.5ºC and 14.5º C, respectively (Climate BC 2020). Snow falls 19 at the OK Ranch from September-May, with peak snowfall from November-March (World Weather Online 2020). The average snowfall at the OK Ranch from November-March 2014-2019 was 13.7 cm with snow falling an average of 17 days per month (World Weather Online 2020). From 2014-2018, SYLVIS, a residuals management consulting and contracting firm, applied biosolids to 22.8 km2 of the upland area where research is conducted, totaling 50.3% of the study area. In 2019, reapplication of biosolids began in pastures that had biosolids applications in 2014 and 2015, in addition to 0.79 km2 of new sites. Biosolids were primarily sourced from Metro Vancouver, in addition to five other regional sources that provided lower amounts. In all years, the biosolids were spread evenly across the surface using a tractor spreader (R. Dionne, personal communication). Grazing was not allowed on pastures where biosolids had been spread until the summer after application. All biosolids applications at the OK Ranch adhere to BC’s Organic Matter Recycling Regulation of 2002. The OK Ranch also has areas of public crown land interspersed throughout and around it. Crown land has not had biosolids applied and is largely ungrazed, with grazed areas containing cattle for only short periods of time. Crown lands were used as control areas. We established nine small mammal trapping sites of 1.1 ha each. Three were in pastures with biosolids spread in 2016, three in pastures with biosolids spread in 2018, and three in control pastures (Figure 2.1, Appendix A.1). Each site had 50 Sherman traps (H.B. Sherman Traps, 3 x 3.5 x 9") located 15 m apart in a 5x10 grid. Each trap was baited with oats, peanut butter, and an apple slice or frozen/thawed corn for moisture. Cotton batting was placed in each trap to provide insulation. 20 Each trapping grid was trapped for four sessions in May-August 2019. Traps were trapped on three consecutive nights, except when weather was rainy and did not allow trapping. In this case, traps were opened again when the weather allowed, no more than 5 days after the last successful trap night. Traps were opened and baited in the evening between 1700-1900 and were checked starting at 0630 the following morning. Each captured small mammal was ear-tagged with an individually numbered self-piercing ear tag (National Band and Tag Company, 3/16″H x 7/16″L). Small mammals were also sexed, aged, weighed, and assessed for their reproductive status. Masses were taken using a spring scale to the nearest 0.1 g. Females were considered sexually mature when they were pregnant, had a perforated vagina, had clearly developed nipples, or were lactating (Goswami et al. 2011, Pinot et al. 2016). Males were considered sexually mature when testes were scrotal (Goswami et al. 2011, Pinot et al. 2016). Non-reproductive deermice were classified as juvenile (≤ 15 g), adult (≥ 18 g), or undetermined (15.1 – 17.9 g). Non-reproductive montane voles were aged in the field based on mass and appearance. When handling was complete, each animal was released at the point of capture. Small mammal capture and handling was approved by the University of British Columbia Animal Care Committee (Certificate #A18-0068) and the British Columbia Ministry of Forests, Lands, Natural Resource Operations, and Rural Development (Permit #WL19-472657). We sampled vegetation in each of the trapping sites. Seven of the nine pastures were sampled three times during the field season, once a month from May – July. We sampled 30 random points in each treatment pasture. Points were selected by walking a random distance into the pasture and selecting which direction to walk via a random compass spin. The distance between each point was selected using a random number generator. All sampling points were 21 plotted using Google Earth. 63 points were later determined to be outside the intended treatment area and were removed from analysis. At each sampling point, a 50 cm x 20 cm rectangular Daubenmire plot was placed on the ground and percent cover within the plot was recorded in 5% increments (Daubenmire 1959). Cover categories were perennial/annual grasses, litter, forbs, shrubs, rocks, woody debris, or bare ground. A meter stick was used to determine vegetative height within the plot. Small mammal population estimates were determined using the Lincoln Peterson index and the Chapman estimator to provide a less biased estimate when working with small sample sizes (Grimm et al. 2014). For each trapping session, we combined trap nights one and two and used this as our capture data and used trap night three for our recapture data. When sites had many captures and recaptures, we calculated deermouse density through spatially explicit capture-recapture modeling using the secr package in RStudio (Efford 2019). However, due to the small number of captures and recaptures overall, we were unable to use this analysis for the majority of our sites (Appendix A.2) and we thus present results from Lincoln-Peterson. We used Generalized Linear Mixed Models using the glmmTMB package version 1.0.0 in R (Brooks et al. 2017) to evaluate the differences in percent grass cover and bare ground between pastures. We first examined collinearity via a Pearson correlation matrix; litter cover was correlated with grass cover (-0.64) and was removed from further analysis. We used a beta-binomial distribution for all models to correct for overdispersion of proportional data (Harrison 2015). We used Treatment and Month as fixed effects to evaluate the effect of seasonality on grass cover and bare ground, with sampling site as a random effect. We used Type II Wald chi-square tests to draw inference from the results (Bolker et al. 2009) using the car package version 3.0-6 in R (Fox and Weisberg 2019). We used Tukey’s Honest Significant Difference testing for 22 post-hoc multiple comparisons of GLMM models using the lsmeans package version 2.30-0 in R (Lenth 2016). We used a Two-Way ANOVA to compare the density of small mammals between month and the two treatments and the control. To assess differences in condition and reproductive status, we used a One-Way ANOVA to analyze the difference in body mass among treatments. We used a G-Test to compare the proportion of reproductive females in the three treatments. We used linear regression to test the effect of grass cover and bare ground on population sizes of deermice and montane voles. A significance level of 0.05 was used for all analyses. Data analysis was performed in RStudio Version 1.2.1335 (R Core Team 2019). 2.3 Results We had 386 small mammal captures over 1,800 trap nights from May-August 2019. Across all four sessions, we captured 251 individual deermice and 19 individual montane voles. In total, 5 montane voles were caught on the 2016 biosolids pastures (4 females, 1 male), 8 females on the 2018 biosolids pastures, and 6 on control pastures (4 females, 2 males). We did not determine population size or conduct seasonal or habitat comparisons for montane voles because of the small sample size. Few deermice were captured in May, with more individuals captured later in the summer across all sites (Table 2.1). Deermouse populations differed by month (F2,22 = 5.28, P = 0.01) and were highest in the late summer with the population in August being significantly larger than the population in June (Figure 2.2, Table 2.1). Deermouse populations also differed by treatment (F2,22 = 5.51, P = 0.01) and were significantly smaller on 2018 biosolids pastures and 2016 biosolids pastures than on the control pastures (Figure 2.2, Table 2.1). Deermouse populations did not differ between the two biosolids treatments. 23 The mass of adult deermice did not significantly differ among the three treatments (Table 2.2). The percentage of reproductive female deermice fluctuated throughout the summer on all three treatments (Table 2.3), with different patterns seen on biosolids pastures versus control pastures. On the control pastures, the percentage of reproductive female deermice was identical in June and August (36.4%) but increased sharply to 88.9% in July (Table 2.3). In contrast, in both biosolids treatments, the number of reproductive females increased from May-August (Table 2.3). However, when all of the adult females captured are totaled across the entire summer, the proportions of reproductive female mice per treatment were not significantly different (G = 1.65, P = 0.44). Grass cover was highest on biosolids treatments (Figure 2.3). Grass cover on 2016 and 2018 biosolids pastures was more than double what it was on control pastures from May-July. Grass cover was significantly higher in both biosolids treatments than on the controls in May, June, and July (Figure 2.3, Table 2.4). Grass cover was significantly higher on 2018 biosolids pastures than 2016 biosolids pastures in July but was not significantly different in May or June. Grass cover lessened in June on 2018 biosolids pastures, while it rose in both the control pastures and 2016 biosolids pastures, causing the significant Treatment x Month interaction. The average amount of bare ground was similar in May at all three treatments (Figure 2.3). Bare ground then decreased on biosolids pastures, while it increased on control pastures in June and July, causing a significant Treatment x Month interaction (Figure 2.3, Table 2.4). However, bare ground did not significantly differ between biosolids pastures or control pastures in May, June, or July (Figure 2.3, Table 2.2). 24 Grass cover was negatively related to the density of deermice in July (Figure 2.4, R2 = 0.52, P = 0.03), with larger deermouse populations on control pastures with lower ground cover. Bare ground was not a significant predictor of deermouse density in July (R2=0.30, P = 0.12). 2.4 Discussion Deermouse populations were larger and vegetation cover was lower on control pastures than on biosolids pastures, fitting with our hypothesis that deermice would prefer control areas because of lower grass cover in these areas. Body mass and the number of reproductive female deermice were the same across all treatments. Grass cover was highest in biosolids applications, while bare ground was the same on all three treatments. Deermouse populations had a negative response to increasing grass cover, a finding that is congruent with other deermouse studies (Kaufman et al. 1988, Stone 2007). We were unable to determine the response of montane voles to grass cover and biosolids because of the small number of montane voles captured. The low number of montane voles is likely due to the cyclic nature of vole populations (Fitzgerald 1977, Pinter 1986, Pearson et al. 2001), rather than being an effect of the biosolids treatments. The OK Ranch had a vole population peak in 2017 (Buers et al. 2019, Ormrod et al. in review) then a sharp population decline to a low in 2018 (Chapter 3). It is likely that the population of montane voles was still recovering during the 2019 trapping season, leading to the low capture numbers. Alternatively, low capture numbers of montane voles could be due to deermice being higher in trappability. Trapping took place at night, when deermice are predominately active (Gilbert et al. 1986), while montane voles are active day or night but are more diurnal in the summer (Drabek 1994). The timing of setting and checking traps at sunrise and sundown may have also disturbed voles, as their diurnal summer activity is characterized by early and late day 25 bimodal peaks (Drabek 1994). Finally, trappability can also be affected by bait type (Harkins et al. 2019). Our traps were baited with oats and peanut butter, which may have been more effective at luring in omnivorous deermice over herbivorous montane voles. We found that deermouse populations at the OK Ranch were larger in control pastures where the grass cover was lower. These results fit with previous studies that show deermice are commonly found in degraded areas (McGee et al. 1982, Kaufman et al. 1988, Stone 2007, Mulligan et al. 2013). Control areas at the OK Ranch did not have the additional nutrient availability that comes from biosolids that allows for higher grass cover (Newman et al. 2014, Avery et al. 2018). Instead, control pastures had sparse ground cover, which makes it easy for deermice to forage and move (Kaufman et al. 1988). The reduction of deermice in response to land application of biosolids is consistent with studies that have used deermice as indicator species during restoration projects to show when a site is returning to its historic state (Stone 2007, Mulligan et al. 2013). Deermouse populations decline as the ecosystem is restored and are replaced by specialist small mammal species that had been pushed out because of land degradation (Mulligan et al. 2013). While we do not have enough evidence to support that the specialist montane vole increased in biosolids pastures in comparison to the control pastures at the OK Ranch, biosolids application led to reduced deermouse populations and increased grass cover, consistent with the observation that these pastures are being restored (Mulligan et al. 2013). Body mass of deermice was the same on the low cover control pastures and both high cover biosolids treatments. These results contrast with a study in Sheep River Provincial Park in Alberta, Canada that found deermice in grazed areas with lower cover were lighter than those in ungrazed areas with higher cover (Bueno et al. 2012). However, the study design and duration 26 and intensity of cattle grazing in that study differed from ours. There is the potential that other factors caused by grazing cattle unrelated to cover, not apparent in our study system, were impacting the body mass of the deermice in the other study. The percentage of reproductive females was also the same across all treatments, although there was some month to month variation. Both reproduction and body mass are complex individual issues that can be driven by several complex factors related to habitat and food (Perrigo 1987, Millar et al. 1991, Arsenault and Owen‐Smith 2002, Smith et al. 2015). Because deermice primarily eat seeds and insects, including grasshoppers (Pearson and Fletcher 2008), we might expect deermice to be heavier and more reproductive on pastures with biosolids treatments where vegetation cover (Avery et al. 2019) and grasshoppers (Gaudreault et al. 2019) were more abundant. However, our results indicate that deermice were not seriously limited by food on the sparser control pastures. Conversely, we did not quantify food availability (Smith et al. 2015), test food quality (Millar et al. 1991, Arsenault and Owen‐Smith 2002), or measure energy costs of foraging (Perrigo 1987), all of which affect body mass and reproductive status. Cattle grazed the study pastures during May, June, and July. However, we did not test for differences in grazing pressure within our study system. Future studies should look into these additional factors to further determine deermice response to biosolids-amended systems. Biosolids applications led to more grass cover on biosolids pastures than on control pastures. This result is in line with previous studies on the OK Ranch that found grass cover was higher in areas with biosolids (Newman et al. 2014, Avery et al. 2019). Deermice responded negatively to this increase of grass cover on biosolids pastures, resulting in comparatively larger populations on control pastures. 27 In contrast, the amount of bare ground was similar on biosolids and control pastures. Deermouse populations showed no relationship to bare ground in any of the treatments. This result is unlike previous studies that found deermice respond positively to bare ground (Kirchner et al. 2011). Bare ground on all three of our treatments was variable but relatively low. It is possible that deermice were not responding to bare ground at the OK Ranch simply because it was rare. Unlike our original predictions, deermice populations were similar in 2016 and 2018 biosolids pastures. We thought deermice would be higher on 2018 biosolids pastures in comparison to 2016 biosolids pastures because of delayed vegetation response to biosolids on the 2018 pastures. However, 2018 biosolids pastures had higher grass cover than 2016 biosolids pastures in July, during the peak of the growing season. Deermice can respond quickly to changes in their environment (Ricketts and Sandercock 2016), which resulted in similarly small populations on both 2016 and 2018 biosolids pastures in 2019. More grass cover and more complex vegetative structure caused by biosolids can also impact small mammal populations in the winter, when they use tunnels beneath the snow for movement, foraging, and evading predators (Fitzgerald 1977, Sera and Early 2003, Korslund and Steen 2006). The amount of snow that falls throughout the winter can also impact vole and deermouse populations. When snow cover is low, small mammals may be unable to create subnivean tunnels and be more at risk of predation and more exposed to harsh climatic elements, which can impact spring population size (Fuller et al. 1969, Fitzgerald 1977). Finally, studies that examine winter impacts of livestock on small mammals, vegetation, and snow cover are scarce, however cattle may have novel impacts on small mammals during the winter, by trampling vegetation, collapsing subnivean tunnels, and compacting snow cover. Future studies 28 should look at how cattle, vegetation cover, and snow depth affect montane vole and deermouse populations throughout the winter to determine if biosolids have year-round impacts on small mammals. The terrestrial application of biosolids increases vegetative cover (McFarland et al. 2010, Avery et al. 2019) and can alter the vegetative community in ways that affect wildlife species inhabiting grasslands (Washburn and Begier 2011, Lawson 2018, Buers et al. 2019, Gaudreault et al. 2019). We have shown that biosolids can also change vegetative structure in ways that alter small mammal communities on grazed arid grasslands of British Columbia. The generalist deermouse had smaller populations on control pastures, likely because of heightened grass cover in these areas, a response that demonstrates biosolids are having a positive restorative effect (Mulligan et al. 2013). Additional small mammal trapping should be conducted over multiple years to get a better picture of cyclic montane vole communities in areas amended with biosolids. These additional years of trapping would allow us to see if populations of specialist montane voles increase in biosolids treatments as deermouse populations decline and grass cover rises. This response from montane voles is important because many species of carnivore rely on voles as prey (Fitzgerald 1977, Sullivan and Sullivan 2018) and could benefit from the increased prey base that biosolids provide. Biosolids on rangelands in British Columbia have been used to increase forage production for cattle. Our results indicate that biosolids may also be useful to restore and support wildlife communities. The next step in this line of research is to determine if biosolids could be used explicitly for conservation purposes within grassland restoration areas. 29 Table 2.1 Deermouse captures and density estimates. Values are n (individuals caught), Lincoln Peterson estimate (deermice/ha) ± 95% CI. “--" is used when density was not calculated because no individuals were captured. Treatment May N density June N density July N density August N density Control Pastures Control A 0 -- 0 -- 7 16.9 ± 21.5 19 27.1 ± 17.4 Control B 3 3.1 ± 1.7 15 17.4 ± 5.9 20 35.4 ± 20.9 37 80.7 ± 55.7 Control C 1 0.9 ± 0.0 5 5.8 ± 3.8 25 52.9 ± 38.3 26 36.1 ± 18.3 2018 Biosolids Pastures 2018 A 0 -- 1 0.9 ± 0.0 10 25.8 ± 32.2 14 34.7 ± 35.1 2018 B 0 -- 7 7.4 ± 3.5 4 4.4 ± 2.8 5 5.0 ± 2.1 2018 C 0 -- 0 -- 4 5.3 ± 0.0 4 4.4 ± 2.8 2016 Biosolids Pastures 2016 A 1 1.8 ± 0.0 0 -- 15 23.1 ± 14.4 16 24.0 ± 18.0 2016 B 0 -- 11 15.1 ± 10.2 13 11.7 ± 2.8 14 20.4 ± 12.1 2016 C 0 -- 2 1.8 ± 0.0 8 8.0 ± 6.2 6 9.8 ± 10.7 Table 2.2 Mass of deermice in relation to biosolids application. Values are x̄ ± 1 SD (n). Mass is in grams. Control Pastures 2018 Biosolids Pastures 2016 Biosolids Pastures F P Adult Males 19.8 ± 2.4 (44) 21.3 ± 2.7 (27) 20.8 ± 3.1 (25) 2.90 0.06 Adult Females 21.3 ± 3.8 (41) 21.2 ± 4.2 (24) 22.4 ± 3.4 (35) 1.16 0.32 30 Table 2.3 Sex ratio and percentages of reproductive male and female deermice per trapping session. N is the total number of individuals captured across all three sites for that month on the treatment. No animals were caught in May on any of the 2018 biosolids pastures. Treatment N % Female % of Males Scrotal % of Adult Females Reproductive Control Pastures May 4 25.0 100.0 0 June 20 57.9 66.7 36.4 July 52 45.1 57.1 88.9 August 82 52.6 0 36.4 2018 Biosolids Pastures June 8 37.5 60.0 33.3 July 18 55.6 57.1 28.6 August 23 54.5 50.0 57.1 2016 Biosolids Pastures May 1a 0 0 -- June 13 66.7 33.3 28.6 July 36 48.5 72.7 70.0 August 36 64.7 75.0 72.7 a. One juvenile male was captured. Table 2.4 Effects of treatment and month on grass cover and bare ground within biosolids and control pastures in 2019. Wald Type II Chi-Square tests were used to draw inference from Generalized Linear Mixed Models. Fixed Effects Wald 2 d.f. p-value Grass Cover Treatment 237.44 2 < 0.01 Month 44.70 2 < 0.01 Treatment * Month 16.64 4 < 0.01 Bare Ground Treatment 0.56 2 0.76 Month 5.09 2 0.08 Treatment * Month 26.61 4 < 0.01 31 Figure 2.1 Locations of nine trapping grids at the OK Ranch, located near Clinton, British Columbia. 32 Figure 2.2 Deermouse densities on the OK Ranch from June-August 2019. May is not included because of low capture numbers across all sites. Each point with bars represents the Lincoln Peterson estimate for deermice/ha and the 95% confidence interval per site per month. The points are jittered on the x-axis to avoid overlap.33 Figure 2.3 Grass cover and bare ground at the OK Ranch in 2019. Values are mean  standard deviation34 Figure 2.4 Deermice poulations were negatively related to grass cover in July at the OK Ranch. Deermice/ha = -1.02 + 76.7 grass; F = 7.61; R2= 0.52; P = 0.03. 35 Chapter 3 Abundance, habitat use, and diet in raptors and owls on biosolids-amended grasslands 3.1 Background Many species of predators, including raptors and owls, eat the same types of prey (Hanski et al. 1991, Sullivan and Sullivan 2018). When predators that feed on the same prey types live in the same area, they must find a way to divide these shared resources. Resource partitioning is an important way that sympatric species can coexist in an area where they share limited resources (Atuo and O’Connell 2017). When sympatric predators are different sizes, prey partitioning can be as simple as smaller species selecting for smaller prey and larger species selecting for larger prey (Bosakowski and Smith 1992, Slagsvold and Sonerud 2007, Pande et al. 2018). When predators are similar sizes, therefore selecting similar-sized prey, partitioning can be based on morphological constraints, such as beak shape. Raptors with wide beaks eat more small mammals, whereas raptors with longer and narrower beaks eat more birds (Slagsvold and Sonerud 2007). When beak morphology and body size are similar, partitioning between species can be further based on where hunting takes place (Gerstell and Bednarz 1999), hunting method (Bosakowski and Smith 1992, Ward et al. 2002, Atuo and O’Connell 2017), or time of day (Aumann 2001). When prey are abundant, more dietary overlap is possible because resources are not as limited (Lack 1946, MacArthur and Levins 1967, Gerstell and Bednarz 1999). When prey become scarce, predators may show more prey partitioning, as specialization reduces competition (Lack 1946, MacArthur and Levins 1967, Gerstell and Bednarz 1999). Predators that specialize in preferred prey types will also respond numerically to reductions in their preferred prey (Boutin et al. 1995). For example, Great Horned Owl (Bubo virginianus) 36 populations are tied to cycling snowshoe hare (Lepus americanus) populations (Houston and Francis 1995, Rohner 1996, Rohner et al. 2001). When snowshoe hare populations are low, Great Horned Owls disperse to different areas, limit breeding, and have lower survival rates, causing Great Horned Owl populations to decline in the years after snowshoe hare populations decline (Houston 1987, Boutin et al. 1995, Houston and Francis 1995). Within grassland ecosystems in North America, many avian and mammalian species prefer similar prey types (Fitzgerald 1977, Hanski et al. 1991, Sullivan and Sullivan 2018). The Short-eared Owl (Asio flammeus), Long-eared Owl (A. otus), and American Kestrel (Falco sparverius) are three raptor and owl species that have similar diets and inhabit similar spaces (Booms et al. 2014, Buers et al. 2019, Marks et al. 2020). Voles (Microtus spp.) are prey that all three of these predators consume extensively (Clark 1975, Sherrod et al. 1978, Poulin et al. 2001, Smith et al. 2017). When these three species overlap, they must live in areas with high available prey densities or partition available resources. Some partitioning occurs through body size, as American Kestrels more commonly eat smaller prey types (i.e. insects) than the larger-bodied owls (Rojas and Stappung 2004). Hunting time also allows these species to partition resources. Long-eared Owls are primarily nocturnal (Marks et al. 2020), Short-eared Owls hunt day and night (Wiggins et al. 2020), and American Kestrels are diurnal (Smallwood and Bird 2020). Hunting method also varies in these species, American Kestrels use exposed perches for sit-and-wait hunting (Smallwood and Bird 2020), while Long-eared and Short-eared Owls hunt “on the wing” a short distance above ground (Marks et al. 2020, Wiggins et al. 2020) Territory sizes and nest density of Short-eared and Long-eared Owls can vary greatly depending on the size of prey populations, so the number of breeding pairs and territory size change from year to year (Houston 2005, Booms et al. 2014, Wiggins et al. 2020). In Manitoba, 37 the average defended territory for Short-eared Owls was 82 ha (Clark 1975). In Alaska, no nesting Short-eared Owls were found during a lemming (Lemmus sibiricus) decline, then two years later, 28 nests (~3-4 pairs/2.5 km2) were located during a lemming peak (Pitelka et al. 1955). Long-eared Owls are less territorial than Short-eared Owls, defending only the area directly around their nest (Marks et al. 2020). In Idaho, Long-eared Owls had a nest density ranging from 0.28-1.55 pairs/km2 (Marks et al. 1986). For both Long-eared and Short-eared Owls, proportion of non-breeding individuals and chance of successfully nesting can be tied to prey abundance (Village 1981, Booms et al. 2014). American Kestrel’s breed every year starting at one year of age (Smallwood and Bird 2020). Nest density can vary greatly between 0.11-25 pairs/ km2 Smallwood and Bird 2020. This variation can be driven by nest site availability, as the installation of nest boxes has been shown to increase American Kestrels breeding densities (Toland and Elder 1987). Unlike the Short-eared Owl, which nests on the ground in grasslands (Clark 1975), American Kestrels are secondary cavity nesters (Aitken et al. 2002), and Long-eared Owls nest in forests alongside grasslands in stick nests built by other species (Marks 1986). These specialized features required for nesting by Long-eared Owls and American Kestrels can limit the number of these species that can nest in a grassland habitat and can increase the variability in the number of nesting pairs across study areas (Newton et al. 1994, Marks et al. 2020, Smallwood and Bird 2020) American Kestrels are considered resident generalists after migration (Poulin et al. 2001). Resident generalists are species that live in an area and consume whatever prey is abundant in that area (Korpimäki 1985, Hanski et al. 1991). American Kestrels have high dietary plasticity, eating whatever is most abundant in their location (Rojas and Stappung 2004, Santillán et al. 2009, Smallwood and Bird 2020). American Kestrels commonly switch between eating small 38 mammals and insects, including grasshoppers (Orthoptera) and beetles (Coleoptera), depending on abundance and time of year (Sherrod et al. 1979, Buers et al. 2019, Smallwood and Bird 2020). Kestrels have also been called carnivorous insectivores because they eat large quantities of insects, but a significant amount of their diet by biomass can be made up of vertebrates (Sarasola et al. 2003, Santillán et al. 2009). American Kestrels are weakly philopatric, commonly choosing to nest in different cavities or boxes within the same general area in multiple years (Steenhof and Peterson 2009, Smallwood and Bird 2020). However, because American Kestrels are dietary generalists, when individuals move to different areas it is based on the availability of suitable nesting locations, not because they are numerically responding to changes in prey density or tracking a specific prey type (Steenhof and Peterson 2009). In contrast to American Kestrels, Short-eared Owls and Long-eared Owls are nomadic specialists, transient species that primarily eat one prey type (Korpimäki 1985, Poulin et al. 2001). When populations of their preferred prey are insufficient, these birds respond numerically, by leaving the area and moving to where their preferred prey are more abundant (Korpimäki 1985, Hanski et al. 1991). The Short-eared Owl is a nomadic vole specialist in North America and Europe (Korpimäki 1985, Booms et al. 2014), while the Long-eared Owl’s nomadism in North America is less understood due to their nocturnal nature and lack of banding and recapture data (Kirk and Hyslop 1998, Houston 2005, Marks et al. 2020). In Europe, the Long-eared Owl is a nomadic vole specialist (Korpimaki and Norrdahl 1991, Tulis et al. 2015), and research indicates nomadism tied to vole cycles in parts of Canada because of increases in abundance when vole populations are high (Hooper and Nyhof 1986, Houston 2005). However, to fully confirm nomadism, Long-eared Owls that breed in one area must be found breeding in a 39 different area a considerable distance away in subsequent years, which has not currently been shown (Houston 2005, Marks et al. 2020). Because some vole populations are cyclic (Myers and Krebs 1974, Birney et al. 1976, Oli and Dobson 1999, Krebs et al. 2002, 2018), Short-eared Owls and nomadic Long-eared Owls lack site-fidelity and instead move from year to year depending on where vole populations are the highest (Korpimäki and Norrdahl 1991, Poulin et al. 2001, Houston 2005, Booms et al. 2014). In the past, Short-eared Owls could easily follow vole population peaks (Booms et al. 2014). However, grassland degradation has reduced habitat area and caused widespread interruptions in vole cycles, which has led to reduced vole populations on degraded sites (Sullivan and Sullivan 2010, Booms et al. 2014). Reduced vole populations are theorized to be a major reason for population declines of the Short-eared Owl, which have been declining in North America since at least 1966 (Ims et al. 2008, Booms et al. 2014). Long-eared Owl populations are also declining in Canada and other parts of their North American range (Kirk and Hyslop 1998, Marks et al. 2020). Like the Short-eared Owl, population declines of the Long-eared Owl have been attributed to habitat loss, including loss of grasslands (Marks et al. 2020). Natural grasslands around the world have been degraded and lost due to factors including urban encroachment, disrupted fire regimes, and expansion of agriculture (Fleischner 1994, Matson et al. 1997, Gayton 2004, Vukovich and Ritchison 2008, Krzic et al. 2013, Newman et al. 2014). In northwestern North America, including Washington, Oregon, Idaho, and British Columbia, more than 90% of semi-arid grasslands have been degraded due to human causes (Sullivan and Sullivan 2018). In British Columbia, grasslands have become degraded since cattle began grazing on the landscape starting in the mid-1800s (Krzic et al. 2013). This grazing pressure from domestic cattle changed natural historic grasslands into rangelands (Fleischner 40 1994, Wikeem and Wikeem 2004, Sullivan and Sullivan 2018). Many rangelands faced heavy grazing pressure, causing a backward shift in the successional level of the ecosystem and reducing overall productivity in these overgrazed areas (Fleischner 1994, Krzic et al. 2013). Today, some ranchers are using chemical or organic fertilizers to restore vegetative cover and increase the amount of forage for their cattle on degraded areas (Larney and Angers 2012). Some ranchers in British Columbia are using biosolids, an organic soil amendment known for quickly and effectively increasing plant growth in degraded areas (Wallace et al. 2009, Larney and Angers 2012, Avery et al. 2018). Biosolids are derived from the solid leftovers at wastewater treatment facilities (Elkhatib and Moharem 2015). After being treated for pathogens they are safe for land application (Larney and Angers 2012, Newman et al. 2014, Wallace et al. 2016, Avery et al. 2018). Just one application of biosolids can have long term impacts on soil quality and vegetation cover (Avery et al. 2019). Although there have been many studies on how biosolids affect soil, vegetation, cattle, and humans (Meyer et al. 2004, Wallace et al. 2009, Borden and Black 2011, Larney and Angers 2012, Lu et al. 2012), there is a shortage of studies on the impacts of biosolids on wildlife, especially as it pertains to animals at higher trophic levels, such as raptors and owls. At the OK Ranch, a cattle ranch near Clinton, British Columbia, biosolids have been spread regularly since 2014. Research into how wildlife has responded to biosolids applications at the OK Ranch began in 2016 (Lawson 2018, Buers et al. 2019, Gaudreault et al. 2019, Ormrod et al. in review, Hodges and Doyle unpublished). American Kestrels and Short-eared Owls at the OK Ranch used biosolids application areas more than control areas (Buers et al. 2019, Ormrod et al. in review), and grasshopper densities were higher in areas with biosolids (Gaudreault et al. 2019). In 2017, there was a natural peak in vole populations at the OK Ranch (Buers et al. 2019, 41 Ormrod et al. in review). This peak may have contributed to Long-eared Owls, Short-eared Owls, and Northern Harriers (Circus hudsonius) being present on the ranch and sharing voles as their primary prey source, with three vole species comprising more than 90% of their diet (Ormrod et al. in review). These three vole species also made up nearly 50% of American Kestrels’ diet in 2017 (Buers et al. 2019). These four species of raptor and owl may have been able to share prey types because biosolids promoted vegetation growth that led to bottom-up impacts on insects (Gaudreault et al. 2019) and potentially voles (Buers et al. 2019). We studied American Kestrels, Short-eared Owls, and Long-eared Owls at the OK Ranch. We assessed abundance, habitat use, and diet composition from 2016–2019, while we assessed diet in relation to rodent availability for only 2019. Our objectives were to: 1) assess whether we observe nomadic species appearing when densities of small mammals are high, 2) examine habitat use of raptors and owls in relation to biosolids application, 3) track diet shifts in the three focal species in relation to the vole cycle, and 4) relate prey selection of raptors and owls to rodent density in 2019. We expected that the abundance of Short-eared Owls and Long-eared Owls would be highest in 2017, during the vole peak, and lowest in 2018 and 2019, in the years of vole decline, while American Kestrel populations will remain stable across all years. We expect areas with biosolids applications will be more productive (Buers et al. 2019, Gaudreault et al. 2019), and therefore raptors and owls will use these pastures more than control areas. We expect that biosolids will bolster the insect and small mammal prey base at the OK Ranch (Buers et al. 2019, Gaudreault et al. 2019, Ormrod et al. in review), leading to raptor and owl species sharing dietary resources. However, due to other factors impacting the prey base, such as vole cycling, we also expect yearly changes in diet for each species, with American Kestrels having the most 42 significant diet changes. In 2019, North American deermice (Peromyscus maniculatus) were far more common than voles on the OK Ranch (chapter 2), so if Short-eared or Long-eared Owls are present we expect they will have to vary their diet to include more non-vole species, such as the abundant deermouse. 3.2 Methods We worked on the OK Ranch, a privately-owned cattle ranch near Jesmond, British Columbia (51°33’N, 122°0’W), from 2016-2019. Our research focused on 45.3 km2 of rangeland that consisted mostly of grasslands dominated by bluebunch wheatgrass (Pseudoroegneria spicata), needle-and-thread grass (Hesperostipa comate), junegrass (Koeleria macrantha), and Nevada bluegrass (Poa secunda juncifolia). However, within these grasslands, there are small woodlots that have a history of logging, and wetlands and ponds that are home to various waterfowl. The British Columbia Ministry of Forests Lands and Natural Resource Operations categorizes the OK Ranch’s general area as interior Douglas fir (Pseudotsuga menziesii) with a very dry climate in a mild grassland phase (Lloyd et al. 1990). The average yearly temperature at the OK Ranch from 2014-2019 was 3.2º C, with January and July averages being -6.5ºC and 14.5º C respectively (Climate BC 2020). Snow falls at the OK Ranch from September-May, with peak snowfall from November-March (World Weather Online 2020). The average snowfall at the OK Ranch from November-March 2014-2019 was 13.7 cm with snow falling an average of 17 days per month (World Weather Online 2020). The average elevation is 1,100 m and the annual precipitation is 401 mm per year (Avery et al. 2019). From 2014-2019, SYLVIS Environmental Ltd. applied biosolids regularly to the OK Ranch. We used GIS data layers provided by SYLVIS to quantify how much of the study area 43 was spread each year and calculated the total amount of biosolids in the study area. Between 2014-2018 biosolids were spread to 22.78 km2 of the 45.3 km2 rangeland study area. In 2016, 24.5% of the pastures had had biosolids applied, in 2017, 34.7%, and 2018, 50.3%. In 2019, the reapplication of biosolids began in pastures that had biosolids applications in 2014 and 2015, and new sites also had biosolids applied, with 52.0% of pastures spread by the end of the season. The OK Ranch also contains areas of publicly owned Crown Land interspersed throughout and around it. Crown land has not been applied with biosolids, with grazed areas containing cattle for only short periods of time. Crown lands were used as control areas. The presence of raptors and owls at the OK Ranch was monitored through the use of “Seen Sheet” data, visual observations of species of interest recorded while driving around the OK Ranch (Hochachka et al. 2000, Aumann 2001, Poulin et al. 2001, Buers et al. 2019). Every researcher in a separate vehicle recorded date, start and end time, and start and end kilometers. Roads on the OK Ranch are primarily single-track dirt or gravel roads that are traveled between 10-30 kph. We used roads and fences to divide the ranch into 75 pasture or woodland sections and classified each section as either biosolids or control. We reevaluated classifications each year using annual GIS spatial layers provided by SYLVIS (Figure 3.1). When species of interest were observed, the time, location, and number of individuals were recorded. Multiple sections could be seen simultaneously from roads, so the section the focal species was first sighted in was recorded as the location for that detection. Observations occurred between 0500 and 2200. To address raptor and owl diets, regurgitated pellets were collected, and prey were identified. Pellets were collected by walking the same fence lines at the OK ranch once a month and collecting pellets at the base of fence posts. We surveyed ~11.8 km of fenceline each month. 44 Pellets were also collected around nest sites after chicks had fledged and under perch or nest trees. Pellets were identified to raptor or owl species based primarily on the location of collection, such as being found near known nest sites. Pellets found along fence lines were identified to species by using morphometrics including length, mass, and shape (Holt et al. 1987, Buers et al. 2019, Ormrod et al. in review). We collected <10 American Kestrel pellets in 2016 and 2019; we did not analyze these pellets. Therefore, we include diet data for American Kestrels for only 2017 and 2018. When skulls were present, small mammal remains were identified to species using features of the dentition and incisive foramina (Nagorson 2002, CalPhotos 2020). Skulls missing key identifying features were identified to genus or family. We counted the minimum number of small mammal individuals per pellets using the number of jaws and pelves found in each pellet. When nondiagnostic fur and bones were found in pellets, we identified remains as “small mammal”. Grasshopper mandibles were identified to species using a reference collection from the ranch (Gaudreault et al. 2019, Buers et al. 2019). The number of grasshoppers per pellet was calculated by dividing the number of individual mandibles per species by two. Any unidentifiable grasshopper remains found in pellets that lacked mandibles were counted as one individual. Beetle remains found in pellets were identified as Meloe spp. based on field observations but were not quantified to individuals. Bird remains in pellets mostly consisted of feathers and bones, with a few feet and beaks; multiple individuals were identified if several beaks were present. We identified birds as Passeriformes or Galliformes based on structures of the feathers when viewed under a compound microscope (Dove and Koch 2011). We counted 45 birds as part of the diet when feathers made up more than 20% of the pellet, as values less than 20% could indicate feathers of one bird spread over multiple pellets or feathers that were ingested while preening (Buers et al. 2019). The Vesper Sparrow (Pooecetes gramineus) was used as the weight for all passerine remains because they were the most common grassland nesting bird in the study area (Buers et al. 2019) and are known prey for Short-eared Owls and American Kestrels (Tomkins 1936, Sherrod et al. 1978). Grass and seeds were found in several pellets each year; however, these remains were minor and were likely incidental ingestion or secondary consumption and were not included in our analyses. To analyze seen sheet data, we present “detections”, the number of times we saw one or more individuals per species, and “counts”, the sum of individuals seen during all detections. For example, if we saw two American Kestrels and then another American Kestrel over a different pasture later, that was two detections and a count of three. To examine how the raptors and owls used the pastures amended with biosolids, we used a chi-square test to compare counts and detections in areas with and without biosolids applications against the proportion of the study area with biosolids-applied pastures in each year. We also calculated Manly standardized selectivity ratios (Manly’s ) using the adehabitatHS package version 0.3.15 (Calenge 2006) in RStudio version 1.2.1335 (R Core Team, 2019) to measure raptor and owl selection of biosolids and control pastures (Manly 2002). Alpha values over 0.5 (1/number of habitat types available) were considered to suggest selection for that pasture type. To analyze the diets of Short-eared Owls, Long-eared Owls, and American Kestrels, we measured the Relative Frequency of Occurrence (RFO), which looks at the proportion of each individual prey type within the pellets (n prey type/total prey consumed). Biomass estimates and Absolute Frequency of Occurrence are presented in Appendix B. Small mammals were captured 46 on four trapping occasions between May-August 2019 on nine sites (Chapter 2). To relate prey selection to small mammal density in 2019, we calculated the number of small mammals on biosolids and control pastures and created a proportion of how many small mammals on the ranch were mice versus voles. We used a chi-square test to compare the number of mice and voles available against the portion in pellets to see if owls and raptors were eating mice and voles in proportion to their availability. We also used Manly’s standardized selectivity ratios (Manly 2002) to measure Short-eared Owl selection of voles over deermice, where alpha values over 0.5 were indicative of selection for that prey type. 3.3 Results From 2016-2019, we detected American Kestrels, Short-eared Owls, and Long-eared Owls 1,355 times (Table 3.1). American Kestrels were abundant, making up 1,253 (92%) of all observations, with ~10-20 nesting pairs each year (Table 3.1). When compared to 2017, American Kestrel detections were reduced by 70% in 2018 and 35% in 2019. Short-eared Owls were abundant in 2017, being detected 95 times with three confirmed and one probable nest location identified (Ormrod et al. in review). Short-eared Owls were also present in 2016 (3 detections) and 2018 (0 detections, but pellets were collected). One Short-eared Owl nest was located in 2019, with three detections (Table 3.1). Long-eared Owls were present in 2017, with one seen sheet record, two confirmed nests, and multiple pellets (Ormrod et al. in review), but we did not confirm the presence of Long-eared Owls in 2016, 2018, or 2019. American Kestrels selected for biosolids over control pastures disproportionately to availability every year from 2016-2019 (Figure 3.2, Table 3.2). Selection for biosolids was strong each year, with alpha values ranging from 0.69 to 0.84 (Figure 3.2). In 2016, Short-eared Owls were seen only three times, all on control areas. When Short-eared Owls were abundant in 47 2017, they used areas with biosolids application significantly more than control areas and nested in pastures with biosolids applications (Ormrod et al. in review). In 2019, we had three Short-eared Owl sightings on biosolids pastures, with one successful breeding pair nesting on a pasture that had biosolids applied in 2016 (Table 3.1). From 2016-2019 we collected and analyzed 272 pellets from Short-eared Owls, Long-eared Owls, and American Kestrels. In 2017, American Kestrel diets were fairly evenly split between insects (50.6%) and voles (47.5%, Figure 3.3, Buers et al. 2019). In 2018, American Kestrel diets were primarily insect-based, with grasshoppers and beetles making up 98.7% of the individual prey in the diet (Figure 3.3, Table 3.3). American Kestrels also ate a few birds (4.1%) in 2017. Long-eared Owl pellets were collected on the OK Ranch in 2017 (Table 3.1). Long-eared Owls were not detected, nor were any pellets collected, in 2018 or 2019. In 2017, voles made up 91.6% of the diet (Ormrod et al. in review, Figure 3.3). Long-eared Owls ate several tundra voles (Microtus oeconomus; Ormrod et al. in review), which were not eaten by any of the other focal species in any year. Voles made up >92% of individual prey in the Short-eared Owl diet in 2017 and 2018 (Figure 3.2, Table 3.4, Ormrod et al. in review). In contrast, voles were only 61.7% of the Short-eared Owl diet in 2019. Deermice largely made up the remainder of the Short-eared Owl diet at 27.7%, which is much higher than all other years, when deermice never made up more than 3.0% of the diet. In 2019, the rodent prey population at the OK Ranch was 93% deermouse and 7% vole. In contrast, out of the small mammal prey Short-eared Owls ate, voles made up 69%, while deermice made up 30% (Table 3.3). Short-eared Owls ate voles disproportionally to deermice (χ2 48 = 266.37, p <0.01) and strongly selected voles over deermice (Manly’s  0.97) in 2019. Short-eared Owls also ate a few birds in 2017 (2.6%) and 2019 (4.3%, Table 3.4). 3.4 Discussion In our first objective, we examined relative abundance of raptors and owls to determine whether we observe nomadic species appearing when densities of small mammals are high. We found Short-eared Owls and Long-eared Owls were most abundant in 2017, during a peak in the vole cycle. Short-eared Owls were less abundant but present in other years, while Long-eared Owls were detected only in 2017. In contrast, American Kestrels were abundant and nested at the Ranch each year. In our second objective, we examined habitat use in relation to biosolids and found American Kestrels selected for pastures with biosolids application more than control pastures in all years. In 2017, when Short-eared Owls were most abundant, they also used biosolids more than control pastures, and when they bred (2017 and 2019) they also nested only in the biosolids pastures. In our third objective, we tracked diet shifts of raptors and owls in relation to the vole cycle. American Kestrels switched from having a nearly even split between eating small mammals and insects during the vole peak in 2017, to eating primarily insects in 2018 after the vole population declined. Long-eared Owls and Short-eared Owls were clear vole specialists, with slight dietary variation between the two species. Finally, although deermice were the most abundant small mammal on the OK Ranch in 2019, we found Short-eared Owls ate far more voles than deermice in 2019. American Kestrels were abundant, with multiple nesting pairs on the ranch from 2016-2019. However, unlike our hypothesis, American Kestrel populations did not remain stable from year to year. In 2016 and 2017, American Kestrel detections were high, with 9.5 detections/100 49 km in 2017. In contrast, this number changed to 4.0 detections/100 km and 4.7 detections/100 km in 2018 and 2019, respectively. Because American Kestrels are dietary generalists (Rojas and Stappung 2004, Santillán et al. 2009), this variation likely has little to do with prey density and more to do with observed changes in nesting location availability. Logging at the OK Ranch increased starting in winter 2017, removing about 70% of the small patches of forest in our study area by 2019 (F. Doyle, pers. comm.). Removal of trees and snags likely limited nest site availability for American Kestrels at the OK Ranch after 2017, which led to lower number of detections. Conversely, known predators of American Kestrels are present on the OK Ranch and include the Red-tailed Hawk (Buteo jamaicensis), Peregrine Falcon (Falco peregrinus), Prairie Falcon (Falco mexicanus), Sharp-shined Hawk (Accipiter striatus), Coopers Hawk (Accipiter cooperii), Northern Goshawk (Accipiter gentilis), and Great-horned Owl (Smallwood and Bird 2020), and may have contributed to American Kestrel decline. In 2017, when detections from six American Kestrel predators were combined, we had 3.0 detections/100 km, with Red-tailed Hawks making up the majority of the detections. Detections of these raptor and owl predators declined in 2018 (2.0 detections/100 km) and 2019 (1.8 detections/100 km). However, avian predators can be a major source of American Kestrel mortality (Farmer et al. 2006) and predators have been known to switch to eating other predators when prey populations decline (O’Donoghue et al. 1995). Because the vole population at the OK Ranch declined after 2017, avian predators may have predated on American Kestrels more heavily in 2018 and 2019, contributing to the decline in American Kestrel detections. As expected, the number of Short-eared and Long-eared Owls inhabiting the OK Ranch changed from year to year and was strongly associated with vole density. During the vole 50 population peak in 2017, Short-eared Owls were seen 95 times (Ormrod et al. in review) versus 6 observations total in the other 3 years. This increase in 2017 is likely due to the nomadic nature of Short-eared Owls (Booms et al. 2014), with the vole population peak attracting high numbers of breeding pairs. Similarly, Poulin et al. (2001) observed Short-eared Owls 604 times during a peak vole year on their study area in Saskatchewan, when in other years they did not record more than two observations. However, in 2019 a pair of Short-eared Owls bred and successfully raised a nest at the OK Ranch, indicating that the OK Ranch is a productive and sustainable environment for Short-eared Owls even in years when vole populations at the Ranch are low. Long-eared Owls were present on the OK Ranch only in 2017, when two nests were found (Ormrod et al. in review). Our results are similar to a study in Saskatchewan that found Long-eared Owls were extremely rare, except in years of vole population peaks (Houston 2005). However, Long-eared Owls are difficult to detect (Kirk and Hyslop 1998, B.C. Conservation Data Centre 2020) and are more nocturnal than Short-eared Owls (Marks et al. 2020). It is possible that Long-eared Owls were present in the forest on the outskirts of our main study area, but we are confident owls did not nest in 2016, 2018, or 2019 in the main study area. American Kestrels used biosolids application areas more than control areas during all four years of study. This pattern of biosolids use was especially notable in 2016 when they disproportionately selected for biosolids treatments when only 24% of the study area had biosolids applied. Areas with biosolids application substantially increase grasshopper populations (Gaudreault et al. 2019), and birds that primarily forage for ground insects prefer areas with biosolids applications over control plots (Washburn and Begier 2011). American Kestrels consume large quantities of grasshoppers throughout their range (Buers et al. 2019, Smallwood and Bird 2020), and likely prefer the biosolids application areas over the control 51 plots because of an increase in grasshopper prey in these areas. In 2017, when American Kestrels ate a considerable number of voles during a vole population peak (Buers et al. 2019), it is likely that the vole populations were highest on the biosolids plots, where American Kestrels were already foraging, because voles prefer high cover areas (Birney et al. 1976) that biosolids applications would provide. We located two Long-eared Owl nests in forested areas where biosolids were not spread, but both were surrounded by high proportions of biosolids-applied pastures (Ormrod et al. in review). However, because of low detections of Long-eared Owls (1 bird seen in 2017), we were unable to confirm potential associations between biosolids and foraging, thus more dedicated research is needed. Observations of Short-eared Owls in 2017 indicated that they used biosolids twice as often as expected if they used pastures proportional to their availability (Ormrod et al. in review). Consistent with this observation, all four Short-eared Owl nests found on the ranch since 2016 have been in plots with biosolids application (Ormrod et al. in review). Because Short-eared Owl observations are tied to higher abundances of voles (Baker and Brooks 1991, Poulin et al. 2001, Booms et al. 2014), and voles prefer areas with high cover (Birney et al. 1976), Short-eared Owls may be selecting biosolids pastures due to the higher abundance of voles in these areas. More research into Short-eared Owl nesting sites on biosolids applied areas is needed to fully determine how and why Short-eared Owls are using these types of sites. In 2017, when the vole population cycle was at its peak, Short-eared and Long-eared Owls at the OK Ranch ate almost entirely voles (Ormrod et al. in review), while American Kestrels’ diet was split almost evenly between insects and voles (Buers et al. 2019). The small vole populations in 2018 resulted in American Kestrels demonstrating their generalist tendencies, as they switched to eating primarily insects, with almost no small mammals in their diet. This 52 response is similar to American Kestrels in South America that changed their diets based on seasonal changes in prey availability, rather than selecting specific prey (Sarasola et al. 2003). In contrast, voles still made up over 90% of Short-eared Owl diets in 2018. There was some diet variation between the vole specialist Long-eared and Short-eared Owls. When Short-eared Owls nested at the OK Ranch (2017 and 2019), they ate a few birds (<5%), which were not eaten by Long-eared Owls (Ormrod et al. in review). Unlike Long-eared Owls, Short-eared Owls sometimes forage during daylight hours (Wiggins et al. 2020). Because passerines are diurnal, they would be available for the cathemeral Short-eared Owl, but not for nocturnal Long-eared Owls, indicating foraging time is a possible method of prey partitioning for these two species. Long-eared Owls ate a few tundra voles in 2017 (Ormrod et al. in review), a resource that was not identified in any raptor or owl species in any other year of study. This finding indicates that Long-eared Owls were either foraging in areas not used by the other species or that tundra voles were more detectable to Long-eared Owls than to Short-eared Owls or American Kestrels. Tundra voles prefer wet landscapes (Batzli and Henttonen 1990), indicating Long-eared Owls may have been foraging over wetland areas at the OK Ranch more than Short-eared Owls and American Kestrels in a type of landscape partitioning. However, there is little information about the preferred foraging areas of Long-eared Owls (Marks et al. 2020), so we cannot confirm this potential association. More research into where and how Long-eared Owls forage is needed to further investigate diet differences between them and the other vole-eating species at the OK Ranch. In 2019, the small mammal prey base at the OK Ranch was 93% deermouse and only 7% vole. However, the Short-eared Owl pair that nested at the OK Ranch in 2019 still selected for 53 voles disproportionately to their availability, with voles making up 61.7% of their diet. This result is similar to a study on wintering Short-eared Owls in Ohio that also found Short-eared Owls disproportionately select for voles over deermice (Colvin and Spaulding 1983). Despite this selection for voles, Short-eared Owls in 2019 were more diverse in their prey selection than the nesting pairs in 2017, with nearly 30% of the diet being deermice and shrews, and 10% consisting of birds and insects. Short-eared Owls are vole specialists and are not known for having a diverse diet in North America (Wiggins et al. 2020). Small mammals made up >90% of the Short-eared Owl diet, with vole species making up >83% of these small mammals, in British Columbia (Wiebe 1991), Iowa (Errington 1937), Massachusetts (Holt 1993), Manitoba (Clark 1975) and Ohio (Colvin and Spaulding 1983). Our finding that voles made up only 61.7% of the Short-eared Owl diet in 2019 indicates that insufficient vole populations at the OK Ranch may have led to Short-eared Owls consuming a more generalist diet. Future studies looking into the diets of American Kestrels, Short-eared Owls, and Long-eared Owls in association with prey density and biosolids should also examine the impacts of other predators on the landscape. Along with additional species of raptors and owls, Common Ravens (Corvus corax), coyotes (Canis latrans), and foxes (Vulpes vulpes) were present on the OK Ranch and are all known vole consumers (MacCracken and Hansen 1987, Hanski et al. 1991, Ormrod et al. in review). Long-tailed (Mustela frenata) and short-tailed (Mustela erminea) weasles are also found on the Ranch and are vole specialists (Fitzgerald 1977). The OK Ranch is also home to a small number of snake species, including garter snakes (Thamnophis spp.), which consume voles (Gregory et al. 1980). All of these vole consumers were foraging in the same grasslands as American Kestrels, Long-eared Owls, and Short-eared Owls, thus competing for 54 dietary resources. Competition may have impacted the type and amount of prey that was eaten by each species and may have led to additional prey partitioning when resources were low. Degradation of grassland habitats and declining raptor and owl populations are a focal conservation concern (Ims et al. 2008, Vucovich and Ritchison, 2008, Booms et al. 2014, Marks et al. 2020). In British Columbia, the Short-eared Owl is a species of special concern, while the Long-eared Owl has declined throughout its North American range (Kirk and Hyslop 1998, B.C. Conservation Data Centre 2020, Marks et al. 2020). In our study, American Kestrels and Short-eared Owls used biosolids pastures more than controls, likely because of an increase in abundance of available prey in these areas. These results show that there is the potential for biosolids to be used as a means of grassland restoration to conserve threatened species of raptors and owls. Further research into how the small mammal prey base responds to biosolids applications, and how Short-eared and Long-eared Owls use biosolids application areas for nesting and foraging, are needed to determine the extent to which biosolids can affect raptor and owl populations. 55 Figure 3.1 Map of biosolids applications at the OK Ranch from 2014-2019. The white outline is the approximate border of the OK Ranch study area. 56 Figure 3.2 American Kestrel detections on biosolids pastures in relation to biosolids availability between 2016-2019. A) The number of American Kestrel detections from 2016-2019 were 607, 332, 99, and 215, respectively. B) Standardized selection ratios (Manly’s ) for American Kestrel detections on biosolids treatments from 2016-2019. Values above the line indicate selection for biosolids and values farther from the line represent stronger selection. 2017 data are from Buers et al. (2019). 57 Figure 3.3 Prey eaten by Short-eared Owls, Long-eared Owls, and American Kestrels at the OK Ranch from 2017-2019. 2017 was a high vole year, while 2018 and 2019 were low vole years. Small mammal refers either to pellets that had fur but no diagnostic bones, or other small mammal species that comprised <3.0% of the diet. The number above each bar is the number of pellets analyzed. Long-eared Owls were detected on the Ranch only in 2017. 2017 data are from Buers et al. (2019) and Ormrod et al. (in review). 58 Table 3.1 Numbers of nests, pellets, and observations of Short-eared Owls, Long-eared Owls, and American Kestrels from 2017-2019a at the OK Ranch. Birds were often sighted in groups, so we present detections per 100 km (number of times that species was seen, regardless of group size), and counts per 100 km (total number of birds observed). 2017 was a high vole year, while 2018 and 2019 were low vole years. Long-eared Owls were detected only in 2017. We drove 3503, 2481, and 4646 survey kilometers in 2017, 2018, and 2019, respectively. a. In 2016, we did not record the kilometers driven, but hours were comparable to other years. Three Short-eared Owls were counted, and American Kestrels were detected 607 times and counted 1138 times. b. Ormrod et al. (in review) c. Buers et al. (2019) d. We did not quantify the exact number of American Kestrel nests in 2018 and 2019, the estimate of 10-20 nests is based on field observations. # of Nests Detected # of Pellets Collected Counts per 100 km Detections per 100 km Short-eared Owl 2017b 3 291 3.2 2.7 2018 0 11 0 0 2019 1 24 0.06 0.06 Long-eared Owl 2017b 2 190 0.03 0.03 American Kestrel 2017c 18 74 13.1 9.5 2018 ~10-20d 50 5.5 4.0 2019 ~10-20d 10 6.3 4.7 59 Table 3.2 Observations of American Kestrels over three years in relation to biosolids applications. 2017 data are from Buers et al. (2019). Birds were often sighted in groups, so we present detections (number of times that species was seen, regardless of group size), and counts (total number of birds observed). Between 2016-2019, 24%, 35%, 50%, 52% of the land had applications of biosolids, respectively. Chi-square testing was used to test for American Kestrel selection of sites with biosolids. sites without biosolids sites with biosolids detections counts group size ?̅? ± SD (maximum) detections counts group size ?̅? ± SD (maximum) detections χ2 p counts χ2 p 2016a 352 740 2.10 ± 2.18 (14) 255 398 1.56 ± 1.01 (6) 99.19 < 0.01 66.02 < 0.01 2017 133 - - 199 - - 83.5 < 0.01 - 2018 21 25 1.19 ± 0.51 (3) 78 111 1.42 ± 0.83 (5) 32.13 < 0.01 53.36 < 0.01 2019 33 34 1.03 ± 0.17 (2) 182 259 1.42 ± 0.98 (7) 91.61 < 0.01 155.17 < 0.01 a. In 2016, 25 detections (count = 45) occurred over pastures that were spread with biosolids during our 2016 field season. We were unable to determine if these pastures had or had not been spread with biosolids at the time the observation took place. An additional 11 American Kestrel detections (count = 29) occurred where birds were either on the edge between a control and biosolids area or moved quickly from one to the other. We have not included these records in our habitat use statistics. 60 Table 3.3 Diet of American Kestrels in 2017 and 2018. In 2017, we had 54 pellets with 101 prey as published in Buers et al. (2019); here, we present the relative frequency of occurrence for comparison to 2018. In 2018, we had 47 pellets with 563 prey. 2017 was a high vole year, while 2018 was a low vole year. Relative frequency of occurrence is the percentage of a given prey type out of all prey recorded. Two pellets in 2018 contained a small number of feathers. 2017 2018 Relative Frequency of Occurrence in n pellets n prey Relative Frequency of Occurrence mammals small mammala - 7 7 1.2 Microtus longicaudus 17.8 0 - - Microtus pennsylvanicus 22.8 0 - - Microtus montanus 6.9 0 - - Sorex spp. 1.0 0 - - birds Pooecetes gramineus 1.0 0 - - insects Pseudochorthippus curtipennis 5.0 40 200 35.5 Conozoa sulcifrons 6.9 35 160 28.4 Anabrus longipes 14.9 27 89 15.8 Meloe spp. 8.9 44 44 7.8 Melanoplus bivitattus 1.0 15 35 6.2 unknown grasshopper - 14 24 4.3 Bruneri brunnea 1.0 2 2 0.4 Anabrus simplex 1.0 1 1 0.2 Melanoplus sanguinipes 2.0 1 1 0.2 Arphia pseudonietana 2.0 - - - Camnula pellucida 7.9 - - - a. These pellets had fur but no diagnostic bones, so we were unable to differentiate species. 61 Table 3.4 Diet of the Short-eared Owl. In 2016, we had 2 pellets with 6 prey: 5 voles, and 1 Meloe spp. In 2017, we had 90 pellets with 194 prey (Ormrod et al. in review); here, we present the relative frequency of occurrence for comparison to 2018 and 2019. In 2018 we had 11 pellets with 34 prey, and in 2019 we had 24 pellets with 47 prey. 2017 was a high vole year, while 2018 and 2019 were low vole years. In 2019, the majority of pellets were collected from a single nest site. Relative frequency of occurrence is the percentage of a given prey type out of all prey recorded. 2017 2018 2019 Relative Frequency of Occurrence in n pellets n prey Relative Frequency of Occurrence in n pellets n prey Relative Frequency of Occurrence mammals Meadow Vole 7.2 3 3 8.8 0 - - Montane Vole 30.4 6 8 23.5 9 13 27.7 Long-tailed Vole 13.9 7 9 26.5 3 3 6.4 Deermouse 0.5 1 1 2.9 12 13 23.6 Microtus spp. 40.7 6 12 35.3 12 13 27.7 Sorex spp. - 1 1 2.9 1 1 2.1 birds passerine 2.6 0 - - 2 2 4.3 insects Meloe spp. - 0 - - 2 2 4.3 Orthoptera 4.6 0 - - 0 - - 62 Chapter 4 Conclusion Degradation of grasslands in North American is an important conservation concern (Blair et al. 2014). Grassland degradation reduces wildlife habitat, dismantles trophic interactions, and negatively impacts grassland communities (Matson et al. 1997, Askins et al. 2007, Booms et al. 2014). To restore wildlife in grasslands, grasslands must be returned to their natural historic state (Askins et al. 2007). Terrestrial biosolids applications used to increase forage for cattle can positively affect soils and vegetation (Newman et al. 2014, (Avery et al. 2018, 2019), potentially having restoration effects on higher trophic levels as well. In this thesis, I looked at how the application of biosolids impacted deermice, voles, and three species of raptors and owls at the OK Ranch, a grazed grassland in southern interior British Columbia. Previous work at the OK Ranch showed that biosolids enhanced Dusky Grouse habitat (Lawson 2018), led to higher densities of grasshoppers (Gaudreault et al. 2019), and pastures with biosolids were disproportionately used by American Kestrels (Buers et al. 2019) and Short-eared Owls (Ormrod et al. in review). My work expanded upon this previous research by looking at the small mammal prey base to determine how vole and deermouse populations respond to biosolids applications. I also studied predators at higher trophic levels by examining the abundance, habitat use, and diet of Short-eared Owls, Long-eared Owls, and American Kestrels to determine how changes in small mammal density and biosolids application affected these three species. In Chapter 2, I found smaller deermouse populations on biosolids pastures in comparison to control pastures. Deermouse populations responded negatively to increased grass cover in the biosolids pastures, where grass cover was significantly higher than in the control pastures. Body mass and the percentage of reproductive deermice was similar on all treatments, although we did not test for food quality, food availability, or energy costs of foraging, all of which affect both 63 reproductive status and body mass (Perrigo 1987, Millar et al. 1991, Arsenault and Owen‐Smith 2002, Smith et al. 2015). Deermice prefer sparse degraded areas (Stone 2007) and deermouse populations decline while vole populations increase once an area enters a phase of restoration (Mulligan et al. 2013). My finding that deermice populations are smaller on biosolids pastures than on control pastures indicates that biosolids may be restoring the landscape. In Chapter 3, I examined the yearly abundance of raptors and owls to see if the abundance of generalist species remained stable across all years, while the abundance of nomadic species fluctuated with the vole cycle. Long-eared Owls were detected only in 2017, during the vole population peak. Short-eared Owls were most abundant in 2017 and less abundant in 2018 and 2019 when vole populations had declined. These results are consistent with previous studies that found Short-eared and Long-eared Owls are abundant when voles are numerous, and nearly absent when voles are scarce (Poulin et al. 2001, Houston 2005). In contrast, generalist American Kestrels were abundant and nesting at the ranch across all years, although they also declined after 2017. This decline was likely due to logging at the OK Ranch that caused a loss of nesting habitat, as this generalist species does not respond numerically to changes in prey density (Sarasola et al. 2003, Steenhof and Peterson 2009). However, predation on American Kestrels by other avian predators also may have contributed to this decline (O’Donoghue et al. 1995, Farmer et al 2006). I also looked at how biosolids influenced habitat use in three raptor and owl species. American Kestrels selected for pastures with biosolids application in all four years of study. Short-eared Owls selected for biosolids pastures in 2017, when they were abundant on the ranch. Similarly, all four Short-eared Owl nests located during this study were on biosolids pastures. American Kestrels may be selecting for areas that have biosolids applied due to increased 64 densities of grasshoppers (Gaudreault et al. 2019) and similarly the Short-eared Owl may be selecting for higher densities of voles in these areas (Buers et al. 2019). Long-eared Owls were not detected on the ranch enough to determine associations with biosolids, however, the Long-eared Owl nests on the Ranch in 2017 were surrounded by pastures with high proportions of biosolids applications, indicating Long-eared Owls may also preferentially forage in these areas (Ormrod et al. in review). Finally, I examined the diet of three focal species and found that American Kestrels ate more voles during the vole population peak in 2017 than in 2018 when they mostly ate insects after the vole population declined. In contrast, voles made up the largest part of Short-eared Owl diets in all four years of study. I also compared the diet of nesting Short-eared Owls in 2019 to rodent density to see how a vole specialist reacts when voles are scarce, but deermice are abundant. I found that while voles still made up over 60% of Short-eared Owl diets, nearly 30% of their diet was deermice, which made up <3% of the diet in all other years. This apparent prey “switch” from a specialist species demonstrates that Short-eared Owls may have needed to adapt more generalist traits in 2019 when the owls successfully bred. In conclusion, I have shown that deermice populations are higher on control sites with lower amounts of cover, indicating that restoration may be occurring on biosolids pastures. The diet and abundance of Short-eared Owls, Long-eared Owls, and American Kestrels were in line with previous research (Poulin et al. 2001, Houston 2005, Santillán et al. 2009, Steenhof and Peterson 2009, Booms et al. 2014), although in 2019 Short-eared Owls had a more varied diet than is typical in other regions of North America (Errington 1937, Clark 1975, Colvin and Spaulding 1983, Wiebe 1991, Holt 1993, Wiggins et al. 2020). American Kestrels and Short-eared Owls preferentially selected for biosolids pastures, potentially because there is more prey 65 in these areas. Future research should focus on additional small mammal trapping to get a better picture of how voles respond to biosolids and to determine if vole populations are larger on biosolids applications. Voles are important because many predators rely on them as a food source (Sullivan and Sullivan 2018) and could benefit if vole populations increased. Follow up research on raptors and owls should examine why Short-eared Owls choose to nest on biosolids pastures. More dedicated research is also needed into where Long-eared Owls forage when nesting near biosolids applications. This research is important because wastewater facilities in Canada produce 660,000 dry tonnes of biosolids per year (Canadian Council of Ministers of the Environment, 2012). Land application in the form of fertilizer is considered the most environmentally friendly way of disposing of biosolids (Singh and Agrawal 2008, Haynes et al. 2009, Kominko et al. 2017). Land application of biosolids may also play a role in ecological restoration. Restoration is typically defined as restoring an ecosystem to a previous historical state (Gann et al. 2019). Full recovery occurs when all attributes of an ecosystem closely resemble what they were before degradation occurred (Gann et al. 2019). Biosolids can move ecosystems towards restoration by having impacts at multiple trophic levels. However, impacts of climate change have made it more challenging to restore grasslands to historical states (Fuhlendorf et al. 2018, Gann et al. 2019). Climate change can alter precipitation patterns (Kulmatiski and Beard 2013), cause invasive vegetation encroachment (Brown et al 2018), and both expand and reduce ranges for wildlife (Staudinger et al. 2013), leading to an inability to fully restore grassland areas. By having multiple impacts on soil, vegetation and wildlife (Lu et al. 2012, Buers et al. 2019, Gaudreault et al. 2019), biosolids may also assist with rehabilitation, in which the goals are ecosystem functioning, rather than full native recovery (Gann et al. 2019). 66 My results indicate that there is the potential for biosolids to be used in a restorative capacity to conserve wildlife. 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Accessed October 7, 2020. https://www.worldweatheronline.com/lang/en-us/clinton-weather-history/british-columbia/ca.aspx Yáñez, J. L., H. Núñez, R. P. Schlatter, and F. M. Jaksić. 1980. Diet and weight of American Kestrels in Central Chile. Auk 97:629–631. 90 Appendices Appendix A: Small mammal trapping and data analysis A.1: Small mammal trapping sites We conducted small mammal trapping within three different treatments on nine different sites. Each site was slightly different in its location, elevation, and plant species composition. A summary of these differences is in Table A1. Table A.1 Site names and descriptions for each small mammal trapping site at the OK Ranch. Poa was the most dominant grass genus at all sites. Six sites had additional grass genus’ that were also abundant. Label Site Name UTM (Zone 10 U) Elevation (m) Abundant Grass Types 2016 A Swan Lake 564391 5677638 1115 Stipa 2016 B Perkins Point 563581 5676818 1118 Stipa 2016 C Poison West 568259 5682549 1206 Nassella, Calamagrostis 2018 A Gate 6 567915 5680128 1250 Nassella 2018 B Poison Lake 568734 5683393 1140 Stipa 2018 C Corrals 564664 5680113 1202 Nassella Control A Racetrack 567507 5679289 1212 Control B Poison East 569568 5682732 1268 Control C Gate 7 567045 5679517 1252 A.2: Small mammal data analysis Mark-recapture models have been used for many years to determine the density of wildlife populations (Efford and Fewster 2013). In recent years, spatially explicit capture-recapture (SECR) models have become popular for density estimation (Borchers 2012). Spatial models are useful because they include information about the location of the animal relative to the location of the traps when determining capture probabilities and density estimates (Borchers 2012). We used SECR models to determine the density of deermice from mark-recapture 91 trapping at the OK Ranch. However, because our captures were low, we were unable to determine density using SECR models for most of our trap sites. Here, we present density estimates for the sites that we had enough captures to use SECR models to determine density. To estimate density, spatial capture histories were made for each of the nine trapping locations. We used the secr package (Efford 2019) in RStudio (R Core Team 2019) to estimate density per hectare of deermice for each trapping grid per session, for a maximum of nine models each month (Table B1). We used a half-normal detection function with a 100 m buffer for each of the models. We did not include covariates on capture probability (i.e. time, individual, behavior) in any of the models because of the low capture numbers. Lincoln Peterson (LP) density estimates and SECR model density estimates were within 10 deermice/ha of each other for most of the treatments per month. Estimates on control sites were the most notably different, with three of the LP estimates being much higher than SECR estimates. However, these higher estimates using LP still fell within the confidence intervals of the SECR estimate. 2018A in July was the only major outlier, with the SECR model estimating a density of 3 mice/ha, while LP estimated the density at 26 mice/ha. However, the LP estimate has extremely wide confidence intervals, which brings the density down to include the lower SECR estimate. When comparing both methods, the SECR models and LP estimates are similar, therefore, we believe the Lincoln Peterson estimates presented in text are accurate. 92 Table A.2 Deermice per hectare estimated via spatially explicit capture-recapture and Lincoln Peterson models. Values are estimates and upper and lower confidence intervals. A single dash (-) indicates where captures were too low to use SECR models for density estimates. Double dashes (- -) indicate when no individuals were captured. Captures in May were too low to perform SECR modeling at any of the sites. Treatment June July August SECR Lincoln Peterson SECR Lincoln Peterson SECR Lincoln Peterson 2016 A - - - - 18 (8-40) 23 (9-38) 31(9-101) 24 (6-42) 2016 B - 15 (5-25) 8 (4-18) 12 (9-14) 19 (8-45) 20 (8-33) 2016 C - 2 (2-2) - 8 (2-14) - 10 (0-21) 2018 A - 1 (1-1) 3 (1-14) 26 (0-58) 25 (7-82) 35 (0-70) 2018 B 8 (3-29) 7 (4-11) - 4 (2-7) 6 (2-18) 5 (3-7) 2018 C - - - - - 5 (5-5) - 4 (2-7) Control A - - - - - 17 (0-38) 16 (4-68) 27 (10-45) Control B 11 (6-21) 17 (11-23) 17 (9-32) 35 (14-56) 38 (17-86) 81 (25-136) Control C - 6 (2-10) 29 (15-59) 53 (15-91) 42 (20-88) 36 (18-54) 93 Appendix B: Raptor and Owl Prey and Diet Analysis B.1: Raptor and Owl Prey We gathered information on the average wet body masses for each species consumed by our three focal raptor and owl species. These wet body masses inform us about the percent biomass each species of prey made up in the raptor or owl diet. Table B.1 Average mass of prey used to calculate biomass in raptor and owl diets. Averages for mammals are all from British Columbia. Averages for insects are local as possible (i.e. primarily western North America). We developed a version of this table for species from the same study area in Buers et al. (2019) and Ormrod et al. (in review). Here, we present species eaten by raptors and owls in 2016, 2018, and 2019. Values are wet masses. Species Common Name Body mass (g) Source Mammals Microtus pennsylvanicus Meadow Vole 34.9 Nagorsen 2005 Microtus longicaudus Long-Tailed Vole 43.7 Nagorsen 2005 Microtus montanus Montane Vole 31.6 Nagorsen 2005 Peromyscus maniculatus Deermouse 21.7 Nagorsen 2005 Microtus spp. mean of 4 Microtus 36.9 -- Small mammal mean of Microtus and mice 33.45 -- Birds Pooecetes gramineus Vesper Sparrow 24.0 Jones and Cornely 2002 Insects Anabrus longipes Long-legged Anabrus 3.75 DeFoliart et al. 1982 94 Species Insects Common Name Body mass (g) Source Bruneri brunnea Bruner slant-faced grasshopper 0.376 Larsen et al. 1988-94 Conozoa sulcifronsa Groove-headed grasshopper 0.483 -- Melanoplus bivittatus Two-striped grasshopper 1.086 Bird 1966 Melanoplus sanguinipes Migratory grasshopper 0.313 Fielding 2004 Pseudochorthippus curtipennis Meadow grasshopper 0.19 Rode 2017 Meloe spp. Oil beetle 0.180 Saul-Gershenz 2017 a We could not find mass data for this species. We used the mean from the other grass-eating grasshoppers. 95 B.2: Raptor and Owl Diet Analysis Along with the Relative Frequency of Occurrence, we also calculated the Absolute Frequency of Occurrence and percent wet biomass of each prey item. Absolute frequency of occurrence shows the percent of pellets containing each prey type. Given that American Kestrels typically expel one pellet every 21 hours (Balgooyen 1971), Short-eared Owls expel about one pellet per meal (Duke et al. 1976), and Long-eared Owls expel about 1.4 pellets per day (Marti 1973), this metric can help to examine the daily diet of raptors and owls when pellets are collected frequently. The percent wet biomass of each prey species indicates how much energy the raptor and owl gained from eating that prey item. This metric can help examine how much prey raptors or owls require when energy needs and expenditure are calculated. Table B.2 Diet of American Kestrels in 2018 showing Absolute Frequency of Occurrence and % biomass. Absolute frequency of occurrence is the percent of pellets containing each prey type. Biomass represents the % of diet that species represents by wet weight. We had 47 pellets with 563 individual prey. Two pellets contained a small number of feathers. Absolute Frequency of Occurrence % biomass mammals small mammala 14.9 30.8 insects Pseudochorthippus curtipennis 85.1 5.0 Conozoa sulficrons 74.5 10.2 Anabrus longipes 57.4 43.9 Meloe spp. 93.6 1.0 Melanoplus bivitattus 31.9 5.0 unknown grasshopper 29.8 3.5 96 Absolute Frequency of Occurrence % biomass Bruneri brunnea 4.3 0.1 Anabrus simplex 2.1 0.5 Melanoplus sanguinipes 2.1 0.04 a. These pellets had fur but no diagnostic bones, so we were unable to differentiate remains to species. 97 Table B.3 Diet of the Short-eared Owl showing Absolute Frequency of Occurrence and % biomass. Absolute frequency of occurrence is the percent of pellets containing each prey type. Biomass represents the percent of diet that species represents by wet weight. In 2017, we had 90 pellets with 194 prey, in 2018 we had 11 pellets with 34 prey, and in 2019 we had 24 pellets with 47 prey. In 2019, 79% of pellets were collected from a single nest site. 2017a 2018 2019 Absolute Frequency of Occurrence % biomass Absolute Frequency of Occurrence % biomass Absolute Frequency of Occurrence % biomass mammals Meadow Vole 15.6 7.2 27.3 8.6 0 0 Montane Vole 58.9 27.7 54.5 20.7 37.5 30.3 Long-tailed Vole 23.3 17.5 64.6 32.3 12.5 9.7 Deermouse 1.1 0.3 9.1 1.8 50.0 20.8 Microtus spp. 53.3 45.3 54.4 36.3 50.0 35.4 Sorex spp. 0 0 9.1 0.3 4.2 0.3 birds passerine 4.4 1.8 0 - 8.3 3.5 insects Meloe spp. 0 0 0 - 8.3 0.03 Orthoptera 3.3 0.09 0 - 0 0
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Effects of biosolids on a grassland community of rodents and birds of prey in British Columbia
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