The Effects of Livestock Grazing and Habitat Type on Plant-Pollinator Communities of British Columbia’S Endangered Shrubsteppe

The Effects of Livestock Grazing and Habitat Type on Plant-Pollinator Communities of British Columbia’s Endangered Shrubsteppe

by Sherri L. Elwell B.Sc. (Hons., Biology), University of Victoria, 2007

Thesis Submitted in Partial Fulfillment of the Requirements for the Degree of Master of Science

in the Department of Biological Sciences Faculty of Science

 Sherri L. Elwell 2012 SIMON FRASER UNIVERSITY Spring 2012

All rights reserved. However, in accordance with the Copyright Act of Canada, this work may be reproduced, without authorization, under the conditions for “Fair Dealing.” Therefore, limited reproduction of this work for the purposes of private study, research, criticism, review and news reporting is likely to be in accordance with the law, particularly if cited appropriately.

Approval

Name: Sherri L. Elwell Degree: Master of Science (Biological Sciences) Title of Thesis: The effects of livestock grazing and habitat type on plant-pollinator communities of British Columbia’s endangered shrubsteppe

Examining Committee: Chair: Bernard D. Roitberg, Professor

Elizabeth Elle Senior Supervisor Associate Professor

David J. Green Supervisor Associate Professor

Jonathan W. Moore Internal Examiner Assistant Professor

Date Defended/Approved: April 16, 2012

Partial Copyright Licence

iii

Abstract

Understanding how anthropogenic disturbances affect plant-pollinator communities is important for their conservation. I investigated how plant-pollinator communities of British Columbia’s endangered shrubsteppe are affected by spring livestock grazing. I surveyed vegetation structure and abundance and diversity of flowering plants and pollinators in four paired grazed/ungrazed sites. Grazing increased percent cover of shrubs and bare soil and decreased grass and forb height. However, flowering plant and pollinator abundance, richness and community composition were unaffected by grazing. Instead, floral and pollinator community composition differed between antelope-brush and big sagebrush habitats. I also compared plant-pollinator interaction network structure between habitats, and found that generalization was greater in big sagebrush than the more endangered antelope-brush habitat. Late- flowering-season networks were more asymmetric and had greater plant generalization. These results suggest differences in network resilience to disturbance between habitats and across the flowering season, and so could be used to inform conservation planning in the region.

Keywords: Biodiversity; pollinator; livestock grazing; community composition; interaction network; shrubsteppe

Acknowledgements

I owe my deepest gratitude to my supervisor, Elizabeth Elle, for her mentorship, support, and encouragement throughout all aspects of my degree. I am also thankful to her for igniting in me what I know will be a lifelong love of bees. I extend a special thank you to David Green for his advice and insight into my research and to Jonathan Moore for his role as public examiner.

This research would not have been possible without the assistance of many wonderful people, to whom I am grateful. I thank Jane Pendray and Taylor Holland for their friendship and wonderful assistance in the field. I also extend my sincere thank you to those who helped me with pollinator identification: Elizabeth Elle, Lisa Neame, Terry Griswold and associates at the USDA Bee Biology and Systematics Lab, Jason Gibbs from Cornell University, and Cory Sheffield from York University. Additionally, I thank The Nature Trust of B.C., B.C. Parks, B.C. Ministry of Forest and Range, Canadian Wildlife Service, and Wade Clifton of the Clifton Ranch for allowing me to conduct my research on their properties. I am also thankful to Anne Skinner, from the B.C. Ministry of Forest and Range, for her help with grazing regime information.

I was truly fortunate to have shared my time in the Elle lab with some amazing labmates. I thank Grahame Gielens, Lisa Neame, Lindsey Button and Julie Wray for being wonderful colleagues and friends and for providing me with plenty of laughs, support and encouragement.

Finally, I wish to extend a heartfelt thank you to my family and Jordy Thomson for their loving support throughout this degree. In particular, I thank my parents, Kathy and Tom Elwell, for showing me the meaning of dedication to both family and profession, and for their friendship, love and unwavering support of my educational goals.

Table of Contents

Approval ...... ii Partial Copyright Licence ...... iii Abstract ...... iv Acknowledgements ...... v Table of Contents ...... vi List of Tables ...... viii List of Figures...... ix

Chapter 1 General introduction ...... 1 References ...... 5

Chapter 2 Shrubsteppe plant and pollinator communities influenced more by habitat type than by livestock grazing ...... 8 Introduction ...... 8 Methods ...... 11 Study area ...... 11 Study sites ...... 11 Vegetation ...... 12 Vegetation structure ...... 12 Flowering plant diversity ...... 12 Pollinator diversity ...... 13 Statistical analysis ...... 14 Vegetation structure ...... 14 Abundance, richness and diversity ...... 14 Community composition ...... 16 Results ...... 17 Vegetation structure ...... 17 Abundance, richness and diversity ...... 18 Flowering plants ...... 18 Pollinators ...... 18 Community composition ...... 19 Flowering plants ...... 19 Pollinators ...... 20 Discussion ...... 21 The effects of livestock grazing ...... 21 Vegetation structure ...... 21 Flowering plants ...... 22 Pollinators ...... 23 The effects of shrubsteppe type...... 25 Management implications ...... 26 Conclusions ...... 27 References ...... 28 Tables ...... 34 Figures ...... 37

Chapter 3 A comparison of plant-pollinator network structure between British Columbia’s endangered shrubsteppe habitats ...... 47 Introduction ...... 47 Methods ...... 51 Study sites ...... 51 Sampling plant-pollinator interactions ...... 52 Quantifying plant-pollinator network structure ...... 53 Statistical analysis ...... 56 Results ...... 56 Discussion ...... 58 Habitat and temporal influences on network structure ...... 58 Network size and generalization ...... 58 Asymmetry ...... 60 Nestedness ...... 61 Caveats to the current network approach ...... 62 Practical implications ...... 63 Conclusions and future directions ...... 64 References ...... 66 Tables ...... 72 Figures ...... 76

Chapter 4 General conclusions ...... 78 The effects of livestock grazing and habitat type on flowering plants and pollinators ...... 78 The plant-pollinator network structure of British Columbia’s endangered shrubsteppe ...... 80 Summary and future directions ...... 82 References ...... 85

Appendices ...... 88 Appendix A Floral unit designations ...... 89 Appendix B Species degree and asymmetry ...... 92 Appendix C Most abundant pollinators and floral resources ...... 104 Appendix D Formulas for network structural properties ...... 105 Appendix E Network structural property values ...... 109

vii

List of Tables

Table 2.1. Site characteristics of focal shrubsteppe sites in the Southern Okanagan Valley, British Columbia. The first two letters of the site abbreviation designate a grazed and ungrazed pair. AUM refers to Animal Unit Month, where 1 AUM is equivalent to the forage removed by one 454 kg cow grazing for one month (Gayton 2003), and 1 AUM/ha is considered sufficient to maintain dry bunchgrass habitat in good range condition (McLean and Marchland 1968)...... 34

Table 2.2. Summary of the total abundance, richness and diversity of pollinator- attractive flowering plants and flower visitors (hereafter pollinators) for eight shrubsteppe study sites in the southern Okanagan, British Columbia...... 35

Table 2.3. The effects of livestock grazing and sample episode on the abundance, richness and diversity of all pollinators, and on pollinator functional groups defined by nesting location or taxonomic (and so resource-based) affiliations. GLMMs were used to investigate grazing impacts on pollinator abundance and actual species richness, while mixed models were used to investigate impacts on pollinator diversity. Simpson’s index of diversity was arcsine square-root transformed for analysis...... 36

Table 3.1. Plant-pollinator interaction network property definitions with brief explanations of their influence on network resilience...... 72

Table 3.2. Characteristics of focal shrubsteppe sites in the southern Okanagan Valley, British Columbia. “U” in the site abbreviation denotes ungrazed and “G” denotes grazed. For more information see Table 2.1...... 73

Table 3.3. The effects of habitat type and period of the flowering season (early, mid, late) on plant-pollinator interaction network structural properties. The effects of habitat on network structure were also generated using full season networks. Bolded values = P < 0.10, * = P < 0.05...... 74

Table 3.4. The identity if the top-10 most functionally important plants and pollinators in antelope-brush and big sagebrush shrubsteppe. Species presented have the highest combined degree and asymmetry...... 75

viii

List of Figures

Figure 2.1. Map of study area in the Southern Okanagan Valley, B.C. The four paired sample sites (grazed and ungrazed) are denoted by different coloured symbols. The WL and SO pairs are located in big sagebrush shrubsteppe, while the OK and HL pairs are in antelope- brush shrubsteppe...... 37

Figure 2.2. Sample-based rarefaction curves, rescaled to individuals, for pollinator species richness in all eight sample sites, paired on the basis of similar environmental characteristics except for the presence of grazing livestock...... 38

Figure 2.3. The number of individuals caught in pan-trap surveys for all grazed and ungrazed sites. The number above each bar represents the taxonomically distinguished groups: species for bees [mining bees (Andrenidae); honeybee (Apis mellifera); bumblebees, digger bees, small carpenter bees (Apidae); plasterer bees (Colletidae); sweat bees (Halictidae); mason bees and leaf cutter bees (Megachilidae)], Syrphid flies and Bombyliid flies; morphospecies for beetles, butterflies, moths and wasps...... 39

Figure 2.4. The effects of livestock grazing on the percent cover of vegetation and ground layers and maximum height of grasses and forbs. Note the different scale for percent cover and height variables. Significant effects are indicated by an asterisk: P < 0.01...... 40

Figure 2.5. Least square means of the natural logarithm of flowering plant abundance and richness between grazed and ungrazed sites, over eight sampling episodes. Floral richness is based on Chao2 richness estimates. Flowering plants were surveyed every two weeks from late March until late July...... 41

Figure 2.6. Least square means of the natural logarithm of flowering plant abundance and richness between antelope-brush and big sagebrush sites, over eight sampling episodes. Floral richness is based on Chao2 richness estimates. Flowering plants were surveyed every two weeks from late March until late July...... 42

Figure 2.7. Least square means of the natural logarithm of total pollinator abundance and richness between grazed and ungrazed sites, over eight sampling episodes. Pollinators were sampled every two weeks from late March until late July...... 43

Figure 2.8. Least square means of the natural logarithm of total pollinator abundance and richness between antelope-brush and big sagebrush sites, over eight sampling episodes. Pollinators were surveyed every two weeks from late March until late July...... 44

Figure 2.9. NMDS of sites in flowering plant species space, over eight sampling episodes. The site management regime (grazed vs. ungrazed) is coded with open and filled symbols, shrubsteppe type is coded by symbol shape and sampling date is coded by colour. The axes are labelled with the traits of floral species that are significantly correlated with the NMDS output...... 45

Figure 2.10. NMDS of sites in pollinator species space, over eight sampling episodes. The site management regime (grazed vs. ungrazed) is coded with open and filled symbols, shrubsteppe type is coded by symbol shape and the sampling date is coded by colour. Pollinators associated with axes were significantly correlated with the NMDS output...... 46

Figure 3.1. Quantitative plant-pollinator interaction networks from antelope-brush and big sagebrush habitats: a/e) Full season networks; b/f) Early season networks; c/g) Middle season networks; and d/h) Late season networks. In each network, rectangles represent pollinator (top row) or plant (bottom row) species, and the lines connecting them represent interactions. The width of each plant rectangle represents how frequently the plant was visited by pollinators, and the width of each pollinator rectangle indicates how frequently a pollinator was collected off of flowering plants. The width of the interaction represents how frequently that interaction was recorded. Pollinators are colour-coded as follows: red = bees (Hymenoptera); green = wasps (Hymenoptera); blue = flies (Diptera); purple = beetles (Coleoptera); yellow = butterflies (Lepidoptera); orange = hummingbird (Trochilidae). Plants in the seasonal sub-networks are colour-coded as follows: light grey = blooming in early and mid season; dark grey = blooming in mid and late season; black = blooming during a single season. Species blooming through two seasons are arranged in the same order to allow comparison. Networks are meant to give an impression of how network interaction change through time, and are not all drawn to the same scale...... 76

Figure 3.2. Changes in plant-pollinator network structural properties across early, mid. and late flowering seasons, including full season values, in antelope-brush and big sagebrush shrubsteppe: a) network size, b) number of plant and pollinator species, c) H2’ specialization index, d) plant and pollinator generality, e) interaction strength asymmetry, f) NODF nestedness. The solid lines connect the least square mean values of each metric across the flowering season for both shrubsteppe types...... 77

Chapter 1

General introduction

Pollinators are an important component of global biodiversity, playing a vital role in maintaining natural ecosystems and agricultural productivity (Kearns et al. 1998; Potts et al. 2010). It is estimated that 87.5% of the world’s flowering plant species, require animal pollinators, primarily insects, for sexual reproduction (Ollerton et al. 2011). Thus in natural ecosystems, pollinator declines could lead to a decrease in pollination service to pollinator-dependent plants which in turn could result in plant population declines (Kearns and Inouye 1997). Such parallel declines, between pollinators and pollinator- dependent plants, have already been reported in the Netherlands and United Kingdom (Biesmeijer et al. 2006). The pollinator declines now reported in many regions of the world are thus raising concern over the health of pollinator populations and the preservation of their functional roles (Kearns et al. 1998; Potts et al. 2010). These reports have emphasized the need to understand how anthropogenic disturbances affect pollinator populations and highlight the importance of their consideration in conservation planning and protection efforts.

Anthropogenic disturbances play a major role in influencing biodiversity patterns worldwide (Dornelas et al. 2011). At present, habitat loss, which is the most commonly studied anthropogenic threat to pollinators, appears to be the most important factor influencing pollinator populations (Winfree et al. 2009; Potts et al. 2010). Habitat altering disturbances, such as fire, logging and livestock grazing, on the other hand, were found not to have an overall significant effect on pollinator communities in a meta-analysis by Winfree et al. (2009), but studies assessing the effects of these disturbances on pollinators are still few. It is apparent in the current literature that the response of pollinator communities to anthropogenic disturbance is quite variable (e.g., Erhardt 1985; Cane et al. 2006; Vulliamy et al. 2006; Winfree et al. 2007), emphasizing that additional

studies are needed to gain a better understanding of how various disturbances affect pollinator communities (Potts et al. 2010; Winfree 2010).

Livestock grazing is one of the most prevalent grassland and shrubsteppe disturbances (Fleischner 1994). Grazing has been shown to directly influence vegetation structure and community composition, soil compactness and nutrient cycling, while indirectly affecting populations of mammals, birds and amphibians (Fleischner 1994; Jones 2000). Although pollinators make a substantial contribution to grassland biodiversity and are important for grassland functioning (Wilson 1987; Gilgert and Vaughan 2011), they have been given less attention in grazing impact studies (Debano 2006; Yoshihara et al. 2008). Furthermore, the studies that have been conducted report a range of pollinator responses to grazing, both positive (Carvell 2002; Vulliamy et al. 2006) and negative (Soderstrom et al. 2001; Debano 2006; Hatfield and LeBuhn 2007; Xie et al. 2008). One commonality found among previous studies is that pollinator communities tend to respond in concert with plant communities, because of their need for floral resources for food and nest provision. The majority of studies assessing the impacts of grazing on pollinator populations come from Europe, with only a few studies previously conducted in North American grasslands, most of which focus exclusively on bees (Sugden 1985; Debano 2006; Hatfield and LeBuhn 2007; Kearns and Oliveras 2009; Kimoto 2010). Continuing to develop an understanding of how all pollinating insects and the flowering plants they interact with respond to grazing pressure will be important for their conservation, particularly as grasslands are among North America’s most threatened ecosystems (Curtin and Western 2008; Peart 2008).

Long term data sets on pollinator populations, particularly solitary native bees and pollinating flies, are fragmentary at best (Potts et al. 2010; Winfree 2010). Most studies to date have relied on making inferences about pollinator communities by comparing pollinator richness, abundance and diversity along gradients of disturbance as a surrogate for change over time (e.g., Kruess and Tscharntke 2002; Cane et al. 2006; Vulliamy et al. 2006; Winfree et al. 2007). Furthermore, it has been argued that quantifying species composition, in addition to diversity, is important for understanding disturbance impacts on pollinators. Studies have shown that even when overall bee abundance and species richness are not negatively affected by disturbance; there can be significant changes in species composition (Cane et al. 2006; Winfree et al. 2007;

Brosi et al. 2008). Thus, multivariate analyses are important statistical tools for determining community-level disturbances, such as livestock grazing. Additionally, over the last decade the study of plant-pollinator interaction networks, a community-based analytical approach, has provided another means of quantifying the structural and functional dynamics of communities (Bascompte and Jordano 2007; Bascompte 2009; Vazquez et al. 2009). Plant-pollinator network analysis can identify which species interact within a community and how those interactions collectively influence community structure (Bascompte and Jordano 2007). It has been found that plant-pollinator networks have particular network-level structural properties that have consequences for community stability and resilience (Memmott et al. 2004; Bascompte and Jordano 2007). Combining multivariate and network approaches when and where possible is likely to provide a more comprehensive view of pollinator communities, with more power to inform conservation planning and management.

Within Canada, the shrubsteppe habitats of the south Okanagan basin, British Columbia, are recognized as some of the most biologically diverse as well as endangered ecosystems in the country. Antelope-brush shrubsteppe in particular, supports a disproportionately high percentage of Canada’s endangered and threatened species and is considered in the top four most endangered ecosystems in the country (Schlute et al. 1995; Dyer and Lea 2003). Over the last century the Okanagan basin has lost 68% of its antelope-brush shrubsteppe and 33% of its big sagebrush shrubsteppe to agricultural and urban development. Much of what remains is grazed by livestock (Lea 2008). The pollinator communities of these habitats are predicted to be very diverse (L. Packer, bee taxonomist, York University, pers. comm.), but have not been extensively inventoried and studies assessing the impacts of anthropogenic disturbances on pollinator communities are lacking. These pollinators and the flowering plants with which they interact are a vital component of shrubsteppe biodiversity, together providing vegetation structure and forage for many of the other species that inhabit these ecosystems (Gilgert and Vaughan 2011). Thus, understanding how plant and pollinator communities are structured and how they are affected by disturbance will be important for effective management and conservation.

In this thesis, I examine the influence of livestock grazing on plant and pollinator communities in British Columbia’s Okanagan Valley and use network analysis to assess

the structural properties of plant-pollinator communities in antelope-brush and big sagebrush habitats. In Chapter 2, I investigate the effects of livestock grazing on floral and pollinator abundance, richness and community composition. I also assess the impacts of grazing on habitat structure, as habitat features other than floral resources, such as vegetation structure and bare soil availability, can impact pollinator populations. Additionally, as two shrubsteppe types, antelope-brush and big sagebrush, were sampled, I investigate whether habitat type influenced the abundance, richness or community composition of flowering plants and pollinators. In Chapter 3, I investigate differences in plant-pollinator network structure between antelope-brush and big sagebrush shrubsteppe that may have consequences for community resilience to disturbance, and assess which plant and pollinator species are functionally important in each habitat. I also examine temporal variability in network structure to investigate how these plant-pollinator networks, and sensitivity to disturbance, change over the course of the flowering season. This thesis contributes to our understanding of the plant-pollinator communities of B.C’s endangered shrubsteppe. Additionally, in a broader context, it contributes to a growing body of research examining how habitat alteration influences pollinator communities, and illustrates how plant-pollinator networks could be useful in elucidating practical implications for conservation planning.

References

Bascompte, J. (2009). Mutualistic networks. Frontiers in Ecology and the Environment 7(8): 429-436.

Bascompte, J. and P. Jordano (2007). Plant-animal mutualistic networks: The architecture of biodiversity. Annual Review of Ecology Evolution and Systematics. Palo Alto, Annual Reviews. 38: 567-593.

Biesmeijer, J. C., S. P. M. Roberts, M. Reemer, R. Ohlemuller, M. Edwards, T. Peeters, A. P. Schaffers, S. G. Potts, R. Kleukers, C. D. Thomas, J. Settele and W. E. Kunin (2006). Parallel declines in pollinators and insect-pollinated plants in Britain and the Netherlands. Science 313(5785): 351-354.

Brosi, B. J., G. C. Daily, T. M. Shih, F. Oviedo and G. Duran (2008). The effects of forest fragmentation on bee communities in tropical countryside. Journal of Applied Ecology 45(3): 773-783.

Cane, J. H., R. L. Minckley, L. J. Kervin, T. H. Roulston and N. M. Williams (2006). Complex responses within a desert bee guild (Hymenoptera : Apiformes) to urban habitat fragmentation. Ecological Applications 16(2): 632-644.

Carvell, C. (2002). Habitat use and conservation of bumblebees (Bombus spp.) under different grassland management regimes. Biological Conservation 103(1): 33-49.

Curtin, C. and D. Western (2008). Grasslands, people, and conservation: Over-the- horizon learning exchanges between African and American pastoralists. Conservation Biology 22(4): 8.

Debano, S. J. (2006). Effects of livestock grazing on aboveground insect communities in semi-arid grasslands of southeastern Arizona. Biodiversity and Conservation 15(8): 2547-2564.

Dornelas, M., C. U. Soykan and K. I. Ugland (2011). Biodiversity and disturbance. Biological Diversity: frontiers in measurment and assessment. A. E. Magurran and B. J. McGill. New York, Oxford University Press: 237-251.

Dyer, O. and E. C. Lea (2003). Status and importance of the Antelope-brush - Needle- and-thread grass plant community in the South Okanagan Valley, British Columbia. Ecosystems at Risk - Antelope Brush Restoration Conference. R. Seaton. Osoyoos, B.C.

Erhardt, A. (1985). Diurnal Lepidoptera - senstive indicators of cultivated and abandoned grassland. Journal of Applied Ecology 22(3): 849-861.

Fleischner, T. L. (1994). Ecological costs of livestock grazing in western North America. Conservation Biology 8(3): 629-644.

Gilgert, W. and M. Vaughan (2011). The value of pollinators and pollinator habitat to rangelands: Connections among pollinators, insects, plant communities, fish, and wildlife. Rangelands 33(3): 6.

Hatfield, R. G. and G. LeBuhn (2007). Patch and landscape factors shape community assemblage of bumble bees, Bombus spp. (Hymenoptera : Apidae), in montane meadows. Biological Conservation 139(1-2): 150-158.

Jones, A. (2000). Effects of cattle grazing on North American arid ecosystems: A quantitative review. Western North American Naturalist 60(2): 155-164.

Kearns, C. A. and D. S. Inouye (1997). Pollinators, flowering plants, and conservation biology - Much remains to be learned about pollinators and plants. Bioscience 47(5): 297-307.

Kearns, C. A., D. W. Inouye and N. M. Waser (1998). Endangered mutualisms: The conservation of plant-pollinator interactions. Annual Review of Ecology and Systematics 29: 83-112.

Kearns, C. A. and D. M. Oliveras (2009). Environmental factors affecting bee diversity in urban and remote grassland plots in Boulder, Colorado. Journal of Insect Conservation 13(6): 655-665.

Kimoto, C. (2010). Effect of livestock grazing on native bees in a pacific northwest bunchgrass prairie. Corvallis Oregon State University. Master of Science in Wildlife Science: 135.

Kruess, A. and T. Tscharntke (2002). Grazing intensity and the diversity of grasshoppers, butterflies, and trap-nesting bees and wasps. Conservation Biology 16(6): 1570-1580.

Lea, T. (2008). Historial (pre-settlement) ecosystems of the Okanagan Valley and Lower Similkameen Valley of British Columbia - pre-European contact to the present. Davidsonia 19(1): 34.

Memmott, J., N. M. Waser and M. V. Price (2004). Tolerance of pollination networks to species extinctions. Proceedings of the Royal Society of London Series B- Biological Sciences 271(1557): 2605-2611.

Ollerton, J., R. Winfree and S. Tarrant (2011). How many flowering plants are pollinated by animals? Oikos 120(3): 321-326.

Peart, B. (2008). A record of The World Temperate Grasslands Conservation Initiative Workshop Hohhot, China - June 28 & 29, 2008. Vancouver, BC, Temperate Grasslands Conservation Initiative.

Potts, S. G., J. C. Biesmeijer, C. Kremen, P. Neumann, O. Schweiger and W. E. Kunin (2010). Global pollinator declines: trends, impacts and drivers. Trends in Ecology & Evolution 25(6): 345-353.

Schlute, A., T. Lea, S. Cannings and P. G. Krannitz (1995). Antelope-brush ecosystems. L. a. P. B.C. Ministry of Environment, Wildlife Branch. Victoria, B.C.: 6.

Soderstrom, B., B. Svensson, K. Vessby and A. Glimskar (2001). Plants, insects and birds in semi-natural pastures in relation to local habitat and landscape factors. Biodiversity and Conservation 10(11): 1839-1863.

Sugden, E. A. (1985). Pollinators of Astragalus monoensis Barneby (Fabaceae) - New host records - Potential impact of sheep grazing. Great Basin Naturalist 45(2): 299-312.

Vazquez, D. P., N. Bluethgen, L. Cagnolo and N. P. Chacoff (2009). Uniting pattern and process in plant-animal mutualistic networks: a review. Annals of Botany 103(9): 1445-1457.

Vulliamy, B., S. G. Potts and P. G. Willmer (2006). The effects of cattle grazing on plant- pollinator communities in a fragmented Mediterranean landscape. Oikos 114(3): 529-543.

Wilson, E. O. (1987). The little things that run the world (the importance and conservation of invertebrates). Conservation Biology 1: 2.

Winfree, R. (2010). The conservation and restoration of wild bees. Year in Ecology and Conservation Biology 2010. R. S. Ostfeld and W. H. Schlesinger. Malden, Wiley- Blackwell. 1195: 169-197.

Winfree, R., R. Aguilar, D. P. Vazquez, G. LeBuhn and M. A. Aizen (2009). A meta- analysis of bees' responses to anthropogenic disturbance. Ecology 90(8): 2068- 2076.

Winfree, R., T. Griswold and C. Kremen (2007). Effect of human disturbance on bee communities in a forested ecosystem. Conservation Biology 21(1): 213-223.

Xie, Z. H., P. H. Williams and Y. Tang (2008). The effect of grazing on bumblebees in the high rangelands of the Eastern Tibetan Plateau of Sichuan. Journal of Insect Conservation 12(6): 695-703.

Yoshihara, Y., B. Chimeddorj, B. Buuveibaatar, B. Lhaquasuren and S. Takatsuki (2008). Effects of livestock grazing on pollination on a steppe in eastern Mongolia. Biological Conservation 141(9): 2376-2386.

Chapter 2

Shrubsteppe plant and pollinator communities influenced more by habitat type than by livestock grazing

Introduction

In North America, grasslands and shrubsteppe are among the continent’s most species-rich and threatened ecosystems. Thus, the continued fragmentation and degradation of grassland ecosystems due to agricultural and urban development is an increasing cause for concern (Curtin and Western 2008; Peart 2008). One of the continent’s most prevalent grassland disturbances is livestock grazing (Fleischner 1994). Consequently, the ecological impacts of livestock grazing in grassland and shrubsteppe ecosystems, particularly the impacts on the vegetative community, have been extensively studied. It is well known that livestock grazing can alter various habitat features including vegetation community structure and composition, soil compactness, bare soil abundance, nutrient cycling and microhabitat temperature and humidity (see Fleischner 1994; Jones 2000 for reviews). Many studies have also indicated that grazers indirectly impact other grassland organisms, such as birds, mammals and amphibians, through structural changes in habitat caused by herbivory and trampling (Fleischner 1994). However, although they constitute a large portion of the animal biomass and have important roles in grassland ecosystem functioning (Wilson 1987; Gilgert and Vaughan 2011), invertebrates, particularly pollinators, have been given less attention in grazing impact studies (Debano 2006; Yoshihara et al. 2008).

Pollination is a vital ecosystem service (Kearns et al. 1998) and deserves thorough consideration in terrestrial ecosystem disturbance studies. Pollinators, of which bees are the primary group, are required for the successful reproduction of an

estimated 87.5% of the world’s flowering plant species (Ollerton et al. 2011). Thus, aside from their immense importance in crop pollination (Klein et al. 2007), pollinator communities are also critically important for the maintenance of natural ecosystems (Kearns et al. 1998). The decline of pollinators and consequent disruption of pollination systems now being reported in many regions of the world (Kearns et al. 1998; Biesmeijer et al. 2006; Potts et al. 2010) emphasize the need to understand how anthropogenic disturbances affect pollinator populations.

Collectively, a range of pollinator responses to grazing have been documented, both positive (Carvell 2002; Vulliamy et al. 2006) and negative (Soderstrom et al. 2001; Kruess and Tscharntke 2002; Debano 2006; Hatfield and LeBuhn 2007; Xie et al. 2008). One commonality found among previous studies is that pollinator communities tend to respond in concert with plant communities, because of their need for floral resources for food and nest provision. Thus, studies of grasslands that depend on a frequent disturbance regime to maintain floral diversity indicate grazing can be beneficial to pollinator communities (e.g. Carvell 2002; Vulliamy et al. 2006), whereas studies of grasslands without an abundance of disturbance-adapted plants, or those under heavy grazing, suggest grazing can negatively impact pollinator communities (e.g. Kruess and Tscharntke 2002; Debano 2006; Xie et al. 2008). Additionally, the response of pollinators to grazing can be affected by impacts on nest sites, with increased availability and compaction of bare soil in areas with historically high grazing tending to increase ground nesting bees (Vulliamy et al. 2006). Collectively, these studies indicate that the impacts of grazing on floral and pollinator communities are not universal and depend on a host of factors, including the type of grazers (e.g. cattle, sheep) and the historical disturbance and current grazing regimes.

The majority of studies assessing the impacts of livestock grazing on pollinator populations have been conducted in Europe, with only a few studies previously completed in North America, all of which were situated in the United States (Sugden 1985; Debano 2006; Hatfield and LeBuhn 2007; Kearns and Oliveras 2009; Kimoto 2010). All but one of these studies focused exclusively on bees. Although bees may be the most important group of pollinators, grasslands and shrubsteppe also support diverse beetle, fly, butterfly and wasp communities that contribute to the pollination of native plants (Kearns et al. 1998; Harmon et al. 2011). Thus, information on how

grazing impacts whole communities of insect pollinators in North America, particularly in more northern grasslands, is lacking.

In the Okanagan Valley, south-central British Columbia, shrubsteppe ecosystems support numerous rare and endangered species, encompass Canada’s only temperate desert (Schluter et al. 1995; Seaton 2003; Wikeem and Wikeem 2004), and due to the hot and dry climate are likely to have a high diversity of bees (O'Toole and Raw 1999; Michener 2000). The pollinator community has not been extensively inventoried and studies assessing the impacts of anthropogenic effects on pollinator communities are lacking. Over the last century the Okanagan basin has lost 68% of its antelope-brush shrubsteppe and 33% of its big sagebrush shrubsteppe to agricultural and urban development. Much of what remains is in semi-natural condition, largely due to livestock grazing (Lea 2008). Grazing in shrubsteppe ecosystems can alter shrub cover, the composition and distribution of herbaceous species and bare soil abundance (Jones 2000; Krannitz 2008). Therefore, I predicted that grazing would indirectly impact pollinator communities by altering the plant community and ground-nesting site availability. Understanding how livestock grazing affects the floral and pollinator communities of these ecosystems will be important for biodiversity preservation and management.

I surveyed flowering plants and pollinators in grazed and ungrazed shrubsteppe sites over the course of an entire flowering season, March-July. I tested my expectation that flowering plant and pollinator abundance, richness, diversity and community composition would be negatively affected by livestock grazing. I also assessed whether different pollinator functional groups were affected similarly by grazing disturbance. Also, as habitat features other than floral resources, such as vegetation height and bare soil availability, can also affect pollinator populations, I assessed the influence of grazing on shrubsteppe vegetation structure. Finally, because the dominant shrubs species, as well as other plant species, vary with elevation in this ecosystem, I also investigated whether shrubsteppe type influenced the abundance, richness, diversity or community composition of flowering plants or pollinators.

Methods

Study area

The shrubsteppe ecosystems of western North America range from the Great Basin in eastern California and Nevada northward through the Columbia Basin and into south central British Columbia (Mack 1981; Gayton 2003). In B.C., shrubsteppe ecosystems occur primarily in the southern Okanagan and Similkameen Valleys, and in the Thompson River Valley around Kamloops (Mack 1981; Krannitz 2008).

Within the Okanagan Valley, shrubsteppe ecosystems occupy the valley floor, benches and lower slopes, ranging from approximately 250 m to 700 m (Wikeem and Wikeem 2004). At slightly higher elevations, a sparse Ponderosa Pine (Pinus ponderosa) over-story accompanies the shrubsteppe vegetation (Nicholson et al. 1991). Antelope-brush (Purshia tridentata), along with rabbit-brush (Chrysothamnus nauseosus), dominate dry sites with sandy soils, and are replaced by big sagebrush (Artemisia tridentata) as elevation and moisture increase. The understory vegetation is characterized by widely spaced bunchgrasses mixed with a variety of wildflowers and a well-developed cryptogamic crust. This region has been subject to increasing anthropogenic disturbance, with a combination of cattle ranching, commercial orchards, vineyards, and urban development (Lea 2008).

Study sites

I chose eight sites, four grazed and four ungrazed, in the southern Okanagan Valley (Figure 2.1, Table 2.1). Gazed and ungrazed sites were paired for similarity in elevation, slope, aspect, and vegetation to improve the strength of comparison of grazing regime. All sites were a minimum of 20 hectares and were connected to contiguous shrubsteppe, grassland or ponderosa pine forest on at least one side. The average distance between sites within a pair was 4.3 km, with an average between-pair distance of 11 km. As the entire Okanagan Valley was historically grazed (Rick Tucker, BC Ministry of Forests and Range, pers. comm.), even ‘ungrazed’ sites have had cattle grazing at some point in the past; grazing regimes as reported by site managers are in Table 2.1. All grazed sites are spring grazed by cattle for roughly one month between the beginning of April and end of June (Table 2.1), and with the exception of SOG, were

grazed by livestock during the sampling season. At the start of research we were informed that SOG was regularly grazed but subsequently it became clear that accurate records since 2004 were lacking.

Within each site I chose an area that would be suitable for pairing, e.g. with similar environmental variables to the paired site. Within these areas, a point was randomly selected on an aerial photo and used as the starting point for a 100 m permanent transect. A single random orientation was used for all transects. Around each permanent transect a 1-ha sampling plot was delineated, within which I conducted all pollinator and vegetation sampling.

Vegetation

Vegetation structure

Pollinator communities can be influenced by vegetation (floral resources and habitat structure) and bare soil availability (nesting sites), both of which are expected to be impacted by livestock grazing (Vulliamy et al. 2006). Therefore, I measured the maximum height and percent cover of vegetation by layer (shrub, grass and forb), as well as the percent cover of ground layers (bare soil, cryptogamic crust and litter). Sampling was conducted in 60 0.5x1 m quadrats, spaced five meters apart, along four 90 m transects spread evenly across each 1-ha plot. Within quadrats, the percent cover of vegetation and ground layers was estimated to the nearest half percentage. I conducted vegetation structure sampling within the same week for sites within a pair, between June 21st and July 7th, 2010.

Flowering plant diversity

I surveyed flowering plants at each site eight times. The first survey at each site coincided with the beginning of the spring bloom (March 2010), after which surveys were continued approximately every two weeks until the end of the flowering season (July 2010). In each hectare, I sampled 30 0.5x1 m quadrats evenly spaced along 90 m of the central permanent transect. In each quadrat, I counted the number of “floral units”, generally an inflorescence, of each wildflower species present (see Appendix A for floral unit designations by species). I assessed forbs only, as grasses do not normally provide

forage for pollinators, and deleted from analysis any forb species apparently unattractive to most pollinators (pers. obs.; mostly small-flowered species like Draba verna).

Pollinator diversity

I used pan-traps (blue, yellow, and white) to collect flying insects (putative pollinators) eight times over the flowering season, concurrent with flowering plant diversity surveys. Thirty 12 oz pan-traps (10 per color in a regular order) were laid out at 3-m intervals along the central transect. On each sampling date, pan-traps were deployed by 8:30 am and collected starting at 5:00 pm, to keep the sampling time between sites consistent (~8.5hr/date). Paired sites were always sampled on the same day to eliminate potential differences between grazing treatments that were due to other factors such as weather or Julian date. Pan-traps were only deployed on warm, sunny days with low to moderate wind. I stored pan-trap samples in 75% ethanol until the specimens could be dried and pinned for identification.

All species were identified to the lowest taxonomic level possible, with a focus on insects observed to be floral visitors (data from netting surveys where insects were collected directly from flowers; Chapter 3). Bees (Hymenoptera) comprised the majority of specimens and were identified to species except for some genera without revised keys (Evylaeus, Nomada, Sphecodes) which were identified to morphospecies. Hoverflies (Syrphidae) and bee flies (Bombyliidae) were also identified to species, as were thick-headed flies (Conopidae). Tachinids, Sarcophagids, and Calliphorid flies were identified to morphospecies and included in the analysis if they were collected off of flowers in other research (Chapter 3), but other Dipterans were not common flower visitors and so were not included. Finally, beetles (Coleoptera), butterflies and moths (Lepidoptera), and wasps (Hymenoptera) were often identified to genus or morphospecies, and were included in the current analysis if also collected in netting surveys. I considered flower visitors to be putative pollinators, as frequent flower visitors often contribute to plant reproduction (Vazquez et al. 2005; Sahli and Conner 2006). Hereafter, all morphospecies will be referred to as species for the purposes of simplicity (specific morphospecies designations are presented in Appendix B, Table B.2).

Statistical analysis

Vegetation structure

The influence of livestock grazing on vegetation structure, measured as the percent cover of vegetation and ground layers (shrub, forb, grass, bare soil, crust, litter) and maximum height of vegetation layers (forb, grass), was analysed with a multivariate analysis of variance (MANOVA) using the GLM procedure of SAS version 9.2 (SAS institute, 2008). The maximum height of shrubs was excluded from the analysis as over half of the quadrats sampled were without shrubs, and I wanted to retain the information on other measured variables. The model included management (grazed vs. ungrazed) as a fixed effect, transect as a random effect, and pair as random blocking effect. Transect nested within block and management was used as the error term when testing for a main effect of management. Because the overall MANOVA was significant (see results) I subsequently performed univariate analyses of variance (ANOVAs) on each variable using the same model. All percent cover variables were arcsine square-root transformed, which is appropriate for proportion data (Sokal and Rohlf 1995), and height measurements were log transformed to reduce heteroscedasticity.

Abundance, richness and diversity

I computed sample-based rarefaction for the pollinator communities of each site to assess approximately how well sampling captured pollinator species richness. I also calculated Simpson’s index of diversity (1-D) and Chao2 richness estimates for the flowering plant and pollinator communities at all sites, using EstimateS (Colwell 2005). Chao2 derives estimates of the true species richness of a community using the occurrence of rare species within samples, specifically the number of species that occur in just one (uniques) or two (duplicates) samples (Colwell and Coddington 1994). Chao2 is robust to small sample sizes (Colwell and Coddington 1994) and is considered appropriate for invertebrate communities as singletons and doubletons are commonly sampled.

I evaluated the effect of grazing on the species richness and abundance of flowering plants and pollinators using generalized linear mixed models (GLMMs) in SAS (PROC GLIMMIX), that included management and sample episode as fixed effects and pair as a random blocking effect. Since abundance and richness are count data, a

Poisson distribution with log link function was used for all models (Zuur et al. 2007). Raw count data was used for floral unit abundance, pollinator abundance and pollinator richness, while Chao2 richness estimates were used for flowering plant richness. Although pollinator sampling as some sites wasn’t sufficient to produces a clear asymptote on the rarefaction curve (Figure 2.2), pollinator Chao2 richness estimates could not be used in the GLMM as each sample episode had only three sub-samples which was not sufficient for richness estimation. The residuals of initial models indicated overdispersion, therefore the data was re-fitted with quasi-Poisson models (Zuur et al. 2007). Flowering plant and pollinator diversity (1-D) was compared between grazed and ungrazed sites using mixed models in SAS (PROC MIXED). All mixed models included management and sample episode as fixed effects and pair as a random blocking effect, with an autoregressive covariance structure.

I also analyzed the effects of grazing on the abundance, richness and diversity of pollinator functional groups using the same models. Differences in nesting substrate and foraging strategies can be used to define functional groups (e.g., Neame et al. 2012), because they influence how pollinators are affected by anthropogenic disturbances (e.g., Cane et al. 2006; Sjodin et al. 2008; Williams et al. 2010). I categorized specimens into five functional groups based on taxonomy and nesting behaviour: above-ground nesting bees (including the introduced honeybee, Apis mellifera); below-ground nesting bees; beetles; wasps; and other pollinators (flies, butterflies and moths). Cleptoparasitic bees, whose nesting biology is dictated by their hosts, were excluded from the analysis because their response to disturbance is not independent of the response of their host species (Williams et al. 2010). Five species of the bee genus Megachile were also excluded as they are known to nest both above and below ground and could not conclusively be placed in either category.

I also used GLMMs and mixed models in SAS to assess whether shrubsteppe type (antelope-brush vs. big sagebrush) influenced flowering plant and total pollinator abundance, richness and diversity over time. In all GLMMs investigating effects on abundance and richness I specified a quasi-Poisson distribution and log link function. Mixed models investigating effects on diversity included an autoregressive covariance structure. All GLMMs and mixed models included shrubsteppe type and sample episode as fixed effects and site nested within habitat as a random effect.

For all models the degrees of freedom were calculated using the Kenwood Roger method and least square (LS) means were computed for all fixed effects. Flowering plant and pollinator diversity values were arcsine square-root transformed to eliminate heteroscedasticity prior to analyses.

Community composition

To explore whether livestock grazing impacts flowering plant or pollinator community composition over time, I performed non-metric multidimensional scaling (NMDS) of sites in species space using PC-ORD 5 (McCune and Mefford 2006). Prior to running the ordinations, I square-root transformed floral unit and pollinator abundances. The square-root transform is appropriate for community data, as it down- weights the effect of single species and allows species of intermediate abundance to contribute more to the overall assemblage pattern (McCune and Grace 2002). Additionally, I removed all species represented by a single individual prior to the pollinator community ordination to reduce noise (McCune and Grace 2002). All flowering plant species were retained in the floral community ordination, as singletons were few and did not influence the stress of the ordination. Sorensen distance was used to generate the dissimilarity matrix of both ordinations. I determined the appropriate number of dimensions for the ordinations using a step-down procedure from six dimensions, using a maximum of 150 random starting configurations. The scree plot of stress values generated from both ordinations suggested that a final three dimensional solution was best. To facilitate interpretation of the ordinations I calculated correlations between the abundance of floral units and pollinator species and the NMDS solutions using SAS (PROC CORR).

To test for the effects of grazing and shrubsteppe type on pollinator and flowering plant community composition, I used permutation-based multivariate analyses of variance (PerMANOVA). PerMANOVA allows the effects of one or more factors on a whole assemblage of species to be tested simultaneously on the basis of any distance measure, using permutation methods (Anderson 2001). PerMANOVAs were performed in R using the adonis function in the vegan package (R Development Core Team, 2011; Oksanen et al. 2011). Models testing the impact of livestock grazing on plant and pollinator communities incorporated management and sample episode as fixed effects,

blocked by pair. Models testing the impact of habitat type included sample episode and shrubsteppe type as fixed effects. For all tests, Sorensen’s distance measure and 4999 random permutations were used. I also tested for the multivariate homogeneity of group dispersions using the betadisper function, as PerMANOVA is sensitive to differences in the dispersion of points within groups (Anderson 2001).

Finally, to assess whether the pollinator and flowering plant communities of the sites sampled were correlated I ran a Mantel test, based on Mantel’s asymptotic approximation, in PC-ORD (McCune and Grace 2002), using Sorensen’s distance and square-root transformed overall abundance data as before. I retained all species sampled for this analysis.

Results

I collected a total of 6317 putative pollinators, comprising 185 bee species, 25 fly species, 11 beetle species, 17 wasp species, and 18 butterfly and moth species (Figure 2.3; Appendix B, Table B.2). The number of pollinators collected varied between sites, from 514 individuals at SOG to 1227 individuals at OKG (Table 2.2). Pollinator species richness varied considerably less, from 139 species at OKG to 198 species at OKU (Chao2). Bees made up 81% of the total individuals caught, followed by wasps and beetles with 8% and 7%, respectively. Ground nesting bees from the families Halictidae and Andrenidae comprised the majority of bees collected (83%), however, the Megachilidae were the most speciose family represented, with 59 species (Figure 2.3). Many of the species collected were uncommon, with 27% of bee species being represented by only one or two individuals.

I surveyed 54 pollinator-attractive wildflower species across the eight sites. Flowering plant richness varied from 5 to 26 species across sites, while floral units varied over almost an order of magnitude (Table 2.2).

Vegetation structure

The overall structure of shrubsteppe vegetation was affected by livestock grazing

(MANOVA, F8,20=13.73, P<0.0001). Univariate tests indicated that grazing increases the

percent cover of shrubs and bare soil, while decreasing the cover of cryptogamic crust and the maximum height of forbs and grasses (Figure 2.4). The percent cover of the grass, forb, and litter layers were unaffected by grazing.

Abundance, richness and diversity

Flowering plants

Although a trend of decreased flowering plant abundance was observed in grazed sites over the last four sampling episodes (Figure 2.5), models revealed no overall influence of grazing on floral abundance, richness or diversity (GLMMs, abundance: F1,4.4=3.30, P=0.076; richness: F1,2.6=2.62; P=0.217; Figure 2.5; Mixed model, diversity: F1,5.7=0.45, P=0.5266). Floral richness increased after the second sampling episode (F7,41.4=2.62, P=0.0001; Figure 2.5), however floral abundance and diversity were unaffected by flowering season stage (abundance: F7,44.8=1.63, P=0.152,

Figure 2.5; diversity: F7,34=2.05, P=0.0765). There was no effect of the interaction between management and sampling episode for floral abundance, richness or diversity (all significances were P>0.1). Shrubsteppe type, like livestock grazing, did not affect flowering plant abundance, richness or diversity (GLMMs, abundance: F1,6.2=3.18,

P=0.123; richness: F1,5.7=4.42, P=0.083; Mixed model, diversity: F1,6=3.87, P=0.097), although there was a trend for all values to be higher in big sagebrush sites throughout the flowering season (Figure 2.6).

Pollinators

Livestock grazing did not affect the abundance, richness or diversity of the overall pollinator assemblage or any pollinator functional group (Table 2.3; Figure 2.7). However, abundance and richness of all pollinator functional groups, as well as the overall pollinator assemblage, significantly increased after early spring sampling (episodes 1 and 2; Table 2.3; Figure 2.7). Pollinator richness and abundance tended to peak during mid season (episodes 3-5), with an additional peak in abundance at the end of the sampling season (episode 8). Overall pollinator diversity, as well as most pollinator functional groups, followed the same pattern with diversity significantly increasing after early spring sampling. However, below-ground bee diversity was unaffected by period of the flowering season (Table 2.3). There was no interaction

between management and time of sampling for the overall pollinator assemblage or any pollinator functional group (all significances were P>0.1). Habitat type did not influence overall pollinator abundance, richness or diversity (GLMMs, abundance: F1,10.4=0.00,

P=0.996; richness: F1,7.2=1.63, P=0.242; Figure 2.8; Mixed model, diversity: F1,8.5=3.01, P=0.112).

Community composition

Flowering plants

The NMDS plot of sites in species space, across all sample dates, suggested that the most influential factor contributing to floral community composition was time of season, not livestock grazing or habitat type. Grazed and ungrazed sites, as well as sites in antelope-brush and big sagebrush, were distributed similarly along both ordination axes (Figure 2.9). Conversely, sites that were sampled early in the season were grouped separately from sites sampled in the middle and end of the flowering season. Correlations between the abundance of flowering plant species and the NMDS solution contributed to the pattern observed. Early flowering plant species, such as yellow bell (Fritillaria pudica) were significantly positively correlated with axis 2, while late flowering species, such as sagebrush mariposa lily (Calochortus macrocarpus), were negatively correlated (Figure 2.9). Along axis 1, species with strong positive correlations were mid-flowering and tended to be present at only a few sites, while species with strong negative correlations were early or late bloomers with a more ubiquitous distribution. The final stress for the ordination was 16.29 and the final instability was 0.00001, through 147 iterations.

PerMANOVAs confirmed two of the patterns visualized in the NMDS. Flowering plant community composition was unaffected by livestock grazing (pseudo-F1,3= 1.29, P= 0.116), but did significantly change over the course of the flowering season (pseudo-

F7,57= 7.65, P= 0.0002). Although not suggested by the NMDS plot, floral community composition also differed between antelope-brush and big sagebrush habitats (pseudo-

F1,7= 3.03, P= 0.0002). There were no differences in the dispersions between sample episodes and shrubsteppe types (sample episode: F7.57= 0.46, P= 0.859; shrubsteppe:

F1,7= 0.11, P= 0.746), therefore confidence can be placed in the differences found.

Pollinators

The NMDS plot of sites in pollinator species space, across all sample dates, suggested that pollinator community composition was also unaffected by livestock grazing, as grazed and ungrazed sites were distributed similarly along both ordination axes (Figure 2.10). Early, middle and late flowering season periods, as well as shrubsteppe types, however, appeared to have differing pollinator community composition. Along axis 1 sample dates early in the flowering season were grouped separately from sample dates in the middle and end of the season (Figure 2.10). While antelope-brush and big sagebrush sites were separated along axis 2 during the middle and late season. Correlations between pollinator species abundances and the NMDS solution revealed that species from the bee genus Andrena are indicative of early- season communities, whereas wasps and species from the bee genera Lasioglossum and Agapostemon are prevalent in late season communities (Figure 2.10). Additionally, the bee genera Eucera, Nomada, Andrena and Cerambycid beetles tended to be strongly positively correlated with axis 2, indicating a greater prevalence in big sagebrush habitats. Apis mellifera, the European honeybee, was strongly, and negatively, correlated with axis 2 indicating a higher occurrence in Antelope-brush habitats. Perdita, Melissodes and Dianthidium species were also found almost exclusively in Antelope-brush habitats, but did not correlate significantly with axis 2. The final stress for the ordination was 15.18 and the final instability was 0.00001, through 139 iterations.

PerMANOVAs confirmed the patterns visualized in the NMDS. Pollinator community composition was not significantly affected by livestock grazing, although there was a trend toward differing community composition between grazed and ungrazed sites (pseudo-F1,3= 1.33, P= 0.066). However, when the site pair with the uncertain recent grazing history was removed from the analysis the trend became significant (pseudo-F1.3= 1.46, P= 0.0362). As visualized in the NMDS, pollinator community composition differed between shrubsteppe habitats and periods of the flowering season (shrubsteppe: pseudo-F1,7= 4.12, P= 0.0002; sample episode: pseudo-

F7,57= 6.81, P= 0.0002). There was no difference in the dispersions between management types (F1,3=0.10, P=0.755), sample episodes (F7,57= 0.95, P= 0.480) or shrubsteppe types (F1,7= 2.02, P= 0.160).

The Mantel test indicated that there was a positive correlation (t= 3.128) between pollinator and flowering plant communities; sites with similar plant community composition are also more likely to have similar pollinator community composition (Mantel statistic r=0.612, P= 0.002).

Discussion

In this study, I found that livestock grazing affected shrubsteppe vegetation structure, but did not significantly influence flowering plant or pollinator abundance, richness, diversity or community composition, although a trend towards differing pollinator community composition was identified. Instead, the composition of both the flowering plant and pollinator community differed significantly between the two shrubsteppe habitats sampled, antelope-brush and big sagebrush. This difference was likely driven by environmental characteristics associated with elevation change. Flowering plant and pollinator community compositions were positively correlated across sites, and along with floral and pollinator abundance, richness and diversity, changed over the course of the flowering season, as was expected. Overall, these findings suggest that flowering plant and pollinator diversity can be maintained under short- duration, low-intensity livestock grazing in the southern Okanagan.

The effects of livestock grazing

Vegetation structure

In agreement with other studies investigating the impacts of livestock grazing on vegetation structure (Jones 2000; Kruess and Tscharntke 2002; Krannitz 2008), my results show that grazing can influence the percent cover and height of vegetation. Shrub cover was greater on grazed sites, complementing findings by Krannitz (2008) which showed that big sagebrush increases with grazing intensity and that although heavy grazing can be detrimental, antelope-brush cover is highest under light grazing pressure. Similarly, Ganskopp et al. (2004) found that the growth of young antelope- brush shrubs can be stimulated by light, spring cattle grazing. In B.C., range managers consider both antelope-brush and big sagebrush as “increasers” and two other common shrub species, rabbit-brush and pasture sage (Artemisia frigida), as “invaders” in

response to livestock grazing (Gayton 2003), also supporting my findings. In contrast, the percent cover of grasses and forbs were unaffected by grazing, but both vegetation layers were shorter at grazed sites. Grasses and forbs could be shorter under grazing disturbance for a variety of reasons, including herbivory, reductions in plant vigor due to herbivory stress (Pond 1960; Krannitz 2008) and changes in species composition (Fleischner 1994). However, the consistency in the percent cover of grasses and forbs between grazed and ungrazed sites suggests that the current intensity of grazing does not negatively affect plant basal diameter or recruitment. Many other grazing impact studies have reported similar responses of the grass and forb layers (e.g. Kruess and Tscharntke 2002; Krannitz 2008; Sjodin et al. 2008).

Trampling by livestock increased the cover of bare soil, while decreasing the cover of cryptogamic crust, a finding previously shown in these (Krannitz 2008) and other semi-arid ecosystems (Anderson et al. 1982; Fleischner 1994; Jones 2000; Vulliamy et al. 2006). Cryptogamic crust, which is important for soil stability (Kleiner and Harper 1972) and moisture retention (Loope and Gifford 1972), is most susceptible to disturbance in the growing season (Memmott et al. 1998; Krannitz 2008) which is likely why spring grazing can be so damaging. The litter layer however, which commonly decreases in cover under grazing (Jones 2000; Sjodin et al. 2008) was unchanged at my sites, likely because of the relatively low grazing pressure.

Flowering plants

Contrary to expectations, changes in vegetation structure under livestock grazing did not extend to changes in pollinator-attractive flowering plant abundance, richness, diversity or community composition. Although other studies have reported similar findings (Sjodin et al. 2008; Vazquez et al. 2008; Batary et al. 2010), the response of floral communities to grazing appears complex, as reports that grazing is positive (Carvell 2002; Vulliamy et al. 2006) and that it is negative (Xie et al. 2008; Yoshihara et al. 2008; Kimoto 2010) have also been made. A number of factors appear to be important in determining how livestock grazing will impact floral communities. Ecosystems with long grazing histories and disturbance-adapted flowering plants often respond positively to grazing, as long as the grazing intensity is at an intermediate level (Carvell 2002; Vulliamy et al. 2006). In contrast, ecosystems without long grazing

histories, such as western North America and South America, are thought to be more likely to respond negatively to livestock grazing (Mack and Thompson 1982; McIntyre et al. 1996; Debano 2006; Vulliamy et al. 2006). This study, and that of Vazquez et al. (2008) from Argentina, indicate that floral communities of grasslands without long grazing histories do not necessarily respond negatively, and emphasize the importance of the current grazing regime in determining vegetation responses.

Although non-significant, there was a trend (P= 0.076) of decreased floral abundance in grazed sites during the later-half of the flowering season (June – July), roughly corresponding with the cessation of grazing. It may be that forbs subject to herbivory by livestock tend to produce fewer flowers, or that grazing decreases the abundance of some species. Yarrow (Achillea millefolium), slender hawksbeard (Crepis atrabarba), triple-nerved daisy (Erigeron subtrinervis), and silky lupine (Lupinus sericeus), all mid-to-late season flowering plants, had at least four times more floral units at ungrazed sites. Therefore, livestock may have had some influence on particular members of the floral community, but the short duration and low-intensity of grazing precluded any significant negative effects on the community as a whole.

Pollinators

The abundance, richness and diversity of pollinators were also unaffected by livestock grazing. As with floral communities, the reported responses of pollinators to grazing disturbance varies widely (e.g., Kruess and Tscharntke 2002; Vulliamy et al. 2006; Sjodin et al. 2008; Xie et al. 2008; Sarospataki et al. 2009). A common thread throughout this and previous studies is that regardless of whether grazing proved positive, negative or neutral, the response of the pollinator community closely mirrored that of the floral community. For bees this is an expected result, as pollen and nectar are food sources for both adults and their offspring (Kearns and Inouye 1997). However, flowering plant abundance and richness have also been shown to influence community structure of butterflies (Erhardt 1985), flies (Hegland and Boeke 2006; Frund et al. 2010), beetles (Volkl et al. 1993; Hegland and Boeke 2006) and wasps (Karem et al. 2010) even though only the adults feed on floral resources. The flowering plant and pollinator communities at my sites were significantly correlated, suggesting that food resources (i.e., floral community composition) are an important determinant of pollinator

community composition. Therefore, it is not surprising that the neutral effects of grazing on flowering plant abundance, richness and diversity were carried through to the pollinator community.

Although floral resources are important in shaping pollinator communities, nesting resources (Kearns and Inouye 1997; Potts et al. 2005) can also be a critical factor. Several studies have found that pollinators of differing functional groups, such as those differing in nesting habit, can respond differently to anthropogenic disturbance (e.g., Cane et al. 2006; Sjodin et al. 2008; Williams et al. 2010; Neame et al. 2012). The abundance and richness of ground nesting bees, for example, can be positively influenced by livestock grazing through the increased availability of compacted bare soil (nesting substrate) caused by livestock trampling (Vulliamy et al. 2006). However, other researchers have suggested that cattle trampling may actually disturb underground nests, leading to detrimental effects on ground nesters (Gess and Gess 1983; Sugden 1985). My results show that bare soil cover can increase under grazing pressure, but that this does not necessarily lead to an increase or decrease in ground nesting bees. Additionally, decreased vegetation height due to grazing has been shown to negatively influence the abundance and richness of butterflies (Kruess and Tscharntke 2002), hoverflies and beetles (Sjodin et al. 2008), presumably because habitat requirements for their young were altered. Although grasses and forbs were shorter at grazed sites, this did not have a significant effect on abundance, richness or diversity of butterflies, flies or beetles in my study.

Although there were no differences in pollinator abundance, richness or diversity between grazed and ungrazed sites, there was a trend towards differing pollinator community composition. Examination of pollinator species abundances indicated that, although the majority of pollinator species did not differ in abundance between grazed and ungrazed sites, those that did were among the most abundant species collected. For example, Buprestidae sp.2 and Lasioglossum pruinosum were roughly three times more abundant in ungrazed sites, whereas Halictus farinosus and H. tripartitus were twice as abundant in grazed sites. All four species were within the top-10 most abundant species sampled (175-830 individuals collected). Due to their high abundance these species are likely to carry a large weight in multivariate analyses (McCune and Grace 2002) and are likely to be driving the trend seen. Therefore, livestock grazing

may influence the relative abundance of some pollinator species, but the duration and intensity of grazing has precluded any changes in overall pollinator abundance, richness and diversity.

The effects of shrubsteppe type

Antelope-brush shrubsteppe is a lower-slope to valley-bottom ecosystem, and as a result is drier, sandier and often more nutrient-poor than the higher elevation big sagebrush shrubsteppe (Nicholson et al. 1991). These environmental differences appear to have significantly influenced the flowering plant community composition at my sites. Although many flowering plant species are present in both ecosystems, many others were sampled primarily or exclusively at low or high elevations. For instance, silky lupine (Lupinus sericeus), Thompson’s paintbrush (Castilleja thompsonii), and lemonweed (Lithospermum ruderale) were found exclusively at higher elevation big sagebrush sites, whereas golden aster (Heterotheca villosa), pale evening-primrose (Oenothera pallida) and brittle prickly-pear cactus (Opuntia fragilis) were found primarily at low elevation antelope-brush sites.

The composition of the pollinator community also differed between shrubsteppe types, perhaps as a response to available floral resources (see Appendix C for the top 10 most abundant flowering plants and pollinators of each shrubsteppe type). The bee genera Eucera, Andrena, Nomada and Cerambycid beetles were all more prevalent in big sagebrush shrubsteppe. Eucera spp. are long-tongued bees that often visit lemonweed, silky lupine, and thread-leaved phacelia (Phacelia linearis), all species more common or exclusively in big sagebrush shrubsteppe. The most common Andrena spp. collected also favoured plants more abundant in big sagebrush sites: desert-parsleys (Lomatium macrocarpum and L. triternatum), and long-flowered mertensia (Mertensia longiflora). Nomada spp. are cleptoparasites, primarily of Andrena, and their habitat choices are likely based on the location of their hosts (T. Griswold, USDA bee lab, Utah State University, pers. comm.). Cerambycid beetles favoured species from the Asteraceae family which were common everywhere, suggesting factors other than floral resources are important in determining their distribution.

Several other pollinator species were collected more frequently at antelope-brush sites than at big sagebrush sites, including honeybees. Honeybee prevalence, like Cerambycid beetles, was likely unrelated to diet. Honeybee abundance is largely determined by the location of managed hives, and in the southern Okanagan Valley most orchards and other crops are on the valley bottoms, adjacent to the remaining antelope-brush shrubsteppe. Dianthidium spp. were collected almost exclusively in antelope-brush habitats but insufficient floral records exist to assess whether diet may drive this pattern. Melissodes spp. were also collected primarily in antelope-brush habitats and visited only a few plant species including golden aster (Heterotheca villosa) which is only found at low elevations. Perdita fallax is a golden aster specialist, and along with Megachile umatillensis which specializes on the provincially Red-listed pale evening-primrose (Oenothera pallida), is found only in low elevation antelope-brush shrubsteppe where these plants occur.

Management implications

Much of the remaining shrubsteppe in the Okanagan is grazed by livestock under regimes that vary widely depending on the productivity of the land as well as the goals of local land managers. Although this variability in grazing regimes could have contributed to a lack of a grazing effect in this study, the long recovery time after disturbance of dry bunchgrass ecosystems (20-40 years; McLean and Tisdale 1972) I expected any grazing to be detrimental. Although much of the Okanagan Valley was severely overgrazed by the early 1900’s, changes in range management, such as the implementation of single-season and rotational grazing, have resulted in considerable improvements to ecosystem health (Bawtree 2005). The grazed areas included in this study, at least in the recent past, were managed with the preservation of biological diversity in mind (Wade Clifton, Clifton Ranch; Anne Skinner, B.C. Ministry of Forest and Range, pers. comm.).

My results indicate that short-term spring cattle grazing with <1 AUM/ha (1 AUM is equivalent to the forage removed by one 454 kg cow grazing for one month) does not negatively impact flowering plant or pollinator abundance, richness or diversity. Given the trends towards differing pollinator community composition and decreased floral abundance identified, I recommend that these grazing regimes be maintained, and not

increased, if the preservation of flowering plants and pollinators is of conservation concern. As the trend of decreased floral resources occurred during the mid-to-late portion of the flowering season, monitoring of grazing impacts may be most valuable within this time frame. Additionally, because antelope-brush and big sagebrush shrubsteppe both support high flowering plant and pollinator diversity, but their community compositions differ, attention to maintaining the health of both habitats under sustainable grazing practice is of great importance.

My research was performed at the community level, therefore provincially-listed rare species such as pale evening-primrose, Lyall’s mariposa lily (Calochortus lyallii), and grand coulee owl-clover (Orthocarpus barbatus) were not my focus. If rare plant species or specialist pollinators are of conservation interest, further work will be needed that aims specifically to assess such species. An encouraging observation is that aside from pale evening-primrose, other plants supporting specialist pollinators [meadow death camas (Zigadenus venenosus) with Andrena astragali; golden aster with Perdita fallax, large-fruited and narrow-leaved desert-parsleys with Andrena microchlora; Phacelia spp. with Dufourea trochantera, Colletes consors and Chelostoma phaceliae] were common and appeared unaffected by low levels of livestock grazing, suggesting grazing may not be detrimental to these specialist pollinators.

Conclusions

Short-term, low-intensity livestock grazing in the southern Okanagan shrubsteppe does negatively influence some aspects of vegetation structure, but does not significantly impact flowering plant or pollinator communities. These results suggest that semi-natural habitats, when managed responsibly, can remain reservoirs of flowering plant and pollinator diversity. This is especially encouraging for habitats, like shrubsteppe, which are biologically diverse but have few remaining undisturbed areas. As anthropogenic pressures continue to increase, the continued effort of land managers to find a balance between biological integrity and economic viability will be vital for the conservation of native plants and pollinators.

References

Anderson, D. C., K. T. Harper and S. R. Rushforth (1982). Recovery of cryptogramic soil crusts from grazing on Utah winter ranges. Journal of Range Management 35(3): 355-359.

Anderson, M. J. (2001). A new method for non-parametric multivariate analysis of variance. Austral Ecology 26(1): 32-46.

Batary, P., A. Baldi, M. Saropataki, F. Kohler, J. Verhulst, E. Knop, F. Herzog and D. Kleijn (2010). Effect of conservation management on bees and insect-pollinated grassland plant communities in three European countries. Agriculture Ecosystems & Environment 136(1-2): 35-39.

Bawtree, A. H. (2005). A History of Range Use in British Columbia. Rangelands 27(6): 4.

Carvell, C. (2002). Habitat use and conservation of bumblebees (Bombus spp.) under different grassland management regimes. Biological Conservation 103(1): 33-49.

Colwell, R. K. (2005). EstimateS: Statistical estimation of species richness and shared species from samples. Version 7.5, User's guide and application published at: http://purl.oclc.org/estimates.

Colwell, R. K. and J. A. Coddington (1994). Estimating terrestrial biodiversity through extrapolation. Philosophical Transactions of the Royal Society of London Series B-Biological Sciences 345(1311): 101-118.

Curtin, C. and D. Western (2008). Grasslands, people, and conservation: Over-the- horizon learning exchanges between African and American pastoralists. Conservation Biology 22(4): 8.

Debano, S. J. (2006). Effects of livestock grazing on aboveground insect communities in semi-arid grasslands of southeastern Arizona. Biodiversity and Conservation 15(8): 2547-2564.

Erhardt, A. (1985). Diurnal Lepidoptera - senstive indicators of cultivated and abandoned grassland. Journal of Applied Ecology 22(3): 849-861.

Fleischner, T. L. (1994). Ecological costs of livestock grazing in western North America. Conservation Biology 8(3): 629-644.

Frund, J., K. E. Linsenmair and N. Bluthgen (2010). Pollinator diversity and specialization in relation to flower diversity. Oikos 119(10): 1581-1590.

Ganskopp, D., T. Svejcar, F. Taylor and J. Farstvedt (2004). Can spring cattle grazing among young bitterbrush stimulate shrub growth? Journal of Range Management 57(2): 161-168.

Gayton, D. (2003). British Columbia Grasslands Monitoring Vegetation Change, FORREX - Forest Research Extension Partnership. 7.

Gess, F. W. and S. K. Gess (1983). Effects of increasing land utilization on species representation and diversity of aculeate wasps and bees in semi-arid areas of southern Africa. Hymenoptera and biodiversity. J. LaSalle and I. D. Gauld, CAB International: 83-114.

Gilgert, W. and M. Vaughan (2011). The value of pollinators and pollinator habitat to rangelands: Connections among pollinators, insects, plant communities, fish, and wildlife. Rangelands 33(3): 6.

Harmon, J. P., A. C. Ganguli and M. J. Solga (2011). An overview of pollination in rangelands: Who, why, and how. Rangelands 33(3): 5.

Hegland, S. J. and L. Boeke (2006). Relationships between the density and diversity of floral resources and flower visitor activity in a temperate grassland community. Ecological Entomology 31(5): 532-538.

Jones, A. (2000). Effects of cattle grazing on North American arid ecosystems: A quantitative review. Western North American Naturalist 60(2): 155-164.

Karem, J., S. A. Woods, F. Drummond and C. Stubbs (2010). The relationships between Apocrita wasp populations and flowering plants in Maine's wild lowbush blueberry agroecosystems. Biocontrol Science and Technology 20(3): 257-274.

Kearns, C. A. and D. S. Inouye (1997). Pollinators, flowering plants, and conservation biology - Much remains to be learned about pollinators and plants. Bioscience 47(5): 297-307.

Kearns, C. A., D. W. Inouye and N. M. Waser (1998). Endangered mutualisms: The conservation of plant-pollinator interactions. Annual Review of Ecology and Systematics 29: 83-112.

Kearns, C. A. and D. M. Oliveras (2009). Environmental factors affecting bee diversity in urban and remote grassland plots in Boulder, Colorado. Journal of Insect Conservation 13(6): 655-665.

Kimoto, C. (2010). Effect of livestock grazing on native bees in a pacific northwest bunchgrass prairie. Corvallis Oregon State University. Master of Science in Wildlife Science: 135.

Klein, A. M., B. E. Vaissiere, J. H. Cane, I. Steffan-Dewenter, S. A. Cunningham, C. Kremen and T. Tscharntke (2007). Importance of pollinators in changing landscapes for world crops. Proceedings of the Royal Society B-Biological Sciences 274(1608): 303-313.

Kleiner, E. F. and K. T. Harper (1972). Environment and community organization in grasslands of Canyonlands National Park. Ecology 53(2): 299-&.

Krannitz, P. G. (2008). Response of antelope bitterbrush shrubsteppe to variation in livestock grazing. Western North American Naturalist 68(2): 138-152.

Kruess, A. and T. Tscharntke (2002). Contrasting responses of plant and insect diversity to variation in grazing intensity. Biological Conservation 106(3): 293-302.

Kruess, A. and T. Tscharntke (2002). Grazing intensity and the diversity of grasshoppers, butterflies, and trap-nesting bees and wasps. Conservation Biology 16(6): 1570-1580.

Lea, T. (2008). Historial (pre-settlement) ecosystems of the Okanagan Valley and Lower Similkameen Valley of British Columbia - pre-European contact to the present. Davidsonia 19(1): 34.

Loope, W. L. and G. F. Gifford (1972). Influence of a soil microfloral crust on select properties of soils under pinyon-juniper in southeastern Utah. Journal of Soil and Water Conservation 27(4): 164-&.

Mack, R. N. (1981). Invasion of Bromus tectorum L into western North America - An ecological chronicle. Agro-Ecosystems 7(2): 145-165.

Mack, R. N. and J. N. Thompson (1982). Evolution in steppe with few large, hoofed mammals. American Naturalist 119(6): 757-773.

McCune, B. and J. B. Grace (2002). Analysis of Ecological Communities. Gleneden Beach, Oregon, U.S.A., MjM Software Design.

McCune, B. and M. J. Mefford (2006). PC-ORD. Multivariate Analysis of Ecological Data. Gleneden Beach, Oregon, U.S.A., MjM Software.

McIntyre, S., G. W. Barrett and H. A. Ford (1996). Communities and Ecosystems. Conservation Biology. I. F. Spellerberg. England, Longman Group Limited: 154- 170.

McLean, A. and L. Marchand (1968). Grassland ranges in the southern interior of British Columbia. Ottawa, Canada , Publication 1319. Canadian Department of Agriculture.

McLean, A. and E. W. Tisdale (1972). Recovery rate of depleted range sites under protection from grazing. Journal of Range Management 25(3): 178-&.

Memmott, K. L., V. J. Anderson and S. B. Monsen (1998). Seasonal grazing impact on cryptogamic crusts in a cold desert ecosystem. Journal of Range Management 51(5): 547-550.

Michener (2000). The Bees of the World. Baltimore, The Johns Hopkins University Press.

Neame, A. L., T. Griswold and E. Elle (2012). Pollinator nesting guilds respond differently to urban habitat fragmentation in an oak-savannah ecosystem. Insect Conservation and Diversity doi:10.1111/j.1752-4598.2012.00187.x.

Nicholson, A., E. Hamilton, W. L. Harper and B. M. Wikeem (1991). Bunchgrass Zone. Ecosystesm of British Columbia. D. Meidinger, Pojar, J. Victoria, B.C, Province of British Columbia.

O'Toole, C. and A. Raw (1999). Bees of the World. London, England, Blandford.

Oksanen, J., F. G. Blanchet, R. Kindt, P. Legendre, R. B. O'ttara, G. L. Simpson, P. Solymos, M. H. H. Stevens and H. Wagner (2011). vegan: Community Ecology Package. R package.

Ollerton, J., R. Winfree and S. Tarrant (2011). How many flowering plants are pollinated by animals? Oikos 120(3): 321-326.

Peart, B. (2008). A record of The World Temperate Grasslands Conservation Initiative Workshop Hohhot, China - June 28 & 29, 2008. Vancouver, BC, Temperate Grasslands Conservation Initiative.

Pond, F. W. (1960). Vigor of idaho fescue in relation to different grazing intensities. Journal of Range Management 13(1): 3.

Potts, S. G., J. C. Biesmeijer, C. Kremen, P. Neumann, O. Schweiger and W. E. Kunin (2010). Global pollinator declines: trends, impacts and drivers. Trends in Ecology & Evolution 25(6): 345-353.

Potts, S. G., B. Vulliamy, S. Roberts, C. O'Toole, A. Dafni, G. Ne'Eman and P. Willmer (2005). Role of nesting resources in organising diverse bee communities in a Mediterranean landscape. Ecological Entomology 30(1): 78-85.

Sahli, H. F. and J. K. Conner (2006). Characterizing ecological generalization in plant- pollination systems. Oecologia 148(3): 365-372.

Sarospataki, M., A. Baldi, P. Batary, Z. Jozan, S. Erdos and T. Redei (2009). Factors affecting the structure of bee assemblages in extensively and intensively grazed grasslands in Hungary. Community Ecology 10(2): 182-188.

Schluter, A., T. Lea, S. Cannings and P. G. Krannitz (1995). Antelope-Brush Ecosystems. L. a. P. B.C. Ministry of Environment, Wildlife Branch. Victoria, B.C.: 6.

Seaton, R., Ed. (2003). Ecosystem at risk: Antelope-brush restoration - proceedings. BC, Canada, Brinkman and Associates Reforestation Ltd.

Sjodin, N. E., J. Bengtsson and B. Ekbom (2008). The influence of grazing intensity and landscape composition on the diversity and abundance of flower-visiting insects. Journal of Applied Ecology 45(3): 763-772.

Sokal, R. R. and F. J. Rohlf (1995). Biometry: the principles and practice of statistics in biological reseach. New York, W.H. Freeman and Company.

Sugden, E. A. (1985). Pollinators of Astragalus monoensis Barneby (Fabaceae) - New host records - Potential impact of sheep grazing. Great Basin Naturalist 45(2): 299-312.

Vazquez, D. P., V. Aschero and E. L. Stevani (2008). Livestock grazing, habitat protection and diversity of bees and wasps in the Central Monte desert. Rev. Soc. Entomol. Argent. 67(3-4): 10.

Vazquez, D. P., W. F. Morris and P. Jordano (2005). Interaction frequency as a surrogate for the total effect of animal mutualists on plants. Ecology Letters 8(10): 1088-1094.

Volkl, W., H. Zwolfer, M. Romstockvolki and C. Schmelzer (1993). Habitat management in calcareous grasslands - Effects on the insect community developing in flower heads of Cynarea. Journal of Applied Ecology 30(2): 307-315.

Vulliamy, B., S. G. Potts and P. G. Willmer (2006). The effects of cattle grazing on plant- pollinator communities in a fragmented Mediterranean landscape. Oikos 114(3): 529-543.

Wikeem, B. and S. Wikeem (2004). The Grasslands of British Columbia. Kamloops, B.C., Grasslands Conservation Council of British Columbia.

Williams, N. M., E. E. Crone, T. H. Roulston, R. L. Minckley, L. Packer and S. G. Potts (2010). Ecological and life-history traits predict bee species responses to environmental disturbances. Biological Conservation 143(10): 2280-2291.

Wilson, E. O. (1987). The little things that run the world (the importance and conservation of invertebrates). Conservation Biology 1: 2.

Xie, Z. H., P. H. Williams and Y. Tang (2008). The effect of grazing on bumblebees in the high rangelands of the Eastern Tibetan Plateau of Sichuan. Journal of Insect Conservation 12(6): 695-703.

Zuur, A. F., E. N. Ieno and G. M. Smith (2007). Analysing ecological data. USA, Springer.

Tables

Site Area Elevation Slope Site Name Grazing regime Description abbr. (ha) (m) (%) HLU Haynes Lease 50 337 9 Ungrazed 30 yrs + Antelope-brush; Ecological Reserve valley bottom HLG Haynes Lease -Calf 43 314 7 Grazed yearly from April Antelope-brush; pasture 1-30th; 14 AUMs valley bottom OKU Kennedy Bench 40 448 3 Ungrazed 40 yrs + Antelope-brush; Antelope-brush valley side bench Conservation Area OKG Mt. Oliver Protected 260.5 503 5 Grazed yearly from April Antelope-brush; Area 15- May 15; 72 AUMs valley side bench WLU White Lake 55 713 7 Ungrazed 12 yrs; rarely Big sagebrush; Biodiversity Ranch and lightly grazed prior bottom of side valley WLG White Lake 170 563 15 Grazed every other year Big sagebrush; lower Biodiversity Ranch from May 15- June 30; slope of side valley 110 AUMs SOU Southern Okanagan 20 883 22 Ungrazed 6 yrs; prior Big sagebrush; mid Grasslands management unknown slope bench Protected Area SOG Southern Okanagan 1850 884 18 Ungrazed 6 yrs ; prior Big sagebrush; mid Grasslands grazing: May 1-31; 160 slope bench Protected Area AUMs

Table 2.2. Summary of the total abundance, richness and diversity of pollinator-attractive flowering plants and flower visitors (hereafter pollinators) for eight shrubsteppe study sites in the southern Okanagan, British Columbia.

Flowering plants Pollinators Site abbr. Total floral unit Diversity Richness Richness Total pollinator Diversity Richness Richness abundance (1-D) (Chao2) (Actual) Abundance (1-D) (Chao2) (Actual) HLU 361 0.3310 13.75 12 745 0.8865 170.11 95 HLG 143 0.2305 5.00 5 598 0.8720 142.10 87 OKU 873 0.5978 20.12 20 725 0.8578 198.17 94 OKG 474 0.4537 23.08 19 1227 0.9040 139.53 98 WLU 729 0.6805 21.63 19 629 0.9671 152.08 108 WLG 645 0.5578 27.15 24 895 0.9462 172.88 130 SOU 1382 0.6805 27.09 26 984 0.9710 171.48 120 SOG 1118 0.5578 21.88 20 514 0.9533 144.21 96

Figures

Haynes Lease (HL) 140 120 100 80 60 40 20 Grazed Ungrazed 0 Mt. Oliver/Kennedy bench (OK) 140 120 100 80 60 40 20 Grazed Ungrazed 0 S. Okanagan Grasslands (SO) 140 120 100 80

Number of Species of Number 60 40 20 Grazed Ungrazed 0 White Lake (WL) 140 120 100 80 60 40 20 Grazed Ungrazed 0 0 200 400 600 800 1000 1200 0 200 400 600 800 1000 1200

Number of Individuals

3500 41 Grazed 3000 Ungrazed

2500

2000

1500

1000 47

17 Number of individuals caught individuals of Number 500 35 59 11

17 18 1 3 5 3 0

Apidae Beetles Wasps ColletidaeHalictidae Syrphidae Andrenidae Bombylidae Apis mellifera Megachilidae Other Diptera

Butterflies/Moths Pollinator type

40 100

Grazed Ungrazed * 80 30

* * 60 * 20 *

10 20

Mean percent cover ± 1 SE ± cover percent Mean

Mean maximum height (cm)1 SE ±height maximum Mean

0 0

Shrub Grass Forb Crust Litter Grass Forb Bare soil

Vegetation and ground layers

LS means of floral unit Grazed

abundance and richness ± 1SEabundance and richness e-1 Ungrazed Abundance Richness

0 1 2 3 4 5 6 7 8

Sample episode

LS means of floral unit Antelope-brush

abundance and richness ± 1SE and richness abundance e-1 Big sagebrush Abundance Richness

1 2 3 4 5 6 7 8

Sample episode

Grazed

LS ofmeans total pollinator

abundance and richness ± 1SE abundance and richness Ungrazed Abundance Richness e0 0 1 2 3 4 5 6 7 8

Sample episode

LS means of total pollinator Antelope-brush

abundance and richness ± 1SE and richness abundance Big sagebrush Abundance Richness e0 1 2 3 4 5 6 7 8 Sample episode

1.5 Early-flowering Management: spp. Grazed 1.0 Ungrazed Shrubsteppe type:

0.5 Antelope-brush Big sagebrush Sample episode: 0.0 1 Axis 2 Axis 2 -0.5 3 4 5 -1.0 6 7 Late-flowering 8 spp. -1.5 -1.5 -1.0 -0.5 0.0 0.5 1.0 1.5 Early/late-flowering spp. Axis1 Mid-flowering spp. Common across sites Infrequent across sites

Cerambycidae 1.5 Andrena Management: Nomada 1.0 Ungrazed Eucera Grazed

0.5 Shrubsteppe type: Antelope-brush Big sagebrush 0.0 Sample episode: 1 Axis 2 -0.5 2 3 -1.0 4 5 6 -1.5 7 8 Apis mellifera -2.0 -1.5 -1.0 -0.5 0.0 0.5 1.0 1.5 2.0 Wasps Axis 1 Andrena Lasioglossum Agapostemon

Chapter 3

A comparison of plant-pollinator network structure between British Columbia’s endangered shrubsteppe habitats

Introduction

The vast majority of flowering plants are dependent on animals, primarily insects, for pollination (Ollerton et al. 2011). Thus, through the facilitation of plant reproduction, pollinators play a vital role in maintaining natural ecosystems and agricultural productivity. Concern over the fate of pollinator communities is rising, however, as reports of pollinator declines have surfaced in numerous places around the world (Kearns et al. 1998; Potts et al. 2010), with some areas also reporting parallel declines in pollinator-dependent plants (Biesmeijer et al. 2006). These reports have highlighted the importance of considering pollinator communities in future conservation and management efforts (Potts et al. 2010).

Over the last decade, the study of pollinators has benefited from taking a community-based analytical approach, made possible by examining plant-pollinator interaction networks (Bascompte 2007; 2009; Vazquez et al. 2009). In contrast to traditional analyses, which focus on quantifying species abundances, plant-pollinator networks provide a more functional perspective by identifying which species interact within a community, how frequently species interact, and how these interactions are structured (Bascompte and Jordano 2007). Recently, studies have shown that network structure can be influenced by anthropogenic disturbances (e.g., Lopezaraiza-Mikel et al. 2007; Aizen et al. 2008; Yoshihara et al. 2008), even when species richness within a community is unaffected (Tylianakis et al. 2007). These results emphasize the importance of an analytical approach that addresses species interactions in ecological

communities, in addition to traditional abundance and richness indices. Although there is still much to be learned about plant-pollinator network structure, its potential use in conservation and management is beginning to be realized (e.g., Gibson et al. 2006; Carvalheiro et al. 2008; Forup et al. 2008).

Plant-pollinator networks have been found to have conserved network-level structural properties that have implications for community resilience (Table 3.1; Memmott et al. 2004; Bascompte and Jordano 2007). In the context of plant-pollinator networks, resilience generally concerns a network’s capacity to resist secondary extinctions following species loss (Memmott et al. 2004; Tylianakis et al. 2010), thus increasing the ability of the community to absorb disturbance and retain essentially the same structure and function. Although many network structural properities can be measured, connectance, generalization, asymmetry and nestedness may be the most useful properties for understanding resilience (Elle et al. in press).

One of the most commonly measured network properties is connectance, a measure of interaction richness, which is calculated as the proportion of realized interactions out of all possible interactions within a network (Jordano et al. 2006). When compared between networks of similar size (i.e. similar species richness), increased connectance indicates increased generalization of the species involved (Tylianakis et al. 2010) and confers higher network resilience through redundancy in interaction partners (Thebault and Fontaine 2010). That is, the more interaction partners each species has, i.e. the more pollinators each plant has and the more plants each pollinator visits, the less likely the loss of an interaction partner will result in secondary population declines or extinctions. However, it is well-established that connectance decreases with increasing network size and is thus inappropriate to compare across networks of different size (Vazquez et al. 2009). Furthermore, connectance is based on binary data and thus does not incorporate the frequency of interactions between species in a network, which is an important component in assessing species generalization and in determining how detrimental the loss of an interaction partner may be. Quantitative metrics of generalization (generality and H2’ specialization index) which evaluate the average number of species each species in the network interacts with, account for heterogeneity in interaction frequency and are thus an appropriate way to compare interaction diversity

among networks of different sizes (Bluthgen et al. 2006; Albrecht et al. 2010; Tylianakis et al. 2010).

Asymmetry and nestedness are two other commonly measured network properties that have implications for network resilience. Plant-pollinator networks are usually asymmetric in terms of degree and interaction strength (Vazquez and Aizen 2004; Bascompte et al. 2006). Asymmetry in degree refers to the tendency of specialized species to interact with more generalized species (Vazquez and Aizen 2004), while asymmetry in interaction strength refers to the tendency of species with strong effects to usually experience weak reciprocal effects from their interaction partners (Bascompte et al. 2006). A generalist plant, for example, may be the dominant pollen source for a number of pollinator species, but does not rely strongly on any one of these species for pollen transfer. Increased asymmetry is thought to confer greater network resilience by contributing to the persistence of more specialized species, since the abundance of their generalist interaction partners tend to be higher and less prone to fluctuation (Bascompte et al. 2006; Bascompte 2009). When calculated separately for each species in a network, interaction strength asymmetry also provides a method to identify which plants and pollinators are strongly relied upon within the network (Hegland et al. 2010) and thus may be good candidates for monitoring programs (Elle et al. in press). Plant-pollinator networks are also usually nested, such that they are organized around a core of interacting generalists, some of which also interact with specialists (Bascompte et al. 2003). Increased nestedness is thought to confer network resilience by increasing the persistence of more specialized species, similar to asymmetry, and by creating an interacting core of generalists that remains intact if more specialized species are lost from the network (Memmott et al. 2004; Fortuna and Bascompte 2006).

The aforementioned structural properties, among others, can be compared between different networks to indicate which may be more sensitive to disturbance or investigate how disturbances influence community structure. For example, many studies have investigated how introduced species, particularly plants, influence network structure (e.g., Lopezaraiza-Mikel et al. 2007; Aizen et al. 2008; Bartomeus et al. 2008; Vila et al. 2009; Kaiser-Bunbury et al. 2010). Other studies have compared the robustness of different networks to species loss, for example those in restored vs.

reference heathland sites (Forup et al. 2008) and lightly grazed vs. heavily grazed rangeland (Yoshihara et al. 2008).

Thus far, studies taking a network approach to plant-pollinator conservation have rarely incorporated the effects of temporal dynamics (e.g., seasonality) on network structure (but see Valdovinos et al. 2009; Hagen and Kraemer 2010). The few studies that have explored the temporal dynamics of plant-pollinator networks (e.g., Basilio et al. 2006; Medan et al. 2006; Olesen et al. 2008) have shown that some network properties have strong temporal dynamics, such as connectance and nestedness. Taking a temporal approach to network analysis could therefore aid in the development of conservation strategies by indicating how network structure changes over time and suggesting when networks may be more sensitive to disturbance.

In North America, grasslands and shrubsteppe are among the continent’s most species-rich and threatened ecosystems. Conservation concerns in these ecosystems include extensive fragmentation and degradation due to agricultural and urban development (Curtin and Western 2008; Peart 2008). Within Canada, the shrubsteppe habitats of the south Okanagan Valley, British Columbia, are recognized as some of the most biologically diverse and threatened habitats. Antelope-brush shrubsteppe in particular, supports a disproportionately high percentage of Canada’s endangered and threatened species and is considered one of the top four most endangered ecosystems in the country (Schlute et al. 1995; Dyer and Lea 2003). Due to its position on valley bottoms and low elevation valley side benches, 68% of antelope-brush habitat has been lost to agriculture and urban development and was still being lost at a rate of 2% per year within the last decade (CDC 2003; Dyer and Lea 2003). Big sagebrush shrubsteppe, which also supports numerous endangered and threatened species, occurs at higher elevations than antelope-brush shrubsteppe and has suffered less from habitat loss and fragmentation (Lea 2008). Fairly little is known about the pollinator communities of this region, though they are hypothesized to be very diverse due to the sub-desert climate. Pollinators and the flowering plants with which they interact are a vital component of shrubsteppe biodiversity, together providing vegetation structure and forage that is vitally important for many species of herbivores, insectivores, granivores and frugivores that also inhabit these ecosystems (Gilgert and Vaughan 2011). Thus,

understanding the structure of the plant-pollinator communities in shrubsteppe habitats is important for their effective management and conservation.

In this study, I describe plant-pollinator interaction networks from antelope-brush and big sagebrush shrubsteppe generated from plant-pollinator interaction sampling completed over an entire flowering season. I investigate differences in network structure between antelope-brush and big sagebrush shrubsteppe that may have consequences for community resilience to disturbance and explore which plant and pollinator species are functionally important in each habitat. I also examine temporal variability in network structure to investigate how these plant-pollinator networks, and sensitivity to disturbance, change over the course of the flowering season. I conclude by summarizing the practical implications of my findings for the conservation of shrubsteppe plant-pollinator communities of the southern Okanagan.

Methods

Study sites

I selected eight sites in the southern Okanagan Valley; four in antelope-brush shrubsteppe and four in big sagebrush shrubsteppe (Table 3.2). Sites were initially

chosen to investigate the impacts of livestock grazing on plant and pollinator communities, thus two sites in each shrubsteppe habitat type are lightly spring grazed for a month’s time on a yearly or bi-yearly basis. Previous analysis has shown this disturbance does not significantly impact flowering plant and pollinator diversity or community composition, although vegetation structure does change with grazing (see Chapter 2). All sites were a minimum of 20 hectares and were connected to contiguous shrubsteppe, grassland or ponderosa pine forest on at least one side. Within each site I selected areas that would be suitable for sampling, i.e. encompassing the most prevalent shrubsteppe vegetation type of the site and excluding less common landscape features such as drainage areas. Within these areas a point was randomly selected on an aerial photo and used as the starting corner for a 1-ha sampling plot, within which I conducted all plant-pollinator interaction sampling.

Sampling plant-pollinator interactions

I sampled plant-pollinator interactions at each site over the entire flowering season (March-July 2010). Flower visitors were collected with a net directly off of flowering plants so that species-level interactions could be identified and their frequency assessed. Plant-species-specific netting is considered more appropriate than transect- based netting in heterogeneous environments like shrubsteppe, as netting effort is allocated more evenly among plant species and is more likely to detect uncommon interactions (Gibson et al. 2011). I conducted netting surveys at roughly one-week intervals, in fair weather conditions between 9:30 and 16:00 hours. During each netting survey, two 10 minute netting bouts were conducted on each flowering plant species in bloom. On occasion, a single netting bout was conducted if a plant species was just beginning to bloom and there were few open flowers. During netting bouts, samplers walked throughout the plot catching all observed flower visitors that came into contact with the reproductive organs of the focal plant species. Sampling bout times (AM, mid- day, PM) for each plant species were varied within and between netting surveys to encompass the flight time of most pollinating insects. I attempted to allocate consistent netting effort to each plant species across all sites, but differences in bloom length precluded complete consistency. Thus, the overall netting effort of each site reflects the flowering plant diversity and phenology of that site. Plant species with very small flowers, such as spring draba (Draba verna), and species present at only a single site in

low abundance (< 10 flowering individuals) were not included in the study. All flower visitors were identified to the lowest taxonomic level possible, primarily species or morphospecies. Bees and wasps (Hymenoptera), beetles (Coleoptera), and flies (Diptera) were collected for identification, whereas butterflies (Lepidoptera) and hummingbirds (Trochilidae) were identified without being captured. The interaction networks resulting from this sampling are most appropriately termed flower-visitor networks, as I did not assess the role each visitor played in plant pollination, but following convention I will call them plant-pollinator networks hereafter.

Each site was surveyed 13 or 14 times across the flowering season, but to facilitate comparison of network structural properties across sites over time, each network was reduced to 12 netting surveys for analysis. The first survey of the flowering season was removed for all networks because of low floral availability and limited pollinator activity. If a site was surveyed an extra time (14 rather than 13 samples), the last netting survey was also removed. Each network was then divided into three seasonal sub-networks (early, mid and late flowering season), each consisting of four netting surveys spanning approximately 35 days.

Quantifying plant-pollinator network structure

I constructed quantitative plant-pollinator interaction matrices for all complete and seasonal sub-networks. In these matrices, rows and columns represent flowering plant and pollinator species, respectively, while cells record the frequency of interactions between each plant and pollinator species. To investigate if plant-pollinator community structure differs between habitat types over the course of the flowering season I calculated the following properties for all networks: number of plant species, number of pollinator species, network size, connectance, plant generality and pollinator generality,

H2’ specialization index, interaction strength asymmetry, and nestedness. All network properties were calculated using the bipartite package of R v.0.95.263 (R Development Core Team, 2011; see Appendix D for network property formulas).

I calculated network size as the sum of all plant and pollinator species in the network, as defined by Dormann et al. (2009). Connectance was calculated as the realized proportion of possible interactions (number of realized interactions/ total number

of possible interactions). Although connectance is ineffective for comparison across networks of different size, as those in this study are, it does provide a rough gauge of sampling effort through the comparison of connectance values with previously published networks of similar size.

I calculated overall plant and pollinator generalization using the generality and vulnerability indices originally derived by Bersier et al. (2002) for food web analysis, which have recently been applied to mutualistic networks by Albrecht et al. (2010). Pollinator generality (same as the generality index) is measured as the mean number of plant species visited by a pollinator species weighted by interaction strength. Where interaction strength is a measure of the dependence of one species on another and is well estimated by interaction frequency (Vazquez et al. 2005). Plant generality (same as the vulnerability index) is measured as the mean number of pollinator species visiting a plant species weighted by interaction strength.

To characterize the degree of network-level generalization, including both plants and pollinators, I used the H2’ specialization index developed by Bluthgen et al. (2006).

H2’ ranges between 0 and 1 for extreme generalization and specialization, respectively. This index is useful for comparison across multiple networks as it is robust to differences in network size and sampling intensity (Bluthgen et al. 2006). The H2’ index characterizes the degree of specialization in a network based on the deviation of a species’ realized number of interactions from that expected from the total number of interactions for that species. The underlying equation is the same as Shannon’s interaction diversity (H2), but the value computed for a given network is standardized against the minimum and maximum possible for the same distribution of interaction totals (Bluthgen et al. 2006; Dormann et al. 2009).

I calculated interaction strength asymmetry (hereafter asymmetry) as per Vazquez et al. (2007), which ranges between -1 and 1. Using this method, independent asymmetry values are calculated for each species in a network and then averaged to obtain a network-level asymmetry value. At the species-level, an asymmetry value close to 1 indicates that a species is strongly relied upon by its interaction partners but does not experience strong reciprocal effects, i.e., it does not rely strongly on any one interaction partner. Conversely, a species with an asymmetry value close to -1 relies

strongly on its interaction partners, but does not exert a strong reciprocal effect. At the network level, a value closer to 1, either positive or negative, indicates high overall asymmetry in the network (reliance between interaction partners is disproportionate), whereas a value closer to 0 indicates that interactions within the network are more symmetric (reliance between interaction partners is similar).

I also used species-level asymmetry values to identify functionally important plants and pollinators. I consider a species to be of high functional importance if it has many interaction partners and is relied strongly upon by those interaction partners. Antelope-brush and big sagebrush habitat differ in plant and pollinator community composition (see Chapter 2), therefore I aimed to identify and compare the top-10 plant and pollinator species with high functional importance in each habitat type. For this assessment I created a cumulative network for each habitat type, combining data across all sites, and ranked species according to asymmetry and species degree (the number of species which a species visits or is visited by). As a species may be able to interact with more species than are available at a single site, combining data across sites provides a more accurate estimate of each species degree and asymmetry and will identify species that are, in general, the most functionally important in each habitat type. Plants and pollinators were ranked separately and those species with a large species degree and high positive asymmetry were identified.

I calculated nestedness using the NODF metric developed by Almeida-Neto et al. (2008), which reduces the potential bias introduced by network size and asymmetry in network dimensions (ratio of plants to pollinators) compared to the previously commonly used nestedness metrics like matrix temperature and discrepancy (Almeida-Neto et al. 2008). The NODF metric is based on two properties of nestedness termed decreasing fill and paired overlap. The metric measures whether the number of interaction partners differs among plants and among pollinators in the matrix (decreasing fill), and whether more specialized species interact with subsets of the species that more generalized species interact with (paired overlap). The NODF metric ranges from 0 to 100 indicating non-nestedness and perfect nestedness, respectively (Almeida-Neto et al. 2008). Perfect nestedness implies that each species interacts only with proper subsets of those species interacting with more generalized species (Bascompte et al. 2003).

Statistical analysis

To investigate whether antelope-brush and big sagebrush habitat differ in network structural properties over the course of the flowering season I used mixed models in SAS 9.2 (Proc MIXED; SAS Institute Inc. 2008). The models included habitat, season and the corresponding two-way interaction as fixed effects and site nested within habitat as a random effect. For comparison, I also used mixed models to assess the impacts of habitat type on the structural properties of full season networks. These models included habitat as a fixed effect and site nested within habitat as a random effect. For all models, least square means were computed for all fixed effects.

Results

I surveyed 48 flowering plant species, 26 of which were present in both habitat types, and collected 264 floral visitor species/morphospecies (hereafter pollinators), 112 of which were present in both habitats, across the eight sites. Overall, I recorded 2480 plant-pollinator interactions, 919 of which were unique. Full season networks ranged in size between 43 and 170 species, while seasonal sub-networks varied between 16 and 79 species (see Appendix E for the properties of each network). Bees were the most prevalent and species-rich pollinator group collected (66.8% of recorded interactions; 153 species), followed by flies (12.7%; 57 species), beetles (9.2%; 16 morphospecies), wasps (7.1 %; 28 morphospecies), butterflies (4.0%; 8 morphospecies) and hummingbirds (<1%; 1 species; Figure 3.1). The most prevalent plant families surveyed were the aster (Asteraceae, 13 species), carrot (Apiaceae, 3 species), and lily (Liliaceae, 3 species) families.

A quantitative plant-pollinator network, including both full season and seasonal sub-networks from both antelope-brush and big sagebrush habitat is shown in Figure 3.1. There was a trend, although non-significant, for networks to increase in size across the flowering season (Table 3.3; Figure 3.2a), driven by a significant increase in the number of pollinator species (Table 3.3; Figure 3.2b). Plant species richness, on the other hand, decreased in richness late in the flowering season (Table 3.3; Figure 3.2b).

Additionally, networks in big sagebrush habitat tended to be larger than those in antelope-brush habitat, but the difference was non-significant (Table 3.3; Figure 3.2a, b).

Network-level specialization, measured by the H2’ specialization index, was unchanged throughout the flowering season, but was significantly higher in antelope- brush networks than in big sagebrush networks (Table 3.3, Figure 3.2c). This was reflected in the results for generality: plant generalization was significantly higher in big sagebrush networks, and although non-significant, pollinators also tended to be more generalized in big sagebrush habitats (Table 3.3; Figure 3.2d). Over time, pollinator generalization remained the same, which is likely driving the consistency in overall network-level specialization, as pollinators make up a large proportion of the species present. Plant generalization increased over the course of the flowering season, but the timing of this increase depended on shrubsteppe type (habitat x season interaction; Table 3.3; Figure 3.2d). In antelope-brush networks plant generalization increased in the mid flowering season and continued into the late season, whereas in big sagebrush habitats the increase in plant generalization occurred late in the flowering season.

Network-level asymmetry was similar between antelope-brush and big sagebrush shrubsteppe networks throughout the flowering season (Table 3.3; Figure 3.2e). At all sites, there were more pollinators with negative asymmetry values (~96%) than plants (~15%; see Appendix B for species-level asymmetry values), which indicates that pollinators in these habitats relied more strongly on the plants they visit for floral resources than the plants relied on them in turn for pollen transfer. As there was approximately four times more pollinator than plant species in these habitats, network- level asymmetry was consequently negative. Network asymmetry also became significantly more negative late in the flowering season (Table 3.3; Figure 3.2e). This difference was a result of a significantly more generalized plant community interacting with a proportionally larger, but similarly generalized pollinator community during that period of the flowering season.

Antelope-brush and big sagebrush shrubsteppe shared six of their top-10 functionally important plant species (Table 3.4). Yarrow (Achillea millefolium) had the largest degree and was the most positively asymmetric plant in the study, interacting with 47 and 51 species in antelope-brush and big sagebrush shrubsteppe, respectively.

Species from the Aster family were prominent in the top-10 lists of both habitats. Bee species comprised 8 of the top-10 functionally important pollinators in both habitats, three species of which were shared between shrubsteppe types. The only introduced bee species collected (Apis mellifera; European honeybee) was included in the top-10 lists for both habitats, while the most functionally important plant species were all native.

The nestedness of networks in antelope-brush and big sagebrush differed significantly with the period of the flowering season (habitat x season interaction; Table 3.3; Figure 3.2f). Nestedness of big sagebrush networks tended to increase across the flowering season, while the nestedness of antelope-brush networks tended to decrease. There was no significant main effect of either habitat type or season (Table 3.3). Overall, nestedness values for networks in both habitats were low, suggesting nestedness may not be a strong structural component of these plant-pollinator communities.

Discussion

Habitat and temporal influences on network structure

Network size and generalization

There were trends in network size, although non-significant, between habitat types and across the flowering season. Big sagebrush networks tended to be larger than antelope-brush networks, and late-season networks tended to be larger than those earlier in the season. It has frequently been proposed that more diverse communities are more stable and resilient to disturbance (Macarthur 1955; Elton 1958; Tilman et al. 1996; McCann 2000). One hypothesis is that increased species richness increases functional redundancy; in other words, it increases the number of species contributing to the same function so that if one species is lost, ecological function (e.g. pollination) may persist because of compensation from other species (Lawton and Brown 1993; Naeem 1998). Additionally, it is thought that the more species present in a community the higher the odds that at least some species contributing to the same function will respond differently to perturbations and thus be able to compensate for the loss of affected species (Elmqvist et al. 2003). Winfree and Kremen (2009), for example, have shown

that diverse pollinator assemblages comprised of species that respond differently to agricultural development are responsible for the maintenance of pollination rates in watermelon. Also, in the network literature, increased species richness and increased connectance (interaction richness) have been shown to increase plant-pollinator community resilience to simulated species loss (Memmott et al. 2004; Thebault and Fontaine 2010). In the context of my study, these results suggests plant-pollinator communities in big sagebrush habitat may be more resilient to species loss than those in antelope-brush habitat, and that all shrubsteppe communities are more resilient late in the flowering season. However, a more concrete comparison of interaction redundancy can be provided by network-level generalization measures.

For more diverse plant-pollinator communities to be more resilient to disturbance through interaction redundancy there should be an increase in the average generalization of the species involved, which I found. Plant-pollinator networks in big sagebrush had more generalized species on average than those in antelope-brush habitat. For example, 71% of plant species in big sagebrush habitats interacted with more than 10 pollinator species, compared to only 48% in antelope-brush habitats. Similarly, 37% of pollinators in big sagebrush habitat interacted with three or more plants, compared to 29% in antelope-brush habitats. Specialized plants and pollinators have long been hypothesized to be more vulnerable to disturbances, such as habitat alteration or fragmentation, than more generalized species because the loss or decline in even one interaction partner could lead to reproductive failure for plants or population declines due to reduced forage for pollinators (Bond 1994; Waser et al. 1996; Aizen et al. 2002). Thus if higher generalization levels can increase network resilience through interaction redundancy then plant-pollinator communities in big sagebrush habitat may be less vulnerable to disturbance. Additionally, I found that late-season communities of both habitats contained plant species that had a wider suite of pollinators (more generalized) on average than those blooming early in the season. Therefore, plants blooming late in the season should be less sensitive to anthropogenic disturbances that influence the abundance of some pollinator species (Waser et al. 1996).

Asymmetry

The asymmetry of antelope-brush and big sagebrush networks was similar through time, and was significantly higher late in the flowering season. Simulation studies suggest that network asymmetry contributes to the resilience of plant-pollinator communities, by contributing to the persistence of more specialized species (Fortuna and Bascompte 2006; Kaiser-Bunbury et al. 2010). Since specialists are more likely to persist under disturbance when they interact with more generalized species, which are often more abundant and less prone to population fluctuation, than when they interact with other specialists. This increase in resilience rests on the assumption that rare and more specialized species are most likely to be lost first from a community, since losing a well-connected generalist can be very detrimental to community structure (Memmott et al. 2004; Kaiser-Bunbury et al. 2010). Although anthropogenic disturbances or other ecological processes can result in the selective decline of abundant and generalized species in a network, such as bumblebee declines in Europe (Goulson et al. 2008), theory and empirical evidence indicate that it is more likely that rare and/or specialized species will be lost before the most functionally important mutualists (Tscharntke et al. 2002; Henle et al. 2004; Biesmeijer et al. 2006; Kaiser-Bunbury et al. 2010). In Britain, for example, pollinators that rely on few plants for their floral resources have experienced the largest declines over the past 30 years (Biesmeijer et al. 2006). If asymmtery in mutualistic networks can promote network resilience, then networks of both shrubsteppe habitats may be more resilient to disturbance late in the flowering season. Specifically, higher network asymmetry paired with higher plant generalization during this time period may contribute to the persistence of more specialized pollinators.

As with most plant-pollinator networks studied, network-level asymmetry values were negative in this study. Pollinators tend to have more negative asymmetry than plants (Vazquez et al. 2007) and are often far more abundant within networks. In the shrubsteppe habitats I studied, most plants were generalists and many were very generalized (> 20 interaction partners), supporting a much more diverse pollinator community that on average interacted with only a few plants. Thus, pollinator species tended to have a stronger reliance on the plants they visited than the plants did on them, suggesting plant-centered conservation and monitoring is likely a good strategy in this region.

Comparing asymmetry and degree between species provides a way to identify which plants and pollinators are most functionally important within each habitat. Although antelope-brush and big sagebrush habitat have different plant and pollinator community composition (see Chapter 2), they do share roughly half of their top-10 most functionally important species. Bees were the most prevalent and generally most functionally important pollinators in both habitats. Sweat bees (Lasioglossum spp., Halictus spp.) and mining bees (Andrena spp.) were the most common functionally important native bees, followed by bumblebees (Bombus spp.), small carpenter bees (Ceratina spp.), and mason bees (Osmia spp.). Even when extended to include the top- 20 pollinators, bees dominated functionally in big sagebrush habitats, along with Cerambycid beetles, while in antelope-brush habitats wasps (Vespids, Chrysidids and Sawflies) and flies (Syrphids and Tachinids) also became quite important.

Introduced plants did not rank highly in terms of functional importance in either habitat. This suggests that although introduced plants are integrated into the pollination networks of these sites, at their current abundance they are unlikely to be attracting a considerable number of pollinator visits away from native species. The one introduced pollinator collected in the study, the European honeybee, was ranked within the top-10 functionally important pollinators in both habitats. Although present in both, honeybees were far more prevalent in antelope-brush habitat, comprising 12.5% of all interactions compared to only 1.5% in big sagebrush habitat, likely because antelope-brush habitat is closer in proximity to the valley bottom orchards that use managed honeybees for pollination services.

Nestedness

Although overall nestedness between the two habitats was similar, there was a trend for nestedness to be larger in big sagebrush networks late in the flowering season. Nestedness in big sagebrush networks increased over time while it decreased in antelope-brush networks. Higher nestedness, like asymmetry, indicates increased tendency of specialist species to interact with generalists, but also indicates an increased tendency for generalists to interact amongst themselves, which together buffer against secondary extinctions (Memmott et al. 2004; Tylianakis et al. 2010). Because networks in both habitats are similarly asymmetric, it is likely an increase in the number

of generalized interactions in big sagebrush habitat that is the cause for the observed differences. That said, networks in both habitats had low nestedness values overall (between 4-11), with all network values falling well below the range of NODF nestedness values commonly reported in the literature (commonly published NODF range: 20-60; e.g., Bosch et al. 2009; Hegland et al. 2010; Sugiura 2010; Chacoff et al. 2012; Vilhena et al. 2012). This suggests nestedness is not a large structural attribute of these communities. Although it was previously suggested that nestedness was a universal property of plant-pollinator networks, recent studies indicate that some plant-pollinator networks are not in fact nested (Ulrich et al. 2009; Joppa et al. 2010; Gibson et al. 2011). NODF nestedness is sensitive to matrix fill (the number of realized interactions between species in a network; Almeida-Neto et al. 2008), as all nestedness metrics are, thus it is possible that higher sampling effort could have produced a more nested structure. However, connectance values of all full and seasonal sub-networks were comparable to networks of similar size in the published literature (Memmott 1999; Olesen et al. 2002; Vazquez and Simberloff 2003; Bezerra et al. 2009; Albrecht et al. 2010), thus I feel confident that my sampling effort was adequate and that nestedness is not a large structural component of these plant-pollinator communities.

Caveats to the current network approach

Several caveats to this study are worth noting. Firstly, I quantified pollinator visitation to flowering plants and not pollination. However, Vazquez et al. (2005) has showed that the most frequent pollinators to a flowering plant species are likely to be the most important pollinators. Secondly, sampling effort was not only related to flowering plant phenology but also species richness. This resulted in equal netting effort among species of similar bloom length, but different netting effort across sites. Therefore big sagebrush sites, which tended to have higher plant richness, had increased overall sampling effort which may have influenced recorded pollinator richness. However, more species-rich floras commonly support higher pollinator diversity (Kevan 1999). Thirdly, these networks exclude flowering plant species that were infrequent and in low abundance, thus do not capture the entire plant-pollinator networks of these sites. Although I feel confident the networks are a good representation of these plant-pollinator communities, further work would be needed to address the pollinator assemblages of rare plant species. Lastly, although taking a sub-network approach can provide

additional valuable information about network structure, it also requires analyzing smaller networks. Extensive simulations recently completed by Dormann et al. (2009) suggest certain network properties, such as H2’ specialization index, are sensitive to asymmetry in network dimensions (ratio of plants to pollinators) when networks are small. The authors suggest networks have a minimum of 50 species, which is the average size of my sub-networks, before such indices are used with confidence. As there was good overall concordance between the influence of habitat on network structure generated through the seasonal and full season approach, I have confidence the sub-network properties generated were representative of each network and were not an artefact of their size.

Practical implications

Both of the shrubsteppe habitats of the Okanagan Valley support diverse plant and pollinator communities. My network analysis suggests that the plant-pollinator communities of the more critically endangered antelope-brush shrubsteppe may be more sensitive to disturbances than those in big sagebrush, such as increased habitat alteration or fragmentation, as they have a tendency to be less diverse and are less generalized on the whole. These may be natural differences characteristic of each plant-pollinator community, or could be a result of the differences in anthropogenic disturbance experienced regionally by these shrubsteppe habitats. If so, my results suggest that further fragmentation and alteration of antelope-brush habitats is likely to have negative effects on plant-pollinator communities. Although their community composition differs, these shrubsteppe habitats support a number of the same plant species and share more than half of their pollinator species when singletons are not considered. Additionally, the habitats share roughly half of their top-10 functionally important plants and pollinators. Thus, the management and protection of one habitat is likely to be very beneficial to the plant-pollinator communities of the other. Protection or restoration of big sagebrush habitat near remaining antelope-brush fragments will likely promote plant and pollinator diversity by reducing habitat isolation. Also, due to the many species present in only one of the shrubsteppe habitat types, an effort towards management and conservation of both habitats will be important for maintaining regional diversity. The networks of both habitats appear to be better buffered against the loss of

pollinators than that of plants, suggesting that monitoring and management of the floral communities may be most important.

Network structural properties also indicated that early-season communities may be more sensitive to disturbance than late-season communities. The majority of remaining antelope-brush and big sagebrush habitat is grazed by livestock at some point on a yearly or bi-yearly basis (Lea 2008). Although low-intensity spring grazing has been found not to negatively affect shrubsteppe plant and pollinator communities (see Chapter 2), grazing at higher intensities may have less impact on plant-pollinator communities if it can be shifted away from early spring to later in the flowering season (June-July). Comparing network structural properties temporally, although rarely pursued in a conservation context, may be a promising additional component of using plant-pollinator networks to address applied ecological questions.

Often conservation aims to preserve endemic or endangered species, but taking a network approach highlights that monitoring and preserving some of the more common, generalist taxa may be more beneficial for preserving overall community diversity and functioning (Dupont et al. 2003; Hegland et al. 2010). Networks allow functionally important species to be identified which not only provide good candidates for monitoring programs but also suggest plants species that are likely to be most beneficial in restoration programs (Hegland et al. 2010; Elle et al. in press). That said, networks also allow specialist species to be identified. There were 12 oligolectic bee species collected in these habits: Andrena astagali, A. microchlora, Colletes consors, Duforea trochantera, Heriades cressoni, Megachile perihirta, M. umatillensis, Osmia californica, O. coloradensis, O. marginipennis, O. montana and Perdita fallax. Only antelope-brush habitat supports all 12 specialist species, as H. cressoni was collected exclusively in antelope-brush habitats and the floral hosts of M. umatillensis and P. fallax only grow in dry, low-elevation shrubsteppe. If preservation of specialist bees is a conservation goal, protection of antelope-brush habitat should be a priority.

Conclusions and future directions

I found that plant-pollinator networks in big sagebrush habitat may be more resilient to disturbance than those in the more critically endangered antelope-brush

shrubsteppe habitat, through a trend towards larger network size and significantly greater network-level generalization. Additionally, late-season communities may also be more resilient than those early in the season as they tended to be larger, were more asymmetric and had more generalized plant species. Comparing plant-pollinator network structure to investigate differences in resilience between threatened and endangered communities has the potential to contribute valuable information to conservation priority decision-making. Additionally, the species-level information that can be deduced from network analysis, such as assessing the functional importance of species, can provide useful information for habitat monitoring and restoration. It is, however, still early in the study of plant-pollinator interaction networks and much still needs to be learned about how strongly structural differences detected through network analysis translate into differences in network resilience in natural communities. Future research should focus on challenging network theory with empirical data to gain a better understanding of how network structural parameters respond to different natural and anthropogenic disturbances and what these responses mean functionally for real communities. The development of networks as research and conservation tools, which are capable of understanding both species- and community-level interactions, will be important for the conservation and sustainability of pollinations systems in the Okanagan and around the world.

References

Aizen, M. A., L. Ashworth and L. Galetto (2002). Reproductive success in fragmented habitats: do compatibility systems and pollination specialization matter? Journal of Vegetation Science 13(6): 885-892.

Aizen, M. A., C. L. Morales and J. M. Morales (2008). Invasive mutualists erode native pollination webs. Plos Biology 6(2): 396-403.

Albrecht, M., M. Riesen and B. Schmid (2010). Plant-pollinator network assembly along the chronosequence of a glacier foreland. Oikos 119(10): 1610-1624.

Almeida-Neto, M., P. Guimaraes, P. R. Guimaraes, R. D. Loyola and W. Ulrich (2008). A consistent metric for nestedness analysis in ecological systems: reconciling concept and measurement. Oikos 117(8): 1227-1239.

Bartomeus, I., M. Vila and L. Santamaria (2008). Contrasting effects of invasive plants in plant-pollinator networks. Oecologia 155(4): 761-770.

Bascompte, J. (2007). Networks in ecology. Basic and Applied Ecology 8(6): 485-490.

Bascompte, J. (2009). Mutualistic networks. Frontiers in Ecology and the Environment 7(8): 429-436.

Bascompte, J. and P. Jordano (2007). Plant-animal mutualistic networks: The architecture of biodiversity. Annual Review of Ecology Evolution and Systematics. Palo Alto, Annual Reviews. 38: 567-593.

Bascompte, J., P. Jordano, C. J. Melian and J. M. Olesen (2003). The nested assembly of plant-animal mutualistic networks. Proceedings of the National Academy of Sciences of the United States of America 100(16): 9383-9387.

Bascompte, J., P. Jordano and J. M. Olesen (2006). Asymmetric coevolutionary networks facilitate biodiversity maintenance. Science 312(5772): 431-433.

Basilio, A. M., D. Medan, J. P. Torretta and N. J. Bartoloni (2006). A year-long plant- pollinator network. Austral Ecology 31(8): 975-983.

Bersier, L. F., C. Banasek-Richter and M. F. Cattin (2002). Quantitative descriptors of food-web matrices. Ecology 83(9): 2394-2407.

Bezerra, E. L. S., I. C. S. Machado and M. A. R. Mello (2009). Pollination networks of oil- flowers: a tiny world within the smallest of all worlds. Journal of Animal Ecology 78: 1096-1101.

Bluthgen, N., F. Menzel and N. Bluthgen (2006). Measuring specialization in species interaction networks. BMC Ecology 6(9).

Bond, W. J. (1994). Do mutualisms matter? Assessing the impact of pollinator and disperser disruption on plant extinction. Philosophical Transactions of the Royal Society of London B Biological Sciences 344(1307): 83-90.

Bosch, J., A. M. M. Gonzalez, A. Rodrigo and D. Navarro (2009). Plant-pollinator networks: adding the pollinator's perspective. Ecology Letters 12(5): 409-419.

Carvalheiro, L. G., E. R. M. Barbosa and J. Memmott (2008). Pollinator networks, alien species and the conservation of rare plants: Trinia glauca as a case study. Journal of Applied Ecology 45(5): 1419-1427.

CDC, B. C. (2003). [Purshia tridentata / Hesperostipa comata Antelope brush / needle- and-thread grass]. Plant Community Account Summary [C3B4APTSC1]. B. C. M. o. S. R. Management. Victoria.

Chacoff, N. P., D. P. Vazquez, S. B. Lomascolo, E. L. Stevani, J. Dorado and B. Padron (2012). Evaluating sampling completeness in a desert plant-pollinator network. Journal of Animal Ecology 81(1): 190-200.

Curtin, C. and D. Western (2008). Grasslands, people, and conservation: Over-the- horizon learning exchanges between African and American pastoralists. Conservation Biology 22(4): 8.

Dormann, C. F., J. Frund, N. Bluthgen and B. Gruber (2009). Indices, graphs and null models: Analyzing bipartite ecological networks. The Open Ecology Journal 2: 7- 24.

Dupont, Y. L., D. M. Hansen and J. M. Olesen (2003). Structure of a plant-flower-visitor network in the high-altitude sub-alpine desert of Tenerife, Canary Islands. Ecography 26(3): 301-310.

Elle, E., S. Elwell and G. Gielens (in press). The use of pollination networks in conservation. Botany.

Elmqvist, T., C. Folke, M. Nystrom, G. Peterson, J. Bengtsson, B. Walker and J. Norberg (2003). Response diversity, ecosystem change, and resilience. Frontiers in Ecology and the Environment 1(9): 488-494.

Elton, C. S. (1958). Ecology of invasions by animals and plants. London, Chapmen & Hall.

Fortuna, M. A. and J. Bascompte (2006). Habitat loss and the structure of plant-animal mutualistic networks. Ecology Letters 9(3): 278-283.

Forup, M. L., K. S. E. Henson, P. G. Craze and J. Memmott (2008). The restoration of ecological interactions: plant-pollinator networks on ancient and restored heathlands. Journal of Applied Ecology 45(3): 742-752.

Gayton, D. (2003). British Columbia Grasslands Monitoring Vegetation Change, FORREX - Forest Research Extension Partnership. 7.

Gibson, R. H., B. Knott, T. Eberlein and J. Memmott (2011). Sampling method influences the structure of plant-pollinator networks. Oikos 120(6): 822-831.

Gibson, R. H., I. L. Nelson, G. W. Hopkins, B. J. Hamlett and J. Memmott (2006). Pollinator webs, plant communities and the conservation of rare plants: arable weeds as a case study. Journal of Applied Ecology 43(2): 246-257.

Gilgert, W. and M. Vaughan (2011). The value of pollinators and pollinator habitat to rangelands: Connections among pollinators, insects, plant communities, fish, and wildlife. Rangelands 33(3): 6.

Goulson, D., G. C. Lye and B. Darvill (2008). Decline and conservation of bumble bees. Annual Review of Entomology. Palo Alto, Annual Reviews. 53: 191-208.

Hagen, M. and M. Kraemer (2010). Agricultural surroundings support flower-visitor networks in an Afrotropical rain forest. Biological Conservation 143(7): 1654- 1663.

Hegland, S. J., J. Dunne, A. Nielsen and J. Memmott (2010). How to monitor ecological communities cost-efficiently: The example of plant-pollinator networks. Biological Conservation 143(9): 2092-2101.

Henle, K., K. F. Davies, M. Kleyer, C. Margules and J. Settele (2004). Predictors of species sensitivity to fragmentation. Biodiversity and Conservation 13(1): 207- 251.

Joppa, L. N., J. M. Montoya, R. Sole, J. Sanderson and S. L. Pimm (2010). On nestedness in ecological networks. Evolutionary Ecology Research 12(1): 35-46.

Jordano, P., J. Bascompte and J. M. Olesen (2006). The ecological consequences of complex topology and nested structure in pollination webs. Plant-Pollinator Interactions: From Specialization to Generalization. N. M. Waser and J. Ollerton. Chicago, The University of Chicago Press: 173-199.

Kaiser-Bunbury, C. N., S. Muff, J. Memmott, C. B. Mueller and A. Caflisch (2010). The robustness of pollination networks to the loss of species and interactions: a quantitative approach incorporating pollinator behaviour. Ecology Letters 13(4): 442-452.

Kearns, C. A., D. W. Inouye and N. M. Waser (1998). Endangered mutualisms: The conservation of plant-pollinator interactions. Annual Review of Ecology and Systematics 29: 83-112.

Kevan, P. G. (1999). Pollinators as bioindicators of the state of the environment: species, activity and diversity. Agriculture Ecosystems & Environment 74(1-3): 373-393.

Krannitz, P. G. (2008). Response of antelope bitterbrush shrubsteppe to variation in livestock grazing. Western North American Naturalist 68(2): 138-152.

Lawton, J. H. and V. K. Brown (1993). Redundancy in ecosystems. Biodiversity and ecosystem function. E. D. Schulze and H. A. Mooney. 99: 255-270.

Lea, T. (2008). Historial (pre-settlement) ecosystems of the Okanagan Valley and Lower Similkameen Valley of British Columbia - pre-European contact to the present. Davidsonia 19(1): 34.

Lopezaraiza-Mikel, M. E., R. B. Hayes, M. R. Whalley and J. Memmott (2007). The impact of an alien plant on a native plant-pollinator network: an experimental approach. Ecology Letters 10(7): 539-550.

Macarthur, R. (1955). Fluctuations of animal populations, and a measure of community stability. Ecology 36(3): 533-536.

Mack, R. N. (1981). Invasion of Bromus tectorum L into western North America - An ecological chronicle. Agro-Ecosystems 7(2): 145-165.

McCann, K. S. (2000). The diversity-stability debate. Nature 405(6783): 228-233.

Medan, D., A. M. Basilio, M. Devoto, N. H. Bartoloni, J. P. Torretta and T. Petanidou (2006). Measuring generalization and connectance in temperate, year-long active systems. Plant-pollinator interactions: From specialization to generalization. N. M. Waser and J. Ollerton. Chicago, The University of Chicago Press: 245-259.

Memmott, J. (1999). The structure of a plant-pollinator food web. Ecology Letters 2(5): 276-280.

Naeem, S. (1998). Species redundancy and ecosystem reliability. Conservation Biology 12(1): 39-45.

Nicholson, A., E. Hamilton, W. L. Harper and B. M. Wikeem (1991). Bunchgrass Zone. Ecosystesm of British Columbia. D. Meidinger, Pojar, J. Victoria, B.C, Province of British Columbia.

Olesen, J. M., J. Bascompte, H. Elberling and P. Jordano (2008). Temporal dynamics in a pollination network. Ecology 89(6): 1573-1582.

Olesen, J. M., L. I. Eskildsen and S. Venkatasamy (2002). Invasion of pollination networks on oceanic islands: importance of invader complexes and endemic super generalists. Diversity and Distributions 8(3): 181-192.

Ollerton, J., R. Winfree and S. Tarrant (2011). How many flowering plants are pollinated by animals? Oikos 120(3): 321-326.

Peart, B. (2008). A record of The World Temperate Grasslands Conservation Initiative Workshop Hohhot, China - June 28 & 29, 2008. Vancouver, BC, Temperate Grasslands Conservation Initiative.

Potts, S. G., J. C. Biesmeijer, C. Kremen, P. Neumann, O. Schweiger and W. E. Kunin (2010). Global pollinator declines: trends, impacts and drivers. Trends in Ecology & Evolution 25(6): 345-353.

Schlute, A., T. Lea, S. Cannings and P. G. Krannitz (1995). Antelope-brush ecosystems. L. a. P. B.C. Ministry of Environment, Wildlife Branch. Victoria, B.C.: 6.

Sugiura, S. (2010). Species interactions-area relationships: biological invasions and network structure in relation to island area. Proceedings of the Royal Society B- Biological Sciences 277(1689): 1807-1815.

Thebault, E. and C. Fontaine (2010). Stability of ecological communities and the architecture of mutualistic and trophic networks. Science 329(5993): 853-856.

Tilman, D., D. Wedin and J. Knops (1996). Productivity and sustainability influenced by biodiversity in grassland ecosystems. Nature (London) 379(6567): 718-720.

Tscharntke, T., I. Steffan-Dewenter, A. Kruess and C. Thies (2002). Characteristics of insect populations on habitat fragments: A mini review. Ecological Research 17(2): 229-239.

Tylianakis, J. M., E. Laliberte, A. Nielsen and J. Bascompte (2010). Conservation of species interaction networks. Biological Conservation 143(10): 2270-2279.

Tylianakis, J. M., T. Tscharntke and O. T. Lewis (2007). Habitat modification alters the structure of tropical host-parasitoid food webs. Nature 445(7124): 202-205.

Ulrich, W., M. Almeida and N. J. Gotelli (2009). A consumer's guide to nestedness analysis. Oikos 118(1): 3-17.

Valdovinos, F. S., R. Ramos-Jiliberto, J. D. Flores, C. Espinoza and G. Lopez (2009). Structure and dynamics of pollination networks: the role of alien plants. Oikos 118(8): 1190-1200.

Vazquez, D. P. and M. A. Aizen (2004). Asymmetric specialization: A pervasive feature of plant-pollinator interactions. Ecology 85(5): 1251-1257.

Vazquez, D. P., N. Bluethgen, L. Cagnolo and N. P. Chacoff (2009). Uniting pattern and process in plant-animal mutualistic networks: a review. Annals of Botany 103(9): 1445-1457.

Vazquez, D. P., C. J. Melian, N. M. Williams, N. Bluthgen, B. R. Krasnov and R. Poulin (2007). Species abundance and asymmetric interaction strength in ecological networks. Oikos 116(7): 1120-1127.

Vazquez, D. P., W. F. Morris and P. Jordano (2005). Interaction frequency as a surrogate for the total effect of animal mutualists on plants. Ecology Letters 8(10): 1088-1094.

Vazquez, D. P. and D. Simberloff (2003). Changes in interaction biodiversity induced by an introduced ungulate. Ecology Letters 6(12): 1077-1083.

Vila, M., I. Bartomeus, A. C. Dietzsch, T. Petanidou, I. Steffan-Dewenter, J. C. Stout and T. Tscheulin (2009). Invasive plant integration into native plant-pollinator networks across Europe. Proceedings of the Royal Society B-Biological Sciences 276(1674): 3887-3893.

Vilhena, A., L. Rabelo, E. Bastos and S. Augusto (2012). Acerola pollinators in the savanna of Central Brazil: temporal variations in oil-collecting bee richness and a mutualistic network. Apidologie 43(1): 51-62.

Waser, N. M., L. Chittka, M. V. Price, N. M. Williams and J. Ollerton (1996). Generalization in pollination systems, and why it matters. Ecology 77(4): 1043- 1060.

Wikeem, B. and S. Wikeem (2004). The Grasslands of British Columbia. Kamloops, B.C., Grasslands Conservation Council of British Columbia.

Winfree, R. and C. Kremen (2009). Are ecosystem services stabilized by differences among species? A test using crop pollination. Proceedings of the Royal Society B-Biological Sciences 276(1655): 229-237.

Tables

Table 3.1. Plant-pollinator interaction network property definitions with brief explanations of their influence on network resilience.

Site Shrubsteppe Area Elevation Slope abbrev. Site name type (ha) (m) (%) HLU Haynes Lease Ecological Antelope-brush 50 337 9 Reserve HLG Haynes Lease -Calf pasture Antelope-brush 43 314 7

OKU Kennedy Bench Antelope-brush Antelope-brush 40 448 3 Conservation Area OKG Mt. Oliver Protected Area Antelope-brush 260.5 503 5

SOU White Lake Biodiversity Ranch Big sagebrush 55 713 7

SOG White Lake Biodiversity Ranch Big sagebrush 170 563 15

WLG Southern Okanagan Grasslands Big sagebrush 20 883 22 Protected Area WLU Southern Okanagan Grasslands Big sagebrush 1850 884 18 Protected Area

Full season Seasonal sub-networks networks Habitat Season Habitat*Season Habitat

F1,6 P F2,12 P F2,12 P F1,6 P Network size 4.07 0.0903 2.87 0.0956 0.09 0.9189 3.62 0.1057

Number of plant spp. 4.18 0.0867 7.87 0.0066* 2.98 0.0891 2.89 0.1401

Number of pollinator spp. 3.97 0.0934 5.01 0.0261* 0.04 0.9575 3.74 0.1012

Plant generality 6.18 0.0474* 15.52 0.0005* 5.37 0.0216* 6.37 0.0451*

Pollinator generality 4.74 0.0724 1.70 0.2235 2.67 0.1099 4.60 0.0757

Specialization 6.36 0.0452* 0.05 0.9479 0.25 0.7839 9.15 0.0232*

Asymmetry 0.06 0.8097 21.33 0.0001* 0.79 0.4764 0.61 0.4634

Nestedness 0.89 0.3829 1.18 0.3394 5.65 0.0186* 0.67 0.4450

Figures

Figure 3.1. Quantitative plant-pollinator interaction networks from antelope- brush and big sagebrush habitats: a/e) Full season networks; b/f) Early season networks; c/g) Middle season networks; and d/h) Late season networks. In each network, rectangles represent pollinator (top row) or plant (bottom row) species, and the lines connecting them represent interactions. The width of each plant rectangle represents how frequently the plant was visited by pollinators, and the width of each pollinator rectangle indicates how frequently a pollinator was collected off of flowering plants. The width of the interaction represents how frequently that interaction was recorded. Pollinators are colour-coded as follows: red = bees (Hymenoptera); green = wasps (Hymenoptera); blue = flies (Diptera); purple = beetles (Coleoptera); yellow = butterflies (Lepidoptera); orange = hummingbird (Trochilidae). Plants in the seasonal sub-networks are colour-coded as follows: light grey = blooming in early and mid season; dark grey = blooming in mid and late season; black = blooming during a single season. Species blooming through two seasons are arranged in the same order to allow comparison. Networks are meant to give an impression of how network interactions change through time, and are not all drawn to the same scale.

120 140 a) Antelope-brush b) Plant 100 Pollinator 120 Big sagebrush

100 80

80 60

60 40

Network sizeNetwork 40 20

Number of speciesNumber 20 0 0.80 c) d) 0.75 10 Plant Pollinator 0.70 8 0.65

0.60 6

0.55 4

0.50 Generality

specialization index 0.45 2

H2' 0.40 -0.25 14 e) f) -0.30 12

-0.35 10 -0.40 8 -0.45 6

Asymmetry -0.50

Nestedness

-0.55 4

-0.60 2 Early Mid Late Full Early Mid Late Full

Network Network

Chapter 4

General conclusions

The reports of pollinator population declines that have surfaced in many places around the world have raised concern over the health of pollinator communities and the preservation of their functional roles (Kearns et al. 1998; Potts et al. 2010). Although still rarely considered in the conservation planning process, pollinators are a vitally important component of most terrestrial ecosystems (Winfree 2010). As anthropogenic pressures on natural ecosystems continue to increase, understanding how habitat-altering disturbances influence pollinator communities will be important for their future conservation and preservation. In this thesis, I assessed the effects of livestock grazing on shrubsteppe flowering plant and pollinator communities (Chapter 2), and used plant- pollinator interaction networks to compare network structure between British Columbia’s endangered shrubsteppe ecosystems (Chapter 3), to investigate network resilience and generate information useful for conservation planning.

The effects of livestock grazing and habitat type on flowering plants and pollinators

Previous studies have shown that the abundance and richness of pollinator populations can be influenced by changes in vegetation structure induced by grazing, such as vegetation height (Kruess and Tscharntke 2002) and bare soil availability (Vulliamy et al. 2006), thus I assessed whether shrubsteppe vegetation structure was influenced by grazing. I found that livestock grazing did affect vegetation structure, by increasing the cover of shrubs and bare soil and decreasing the height of the grass and forb layers. Similar responses to livestock grazing have been reported by many other studies (Anderson et al. 1982; Fleischner 1994; Jones 2000; Kruess and Tscharntke 2002; Vulliamy et al. 2006; Krannitz 2008). Cattle often cause decreases in grass and

forb height through direct herbivory and/or reductions in plant vigor due to herbivory stress (Pond 1960; Fleischner 1994; Krannitz 2008). Additionally, trampling by cattle often increases bare soil availability (Fleischner 1994). In these shrubsteppe ecosystems increases in bare soil come at the expense of cryptogamic crust cover, a layer important for soil moisture retention. Spring grazing is reputed to be the most destructive time of year to graze dry shrubsteppe habitat as it is the primary growing season for many grass, forb and cryptogamic crust species (Gayton 2003; Krannitz 2008).

Contrary to my expectations, the changes in vegetation structure imposed by livestock grazing did not extend to significant changes in flowering plant or pollinator abundance, richness or community composition. There was a trend of decreased floral abundance in grazed sites during the later-half of the flowering season (June-July), roughly corresponding with the cessation of grazing. It may be that forbs subject to cattle herbivory tend to produce fewer flowers, or that grazing decreases the abundance of some species. There was also a trend towards differing pollinator community composition between grazed and ungrazed sites. Thus, livestock may have had some influence on some members of the plant-pollinator community, but the duration and intensity of grazing precluded any significant negative effects on the community as a whole. It is predicted that livestock grazing in ecosystems without long grazing histories, such as those west of the Rocky Mountains in North America (including my study sites), will be harmful to plants and pollinators. But my work, as well as that of Vazquez et al. (2008) from Argentina, shows that floral and pollinator communities of habitats without long grazing histories do not necessarily respond negatively to grazing pressure. My results therefore contribute to the growing body of literature indicating that the current grazing regime is an important determinant of plant and pollinator responses (e.g., Pond 1960; Carvell 2002; Vulliamy et al. 2006; Krannitz 2008; Sjodin et al. 2008; Xie et al. 2008).

Floral and pollinator communities were significantly correlated at my sites, suggesting that pollen and nectar resources (i.e. floral community composition) were a major determinant of pollinator community composition. Thus, it is perhaps not surprising that the non-significant effects of grazing on the flowering plant community were carried through to the pollinator community. My results suggest that semi-natural

habitats, like rangeland, when managed responsibly, can remain reservoirs of flowering plant and pollinator diversity.

Both antelope-brush and big sagebrush shrubsteppe were found to have diverse plant and pollinator communities, though they differed in their community composition. Differences in flowering plant community composition between habitats is likely driven by environmental differences associated with elevation change, as antelope-brush habitats are drier and lower in elevation than big sagebrush habitats. Although many flowering plant species were present in both ecosystems, many others were sampled primarily or exclusively at low or high elevations. A similar pattern was found with pollinators. Bee genera that were found to be more prevalent in one shrubsteppe type were found to frequently visit flowering plant species with a similar distribution. For example, the bee genera Andrena and Eucera which were more prevalent in big sagebrush habitat, primarily visited flowering plants found only (or more prevalently) in big sagebrush habitat. As both antelope-brush and big sagebrush shrubsteppe support high pollinator diversity, but the composition of their floral and pollinator communities differ, attention to maintaining the health of both habitats under sustainable grazing practices is of high conservation importance.

The plant-pollinator network structure of British Columbia’s endangered shrubsteppe

Given the endangered status of both shrubsteppe habitats and the differences in their plant and pollinator community composition, I also investigated differences in plant- pollinator network structure between habitats. I found that plant-pollinator networks of the two shrubsteppe habitats were different in the average generalization of their constituent species. Big sagebrush networks were significantly more generalized overall, with more generalized plants and a trend towards more generalized pollinators. Additionally, big sagebrush networks tended to be larger (more species-rich). Networks with more interacting species that are, on the whole, more generalized are thought to be more stable and resilient to disturbances, such as habitat alteration, through increased interaction redundancy (Memmott et al. 2004; Thebault and Fontaine 2010). For example, plants that are visited by many pollinator species should be less likely to

experience population decline or local extinction from the loss of a pollinator than those species that interact with few. Network analysis also indicated that big sagebrush and antelope-brush habitats shared roughly half of their top-ten most functionally important flowering plants and pollinators. Thus, protection of one habitat is likely to be beneficial to the plant-pollinator communities of the other; protecting and restoring big sagebrush habitat near remaining antelope-brush fragments could reduce the negative effects associated with this habitats excessive fragmentation. As was found in Chapter 2, network analysis also indicated that many species are more prevalent or only found in one shrubsteppe type, thus preservation of both habitats will be important in maintaining regional diversity.

I also used plant-pollinator networks to investigate how community interactions changed across the flowering season. Plant-pollinator networks late in the flowering season tended to be larger, were more asymmetric, and had greater plant generalization than those in the spring. Thus, plants flowering late in the season should be less susceptible to fluctuations in population sizes of their pollinators than those flowering early in the season, because they have more pollinators to rely on for pollen receipt (Waser et al. 1996). Although I found that low-intensity spring grazing did not negatively affect these plant and pollinator communities, my network analysis suggests that the potential impact of higher-intensity grazing in this region could be minimized if it can be shifted away from early spring to later in the flowering season. Comparing network structural properties temporally, although rarely pursued in a conservation context, may be a promising additional component of using plant-pollinator networks to address applied ecological questions.

Plant-pollinator interaction networks can provide a more functional perspective of communities than traditional biodiversity sampling, by identifying which species interact within a community, how those interactions are structured and what that structure may mean for community stability (Bascompte and Jordano 2007). However, there is still much to be learned about plant-pollinator community structure and network analysis, particularly in relation to their use as a tool for management and conservation (Tylianakis et al. 2010; Elle et al. in press). Future research should work towards gaining a deeper understanding of the functional consequences of network structure in real communities and should focus on identifying what network structural properties are influenced by

different anthropogenic disturbances. Although a push towards this aim can be seen in the current literature (e.g., Forup and Memmott 2005; Aizen et al. 2008; Bartomeus et al. 2008; Yoshihara et al. 2008; Hagen and Kraemer 2010), there is still much to be understood about practically applying the information gained from networks in conservation planning (Tylianakis et al. 2010; Elle et al. in press). Additionally, identifying how plant-pollinator networks can be sampled effectively, but also cost efficiently, is necessary if they are to be widely adopted as a management tool (Hegland et al. 2010; Tylianakis et al. 2010). These are laudable aims because the development of networks, which improve understanding of both species- and community-level interactions, as research and conservation tools will be important for the conservation and sustainability of pollinations systems.

Summary and future directions

Low-intensity, spring livestock grazing does not negatively affect plant and pollinator abundance, richness or community composition, although trends towards decreased floral abundance and differing pollinator community composition were found in grazed sites. My results suggest that rangelands can maintain grassland flowering plant and pollinator diversity when responsibly managed. This is heartening news for pollinator conservation given the ever-increasing threat of human-induced disturbance. However, given the trends observed, I recommend that the current grazing regimes of these areas be maintained and not increased. Although some aspects of vegetation structure are influenced by low-intensity spring grazing, the disturbance is minor from the flowering plant and pollinator perspective, suggesting the grazing regimes implemented at these sites could act as a model for private land owners in the region. Additionally, since network analysis indicated early-season plant-pollinator communities may be less resilient to disturbance than late season communities, the potential impacts of higher- intensity grazing could be minimized if it can be shifted away from early spring to later in the flowering season.

Grasslands are among the ecosystems predicted to experience the largest losses in biodiversity over the next century, particularly due to their sensitivity to land- use change (Sala et al. 2000), thus community-based monitoring and analyses of plants

and pollinators in the southern Okanagan is likely to be important for their conservation. Although the monitoring of both taxa would be preferable, under monetary constraints monitoring flowering plant diversity and community composition may be a decent surrogate for monitoring pollinator communities. Flowering plant and pollinator community compositions were correlated across my sites and network analysis indicated that pollinator species tended to rely more strongly on the plant species they visited for pollen and nectar than the plants relied on them, on an individual species basis, for pollination. Thus, if shrubsteppe flowering plant communities are doing well, it is likely that the pollinator communities are also doing well. That said, long-term data sets on pollinator populations are few (Potts et al. 2010) and need to be initiated now to better assess the changes in pollinator populations over the next few decades due to increased habitat loss, alteration and climate change. I believe the southern Okanagan is an appropriate region to begin a pollinator monitoring program within B.C. because of the region’s high pollinator diversity, endangered ecosystems and increasing urban and agricultural development. Due to the overlap of early and late season pollinators, data collection during mid flowering season (late May - early June) would likely be sufficient for such a monitoring program.

Both antelope-brush and big sagebrush shrubsteppe of the southern Okanagan Valley are not only Red-listed in British Columbia, but are considered globally imperilled (B.C. Ministry of Environment) thus effective conservation of these habitats and their flowering plant and pollinator communities should be a Canadian conservation priority. The conservation of flowering plants and pollinators will be critical to the successful preservation of both shrubsteppe habitats, as so many other species depend on flowering plants and insects for food resources and vegetative habitat structure (Gilgert and Vaughan 2011). Given that the plant-pollinator communities of antelope-brush habitat may be more sensitive to disturbance than those in big sagebrush habitat and given the highly fragmented state of remaining antelope-brush shrubsteppe, prioritizing the monitoring, restoration and conservation of remaining antelope-brush habitat will be highly important in preserving biodiversity in the southern Okanagan region. Big sagebrush shrubsteppe, although also Red-listed, is approximately twice as abundant as antelope-brush shrubsteppe (Lea 2008) and is also far less fragmented. The plant- pollinator communities of big sagebrush shrubsteppe tend to be more diverse and may

be more resilient to disturbance through increased network-level generalization than those in antelope-brush shrubsteppe. Large contiguous tracts of big sagebrush habitat still exist in the Okanagan and, although much of this habitat is grazed by livestock attention towards responsible grazing practices, such as those studied here, will be one of the most important factors in maintaining the integrity of this ecosystem. The proposal to create a national park in the south Okanagan – lower Similkameen Valleys, which would have enforced adaptive management of livestock grazing within park boundaries and connected many present day protected areas (Parks Canada 2010), is not currently supported by the government of British Columbia (Parks Canada 2012). Thus, the continued effort of land managers and conservation practitioners to use community- based monitoring and analyses to find a balance between biological integrity and economic viability in this region will be vital for the conservation of shrubsteppe pollination systems.

References

Aizen, M. A., C. L. Morales and J. M. Morales (2008). Invasive mutualists erode native pollination webs. Plos Biology 6(2): 396-403.

Anderson, D. C., K. T. Harper and S. R. Rushforth (1982). Recovery of cryptogramic soil crusts from grazing on Utah winter ranges. Journal of Range Management 35(3): 355-359.

Bartomeus, I., M. Vila and L. Santamaria (2008). Contrasting effects of invasive plants in plant-pollinator networks. Oecologia 155(4): 761-770.

Bascompte, J. and P. Jordano (2007). Plant-animal mutualistic networks: The architecture of biodiversity. Annual Review of Ecology Evolution and Systematics. Palo Alto, Annual Reviews. 38: 567-593.

B.C. Ministry of Environment. B.C. Species and Ecosystem Explorer. Retrieved Februrary 23, 2012 from http://www.env.gov.bc.ca/atrisk/red-blue.htm.

Carvell, C. (2002). Habitat use and conservation of bumblebees (Bombus spp.) under different grassland management regimes. Biological Conservation 103(1): 33-49.

Elle, E., S. Elwell and G. Gielens (in press). The use of pollination networks in conservation. Botany.

Fleischner, T. L. (1994). Ecological costs of livestock grazing in western North America. Conservation Biology 8(3): 629-644.

Forup, M. L. and J. Memmott (2005). The restoration of plant-pollinator interactions in hay meadows. Restoration Ecology 13(2): 265-274.