Wildlife Use of Native Perennial and Exotic Annual Grasslands: A Comparative Study

Hi there! Here I will discuss the general experimental design and questions I am asking with my study comparing wildlife use of paired restored native perennial grasslands (hereafter, “restored”) and unrestored exotic annual grasslands (hereafter “control”).

Introduction and Significance

Intact native grasslands are some of the rarest ecosystems in the world, and wildlife species associated with grasslands have declined precipitously over the last 200 years (Sampson and Knopf 1994). California’s grasslands are one of the most invaded rangeland systems in the world, with native species replaced almost entirely by annual Mediterranean forbs and grasses (Heady et al. 1992, Hamilton et al. 2002). Grassland restoration is currently being undertaken to restore biodiversity and ecosystem services that have been lost due to invasion by exotic weedy species, including floral and faunal native species richness, forage for livestock, and aesthetic value. However, restoration projects are expensive and time-consuming, and wildlife monitoring following restoration is not often conducted due to limited resources (Morisson 2002, Gill et al. 2006). For example, many hundreds of hectares of native grassland in Northern California have been restored for the purpose of habitat and forage for the threatened Swainson’s hawk (Buteo swainsonii). Swainson’s hawks feed on rodents such as ground squirrels, gophers, and field mice, as well as grasshoppers, locusts, and other insects that live in open fields. High quality foraging habitat in close proximity to nesting sites is particularly important for these birds, but habitat fragmentation and destruction has reduced suitable nesting sites adjacent to native grasslands (Detrich 1986). Hundreds of additional acres are slated for restoration, but neither restoration contractors nor wildlife agencies know whether past efforts have sustained or bolstered hawk populations by increasing suitable foraging habitat, and follow-up monitoring is not currently planned (Jennifer Burt, AECOM, personal communication 2013). In most cases when monitoring is conducted, only one or a few taxa are included on a limited spatiotemporal basis, the results of which may not be applicable across years or locations (Hurlbert 1984, Hayes and Holl 2003).

In California the diverse assemblage of wildlife that reflects the state’s great variety of climates and habitats is now greatly reduced in abundance and diversity (Jameson and Peeters 2002), largely due to habitat destruction and fragmentation (Noss et al. 1995, Aronson and Falk 2002, Krausman 2002). At a global scale, studies have shown that habitat loss is a greater immediate threat to biodiversity than climate change (Sala et al. 2000, Jetz et al. 2007). One of the goals of habitat restoration and management is the provision of wildlife habitat and increased biodiversity, which theoretically extends to impacts on food webs (Vander Zanden et al. 2006). Beschta and Ripple (2009) concluded, for example, that the removal of large predators in the western United States has resulted in drastic changes to native plant communities due to the release of large ungulates and a switch to alternative stable states, and that restoration of native flora is necessary for the recovery of former ecosystem services. Moreover, while the restoration of historic plant community assemblages is assumed to provide increased resource availability for other trophic levels, how well this objective is met is rarely monitored (Boyd and Svejcar 2009). As budgets constrict and demands for accountability and results heighten, it becomes imperative that the presumed positive relationship between restoration and ecosystem services be better documented (Benayas et al. 2009). More information is needed regarding wildlife needs, species abundances and distributions; vegetation and habitat requirements; and impacts of exotic species on wildlife (Morrison 2002).

Behaviors, distribution, and resource needs of wildlife may change dramatically throughout the year – especially in regions with strong seasonal signals, like California’s Mediterranean climate – and this variation is often overlooked in research. Many studies commonly use averaged values, or sample only during very narrow time periods, which are insensitive to seasonal or yearly changes in wildlife abundance or habitat use (Schooley 1994). Lumping species into guilds may also mask species differences in response to restoration (Majer 2009). Without species-specific monitoring of these changes over time, restoration plans may fall short of their targets, and could fail to account for important unintended interactions between different trophic levels and species. For example, Siemann (2007) found that experimental manipulation of plant community composition directly controlled diversity of consumers by determining available forage resources, but also indirectly affected diversity via herbivore-parasite-predator interactions.

Some studies have also shown fine-scale spatial structure of vegetation to be particularly important in determining patterns of habitat occupancy (see review by Morrison et al. 1998). However, scientific evidence backing increased wildlife habitat as it correlates with native plant restoration still lags behind single-focused monitoring of vegetation (Majer 2009). The invasion by Mediterranean annual grasses results in a sweeping transition from spatially and physically diverse native grasslands with patches of bunchgrasses and open interspaces to simplified structural monocultures of annual grasses and forbs (Heady 1977, Jackson 1985). Spatial heterogeneity (or lack thereof) in plant communities is an important driver of biodiversity and should be incorporated into empirical studies to assist in the explanation of observed responses (Seabloom et al. 2005). While restoration of native plant species in and of itself is inherently valuable for the conservation and protection of native flora, habitat preference may be more influenced by vegetation structure than community composition. Gill et al. (1986) found that community structure affected shrub-land and grassland bird species behavior more so than plant species composition. However, Goerrison (1996) found that grassland specialist birds responded negatively to invasion by non-native plants. However, the majority of birds observed in grasslands are not obligate grassland species, but do utilize grasslands to forage. Thus, grasslands are an important habitat resource for non-obligate bird species, and understanding how raptors respond to grassland structure and composition is necessary for managing natural landscapes as raptor foraging grounds.

Finally, restoration projects commonly sample only at a single one location thereby limiting inference (Morrison et al. 1998, Morrison 2002). It is important, insofar as resources are available, to replicate studies at multiple sites within the area of interest to test for the generality of results (Hayes and Holl 2003). Large-scale restorations – like those utilized in this study – represent unique natural experiments (Majer 2009). Paired restored and unrestored grassland restorations represent powerful replicated natural experiments in which to link structural and compositional heterogeneity with wildlife biodiversity.  The current and proposed components of this research are therefore unique in being multi-season, multi-species, and multi-trophic across four landscape-scale locations in California’s heavily invaded Central Valley, where less than 2% of native grasslands still exist somewhat intact (Huenneke 1989, Noss et al. 1995, Olson and Cox 2014). In addition, causal mechanisms (horizontal and vertical vegetation structure, availability and accessibility of nesting sites, food, and other resources) underlying wildlife responses to restoration (or invasion) have not been identified. This research will identify and inform not only future restorations, but current grassland management, both restored and unrestored.

Specifically, this research addresses information needs in California grasslands research, and will:

  • Compare relative utilization and diversity at paired restored and unrestored grasslands for a suite of interacting wildlife species.
  • Describe relationships between multiple trophic levels in each grassland habitat.
  • Identify mechanisms driving wildlife responses to structural and community composition.
  • Determine prevalence of zoonotic pathogens carried by in-situ vectors in these systems.

Study Design and Methods

The current proposed research will entail detailed monitoring of relative wildlife utilization and diversity in paired restored (native perennial) grasslands, and unrestored (exotic annual) grasslands (i.e., “controls”). This proposal builds upon two seasons of vertebrate mammal monitoring in paired grasslands at four locations across the Central Valley (Table 1).

Sampling Location City County GPS Coordinates
Bufferlands Elk Grove Sacramento 38º 26’ N, 121º 29’ W
Citrona Farms Winters Yolo 38º 39’ N, 122º 00’ W
Long Ranch Zamora Yolo 38º 50’ N, 121º 55’ W
Yolo Land & Cattle Co. Esparto Yolo 38º 36’ N, 121º 03’ W

These locations were chosen for the availability of well-paired restored and unrestored (control) sites with similar soil types, land-use history, management. Paired sites within each location have similar livestock grazing regimes, and owners cooperate in the removal of livestock grazing during future trapping and survey periods. All data will be collected over a 30-day period in each season: April (spring), July (summer), October (fall), and January (winter) over a one- to two-year period.

To monitor multiple species, a combination of live trapping, camera trapping, and visual encounter surveys (VES) began in April 2014, while raptor surveys and monitoring for fire ants began in July 2014 (Table 2).

Methods Groups Monitored
Capture/mark/release (Sherman live traps) Mice, voles, other small rodent species
Visual encounter surveys Voles, mice, snakes, lizards amphibians, fire ants
Camera traps Lagomorphs, wild ungulates (deer), mesopredators (opossum, raccoon, bobcat), large predators (coyote, mountain lion)
Timed surveys Raptors

Small rodents are caught in small Sherman live traps, sexed, weighed, identified to species, and eartagged with a small self-piercing tag (Nietfeld et al. 1994). Twenty-five fecal samples from each location per season (over four seasons) will be collected beginning in October 2014 to test for the presence of Cryptosporidium and Giardia in the non-native house mouse, Mus musculus, and the native deer mouse, Peromyscus maniculatus. A total of 400 fecal samples will be randomly selected and distributed between the two site types. These samples will be stored in vials and refrigerated until analysis, which will occur within one week of collection. Rodents are then immediately released, with the entire process taking less than three minutes. Live traps are set at dusk and opened at dawn to minimize exposure and reduce stress.

Camera traps are effective for surveying both large and small mammals (Trolle et al. 2007, De Bondi et al. 2010), can be applied over long time periods, may capture elusive species (Kelly 2008), and eliminates the need to handle animals (Silveira et al. 2003, Kelly and Holub 2008). For rodent species, camera traps are baited with non-toxic wax blocks designed for monitoring small rodents, and these cameras often opportunistically capture snakes, lizards, and grassland birds. Blocks serve as an attractant to remote-triggered cameras which are set to record visits by rodent species.

Coverboards of two materials (plywood and galvanized metal roofing) have been present on site since January 2014 and are monitored across all locations in the morning at least weekly to conduct visual encounter surveys (VES) for reptiles, amphibians, and small rodents. This provides two habitat choices that may be preferred by different species in different seasons. A total of 12 coverboard points are located at each site and 24 at each location, for a total of 96 points across all locations. Voles (Microtus californicus) utilize coverboards at relatively high rates, and VES appears to be a more reliable method for surveying voles than live trapping methods, which have had extremely low capture rates (<<1%). Coverboards are surveyed between six to eight times during each 30-day sampling period. VES is relatively inexpensive, easy to implement, limits site and wildlife disturbance, and allows for detection of multiple species simultaneously (Fellers and Drost 1994, Kinkead 2006, Mills et al. 2013).

Timed surveys (1 hour per site) to monitor raptor presence by species, diversity, and foraging behavior were conducted for five days at each location in July 2014. Surveyors stood at the edge of each site type and observed raptors with binoculars to determine species, and recorded time spent in the habitat, time spent foraging, number of attacks, and number of apparently successful attacks.

The number of live traps infested with fire ants will be used as an estimate of fire ant abundance within each site. All traps will be inspected at dusk when baited, and recorded if infested. Number of infested traps can be compared between the two sites to determine if invasion is higher in unrestored or restored grasslands. Fire ants eat both seeds and may kill small rodents, thereby indirectly and directly impacting rodent populations. Additionally, data from trapping the following morning can be used to determine if small rodents avoid infested traps (Allen et al. 1994, Holtcamp et al. 1997), as was observed for several mice trapped in the summer of 2014.

For live and camera trapping methods, four 150-m transects are located at each site, with live traps every 15 meters (11 traps/transect) and several camera traps every 50 m aimed at various angles to capture both rodents and larger species (8 cameras/transect). Transects run approximately parallel to each other and are separated by 50 m to allow for independence between transects for at least small rodents, as validated by data from the first two seasons. Paired coverboards are located along a parallel transect approximately 5 meters from trapping transects, with three pairs of coverboards at 75 m intervals along each transect. Averaging seven trapping days and/or nights per season and site over the two-year period allows for approximately 20,000 live trap nights, and 14,000 camera trap days and nights each. Between 9,000 and 12,000 coverboard surveys will be conducted over a two-year period. Adding in two-hour raptor surveys for seven days at each site for at least four additional seasons would add over 3,000 survey hours to the current survey effort of 200 hours. Continued monitoring of fire ant invasion will result in 14,784 samples. Order of live trapping and surveys is randomly assigned to locations each season, subject to logistical constraints (e.g., accessibility after heavy rains).

Statistical Approach and Analysis

The two “treatments” in this natural experiment are restored native grassland and unrestored grassland sites nested within each of four locations. A general indexing (GI) approach as described by Engeman 2005 (see also Engeman and Whisson 2006) will be used to make relative comparisons of wildlife utilization at each site type. While population size could be estimated at each site, differences in site size, vegetation, and other factors could confound comparisons between site type (i.e., restored vs. control). Furthermore, density-estimation procedures using line transect and mark-capture methods are expensive to implement for a wide range of species, and it unlikely that unrealistic statistical assumptions will be met. A 5-year literature review by McKelvey and Pearson (2001) found that 98% of studies resulted in too few data for valid estimates of population abundance. Indices are a desirable alternative and can be used across a wide variety of sampling methods, allowing for relative comparisons to be made between and within populations (Caughley and Sinclair 1994, Krebs 1999, Engeman 2003). Rather than investing considerable time, effort, and money into common but less reliable estimates of absolute density or abundance (Caughley and Sinclair 1994), a GI approach is most efficient and informative to evaluate habitat use. Using this method an index and variance component will be calculated based on counts of species at each trapping point or observation station.

Conclusion

Any management action will have cascading effects on multiple taxa, even if the intended result is focused on one species. Restoration ecology is a rapidly evolving discipline, but still lacks in thoroughly integrating multi-trophic interactions into large-scale experiments (Kardol and Wardle 2010). Grassland restorations affect not just abiotic factors like soil moisture plant community composition and structure, but effects trickle down (or up) to other trophic levels, from raptors to insects. Changes in each trophic level will have subsequent impacts on other levels, and even, perhaps, further-reaching impacts on zoonotic diseases and human health. Monitoring of organismal responses – and changes in multi-trophic interactions – to grassland restoration is necessary to inform management and better appropriate the ever-decreasing funds available for restoration and conservation.

Photo Credit for Featured Image: Ryan Bourbour, U.C. Davis, taken at the Esparto sampling location

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