University of Nevada, Reno
Responses of Above- and Belowground Forb and Plant Species Diversity to Grazing
Exclusion and Fire in the Northern Great Basin Sagebrush Steppe
A thesis submitted in partial fulfillment of the
requirements for the degree of Master of Science in
Animal and Rangeland Science
By
Mariel T. Boldis
Dr. Barry Perryman/Thesis Advisor
May 2020
THE GRADUATE SCHOOL
We recommend that the thesis prepared under our supervision by
Entitled
be accepted in partial fulfillment of the requirements for the degree of
, Advisor
, Committee Member
, Graduate School Representative
David W. Zeh, Ph.D., Dean, Graduate School
i
ABSTRACT
The legacy effects of improper grazing regimes in the pre-1936 Taylor Grazing
Act era and historical fire suppression have contributed to an overall decrease in native deep-rooted perennial bunch grasses and forbs, an increase in annual invasive grasses, and greater sagebrush dominance. Although not as widely used as perennial bunchgrasses, forbs of the Intermountain west have also been tested for use in rehabilitation purposes in the Intermountain West. Forbs provide the majority of plant species richness in stable-state sagebrush systems of the Northern Great Basin, are important seasonal food sources for wildlife like the Greater sage-grouse, provide erosion control through rapid establishment, and help prevent soil-nutrient loss. We assessed differences in above- and belowground diversity between burned and adjacent unburned areas, and between grazed and long-term grazing excluded areas, using soil seed bank and aboveground cover attributes in order to provide insight into ecological potentials of sagebrush sites in the northern Great Basin.
Based on soil texture, elevation, species richness and composition of the seed bank in burned areas at each site, aboveground diversity (Effective S) increased as the seed bank became more diverse and was likely dominated by annual herbaceous species (except Claypan 14-16 #1). Below 1660-m in elevation (Loamy Slope 10-14), non- native annual grasses and forbs generally dominated the seed bank, suggesting that in the event of disturbance, aboveground cover may recover into a non-native annual herbaceous community. At sites between 1660—1970-m in elevation (Loamy 14-16,
Gravelly North Slope 14-18, Claypan 14-16 #2, Claypan 10-14), annual herbaceous plants ii dominated the seed bank, and the composition of native forbs and native perennial grasses increased while the composition of non-native forbs and grasses decreased with increasing elevation (except at Claypan 14-16 #2), suggesting predicted aboveground diversity (Effective S) would generally have a native, mixed annual-perennial herbaceous plant community at higher elevations. Claypan 14-16 #2 did not follow this trend with
91% of total seed density in the burned area characterized as annual, composed mostly by cheatgrass and non-native forbs. The irregularity exhibited by the seed bank composition at Claypan 14-16 #2 can be related to the availability of soil-moisture around the time of the fire. The years following the Holloway fire marked the end of a 9- year negative-phase Pacific Decadal Oscillation (PDO) weather cycle, reoccurring with the El Niño/Southern Oscillation which coincided with normal- to below-normal 30-year annual precipitation averages in 2011 and 2012, and below-normal averages following the fire at all sites (2013). Weather cycles that promote a seed bank that is dominated by annual non-native species and short-lived native perennial grasses, while underrepresenting sagebrush, and long-lived native deep-rooted perennial bunchgrasses and forbs can pose a risk for site potential, influencing structural shifts in plant communities at low and mid-elevations. At elevations above 1970-m (Claypan 14-
16 #1), native perennial grasses and native annual sage-grouse forbs dominated the seed bank, suggesting that aboveground plant communities at this elevation or higher are more likely to recover as a native perennial grass and forb communities post-fire. As elevation increased, more native annual forb species, including sage-grouse forbs (M. gracillis, C. parvifolia, G. decipiens), and fewer non-native annual forbs (C. testiculata, S. iii altissimum) contributed to the seed bank, suggesting there may be selective influences that can provide certain native annual sage-grouse forbs a competitive edge with cheatgrass and other non-native invasive grasses and forbs, which may contribute to the long-term persistence of native forb populations.
In grazing and grazing-excluded sites, there were generally few to no differences in richness (S) and evenness (J’) above- and belowground communities, indicating that diversity (H’) since pre-Taylor Grazing Act conditions is similar to 82-years since exclosures have been in place. Lack of differentiation in belowground diversity (H’) in fire affected areas and grazing/grazing excluded areas, suggests annual and perennial forb and grass communities are more or less similar across the landscape, making it difficult to assess how sites would cluster each year when forb germination depends on availability of soil-moisture. However, one year of data is not enough to suggest grazing pressure and fire effects do or do not drive biologically significant differences in above- and belowground diversity. Regardless of elevation, when only a few annual species are likely to contribute to total cover, sites can exhibit lower diversity (H’), more similar above- and belowground communities, and truncated post-fire and herbivory recovery of long-lived native perennial forbs, grasses, and sagebrush, negatively impacting sage- grouse survival long-term. In such a case, reduction in fine fuels through targeted grazing followed by mechanical reduction of decadent sagebrush can help reduce high- intensity wildfire risk and subsequently maintain more even plant communities. iv
ACKNOWLEDGEMENTS
I would like to thank my advisor Barry Perryman for listening to my curiosities about forbs and creating a great project when I inquired about my interest in graduate school. Barry has always been patient and provided me with guidance and constructive criticism that has improved my technical writing and analytical skills, allowing me to grow as a researcher. I also want to thank Brad Shultz for his help in the field. Without his general appreciation and knowledge for the northern Great Basin, we wouldn’t have found our Holloway fire sites the way we did. I would like to thank my committee for their guidance and support, especially when I needed to reschedule my defense. I also want to thank my committee members Beth Leger and Juan Solomon who provided guidance through my proposal and were there for any questions I had about my work. I would like to thank Kevin Shoemaker for his help in my data analysis. I truly benefited from the one- on-one help when learning how to organize my data and perform simple and complex analyses. I would like to thank Jerry Tiehm for his passion and knowledge in plant taxonomy, and for helping me identify many dead, flattened, growing, and photographic images of desert plants. I’ve learned to be a better plant collector and plant educator through his guidance and tips for identifying plant families and genera. I would like to thank Jon Cerri and his family who opened their home to me when I worked long days in the field last summer. I also want to say thank you to Bo and Fred who helped me set up in the greenhouse. Finally, thank you to my best friends and family who have been there for me through this whole journey. v
TABLE OF CONTENTS
Abstract...... i
Acknowledgements...... iv
Table of Contents...... v
List of Tables...... vi
List of Figures...... x
Chapter I
Literature Review...... 1
Chapter II
Aboveground cover and belowground seed bank diversity of forbs 6 years post-fire in the northern Great Basin...... 30
Chapter III
Above- and belowground species diversity in response to 82 years of grazing exclusion in the Northern Great Basin...... 110 vi
LIST OF TABLES
Chapter II
Table 1: Locations of sites along burn boundaries of the 2012 Holloway fire in Humboldt
Co., NV, 30-year annual precipitation (mm) values, 30-year annual mean temperatures
(°C), elevations, ecological sites, and ecological site description (ESD) associated within major land resource area (MLRA) 23 Malheur High Plateau, plant community characteristics, and soil taxonomic class (USDA-NRCS, 2006).
Table 2: Percent (%) composition by species for aboveground cover in unburned and burned, 6 years after the August 2012 Holloway fire. Species are grouped into life forms.
Forb consumed by sage-grouse are followed by an asterisk *. Only species that composed at least 1% of cover for either unburned or burned sites are listed individually; all remaining species are grouped together for simplicity.
Table 3: Percent (%) composition by species for the seed bank in unburned and burned
6 years after the August 2012 Holloway fire. Species are grouped into life forms (annual forbs, annual grasses, and perennial forbs). Species are grouped into life forms. Forb consumed by sage-grouse are followed by an asterisk *. Only species that composed at least 1% of the seed bank for either unburned or burned sites are listed individually; all remaining species are grouped together for simplicity.
Table 4: Aboveground cover of plant functional groups compared between paired burned (B) and unburned (UB) plots for each site. All variables are reported by site as mean percent cover (%), standard error (SE) (±), and mean difference of % cover (B—
UB). vii
Table 5: Species that germinated in the soil seed bank assay organized into plant functional groups and compared between burned (B) and unburned (UB) areas for each site. All variables are reported as mean density of seeds per treatment (seeds•m-2), standard error (SE) (±), and mean difference of mean density of seeds (B – UB).
Table 6: Sites, elevation (m), treatment (trt), and Bray-Curtis (BC) similarity values (1—
BC dissimilarity) between the seed bank and aboveground vegetation for overall site similarity (BC similarity) and for forbs only (BC forbs similarity).
Table 7: Sites, and percent species composition in aboveground percent cover and belowground seed density, by life habit (annual, perennial) in unburned (UB) and burned (B) areas for overall species and the forb group only.
Table 8: Sites, elevations (m), treatment (trt), Shannon-Weaver diversity index values
(H’), Hutcheson’s t-statistic (t), probability values (P), Effective Species (Effect S), Species richness (S), and Shannon’s Evenness (J’) between paired burned (B) and unburned (UB) treatments in aboveground and soil seed bank diversity and forb diversity.
Table 9: Summary of linear regression analysis for variables: belowground Effective species (below S) in unburned, and below S in burned predicting aboveground overall diversity and aboveground forb diversity. Regression estimates, standard errors (SE), and probability values (P) are reported.
Chapter III
Table 1: The locations of three grazing exclosures, also known as the Nevada Plots, and precipitation zone (PZ), elevations, plant community characteristics, ecological sites, and viii ecological site description (ESD) associated within major land resource area (MLRA) 24
Humboldt Area and MLRA 25 Owyhee High Plateau (USDA-NRCS, 2006).
Table 2: Percent (%) composition by species (annual forbs, annual grasses, perennial forbs, shrubs, and perennial grasses) in aboveground cover in ungrazed and grazed areas, and difference (Grazed – Ungrazed). Species are grouped into life forms. Only species that composed at least 1% of the canopy for either ungrazed or grazed sites are listed individually; all remaining species are grouped together for simplicity.
Table 3: Percent (%) composition by species (annual forbs, annual grasses, perennial forbs, shrubs, and perennial grasses) for the seed bank in ungrazed (UG) and grazed (G) areas, and difference (Grazed – Ungrazed). Species are grouped into life forms. Only species that composed at least 1% of the seed bank for either ungrazed or grazed sites are listed individually; all remaining species are grouped together for simplicity.
Table 4: Aboveground cover of plant functional groups compared between paired ungrazed (UG) and grazed (G) areas for each site. All variables are reported by site as mean percent cover (%) for UG and G, standard error (SE) (±), and mean difference % cover (G – UG).
Table 5: Plant functional groups compared between ungrazed (UG) and grazed (G) areas for each site in the seed bank. All variables are reported as mean number of seeds•m2
(density), standard error (SE) (±), and difference of mean seeds•m2 (G – UG).
Table 6: Sites, elevations (m), treatment (trt), Shannon-Weaver diversity index values
(H’), Effective species (Effective S), richness (S), evenness (J’), Hutcheson’s t-statistic (t), ix and probability values (P) for ungrazed (UG) and grazed (G) plots in aboveground and belowground for overall site diversity. x
LIST OF FIGURES
Chapter II
Figure 1: Six research sites selected along burn boundaries of the 2012 Holloway fire in the Bilk Creek, Trout Creek, and Montana Mountains located in Humboldt County, NV.
Sites identified were within the Holloway fire burn boundary by association with apparent fire scarring (Bureau of Land Management, 2019) and within the Nevada
Department of Wildlife Lone Willow sage-grouse Population Management Unit.
Figure 2: Mean relative percent aboveground cover of all species recorded across all six sites. Proportion of species identified as forb, grass, or shrub are indicated by shade. Poa canbyi (POCA) and Poa cusickii (POCU3) are recognized separately from Poa secunda
(POSE) based on phenology and presence that was unique within ecological sites.
Figure 3: Mean relative density of all species that germinated from the soil seed bank samples across sites. Proportion of species identified as forb, grass, or shrub are indicated by shade. The species, Poa canbyi (POCA), is recognized as unique from Poa secunda (POSE) based on phenology unique to ecological site descriptions.
Figure 4: Plant species shared between above- and belowground plant communities across sites ascending from lowest (1690 m) to highest elevation (1997 m) in unburned and burned areas. Aboveground cover reported as mean relative percent (%) cover.
Belowground reported as mean relative density of seeds that germinated from the soil seed bank.
Figure 5: a) Mean percent (%) aboveground cover of observed species by site, categorized into 9 functional groups and compared between unburned (UB) and burned xi
(B) treatments across ecological sites ascending from lowest (1690 m) to highest elevation (1997 m). b) Mean density of plant species (seeds•m2) that germinated in the soil seed bank at each site, categorized into 8 functional groups and compared between
UB and B. Means and standard error (SE) (±) bars reported.
Figure 6: a) Mean percent (%) cover of cheatgrass between unburned (UB) and burned
(B) areas across sites in order from lowest (1690 m) to highest elevation (1997 m). b)
Mean density (seeds•m2) of cheatgrass seeds that germinated in the soil seed bank at each site between UB and B areas across sites in order from lowest (1690 m) to highest elevation (1997 m). Means and standard error (SE) (±) bars reported.
Figure 7: a.1) A partitioning around medoids (PAM) cluster analysis using Bray-Curtis
(BC) dissimilarity that compared aboveground (a) relative composition of forbs to belowground (b) relative composition of forbs across sites in unburned areas. b.1) the same analysis as in unburned that compared aboveground (a) relative composition of forbs to belowground (b) relative composition of forbs across sites in burned areas. Sites within cluster group (1—3) indicate they are more similar to each other than the sites outside their cluster group.
Figure 8: Plant species identified as indicators in aboveground communities across ecological sites. Pearson’s correlation residuals extracted from chi-squared analysis were significant at ±2. Species with positive residuals indicate a preference for associating with site. Species with negative residuals indicate a preference of disassociation with site. xii
Figure 9: Plant species identified as indicators in belowground communities across ecological sites. Pearson’s correlation residuals extracted from chi-squared analysis were significant at ±2. Species with positive residuals indicate a preference for associating with the seed bank within the site. Species with negative residuals indicate a preference of disassociation with the seed bank within site.
Figure 10: a) Comparison of Shannon-Weaver diversity index (H) measuring aboveground diversity between unburned (UB) and burned (B) areas across ecological sites in order by ascending elevation. b) Comparison of Shannon-Weaver diversity index
(H) measuring belowground diversity between unburned (UB) and burned (B) areas across ecological sites in order by ascending elevation. Error bars indicate 95% confidence intervals. Higher values indicate greater diversity while lower values indicate lower diversity relative to overall diversity by site.
Figure 11: a) Comparison of Shannon-Weaver diversity index (H) for forbs only, measuring aboveground diversity between unburned (UB) and burned (B) areas across ecological sites in order by ascending elevation. b) Comparison of Shannon-Weaver diversity index (H) for forbs only, measuring belowground diversity between unburned
(UB) and burned (B) areas across ecological sites in order by ascending elevation. Error bars indicate 95% confidence intervals. Higher values indicate greater forb diversity while lower values indicate lower forb diversity relative to overall forb diversity by site.
Figure 12: Linear model regression of how well the overall belowground diversity can predict aboveground diversity within unburned (UB) and burned areas (B). R2 values are xiii for the linear trendline for the single regression between number of effective species belowground and number of effective species aboveground.
Figure 13: Linear model regression of how well belowground forb diversity can predict aboveground forb diversity within unburned (UB) and burned areas (B). R2 values are for the linear trendline for the single regression between number of effective forb species belowground and number of effective forb species aboveground.
Chapter III
Figure 1: Mean relative percent (%) aboveground cover of all species recorded across all six sites. Proportion of species identified as forb, grass, or shrub are indicated by shaded bar.
Figure 2: Mean relative density (individuals/m2) of all species recorded across all sites in the seed bank. Proportion of species identified as forb, grass, or shrub are indicated by shaded bar.
Figure 3: Six plant species shared between the seed bank and aboveground cover reported as mean relative percent (%) cover for aboveground and mean relative density
(seeds•m2) for belowground.
Figure 4: a.) Mean percent (%) aboveground cover of observed species categorized into
9 functional groups and compared between ungrazed (UG) and grazed (G) areas across ecological sites ascending from lowest (1339 m) to highest elevation (1799 m). b) Mean density (seeds•m2) of species that germinated in the soil seed bank categorized into 9 xiv functional groups and compared between UG and G areas. Means and standard error
(SE) (±) bars reported.
Figure 5: a.) Mean (%) aboveground cover of cheatgrass between ungrazed (UG) and grazed (G) areas in each site in order from lowest (1339 m) to highest elevation (1799 m). Dinner Station was the only site without observed cheatgrass cover in the canopy for both treatments. b) Mean density (seeds•m2) of cheatgrass that germinated in the seed bank between UG and G areas in each site. Means and standard error (SE) (±) bars reported.
Figure 6: Plant species identified as indicators in aboveground communities across
Paradise Valley 1 (PV1), Paradise Valley 2 (PV2), and Dinner Station (DS). Pearson’s correlation residuals extracted from chi-squared analysis were significant at ±2. Species with positive residuals show there is a preference for associating in the canopy within site. Species with negative residuals indicate a preference of disassociation in the canopy within site.
Figure 7: Plant species identified as indicators in belowground communities across
Paradise Valley 1 (PV1), Paradise Valley 2 (PV2), and Dinner Station (DS). Pearson’s correlation residuals extracted from chi-squared analysis were significant at ±2. Species with positive residuals indicate a preference for associating with the seed bank within the site. Species with negative residuals indicate a preference of disassociation with the seed bank within site.
Figure 8: a) Comparison of Shannon-Weaver diversity index (H’) measuring aboveground diversity between ungrazed (UG) and grazed (G) areas across Paradise Valley 1 (PV1), xv
Paradise Valley 2 (PV2), and Dinner Station (DS) in order by ascending elevation. b)
Comparison of Shannon-Weaver diversity index (H’) measuring belowground diversity between ungrazed (UG) and grazed (G) areas across Paradise Valley 1 (PV1), Paradise
Valley 2 (PV2), and Dinner Station (DS) in order by ascending elevation. Error bars indicate 95% confidence intervals. Higher values indicate greater diversity (H’) while lower values indicate lower diversity by site. 1
CHAPTER I: Literature Review
DESERT SEED BANKS: Seed banks of arid and semi-arid ecosystems are highly variable and dependent on the interactions of multiple physical and environmental characteristics such as soil, elevation, climate, fire, grazing, time, and weather variability (Miller et al.
2011). Most viable seeds will germinate in the upper 2-cm of soil with most seed from annual species unable to germinate below the 1-cm depth (Kemp 1989). Multiple seed bank studies have focused on seed banks in hot deserts such as the Sonoran, Mojave, and Chihuahua Deserts (Reichman 1984; Kemp 1989; Cuello et al. 2019; Filazzola et al.
2019), and in cold deserts like the Great Basin and Colorado Plateau (Hassan and West
1986; Pekas and Schupp 2013; Haight et al. 2019). Both warm and cold desert seed banks exhibit annual production and spatial variability, but at different scales (Kemp
1989) under differing conditions like fire (Hassan and West 1986; Allen et al. 2008) and grazing pressures (Courtois et al. 2004; Davies et al. 2010). Generally, plant species, especially annual herbaceous plants, in seed banks are patchier than aboveground plant distributions (Kemp 1989), making it difficult to predict plant community composition in the short- and long-term (Kemp 1989; Osem et al. 2006; Pekas and Schupp 2013).
Persistent seed banks contain seeds that are viable for more than one calendar year, typically remaining unchanged through time (Filazzola et al. 2019) in warm and cold deserts (Freas and Kemp 1983). It has been previously believed that cold deserts contain seed banks with greater proportions of perennial seeds (Kemp 1989; Guo et al. 1999;
Pekas and Schupp 2013), than do hot desert seed banks (Marone et al. 1998). However, 2 viable seed banks in the Great Basin are typically dominated by forbs (Germain et al.
2018) that are generally annual and native at higher elevations (Martyn et al. 2016) and non-native at lower elevations (Chambers et al. 2017), acting as a source for aboveground plant recovery, dependent on factors such as soil, elevation, climate, timing and intensity of grazing (Freas and Kemp 1983; Kemp 1989; Miller et al. 2011). In addition to the seed bank, Great Basin sagebrush systems are known to contain the highest percentage of annual native forb cover compared to other cold deserts
(Pennington et al. 2017), suggesting dormancy adaptations (Jurado and Flores, 2005) and post-disturbance competition (Chambers et al. 2017) by annual species facilitates their success. The presence of shrubs may influence seed presence and dominance.
Annual and perennial seed densities are usually greater beneath shrub canopies than in- between interspaces (Marone et al. 1998; Allen et al. 2008; Barga and Leger 2018;
Chiquoine and Abella 2018; Filazzola et al. 2019; Funk et al. 2019), suggesting shrubs are an integral part of maintaining the presence of annuals and perennials, especially forbs, year after year.
A greater number of forbs in the seed bank can increase the representation of forb cover aboveground, but establishment may be limited by the effects of fire and grazing pressure. Many sagebrush systems are becoming more susceptible to conversion from sagebrush dominant to annual grassland dominant by factors including
(but not limited to) alterations in fire regimes, historic grazing practices, introduction of non-native plant species, change in land use, and environmental effects like temperature and soil-moisture availability, exacerbated by drought (Crawford et al. 3
2004; Davies et al. 2011; Snyder et al. 2019). Lower elevation plant communities less than 1600-m with southern exposures typically exhibit lower recovery rates (Mata-
Gonzalez et al. 2018) and lower ecological resistance and resilience after intermediate disturbance (Wilkinson 1999; Sherrill and Romme 2012), making lower elevation communities more susceptible to invasive annual grasses (Davies et al. 2011). In turn, the seed bank at lower elevations contains less perennial species inputs (Ripplinger et al.
2015) and annual species continue to increase in the seed bank (Chambers et al. 2017;
Cuello et al. 2019).
FORB SEED BANKS AND ABOVE-GROUND RELATIONSHIPS: Forbs account for a major component of total plant species composition in sagebrush ecosystems of the Northern
Great Basin (Kemp and Smartt 1987; Guo et al. 1999; Pennington et al. 2017). Forbs increase community diversity and resiliency (Ellsworth et al. 2016), are important seasonal food sources for wildlife like the Greater sage-grouse (Barnett and Crawford
1994; Crawford et al. 2004; Dumroese et al. 2016), provide stability to soil through rapid establishment, promoting soil aggregation that helps prevent soil-nutrient loss (Shaw et al. 2005; Walker and Shaw 2005), and forbs help decrease the spread of non-native invasive annual grasses (Wirth and Pyke 2003). Forbs are often underrepresented in the scientific literature and most native forb species have been poor candidates for mass seed production due to specialized requirements for growth, and limited technologies and infrastructure (Pakeman and Small 2005; Adair et al. 2006; Rawlins et al. 2009;
Jones 2019). Since the Northern Great Basin typically contains the highest percentage of 4 native annual forb cover (Pennington et al. 2017) compared to other cold deserts, annual forbs may dominate the seedbank year-round (Chiquoine and Abella 2018), and above-ground vegetation during above-average precipitation years (Pennington et al.
2019) or after fire (Wirth and Pyke 2003; Wrobleski and Kauffman 2003; Chambers et al.
2017; Swanson et al. 2018) due to changes in soil chemistry and soil surface. When annual forbs and grasses dominate both above- and belowground communities (Kinloch and Friedel 2005; Osem et al. 2006; Pekas and Schupp 2013; Heydari et al. 2017), similarity between above- and belowground tends to be greater than in perennial dominated or annual-perennial mixed communities (Thompson and Grime 1979;
Milberg 1995; Osem et al. 2006; Sanaei and Ali 2019). Annual forbs and grasses are able to add seeds belowground year to year (Kemp and Smartt 1987; Osem et al. 2006; Pekas and Schupp 2013), while most perennial forbs are subject to episodic recruitment (Kemp and Smartt 1987; Maron et al. 2019). Although relatively unknown, seed viability of perennial forbs in the Great Basin may exhibit similar survival characteristics indicated in viable seeds of perennial grasses and sagebrush in the Great Basin, typically persisting for less than a year (Humphrey and Schupp 1999; Maier et al. 2001; Hourihan et al.
2018). Greater dissimilarity exhibited between above- and belowground in perennial dominated and annual-perennial mixed plant communities may depend on the organizational level used in comparisons (i.e. functional groups versus species) (Pekas and Schupp 2013), capacity for incident radiation (Davies et al. 2007; Pennington et al.
2019), soil chemistry, especially after fire (Rau et al. 2008; Blank et al. 2017), and variation in soil texture (Davies et al. 2007; Haight et al. 2019) that favors germination of 5 certain species. Soils with higher sand content typically contains less native annual and perennial forb cover (Pennington et al. 2017) and less cheatgrass cover (Haight et al.
2019). In contrast, soils with a higher clay content typically contain more native perennial forb and grass cover and less cheatgrass and non-native forb cover (Davies et al. 2007; Kachergis et al. 2012). Annual and perennial forb cover species richness also increases with greater percent silt in soil texture (Pennington et al. 2017), but the presence of non-native invasive cheatgrass above- and belowground can dampen forb richness all together (Haight et al. 2019). Cover differences influenced by soil texture suggests that if there is high relative presence of specific species functional groups in the seed bank, successful germination may be more or less likely under certain soil textures.
Variability in seed dispersal characteristics and seed survival strategies between annual and perennial forbs help buffer against weather variability (Chambers 2000;
Snyder et al. 2019). Cuello et al. (2019) found lower germination and lower reproductive success for high-survival forb species during dry periods, which will proliferate during wet years. This phenomenon is common for both forb and insect populations that fluctuate during drought (Fischer et al. 1996) and wet cycles (Benson et al. 2003). In addition, shrubs are integral in maintaining belowground seed stores (Hassan and West
1986) where it can be cooler than interspaces, important for the dormancy of many desert-adapted species (Kildisheva et al. 2019). Barga and Leger (2018) found that rare native species richness increased in the seed bank beneath canopies of big sagebrush while densities of native annual forbs were highest beneath rubber rabbitbrush, 6 suggesting shrubs can help facilitate nearly 90% of forb seeds stored belowground
(Marone et al. 1998). Seed bank density and composition beneath shrub canopies may also be higher or lower when compared by functional groups versus species.
Consequently, non-native annual herbaceous species can be overrepresented above- and belowground due to high, annual seed production (Kemp and Smartt 1987; Guo et al. 1999; Pekas and Schupp 2013).
Sagebrush landscapes are composed of ecological sites, which are a land-type classification system developed by the Natural Resource Conservation Service (NRCS) to help distinguish areas with different soil, climate, and vegetation characteristics
(Duniway et al. 2010). Forb responses are sensitive to shifts between and among ecological sites. Species composition correlates strongly to different soil textures between ecological sites, historic use (i.e. grazing pressure, fire frequency) and environmental variation (i.e weather patterns, soil-moisture) impact species composition within ecological sites (Kachergis et al. 2012). Where non-native invasive annual grasses like cheatgrass are present, especially at higher elevations, managing the presence of native annual forbs that have ecological niches overlapping with cheatgrass can help reduce cheatgrass aboveground cover and seed bank inputs (Davy and Rinella
2019). With a better understanding of belowground forb species composition in the
Northern Great Basin, we may better understand the potential for re-establishing native species post-disturbance (Hassan and West 1986; Guo et al. 1999).
7
THE IMPORTANCE OF FORBS FOR SAGE-GROUSE: Forbs are important for sagebrush-obligate species like the Greater sage-grouse (Centrocereus urophasianus), relying on native forbs in the spring and early summer for forage and habitat selection for lekking, nesting, and brood-rearing (Crawford et al. 2004; Aldridge and Boyce 2007; Pennington et al. 2016). The diet of pre-nesting hens includes up to 50% of their body weight in forbs (Barnett and Crawford 1994) and native annual forbs are integral in the early stages of chick survival (Dumroese et al. 2016). Over 34 genera of forbs including
Astragalus ssp., Crepis ssp., and Taraxacum ssp. and 12 genera of forbs classified as most preferred forage, are linked directly to sage-grouse diets, and indirectly by attracting over 41 families of insects (Drut et al. 1994; Dumroese et al. 2016). Although
Arkle et al. (2014) indicated a negative association between the presence of sage-grouse and non-native forbs (Arkle et al. 2014), sage-grouse select for moderate sagebrush and bunchgrass cover that include an abundance of both native and non-native forbs, important for nest success and chick survival (Klebenow 1969; Crawford et al. 2004;
Aldridge and Boyce 2007; Dumroese et al. 2015; Pennington et al. 2016). In the Great
Basin, sage-grouse are found in low sagebrush, black sagebrush, three-tip sagebrush, and large, intact, big sagebrush dominant habitat (Hagen et al. 2011; Miller et al. 2011;
Arkle et al. 2014; Bates and Davies 2019). Optimal heights of perennial grass (≥ 18-cm) and shrub cover (40-80 cm) define management criteria for sage-grouse habitat
(Crawford et al. 2004; Schroff et al. 2018); however, forb and insect abundance are also key indicators for assessing the quality of habitat, especially as threats to entire sagebrush communities also threaten the forage key to sage-grouse survival (Wenninger 8 and Inouye 2008; Gregg and Crawford 2009; Pennington et al. 2016). In recent years, land managers have responded by prioritizing connectivity of the landscape within priority management unit (PMU) areas defined by the Nevada Department of Wildlife
(NDOW). Rehabilitation potential of sagebrush habitat typically increases with elevation, based on lower annual temperatures, greater precipitation, and greater soil-moisture availability (Davies et al. 2007; Chambers 2000; Roundy et al. 2018; Pennington et al.
2019). Low post-disturbance rehabilitation potential can negatively impact presence of sage-grouse in some burned areas for at least 20 years (Nelle et al. 2000; Arkle et al.
2014); however, habitat with high post-disturbance rehabilitation potential can exhibit native perennial grass, forb, and some non-sagebrush shrub cover recovery (Crawford et al. 2004; Bates and Davies 2019; Davies et al. 2016) beneficial to sage-grouse. As long there is adequate plant cover effectively disguising sage-grouse, lowering risk of predation, and availability of forb food resources are high, sage-grouse will likely use the habitat (Fischer et al. 1996; Crawford et al. 2004; Dumroese et al. 2015; Germain et al.
2018). With a continued focus on specific forb responses in above- and belowground communities to changing weather patterns following disturbance (Wirth and Pyke 2003;
Dahlgren et al. 2015; Barga and Leger 2018), we may better understand how lifeform
(annual versus perennial) and species-specific forb responses can become candidates for use in sage-grouse habitat rehabilitation (Dumroese et al. 2016; Jones 2019; Pennington et al. 2019).
9
FIRE AND GRAZING EFFECTS ON FORBS: Historical disturbance has been considered an integral component in maintaining native plant communities in the Great Basin.
Disturbance was oftentimes in the form of fire by Native Americans (Miller et al. 2011), and post-European settlement in the mid-1800s changed the way lands were managed, shifting composition and structure of sagebrush systems through a combination of inappropriate grazing by livestock, fire suppression, and climate change (Chambers et al.
2014). The legacy effects observed today have changed fire regimes, shifting plant communities from native perennial dominated to non-native and native annual dominated (Balch et al. 2013; Bates and Davies 2014; Chambers et al. 2017; Haight et al.
2019).
Controlled grazing can promote species diversity, by increasing richness and evenness between above- and belowground plant communities during highly productive years (Osem et al. 2006; Zhao et al. 2011; Agra and Ne'eman 2012). Based on the principles of the Intermediate Disturbance Hypothesis (Wilkinson 1999), light to moderate grazing during certain stages in plant phenology can increase biomass and aboveground plant diversity (Davy and Rinella 2019). Germain et al. (2018) found aboveground plant diversity was greatest in low densities of sagebrush cover that were moderately grazed. Under certain targeted grazing conditions during the growing season, decreases in native perennial forb and grass cover and density can be greatest
(Davies et al. 2009; Pennington et al. 2017) and recovery can be slow if soil-moisture near the surface is depleted (Maier et al. 2001; Ryel et al. 2010). In areas devoid of invasive annual grasses, rest-rotation grazing can increase native forb cover (Clark et al. 10
2018; Bates et al. 2019) suggesting grazed rather than grazing-excluded habitats would be preferred by sage-grouse (Neel 1980; Crawford et al. 2004; Dahlgren et al. 2015;
Ellsworth et al. 2016). Few observed differences in diversity (richness and evenness) in above- and belowground, grazed and grazing excluded areas (Courtois et al. 2004), indicates that other factors besides grazing pressure, like soil-moisture availability
(Mitchell et al. 2017) and aboveground resource acquisition by shrub and deep-rooted perennials (Rigge et al. 2019) are driving above- and belowground diversity (Barga and
Leger 2018). However, timing, duration, and intensity of grazing in relation to temperature and precipitation of an ecological site, needs consideration when analyzing plant composition and diversity measures (Schmelzer 2014).
The seed bank in arid regions can be important for the recovery of fragmented ecosystems (Caballero et al. 2008; Hassan and West 1986). Understanding how fire characteristics affect the seed bank in arid regions can provide insight into plant community response for fire rehabilitation purposes. Fire and fire exclusion affect sagebrush-dominated communities differently in time and space (Baker 2006). Since most sagebrush species do not re-sprout after fire and residual seed stores are often burned off (Hassan and West 1986 ; Welch and Criddle 2003), recovery after fire can be slow (Miller et al. 2013; Ellsworth et al. 2016; Mahood and Balch 2019).
Wyoming big sagebrush (Artemisia tridentata [Nutt.] ssp. wyomingensis Beetle &
Young) communities are estimated to burn every 50-100+ years, taking 50-100+ years to recover post-fire (Miller et al. 2011), while low sagebrush (Artemisia arbuscula Nutt.) is estimated to burn every 100+ years, taking 100+ years to recover (Miller and Rose 11
1999). Fire has generally been more common in mountain big sagebrush (Artemisia tridentata [Nutt.] ssp. vaseyana [Rydb.] Beetle) communities, estimated to burn every
35+ years and recovers within 20-100 years post-fire (Lesica et al. 2007). Other sagebrush sites used by sage-grouse such as three-tip sagebrush (Artemisia tripartita
Rydb.), black sagebrush (Artemisia nova A. Nelson) and Alkali (early) sagebrush
(Artemisia arbuscula [Nutt.] ssp. longiloba [Osterh.] L.M. Shultz) communities have not been subject to fire-history research as heavily (Baker 2006; Lesica et al. 2007; Miller et al. 2011). Although three-tip sagebrush can re-sprout in some cases, stands do not always recover post-fire (Lesica et al. 2007), as previously identified (Beetle and Johnson
1982; Shariff 1988).
Fire intensity tends to be greatest below shrub cover (Hassan and West 1986) putting viable seed stores at risk of mortality (Allen et al. 2008; Barga and Leger 2018;
Hassan and West 1986; Marone et al. 1998), affecting plant response post-fire. This puts lower elevation, south-facing shrub-dominated communities in danger of annual grassland conversion, burning more frequently and at higher intensities (Balch et al.
2013; Hassan and West 1986; Mata-Gonzalez et al. 2018; Riginos et al. 2019; Swanson et al. 2018). Seeding deep-rooted perennial bunchgrasses can improve recovery post- disturbance (Davies et al. 2011), but seeding native perennial forbs with known growth responses below burned shrub mounds (Parkinson et al. 2013) can benefit sage grouse during the reproductive and brooding season (Wirth and Pyke 2003; Wrobleski and
Kauffman 2003). In addition, fire can trigger germination of surviving annual and perennial forbs in the seed bank (Riginos et al. 2019), increasing above- and 12 belowground diversity (Heydari et al. 2017) and extending their growing season. Post- fire, native forb cover may increase initially (Marone et al. 1998; Wrobleski and
Kauffman 2003; Ripplinger et al. 2015; Pennington et al. 2017), with perennial forb and grass cover persisting most over the long-term (Porensky et al. 2018; Riginos et al.
2019), possibly due to seed carry-over in the soil or surviving residuals (Clark et al.
2018). Time-since-fire can also impact above- and belowground native species richness differently. Habitat affected by fire more than 10 years ago can have greater native herbaceous species richness aboveground than in areas affected by fire less than 10 years ago, while the seed bank had greater native species richness when affected by fire less than 10 years ago (Barga and Leger 2018).
Forb response to time-since-fire and diversity characteristics of above- and belowground communities in burned and unburned areas are not well understood, especially across a region with a mix of sagebrush species (Barga and Leger 2018;
Pennington et al. 2019). Filling these knowledge gaps can provide insight into the ecological potentials of burned and unburned sagebrush sites (Chambers et al. 2014;
Shriver et al. 2019) More than 12 years post-fire in Wyoming and mountain big sagebrush communities (Crawford et al. 2004; Riginos et al. 2019), annual and perennial forb cover contribute the most to diversity, similar to sites a few years post-fire (Samuel and Hart 1994). However, post-fire plant response may vary based on pre-fire habitat composition. Some Wyoming big sagebrush communities with a balanced proportion of functional groups (i.e., shrubs, forbs, perennial bunchgrasses) do not experience increases in forb cover following fire (Fischer et al. 1996), and no difference in forb 13 cover, density, or frequency has been observed between different aged burns (Nelle et al. 2000; Wrobleski and Kauffman 2003). However, fire may have an effect on forb phenology, with some species growing later in the season and flowering at greater rates after fire (Wrobleski and Kauffman 2003).
Long-term grazing exclusion (since 1936) followed by fire, may increase densities of invasive, non-native annual forbs and grasses in sage-grouse habitat (Davies et al.
2009; Pennington et al. 2017). Annual forbs are known for their ability to proliferate and store a seed bank following both grazing exclusion and heavy grazing use (Davies et al.
2009), causing the overall species composition to remain annual dominated. In burned areas managed with moderate, 30-40% grazing utilization, perennial forb cover increased (Davies et al. 2009). In almost all cases of rest from grazing post-fire, there was no difference in density or frequency of perennial forbs when compared to areas that grazed immediately post-fire (Clark et al. 2018); however, grazing practices before and after fire can increase resilience and resistance against non-native annual grasses
(Davies et al. 2010; Schmelzer 2014; Davies et al. 2016; Davies et al. 2018). Soil moisture availability pre- and post-fire can influence the eventual species composition of a site
(Ryel et al. 2010; Cline et al. 2018). For instance, Davy and Rinella (2019) observed no native annual forb cover during a drought year regardless of grazing pressure, while native perennial grasses and shrubs are likely not producing viable seeds or recruiting into adult populations without at least three years of above-normal precipitation (Maier et al. 2001; Ryel et al. 2010; Hourihan et al. 2018). 14
Overall, both fire and grazing practices are both valuable tools that can help maintain and rehabilitate sagebrush ecosystems in the Great Basin. However, understanding how the duration, intensity, and timing of applying fire or grazing pressures to sagebrush habitat is important in order to maximize benefits (i.e. forb production for sage-grouse), increase diversity of native species (Dumroese et al. 2015), and minimize long-term losses of sagebrush (Porensky et al. 2018). 15
REFERENCES
Adair, R., R. C. Johnson, B. Hellier, and W. Kaiser. 2006. collecting tapertip onion (Allium
acuminatum hook.) in the great basin using traditional and GIS methods. Native
Plants Journal. 7:141-148.
Aldridge, C. L., and M. S. Boyce. 2007. Linking occurrence and fitness to persistence:
Habitat-based approach for endangered Greater Sage-Grouse. Ecological
Applications. 17:508-526.
Allen, E. A., J. C. Chambers, and R. S. Nowak. 2008. Effects of a spring prescribed burn on
the soil seed bank in sagebrush steppe exhibiting pinyon-juniper expansion.
Western North American Naturalist. 68:265-277.
Arkle, R. S., D. S. Pilliod, S. E. Hanser, M. L. Brooks, J. C. Chambers, J. B. Grace, K. C.
Knutson, D. A. Pyke, J. L. Welty, and T. A. Wirth. 2014. Quantifying restoration
effectiveness using multi-scale habitat models: Implications for sage-grouse in
the Great Basin. Ecosphere. 5.
Agra, H., and G. Ne'eman. 2012. Composition and diversity of herbaceous patches in
woody vegetation: The effects of grazing, soil seed bank, patch spatial properties
and scale. Flora. 207:310-317.
Baker, W. L. 2006. Fire and restoration of sagebrush ecosystems. Wildlife Society
Bulletin. 34:177-185.
Balch, J. K., B. A. Bradley, C. M. D'Antonio, and J. Gomez-Dans. 2013. Introduced annual
grass increases regional fire activity across the arid western USA (1980-2009).
Global Change Biology. 19:173-183. 16
Barga, S., and E. A. Leger. 2018. Shrub cover and fire history predict seed bank
composition in Great Basin shrublands. Journal of Arid Environments.
Barnett, J. K., and J. A. Crawford. 1994. Prelaying nutrition of sage grouse hens in
Oregon. Journal of Range Management. 47:114-118.
Bates, J. D., and K. W. Davies. 2014. Cattle grazing and vegetation succession on burned
sagebrush steppe. Rangeland Ecology & Management. 67:412-422.
Bates, J. D., and K. W. Davies. 2019. Characteristics of intact Wyoming big sagebrush
associations in southeastern Oregon. Rangeland Ecology & Management. 72:36-
46.
Bates, J. D., K. W. Davies, J. Bournoville, C. Boyd, R. O’Connor, and T. J. Svejcar. 2019.
Herbaceous biomass response to prescribed fire in juniper-encroached
sagebrush steppe. Rangeland Ecology & Management. 72:28-35.
Beetle, A. A., and K. L. Johnson. 1982. Sagebrush in Wyoming. Agricultural Experiment
Station, University of Wyoming.
Benson, L., B. Linsley, J. Smoot, S. Mensing, S. Lund, S. Stine, and A. Sarna-Wojcicki.
2003. Influence of the Pacific Decadal Oscillation on the climate of the Sierra
Nevada, California and Nevada. Quaternary Research. 59:151-159.
Blank, R. R., C. Clements, T. Morgan, and D. Harmon. 2017. Sagebrush wildfire effects on
surface soil nutrient availability: A temporal and spatial study. Soil Science
Society of America Journal. 81:1203-1210. 17
Caballero, I., J. M. Olano, A. Escudero, and J. Loidi. 2008. Seed bank spatial structure in
semi-arid environments: Beyond the patch-bare area dichotomy. Plant Ecology.
195:215-223.
Chambers, J. C. 2000. Seed movements and seedling fates in disturbed sagebrush steppe
ecosystems: Implications for restoration. Ecological Applications. 10:1400-1413.
Chambers, J. C., D. I. Board, B. A. Roundy, and P. J. Weisberg. 2017. Removal of
perennial herbaceous species affects response of Cold Desert shrublands to fire.
Journal of Vegetation Science. 28:975-984.
Chambers, J. C., D. A. Pyke, J. D. Maestas, M. Pellant, C. S. Boyd, S. B. Campbell, S.
Espinosa, D. W. Havlina, K. E. Mayer, and A. Wuenschel. 2014. Using resistance
and resilience concepts to reduce impacts of invasive annual grasses and altered
fire regimes on the sagebrush ecosystem and greater sage-grouse: A strategic
multi-scale approach. Gen. Tech. Rep. RMRS-GTR-326. Fort Collins, CO: US
Department of Agriculture, Forest Service, Rocky Mountain Research Station. 73
p. 326.
Chiquoine, L. P., and S. R. Abella. 2018. Soil seed bank assay methods influence
interpretation of non-native plant management. Applied Vegetation Science.
21:626-635.
Clark, P. E., C. J. Williams, P. R. Kormos, and F. B. Pierson. 2018. Postfire grazing
management effects on mesic sagebrush-steppe vegetation: Mid-summer
grazing. Journal of Arid Environments. 151:104-112. 18
Cline, N. L., B. A. Roundy, S. Hardegree, and W. Christensen. 2018. Using germination
prediction to inform seeding potential: II. Comparison of germination predictions
for cheatgrass and potential revegetation species in the Great Basin, USA.
Journal of Arid Environments. 150:82-91.
Courtois, D. R., B. L. Perryman, and H. S. Hussein. 2004. Vegetation change after 65
years of grazing and grazing exclusion. Journal of Range Management. 57:574-
582.
Crawford, J. A., R. A. Olson, N. E. West, J. C. Mosley, M. A. Schroeder, T. D. Whitson, R. F.
Miller, M. A. Gregg, and C. S. Boyd. 2004. Ecology and management of sage-
grouse and sage-grouse habitat. 57:19.
Cuello, W. S., J. R. Gremer, P. C. Trimmer, A. Sih, and S. J. Schreiber. 2019. Predicting
evolutionarily stable strategies from functional responses of Sonoran Desert
annuals to precipitation. Proceedings of the Royal Society B-Biological Sciences.
286.
Dahlgren, D. K., R. T. Larsen, R. Danvir, G. Wilson, E. T. Thacker, T. A. Black, D. E. Naugle,
J. W. Connelly, and T. A. Messmer. 2015. Greater sage-grouse and range
management: Insights from a 25-year case study in Utah and Wyoming.
Rangeland Ecology & Management. 68:375-382.
Davies, K. W., J. D. Bates, C. S. Boyd, and T. J. Svejcar. 2016. Prefire grazing by cattle
increases postfire resistance to exotic annual grass (Bromus tectorum) invasion
and dominance for decades. Ecology and Evolution. 6:3356-3366. 19
Davies, K. W., J. D. Bates, and R. F. Miller. 2007. Environmental and vegetation
relationships of the Artemisia tridentata spp. wyomingensis alliance. Journal of
Arid Environments. 70:478-494.
Davies, K. W., J. D. Bates, T. J. Svejcar, and C. S. Boyd. 2010. Effects of long-term
livestock grazing on fuel characteristics in rangelands: An example from the
sagebrush steppe. Rangeland Ecology & Management. 63:662-669.
Davies, K. W., C. S. Boyd, and J. D. Bates. 2018. Eighty years of grazing by cattle modifies
sagebrush and bunchgrass structure. Rangeland Ecology & Management. 71:275-
280.
Davies, K. W., C. S. Boyd, J. L. Beck, J. D. Bates, T. J. Svejcar, and M. A. Gregg. 2011.
Saving the sagebrush sea: An ecosystem conservation plan for big sagebrush
plant communities. Biological Conservation. 144:2573-2584.
Davies, K. W., T. J. Svejcar, and J. D. Bates. 2009. Interaction of historical and
nonhistorical disturbances maintains native plant communities. Ecological
Applications. 19:1536-1545.
Davy, J. S., and M. J. Rinella. 2019. Targeted grazing for native forbs in annual
grasslands. Rangeland Ecology & Management. 72:501-504.
Drut, M. S., W. H. Pyle, and J. A. Crawford. 1994. Technical note: Diets and food
selection of sage grouse chicks in Oregon. Journal of Range Management. 47:90-
93. 20
Dumroese, R. K., T. Luna, J. R. Pinto, and T. D. Landis. 2016. Forbs: Foundation for
restoration of monarch butterflies, other pollinators, and greater sage-grouse in
the western United States. Natural Areas Journal. 36:499-511.
Dumroese, R. K., T. Luna, B. A. Richardson, F. F. Kilkenny, and J. B. Runyon. 2015.
Conserving and restoring habitat for greater sage-grouse and other sagebrush-
obligate wildlife: The crucial link of forbs and sagebrush diversity. Native Plants
Journal. 16:276-299.
Duniway, M. C., B. T. Bestelmeyer, and A. Tugel. 2010. Soil processes and properties that
distinguish ecological sites and states. Rangelands. 32:9-15.
Ellsworth, L. M., D. W. Wrobleski, J. B. Kauffman, and S. A. Reis. 2016. Ecosystem
resilience is evident 17 years after fire in Wyoming big sagebrush ecosystems.
Ecosphere. 7.
Filazzola, A., A. R. Liczner, M. Westphal, and C. J. Lortie. 2019. Shrubs indirectly increase
desert seedbanks through facilitation of the plant community. Plos One. 14.
Fischer, R. A., K. P. Reese, and J. W. Connelly. 1996. An investigation on fire effects
within xeric sage grouse brood habitat. Journal of Range Management. 49:194-
198.
Freas, K. E., and P. R. Kemp. 1983. Some relationships between environmental reliability
and seed dormancy in desert annual plants. Journal of Ecology. 71:211-217.
Funk, F. A., A. Loydi, G. Peter, and R. A. Distel. 2019. Effect of grazing and drought on
seed bank in semiarid patchy rangelands of northern Patagonia, Argentina.
International Journal of Plant Sciences. 180:337-344. 21
Germain, S. J., R. K. Mann, T. A. Monaco, and K. E. Veblen. 2018. Short-term
regeneration dynamics of Wyoming big sagebrush at two sites in northern Utah.
Western North American Naturalist. 78:7-16.
Gregg, M. A., and J. A. Crawford. 2009. Survival of greater sage-grouse chicks and broods
in the northern Great Basin. Journal of Wildlife Management. 73:904-913.
Guo, Q. F., P. W. Rundel, and D. W. Goodall. 1999. Structure of desert seed banks:
Comparisons across four North American desert sites. Journal of Arid
Environments. 42:1-14.
Hagen, C. A., M. J. Willis, E. M. Glenn, and R. G. Anthony. 2011. Habitat selection by
greater sage-grouse during winter in southeastern Oregon. Western North
American Naturalist. 71:529-538.
Haight, J. D., S. C. Reed, and A. M. Faist. 2019. Seed bank community and soil texture
relationships in a cold desert. Journal of Arid Environments. 164:46-52.
Hassan, M. A., and N. E. West. 1986. Dynamics of soil seed pools in burned and
unburned sagebrush semi-deserts. Ecology. 67:269-272.
Heydari, M., R. Omidipour, M. Abedi, and C. Baskin. 2017. Effects of fire disturbance on
alpha and beta diversity and on beta diversity components of soil seed banks and
aboveground vegetation. Plant Ecology and Evolution. 150:247-256.
Hourihan, E., B. W. Schultz, and B. L. Perryman. 2018. Climatic influences on
establishment pulses of four Artemisia species in Nevada. Rangeland Ecology &
Management. 71:77-86. 22
Humphrey, L. D., and E. W. Schupp. 1999. Temporal patterns of seedling emergence and
early survival of Great Basin perennial plant species. Great Basin Naturalist.
59:35-49.
Jones, T. A. 2019. Native seeds in the marketplace: Meeting restoration needs in the
Intermountain West, United States. Rangeland Ecology & Management. 72:1017-
1029.
Jurado, E., and J. Flores. 2005. Is seed dormancy under environmental control or bound
to plant traits? Journal of Vegetation Science 16:559-564.
Kachergis, E., M. E. Fernandez-Gimenez, and M. E. Rocca. 2012. Differences in plant
species composition as evidence of alternate states in the sagebrush steppe.
Rangeland Ecology & Management. 65:486-497.
Kemp, P. R. 1989. Seed banks and vegetation processes in deserts. In: M. A. Leck (ed.).
Ecology of Soil Seedbanks. p. 257 - 281.
Kemp, P. R., and R. A. Smartt. 1987. Seed banks and vegetation processes-deserts.
American Journal of Botany. 74:636-636.
Kildisheva, O. A., T. E. Erickson, M. D. Madsen, K. W. Dixon, and D. J. Merritt. 2019. Seed
germination and dormancy traits of forbs and shrubs important for restoration
of North American dryland ecosystems. Plant Biology. 21:458-469.
Kinloch, J. E., and M. H. Friedel. 2005. Soil seed reserves in arid grazing lands of central
Australia. Part 1: Seed bank and vegetation dynamics. Journal of Arid
Environments. 60:133-161. 23
Klebenow, D. A. 1969. Sage grouse nesting and brood habitat in Idaho. Journal of
Wildlife Management. 33:649-662.
Lesica, P., S. V. Cooper, and G. Kudray. 2007. Recovery of big sagebrush following fire in
southwest Montana. Rangeland Ecology & Management. 60:261-269.
Mahood, A. L., and J. K. Balch. 2019. Repeated fires reduce plant diversity in low-
elevation Wyoming big sagebrush ecosystems (1984-2014). Ecosphere. 10.
Maier, A. M., B. L. Perryman, R. A. Olson, and A. L. Hild. 2001. Climatic influences on
recruitment of 3 subspecies of Artemisia tridentata. Journal of Range
Management. 54:699-703.
Marone, L., B. E. Rossi, and M. E. Horno. 1998. Timing and spatial patterning of seed
dispersal and redistribution in a South American warm desert. Plant Ecology.
137:143-150.
Martyn, T. E., J. B. Bradford, D. R. Schlaepfer, I. C. Burke, and W. K. Lauenroth. 2016.
Seed bank and big sagebrush plant community composition in a range margin for
big sagebrush. Ecosphere. 7.
Mata-Gonzalez, R., C. M. Reed-Dustin, and T. J. Rodhouse. 2018. Contrasting effects of
long-term fire on sagebrush steppe shrubs mediated by topography and plant
community. Rangeland Ecology & Management. 71:336-344.
Milberg, P. 1995. Soil seed bank after 18 years of succession from grassland to forest.
Oikos. 72:3-13.
Miller, R. F., J. C. Chambers, D. A. Pyke, F. B. Pierson, and C. J. Williams. 2013. A review
of fire effects on vegetation and soils in the Great Basin Region: Response and 24
ecological site characteristics. Gen. Tech. Rep. RMRS-GTR-308. Fort Collins, CO:
US Department of Agriculture, Forest Service, Rocky Mountain Research Station.
126 p. 308.
Miller, R. F., S. T. Knick, D. A. Pyke, C. W. Meinke, S. E. Hanser, M. J. Wisdom, and A. L.
Hild. 2011. Characteristics of sagebrush habitats and limitations to long-term
conservation. Greater Sage-Grouse: Ecology and Conservation of a Landscape
Species and Its Habitats:145-184.
Miller, R. F., and J. A. Rose. 1999. Fire history and western juniper encroachment in
sagebrush steppe. Journal of Range Management. 52:550-559.
Mitchell, R. M., J. D. Bakker, J. B. Vincent, and G. M. Davies. 2017. Relative importance
of abiotic, biotic, and disturbance drivers of plant community structure in the
sagebrush steppe. Ecological Applications. 27:756-768.
Neel, L. A. 1980. Sage grouse response to grazing management in Nevada. Master’s
Thesis. University of Nevada Reno, Reno, NV.
Nelle, P. J., K. P. Reese, and J. W. Connelly. 2000. Long-term effects of fire on sage
grouse habitat. Journal of Range Management. 53:586-591.
Osem, Y., A. Perevolotsky, and J. Kigel. 2006. Similarity between seed bank and
vegetation in a semi-arid annual plant community: The role of productivity and
grazing. Journal of Vegetation Science. 17:29-36.
Pakeman, R. J., and J. L. Small. 2005. The role of the seed bank, seed rain and the timing
of disturbance in gap regeneration. Journal of Vegetation Science. 16:121-130. 25
Parkinson, H., C. Zabinski, and N. Shaw. 2013. Impact of native grasses and cheatgrass
(Bromus tectorum) on Great Basin forb seedling growth. Rangeland Ecology &
Management. 66:174-180.
Pekas, K. M., and E. W. Schupp. 2013. Influence of aboveground vegetation on seed
bank composition and distribution in a Great Basin Desert sagebrush community.
Journal of Arid Environments. 88:113-120.
Pennington, V. E., J. B. Bradford, K. A. Palmquist, R. R. Renne, and W. K. Lauenroth.
2019. Patterns of big sagebrush plant community composition and stand
structure in the western United States. Rangeland Ecology & Management.
72:505-514.
Pennington, V. E., K. A. Palmquist, J. B. Bradford, and W. K. Lauenroth. 2017. Climate
and soil texture influence patterns of forb species richness and composition in
big sagebrush plant communities across their spatial extent in the western US.
Plant Ecology. 218:957-970.
Pennington, V. E., D. R. Schlaepfer, J. L. Beck, J. B. Bradford, K. A. Palmquist, and W. K.
Lauenroth. 2016. Sagebrush, greater sage-grouse, and the occurrence and
importance of forbs. Western North American Naturalist. 76:298-312.
Porensky, L. M., J. D. Derner, and D. W. Pellatz. 2018. Plant community responses to
historical wildfire in a shrubland-grassland ecotone reveal hybrid disturbance
response. Ecosphere. 9. 26
Rau, B. M., J. C. Chambers, R. R. Blank, and D. W. Johnson. 2008. Prescribed fire, soil,
and plants: Burn effects and interactions in the central great basin. Rangeland
Ecology & Management. 61:169-181.
Rawlins, J. K., V. J. Anderson, R. Johnson, and T. Krebs. 2009. Optimal seeding depth of
five forb species from the Great Basin. Native Plants Journal. 10:32-42.
Reichman, J. O. 1984. Spatial and temporal variation of seed distributions in Sonoran
Desert soils. 1 p.
Rigge, M., C. Homer, B. Wylie, Y. X. Gu, H. Shi, G. Xian, D. K. Meyer, and B. Bunde. 2019.
Using remote sensing to quantify ecosystem site potential community structure
and deviation in the Great Basin, United States. Ecological Indicators. 96:516-
531.
Riginos, C., K. E. Veblen, E. T. Thacker, K. L. Gunnell, and T. A. Monaco. 2019.
Disturbance type and sagebrush community type affect plant community
structure after shrub reduction. Rangeland Ecology & Management. 72:619-631.
Ripplinger, J., J. Franklin, and T. C. Edwards. 2015. Legacy effects of no-analogue
disturbances alter plant community diversity and composition in semi-arid
sagebrush steppe. Journal of Vegetation Science. 26:923-933.
Roundy, B. A., J. C. Chambers, D. A. Pyke, R. F. Miller, R. J. Tausch, E. W. Schupp, B. Rau,
and T. Gruell. 2018. Resilience and resistance in sagebrush ecosystems are
associated with seasonal soil temperature and water availability. Ecosphere. 9. 27
Ryel, R. J., A. J. Leffler, C. Ivans, M. S. Peek, and M. M. Caldwell. 2010. Functional
differences in water-use patterns of contrasting life forms in Great Basin
steppelands. Vadose Zone Journal. 9:548-560.
Samuel, M. J., and R. H. Hart. 1994. 61 years of secondary succession on rangelands of
the Wyoming high-plains. Journal of Range Management. 47:184-191.
Sanaei, A., and A. Ali. 2019. What is the role of perennial plants in semi-steppe
rangelands? Direct and indirect effects of perennial on annual plant species.
Ecological Indicators. 98:389-396.
Schmelzer, L. 2014. CASE STUDY: Reducing cheatgrass (Bromus tectorum L.) fuel loads
using fall cattle grazing. The Professional Animal Scientist. 30:270 - 278.
Schroff, S. R., K. A. Cutting, C. A. Carr, M. R. Frisina, L. B. McNew, and B. F. Sowell. 2018.
Characteristics of shrub morphology on nest site selection of Greater Sage-
Grouse (Centrocercus urophasianus) in high-elevation sagebrush habitat. Wilson
Journal of Ornithology. 130:730-738.
Shariff, A. R. 1988. The vegetation, soil moisture and nitrogen responses to three big
sagebrush (Artemisia tridentata Nutt.) control methods: University of Wyoming.
Shaw, N. L., S. M. Lambert, A. M. DeBolt, and M. Pellant. 2005. Increasing native forb
seed supplies for the Great Basin. In: Dumroese, RK; Riley, LE; Landis, TD, tech.
coords. 2005. National proceedings: Forest and Conservation Nursery
Associations—2004; 2004 July 12–15; Charleston, NC; and 2004 July 26–29;
Medford, OR. Proc. RMRS-P-35. Fort Collins, CO: US Department of Agriculture,
Forest Service, Rocky Mountain Research Station. p. 94-102 35. 28
Sherrill, K. R., and W. H. Romme. 2012. Spatial variation in postfire cheatgrass: Dinosaur
National Monument, USA. Fire Ecology 8:38-56.
Shriver, R. K., C. M. Andrews, R. S. Arkle, D. M. Barnard, M. C. Duniway, M. J. Germino,
D. S. Pilliod, D. A. Pyke, J. L. Welty, and J. B. Bradford. 2019. Transient population
dynamics impede restoration and may promote ecosystem transformation after
disturbance. Ecology Letters. 22:1357-1366.
Snyder, K. A., L. Evers, J. C. Chambers, J. Dunham, J. B. Bradford, and M. E. Loik. 2019.
Effects of changing climate on the hydrological cycle in cold desert ecosystems of
the Great Basin and Columbia Plateau. Rangeland Ecology & Management. 72:1-
12.
Swanson, J. C., P. J. Murphy, S. R. Swanson, B. W. Schultz, and J. K. McAdoo. 2018. Plant
community factors correlated with Wyoming big sagebrush site responses to
fire. Rangeland Ecology & Management. 71:67-76.
Thompson, K., and J. P. Grime. 1979. Seasonal-variation in the seed banks of herbaceous
species in 10 contrasting habitats. Journal of Ecology. 67:893-921.
Walker, S. C., and N. L. Shaw. 2005. Current and potential use of broadleaf herbs for
reestablishing native communities. In: Shaw, Nancy L.; Pellant, Mike; Monsen,
Stephen B., comps. 2005. Sage-grouse habitat restoration symposium
proceedings; 2001 June 4-7, Boise, ID. Proc. RMRS-P-38. Fort Collins, CO: US
Department of Agriculture, Forest Service, Rocky Mountain Research Station. p.
56-61. 29
Welch, B. L., and C. Criddle. 2003. Countering misinformation concerning big sagebrush.
Research Paper RMRS-RP-40. Ogden, UT: US Department of Agriculture, Forest
Service, Rocky Mountain Research Station. 28 p. 40.
Wenninger, E. J., and R. S. Inouye. 2008. Insect community response to plant diversity
and productivity in a sagebrush-steppe ecosystem. Journal of Arid Environments.
72:24-33.
Wilkinson, D. M. 1999. The Disturbing History of Intermediate Disturbance. Oikos.
84:145-147.
Wirth, T. A., and D. A. Pyke. 2003. Restoring forbs for sage grouse habitat: Fire,
microsites, and establishment methods. Restoration Ecology. 11:370-377.
Wrobleski, D. W., and J. B. Kauffman. 2003. Initial effects of prescribed fire on
morphology, abundance, and phenology of forbs in big sagebrush communities
in southeastern Oregon. Restoration Ecology. 11:82-90.
Zhao, L. P., J. S. Su, G. L. Wu, and F. Gillet. 2011. Long-term effects of grazing exclusion
on aboveground and belowground plant species diversity in a steppe of the
Loess Plateau, China. Plant Ecology and Evolution. 144:313-320. 30
CHAPTER II: Aboveground cover and belowground seed bank diversity of forbs 6 years
post-fire in the northern Great Basin
INTRODUCTION
Historically, plant species selected for fire rehabilitation purposes in the Intermountain
West have been native perennial grasses and shrubs rather than forbs (Parkinson et al.
2013). Deep-rooted perennial bunchgrasses are useful in all scenarios but especially important for lower elevations most susceptible to invasive annual grasses (Davies et al.
2011), with southern exposures, and lower recovery rates (Mata-Gonzalez et al. 2018).
Although not as widely used as perennial bunchgrasses, forbs of the Intermountain west have also been tested for use in fire rehabilitation. Forbs increase community diversity and resiliency (Ellsworth et al. 2016), providing the majority of plant species richness in stable-state sagebrush systems of the Northern Great Basin (Kemp and Smartt 1987;
Guo et al. 1999; Pennington et al. 2017). They are important seasonal food sources for wildlife like the Greater sage-grouse (Barnett and Crawford 1994; Crawford et al. 2004;
Dumroese et al. 2016). Forbs are also useful in erosion control through rapid establishment in poor quality soils, promoting soil aggregation that helps prevent soil- nutrient loss (Shaw et al. 2005; Walker and Shaw 2005). Forbs consumed by Greater sage-grouse such as the native perennials tapertip hawksbeard (Crepis acuminata Nutt.) and tapertip onion (Allium acuminatum Hook.) (Luna et al. 2018) have been studied as candidate species for use in sage-grouse habitat rehabilitation (Jones 2019). Presently, most native forb species have been poor candidates for mass seed production due to 31 specialized requirements for growth, and limited technologies and infrastructure
(Pakeman and Small 2005; Adair et al. 2006; Rawlins et al. 2009; Jones 2019). However, improving technologies crucial to conserving biodiversity requires understanding of above- and belowground relationships in forb species richness and composition.
Forb responses before, during, and after fire varies by species, climate, and soil conditions (Davies et al. 2007; Barga and Leger 2018). Post-fire forb responses have been well documented in Great Basin Wyoming big sagebrush (Artemisia tridentata
Nutt. ssp. wyomingensis Beetle & Young) and basin big sagebrush (Artemisia tridentata
Nutt. ssp. tridentata) communities (West and Hassan 1985; Kulpa et al. 2012; Chambers et al. 2014b; Ellsworth et al. 2016; Mata-Gonzalez et al. 2018; Roundy et al. 2018;
Swanson et al. 2018; Pennington et al. 2019; Rigge et al. 2019). Fewer studies discuss post-fire forb response in Great Basin low sagebrush (Artemisia arbuscula Nutt. ssp. arbuscula) and mountain big sagebrush (Artemisia tridentata Nutt. ssp. vaseyana
[Rydb.] Beetle) communities (Davies et al. 2011; Bates et al. 2017; Chambers et al. 2017;
Davies et al. 2017; Pennington et al. 2017; Bates et al. 2019; Pennington et al. 2019;
Riginos et al. 2019). Even fewer studies have analyzed forb responses in three-tip sagebrush (Artemisia tripartita Rydb.) and Alkali (early) low sagebrush (Artemisia arbuscula Nutt. ssp. longiloba [Osterh.] L.M. Shultz) communities (Fischer et al. 1996;
Lesica et al. 2007; Lowe et al. 2009; Kachergis et al. 2012). Forb response to time-since- fire and diversity characteristics of above- and belowground communities in burned and unburned areas are not well understood, especially across a region with a mix of sagebrush species (Barga and Leger 2018; Pennington et al. 2019). Filling these 32 knowledge gaps will further our understanding of above- and belowground relationships in northern Great Basin fire-affected sagebrush communities. Understanding these relationships can provide insight into the ecological potentials of burned and unburned sagebrush sites (Chambers et al. 2014b; Shriver et al. 2019) and provide information about the presence of native forbs in the seed bank prior to rehabilitation plantings or post-fire rehabilitation seedings. We assessed plant recovery of burned areas in comparison with adjacent unburned areas using the soil seed bank and aboveground cover attributes of the 2012 Holloway fire in Humboldt County, NV. This study addressed three questions: 1) How does aboveground cover and belowground seed densities of overall species and the forb species group differ between burned and unburned areas? 2) Relatively, how does aboveground plant (%) cover (total species and forb group) correlate to seed density counts belowground in burned and unburned areas? 3) Can species diversity in the germinable soil seed bank predict species diversity in aboveground plant cover?
METHODS
STUDY AREA: This research was conducted in Humboldt County, NV along burn boundaries of the Holloway fire (August 2012), that burned approximately 291 km2
(Figure 1). The burned area encompassed partitions of the Trout Creek Mountains to the north, the Bilk Creek Mountains to the west, and the Montana Mountains to the east;
Major Land Resource Area (MLRA) 23 Malheur High Plateau (USDA-NRCS, 2006). Six sites were selected in May 2018 based on the following requirements: 1) Sites fell within 33 the Nevada Department of Wildlife (NDOW) Lone Willow sage-grouse Population
Management Unit (PMU); 2) Sites contained paired burned and unburned areas that were clearly identifiable and accessible (burn boundaries delineated by fire lines or by the road) and; 3) Were characterized as important sage-grouse habitat that had not been seeded for fire rehabilitation purposes (Klebenow 1969; Fischer et al. 1996; Lowe et al. 2009; Hagen et al. 2011; Schroff et al. 2018). Five unique ecological plant communities including the dominant sagebrush species were identified at six sites
(Table 1). The nearest Western Region Climate Center (WRCC) weather stations to sites were Disaster Peak, Orovada 4, and Denio Junction; however, they were not representative of the environmental variation that occurs in each site and period records were missing observations; thus, climate information reported in Table 1 originated from the Parameter-elevation Regressions on Independent Slopes Model
(PRISM data) (PRISM 2004).
Unburned areas were identified as late-seral shrubland sites based on mean basal diameter of shrubs (50-145 mm) (Perryman and Olson 2000), presence of decadent shrubs, and low abundance of perennial herbaceous cover (Miller et al. 2013;
Ripplinger et al. 2015). Sagebrush species at sites included Wyoming big sagebrush, mountain big sagebrush, three-tip sagebrush, low sagebrush and Alkali (early) low sagebrush (USDA NRCS, 2019). Loamy Slope 10-14 (41.725 N, 118.404 W), Gravelly
North Slope 14-18 (41.722 N, 118.388 W), and Loamy 14-16 (41.726 N, 118.397 W) located in the Bilk Creek Mountains approximately 30 miles west of an unmarked road
(41.629 N, 118.456 W), were accessed by State Route (SR) NV-41 approximately 41.4 34 miles northwest of SR NV-95. Claypan 10-14 (41.860 N, 118.107 W), Claypan 14-16
(41.780 N, 118.112 W), and Claypan 14-16 #2 (41.911 N, 118.131 W) located in the
Montana Mountains, were approximately 10—30 miles northwest of an unmarked road
(41.691 N, 117.991 W), and accessed by SR NV-293 Kings River Road 16.5 miles west of
SR NV-95. Photos of sites and plots can be found in Appendix A.
EXPERIMENTAL DESIGN: Each of the six sites were treated as a separate, paired assessment for point-in-time comparisons between multiple paired transects across a large non-replicable sampling area. Each site contained two, unfenced paired sample plots (unburned and burned), 1.6 ha (100 x 160-m) in size, with similar aspect, slope, and ecological site characteristics. A systematic sampling method in each plot was used by randomly placing four 50-m long parallel transects spaced 20-m apart, perpendicular to the slope (2 to 15 percent). Each transect served as a replication within each site.
Edge effects from burn lines, roads, and cattle use were excluded by placing transects
30-m away from the immediate area along the unmarked perimeter of each plot
(Braithwaite and Mallik 2012).
SEEDBANK AND VEGETATION SAMPLING: Field sampling took place following seed production of most annual forbs and grasses (2 to 26 July 2018). Percent canopy cover and species composition for herbaceous plants (annual grass, forbs) were collected using Daubenmire frames (Daubenmire 1959) with 0.1-m2 (20-cm x 50-cm) quadrats placed at 10-m intervals along each 50-m transect. Percent basal cover and species 35 composition for shrubs and perennial grass were collected using line-intercept (Canfield
1941) along each 50-m transect. Seed bank assay samples were collected within the center of each Daubenmire quadrat from the top 5.1-cm of soil (196.2 cm3) using a 7-cm diameter Ames hand bulb tulip planter (Barga and Leger 2018) for a total of five samples at each transect, 20 samples per plot (unburned and burned), totaling to 40 samples at each site (Schmelzer 2014). For each site, soil texture class from the top 0 to 30-cm was assessed from a 43.2-cm x 30.5-cm soil pit in burned and unburned areas using the soil texture flow diagram in the field (Thien 1979) and referenced with ecological site descriptions (USDA-NRCS, 2006).
SEEDBANK GREENHOUSE ASSESSMENT: Seed bank assay samples were processed in the greenhouse at the University of Nevada-Reno beginning November 2018 according to
Barga and Leger (2018) and Espeland et al. (2010). Two-hundred and forty 16.5 cm (L) x
10.2 cm (W) x 5.1 cm (H) mini seed garden trays (Barga and Leger 2018) were filled with a 1-cm layer of vermiculite at the base and overlaid with black landscape fabric (18-cm x
12-cm). A 1:2 ratio of sterilized sand (200 mL) was added to each of the 240 field soil samples (400 mL) and transferred to the mini seed garden trays (Barga and Leger 2018).
The prepared mini trays with their soil samples were placed in equal parts at random assignment on four 10-m x 10-m greenhouse tables (60 trays per table). Each greenhouse table was covered with a blue medium duty tarp (1.8-m x 2.4-m) and overlaid with polyester quilt batting (10-m x 10-m) (Espeland et al. 2010; Barga and
Leger 2018). 36
The seed bank assay consisted of four emergence cycles (November 2018 to July
2019) with each cycle lasting until zero new seedlings emerged over the course of seven days (Barga and Leger 2018). During each cycle, plants were counted, identified, and discarded, and the soils were mixed in the trays to trigger germination of remaining seeds. Prior to the start of each cycle, the trays were randomly rotated to a different table, promoting even distribution of greenhouse conditions. Watering was adjusted for each cycle based on plant response to ambient greenhouse temperatures (15 °C and 18
°C) (Ball and Miller 1989; Espeland et al. 2010; Pekas and Schupp 2013; Barga and Leger
2018).
For the first 60-day cycle (12 November 2018 to 11 January 2019), trays were watered twice per day using aerial misters (delivery rate 49-L/hour), five minutes per session, four times per week. The second cycle was 46 days long (2 February to 3 March
2019) and watering sessions increased to three times per day. The third cycle was a 31- day dry period (5 May to 5 June 2019), followed by a final cycle that was 39-days long (6
June to 16 July 2019). At the beginning of cycle four, 3-mL of gibberellic acid (GA) (250 mg/L) was applied to each tray with a 3-mL pipette followed by daily watering sessions, two times per day, 15 minutes per session.
STATISTICAL ANALYSIS: All statistical analyses were conducted using Program R with additional associated statistical packages (Team 2018). Paired t-tests analyzed differences in plant cover and seed density between unburned and burned areas
(Coulloudon et al. 1999) using the PairedData package (Champely and Champely 2018). 37
Potential correlations between relative aboveground plant cover and relative belowground plant density in burned and unburned areas were assessed using the vegan package (Oksanen et al. 2013). The linear relationship between plant species diversity in above- and belowground communities were analyzed using the R Stats package (Team 2018). Figures were produced with the package, ggplot2 (Venables and
Ripley 2002; Wickham 2016).
Means and standard errors were calculated for aboveground cover (canopy and basal cover), and for belowground plant densities (seed bank assay) across each transect
(four unburned and four burned) at each site, by lifeform (forb, grass, or shrub) and by known status (annual, perennial, native or non-native) (USDA NRCS, 2019). A measure of similarity (Bray-Curtis dissimilarity index) (Gardener 2014) was calculated for each site to determine how similar/dissimilar the germinable soil seed bank was to aboveground plant cover. Bray-Curtis index values were calculated with combined relative canopy and basal cover (aboveground) and relative plant density (belowground) from log- transformed abundances log(% + 1), which increased the contribution of less abundant species (Gardener 2014; Greenacre 2017). Relativizing abundance allows comparisons between aboveground plant cover and belowground plant density by standardizing and constraining abundance (0 and 1) at different scales (Gardener 2014). Bray-Curtis index values were analyzed with partition around medoids (PAM) analysis to identify sites with forb communities that are similar to one another (Kaufman and Rousseeuw 1990;
Van der Laan et al. 2003). Clusters were then selected based on average silhouette width (Kaufman and Rousseeuw 1990). 38
Plant species preference for above- and belowground communities across sites were assessed through a chi-squared test using combined (canopy and basal) percent cover, and density (plant counts) from the seed bank. Pearson residuals ()*+) with values
> ±2 were extracted and considered significant under a normal distribution (De Cáceres et al. 2010; Gardener 2014).
Finally, species richness (number of species) and diversity (richness and evenness) assessments were completed for above- and belowground plant communities. Percent canopy and basal cover were combined as abundance (evenness) estimates for the aboveground plant community (Gardener 2014), and plant density from the seed bank assay were abundance (evenness) estimates for the belowground community. A Shannon-Weaver (Shannon and Weaver 1949) index value (H’) was calculated for all plant and all forb species in each site, and a modified t-test (Hutcheson
1970; Heip and Engels 1974; Gardener 2014) analyzed each paired set (unburned and burned) of index values. Effective species was calculated from Shannon-Weaver index values for intuitive comparisons of diversity that would not be apparent when using the raw index (H’) itself (Jost 2006). Effective species from the seed bank assay were used in a linear regression model to predict diversity in aboveground plant communities
(Macarthur 1965).
To meet assumptions for parametric analyses, data were log-transformed and tested for normality using Shapiro-Wilk (p > 0.05). Correlation tests (Spearman’s (ρ) > ±
0.75) were used on mean aboveground cover and seed density groups when unburned 39 results were significantly different from burned (Garcia 2011). Results are reported using untransformed data, and all differences were determined at P = 0.05.
RESULTS
ABOVEGROUND AND SEED BANK SPECIES COMPOSITION: In the aboveground communities,
48 species from 41 genera in 15 families were identified (Figure 2). In the belowground seed bank assay, 20 genera in 13 families were identified and only 16 genera were identified to the species level (Figure 3). Across all sites, the above- and belowground plant communities shared 11 native species (4 annuals, 7 perennials) and 4 non-native species (4 annuals and no perennials) (Figure 4). Of the native species shared between above- and belowground, 6 were forbs, and of the non-native species, 2 were forbs.
Four forbs (< 5% total cover) could not be identified to genus or species in aboveground cover and were not included in the diversity assessments. All encountered shrubs or half-shrubs were native in above- and belowground plant communities. Of the 16 species identified in the seed bank assay, 29% were forbs, and 95% of those forbs were annual. Fifty-six percent of annual forbs in the seed bank assay were native, and 100% of perennial forbs were native. Percent species composition by functional group at each site were recorded for aboveground cover (Table 2) and for the belowground seed bank assay (Table 3).
ABOVEGROUND COVER: All aboveground results are reported in Table 4 as mean percent cover and visualized in Figure 5a and 6a. Six out of seven forb groups were different 40 between paired burned and unburned areas for one or more of the six sites (Table 4a—
4e, 4g), but the differences varied at each site. Cheatgrass, perennial grass, and shrubs were also different between burn treatments for one or more sites (Table 4h—4j). Each site had between 4—9 forbs known to be consumed by Greater sage-grouse (Dumroese et al. 2016; Dumroese et al. 2015) in the aboveground cover (Figure 4g). The burned area at the Loamy Slope 10-14 site had the most non-native and annual species cover and the least native species group cover. In burned areas, sage-grouse forb cover was positively correlated (ρ > ± 0.75) with native and perennial forb cover in the highest elevation site (Claypan 14-16 #1). The Claypan 14-16 #1 site also had the fewest annual and non-native forbs (Table 4c, 4e), and the least amount of cheatgrass cover (Table 4h).
Overall, forb cover was greater in burned areas except at the Claypan 14-16 #2 site, which had 3.1% more forb cover in unburned areas than in burned areas (Table 4a).
The most disparate differences in forb cover occurred at the highest (1997-m) and lowest elevation (1690-m) ecological sites, where Claypan 14-16 #1 and Loamy Slope 10-
14 had 4.6% and 11.2% more forb cover in burned areas than in unburned areas, respectively (Table 4a, Figure 5a). Twenty-five percent of all native forb cover was perennial, and all perennial forbs were native and correlated (ρ > ± 0.75) to native forb cover and sage-grouse forb cover. In burned areas across sites, the Claypan 14-16 #1 site had the greatest amount (9.0—10.1%) and the greatest differences from unburned areas (4.7—6.1%) for perennial forb cover, native forb cover, and sage-grouse forb cover (Table 4b, 4d, 4g). In contrast, the unburned area at the Claypan 14-16 #2 site had the most perennial forb cover (6.5%) and native forb cover (5.7%) of all sites (Table 4b, 41
4d). In burned areas, Claypan 14-16 #2 was the only site with lower perennial forb cover
(3.0%), lower native forb cover (2.5%, and lower sage-grouse forb cover (4.3%) than in unburned areas (Table 4b, 4d, 4g). Although not different, Loamy Slope 10-14 also had lower perennial forb cover, lower native forb cover, and lower sage-grouse forb cover in burned areas. Although not different but biologically important, these same sites
(Claypan 14-16 #2 and Loamy Slope 10-14) also had the most cheatgrass cover in burned areas at 20.3% and 39.9% respectively (Table 4h, Figure 6a). Cheatgrass cover was greater in burned areas than in unburned areas across all sites (2.1—39.9%), where
Claypan 14-16 #1 was the only site with no cheatgrass cover detected in the unburned area and was 2.1% greater in the burned area (Table 4h, Figure 6a).
Annual forb cover was 2.3% greater in the burned area at Claypan 10-14, 16.8% greater in the burned area at Loamy Slope 10-14 than in unburned areas, and 5% greater in the unburned area at Loamy 14-16 (Table 4c). For the Claypan 10-14 site, annual forb cover was all native while at the Loamy Slope 10-14 site, five out of 11 forb species were non-native. Non-native forb cover was only observed at half the sites
(Claypan 10-14, Loamy 14-16, Loamy Slope 10-14) and more abundant in burned than in unburned areas (Table 4e). Across sites, the lowest elevation site (Loamy Slope 10-14) had the greatest mean cover (20.5%) of non-native forbs in the burned area compared to the unburned area (Table 4e)
Perennial grass basal cover was greater in burned than in unburned areas, especially at the Claypan 14-16 #2 (1.0%), Claypan 10-14 (2.8%), and Loamy 14-16 (3.0%) sites (Table 4i, Figure 5a). As expected, basal shrub cover was up to 4% lower in burned 42 than in unburned areas especially at Claypan 14-16 #2 that had 0.1% shrub cover in the burned area and at Loamy 14-16 that had no shrub cover in the burned area (Table 4j).
SEED BANK ASSAY DENSITIES: Five out of six forb groups (Table 5, Figure 5b) in the seed bank assay were highly correlated (ρ > ± 0.75), and different between paired burned and unburned areas at only one site (Loamy Slope 10-14). Cheatgrass density (seeds•m-2) was higher in burned areas at four out of six sites (Claypan 14-16 #1 & #2, Gravelly
North Slope 14-18, Loamy slope 10-14) (Table 5g, Figure 6b) and perennial grass density was higher in the burned area only the Claypan 10-14 site (Table 5h). There were generally no differences in plant groups between burned and unburned areas across sites, except at Loamy Slope 10-14. In burned areas, the density of non-native forbs at the Loamy Slope 10-14 site was 5-55 times greater (Table 5e), and cheatgrass densities
3-10 times greater compared to all other sites (Table 5g, Figure 6b). In addition, the burned area at Loamy Slope 10-14 had higher seed densities of forbs that were all annual (Table 5a, 5c), native forbs (Table 5b), and sage-grouse forbs (Table 5f).
Across sites in burned areas, Claypan 10-14 had the greatest densities of native forbs (Table 5b), sage-grouse forbs (Table 5f), and perennial grass (Table 5h) even with
456 seeds of non-native forb, bur buttercup (Ceratocephala testiculata [Crantz] Roth.) and 1133 seeds of cheatgrass (Bromus tectorum L.) present in the seed bank. The unburned areas at Claypan 10-14 and Loamy Slope 10-14 had similar densities of non- native forbs in unburned areas (Table 5e); however, non-native forb seed densities were much greater in the burned area at Loamy Slope 10-14 than at Claypan 10-14 (Table 5e). 43
Native perennial forbs germinated from the seed assay in low densities at two sites
(Claypan 10-14 and Claypan 14-16 #1) but were not different between treatments
(Table 5d). Shrubs also germinated at low densities in both burned and unburned areas across sites (Table 5i).
ABOVEGROUND COVER AND SEED BANK SIMILARITY: The Bray-Curtis (BC) similarity (1—BC) index showed that the seed bank assay and aboveground cover were relatively low to moderately similar by site within fire treatment (Table 6). Above- and belowground burned communities were generally more similar (17.5—58.0%) than unburned communities (6.9—27.5%) and varied in percent annual and perennial species composition (Table 7). At Both Claypan 14-16 sites had the lowest similarity in unburned areas (Claypan 14-16 #1, BC = 6.9%; Claypan 14-16 #2, BC = 8.1%), and Gravelly North
Slope 14-18 had the lowest similarity in burned areas (BC = 17.5%). The lowest elevation site (Loamy Slope 10-14) was the most similar in both unburned (BC = 27.5%) and burned areas (BC = 58.0%). The unburned and burned areas for both aboveground and the seed bank at the Claypan 14-16 #1 site were dominated by perennial species (51—
96%), but similarity was below 21%. Perennial species dominated unburned areas and annual species dominated burned areas for both aboveground and the seed bank at the
Loamy Slope 10-14 and Claypan 14-16 #2 sites. However, in both unburned and burned areas, the low elevation site (Loamy Slope 10-14) was more similar (28—58%) than
Claypan 14-16 #2 (8—34%). 44
Compared to overall species, forbs were more dissimilar between the seed bank and aboveground vegetation across sites (Table 6). All sites (except Loamy Slope 10-14) were dominated by perennial forbs in the aboveground cover, while the seed bank assay was dominated by annual forbs (Table 7). Across sites, forbs in above- and belowground communities were more similar in burned (0.0—45.5%) than in unburned areas (0.0—18.5%). The Claypan 10-14 site had the most similar forb community in unburned areas (BC = 18.5%), while the Loamy Slope 10-14 site had the most similar forb community in burned areas (BC = 45.5%). In both burned and unburned areas,
Gravelly North Slope 14-18 was the only site that did not share forb species between the seed bank assay and the aboveground plant community. In burned areas only, Claypan
14-16 #2 shared no forb species between the seed bank assay and the aboveground community.
ECOLOGICAL SITE SIMILARITY: The first two components of the partitioning around medoids (PAM) cluster analysis explained 70% of variation in unburned areas and 63% in burned areas (Figure 7a.1). The silhouette plot criterion suggested three groups were similar to one another in above- and belowground forb communities. In unburned areas, aboveground forbs in Group 1 (Claypan 14-16 #1, Claypan 14-16 #2, Claypan 10-
14 sites) were most similar (50—71%), and aboveground forbs in Group 2 (Loamy slope
10-14, Gravelly North Slope 14-18, Loamy 14-16 sites) were most similar (39—54%).
Group 3 contained all six sites, indicating that all sites in unburned areas had similar belowground forb communities (22—67%). 45
In burned areas, Groups 1, 2, and 3 were clustered the same as in unburned areas (Figure 7b.1). Similarity between sites in burned areas within Groups 2 (12—37%) and 3 (56—89%) were more variable than in unburned areas. Aboveground forbs in burned areas of Group 1 (Claypan 10-14, Claypan 14-16 #1, and Claypan 14-16 #2) were most similar (22—37%). Two sites within Group 2 (Gravelly North Slope 14-18 and
Loamy 14-16) had aboveground forb communities more similar to each other (51%) than with the third site in Group 2, Loamy slope 10-14 (12—28%). Group 3 contained all six sites, indicating that all sites in burned areas had similar belowground forb communities (Figure 7b.1). However, the three highest elevation sites within Group 3
(Claypan 10-14, Claypan 14-16 #1, Claypan 14-16 #2) had belowground forb communities most similar to each other (77—89%), while the three lowest elevation sites within the same group (Loamy slope 10-14, Gravelly North Slope 14-18, Loamy 14-
16) were most similar to each other (56—69%). Between the two subgroups within
Group 3, all six sites shared 40—59% of the belowground forb community.
INDICATOR SPECIES ASSOCIATION ANALYSIS: For aboveground affinities, six species indicated a significant site preference (Figure 8). Species preference was based on its abundance at a site and given an importance value ()*+ = 2). Species either had a positive association and were more abundant at a site ()*+ > 2) or had a negative association and were less abundant or avoided the site altogether ()*+ > -2). 46
All aboveground indicator species were forbs (four perennials, and two annuals).
One or more native species were positively associated with all sites (except Loamy Slope
10-14), two or more native species were positively associated with three sites, and the non-native species Sysimbrium altissimum L. was negatively associated with sites that were positively associated with native species. The native forb Crepis acuminata Nutt. was positively associated ()*+ = 18.0) with the only site (Claypan 10-14) that also had the most positive associations with other native species. Loamy Slope 10-14 was also the only site positively associated with S. altissimum ()*+ = 19.1). The native forb Lupinus argenteus Pursh. was the only species positively associated with Loamy 14-16 ()*+ =
15.3) and Gravelly North Slope 14-18 ()*+ = 6.0), while all other species negatively associated with these same sites. The forbs Balsamorhiza hookeri (Hook.) Nutt. ()*+ =
13.7) and Senecio integerrimus Nutt. ()*+ = 11.9) had strong positive associations with
Claypan 14-16 #1. The similar (Claypan 14-16 #2) site had only moderate, but positive associations with S. integerrimus ()*+ = 2.7) and Lomatium L. ssp ()*+ = 5.1).
In the seed bank assay, four species indicated a significant site preference (Figure
9). Two indicator species were forbs, two were grasses, and three out of these four species were native (Figure 9). Claypan 10-14 was the only site to have a positive association in the seed bank with more than one species, Lithophragma (Nutt.) Torr. &
A. Gray ()*+ = 13.5) and Poa secunda J. Presl ()*+ = 4.2). Claypan 14-16 #2 was the only site to be negatively associated with P. secunda ()*+ = -5.4) in the seed bank, but Poa canbyi [(Scribn.) Howell] had a positive preference for the site ()*+ = 19.9). The non- 47
native forb C. testiculata was only positively associated with Loamy Slope 10-14 ()*+ =
13.1), and all native species were negatively associated with the seed bank in this same site.
ABOVEGROUND DIVERSITY: In unburned areas, native perennial forbs (Phlox ssp., Lupinus ssp., S. integerrimus, C. acuminata) dominated 41% of total percent cover, followed by deep-rooted native perennial bunchgrasses (F. idahoensis, A. thurberianum), and sagebrush (Artemisia ssp.), each plant group composed 13% of total percent cover across sites. In burned areas, non-native annual grass and forb (B. tectorum, S. altissimum) dominated 35% of total percent cover, followed by native perennial forbs
(B. sagittata, L. argenteus), composing 18% of total percent cover.
The overall diversity (Figure 10a) and forb diversity (Figure 11a) between unburned and burned areas across sites were variable (Table 8). Aboveground communities in unburned areas were more diverse than burned areas for both overall and for forbs at Loamy Slope 10-14 (P < 0.0001; P = 0.0005), Claypan 14-16 #1 (P =
0.0009; P < 0.0001), and Claypan 14-16 #2 (P = 0.0002; P < 0.0001). Claypan 10-14 was the only site more diverse in burned areas both overall (P = 0.0007) and for forbs (P <
0.0001). Claypan 10-14 had the same species richness in aboveground for overall species, but evenness affected which treatment was more or less diverse, while the forb group was richer and more even in burned than in the unburned area. At Loamy 14-16, overall diversity was not different between treatments, but forbs were more diverse 48
(richer and more even) in burned areas (P = 0.005) than in unburned areas (Table 8,
Figure 11a). Within sites, unburned richness and evenness were affected mostly by
Artemisia ssp., perennial forbs Lupinus L. ssp., Phlox L. ssp., and Festuca idahoensis
Elmer., a perennial grass. Richness and evenness in burned areas was mostly affected by perennial forbs Lupinus ssp., Balsamorhiza hookeri [Hook.] Nutt., annual forb
Sisymbrium altissimum L., and Bromus tectorum L., an annual grass.
BELOWGROUND DIVERSITY: Overall site diversity between unburned and burned areas was different at two sites (Table 8, Figure 10b), and forb diversity was only different at
Loamy Slope 10-14 (Figure 11b). Site diversity was greater in unburned areas at Gravelly
North Slope 14-18 (P = 0.0001) and at Claypan 10-14 (P = 0.02) than in burned areas.
Loamy Slope 10-14 had a more diverse forb community in the unburned area (P = 0.004) than in the burned area.
In unburned areas, the seed bank was dominated by seven species, while six species dominated in burned areas. Native perennial grasses (E. elymoides, P. secunda, and P. canbyi) dominated 45% of total seed density in unburned areas, followed by annual native forb species (M. gracillis, C. parvifolia) and annual non-native forb and grass (C. testiculata, B. tectorum), each plant group composing 22% of total seed density. In burned areas, annual non-native forb and grass (C. testiculata, B. tectorum) dominated 65% of total seed density, followed by native perennial grasses (E. elymoides,
P. secunda) that composed 18% of total seed density, and annual native sage-grouse forbs (M. gracillis, C. parvifolia) that composed 12% of total seed density. In burned 49 areas, more native annual sage-grouse forb species (M. gracillis, C. parvifolia,
Cryptantha ssp., Lithophragma ssp.) were recorded in the seed bank than in unburned areas (except Claypan 14-16 #1).
PREDICTING ABOVEGROUND DIVERSITY FROM THE SEED BANK: In order to better understand if seed bank assays can predict aboveground cover, a measure of Effective species was reported, which represented “true” biological diversity (Jost 2006). Measures of diversity including Shannon’s index, Effective species, richness, and Shannon’s Evenness across ecological sites are reported in Table 9. The relationships between above- and belowground diversity between unburned and burned areas were explained by an adjusted R2 of 81% (Table 9). As the belowground community became more diverse, aboveground diversity increased 2.30 species in burned areas (R2 = 0.86, P = 0.0002) and decreased 0.10 species in unburned areas (P = 0.72) (Table 9, Figure 12).
Forb diversity from the seed bank assay was much less predictable with an adjusted R2 that explained only 2% of the aboveground forb diversity (Table 9). As diversity increased belowground, aboveground forb diversity increased 0.29 species in burned areas (R2 = 0.29, P = 0.21) and decreased 0.08 species in unburned areas (P =
0.62) (Table 9, Figure 13).
DISCUSSION
PREDICTING ABOVEGROUND DIVERSITY FROM THE SEED BANK: Based on soil texture, elevation, species richness and composition of the seed bank in burned areas at each 50 site, aboveground diversity (Effective S) increased as the seed bank became more diverse, and was likely dominated by annual herbaceous species (except Claypan 14-16
#1). Although the seed bank was generally dominated by annual herbaceous species, the composition of annual or perennial species that were native and non-native suggested that different potentials existed for aboveground cover across sites. As elevation increased, our sites contained greater amounts of native perennial grasses, native annual forbs, and lower amounts of non-native annual herbaceous species in the seed bank, suggesting that predicted aboveground diversity (Effective S) would generally contain a native, mixed annual-perennial herbaceous plant community at higher elevations, where soil-moisture availability is also greater (Pennington et al.
2019).
Below 1660-m in elevation (Loamy Slope 10-14), non-native annual grasses and forbs generally dominated the seed bank, suggesting that in the event of disturbance, aboveground cover may recover into a non-native annual herbaceous community.
Previous studies have suggested that the composition of aboveground non-native herbaceous cover before fire plays a strong role in determining post-fire plant communities (Swanson et al. 2018); however, our unburned Loamy Slope 10-14 site was characterized as a high quality, intact late-seral Wyoming big sagebrush community with low non-native cover presence, while the adjacent burned area was dominated by annual non-native cover, indicating belowground seed bank composition is important to consider for pre-fire plant community management where soil-moisture and nutrient availability may be low for extended periods of time (Shinneman and Baker 2009), and 51 temporarily increase after wildfire (Rau et al. 2008; Blank et al. 2017). In addition, it is well documented that lower elevation plant communities typically exhibit lower recovery rates (Mata-Gonzalez et al. 2018) and lower ecological resistance and resilience after intermediate disturbance (Wilkinson 1999; Sherrill and Romme 2012), making lower elevation communities more susceptible to invasive annual grasses
(Davies et al. 2011). In turn, the seed bank at lower elevations contains fewer perennial species inputs (Ripplinger et al. 2015) and annual species continue to increase in the seed bank (Chambers et al. 2017; Cuello et al. 2019).
At sites between 1660—1970-m in elevation (Loamy 14-16, Gravelly North Slope
14-18, Claypan 14-16 #2, Claypan 10-14), annual herbaceous plants dominated the seed bank, and the composition of native forbs and native perennial grasses increased while the composition of non-native forbs and grasses decreased with increasing elevation
(except at Claypan 14-16 #2), suggesting predicted aboveground diversity (Effective S) would generally have a native, mixed annual-perennial herbaceous plant community at higher elevations, where soil-moisture availability also tends to be greater (Pennington et al. 2019). Claypan 14-16 #2 did not follow this trend with 91% of total seed density in the burned area characterized as annual, composed mostly by cheatgrass and non- native forbs, indicating that the burned area of this mid-elevation site may exhibit increases in non-native herbaceous cover in the event of another disturbance during a period of low soil-moisture availability. In explaining the irregularity exhibited by the seed bank composition at Claypan 14-16 #2, available soil-moisture around the time of the fire was not optimal for native deep-rooted perennial recruitment, often requiring 52 at least three consecutive years of additional soil- moisture (Maier et al. 2001; Hourihan et al. 2018) at shallow depths (Peek et al. 2005) earlier in the growing season (Ryel et al.
2010). The years following the Holloway fire marked the end of a 9-year negative-phase
Pacific Decadal Oscillation (PDO) weather cycle, reoccurring with the El Niño/Southern
Oscillation (Mantua and Hare 2002; Benson et al. 2003). This coincided with normal- to below-normal 30-year annual precipitation averages in 2011 and 2012, and below- normal averages following the fire at all sites (2013); from the Parameter-elevation
Regressions on Independent Slopes Model (PRISM data) (PRISM 2004). Weather cycles that promote a seed bank that is dominated by annual non-native species, short-lived native perennial grasses, and underrepresenting sagebrush, and long-lived native deep- rooted perennial bunchgrasses and forbs can pose a risk for site potential, influencing structural shifts in plant communities at low and mid-elevations as the Great Basin becomes warmer and more arid (Snyder et al. 2019).
At elevations above 1970-m (Claypan 14-16 #1), native perennial grasses and native annual sage-grouse forbs dominated the seed bank, suggesting that aboveground plant communities at this elevation or higher are more likely to recover as a native perennial grass and forb communities post-fire. The persistent presence of native forbs and native perennial grasses in the burned seed bank, especially at higher elevations, increases the probability of increasing site resistance and resilience during a positive- phase PDO weather cycle (Chambers et al. 2014b; Shriver et al. 2019).
In unburned areas, a negligible decrease in aboveground diversity (Effective S), suggests belowground diversity (Effective S) has little to no impact on diversity (Effective 53
S) aboveground (P = 0.72) (Figures 9). Based on species richness and composition of the seed bank in unburned areas, aboveground diversity (Effective S) was dominated by either annual or perennial herbaceous and sagebrush species. Elevation and soil texture did not seem to influence unburned species composition, especially since the lowest elevation site (Loamy Slope 10-14) and two of three high-elevation Claypan sites
(Claypan 14-16 #1 & 2) were dominated by perennial species, suggesting long-term PDO weather cycles and soil-moisture availability affect the persistence of unburned cover
(Ryel et al. 2010; Hourihan et al. 2018). Soil-nutrient resources are typically allocated in late-seral unburned shrubland sites (Ripplinger et al. 2015) that can induce competitive suppression (Chambers et al. 2014b) of most species accumulating in the seed bank
(Allen et al. 2008; Pekas and Schupp 2013; Barga and Leger 2018; Filazzola et al. 2019), subsequently lowering predicted aboveground diversity (Effective S). Lower aboveground diversity (Effective S) can put unburned sites (especially lower elevations) at risk of transitioning from one state to an alternative, less resilient, annual grass dominated state (Chambers et al. 2014a; Mitchell et al. 2017), putting any viable perennial species in the seed bank at risk of mortality in a wildfire.
Forb diversity (Effective S) from the seed bank assay was less predictable, but followed the same trend as in overall diversity, increasing in burned areas (R2 = 0.29, P =
0.21) and decreasing in unburned areas (P = 0.62). As elevation increased, more native annual forb species, including sage-grouse forbs (M. gracillis, C. parvifolia, G. decipiens), and fewer non-native annual forbs (C. testiculata, S. altissimum) contributed to the seed bank, suggesting predicted aboveground forb diversity (Effective S) in burned and 54 unburned areas were dominated by native annual forbs as elevation increases. Similar to findings by Leger (2008) for a native perennial grass (squirreltail), the persistent presence of native annual sage-grouse forbs (M. gracillis, C. parvifolia, G. decipiens) in the seed bank across sites, especially in the presence of cheatgrass at various seed densities, suggests there may be selective influences that can provide certain native annual sage-grouse forbs a competitive edge, which may contribute to the long-term persistence of native forb populations. Since non-native annual forbs dominate the burned and unburned forb seed bank of the lowest elevation site (Loamy Slope 10-14), aboveground forb diversity (Effective S) would likely be non-native. In addition to annual invasive grasses, high relative composition of annual non-native forbs in the seed bank can dampen germination and recruitment of native perennial species aboveground, and help annual species dominate aboveground cover (Balch et al. 2013), leading to mortality of an entire cohort of surviving native perennial herbaceous seeds or seedlings.
Entire functional groups (native perennial forbs and deep-rooted perennial grasses) were generally missing from the seed bank assay, likely because most perennial species in the Great Basin are subject to episodic recruitment, and viable seeds typically persist for less than a year (Kemp and Smartt 1987; Maier et al. 2001; Hourihan et al.
2018). In addition, the absence of native perennial forbs in the seed bank may coincide with the expected loss of entire forb species due to increasing aridity (Nowak et al.
2017). It is also possible that we failed to stimulate the germination of perennial forb species, as they are notoriously dormant (Jurado and Flores 2005; Kildisheva et al. 55
2019a; Kildisheva et al. 2019b). Subsequently, diversity (Effective S) predicted for aboveground cover in burned and unburned areas can be greater during years where various native perennial forbs and deep-rooted native perennial grass species are also present in the seed bank. If low availability of perennial forbs and deep-rooted perennial bunchgrasses in the seed bank persists (Luna et al. 2018), resistance and resilience of spring and summer sage-grouse habitat can decrease overtime (Chambers et al. 2014a).
In addition, the abundance of non-native annual species will continue to persist and increase in the seed bank, leading to fewer annual sage-grouse forbs like M. gracillis and
C. parvifolia (Chambers et al. 2014b). Where annual non-native grasses and forbs become the ecologically dominant lifeforms in mixed annual-perennial herbaceous understories, controlled year-round grazing or prescribed fire may be needed in order to reduce fine fuel litter, reduce non-native seed carryover (Perryman et al. 2018), and to help native perennial seeds and seedlings recruit into adult plant populations (Maier et al. 2001; Hourihan et al. 2018).
BELOWGROUND DIVERSITY: Diversity (H’) for overall species in the seed bank assay in burned and unburned areas was different at three of six sites, and species richness never varied by more than four species across sites. For the forb group, diversity (H’) in the seed bank assay in burned and unburned areas was different at only one site, and forb species richness never varied more than four species across sites. Richness (S) and evenness (J’) did not increase or decrease linearly in burned or unburned areas as elevation increased. Although resistance and resilience increases with higher soil- 56 moisture availability closely associated with elevation (Chambers et al. 2014b; Shriver et al. 2019), site potential is affected by disturbance and management treatments that can increase or decrease resource availability, altering above- and belowground species composition (Bates et al. 2017; Chambers et al. 2007; Cline et al. 2018). In our study, species richness was the same between burned and unburned areas at Gravelly North
Slope 14-18 (S = 11) and Claypan 10-14 (S = 10); however, unburned areas were more even at both sites (J’ = 0.8). In addition, Loamy Slope 10-14 also indicated a more even
(J’ = 0.7) forb community in the unburned seed bank, despite lower species richness (S =
5). With more non-native annual grass and forb species contributing to total belowground density in burned (Gravelly North Slope 14-18, Claypan 10-14) than in unburned areas, sites can exhibit lower species diversity (H’) in burned areas (Bansal and Sheley 2016; Davies et al. 2011).
In unburned areas, the seed bank was dominated by seven species and were mostly native annual and perennial species. In burned areas, six species dominated in burned areas where most were annual non-native forbs and grasses. In burned areas, more native annual sage-grouse forb species (M. gracillis, C. parvifolia, Cryptantha ssp.,
Lithophragma ssp.) were recorded in the seed bank than in unburned areas (except
Claypan 14-16 #1), suggesting low-intensity late-fall and winter prescribed fires can be useful in maintaining annual sage-grouse food sources, and for increasing perennial sage-grouse food sources at higher elevations (Claypan 14-16 #1) (Bates et al. 2017).
However, few species generally dominating total seed bank density across sites explains why overall species and forb group diversity (H’) was generally not different between 57 burned and unburned areas. Sagebrush species have been subject to episodic recruitment (Hourihan et al. 2018) following positive PDO-phase weather patterns especially in April May, and June. Successful germination and viable seed production of annual and perennial forbs may also be subject to similar weather patterns; suggesting, years with negative PDO-phase precipitation patterns in April, May, or June may exhibit low forb diversity (H’) in the seed bank. However, this assumption may not be made without long-term monitoring of forb populations following precipitation and long-term
PDO weather cycles.
ABOVEGROUND DIVERSITY: In unburned areas, aboveground cover was dominated by six species (no non-native annual species), while four species dominated in burned areas, suggesting the relative presence of non-native cover in burned areas, which is strongly influenced by soil-moisture (Mitchell et al. 2017), soil chemistry (Blank et al. 2017), and time-since fire (Mahood and Balch 2019), leads to lower species richness, especially if areas were affected by high-intensity fire (Heydari et al. 2017). Aboveground cover for overall species (except Claypan 10-14) and the forb group (except Loamy 14-16, Claypan
10-14) was generally richer and/or more even in unburned than in burned areas, suggesting a negative PDO-phase pre- and post-fire weather cycle may have impacted the lack of production of herbaceous perennial species, and subsequently affecting potential recovery post-fire where a temporary rapid release of soil-nutrients are available for surviving seedlings and seeds (Rau et al. 2008). Unburned areas contained more native perennial shrub and herbaceous cover than in burned areas, indicating 58 native perennial shrub and herbaceous cover can help maintain native aboveground richness and evenness while suppressing non-native annual herbaceous richness (Sanaei and Ali 2019) and dominance (Bansal and Sheley 2016). Compared to burned areas, unburned plant communities were more even (J’) due to fewer species dominating aboveground cover. Subsequently, species diversity overall (H’) and forb group species diversity (H’) were more different between burned and unburned areas in aboveground cover than in belowground diversity (H’).
ABOVEGROUND AND BELOWGROUND SIMILARITY: In general, above- and belowground communities across sites (burned and unburned) were moderate to low in similarity
(1—Bray-Curtis (BC)), also suggested in previous Great Basin research (Pekas and
Schupp 2013; Barga and Leger 2018). Above- and belowground unburned areas were less similar (BC) than burned areas. For the forb group, unburned areas were less similar
(BC) than burned areas in three of six sites, and only one site (Gravelly North Slope 14-
18) shared no forb similarity. Gravelly North Slope 14-18 had low shared above- and belowground forb richness and low cover and seed density of shared forbs. Similar to findings in Martyn et al. (2016) and Pekas and Schupp (2013), the sites exhibiting more dissimilar above- and belowground communities (overall and forb group) in burned and unburned areas, also had belowground species composition driven by an overrepresentation of short-lived native perennial grasses (P. secunda, E. elymoides), native and non-native annual forbs (C. testiculata, M. gracillis, C. parvifolia), and an underrepresentation of sagebrush, perennial forbs, and long-lived native deep-rooted 59 perennial bunchgrasses compared to the established aboveground plant community.
Over- and underrepresentation of species and functional groups in the seed bank compared to aboveground suggests available soil-moisture around the time of the fire was not optimal for native deep-rooted perennial recruitment, often requiring at least three consecutive years of additional soil-moisture (Maier et al. 2001; Hourihan et al.
2018) at shallow depths (Peek et al. 2005) earlier in the growing season (Ryel et al.
2010). In addition to the dry conditions that ensued with a negative-phase PDO weather cycle the year before, the year of, and the year following the Holloway fire, immediate post-fire soil conditions that typically have elevated levels of available nutrients (Blank et al. 2017) may have promoted higher post-fire above- and belowground density of non-native invasive annuals like cheatgrass. Higher than expected above- and belowground density of cheatgrass and additional non-native annuals explain the high similarity value (BC) in the burned area at the mid-elevation site, Claypan 14-16 #2. The combination of consecutive negative-phase PDO cycles pre- and post-fire, the loss of adult native perennial species, and the elevated availability of soil-nutrients post- wildfire, may have promoted above- and belowground communities dominated by annual non-native species and short-lived native perennial grasses while underrepresenting sagebrush and long-lived native deep-rooted perennial bunchgrasses and forbs. Subsequently, native perennial plant recovery was truncated, especially at the low elevation site, Loamy Slope 10-14, and mid-elevation site, Claypan 14-16 #2, shifting site potential espeically as the Great Basin becomes warmer and more arid
(Snyder et al. 2019). 60
Annual plant species with “weedy” characteristics can increase above- and belowground similarity (Hopfensperger 2007; Martyn et al. 2016). Sites with greater than 22% similarity (BC) between above- and belowground communities (burned and unburned) (Table 8), had more than three shared species contributing most to total relative cover or density, two of which were non-native: the annual grass, cheatgrass (B. tectorum), and the annual forb, bur buttercup (C. testiculata) (except Claypan 14-16 #2).
Claypan 14-16 #2 shared high relative density and cover of cheatgrass in above- and belowground burned communities; however, instead of sharing bur buttercup, the native perennial grass, squirreltail (E. elymoides), contributed to burned above- and belowground similarity (BC) (Figure 4). Though cheatgrass typically has low establishment, biomass, and seed production in mid- to high-elevation sites due to colder and shorter growing seasons (Chambers et al. 2007), factors such as accumulated fine fuels of ungrazed annual and perennial grasses (Davies et al. 2016), severity of fire
(Mahood and Balch 2019), lower precipitation the year before fire (Shinneman and
Baker 2009), and available soil moisture in warmer temperatures (Cline et al. 2018) around the time of the fire, can influence the eventual composition of higher elevation burned above- and belowground communities. Cheatgrass and bur buttercup emerged in the seed bank assay from every site, indicating non-native invasive annuals can increase above- and belowground post-fire or move into high-elevation burned sites that had little to no cheatgrass or bur buttercup cover in unburned sites (Claypan 14-16
#1 & #2), subsequently increasing similarity (BC) between the burned seed bank and aboveground cover overtime. More cheatgrass and bur buttercup are likely to 61 germinate in warm, wet fall and spring seasons (Roundy et al. 2018), exploiting shallow soil-moisture more successfully than native perennial seedlings (Melgoza et al. 1990). In turn, increasing above- and belowground similarity due to the presence of non-native invasive annual species, can provide early warning indicators that anticipate ecosystem transformations in unburned areas (Kachergis et al. 2012), or a threshold crossing
(Chambers et al. 2017) that can be brought on by another fire in burned areas (Mahood and Balch 2019). When only a few annual species are likely to contribute to total cover, especially in burned areas, sites can exhibit lower diversity (H’), more similar above- and belowground communities, and truncated post-fire recovery of long-lived native perennial forbs, grasses, and sagebrush (Shriver et al. 2019), negatively impacting sage- grouse survival long-term (Arkle et al. 2014).
Annual herbaceous species are known to possess seed dormancy traits (Jurado and Flores 2005), increasing belowground seed densities. Across all sites, non-native annuals (cheatgrass, bur buttercup), native annual forb, slender phlox (M. gracillis), and native perennial grass, sandberg bluegrass (P. secunda), were more abundant belowground (relative density) than in aboveground (relative cover) in both unburned and burned areas. In addition, the high relative composition of annual forb species shared between above- and belowground communities, contributed to more similar above- and belowground forb communities in burned than in unburned areas.
Furthermore, relative belowground density (burned and unburned areas) was greater for sandberg bluegrass than squirreltail across all sites; however, squirreltail was present in both above- and belowground communities while sandberg bluegrass was 62 mostly observed belowground. Many viable seeds of sandberg bluegrass and squirreltail can be produced through self-pollination in early spring (Kellogg 1987; Beckstead et al.
1996) prior to soil-moisture depletion (Ryel et al. 2010), contributing to the temporary seed bank more often than most perennial grasses. Although some functional traits are not different between sandberg bluegrass and squirreltail (Solomon 2019), greater presence of squirreltail aboveground than sandberg bluegrass could be due to its tolerance to compete and establish more successfully in the presence of invasive annual grasses (Arredondo et al. 1998; Leger 2008; Ferguson et al. 2015). While the native annual sage-grouse forb, slender phlox, was not observed aboveground at most sites
(burned or unburned), slender phlox contributed greater than 10% of total relative density belowground at all sites in burned areas, and in unburned mid- to high elevation sites (Gravelly North Slope 14-18, Claypan 14-16 #1 and #2, Claypan 10-14). The persistence of slender phlox in burned seed banks despite low presence aboveground, suggests the forb could have desiccated aboveground prior to data collection, or has dormancy traits and competitive tolerance with non-native annuals (cheatgrass, bur buttercup) (Rowe and Leger 2011), allowing slender phlox to persist in the seed bank.
Previous studies have suggested soil texture drives aboveground plant and forb similarity (richness and composition) across sites (Davies et al. 2007; Pennington et al.
2017); and based on our results, assessing similarity (BC) between above- and belowground (burned and unburned) communities is most valuable when relating it to both soil texture and elevation. Above- and belowground similarity (BC) for overall species and the forb group did not decrease in burned or unburned areas as elevation 63 increased. Lack of a decreasing trend in above- and belowground similarity with increasing elevation for the forb group at all three Claypan sites, suggests differences in species richness and composition could be influenced by having a more north- or south- facing aspect (Kulpa et al. 2012), or slope and depth to clay accumulation layer could be higher or lower in sites regardless of elevation (Davies et al. 2007). Finer soil textures
(clays) can facilitate an increase in cover of native perennial forbs and grasses
(Pennington et al. 2017; Haight et al. 2019), and at our sites, there was high similarity between the three Claypan sites in both burned and unburned areas (Figure 7).
Meanwhile, coarser soil textures (loams) further increase cover of native and non-native annual herbaceous species (Davies et al. 2007; Haight et al. 2019), and at our sites, there was high similarity between the three Loamy sites in burned and unburned areas.
Although possibly at different depths (Davies et al. 2007), the three high-elevation
Claypan sites contained an argillic (clay) soil horizon that may serve as a moisture reservoir, causing water to pool in the upper horizon before becoming thoroughly wetted (Davies et al. 2007). Pooled soil-water in the upper argillic horizon can facilitate native perennial herbaceous seedling survival during the spring even when annual
(native and non-native) forbs and grasses are present in the seed bank. The argillic and deeper soil layers then become the maintenance pool for mid- and late-season growth of surviving deep-rooted perennial plants once soil-moisture in the shallower horizons is depleted (Ryel et al. 2010; Hourihan et al. 2018).
Subsequently, less annual herbaceous species recruit successfully aboveground in high-elevation sites with high clay content soils. In contrast to aboveground forb 64 cover, similar soil textures did not define unique forb communities belowground across sites. Lack of differentiation in the seed bank between ecological sites suggests annual and perennial forb communities are more or less similar across the landscape (Arkle et al. 2014), and it is difficult to assess how sites would change over time when forb germination depends on the availability of soil moisture (Mitchell et al. 2017).
IMPLICATIONS
There is wide agreement among Greater Sage-Grouse biologists and resource managers that forbs are an important diet and habitat component. Altered fire-regimes, primarily caused by non-native invasive annual grasses like cheatgrass, are shifting potential of sage-grouse habitat outside of its historic range in variation. In turn, various ecological sites within a management unit area are becoming less resilient and resistant to disturbance. Maintaining and increasing native perennial bunchgrasses and forbs in fire- affected landscapes will require extending the fire cycle, especially at lower elevations where site potential is lower than at higher elevations. Sites at higher elevations have cooler temperatures and greater soil-moisture availability for a longer period of time making it easier to increase native forb and grass diversity in burned and unburned areas using controlled grazing, mechanical shrub reduction, seeding (no seeding), or low intensity prescribed fire in the fall season. However, considering the richness and evenness of native perennial grasses and native annual and perennial forbs in relation to non-native annual forbs and grasses in the seed bank at any elevation, is important prior to any shrub or annual grass reduction treatments in unburned areas, and prior to 65 applying post-fire rehabilitation methods in burned areas. Furthermore, the timing of fire in relation to a positive- or negative-phase PDO weather cycle may influence immediate and future plant community recovery potentials, especially as the Great
Basin becomes more arid. A fire during a negative-phase PDO can further increase the cumulative and complete loss of perennial herbaceous grasses and forbs, as indicated by a greater than 50% loss of forb species since the Glacial Maximum period (Nowak et al. 2017). The loss of forbs above- and belowground may leave potential gaps in resource use, facilitating the establishment of invasive annual species, like cheatgrass, in areas that were otherwise devoid of non-native annuals. Without adults to produce viable seeds, recovery potential is truncated and at risk of conversion.
Relating above- and belowground similarity values (BC) to belowground diversity
(H’) and composition at various elevations, can also provide managers with an idea of what plant groups to expect aboveground after disturbance. In addition, preventative wildfire treatments and post-fire rehabilitation methods should be timed with respect to yearly precipitation and long-term weather cycles (positive-phase PDO) or designed to replicate similar conditions for native perennial forb and grass recruitment. This approach should be more effective in meeting sage-grouse habitat diversity and aboveground cover goals, allowing valuable and limited resources to be maximized for successful pre-fire management and post-fire rehabilitation.
LIMITATIONS 66
We document diversity relationships between above- and belowground communities in burned and adjacent unburned areas across various ecological sites. Collecting a minimum of three additional years of above- and belowground species abundance data, could help identify recruitment patterns in native annual and perennial herbaceous species between burned and unburned areas. However, given that positive- and negative-phase PDO weather cycles last longer than three years and positive-phase PDO during the months of April, May, and June coincide with sagebrush recruitment (Maier et al. 2001; Hourihan et al. 2018), it is likely native perennial forbs and grasses follow similar patterns. In addition, soil chemistry immediately post-fire due to rapid decomposition and the subsequent release of nutrients, heavily influenced plant community recovery in burned areas over time (Blank et al. 2017). Measuring how nutrients in the first 10-cm of soil, changes during the growing season each year can help us better understand above- and belowground forb community composition. In addition, some forb species collected from the aboveground cover were unable to be identified to genus or species due to immense desiccation, removing identifiable features of plant. Collecting aboveground cover for forbs earlier in the season at lower elevations can help improve likelihood of proper forb identification and use in diversity assessments. There are also limitations associated with using modeled climate data, instead of data from nearby weather stations; however, the nearest weather stations were not representative of environmental variation in sites and were missing observations for some years. In addition to precipitation patterns, variables including fire intensity, grazing history and intensity may have impacted our results (Freas and 67
Kemp 1983; Grigore and Tramer 1996; Osem et al. 2006; Kildisheva et al. 2019a;
Kildisheva et al. 2019b), especially where the mid-high elevation site (Claypan 14-16 #2) had more cheatgrass cover and densities belowground in burned areas than in the other two high-elevation Claypan sites. The application timing of gibberellic acid (GA), timing of tray cycle rotation, length of time for each cycle, and lack of cold nights (<15°C) may have affected germination of perennial species. Temperature is critical for germination timing of forb species (Boyd and Lemos 2013), suggesting the use of growth chambers for studies in the future may increase the germination rates or the number of species germinating from the seed bank. In addition, a more uniform length of time for each cycle could affect the timing and germination of certain species. 68
REFERENCES
Adair, R., R. C. Johnson, B. Hellier, and W. Kaiser. 2006. Collecting tapertip onion (Allium
acuminatum Hook.) in the Great Basin using traditional and GIS methods. Native
Plants Journal. 7:141-148.
Allen, E. A., J. C. Chambers, and R. S. Nowak. 2008. Effects of a spring prescribed burn on
the soil seed bank in sagebrush steppe exhibiting pinyon-juniper expansion.
Western North American Naturalist. 68:265-277.
Arkle, R. S., D. S. Pilliod, S. E. Hanser, M. L. Brooks, J. C. Chambers, J. B. Grace, K. C.
Knutson, D. A. Pyke, J. L. Welty, and T. A. Wirth. 2014. Quantifying restoration
effectiveness using multi-scale habitat models: Implications for sage-grouse in
the Great Basin. Ecosphere 5.
Arredondo, J. T., T. A. Jones, and D. A. Johnson. 1998. Seedling growth of Intermountain
perennial and weedy annual grasses. Journal of Range Management. 51:584-589.
Balch, J. K., B. A. Bradley, C. M. D'Antonio, and J. Gomez-Dans. 2013. Introduced annual
grass increases regional fire activity across the arid western USA (1980-2009).
Global Change Biology. 19:173-183.
Ball, D. A., and S. D. Miller. 1989. A Comparison of techniques for estimation of arable
soil seedbanks and their relationship to weed flora. Weed Research. 29:365-373.
Bansal, S., and R. L. Sheley. 2016. Annual grass invasion in sagebrush steppe: The
relative importance of climate, soil properties and biotic interactions. Oecologia.
181:543-557. 69
Barga, S., and E. A. Leger. 2018. Shrub cover and fire history predict seed bank
composition in Great Basin shrublands. Journal of Arid Environments. 154:40-50.
Barnett, J. K., and J. A. Crawford. 1994. Prelaying nutrition of sage grouse hens in
Oregon. Journal of Range Management. 47:114-118.
Bates, J. D., K. W. Davies, J. Bournoville, C. Boyd, R. O’Connor, and T. J. Svejcar. 2019.
Herbaceous biomass response to prescribed fire in juniper-encroached
sagebrush steppe. Rangeland Ecology & Management. 72:28-35.
Bates, J. D., K. W. Davies, A. Hulet, R. F. Miller, and B. Roundy. 2017. Sage-grouse
groceries: Forb response to pinon-juniper treatments. Rangeland Ecology &
Management. 70:106-115.
Beckstead, J., S. E. Meyer, and P. S. Allen. 1996. Effects of afterripening on cheatgrass
(Bromus tectorum) and squirreltail (Elymus elymoides) germination. Wild Land
Shrub and Arid Land Restoration Symposium: Proceedings: DIANE Publishing. p.
165.
Benson, L., B. Linsley, J. Smoot, S. Mensing, S. Lund, S. Stine, and A. Sarna-Wojcicki.
2003. Influence of the Pacific Decadal Oscillation on the climate of the Sierra
Nevada, California and Nevada. Quaternary Research 59:151-159.
Blank, R. R., C. Clements, T. Morgan, and D. Harmon. 2017. Sagebrush wildfire effects on
surface soil nutrient availability: A temporal and spatial study. Soil Science
Society of America Journal. 81:1203-1210.
Boyd, C. S., and J. A. Lemos. 2013. Freezing stress influences emergence of germinated
perennial grass seeds. Rangeland Ecology & Management. 66:136-142. 70
Braithwaite, N. T., and A. U. Mallik. 2012. Edge effects of wildfire and riparian buffers
along boreal forest streams. Journal of Applied Ecology. 49:192-201.
Bureau of Land Management. 2019. BLM national fire perimeters polygon. Retrieved
from: https://catalog.data.gov/dataset/blm-national-fire-perimeters-polygon
Canfield, R. H. 1941. Application of the line interception method in sampling range
vegetation. Journal of Forestry. 39:388-394.
Chambers, J. C., D. I. Board, B. A. Roundy, and P. J. Weisberg. 2017. Removal of
perennial herbaceous species affects response of Cold Desert shrublands to fire.
Journal of Vegetation Science. 28:975-984.
Chambers, J. C., R. F. Miller, D. I. Board, D. A. Pyke, B. A. Roundy, J. B. Grace, E. W.
Schupp, and R. J. Tausch. 2014a. Resilience and resistance of sagebrush
ecosystems: Implications for state and transition models and management
treatments. Rangeland Ecology & Management 67:440-454.
Chambers, J. C., D. A. Pyke, J. D. Maestas, M. Pellant, C. S. Boyd, S. B. Campbell, S.
Espinosa, D. W. Havlina, K. E. Mayer, and A. Wuenschel. 2014b. Using resistance
and resilience concepts to reduce impacts of invasive annual grasses and altered
fire regimes on the sagebrush ecosystem and greater sage-grouse: A strategic
multi-scale approach. Gen. Tech. Rep. RMRS-GTR-326. Fort Collins, CO: US
Department of Agriculture, Forest Service, Rocky Mountain Research Station. 73
p. 326. 71
Chambers, J. C., B. A. Roundy, R. R. Blank, S. E. Meyer, and A. Whittaker. 2007. What
makes Great Basin sagebrush ecosystems invasible by Bromus tectorum?
Ecological Monographs. 77:117-145.
Champely, S., and M. S. Champely. 2018. Package ‘PairedData’.
Cline, N. L., B. A. Roundy, S. Hardegree, and W. Christensen. 2018. Using germination
prediction to inform seeding potential: II. Comparison of germination predictions
for cheatgrass and potential revegetation species in the Great Basin, USA.
Journal of Arid Environments. 150:82-91.
Coulloudon, B., K. Eshelman, J. Gianola, N. Habich, L. Hughes, C. Johnson, M. Pellant, P.
Podborny, A. Rasmussen, and B. Robles. 1999. Sampling vegetation attributes.
BLM Technical Reference:1734-1734.
Crawford, J. A., R. A. Olson, N. E. West, J. C. Mosley, M. A. Schroeder, T. D. Whitson, R. F.
Miller, M. A. Gregg, and C. S. Boyd. 2004. Ecology and management of sage-
grouse and sage-grouse habitat. Journal of Range Management. 57:2-19.
Cuello, W. S., J. R. Gremer, P. C. Trimmer, A. Sih, and S. J. Schreiber. 2019. Predicting
evolutionarily stable strategies from functional responses of Sonoran Desert
annuals to precipitation. Proceedings of the Royal Society B-Biological Sciences.
286.
Daubenmire, R.F. 1959. Canopy coverage method of vegetation analysis. Northwest
Science. 33:43-64. 72
Davies, K. W., J. D. Bates, C. S. Boyd, and T. J. Svejcar. 2016. Prefire grazing by cattle
increases postfire resistance to exotic annual grass (Bromus tectorum) invasion
and dominance for decades. Ecology and Evolution. 6:3356-3366.
Davies, K. W., J. D. Bates, and A. Hulet. 2017. Attempting to restore mountain big
sagebrush (Artemisia tridentata ssp vaseyana) four years after fire. Restoration
Ecology. 25:717-722.
Davies, K. W., J. D. Bates, and R. F. Miller. 2007. Environmental and vegetation
relationships of the Artemisia tridentata spp. wyomingensis alliance. Journal of
Arid Environments. 70:478-494.
Davies, K. W., C. S. Boyd, J. L. Beck, J. D. Bates, T. J. Svejcar, and M. A. Gregg. 2011.
Saving the sagebrush sea: An ecosystem conservation plan for big sagebrush
plant communities. Biological Conservation. 144:2573-2584.
De Cáceres, M., P. Legendre, and M. Moretti. 2010. Improving indicator species analysis
by combining groups of sites. Oikos. 119:1674-1684.
Dumroese, R. K., T. Luna, J. R. Pinto, and T. D. Landis. 2016. Forbs: Foundation for
restoration of monarch butterflies, other pollinators, and greater sage-grouse in
the western United States. Natural Areas Journal. 36:499-511.
Dumroese, R. K., T. Luna, B. A. Richardson, F. F. Kilkenny, and J. B. Runyon. 2015.
Conserving and restoring habitat for greater sage-grouse and other sagebrush-
obligate wildlife: The crucial link of forbs and sagebrush diversity. Native Plants
Journal. 16:276-299. 73
Ellsworth, L. M., D. W. Wrobleski, J. B. Kauffman, and S. A. Reis. 2016. Ecosystem
resilience is evident 17 years after fire in Wyoming big sagebrush ecosystems.
Ecosphere. 7:1-12.
Espeland, E. K., L. B. Perkins, and E. A. Leger. 2010. Comparison of seed bank estimation
techniques using six weed species in two soil types. Rangeland Ecology &
Management. 63:243-247.
Ferguson, S. D., E. A. Leger, J. Li, and R. S. Nowak. 2015. Natural selection favors root
investment in native grasses during restoration of invaded fields. Journal of Arid
Environments. 116:11-17.
Filazzola, A., A. R. Liczner, M. Westphal, and C. J. Lortie. 2019. Shrubs indirectly increase
desert seedbanks through facilitation of the plant community. Plos One. 14.
Fischer, R. A., K. P. Reese, and J. W. Connelly. 1996. An investigation on fire effects
within xeric sage grouse brood habitat. Journal of Range Management. 49:194-
198.
Garcia, E. 2011. A tutorial on correlation coefficients. June, 6, 2014.
Gardener, M. 2014. Community ecology: analytical methods using R and Excel. Pelagic
Publishing Ltd.
Greenacre, M. 2017. Ordination with any dissimilarity measure: A weighted Euclidean
solution. Ecology. 98:2293-2300.
Grigore, M. T., and E. J. Tramer. 1996. The short-term effect of fire on Lupinus perennis
(L.). Natural Areas Journal. 16:41-48. 74
Guo, Q. F., P. W. Rundel, and D. W. Goodall. 1999. Structure of desert seed banks:
comparisons across four North American desert sites. Journal of Arid
Environments. 42:1-14.
Hagen, C. A., M. J. Willis, E. M. Glenn, and R. G. Anthony. 2011. Habitat selection by
greater sage-grouse during winter in southeastern Oregon. Western North
American Naturalist. 71:529-538.
Haight, J. D., S. C. Reed, and A. M. Faist. 2019. Seed bank community and soil texture
relationships in a cold desert. Journal of Arid Environments. 164:46-52.
Heip, C., and P. Engels. 1974. Comparing species diversity and evenness indices. Journal
of the Marine Biological Association of the United Kingdom. 54:559-563.
Heydari, M., R. Omidipour, M. Abedi, and C. Baskin. 2017. Effects of fire disturbance on
alpha and beta diversity and on beta diversity components of soil seed banks and
aboveground vegetation. Plant Ecology and Evolution. 150:247-256.
Hopfensperger, K. N. 2007. A review of similarity between seed bank and standing
vegetation across ecosystems. Oikos. 116:1438-1448.
Hourihan, E., B. W. Schultz, and B. L. Perryman. 2018. Climatic influences on
establishment pulses of four Artemisia species in Nevada. Rangeland Ecology &
Management. 71:77-86.
Hutcheson, K. 1970. A test for comparing diversities based on the Shannon formula.
Journal of Theoretical Biology. 29:151-154. 75
Jones, T. A. 2019. Native seeds in the marketplace: Meeting restoration needs in the
Intermountain west, United States. Rangeland Ecology & Management. 72:1017-
1029.
Jost, L. 2006. Entropy and diversity. Oikos. 113:363-375.
Jurado, E., and J. Flores. 2005. Is seed dormancy under environmental control or bound
to plant traits? Journal of Vegetation Science. 16:559-564.
Kachergis, E., M. E. Fernandez-Gimenez, and M. E. Rocca. 2012. Differences in plant
species composition as evidence of alternate states in the sagebrush steppe.
Rangeland Ecology & Management. 65:486-497.
Kaufman, L., and P. J. Rousseeuw. 1990. Partitioning around medoids (program PAM).
Finding groups in data: An introduction to cluster analysis. 344:68-125.
Kellogg, E. A. 1987. Apomixis in the Poa secunda complex. American Journal of Botany.
74:1431-1437.
Kemp, P. R., and R. A. Smartt. 1987. Seed banks and vegetation processes-deserts.
American Journal of Botany. 74:636-636.
Kildisheva, O. A., T. E. Erickson, A. T. Kramer, J. Zeldin, and D. J. Merritt. 2019a.
Optimizing physiological dormancy break of understudied cold desert perennials
to improve seed-based restoration. Journal of Arid Environments. 170.
Kildisheva, O. A., T. E. Erickson, M. D. Madsen, K. W. Dixon, and D. J. Merritt. 2019b.
Seed germination and dormancy traits of forbs and shrubs important for
restoration of North American dryland ecosystems. Plant Biology. 21:458-469. 76
Klebenow, D. A. 1969. Sage grouse nesting and brood habitat in Idaho. Journal of
Wildlife Management. 33:649-662.
Kulpa, S. M., E. A. Leger, E. K. Espeland, and E. M. Goergen. 2012. Postfire seeding and
plant community recovery in the Great Basin. Rangeland Ecology &
Management. 65:171-181.
Leger, E. A. 2008. The adaptive value of remnant native plants in invaded communities:
An example from the Great Basin. Ecological Applications. 18:1226-1235.
Lesica, P., S. V. Cooper, and G. Kudray. 2007. Recovery of big sagebrush following fire in
southwest Montana. Rangeland Ecology & Management. 60:261-269.
Lowe, B. S., D. J. Delehanty, and J. W. Connelly. 2009. Greater sage-grouse Centrocercus
urophasianus use of three-tip sagebrush relative to big sagebrush in south-
central Idaho. Wildlife Biology. 15:229-236.
Luna, T., M. R. Mousseaux, and R. K. Dumroese. 2018. Common native forbs of the
northern Great Basin important for Greater Sage-grouse. Gen. Tech. Rep. RMRS-
GTR-387. Fort Collins, CO: US Department of Agriculture, Forest Service, Rocky
Mountain Research Station; Portland, OR: US Department of the Interior, Bureau
of Land Management, Oregon-Washington Region. 76 p. 387.
Macarthur, R. H. 1965. Patterns of species diversity. Biological Reviews. 40:510-533.
Mahood, A. L., and J. K. Balch. 2019. Repeated fires reduce plant diversity in low-
elevation Wyoming big sagebrush ecosystems (1984-2014). Ecosphere. 10:1-19. 77
Maier, A. M., B. L. Perryman, R. A. Olson, and A. L. Hild. 2001. Climatic influences on
recruitment of 3 subspecies of Artemisia tridentata. Journal of Range
Management. 54:699-703.
Mantua, N. J., and S. R. Hare. 2002. The Pacific decadal oscillation. Journal of
Oceanography. 58:35-44.
Martyn, T. E., J. B. Bradford, D. R. Schlaepfer, I. C. Burke, and W. K. Lauenroth. 2016.
Seed bank and big sagebrush plant community composition in a range margin for
big sagebrush. Ecosphere. 7.
Mata-Gonzalez, R., C. M. Reed-Dustin, and T. J. Rodhouse. 2018. Contrasting effects of
long-term fire on sagebrush steppe shrubs mediated by topography and plant
community. Rangeland Ecology & Management. 71:336-344.
Melgoza, G., R. S. Nowak, and R. J. Tausch. 1990. Soil-water exploitation after fire -
competition between Bromus tectorum (cheatgrass) and 2 native species.
Oecologia. 83:7-13.
Miller, R. F., J. C. Chambers, D. A. Pyke, F. B. Pierson, and C. J. Williams. 2013. A review
of fire effects on vegetation and soils in the Great Basin Region: Response and
ecological site characteristics. Gen. Tech. Rep. RMRS-GTR-308. Fort Collins, CO:
US Department of Agriculture, Forest Service, Rocky Mountain Research Station.
126 p. 308.
Mitchell, R. M., J. D. Bakker, J. B. Vincent, and G. M. Davies. 2017. Relative importance
of abiotic, biotic, and disturbance drivers of plant community structure in the
sagebrush steppe. Ecological Applications. 27:756-768. 78
Nowak, R. S., C. L. Nowak, and R. J. Tausch. 2017. Vegetation dynamics during last
35,000 years at a cold desert locale: Preferential loss of forbs with increased
aridity. Ecosphere. 8.
Oksanen, J., F. G. Blanchet, R. Kindt, P. Legendre, P. R. Minchin, R. O’hara, G. L. Simpson,
P. Solymos, M. H. H. Stevens, and H. Wagner. 2013. Package ‘vegan’. Community
ecology package, version 2:1-295.
Osem, Y., A. Perevolotsky, and J. Kigel. 2006. Similarity between seed bank and
vegetation in a semi-arid annual plant community: The role of productivity and
grazing. Journal of Vegetation Science. 17:29-36.
Pakeman, R. J., and J. L. Small. 2005. The role of the seed bank, seed rain and the timing
of disturbance in gap regeneration. Journal of Vegetation Science. 16:121-130.
Parkinson, H., C. Zabinski, and N. Shaw. 2013. Impact of native grasses and cheatgrass
(Bromus tectorum) on Great Basin forb seedling growth. Rangeland Ecology &
Management. 66:174-180.
Peek, M. S., A. J. Leffler, C. Y. Ivans, R. J. Ryel, and M. M. Caldwell. 2005. Fine root
distribution and persistence under field conditions of three co-occurring Great
Basin species of different life form. New Phytologist. 165:171-180.
Pekas, K. M., and E. W. Schupp. 2013. Influence of aboveground vegetation on seed
bank composition and distribution in a Great Basin Desert sagebrush community.
Journal of Arid Environments. 88:113-120.
Pennington, V. E., J. B. Bradford, K. A. Palmquist, R. R. Renne, and W. K. Lauenroth.
2019. Patterns of big sagebrush plant community composition and stand 79
structure in the western United States. Rangeland Ecology & Management.
72:505-514.
Pennington, V. E., K. A. Palmquist, J. B. Bradford, and W. K. Lauenroth. 2017. Climate
and soil texture influence patterns of forb species richness and composition in
big sagebrush plant communities across their spatial extent in the western US.
Plant Ecology. 218:957-970.
Perryman, B. L., and R. A. Olson. 2000. Age-stem diameter relationships of big sagebrush
and their management implications. Journal of Range Management. 53:342-346.
Perryman, B. L., B. W. Schultz, J. K. McAdoo, R. L. Alverts, J. C. Cervantes, S. Foster, G.
McCuin, and S. Swanson. 2018. Viewpoint: An alternative management paradigm
for plant communities affected by invasive annual grass in the Intermountain
West. Rangelands. 40:77-82.
PRISM Climate Group, Oregon State University, 2004. Available at.
http://prism.oregonstate.edu, Accessed date: 11 December 2019.
Rau, B. M., J. C. Chambers, R. R. Blank, and D. W. Johnson. 2008. Prescribed fire, soil,
and plants: Burn effects and interactions in the central Great Basin. Rangeland
Ecology & Management. 61:169-181.
Rawlins, J. K., V. J. Anderson, R. Johnson, and T. Krebs. 2009. Optimal seeding depth of
five forb species from the Great Basin. Native Plants Journal. 10:32-42.
Rigge, M., C. Homer, B. Wylie, Y. X. Gu, H. Shi, G. Xian, D. K. Meyer, and B. Bunde. 2019.
Using remote sensing to quantify ecosystem site potential community structure 80
and deviation in the Great Basin, United States. Ecological Indicators. 96:516-
531.
Riginos, C., K. E. Veblen, E. T. Thacker, K. L. Gunnell, and T. A. Monaco. 2019.
Disturbance type and sagebrush community type affect plant community
structure after shrub reduction. Rangeland Ecology & Management. 72:619-631.
Ripplinger, J., J. Franklin, and T. C. Edwards. 2015. Legacy effects of no-analogue
disturbances alter plant community diversity and composition in semi-arid
sagebrush steppe. Journal of Vegetation Science. 26:923-933.
Roundy, B. A., J. C. Chambers, D. A. Pyke, R. F. Miller, R. J. Tausch, E. W. Schupp, B. Rau,
and T. Gruell. 2018. Resilience and resistance in sagebrush ecosystems are
associated with seasonal soil temperature and water availability. Ecosphere 9:1-
27.
Rowe, C. L. J., and E. A. Leger. 2011. Competitive seedlings and inherited traits: A test of
rapid evolution of Elymus multisetus (big squirreltail) in response to cheatgrass
invasion. Evolutionary Applications. 4:485-498.
Ryel, R. J., A. J. Leffler, C. Ivans, M. S. Peek, and M. M. Caldwell. 2010. Functional
differences in water-use patterns of contrasting life forms in Great Basin steppe
lands. Vadose Zone Journal. 9:548-560.
Sanaei, A., and A. Ali. 2019. What is the role of perennial plants in semi-steppe
rangelands? Direct and indirect effects of perennial on annual plant species.
Ecological Indicators. 98:389-396. 81
Schmelzer, L. 2014. Case study: Reducing cheatgrass (Bromus tectorum L.) fuel loads
using fall cattle grazing. The Professional Animal Scientist. 30:270 - 278.
Schroff, S. R., K. A. Cutting, C. A. Carr, M. R. Frisina, L. B. McNew, and B. F. Sowell. 2018.
Characteristics of shrub morphology on nest site selection of Greater Sage-
Grouse (Centrocercus urophasianus) in high-elevation sagebrush habitat. Wilson
Journal of Ornithology. 130:730-738.
Shannon, C. E., and W. Weaver. 1949. The mathematical theory of communication.
Univ. Illinois Press. Urbana. 117 p.
Shaw, N. L., S. M. Lambert, A. M. DeBolt, and M. Pellant. 2005. Increasing native forb
seed supplies for the Great Basin. In: Dumroese, RK; Riley, LE; Landis, TD, tech.
coords. 2005. National proceedings: Forest and Conservation Nursery
Associations—2004; 2004 July 12–15; Charleston, NC; and 2004 July 26–29;
Medford, OR. Proc. RMRS-P-35. Fort Collins, CO: US Department of Agriculture,
Forest Service, Rocky Mountain Research Station. p. 94-102 35.
Sherrill, K. R., and W. H. Romme. 2012. Spatial variation in postfire cheatgrass: Dinosaur
National Monument, USA. Fire Ecology. 8:38-56.
Shinneman, D. J., and W. L. Baker. 2009. Environmental and climatic variables as
potential drivers of post-fire cover of cheatgrass (Bromus tectorum) in seeded
and unseeded semiarid ecosystems. International Journal of Wildland Fire.
18:191-202.
Shriver, R. K., C. M. Andrews, R. S. Arkle, D. M. Barnard, M. C. Duniway, M. J. Germino,
D. S. Pilliod, D. A. Pyke, J. L. Welty, and J. B. Bradford. 2019. Transient population 82
dynamics impede restoration and may promote ecosystem transformation after
disturbance. Ecology Letters. 22:1357-1366.
Snyder, K. A., L. Evers, J. C. Chambers, J. Dunham, J. B. Bradford, and M. E. Loik. 2019.
Effects of changing climate on the hydrological cycle in cold desert ecosystems of
the Great Basin and Columbia Plateau. Rangeland Ecology & Management. 72:1-
12.
Solomon, J. K. Q. 2019. Characterization of adult functional traits of local populations
and cultivars of sandberg bluegrass and bottlebrush squirreltail perennial
bunchgrasses. Plants. 8:1-21.
Swanson, J. C., P. J. Murphy, S. R. Swanson, B. W. Schultz, and J. K. McAdoo. 2018. Plant
community factors correlated with Wyoming big sagebrush site responses to
fire. Rangeland Ecology & Management. 71:67-76.
Team, R. C. 2018. R: A language and environment for statistical computing. Vienna,
Austria: R Foundation for Statistical Computing.
Thien, S. J. 1979. A flow diagram for teaching texture-by-feel analysis. Journal of
Agronomic Education. 8.
United States Department of Agriculture (USDA), Natural Resources Conservation
Service (NRCS). 2019. The PLANTS Database (http://plants.usda.gov, 1 October
2019). National Plant Data Team, Greensboro, NC 27401-4901 USA
United States Department of Agriculture (USDA), Natural Resources Conservation
Service (NRCS). 2006. Land resource regions and major land resource areas of 83
the United States, the Caribbean, and the Pacific Basin. U.S. Department of
Agriculture Handbook. 296.
Van der Laan, M. J., K. S. Pollard, and J. Bryan. 2003. A new partitioning around medoids
algorithm. Journal of Statistical Computation and Simulation 73:575-584.
Venables, W. N., and B. D. Ripley. 2002. Random and mixed effects. In Modern applied
statistics with S (pp. 271-300). Springer, New York, NY.
Walker, S. C., and N. L. Shaw. 2005. Current and potential use of broadleaf herbs for
reestablishing native communities. In: Shaw, Nancy L.; Pellant, Mike; Monsen,
Stephen B., comps. 2005. Sage-grouse habitat restoration symposium
proceedings; 2001 June 4-7, Boise, ID. Proc. RMRS-P-38. Fort Collins, CO: US
Department of Agriculture, Forest Service, Rocky Mountain Research Station. p.
56-61.
West, N. E., and M. A. Hassan. 1985. Recovery of sagebrush-grass vegetation following
wildfire. Journal of Range Management. 38:131-134.
Wickham, H. 2016. ggplot2: elegant graphics for data analysis. Springer.
Wilkinson, D. M. 1999. The disturbing history of intermediate disturbance. Oikos.
84:145-147. 84
Table 1 Locations of sites along burn boundaries of the 2012 Holloway fire in Humboldt Co., NV, 30-year annual precipitation (mm) values, 30-year annual mean temperatures (°C), elevations, ecological sites, and ecological site description (ESD) associated within major land resource area (MLRA) 23 Malheur High Plateau, plant community characteristics, and soil taxonomic class (USDA-NRCS, 2006). Ecological Site Lat (N) Long (W) *30-yr *30-yr Elevation MLRA/ESD Plant Community Soil Taxonomic Class Annual Annual (m) Precip Mean (mm) Temp °C Loamy Slope 41.72466 -118.4035 355 8.8 1690 023XY039NV Wyoming big Fine, smectitic, mesic 10-14 sagebrush/bluebunch Aridic Argixerolls wheatgrass Loamy 14-16 41.72649 -118.3970 355 8.8 1732 023XY007NV Mountain big Loamy-skeletal, mixed, sagebrush/Idaho fescue- superactive, frigid Pachic bluebunch wheatgrass Argixerolls Gravelly North 41.72219 -118.3877 380 8.4 1807 023XY053NV 3-tip sagebrush-Idaho Loamy-skeletal, mixed, Slope 14-18 Fescue superactive, frigid Pachic Haploxerolls Claypan 14-16 41.91105 -118.1310 472 8.2 1861 023XY017NV Low sagebrush/Idaho Clayey, smectitic, frigid #2 fescue-bluebunch Aridic Lithic Argixerolls wheatgrass Claypan 10-14 41.85983 -118.1068 509 7.6 1931 023XY031NV Alkali sagebrush/Idaho Clayey, smectitic, frigid fescue-bluebunch Lithic Xeric Haplargids wheatgrass Claypan 14-16 41.7797 -118.1116 525 6.6 1997 023XY017NV Low sagebrush/Idaho Fine, smectitic, frigid #1 fescue-bluebunch Aridic Argixerolls wheatgrass *30-year (1981-2010) annual precipitation and mean annual temperature provided by the Parameter-elevation Regressions on Independent Slopes Model (PRISM) Climate Group, Oregon State University, http://prism.oregonstate.edu, created 4 Feb 2004
85
Table 2 Percent (%) composition by species for aboveground cover in unburned and burned, 6 years after the August 2012 Holloway fire. Species are grouped into life forms. Forb consumed by sage-grouse are followed by an asterisk *. Only species that composed at least 1% of cover for either unburned or burned sites are listed individually; all remaining species are grouped together for simplicity. Ecological Site Genus / Species Common Name Unburned (%) Burned (%) Difference (%) Loamy Slope 10-14 Annual forbs 6.2 44.5 38.3 Ceratocephala testiculata bur buttercup 2.2 4.1 1.9 Eriastrum wilcoxii* Wilcox's woollystar 2.7 0.0 -2.7 Sisymbrium altissimum tall tumblemustard 0.0 40.1 40.1 Other species 1.3 0.3 -1.0 Annual grasses 1.3 40.7 39.4 Bromus tectorum cheatgrass 1.3 40.7 39.4 Perennial forbs 22.9 4.4 -18.5 Ionactis alpina* lava aster 0.0 1.8 1.8 Lupinus argenteus* silvery lupine 14.6 1.7 -12.9 Penstemon kingii* King’s beardtongue 3.5 0.0 3.5 Phlox hoodii spiny phlox 4.4 0.0 -4.4 Other species 0.4 0.9 0.5 Perennial grasses 49.6 12.3 -37.3 Achnatherum thurberianum Thurber’s needlegrass 16.1 2.7 -13.4 Elymus elymoides squirreltail 5.4 1.3 -4.1 Festuca idahoensis Idaho fescue 11.1 0.3 -10.8 Leymus cinereus basin wildrye 1.4 1.7 0.3 Poa secunda sandberg bluegrass 11.9 3.8 -8.1 Pseudoroegneria spicata bluebunch wheatgrass 3.0 2.5 -0.5 Other species 0.7 0.0 -0.7 Shrubs 19.9 <0.1 -19.8 Artemisia tridentata ssp. vaseyana Mountain big sagebrush 18.9 0.0 -18.9 Artemisia tridentata ssp. wyomingensis Wyoming big sagebrush 1.0 0.0 -1.0 Other species 0.0 <0.1 <0.1 Loamy 14-16 Annual forbs 28.1 14.4 -13.7 Ceretocephala testiculata bur buttercup 1.8 0.5 -1.3 Gayophytum decipiens* deceptive groundsmoke 0.0 3.2 3.2 Sisymbrium altissimum tall tumblemustard 0.0 5.9 5.9 Uknown species 25.9 0.8 -25.1 Other species 0.4 4.0 3.6 Annual grasses 14.2 12.1 -2.1 Bromus tectorum cheatgrass 13.8 12.1 -1.7 Other species 0.4 0.0 0.4 Perennial forbs 37.2 45.4 8.2 Astragalus iodanthus* violet milkvetch 0.4 3.2 2.8 Lupinus argenteus* silvery lupine 23.7 34.4 10.7 Phlox hoodii spiny phlox 9.2 1.6 -7.6 Phlox longifolia* longleaf phlox 1.8 3.0 1.2 86
Senecio integerrimus* western groundsel 2.1 0.0 -2.1 Trifolium gymnocarpon* hollyleaf clover 0.0 1.9 1.9 Other species 0.0 1.3 1.3 Perennial grasses 18.2 31.9 13.7 Achnatherum thurberianum Thurber’s needlegrass 5.2 10.6 5.4 Achnatherum webberi Webber’s needlegrass 4.1 0.0 -4.1 Elymus elymoides squirreltail 3.4 1.6 -1.8 Festuca idahoensis Idaho fescue 1.9 6.8 4.9 Poa cusickii Cusick’s bluegrass 1.5 2.8 1.3 Pseudoroegneria spicata bluebunch wheatgrass 1.5 10.1 8.6 Other species 0.6 0.0 -0.6 Shrubs 9.4 0.0 -9.4 Artemisia tridentata ssp. vaseyana Mountain big sagebrush 8.7 0.0 -8.7 Other species 0.7 0.0 -0.7 Gravelly North Slope 14-18 Annual forbs 5.2 1.1 -4.1 Eriastrum wilcoxii* Wilcox's woollystar 3.2 0.5 -2.7 Collomia linearis* tiny trumpet 1.2 0.3 -0.9 Other species 0.8 0.3 -0.5 Annual grasses 1.2 19.3 18.1 Bromus tectorum cheatgrass 1.2 19.3 18.1 Perennial forbs 26.9 42.5 15.6 Astragalus iodanthus* violet milkvetch 0.0 3.1 3.1 Frtillaria atropurpurea spotted fritillary 2.4 0.0 -2.4 Lupinus argenteus* silvery lupine 4.8 24.9 20.1 Phlox diffusa spreading phlox 4.4 6.1 1.7 Phlox hoodii spiny phlox 14.5 6.6 -7.9 Other species 0.8 1.8 1.0 Perennial grasses 35.0 15.1 -19.9 Achnatherum thurberianum Thurber’s needlegrass 7.7 2.9 -4.8 Festuca idahoensis Idaho fescue 25.2 10.7 -14.5 Poa cusickii Cusick’s bluegrass 1.2 0.9 -0.3 Other species 0.9 0.6 -0.3 Shrubs 23.3 0.1 -23.2 Artemisia tripartita three-tip sagebrush 20.3 0.0 -20.3 Artemisia tridentata ssp. vaseyana Mountain big sagebrush 1.5 0.0 -1.5 Artemisia arbuscula low sagebrush 1.5 0.0 -1.5 Other species 0.0 0.1 0.1 Claypan 14-16 #2 Annual forbs 13.1 1.0 -12.1 Epilobium brachycarpum* annual willowherb 4.5 0.3 -4.2 Lomatium ssp.* desert parsely 8.1 0.0 -8.1 Other species 0.5 0.7 0.2 Annual grasses 0.2 50.5 50.3 Bromus tectorum cheatgrass 0.2 50.5 50.3 Perennial forbs 74.3 29.0 -45.3 87
Balsamorhiza hookeri* Hooker’s balsamroot 7.6 0.0 -7.6 Eremogone kingii* King’s sandwort 2.9 1.9 -1.0 Eriogonum sphaerocephalum wild buckwheat 1.3 0.0 -1.3 Lupinus ssp.* lupine 3.4 22.5 -19.1 Phlox hoodii spiny phlox 37.6 0.0 -37.6 Phlox longifolia* longleaf phlox 9.4 1.2 -8.2 Polygonum douglasii Douglas’s knotweed 0.0 1.2 1.2 Senecio integerrimus* western groundsel 9.9 2.2 -7.7 Trifolium gymnocarpon* hollyleaf clover 1.3 0.0 -1.3 Other species 0.9 0.0 -0.9 Perennial grasses 3.7 11.0 7.3 Achnatherum thurberianum Thurber’s needlegrass 0.0 1.2 1.2 Elymus elymoides squirreltail 1.5 4.3 2.8 Elymus lanceolatus thickspike wheatgrass 0.0 1.1 1.1 Leymus cinereus basin wildrye 0.0 3.3 3.3 Poa secunda sandberg bluegrass 1.5 0.1 -1.4 Other species 0.7 1.0 0.3 Shrubs 8.7 7.3 -1.4 Artemisia arbuscula low sagebrush 8.7 0.1 -8.6 Chrysothamnus viscidiflorus yellow rabbitbrush 0.0 7.2 7.2 Claypan 10-14 Annual forbs 8.6 22.7 14.1 Ceratocephala testiculata bur buttercup 1.1 2.6 1.5 Gayophytum decipiens* deceptive groundsmoke 0.4 5.2 4.8 Lomatium ssp.* desert parsley 0.0 6.7 6.7 Microsteris gracillis* slender phlox 7.1 4.5 -2.6 Sisymbrium altissimum tall tumblemustard 0.0 1.3 1.3 Uknown species 0.0 2.4 2.4 Annual grasses 3.6 5.8 2.2 Bromus tectorum cheatgrass 3.6 5.8 2.2 Perennial forbs 68.5 55.4 -13.1 Balsamorhiza hookeri* Hooker’s balsamroot 12.8 16.5 3.7 Crepis acuminata* tapertip hawksbeard 34.1 12.9 -21.2 Eremogone kingii* King’s sandwort 6.0 7.1 1.1 Lupinus ssp.* lupine 9.6 0.2 -9.4 Penstemon kingii* King’s penstemon 2.5 0.0 -2.5 Phlox hoodii spiny phlox 2.1 10.1 8.0 Phlox longifolia* longleaf phlox 1.4 8.4 7.0 Other species 0.0 0.2 0.2 Perennial grasses 6.4 16.2 9.8 Elymus elymoides squirreltail 2.2 8.1 5.9 Festuca idahoensis Idaho fescue 1.7 0.2 -1.5 Poa canbyi Canby’s bluegrass 1.1 4.1 3.0 Pseudoroegneria spicata bluebunch wheatgrass 1.0 3.8 2.8 Other species 0.4 0.0 0.4 88
Shrubs 12.9 0.1 -12.8 Artemisia arbuscula ssp. longiloba Alkali (early) sagebrush 12.9 0.0 -12.9 Other shrubs 0.0 0.1 0.1 Claypan 14-16 Annual forbs 4.8 0.7 -4.1 Microsteris gracillis* slender phlox 2.9 0.0 -2.9 Cordylanthus ramosus bushy bird’s beak 1.3 0.0 -1.3 Other species 0.0 0.2 0.2 Uknown species 0.6 0.5 -0.1 Annual grasses 0.0 4.0 4.0 Bromus tectorum cheatgrass 0.0 4.0 4.0 Perennial forbs 65.7 74.5 8.8 Balsamorhiza hookeri* Hooker’s balsamroot 13.5 38.2 24.7 Eremogone kingii* King’s sandwort 3.2 0.2 -3.0 Lupinus* lupine 15.4 0.0 -15.4 Phlox hoodii spiny phlox 0.0 3.6 3.6 Phlox longifolia* longleaf phlox 8.0 15.2 7.2 Senecio integerrimus* western groundsel 21.5 12.3 -9.2 Trifolium gymnocarpon* hollyleaf clover 3.5 5.0 1.5 Other species 0.6 0.0 -0.6 Perennial grasses 16.6 20.5 3.9 Festuca idahoensis Idaho fescue 12.3 2.8 -9.5 Elymus elymoides squirreltail 1.4 11.1 9.7 Poa canbyi Canby’s bluegrass 2.7 5.8 3.1 Other species 0.2 0.8 0.6 Shrubs 12.9 0.3 -12.6 Artemisia arbuscula low sagebrush 12.7 0.0 -12.7 Other species 0.2 0.3 0.1 89
Table 3 Percent (%) composition by species for the seed bank in unburned and burned 6 years after the August 2012 Holloway fire. Species are grouped into life forms (annual forbs, annual grasses, and perennial forbs). Species are grouped into life forms. Forb consumed by sage-grouse are followed by an asterisk *. Only species that composed at least 1% of the seed bank for either unburned or burned sites are listed individually; all remaining species are grouped together for simplicity. Ecological Site Genus / Species Common Name Unburned (%) Burned (%) Difference (%) Loamy Slope 10-14 Annual forbs 29.3 28.6 -0.7 Ceratocephala testiculata bur buttercup 17.3 22.1 4.8 Collinsia parvifolia* blue-eyed Mary 4.2 0.6 -3.6 Gayophytum decipiens* deceptive groundsmoke 1.2 0.0 -1.2 Microsteris gracilis* slender phlox 4.2 2.5 -1.7 Sisymbrium altissimum tall tumblemustard 2.4 3.1 0.7 Other species 0.0 0.3 0.3 Annual grasses 4.8 65.6 60.8 Bromus tectorum cheatgrass 4.8 65.6 60.8 Perennial grasses 64.3 5.3 -59.0 Elymus elymoides squirreltail 29.2 0.9 -28.3 Poa secunda sandberg bluegrass 35.1 4.4 -30.7 Shrubs 1.8 0.4 -1.4 Artemisia tridentata ssp. big sagebrush 1.8 0.4 -1.4 Loamy 14-16 Annual forbs 24.1 26.0 1.9 Ceratocephala testiculata bur buttercup 11.5 0.9 -10.6 Collinsia parvifolia* blue-eyed Mary 4.0 3.4 -0.6 Gayophytum decipiens* deceptive groundsmoke 5.2 0.0 -5.2 Microsteris gracilis* slender phlox 2.3 13.2 10.9 Sisymbrium altissimum tall tumblemustard 0.0 6.8 6.8 Cryptantha fiddleneck 1.1 1.7 0.6 Annual grasses 47.1 44.0 -3.0 Bromus tectorum cheatgrass 37.9 44.0 6.1 Vulpia bromoides six-weeks fescue 9.2 0.0 -9.2 Perennial grasses 27.6 29.5 1.9 Elymus elymoides squirreltail 1.7 5.6 3.9 Poa secunda sandberg bluegrass 25.9 23.9 -2.0 Shrubs 1.1 0.4 -0.7 Artemisia ssp. sagebrush 1.1 0.4 -0.7 Gravelly North Slope 14-18 Annual forbs 32.9 10 -22.9 Ceratocephala testiculata bur buttercup 7.5 2.8 -4.7 Collinsia parvifolia* blue-eyed Mary 6.6 0.3 -6.3 Gayophytum decipiens* deceptive groundsmoke 1.9 0.0 -1.9 Microsteris gracilis* slender phlox 9.4 3.5 -5.9 90
Sisymbrium altissimum tall tumblemustard 0.0 1.0 1.0 Cryptantha fiddleneck 6.6 2.1 -4.5 Other species 0.9 0.3 -0.6 Annual grasses 28.3 62.0 33.7 Bromus tectorum cheatgrass 28.3 61.7 33.4 Vulpia bromoides six-weeks fescue 0.0 0.3 0.3 Perennial grasses 35.8 25.0 -10.8 Elymus elymoides squirreltail 6.6 2.4 -4.2 Poa secunda sandberg bluegrass 29.2 22.6 -6.6 Shrubs 6.6 2.4 -4.2 Artemisia tripartita three-tip sagebrush 6.6 0.0 -6.6 Artemisia ssp. sagebrush 0.0 2.4 2.4 Claypan 14-16 #2 Annual forbs 14.0 12.1 -1.9 Ceratocephala testiculata bur buttercup 0.7 1.0 0.3 Collinsia parvifolia* blue-eyed Mary 7.4 3.6 -3.8 Microsteris gracillis* slender phlox 5.2 7.5 2.3 Other species 0.7 0.0 -0.7 Annual grasses 0.7 79.1 78.4 Bromus tectorum cheatgrass 0.7 79.1 78.4 Perennial grasses 85.1 8.8 -76.3 Elymus elymoides squirreltail 11.1 4.1 -7.0 Poa canbyi Canby’s bluegrass 64.4 3.2 -61.2 Poa secunda sandberg bluegrass 9.6 1.5 -8.1 Claypan 10-14 Annual forbs 51.6 45.2 -6.4 Ceratocephala testiculata bur buttercup 7.9 7.0 -0.9 Collinsia parvifolia* blue-eyed Mary 33.4 22.9 -10.5 Epilobium brachycarpon* annual willowherb 1.7 0.0 -1.7 Microsteris gracillis* slender phlox 8.3 15.3 7.0 Other species 0.3 0.0 -0.3 Annual grasses 9.3 17.5 8.2 Bromus tectorum cheatgrass 9.3 17.5 8.2 Perennial forbs 15.2 2.2 -13.0 Lithophragma woodland star 15.2 2.0 -13.2 Trifolium gymnocarpon* hollyleaf clover 0.0 0.2 0.2 Perennial grasses 23.5 34.7 11.2 Elymus elymoides squirreltail 8.3 3.8 -4.5 Poa secunda sandberg bluegrass 15.2 30.9 15.7 Shrubs 0.3 0.3 0.0 Artemisia sagebrush 0.3 0.3 0.0 91
Claypan 14-16 Annual forbs 34.4 3.3 -31.1 Ceratocephala testiculata bur buttercup 0.0 1.0 1.0 Collinsia parvifolia* blue-eyed Mary 26.3 2.3 -24.0 Microsteris gracilis* slender phlox 7.1 0.0 -7.1 Sisymbrium altissimum tall tumblemustard 1.0 0.0 -1.0 Annual grasses 5.1 36.2 31.1 Bromus tectorum cheatgrass 5.1 36.2 31.1 Perennial forbs 0.0 0.3 0.3 Lupinus lupine 0.0 0.3 0.3 Perennial grasses 59.6 51.2 -8.4 Elymus elymoides squirreltail 1.0 7.0 6.0 Poa secunda sandberg bluegrass 58.6 43.9 -14.7 Other species 0.0 0.3 0.3 Shrubs 1.0 0.0 -1.0 Artemisia tridentata big sagebrush 1.0 0.0 -1.0 92
Table 4 Aboveground cover of plant functional groups compared between paired burned (B) and unburned (UB) plots for each site. All variables are reported by site as mean percent cover (%), standard error (SE) (±), and mean difference of % cover (B—UB). Variable Site UB ± SE (%) B ± SE (%) Difference (%) P(T<=t) a. Forb Claypan 14-16 #1a 3.6 ± 0.7 8.2 ± 1.3 4.6 0.02* Claypan 14-16 #2c 5.6 ± 0.6 2.5 ± 0.5 -3.1 0.03* Claypan 10-14 4.2 ± 0.6 4.5 ± 0.6 0.3 0.78 GNS 14-18 2.3 ± 0.3 5.8 ± 2.3 3.5 0.21 Loamy 14-16+ 4.4 ± 0.8 4.4 ± 0.7 0.1 0.98 Loamy Slope 10-14b 2.0 ± 0.4 13.2 ± 1.9 11.2 0.01* b. Perennial Forb Claypan 14-16 #1a 4.2 ± 0.7 9.0 ± 1.2 4.7 0.04* Claypan 14-16 #2c 6.5 ± 1.0 3.0 ± 0.6 -3.4 0.05* Claypan 10-14+ 5.9 ± 1.0 6.7 ± 1.8 0.8 0.87 GNS 14-18 3.9 ± 1.0 6.4 ± 2.3 2.5 0.46 Loamy 14-16 4.5 ± 1.1 6.8 ± 2.0 2.2 0.46 Loamy Slope 10-14 4.3 ± 1.6 2.1 ± 0.8 -2.2 0.51 c. Annual Forb Claypan 14-16 #1 1.4 ± 0.7 0.5 ± 0.3 -0.9 0.16 Claypan 14-16 #2+ 3.2 ± 0.8 0.6 ± 0.1 -2.6 0.09 Claypan 10-14+ 1.2 ± 0.3 3.5 ± 0.7 2.3 0.03* GNS 14-18+ 0.9 ± 0.2 0.4 ± 0.2 -0.5 0.57 Loamy 14-16+ 7.4 ± 0.6 2.4 ± 0.6 -5.0 0.01* Loamy Slope 10-14b 0.9 ± 0.1 17.7 ± 2.9 16.8 0.003** d. Native Forb Claypan 14-16 #1a 3.7 ± 0.6 9.0 ± 1.2 5.2 0.03* Claypan 14-16 #2+ c 5.7 ± 0.6 2.5 ± 0.5 -3.2 0.02* Claypan 10-14+ 4.7 ± 0.6 5.0 ± 0.8 0.3 0.83 GNS 14-18 2.4 ± 0.3 5.8 ± 2.3 3.4 0.22 Loamy 14-16 4.3 ± 1.1 6.0 ± 2.2 1.8 0.62 Loamy Slope 10-14 3.8 ± 1.6 2.1 ± 0.8 -1.7 0.46 e. Non-Native Forb Claypan 14-16 #1 ------Claypan 14-16 #2 ------Claypan 10-14 0.4 ± 0.1 2.3 ± 0.9 1.9 0.065 GNS 14-18 ------Loamy 14-16 0.6 ± 0.3 2.5 ± 1.7 1.9 0.762 Loamy Slope 10-14b 0.6 ± 0.1 20.5 ± 4.4 19.9 0.003** f. Native Annual Forb Claypan 14-16 #1+ 1.3 ± 0.7 0.1 ± 0.1 -1.2 0.09 Claypan 14-16 #2+ 3.5 ± 1.0 0.6 ± 0.1 -2.9 0.09 Claypan 10-14 1.6 ± 0.5 2.8 ± 0.8 1.2 0.13 GNS 14-18+ 0.8 ± 0.2 0.4 ± 0.2 -0.4 0.48 Loamy 14-16+ 0.1 ± 0.1 1.0 ± 0.4 0.9 0.18 Loamy Slope 10-14+ 0.8 ± 0.3 0.0 ± 0.0 -0.8 0.06 g. Sage-Grouse Forbs Claypan 14-16 #1a 4.0 ± 0.6 10.1 ± 1.1 6.1 0.001** Claypan 14-16 #2+ c 7.2 ± 0.7 4.3 ± 1.2 -2.8 0.13 Claypan 10-14 5.7 ± 1.0 5.7 ± 1.0 0.1 0.99 GNS 14-18 3.9 ± 1.0 9.2 ± 4.9 5.4 0.42 Loamy 14-16 4.4 ± 1.2 7.3 ± 1.8 2.8 0.30 Loamy Slope 10-14 4.3 ± 1.6 1.6 ± 0.9 -2.7 0.29 h. Cheatgrass Claypan 14-16 #1a 0.0 ± 0.0 2.1 ± 0.7 2.1 0.003** Claypan 14-16 #2+ 0.1 ± 0.1 20.3 ± 3.8 20.1 0.82 Claypan 10-14 1.3 ± 0.3 3.4 ± 1.5 2.1 0.29 GNS 14-18 0.4 ± 0.2 9.5 ± 3.2 9.1 0.98 Loamy 14-16 4.9 ± 1.7 5.6 ± 0.9 0.8 0.57 Loamy Slope 10-14+ 0.4 ± 0.4 39.9 ± 6.3 39.5 0.94 i. Perennial Grass Claypan 14-16 #1a 1.8 ± 0.1 2.7 ± 0.4 0.8 0.06 Claypan 14-16 #2 0.7 ± 0.1 1.0 ± 0.1 0.2 0.04* Claypan 10-14 0.6 ± 0.1 2.8 ± 0.4 2.2 0.003** GNS 14-18 2.9 ± 0.4 4.3 ± 0.9 2.5 0.30 Loamy 14-16 1.1 ± 0.1 3.0 ± 0.3 1.8 0.009** Loamy Slope 10-14 2.4 ± 0.3 2.1 ± 0.3 -0.3 0.52 j. Shrub Claypan 14-16 #1 3.5 ± 0.8 0.1 ± 0.1 -3.3 0.41 Claypan 14-16 #2 4.8 ± 0.8 0.1 ± 0.1 -3.3 0.006** Claypan 10-14+ 4.0 ± 0.6 0.0 ± 0.0 -4.0 0.33 GNS 14-18+ 4.4 ± 1.3 0.1 ± 0.1 -4.3 0.37 Loamy 14-16+ 2.2 ± 0.8 0.0 ± 0.0 -2.2 0.002** Loamy Slope 10-14 2.8 ± 0.3 0.0 ± 0.0 -2.8 0.333 + locations with variables that did not assume normality on log-transformed means. * Treatment has a significant effect on cover group within site, p < 0.05 ** Treatment has a significant effect on cover group within site, p < 0.01 a, b, c Significant variables containing the same letter are correlated within site (Spearman’s correlation coefficient > ± 0.75) Variables: GNS = Gravelly North Slope 93
Table 5 Species that germinated in the soil seed bank assay organized into plant functional groups and compared between burned (B) and unburned (UB) areas for each site. All variables are reported as mean density of seeds per treatment (seeds•m-2), standard error (SE) (±), and mean difference of mean density of seeds (B – UB). Variable Site UB ± SE B ± SE Difference P(T<=t) (seeds•m-2) (seeds•m-2) (seeds•m-2) a. Forb Claypan 14-16 #1 269 ± 91 198 ± 100 -72 0.42 Claypan 14-16 #2 106 ± 28 326 ± 55 219 0.07 Claypan 10-14 682 ± 327 880 ± 170 199 0.44 GNS 14-18 112 ± 10 124 ± 49 12 0.75 Loamy 14-16 214 ± 51 225 ± 47 12 0.89 Loamy Slope 10-14a 166 ± 43 990 ± 305 824 0.01* b. Native Forb Claypan 14-16 #1 273 ± 89 228 ± 118 -46 0.44 Claypan 14-16 #2 111 ± 27 326 ± 55 215 0.07 Claypan 10-14 768 ± 350 979 ± 308 211 0.63 GNS 14-18 100 ± 11 93± 41 -7 0.41 Loamy 14-16 156 ± 43 273 ± 88 111 0.48 Loamy Slope 10-14a 89 ± 15 200 ± 46 26 0.03* c. Annual Forb Claypan 14-16 #1 269 ± 91 198 ± 100 -72 0.42 Claypan 14-16 #2 106 ± 28 326 ± 55 219 0.07 Claypan 10-14 674 ± 355 981 ± 227 307 0.33 GNS 14-18 112 ± 10 124 ± 49 12 0.72 Loamy 14-16 214 ± 51 225 ± 47 12 0.89 Loamy Slope 10-14a 166 ± 43 990 ± 305 824 0.01* d. Perennial Forb Claypan 14-16 #1 0 ± 0 13 ± 13 13 0.391 Claypan 14-16 #2 ------Claypan 10-14+ 599 ± 325 143 ± 126 -456 0.415 GNS 14-18 ------Loamy 14-16 ------Loamy Slope 10-14 ------e. Non-Native Forb Claypan 14-16 #1+ 13 ± 13 39 ± 25 26 0.31 Claypan 14-16 #2+ 13 ± 13 52 ± 52 39 0.87 Claypan 10-14+ 326 ± 291 456 ± 314 130 0.81 GNS 14-18+ 104 ± 60 143 ± 94 39 0.50 Loamy 14-16 260 ± 122 163 ± 16 -98 0.62 Loamy Slope 10-14a 326 ± 157 2175 ± 694 1849 0.01* f. Sage-Grouse Forbs Claypan 14-16 #1 273 ± 89 228 ± 118 -046 0.44 Claypan 14-16 #2 111 ± 27 326 ± 55 215 0.07 Claypan 10-14 814 ± 406 1055 ± 347 241 0.61 GNS 14-18 90 ± 11 111 ± 57 21 0.52 Loamy 14-16 169 ± 44 319 ± 124 150 0.52 Loamy Slope 10-14a 89 ± 15 267 ± 72 178 0.02* g. Cheatgrass Claypan 14-16 #1+ 65 ± 33 1419 ± 497 1354 0.005** Claypan 14-16 #2+ 13 ± 13 4245 ± 1641 4232 0.01* Claypan 10-14 365 ± 189 1133 ± 549 768 0.19 GNS 14-18+ 391 ± 305 2305 ± 350 1914 0.04* Loamy 14-16+ 859 ± 377 1341 ± 589 482 0.36 Loamy Slope 10-14a+ 104 ± 56 11304 ± 1180 11199 0.02* h. Perennial Grass Claypan 14-16 #1 573 ± 97 996 ± 236 423 0.30 Claypan 14-16 #2 499 ± 63 198 ± 46 -302 0.08 Claypan 10-14 501 ± 173 1393 ± 468 892 0.04* GNS 14-18 384 ± 105 553 ± 72 169 0.27 Loamy 14-16 560 ± 278 449 ± 121 -111 0.88 Loamy Slope 10-14 703 ± 147 703 ± 286 0 0.74 i. Shrubs Claypan 14-16 #1+ 0 ± 0 13± 13 13 0.39 Claypan 14-16 #2 0 ± 0 0 ± 0 0 -- Claypan 10-14+ 0 ± 0 26 ± 15 26 0.18 GNS 14-18+ 39 ± 25 0 ± 0 -39 0.18 Loamy 14-16+ 52 ± 21 130 ± 78 78 0.89 Loamy Slope 10-14+ 39 ± 39 65 ± 65 26 0.96 + locations with variables that did not assume normality on log-transformed means. Variables: GNS = Gravelly North Slope, GRSG = Greater sage-grouse * Treatment has a significant effect on cover group within site, p < 0.05 ** Treatment has a significant effect on cover group within site, p < 0.01 a Significant variables containing the same letter are correlated within site (Spearman’s correlation coefficient > ± 0.75) 94
Table 6 Sites, elevation (m), treatment (trt), and Bray-Curtis (BC) similarity values (1—BC dissimilarity) between the seed bank and aboveground vegetation for overall site similarity (BC similarity) and for forbs only (BC forbs similarity).
Site Elevation (m) Trt (BC) similarity (BC) forbs similarity Loamy Slope 10-14 1690 UB 0.275 0.137 B 0.580 0.455 Loamy 14-16 1732 UB 0.219 0.095 B 0.229 0.152 GNS 14-18 1807 UB 0.128 — B 0.175 — Claypan 14-16 #2 1861 UB 0.081 0.063 B 0.343 — Claypan 10-14 1931 UB 0.230 0.185 B 0.272 0.158 Claypan 14-16 #1 1997 UB 0.069 0.013 B 0.209 0.015
95
Table 7 Sites, and percent species composition in aboveground percent cover and belowground seed density, by life habit (annual, perennial) in unburned (UB) and burned (B) areas for overall species and the forb group only.
Overall (%) Forbs (%) Above Below Above Below Sites Life Habit UB B UB B UB B UB B
Claypan 14-16 #1 Annual 4 5 39 49 4 0.5 34 12 Perennial 96 95 61 51 66 75 0 0.3
Claypan 14-16 #2 Annual 13 52 15 91 13 2 14 12
Perennial 87 48 85 9 74 28 0 0 Claypan 10-14 Annual 12 28 61 63 9 23 52 45 Perennial 88 72 39 37 69 55 15 2 GNS 14-18 Annual 6 21 62 72 5 1 34 10 Perennial 94 79 38 28 27 42 0 0
Loamy 14-16 Annual 16 23 71 70 2 10 24 26 Perennial 65 77 29 30 37 45 0 0 Loamy Slope Annual 7 85 34 94 5 44 29 29
Perennial 93 15 66 6 23 3 0 0
96
Table 8 Sites, elevations (m), treatment (trt), Shannon-Weaver diversity index values (H’), Hutcheson’s t-statistic (t), probability values (P), Effective Species (Effect S), Species richness (S), and Shannon’s Evenness (J’) between paired burned (B) and unburned (UB) treatments in aboveground and soil seed bank diversity and forb diversity. Overall Diversity Aboveground Belowground Site Elev (m) Trt H’ t P Effect S S J’ H’ t P Effect S S J’ Loamy Slope 10-14 1690 UB 2.4 9.2 2.2 E-17** 10.6 18.0 0.8 1.3 1.0 0.30 3.5 9.0 0.6 B 1.4 4.2 12.0 0.6 1.1 2.9 10.0 0.5 Loamy 14-16 1732 UB 2.3 0.3 0.77 9.6 19.0 0.8 1.7 0.6 0.54 5.6 10.0 0.8 B 2.2 9.2 19.0 0.8 1.6 4.8 9.0 0.7 GNS 14-18 1807 UB 2.2 1.8 0.08 9.4 18.0 0.8 1.9 3.9 0.0001** 6.8 11.0 0.8 B 2.0 7.6 20.0 0.7 1.2 3.4 11.0 0.5 Claypan 14-16 #2 1861 UB 2.1 3.7 0.0002** 8.5 18.0 0.7 1.3 1.2 0.22 3.7 7.0 0.7 B 1.7 5.2 19.0 0.6 0.8 2.3 7.0 0.4 Claypan 10-14 1931 UB 2.2 3.4 0.0007** 8.8 18.0 0.8 1.9 2.3 0.02* 6.6 10.0 0.8 B 2.5 12.4 18.0 0.9 1.7 5.6 10.0 0.7 Claypan 14-16 #1 1997 UB 2.2 3.3 0.0009** 9.2 16.0 0.8 1.1 1.5 0.15 3.1 7.0 0.6 B 1.9 6.8 14.0 0.7 1.3 3.7 8.0 0.6 Forb Diversity Aboveground Belowground Loamy Slope 10-14 1690 UB 1.5 3.7 0.0005** 4.5 8.0 0.7 1.2 3.0 0.004** 3.3 5.0 0.7 B 0.8 2.2 6.0 0.4 0.8 2.2 6.0 0.4 Loamy 14-16 1732 UB 1.2 2.8 0.005** 3.2 7.0 0.6 1.4 0.4 0.67 3.9 5.0 0.8 B 1.7 5.3 14.0 0.6 1.3 3.5 5.0 0.8 GNS 14-18 1807 UB 1.6 1.4 0.15 5.0 8.0 0.8 1.7 1.0 0.31 5.4 7.0 0.9 B 1.3 3.8 11.0 0.6 1.5 4.6 6.0 0.8 Claypan 14-16 #2 1861 UB 1.8 5.0 5.4 E-06** 6.1 11.0 0.8 1.0 0.8 0.42 2.8 4.0 0.7 B 0.9 2.4 6.0 0.5 0.9 2.4 3.0 0.8 Claypan 10-14 1931 UB 1.7 4.3 3.5 E-05** 5.5 10.0 0.7 1.3 1.9 0.06 3.7 6.0 0.7 B 2.2 8.6 12.0 0.9 1.2 3.3 6.0 0.7 Claypan 14-16 #1 1997 UB 1.8 4.8 2.4 E-06** 6.1 9.0 0.8 0.7 0.5 0.62 2.1 4.0 0.5 B 1.3 3.8 7.0 0.7 0.9 2.3 4.0 0.6 Site: GNS = Gravelly North Slope *p < .05 **p < .01 97
Table 9 Summary of linear regression analysis for variables: belowground Effective species (below S) in unburned, and below S in burned predicting aboveground overall diversity and aboveground forb diversity. Regression estimates, standard errors (SE), and probability values (P) are reported. Overall Aboveground Forbs Aboveground
Variables Estimate SE P Estimate SE P
Below S:Unburned -0.10 0.27 0.72 -0.08 0.16 0.62 Below S:Burned 2.30 0.36 0.0002** 0.27 0.19 0.21 R2 0.86 0.29 Adj. R2 0.81 0.02 F 17.01** 1.09
*p < .05 **p < .01
98
Figure 1 Six research sites selected along burn boundaries of the 2012 Holloway fire in the Bilk Creek, Trout Creek, and Montana Mountains located in Humboldt County, NV. Sites identified were within the Holloway fire burn boundary by association with apparent fire scarring (Bureau of Land Management, 2019) and within the Nevada Department of Wildlife Lone Willow sage-grouse Population Management Unit.
99
Figure 2 Mean relative percent aboveground cover of all species recorded across all six sites. Proportion of species identified as forb, grass, or shrub are indicated by shade. Poa canbyi (POCA) and Poa cusickii (POCU3) are recognized separately from Poa secunda (POSE) based on phenology and presence that was unique within ecological sites.
Plant codes: ACNEN2 = Achnatherum nelsonii (western needlegrass), ACWE3 = Achnatherum webberi (Webber needlegrass), ACTH7 = Acnatherum thurberium (Thurber’s needlegrass), ALAC4 = Allium acuminatum (tapertip onion), ARARL = Artermisia arbuscula ssp. longiloba (Alkali sagebrush), ARARA = Artemisia tridentata ssp. arbuscula (low sagebrush), ARTRV = Artemisia tridentata ssp. vayesana (mountain big sagebrush), ARTRW8 = Artemisia tridentata ssp. wyomingensis (Wyoming big sagebrush), ASIO = Astragalus iodanthus (Humboldt river milkvetch), ARTR4 = Artemisia tripartita (three-tip sagebrush), BAHO = Balsamorhiza hookeri (Hooker’s balsamroot), BASA3 = Balsamorhiza sagittata (arrowleaf balsamroot), BRTE = Bromus tectorum (cheatgrass), CETE5 = Ceratocephala testiculata (bur buttercup), CHDO = Chaenactis douglasii (Douglas’ dustymaiden), CHVIP4 = Chrysothamnus viscidiflorus ssp. puberulus (yellow rabbitbrush), CHVI8 = Chrysothamnus viscidiflorus (yellow rabbitbrush), COLI2 = Collomia linearis (tiny trumpet), CORA5 = Cordylanthus ramosus (bushy bird’s beak), CRAC2 = Crepis acuminata (tapertip hawksbeard), DEPI = Descurainia pinnata (western tansymustard), ELEL5 = Elymus elymoides (squirreltail), ELLA3 = Elymus lanceolatus (thickspike wheatgrass), EPBR3 = Epilobium brachycarpum (annual willowherb), ERKI2 (ARKIK) = Eremogone (Arenaria) kingii (King’s sandwort), ERWI = Eriastrum wilcoxii (Wilcox’s woolystar), ERSP7 = Eriogonum sphaerocephalum (rock buckwheat), FEID = Festuca idahoensis (Idaho fescue), FRAT = Fritillaria atropurpurea (spotted fritillary), GADE2 = Gayophytum decipiens (deceptive groundsmoke), IOAL = Ionactis alpina (lava aster), LECI4 = Leymus cinerius (basin wildrye), LETR5 = Leymus triticoides (creeping wildrye), LOMAT = Lomatium ssp. (desert parsley), LUAR = Lupinus argenteus (silver lupine), LUPIN = Lupinus ssp. (lupine), MAEX = Madia exigua (small tarweed), MEAL6 = Mentzelia albicaulis (whitestem blazingstar), MIGR = Microsteris gracillis (slender phlox), PEKI = Penstemon kingii (King’s beardtongue), PHDI3 = Phlox diffusa (spreading phlox), PHHO = Phlox hoodii (cushion phlox), PHLO2 = Phlox longifolia (longleaf phlox), PLSP7 = Pleiacanthus spinosus (thorn skeletonweed), POCA = Poa canbyi (Canby’s bluegrass), POCU3 = Poa cusickii (Cusick’s bluegrass), POSE = Poa secunda (sandberg bluegrass), PODOJ2 = Polygonum douglasii ssp. johnstonii (Johnston’s knotweed), PSSP6 = Pseudoroegneria spicata (bluebunch wheatgrass), PUTR2 = Purshia tridentata (antelope bitterbrush), SEIN2 = Senecio integerrimus (lambstongue ragwort), SIAL2 = Sisymbrium altissimum (tall tumblemustard), TECA2 = Tetradymia canescens (smooth horsebrush), TRGY = Trifolium gymnocarpon (hollyleaf clover), VUBR = Vulpia bromoides (six-weeks fescue), ZIVE = Zigadenus venenosus (meadow deathcamas), uknown = uknown genera and species of annual forb 100
Figure 3 Mean relative density of all species that germinated from the soil seed bank samples across sites. Proportion of species identified as forb, grass, or shrub are indicated by shade. The species, Poa canbyi (POCA), is recognized as unique from Poa secunda (POSE) based on phenology unique to ecological site descriptions.
Plant codes: ARABI2 = Arabis ssp. (rockcress), ARTEM = Artemisia ssp. (sagebrush), ARTRW8 = Artemisia tridentata ssp. wyomingensis (Wyoming big sagebrush), ARTR4 = Atremisia tripartita (three-tip sagebrush), BRTE = Bromus tectorum (cheatgrass), CETE5 = Ceratocephala testiculata (bur buttercup), COPA3 = Collinsia parvifolia (blue-eyed Mary), CRYPT = Cryptantha ssp. (fiddleneck), ELEL5 = Elymus elymoides (squirreltail), EPBR3 = Epilobium brachycarpum (annual willowherb), GADE2 = Gayophytum decipiens (deceptive groundsmoke), GNPA = Gnaphalium palustre (western marsh cudweed), JUBU = Juncus bufonius (toad rush), LITHO2 = Lithophragma (woodland star), LOMAT = Lomatium ssp. (desert parsley), LUPIN = Lupinus ssp. (lupine), MIGR = Microsteris gracillis (slender phlox), NEMOP = Nemophila ssp. (baby blue eyes), POCA = Poa canbyi (Canby’s bluegrass), POSE = Poa secunda (sandberg bluegrass), SIAL2 = Sisymbrium altissimum (tall tumblemustard), TECA2 = Tetradymia canescens (smooth horsebrush), TRGY = Trifolium gymnocarpon (hollyleaf clover), VUBR = Vulpia bromoides (six-weeks fescue) 101
Figure 4 Plant species shared between above- and belowground plant communities across sites ascending from lowest (1690 m) to highest elevation (1997 m) in unburned and burned areas. Aboveground cover reported as mean relative percent (%) cover. Belowground reported as mean relative density of seeds that germinated from the soil seed bank.
Plant codes: BRTE = Bromus tectorum (cheatgrass), POSE = Poa secunda (sandberg bluegrass), ELEL5 = Elymus elymoides (squirreltail), MIGR = Microsteris gracillis (slender phlox), CETE5 = Ceratocephala testiculata (bur buttercup), SIAL2 = Sisymbrium altissimum (tall tumblemustard), POCA = Poa canbyi (Canby bluegrass), LUPIN = Lupinus ssp. (lupine), GADE2 = Gayophytum decipiens (deceptive groundsmoke), ARTR4 = Atremisia tripartita (three-tip sagebrush), LOMAT = Lomatium ssp. (desert parsley), VUBR = Vulpia bromoides (six-weeks fescue), EPBR3 = Epilobium brachycarpum (annual willowherb), ARTRW8 = Artemisia tridentata ssp. wyomingensis (Wyoming big sagebrush), TECA2 = Tetradymia canescens (smooth horsebrush) 102
Figure 5 a) Mean percent (%) aboveground cover of observed species by site, categorized into 9 functional groups and compared between unburned (UB) and burned (B) treatments across ecological sites ascending from lowest (1690 m) to highest elevation (1997 m). b) Mean density of plant species (seeds•m2) that germinated in the soil seed bank at each site, categorized into 8 functional groups and compared between UB and B. Means and standard error (SE) (±) bars reported. a)
**
**
*
** **
** * * * * * * ** * ** *
b)
*
* *
• * *
*
* Treatment has a significant effect on cover group within site, p < 0.05 ** Treatment has a significant effect on cover group within site, p < 0.01 Variables: GNS = Gravelly North Slope, AF = annual forb, GRSG = Greater sage-grouse, NAF = native annual forb, NF = native forb, PG = perennial grass 103
Figure 6 a) Mean percent (%) cover of cheatgrass between unburned (UB) and burned (B) areas across sites in order from lowest (1690 m) to highest elevation (1997 m). b) Mean density (seeds•m2) of cheatgrass seeds that germinated in the soil seed bank at each site between UB and B areas across sites in order from lowest (1690 m) to highest elevation (1997 m). Means and standard error (SE) (±) bars reported. a) b)
*
*
*
** **
* Treatment has a significant effect on cover group within site, p < 0.05 ** Treatment has a significant effect on cover group within site, p < 0.01 Variables: GNS = Gravelly North Slope, UB = unburned, B = burned 104
Figure 7 a.1) A partitioning around medoids (PAM) cluster analysis using Bray-Curtis (BC) dissimilarity that compared aboveground (a) relative composition of forbs to belowground (b) relative composition of forbs across sites in unburned areas. b.1) the same analysis as in unburned that compared aboveground (a) relative composition of forbs to belowground (b) relative composition of forbs across sites in burned areas. Sites within cluster group (1—3) indicate they are more similar to each other than the sites outside their cluster group.
a.1)
Group 1 Loamy Slope 10-14 (a) GNS Group 2 14-18 (a) + Group 3
Loamy Slope GNS Loamy 14-16 (a) 10-14 14-18 (b) (b) Claypan Loamy 14-16 (b) 14-16 #2 (b) Claypan 10-14 (b)
Claypan 14-16 #1 (b)
Claypan 14-16 #2 (a)
Claypan 14-16 #1 (a)
Claypan 10-14 (a)
b.1)
Loamy Slope 10-14 (a) Loamy 14-16 (a)
GNS 14-18 (a) Loamy Slope 10-14 (b) Loamy 14-16 (b) GNS 14-18 (b)
Claypan 10-14 (b) Claypan 10-14 (a) Claypan 14-16 #1 (b) Claypan Claypan 14-16 #2 14-16 #1 (a) (b)
Claypan 14-16 #2 (a)
Variables: (a) = aboveground, (b) = belowground, GNS = Gravelly North Slope 105
Figure 8 Plant species identified as indicators in aboveground communities across ecological sites. Pearson’s correlation residuals extracted from chi-squared analysis were significant at ±2. Species with positive residuals indicate a preference for associating with site. Species with negative residuals indicate a preference of disassociation with site.
Claypan 10-14 Claypan Claypan Gravelly Loamy Loamy 14-16 #1 14-16 #2 North Slope 14-16 Slope 10-14 14-18
Plant codes: BAHO = Balsamorhiza hookeri (Hook.) Nutt., CRAC2 = Crepis acuminata Nutt., LOMAT = Lomatium Raf., LUAR3 = Lupinus argenteus Pursh, SEIN2 = Senecio integerrimus Nutt., SIAL2 = Sisymbrium altissimum L. 106
Figure 9 Plant species identified as indicators in belowground communities across ecological sites. Pearson’s correlation residuals extracted from chi-squared analysis were significant at ±2. Species with positive residuals indicate a preference for associating with the seed bank within the site. Species with negative residuals indicate a preference of disassociation with the seed bank within site.
Claypan 10-14 Claypan Claypan Gravelly Loamy Loamy 14-16 #1 14-16 #2 North Slope 14-16 Slope 10-14 14-18
Plant codes: CETE5 = Ceratocephala testiculata (Crantz) Roth., LITHO2 = Lithophragma (Nutt.) Torr. & A. Gray, POCA = Poa canbyi (Scribn.) Howell, POSE = Poa secunda J. Presl. Sites: GNS = Gravelly North Slope, L.Slope = Loamy Slope 107
Figure 10 a) Comparison of Shannon-Weaver diversity index (H) measuring aboveground diversity between unburned (UB) and burned (B) areas across ecological sites in order by ascending elevation. b) Comparison of Shannon-Weaver diversity index (H) measuring belowground diversity between unburned (UB) and burned (B) areas across ecological sites in order by ascending elevation. Error bars indicate 95% confidence intervals. Higher values indicate greater diversity while lower values indicate lower diversity relative to overall diversity by site. a) b)
Variables: GNS = Gravelly North Slope 108
Figure 11 a) Comparison of Shannon-Weaver diversity index (H) for forbs only, measuring aboveground diversity between unburned (UB) and burned (B) areas across ecological sites in order by ascending elevation. b) Comparison of Shannon-Weaver diversity index (H) for forbs only, measuring belowground diversity between unburned (UB) and burned (B) areas across ecological sites in order by ascending elevation. Error bars indicate 95% confidence intervals. Higher values indicate greater forb diversity while lower values indicate lower forb diversity relative to overall forb diversity by site. a) b)
Variables: GNS = Gravelly North Slope 109
Figure 12 Linear model regression of how well the overall belowground diversity can predict aboveground diversity within unburned (UB) and burned areas (B). R2 values are for the linear trendline for the single regression between number of effective species belowground and number of effective species aboveground.
R2 = 0.86 R2 = 0.86 P = 0.72 P = 0.0002
Figure 13 Linear model regression of how well belowground forb diversity can predict aboveground forb diversity within unburned (UB) and burned areas (B). R2 values are for the linear trendline for the single regression between number of effective forb species belowground and number of effective forb species aboveground.
R2 = 0.29 R2 = 0.29 P = 0.62 P = 0.21
110
CHAPTER III: Above- and belowground species diversity in response to 82 years of
grazing exclusion in the Northern Great Basin
INTRODUCTION
Herbivory is a well-known driver of plant community dynamics in the Great Basin
(Courtois et al. 2004; Davies et al. 2010; Bates and Davies 2014; Davies et al. 2016a). The composition of sagebrush communities post-European settlement in the mid-1800s changed through a combination of inappropriate grazing by livestock, fire suppression, introduction of invasive species like cheatgrass, and climate change (Chambers et al.
2007; Davies et al. 2009; Chambers et al. 2014; Morris and Rowe 2014). The U.S.
Department of Interior, Division of Grazing initially assessed the effects of inappropriate grazing through the establishment of 28 sites selected in nine Nevada counties from
1936 to 1939 (Courtois et al. 2004). These sites, known as the Nevada Plots, were established to help us understand how aboveground cover differed in areas that continued grazing in comparison to the exclosures that removed grazing completely.
Today, the legacy effects of improper grazing regimes in the pre-1936 Taylor Grazing Act era have been compounded by historical fire suppression that have contributed to an overall decrease in native deep-rooted perennial bunch grasses and forbs, an increase in annual invasive grasses, and greater sagebrush dominance, especially in lower elevation
Wyoming big sagebrush communities (Morris and Rowe 2014). This shift in plant community composition has contributed to a buildup of fine fuel loads in grazing- excluded and fire-suppressed areas, lowering resilience to recovery by native species 111
(Chambers et al. 2014), and lowering resistance to disturbance, making these systems more susceptible to dominance by invasive annual grasses (Davies et al. 2011; Sherrill and Romme 2012; Davies et al. 2016a). As the Great Basin becomes more arid (Snyder et al. 2019) and invasive annual grasses continue to spread into undisturbed plant communities (Chambers et al. 2007), fire-return interval regimes are altered (Balch et al.
2013), influencing plant communities to become less dominated by native species, unless targeted grazing to reduce fine fuels or other treatments, like herbicide, are applied (Schmelzer 2014; Perryman et al. 2018).
While the effects of moderate grazing and grazing exclusion on aboveground shrub and annual and perennial herbaceous forb and grass cover have been assessed
(Courtois et al. 2004; Davies et al. 2010; Davies et al. 2018; Davy and Rinella 2019), the effects of herbivory and long-term grazing exclusion on the seed banks of Great Basin plant communities are not as well-known. The viable seed bank in the Great Basin is typically dominated by native and non-native annual species (Martyn et al. 2016;
Germain et al. 2018), which may act as a source for aboveground plant recovery. Seed banks are highly variable and dependent on factors such as soil, elevation, climate, timing and intensity of grazing (Freas and Kemp 1983; Kemp 1989; Miller et al. 2011). In addition to the seed bank, Great Basin sagebrush systems are known to contain the highest percentage of annual native forb cover compared to other cold deserts
(Pennington et al. 2017). Relatively high contribution of annual forbs in above- and belowground communities suggests diversity (richness and evenness) can increase aboveground during wet years (Cuello et al. 2019), moderately decline aboveground 112 during drier years (Munson et al. 2013) and persist belowground during an extended drought (Jurado and Flores 2005) due to annual seed production (Kemp 1989) and dormancy adaptations (Freas and Kemp 1983).
Under certain targeted grazing conditions during the growing season, decreases in native perennial forb and grass cover (Davies et al. 2009; Pennington et al. 2017) can be high; however, controlled grazing may increase species diversity, richness and evenness in above- and belowground plant communities during highly productive years
(Osem et al. 2006; Zhao et al. 2011; Agra and Ne'eman 2012). However, differences in species richness between grazed and ungrazed areas in shrub-dominated systems are not always distinct (Barga and Leger 2018; Davies et al. 2018; Tessema et al. 2012).
Additional factors such as soil-moisture availability (Mitchell et al. 2017) and complete utilization of resources by aboveground shrub and deep-rooted perennial cover (Rigge et al. 2019) may drive above- and belowground diversity more than grazing pressures.
Areas with light to moderate grazing where decadent sagebrush cover occurs in low densities, exhibit the greatest aboveground diversity of native perennial species
(Wilkinson 1999; Courtois et al. 2004; Germain et al. 2018), especially in areas devoid of cheatgrass and where the native annual and perennial forb cover is high. These sites attract wildlife such as sage-grouse, who often prefer grazed over grazing-excluded habitat (Neel 1980; Miller et al. 2011; Dahlgren et al. 2015). Long-term grazing exclusion
(prior to 1936) increases decadent shrub and herbaceous litter cover (Davies et al.
2009), increasing the chances of wildfire destroying sagebrush habitat at risk of crossing 113 a threshold (Davies et al. 2016a) and becoming a non-native annual grassland (Davies et al. 2009; Pennington et al. 2017).
Soil texture may also play a potential role in areas with a history of heavy, uncontrolled grazing. Above- and belowground species composition can be highly dissimilar within the same ecological site gradient (Davies et al. 2007; Haight et al. 2019) if they contain differing soil textures. Finer soil textures with a higher clay content may facilitate an increase in cover of native perennial forb and grass species (Pennington et al. 2017; Haight et al. 2019), and contain less non-native invasive annuals like cheatgrass that would otherwise facilitate more similar above- and belowground communities.
Understanding diversity and similarity between above- and belowground sagebrush-dominated plant communities can provide insight into the ecological potentials of grazed and long-term grazing excluded areas in the northern Great Basin.
We assessed differences in above- and belowground diversity between grazed and ungrazed areas using soil seed bank and aboveground cover attributes in three of the original 28 Nevada Plots in Humboldt and Elko County, NV. This study addressed two questions: 1) How does aboveground cover and belowground seed densities of overall species differ between pre-Taylor Grazing Act era, long-term grazing exclusion and post-
Taylor Grazing Act grazed areas? 2) How do relative abundances (percent cover and seed density) of aboveground plant cover correlate to the germinable soil seed bank in grazed and long-term grazing excluded areas?
METHODS 114
STUDY AREA: This research was conducted in three grazing exclosures in Northern
Nevada originally established between 1936 and 1939 by the U.S. Department of the
Interior, Division of Grazing, known as the Nevada Plots. The first exclosure, Paradise
Valley #1 (41.372 N, 117.564 W) is a Wyoming big sagebrush and squirreltail plant community located in Paradise Valley, NV, north of Winnemucca in Humboldt Co., west of US-95 North. Paradise Valley #2 (41.562 N, 117.545 W) is a Wyoming big sagebrush/low sagebrush and sandberg bluegrass plant community located approximately 22.2 km north of Paradise Valley #1 on Hinkey Summit Road, and Dinner
Station (41.140 N, 115.849 W) is a Wyoming big sagebrush and basin wildrye plant community located in Elko Co., northwest of Elko, NV, east of Hwy 225 (Table 1). Grazing pressure at Paradise Valley #1 was unknown until 1986, and from 1986 to last known current use has been 624 cattle Animal Unit Monthly’s (AUMs) grazing in winter/spring
(Courtois et al. 2004). Paradise Valley #2 has had spring cattle since 1919, and the last known current use was spring cattle at low stocking rates (Courtois et al. 2004). Dinner
Station had summer cattle before 1950 with a 30% reduction in use, concurrent with spring sheep before 1965. From 1966-1986, cattle grazed the area in spring/fall, and from 1987 to last known current use has been spring through fall cattle use and spring/early summer use by sheep (Courtois et al. 2004). The grazing enclosures were selected based on sagebrush plant community types characterized as important sage- grouse habitat (Klebenow 1969; Fischer et al. 1996; Lowe et al. 2009; Hagen et al. 2011;
Schroff et al. 2018), and the exclosures had not been compromised by structural decay at the time of study. 115
EXPERIMENTAL DESIGN: Each of the three sites were treated as a separate, paired assessment for point-in-time comparisons between multiple paired transects across a large non-replicable sampling area (Schmelzer 2014; Davies and Gray 2015). Each site contained a 3-strand, barbwire fenced-in exclosure, 1.6 ha (100 x 160-m) in size, designated as the ungrazed treatment. An adjacent unfenced sample plot, 1.6 ha (100 x
160 m) in size, with similar aspect, slope, and ecological site characteristics was designated as the grazed treatment. A systematic sampling method in each plot was used by randomly placing four 50-m long parallel transects spaced 20-m apart, perpendicular to the slope (2 to 10 percent). Edge effects from fire lines, roads, and cattle use were excluded by placing transects 50-m away from the immediate area surrounding the exclosure (Porensky and Young 2016).
SEEDBANK AND VEGETATION SAMPLING: Field sampling took place following the seed production of most annual forbs and grasses (June to July 2018). Percent canopy cover and species composition for herbaceous plants (annual grass, forbs) were collected using Daubenmire (Daubenmire 1959) with 0.1-m2 (20-cm x 50-cm) quadrats at 10-m intervals along each 50-m transect. Percent basal cover and species composition for shrubs and perennial grass were collected using line-intercept (Canfield 1941) along each 50-m transect. Seed bank assay samples were collected within the center of each
Daubenmire quadrat from the top 5.1-cm of soil (196.2 cm3) using a 7-cm diameter
Ames hand bulb tulip planter (Barga and Leger 2018) for a total of five samples at each 116 transect, 20 samples per plot (ungrazed and grazed), totaling to 40 samples at each site
(Schmelzer 2014). For each site, soil texture class from the top 0 to 30-cm was assessed from a 43.2-cm x 30.5-cm soil pit in grazed and ungrazed areas using the soil texture flow diagram in the field (Thien 1979), which was referenced with ecological site descriptions (USDA-NRCS, 2006).
SEEDBANK GREENHOUSE ASSESSMENT: Seed bank assay samples were processed in the greenhouse at the University of Nevada-Reno beginning November 2018 (Espeland et al.
2010; Barga and Leger 2018). One-hundred and twenty 16.5 cm (L) x 10.2 cm (W) x 5.1 cm (H) mini seed garden trays (Barga and Leger 2018) were filled with a 1-cm layer of vermiculite at the base and overlaid with black landscape fabric (18-cm x 12-cm).
Sterilized sand was added in a 1:2 ratio (200 mL) to each of the 140 field soil samples
(400 mL), mixed homogeneously, and transferred to the mini seed garden trays (Barga and Leger 2018). Each greenhouse table was covered with a blue medium duty tarp (1.8- m x 2.4-m) that was overlaid by a layer of polyester quilt batting (10-m x 10-m)
(Espeland et al. 2010; Barga and Leger 2018). The mini trays that contained soil samples were placed at random on four 10-m x 10-m greenhouse tables (35 trays per table).
The seed bank assay had four emergence cycles (November 2018 to July 2019) and each cycle lasted until zero new seedlings emerged over the course of seven days
(Barga and Leger 2018). Between each cycle, plants were counted, identified, and discarded, and the soil samples were mixed in the trays to trigger germination of remaining seeds before the subsequent cycle began. Prior to the start of each cycle, the 117 trays were randomly rotated to a different table that helped encourage uniform greenhouse conditions across trays. Based on plant response to ambient greenhouse temperatures (15 °C and 18 °C), watering frequency and length of time was adjusted for each cycle (Ball and Miller 1989; Espeland et al. 2010; Pekas and Schupp 2013; Barga and Leger 2018).
For the first 60-day cycle (12 November 2018 to 11 January 2019), trays were watered two times per day using aerial misters (delivery rate 49-L/hour), five minutes per session, four times per week. The second cycle was 46 days long (2 February to 3
March 2019) and watering sessions were increased to three times per day, and remained at five minutes per session, four times per week. The third cycle was a 31-day dry period (5 May to 5 June 2019), followed by a final cycle that was 39-days long (6
June to 16 July 2019). Three milliliters of gibberellic acid (GA) (250 mg/L) was applied to each tray with a 3-mL pipette at the beginning of cycle four, followed by regular watering sessions that were seven days per week, two times per day, and 15 minutes per session.
STATISTICAL ANALYSIS: All statistical analyses were conducted using Program R and Excel
(Team 2018). Paired t-tests were applied using the PairedData package (Champely and
Champely 2018) to analyze differences in plant cover and seed density between grazed and ungrazed areas (Coulloudon et al. 1999). The vegan package (Oksanen et al. 2013) assessed correlations between relative aboveground plant cover (aboveground) and 118 relative belowground plant density in grazed and ungrazed areas. The package, ggplot2
(Venables and Ripley 2002; Wickham 2016) was used to produce figures.
Means and standard errors were calculated for aboveground cover (canopy and basal cover), and for belowground seed densities across each transect at each site, and reported as means and standard errors in tables by treatment (grazed and ungrazed) per lifeform (forb, grass, or shrub) and by known status (annual, perennial, native or non-native) (USDA NRCS, 2019). The Bray-Curtis dissimilarity index (Gardener 2014) was calculated for each site by treatment to determine how similar/dissimilar the belowground seed bank correlated to aboveground plant cover in grazed and ungrazed areas. Bray-Curtis index values were calculated with the combined values of aboveground relative canopy and basal cover and belowground relative plant density from the log-transformed abundance values (log(% + 1)), which increased the contribution of less abundant species (Gardener 2014; Greenacre 2017). Relativizing abundance values standardized and constrained abundance (0 and 1) at different scales which allowed comparisons between aboveground plant cover and belowground plant density (Gardener 2014).
A chi-squared tests assessed plant species preference for above- and belowground communities across sites, using combined (canopy and basal) percent cover, and belowground seed density (plant counts). Pearson residuals ()*+) with values
> ±2 were considered significant and had a normal distribution (De Cáceres et al. 2010;
Gardener 2014). 119
Finally, in order to assess diversity for above- and belowground plant communities, the number of species (species richness) and richness and evenness
(diversity) assessments were performed. For the aboveground plant community, percent canopy and basal cover were combined as abundance (evenness) estimates
(Gardener 2014). For the belowground plant community, plant density from the seed bank assay were calculated as abundance (evenness) estimates. A Shannon-Weaver
(Shannon and Weaver 1949) index value (H’) was calculated for all plant and all forb species in each treatment at each site. To analyze each paired (grazed and ungrazed) index values (H’), a modified t-test (Hutcheson 1970; Heip and Engels 1974; Gardener
2014) was applied. From Shannon-Weaver index values, Effective species was calculated, which provided comparisons of diversity that were more apparent than if using the raw index (H’) itself (Jost 2006).
Assumptions for parametric analyses were met by log-transforming data. The transformed dataset was tested for normality using Shapiro-Wilk (p > 0.05). When grazed area results were significantly different from the ungrazed area, the non- parametric correlation test (Spearman’s (ρ) > ± 0.75) was used on mean aboveground cover and seed density groups (Garcia 2011). Results are reported as means by treatment with standard errors using untransformed data, and all differences were determined at P = 0.05.
RESULTS 120
ABOVEGROUND AND SEED BANK SPECIES COMPOSITION: In total, we identified 31 species of
19 genera in 13 families for the above-ground community (Figure 1), and 23 species of
14 genera in 9 families in the seed bank (Figure 2). Across all sites, aboveground cover and the seedbank shared only three native species (all perennials) and three non-native species (all annuals) (Figure 3). Of the native species shared between above- and belowground plant communities, two were grasses and one was a shrub. Of the non- native species shared, two were grasses and one was a forb. One perennial grass, labeled PG1 (< 3% total density), could not be identified to genus or species in the seed bank assay but was included in the BC similarity and indicator species assessments only.
Percent composition by functional group and species were recorded for aboveground cover (Table 2) and the belowground seed bank (Table 3).
ABOVEGROUND COVER: All aboveground results are reported in Table 4 as mean percent cover by treatment (ungrazed and grazed). Dinner Station and Paradise Valley 2 were the only sites with forbs (5—6 per site) consumed by Greater sage-grouse (GRSG)
(Dumroese et al. 2015; Dumroese et al. 2016) in the aboveground cover (Figure 4a).
Sage-grouse forbs and perennial grass were correlated (ρ > ± 0.75) and different between grazed and the ungrazed exclosure at Paradise Valley 2 (Table 4). Sage-grouse forb cover was greater inside the ungrazed exclosure (Table 4g) while perennial grass was greater in the grazed area (Table 4i) at Paradise Valley #2. Although not different, all other forb cover groups were lower in the grazed area than the ungrazed exclosure at the same site (Paradise Valley 2). Cheatgrass cover did not differ significantly between 121 any grazed area and ungrazed exclosures but was 5% lower in the grazed area at
Paradise Valley #2 and 1.3% higher in the grazed area at Paradise Valley #1 (Table 4h).
Dinner Station was the only site where no annual grass cover was recorded for either treatment (Figure 5a). The lowest elevation site (Paradise Valley #1) was the only site where no perennial or native forb cover was recorded in grazed or ungrazed areas, and all annual forb cover was non-native and 1% greater in the grazed area (Table 4e) than in the ungrazed exclosure. Although not different, Dinner Station had 2% more native forb cover (annual and perennial) in the grazed area (Table 4d), than in the ungrazed exclosure. Also not different, shrub cover was up to 1.3% lower across all sites in grazed areas than in the ungrazed exclosures.
SEED BANK ASSAY DENSITIES: Four of ten plant groups in the seed bank assay were different between paired grazed and ungrazed areas at two of three sites (Figures 4b,
5b). Forb, annual forb, and perennial grass at both Paradise Valley Sites were highly correlated (ρ > ± 0.75) and different between grazed areas and in ungrazed exclosures
(Table 5). All forbs were annual at Paradise Valley #1, while Paradise Valley #2 was the only site that germinated perennial forbs. Both Paradise Valley #1 and Paradise Valley
#2 sites had greater seed density (seeds•m-2) in grazed areas for forbs (Table 5a), annual forbs (Table 5c), non-native forbs (Table 5e), and perennial grass (Table 5i), respectively
(Figure 4b). In the grazed area, greater densities of non-native forbs at Paradise Valley
#2 were correlated (ρ > ± 0.75) with lower densities of sage-grouse forbs, than in the ungrazed exclosure. Cheatgrass seed densities were lower in grazed areas at Paradise 122
Valley #2 and Dinner Station (Table 5h), and higher at Paradise Valley #1 (Table 5h) compared to the ungrazed exclosures (Figure 5b). The lowest elevation site (Paradise
Valley #1) was the only site where no shrubs germinated from the seed bank assay for either treatment. Although not different, shrub seed densities were lower in grazed areas at Paradise Valley #2 and Dinner Station (Table 5j). In general, Dinner Station was the only site that had lower seed densities across all plant groups in grazed areas than in the ungrazed exclosure (Figures 4b, 5b).
ABOVEGROUND COVER AND SEED BANK SIMILARITY: The Bray-Curtis (BC) similarity (1—BC) index showed that the seed bank assay and aboveground cover at Paradise Valley #2 were similar in both the ungrazed exclosure (41%) and in the grazed area (40%).
Meanwhile, Dinner Station was the least similar in both the ungrazed exclosure (7%) and in the grazed area (5%). At Paradise Valley #1, above- and belowground communities were moderately similar in the ungrazed exclosure (26%) and in the grazed area (26%).
INDICATOR SPECIES ASSOCIATION ANALYSIS: For aboveground affinities, seven species had a significant site preference (Figure 6). Species preference was based on its abundance at a site and given an importance value ()*+ = 2). Species either had a positive association and were more abundant at a site ()*+ > 2) or had a negative association and were less abundant or avoided the site altogether ()*+ > -2). 123
Three of the seven species indicators aboveground were shrubs (all native sagebrush), three species were forbs (two perennial, one annual, all were native), and one species was an annual non-native grass (Figure 6). All sites had positive associations with at least one sagebrush species, and Dinner Station was the only site with a positive affinity for rabbitbrush (Ericameria nauseosa [Pall. ex Pursh] G.L. Nesom & Baird). In addition, Dinner Station had the strongest positive association with the mat-forming forb, cushion phlox (Phlox hoodii Richardson), while Paradise Valley #1 had the strongest association with Wyoming big sagebrush (A. tridentata [Nutt.] ssp. wyomingensis Beetle
& Young) amid all other species that exhibited a negative preference for the site. The greatest number of species preferred the Paradise Valley #2 site and associated with one perennial forb, tapertip onion (Allium acuminatum Hook.), an annual forb, rough eyelashweed (Blepharipappus scaber Hook.), Wyoming big sagebrush, and a non-native annual grass, six-weeks fescue (Vulpia bromoides [L.] Gray).
In the seed bank assay, seven species (two native annual forbs, two non-native annual forbs, two non-native annual grasses, and one native perennial grass) indicated a significant site preference (Figure 7). Native annual forb, blue-eyed Mary (Collinsia parvifolia Lindl.) was positively associated in the seed bank with Dinner Station, while the other native annual forb, bristly mousetail (Myosurus apetalus [C.] Gay) was positively associated with Paradise Valley #2. The lower elevation sites (Paradise Valley
#1 and Paradise Valley #2) were positively associated with the invasive annual grass, cheatgrass (Bromus tectorum L.). However, Paradise Valley #1 was also positively associated with a perennial grass (unknown genus) while Paradise Valley #2 indicated a 124 positive affinity for the non-native annual grass, six-weeks fescue. Most of the forbs (3) had a positive affinity for the seed bank at Paradise Valley #2, but two were non-native annual forbs, spring draba (Draba verna L.), and jagged chickweed (Holosteum umbellatum L.).
ABOVEGROUND AND SEED BANK ASSAY DIVERSITY: The overall site diversity between ungrazed and grazed areas was only different in the seed bank at Paradise Valley #2, and not different in aboveground cover (Table 6). In aboveground cover, richness was higher in ungrazed exclosures than grazed areas at Paradise Valley 2 (S = 16) and Dinner
Station (S = 17), but the grazed areas were more even (J’ = 0.8; J’ = 0.9). Hence, diversity
(effective species) was greater in the grazed area at Paradise Valley #2 (H’ = 2.1, P =
0.06), and greater in the ungrazed exclosure at Dinner Station (H’ = 2.1, P = 0.38) (Figure
8a). Although also not different, Paradise Valley 2 was richer (S = 8) and more even (J’ =
0.8) in the grazed area than in the ungrazed exclosure (H’ = 1.6, P = 0.67).
Despite belowground species richness being the same (S = 12) between treatments at Paradise Valley #2, the grazed area was more even (H’ = 1.9, P = 0.02) than in the ungrazed exclosure (H’ = 1.8, P = 0.02) (Figure 8b). Subsequently, effective species richness was higher in the grazed area (Effective S = 6.7) than in the ungrazed exclosure (Effective S = 6.1). Although not different, richness and evenness were greater belowground in the ungrazed exclosure at Dinner Station than in the grazed area, while the grazed area at Paradise Valley 1 was richer than in the ungrazed exclosure.
125
DISCUSSION
ABOVE- AND BELOWGROUND DIVERSITY: In general, there were no differences in richness and evenness between grazed and ungrazed areas in both above- and belowground communities across sites except in the belowground community at Paradise Valley 2.
Similar to Courtois et al. (2004), no difference in aboveground cover diversity (H’) was generally found in grazed or ungrazed areas. However, unlike in our study that identified no difference aboveground at Paradise Valley #2, Courtois et al. (2004) indicated a more diverse aboveground plant community with greater species richness in the ungrazed exclosure. The densities of the grazed seed bank at Paradise Valley #2 contained a relatively even combination of native and non-native annual forbs (C. testiculata, D. verna, H. umbellatum, M. apetalus) and non-native annual grasses (B. tectorum, V. bromoides), compared to the ungrazed exclosure, where seed densities of non-native annual grasses composed 63% of total species composition. Since aboveground recruitment of annual forb and grasses depend on the availability of soil-moisture each year (Mitchell et al. 2017), long-term diversity (H’) in the seed bank at Paradise Valley #2 could actually be less different. Although minute, the slight to no changes in diversity
(H’) between grazed and ungrazed areas in both above- and belowground communities at our sites suggests that species composition in most plant communities vary year to year across the landscape and appear to fall within a natural range of variation based on the amount and timing of precipitation (Humphrey and Schupp 1999; Courtois et al.
2004; Boyte et al. 2016), suggesting a longer-term assessment of diversity is needed to define normal variation. 126
Generally few differences in richness (S) and evenness (J’) in above- and belowground, grazed and grazing excluded areas indicates that diversity (H’) since pre-
Taylor Grazing Act conditions are similar at 65-years (Courtois et al. 2004) and at 82- years since exclosures have been in place. In other Wyoming big sagebrush communities in the Great Basin, where grazing was excluded for six years, no difference in aboveground diversity (H’) was observed between grazed and ungrazed areas (Davies et al. 2018), suggesting other factors besides grazing pressure, like soil-moisture availability (Mitchell et al. 2017) and aboveground resource acquisition by shrub and deep-rooted perennials (Rigge et al. 2019) are driving any change in richness (S) and evenness (J’) in above- and belowground, grazed and ungrazed communities.
ABOVEGROUND AND BELOWGROUND SIMILARITY: In general, aboveground cover and belowground seed densities were similar between ungrazed exclosures and adjacent grazed areas. In addition, similarity comparisons between above- and belowground communities indicated no difference between ungrazed exclosures and adjacent grazed areas which may be driven by low numbers of shared species between above- and belowground areas, an overrepresentation of a few annual (native and non-native) forb and grass species and an underrepresentation of big sagebrush and native perennial forbs and grasses in the seed bank (Martyn et al. 2016). Courtois et al. (2004) indicated that species diversity in aboveground cover were similar between grazed areas and areas that had excluded grazing for 65 years, supporting our conclusions that diversity has not changed 85 years since grazing exclusion in the same areas. Time since 127 disturbance can be a mechanism that drives community composition, where above- and belowground plant community similarity can increase as time since disturbance increases (Hopfensperger 2007); however, the type of disturbance, like grazing, and plant community type should be considered. In Osem et al. (2006), similarity increased between grazing and grazing excluded areas (26 years) based on the idea that lower plant productivity at a site was correlated to greater availability of soil nutrients; however, productivity is a function of seasonal weather patterns and soil characteristics, and additional factors (Pennington et al. 2017; Haight et al. 2019; Snyder et al. 2019), which may be linked to more similar above- and belowground communities (Kinloch and
Friedel 2005), but are not known to be directly correlated. Although we did not measure productivity or litter cover, but observed considerable amounts of standing dead cover and litter at our sites (Appendix B), standing aboveground herbaceous cover is often greater in ungrazed than grazed areas (Davies et al. 2010; Davies et al. 2016b; Davies et al. 2018), since failure to remove overgrown, dried herbaceous cover can facilitate an increase in a fine fuel litter layer (Davies et al. 2010) and enable germination of cheatgrass seeds (Perryman et al. 2018) and other non-native annuals (V. bromoides, C. testiculata). Subsequently, higher non-native annual herbaceous cover can increase above- and belowground similarity (BC) and lower species evenness in both grazed and grazing excluded areas over time. High relative similarity index values and low richness and evenness above- and belowground in ungrazed exclosures can be indicative of an accumulation of non-native herbaceous species which can lead to greater risk of fire- induced mortality of native perennial vegetation in ungrazed compared to grazed areas 128
(Davies et al. 2010). Targeted grazing (Davies et al. 2016b; Perryman et al. 2018) followed by mechanical reduction of decadent sagebrush can help reduce high-intensity wildfire risk (Davies et al. 2018) and support native species recovery through seeding or additional methods that facilitate more diverse above- and belowground communities
(Seefeldt and McCoy 2003; Barga and Leger 2018).
In ungrazed areas that appear to have little to no cheatgrass cover aboveground, like at Dinner Station, the persistent presence of cheatgrass observed in patches near the fence line of the exclosure and along disturbance trails can be indicative of its growing belowground presence in the ungrazed exclosure, suggesting that grazing exclusion for 82 years has not prevented cheatgrass invasion in the seed bank. The presence of cheatgrass in the seed bank is an important factor to consider when assessing the quality of intact sagebrush habitat especially since accumulation of seeds by cheatgrass may be the only sign when evaluating the risk of non-native species invasion and changes in plant diversity. Quality sagebrush habitat may be at risk of conversion before aboveground indicators tell us otherwise, especially if wildfire burns through the area during a long-term drought (Benson et al. 2003; Svejcar et al. 2017;
Snyder et al. 2019) and the perennial seed bank is depleted, and adjacent native perennial plants fail to produce viable seeds (Miller et al. 2011; Munson et al. 2013) for community recovery. In addition, annual and perennial sage-grouse forb cover can be higher in ungrazed than in grazed areas, like in Paradise Valley #2; suggesting grazing exclusion near and around sage-grouse nesting areas during spring season can be beneficial for the increased abundance of food sources (Crawford et al. 2004; Dahlgren 129 et al. 2015). However, these same sage-grouse food sources in the seed bank can be at risk of burning off, especially under dense decadent sagebrush cover and herbaceous litter (Davies et al. 2010), where once burned, are unable to take advantage of elevated soil-nutrients post-disturbance (Blank et al. 2017).
Indicator species were identified by site and were unable to be identified by treatment in above- and belowground communities without at least an additional year of data; however, the number of non-native annual grass or forb species indicating a positive affinity for a site can provide early warning indicators that can anticipate ecosystem transformations (Kachergis et al. 2012). In addition, stronger affinities of native species in aboveground cover like rabbitbrush and cushion phlox, known to increase with disturbance, can be indicative of sites with low ecological potential
(Chambers et al. 2014). A longer-term assessment is needed in order to identify indicator species that persist or are transient as they relate to long-term grazing exclusion and variation in grazing pressure in response to environmental variation affecting plant productivity year to year.
IMPLICATIONS
Long-term grazing by cattle influence the structure of aboveground cover by readily consuming perennial bunchgrasses and generally avoiding sagebrush, while grazing exclusion has increased fine fuel litter layers, the amount of decadent sagebrush, and has protected the soil from compaction (Davies et al. 2018; Perryman et al. 2018).
Although one site indicated a more even belowground community in grazed areas, 130 grazing has not conclusively increased species richness or evenness in above- and belowground communities compared to grazing excluded areas. If differences in richness and evenness are detectable, one year of data is not enough to indicate grazing and grazing exclusion has no discernible impact on diversity. In addition, results were very site specific and difficult to generalize as a whole. Generally, moderate grazing can decrease the probability of fire-induced mortality of native perennial forbs and grasses by reducing the amount of dry fuel accumulating in an area, suggesting grazing will protect existing diversity above- and belowground. In turn, resistance and resilience of a site increases. With cheatgrass present in the seed bank in grazed and ungrazed exclosures across all sites, grazing will decrease the risk of dominance by non-native annual grasses after disturbance. In addition, preventative wildfire treatments and rehabilitation methods addressing heavy grazing over-use, should be timed with respect to yearly precipitation for native perennial forb and grass recruitment, especially since viable seeds are produced during above-average soil-moisture availability. This approach should be more effective in meeting sage-grouse habitat diversity and aboveground cover goals, allowing valuable and limited resources to be maximized.
Nevertheless, our data suggest the effects of moderate grazing on sagebrush dominated habitat contribute in some ways to protecting native plant diversity by increasing evenness in above- and belowground communities, contributing to the persistence of shrub-dominated systems with a mixture of perennial herbaceous and native forb species.
131
LIMITATIONS
We document diversity relationships between above- and belowground communities in grazed areas and adjacent ungrazed exclosures across three ecological sites similar in topography, vegetation, and soils. Although exclosures do not represent the "original" condition prior to the introduction of domestic livestock, these exclosures do serve as indicators of change in vegetation and diversity from pre-Taylor grazing over-use, indicating what the above- and belowground communities would look like without grazing pressure. In addition, many of the original exclosures no longer exist or have been compromised and fallen to disrepair, explaining why only three sites were selected. If possible, additional existing sites and additional years of above- and belowground species abundance data, could help identify recruitment patterns in native annual and perennial herbaceous species between grazed and ungrazed areas across a larger spatial and ecological site gradient. Although all sites were identified as Loamy 8-
10 ecological sites, Dinner Station generally receives more annual precipitation than the other two exclosures (PRISM, 2004). Paradise Valley #2 was also closer to the foothills, and contained more species than Paradise Valley #1, suggesting similar ecological sites have different potentials based on greater availability of soil-moisture closer to the mountain, or land-use history. There are also limitations associated with using modeled climate data, instead of data from nearby weather stations; however, the nearest weather stations were not able to capture the inherent variation within sites. The timing of the application of gibberellic acid (GA), duration of the rotation of tray cycles, and lack of cold nights (<15°C) may have affected germination of perennial species. 132
Temperature is critical for germination timing of forb species (Boyd and Lemos 2013), suggesting that the use of growth chambers for studies in the future may promote the number and frequency of species germinating from the seed bank.
133
REFERENCES
Agra, H., and G. Ne'eman. 2012. Composition and diversity of herbaceous patches in
woody vegetation: The effects of grazing, soil seed bank, patch spatial properties
and scale. Flora. 207:310-317.
Balch, J. K., B. A. Bradley, C. M. D'Antonio, and J. Gomez-Dans. 2013. Introduced annual
grass increases regional fire activity across the arid western USA (1980-2009).
Global Change Biology. 19:173-183.
Ball, D. A., and S. D. Miller. 1989. A comparison of techniques for estimation of arable
soil seedbanks and their relationship to weed flora. Weed Research. 29:365-373.
Barga, S., and E. A. Leger. 2018. Shrub cover and fire history predict seed bank
composition in Great Basin shrublands. Journal of Arid Environments. 154:40-50.
Bates, J. D., and K. W. Davies. 2014. Cattle grazing and vegetation succession on burned
sagebrush steppe. Rangeland Ecology & Management. 67:412-422.
Benson, L., B. Linsley, J. Smoot, S. Mensing, S. Lund, S. Stine, and A. Sarna-Wojcicki.
2003. Influence of the Pacific Decadal Oscillation on the climate of the Sierra
Nevada, California and Nevada. Quaternary Research. 59:151-159.
Blank, R. R., C. Clements, T. Morgan, and D. Harmon. 2017. Sagebrush wildfire effects on
surface soil nutrient availability: A temporal and spatial study. Soil Science
Society of America Journal. 81:1203-1210.
Boyd, C. S., and J. A. Lemos. 2013. Freezing stress influences emergence of germinated
perennial grass seeds. Rangeland Ecology & Management. 66:136-142. 134
Boyte, S. P., B. K. Wylie, and D. J. Major. 2016. Cheatgrass percent cover change:
Comparing recent estimates to climate change−driven predictions in the
Northern Great Basin. Rangeland Ecology & Management. 69:265-279.
Canfield, R. H. 1941. Application of the line interception method in sampling range
vegetation. Journal of Forestry. 39:388-394.
Chambers, J. C., R. F. Miller, D. I. Board, D. A. Pyke, B. A. Roundy, J. B. Grace, E. W.
Schupp, and R. J. Tausch. 2014. Resilience and resistance of sagebrush
ecosystems: Implications for state and transition models and management
treatments. Rangeland Ecology & Management. 67:440-454.
Chambers, J. C., B. A. Roundy, R. R. Blank, S. E. Meyer, and A. Whittaker. 2007. What
makes Great Basin sagebrush ecosystems invasible by Bromus tectorum?
Ecological Monographs. 77:117-145.
Champely, S., and M. S. Champely. 2018. Package ‘PairedData’.
Coulloudon, B., K. Eshelman, J. Gianola, N. Habich, L. Hughes, C. Johnson, M. Pellant, P.
Podborny, A. Rasmussen, and B. Robles. 1999. Sampling vegetation attributes.
BLM Technical Reference:1734-1734.
Courtois, D. R., B. L. Perryman, and H. S. Hussein. 2004. Vegetation change after 65
years of grazing and grazing exclusion. Journal of Range Management. 57:574-
582.
Crawford, J. A., R. A. Olson, N. E. West, J. C. Mosley, M. A. Schroeder, T. D. Whitson, R. F.
Miller, M. A. Gregg, and C. S. Boyd. 2004. Ecology and management of sage-
grouse and sage-grouse habitat. Journal of Range Management. 57:2-19. 135
Cuello, W. S., J. R. Gremer, P. C. Trimmer, A. Sih, and S. J. Schreiber. 2019. Predicting
evolutionarily stable strategies from functional responses of Sonoran Desert
annuals to precipitation. Proceedings of the Royal Society B-Biological Sciences.
286.
Dahlgren, D. K., R. T. Larsen, R. Danvir, G. Wilson, E. T. Thacker, T. A. Black, D. E. Naugle,
J. W. Connelly, and T. A. Messmer. 2015. Greater sage-grouse and range
management: Insights from a 25-year case study in Utah and Wyoming.
Rangeland Ecology & Management. 68:375-382.
Daubenmire, R.F. 1959. Canopy coverage method of vegetation analysis. Northwest
Science. 33:43-64.
Davies, K. W., J. D. Bates, and C. S. Boyd. 2016a. Effects of intermediate-term grazing
rest on sagebrush communities with depleted understories: Evidence of a
threshold. Rangeland Ecology & Management. 69:173-178.
Davies, K. W., J. D. Bates, C. S. Boyd, and T. J. Svejcar. 2016b. Prefire grazing by cattle
increases postfire resistance to exotic annual grass (Bromus tectorum) invasion
and dominance for decades. Ecology and Evolution. 6:3356-3366.
Davies, K. W., J. D. Bates, and R. F. Miller. 2007. Environmental and vegetation
relationships of the Artemisia tridentata spp. wyomingensis alliance. Journal of
Arid Environments. 70:478-494.
Davies, K. W., J. D. Bates, T. J. Svejcar, and C. S. Boyd. 2010. Effects of long-term
livestock grazing on fuel characteristics in rangelands: An example from the
sagebrush steppe. Rangeland Ecology & Management. 63:662-669. 136
Davies, K. W., C. S. Boyd, and J. D. Bates. 2018. Eighty years of grazing by cattle modifies
sagebrush and bunchgrass structure. Rangeland Ecology & Management. 71:275-
280.
Davies, K. W., C. S. Boyd, J. L. Beck, J. D. Bates, T. J. Svejcar, and M. A. Gregg. 2011.
Saving the sagebrush sea: An ecosystem conservation plan for big sagebrush
plant communities. Biological Conservation. 144:2573-2584.
Davies, K. W., T. J. Svejcar, and J. D. Bates. 2009. Interaction of historical and
nonhistorical disturbances maintains native plant communities. Ecological
Applications. 19:1536-1545.
Davy, J. S., and M. J. Rinella. 2019. Targeted grazing for native forbs in annual
grasslands. Rangeland Ecology & Management. 72:501-504.
De Cáceres, M., P. Legendre, and M. Moretti. 2010. Improving indicator species analysis
by combining groups of sites. Oikos. 119:1674-1684.
Dumroese, R. K., T. Luna, J. R. Pinto, and T. D. Landis. 2016. Forbs: Foundation for
restoration of monarch butterflies, other pollinators, and greater sage-grouse in
the western United States. Natural Areas Journal. 36:499-511.
Dumroese, R. K., T. Luna, B. A. Richardson, F. F. Kilkenny, and J. B. Runyon. 2015.
Conserving and restoring habitat for greater sage-grouse and other sagebrush-
obligate wildlife: the crucial link of forbs and sagebrush diversity. Native Plants
Journal. 16:276-299. 137
Espeland, E. K., L. B. Perkins, and E. A. Leger. 2010. Comparison of Seed Bank Estimation
Techniques Using Six Weed Species in Two Soil Types. Rangeland Ecology &
Management. 63:243-247.
Fischer, R. A., K. P. Reese, and J. W. Connelly. 1996. An investigation on fire effects
within xeric sage grouse brood habitat. Journal of Range Management. 49:194-
198.
Freas, K. E., and P. R. Kemp. 1983. Some relationships between environmental reliability
and seed dormancy in desert annual plants. Journal of Ecology. 71:211-217.
Garcia, E. 2011. A tutorial on correlation coefficients.
Gardener, M. 2014. Community ecology: analytical methods using R and Excel. Pelagic
Publishing Ltd.
Germain, S. J., R. K. Mann, T. A. Monaco, and K. E. Veblen. 2018. Short-term
regeneration dynamics of Wyoming big sagebrush at two sites in northern Utah.
Western North American Naturalist. 78:7-16.
Greenacre, M. 2017. Ordination with any dissimilarity measure: A weighted Euclidean
solution. Ecology. 98:2293-2300.
Hagen, C. A., M. J. Willis, E. M. Glenn, and R. G. Anthony. 2011. Habitat selection by
greater sage-grouse during winter in southeastern Oregon. Western North
American Naturalist. 71:529-538.
Haight, J. D., S. C. Reed, and A. M. Faist. 2019. Seed bank community and soil texture
relationships in a cold desert. Journal of Arid Environments. 164:46-52. 138
Heip, C., and P. Engels. 1974. Comparing species diversity and evenness indices. Journal
of the Marine Biological Association of the United Kingdom. 54:559-563.
Hopfensperger, K. N. 2007. A review of similarity between seed bank and standing
vegetation across ecosystems. Oikos. 116:1438-1448.
Humphrey, L. D., and E. W. Schupp. 1999. Temporal patterns of seedling emergence and
early survival of Great Basin perennial plant species. Great Basin Naturalist.
59:35-49.
Hutcheson, K. 1970. A test for comparing diversities based on the Shannon formula.
Journal of Theoretical Biology. 29:151-154.
Jost, L. 2006. Entropy and diversity. Oikos 113:363-375.
Jurado, E., and J. Flores. 2005. Is seed dormancy under environmental control or bound
to plant traits? Journal of Vegetation Science. 16:559-564.
Kachergis, E., M. E. Fernandez-Gimenez, and M. E. Rocca. 2012. Differences in plant
species composition as evidence of alternate states in the sagebrush steppe.
Rangeland Ecology & Management. 65:486-497.
Kinloch, J. E., and M. H. Friedel. 2005. Soil seed reserves in arid grazing lands of central
Australia. Part 1: Seed bank and vegetation dynamics. Journal of Arid
Environments. 60:133-161.
Kemp, P. R. 1989. Seed banks and vegetation processes in deserts. In: M. A. Leck (ed.).
Ecology of Soil Seedbanks. p. 257 - 281.
Klebenow, D. A. 1969. Sage grouse nesting and brood habitat in Idaho. Journal of
Wildlife Management. 33:649-662. 139
Lowe, B. S., D. J. Delehanty, and J. W. Connelly. 2009. Greater sage-grouse Centrocercus
urophasianus use of threetip sagebrush relative to big sagebrush in south-central
Idaho. Wildlife Biology. 15:229-236.
Martyn, T. E., J. B. Bradford, D. R. Schlaepfer, I. C. Burke, and W. K. Lauenroth. 2016.
Seed bank and big sagebrush plant community composition in a range margin for
big sagebrush. Ecosphere 7.
Miller, R. F., S. T. Knick, D. A. Pyke, C. W. Meinke, S. E. Hanser, M. J. Wisdom, and A. L.
Hild. 2011. Characteristics of sagebrush habitats and limitations to long-term
conservation. Greater Sage-Grouse: Ecology and Conservation of a Landscape
Species and Its Habitats:145-184.
Mitchell, R. M., J. D. Bakker, J. B. Vincent, and G. M. Davies. 2017. Relative importance
of abiotic, biotic, and disturbance drivers of plant community structure in the
sagebrush steppe. Ecological Applications. 27:756-768.
Morris, L. R., and R. J. Rowe. 2014. Historical land use and altered habitats in the Great
Basin. Journal of Mammalogy. 95:1144-1156.
Munson, S. M., E. H. Muldavin, J. Belnap, D. P. Peters, J. P. Anderson, M. H. Reiser, K.
Gallo, A. Melgoza-Castillo, J. E. Herrick, and T. A. Christiansen. 2013. Regional
signatures of plant response to drought and elevated temperature across a
desert ecosystem. Ecology. 94:2030-2041.
Neel, L. A. 1980. Sage grouse response to grazing management in Nevada. Master’s
thesis, University of Nevada, Reno, NV, USA. 140
Oksanen, J., F. G. Blanchet, R. Kindt, P. Legendre, P. R. Minchin, R. O’hara, G. L. Simpson,
P. Solymos, M. H. H. Stevens, and H. Wagner. 2013. Package ‘vegan’. Community
ecology package, version 2:1-295.
Osem, Y., A. Perevolotsky, and J. Kigel. 2006. Similarity between seed bank and
vegetation in a semi-arid annual plant community: The role of productivity and
grazing. Journal of Vegetation Science. 17:29-36.
Pekas, K. M., and E. W. Schupp. 2013. Influence of aboveground vegetation on seed
bank composition and distribution in a Great Basin Desert sagebrush community.
Journal of Arid Environments. 88:113-120.
Pennington, V. E., K. A. Palmquist, J. B. Bradford, and W. K. Lauenroth. 2017. Climate
and soil texture influence patterns of forb species richness and composition in
big sagebrush plant communities across their spatial extent in the western US.
Plant Ecology. 218:957-970.
Perryman, B. L., B. W. Schultz, J. K. McAdoo, R. L. Alverts, J. C. Cervantes, S. Foster, G.
McCuin, and S. Swanson. 2018. Viewpoint: An alternative management paradigm
for plant communities affected by invasive annual grass in the Intermountain
West. Rangelands. 40:77-82.
Porensky, L. M., and T. P. Young. 2016. Development of edge effects around
experimental ecosystem hotspots is affected by hotspot density and matrix type.
Landscape ecology. 31:1663-1680.
PRISM Climate Group, Oregon State University, 2004. Available at.
http://prism.oregonstate.edu, Accessed date: 11 December 2019. 141
Rigge, M., C. Homer, B. Wylie, Y. X. Gu, H. Shi, G. Xian, D. K. Meyer, and B. Bunde. 2019.
Using remote sensing to quantify ecosystem site potential community structure
and deviation in the Great Basin, United States. Ecological Indicators. 96:516-
531.
Schmelzer, L. 2014. Case study: Reducing cheatgrass (Bromus tectorum L.) fuel loads
using fall cattle grazing. The Professional Animal Scientist. 30:270 - 278.
Schroff, S. R., K. A. Cutting, C. A. Carr, M. R. Frisina, L. B. McNew, and B. F. Sowell. 2018.
Characteristics of shrub morphology on nest site selection of Greater Sage-
Grouse (Centrocercus urophasianus) in high-elevation sagebrush habitat. Wilson
Journal of Ornithology. 130:730-738.
Seefeldt, S. S., and S. D. McCoy. 2003. Measuring plant diversity in the tall threetip
sagebrush steppe: Influence of previous grazing management practices.
Environmental Management. 32:234-245.
Shannon, C. E., and W. Weaver. 1949. The mathematical theory of communication.
Univ. Illinois Press. Urbana. 117 p.
Sherrill, K. R., and W. H. Romme. 2012. Spatial variation in postfire cheatgrass: Dinosaur
national monument, USA. Fire Ecology. 8:38-56.
Snyder, K. A., L. Evers, J. C. Chambers, J. Dunham, J. B. Bradford, and M. E. Loik. 2019.
Effects of changing climate on the hydrological cycle in cold desert ecosystems of
the Great Basin and Columbia Plateau. Rangeland Ecology & Management. 72:1-
12. 142
Svejcar, T., C. Boyd, K. Davies, E. Hamerlynck, and L. Svejcar. 2017. Challenges and
limitations to native species restoration in the Great Basin, USA. Plant Ecology.
218:81-94.
Team, R. C. 2018. R: A Language and Environment for Statistical Computing. Vienna,
Austria: R Foundation for Statistical Computing.
Tessema, Z. K., W. F. de Boer, R. M. T. Baars, and H. H. T. Prins. 2012. Influence of
grazing on soil seed banks determines the restoration potential of aboveground
vegetation in a semi-arid savanna of Ethiopia. Biotropica. 44:211-219.
Thien, S. J. 1979. A flow diagram for teaching texture-by-feel analysis. Journal of
Agronomic Education. 8.
United States Department of Agriculture (USDA), Natural Resources Conservation
Service (NRCS). 2019. The PLANTS Database (http://plants.usda.gov, 1 October
2019). National Plant Data Team, Greensboro, NC 27401-4901 USA
United States Department of Agriculture (USDA), Natural Resources Conservation
Service (NRCS). 2006. Land resource regions and major land resource areas of
the United States, the Caribbean, and the Pacific Basin. U.S. Department of
Agriculture Handbook. 296.
Venables, W. N., and B. D. Ripley. 2002. Random and mixed effects. In Modern applied
statistics with S (pp. 271-300). Springer, New York, NY.
Wickham, H. 2016. ggplot2: elegant graphics for data analysis. Springer.
Wilkinson, D. M. 1999. The disturbing history of intermediate disturbance. Oikos 84:145-
147. 143
Zhao, L. P., J. S. Su, G. L. Wu, and F. Gillet. 2011. Long-term effects of grazing exclusion
on aboveground and belowground plant species diversity in a steppe of the
Loess Plateau, China. Plant Ecology and Evolution. 144:313-320. 144
Table 1 The locations of three grazing exclosures, also known as the Nevada Plots, and precipitation zone (PZ), elevations, plant community characteristics, ecological sites, and ecological site description (ESD) associated within major land resource area (MLRA) 24 Humboldt Area and MLRA 25 Owyhee High Plateau (USDA-NRCS, 2006).
Site Lat (N) Long (W) PZ (cm) Elevation Plant Community Ecological Site MLRA/ESD (m)
Paradise 41.37171 -117.5643 20-25 1339 Wyoming big Loamy 8-10 024XY005NV Valley 1 sagebrush/squirreltail Paradise 41.56220 -117.5453 20-25 1472 Wyoming big sagebrush/low Loamy 8-10 024XY005NV Valley 2 sagebrush/sandberg bluegrass
Dinner 41.13792 -115.8489 20-25 1799 Wyoming big sagebrush/basin Loamy 8-10 025XY019NV Station wildrye
145
Table 2 Percent (%) composition by species (annual forbs, annual grasses, perennial forbs, shrubs, and perennial grasses) in aboveground cover in ungrazed and grazed areas, and difference (Grazed – Ungrazed). Species are grouped into life forms. Only species that composed at least 1% of the canopy for either ungrazed or grazed sites are listed individually; all remaining species are grouped together for simplicity. Site Genus / Species Common Name Ungrazed Grazed Difference (%) (%) (%) Paradise Valley 1 Annual forbs 2.9 8.0 5.1 Alyssum desertorum desert madwort 2.9 8.0 5.1 Annual grasses 18 25.3 7.3 Bromus tectorum cheatgrass 18 25.3 7.3 Shrubs 40.2 28.7 -11.5 Artemisia tridentata ssp. big sagebrush 39.2 26.7 -12.5 Ericameria nauseosa rubber rabbitbrush 0.0 1.2 1.2 Grayia spinosa spiny hopsage 1.0 0.8 -0.2 Perennial grasses 38.8 38.1 -0.7 Elymus elymoides squirreltail 26.5 17.7 -8.8 Poa secunda sandberg bluegrass 12.3 19.9 7.6 Other species 0.0 0.5 0.5 Paradise Valley 2 Annual forbs 24.0 8.4 -15.6 Blepharipappus scaber rough eyelashweed 16.3 5.2 -11.1 Brassica ssp. mustard 6.7 0.0 -6.7 Ceratocephala testiculata bur buttercup 1.0 2.8 1.8 Other species 0.0 0.4 0.4 Annual grasses 49.5 50.5 1.0 Bromus racemosus bald brome 2.0 1.5 -0.5 Bromus tectorum cheatgrass 34.7 25.2 -9.5 Taeniatherum caput-medusae medusahead 1.4 0.0 -1.4 Vulpia bromoides six-weeks fescue 11.4 23.8 12.4 Perennial forbs 10.2 17.6 7.4 Allium acuminatum tapertip onion 0.0 11.2 11.2 Balsamorhiza hookeri Hooker’s balsamroot 4.9 0.0 -4.9 Penstemon kingii King’s beardtongue 0.0 1.8 1.8 Phlox longifolia long-leaf phlox 0.8 4.6 3.8 Trifolium gymnocarpon hollyleaf clover 4.3 0.0 -4.3 Other species 0.2 0.0 -0.2 Shrubs 14.2 15.3 1.1 Artemisia arbuscula low sagebrush 12.3 10.0 -2.3 Artemisia ludoviciana white sagebrush 1.2 4.2 3.0 Artemisia tridentata ssp. big sagebrush 0.7 1.1 0.4 Perennial grasses 2.4 8.2 5.8 Poa secunda sandberg bluegrass 2.0 7.7 5.7 Other species 0.4 0.5 0.1 Dinner Station Annual forbs 3.6 3.3 -0.3 Epilobium brachycarpum annual willowherb 2.1 3.3 1.2 Other species 1.5 0.0 -1.5 Perennial forbs 37.0 42.9 5.9 Cymoptrus ibapensis Ibapah springparsley 1.0 0.0 -1.0 Mertensia oblongifolia oblongleaf bluebells 9.8 9.8 0.0 Penstemon kingii King’s beardtongue 2.6 2.8 0.2 Phlox hoodii cushion phlox 23.1 30.3 7.2 Other species 0.5 0.0 -0.5 Shrubs 45.3 36.1 -9.2 Artemisia tridentata ssp. big sagebrush 15.2 19.1 3.9 Chrysothamnus viscidiflorus yellow rabbitbrush 4.4 3.0 -1.4 Ericameria nauseosa rubber rabbitbrush 25.7 14.0 -11.7 Perennial grasses 14.1 17.7 3.6 Elymus elymoides squirreltail 1.2 2.8 1.6 Elymus lanceolatus thickspike wheatgrass 0.0 3.7 3.7 Leymus cinereus basin wildrye 2.9 0.0 -2.9 Leymus triticoides creeping wildrye 8.8 7.0 -1.8 Poa secunda sandberg bluegrass 1.2 4.2 3.0 146
Table 3 Percent (%) composition by species (annual forbs, annual grasses, perennial forbs, shrubs, and perennial grasses) for the seed bank in ungrazed (UG) and grazed (G) areas, and difference (Grazed – Ungrazed). Species are grouped into life forms. Only species that composed at least 1% of the seed bank for either ungrazed or grazed sites are listed individually; all remaining species are grouped together for simplicity. Site Genus / Species Common Name Seed bank Seed bank Difference Ungrazed (%) Grazed (%) (%) Paradise Valley 1 Annual forbs 17.4 37.5 20.1 Ceratocephala testiculata 1.6 24.8 23.2 Draba verna 3.8 0.6 -3.2 Gayophytum decipiens 1.6 0.6 -1.0 Microsteris gracillis 4.4 3.7 -0.7 Myosurus apetalus 4.9 5.3 0.4 Nemophila ssp. 1.1 0.0 -1.1 Other species 0.0 2.5 2.5 Annual grasses 67.2 50.2 -17.0 Bromus tectorum 57.4 48.0 -9.4 Vulpia bromoides 9.8 2.2 -7.6 Perennial grasses 3.2 0.9 -2.3 Elymus elymoides 2.7 0.3 -2.4 Other species 0.5 0.6 0.1 Uknown perennial grasses 12.0 11.5 -0.5 Paradise Valley 2 Annual forbs 32.5 63.5 31.0 Ceratocephala testiculata 7.8 16.4 8.6 Collinsia parvifolia 1.5 0.4 -1.1 Draba verna 5.6 13.7 8.1 Holosteum umbellatum 4.2 12.6 8.4 Microsteris gracillis 3.6 0.1 -3.5 Myosurus apetalus 9.8 20.3 10.5 Annual grasses 62.7 33.8 -28.9 Bromus tectorum 39.6 13.6 -26.0 Vulpia bromoides 23.1 20.2 -2.9 Perennial forbs 0.0 1.3 1.3 Lithophragma ssp. 0.0 1.3 1.3 Shrubs 3.5 0.5 -3.0 Artemisia tridentata ssp. 3.5 0.5 -3.0 Perennial grasses 1.1 0.8 -0.3 Poa secunda 1.1 0.4 -0.7 Other species 0.0 0.4 0.4 Unknown perennial grasses 0.2 0.0 -0.2 Dinner Station Annual forbs 88.8 90.9 2.1 Ceratocephala testiculata 2.9 2.2 -0.7 Collinsia parvifolia 63.7 71.6 7.9 Microsteris gracillis 21.6 16.4 -5.2 Other species 0.6 0.7 0.1 Annual grasses 5.3 2.2 -3.1 Bromus tectorum 3.5 0.0 -3.5 Vulpia bromoides 1.8 0.0 -1.8 Juncus bufonius 0.0 2.2 2.2 Perennial forbs 2.9 4.5 1.6 Cryptantha ssp. 0.0 3.0 3.0 Nemophila ssp. 2.3 1.5 -0.8 Other species 0.6 0.0 -0.6 Shrubs 2.3 2.2 -0.1 Artemisia tridentata ssp. 2.3 2.2 -0.1 Unknown perennial grasses 0.6 0.0 -0.6
147
Table 4 Aboveground cover of plant functional groups compared between paired ungrazed (UG) and grazed (G) areas for each site. All variables are reported by site as mean percent cover (%) for UG and G, standard error (SE) (±), and mean difference % cover (G – UG). Variable Site UG ± SE (%) G ± SE (%) Difference P(T<=t) (%) a. Forb Paradise Valley 1 0.5 ± 0.2 1.4 ± 0.4 0.9 0.278 Paradise Valley 2 5.9 ± 1.5 3.7 ± 0.9 -2.1 0.316 Dinner Station 2.6 ± 0.8 4.6 ± 0.9 2.0 0.300 b. Perennial Forb Paradise Valley 1 ------Paradise Valley 2 5.9 ± 0.5 4.2 ± 1.0 -1.7 0.263 Dinner Station+ 3.1 ± 0.8 5.4 ± 1.1 2.3 -- c. Annual Forb Paradise Valley 1 0.5 ± 0.2 1.4 ± 0.4 0.9 0.278 Paradise Valley 2 4.9 ± 3.2 2.0 ± 0.7 -2.9 0.903 Dinner Station+ 1.9 ± 0.8 3.5 ± 1.1 1.6 -- d. Native Forb Paradise Valley 1 ------Paradise Valley 2 6.9 ± 1.3 4.2 ± 1.0 -2.7 0.234 Dinner Station 2.7 ± 0.8 4.6 ± 0.9 1.9 0.324 e. Non-Native Forb Paradise Valley 1 0.5 ± 0.2 1.4 ± 0.4 0.9 0.278 Paradise Valley 2 3.7 ± 2.0 2.0 ± 0.7 -1.6 0.976 Dinner Station+ 0.3 ± 0.1 0.0 ± 0.0 -0.3 -- f. Native Annual Forb Paradise Valley 1 ------Paradise Valley 2+ 4.8 ± 4.8 0.6 ± 0.1 -4.8 -- Dinner Station 1.9 ± 0.8 2.8 ± 0.8 1.6 0.552 g. GRSG Forbs Paradise Valley 1 ------Paradise Valley 2a 7.1 ± 1.2 2.8 ± 0.3 -4.3 0.007** Dinner Station 3.0 ± 0.9 4.6 ± 0.9 1.6 0.416 h. Cheatgrass Paradise Valley 1+ 3.1 ± 0.6 4.4 ± 0.6 1.3 0.391 Paradise Valley 2 22.1 ± 6.3 17.1 ± 2.6 -5.0 0.615 Dinner Station ------i. Perennial Grass Paradise Valley 1 3.4 ± 0.4 2.8 ± 0.8 -0.5 0.463 Paradise Valley 2a 0.9 ± 0.2 3.9 ± 1.1 2.9 0.004** Dinner Station 1.0 ± 0.3 1.2 ± 0.3 0.1 0.443 j. Shrub Paradise Valley 1 4.5 ± 1.2 3.2 ± 0.8 -1.3 0.341 Paradise Valley 2 5.9 ± 0.6 4.9 ± 1.4 -0.9 0.339 Dinner Station 4.1 ± 0.6 3.2 ± 0.2 -0.9 0.337 + locations with variables that did not assume normality on log-transformed means. * Treatment has a significant effect on cover group within site, p < 0.05 ** Treatment has a significant effect on cover group within site, p < 0.01 a Significant variables containing the same letter are correlated within site (Spearman’s correlation coefficient > ± 0.75) Variables: GRSG = Greater sage-grouse
148
Table 5 Plant functional groups compared between ungrazed (UG) and grazed (G) areas for each site in the seed bank. All variables are reported as average number of seeds•m2 (density), standard error (SE) (±), and mean difference of average seeds•m2 (G – UG). Variable Site UG ± SE G ± SE Difference P(T<=t) (seeds•m2) (seeds•m2) (seeds•m2) a. Forb Paradise Valley 1c 111 ± 27 379 ± 99 268 0.016* Paradise Valley 2a 591 ± 100 1697 ± 446 1105 0.018* Dinner Station 721 ± 204 508 ± 120 -213 0.404 b. Perennial Forb Paradise Valley 1 0 ± 0 0 ± 0 0 ± 0 -- Paradise Valley 2+ 0 ± 0 169 ± 169 169 0.391 Dinner Station 0 ± 0 0 ± 0 0 ± 0 -- c. Annual Forb Paradise Valley 1c 111 ± 27 379 ± 99 268 0.016* Paradise Valley 2a 591 ± 100 1722 ± 435 1131 0.014* Dinner Station+ 721 ± 204 508 ± 120 -213 0.404 d. Native Forb Paradise Valley 1 111 ± 27 161 ± 40 50 0.350 Paradise Valley 2 450 ± 166 1524 ± 1034 1073 0.232 Dinner Station 794 ± 176 569 ± 91 -226 0.277 e. Non-Native Forb Paradise Valley 1 39 ± 25 905 ± 277 866 0.054* Paradise Valley 2b 859 ± 310 1817 ± 410 957 0.041* Dinner Station+ 78 ± 62 39 ± 25 -39 0.908 f. Native Annual Forb Paradise Valley 1 111 ± 27 161 ± 40 50 0.350 Paradise Valley 2 450 ± 166 1472 ± 1057 1021 0.642 Dinner Station 794 ± 176 569 ± 91 -226 0.277 g. GRSG Forbs Paradise Valley 1 104 ± 56 169 ± 54 65 0.326 Paradise Valley 2b 247 ± 77 65 ± 33 -182 0.207 Dinner Station 951 ± 256 768 ± 192 -182 0.592 h. Cheatgrass Paradise Valley 1 1367 ± 406 2018 ± 392 651 0.383 Paradise Valley 2 2839 ± 269 1706 ± 303 -1133 0.089 Dinner Station 78 ± 45 0 ± 0 -78 0.064 i. Perennial Grass Paradise Valley 1c 111 ± 27 379 ± 99 268 0.016* Paradise Valley 2a 591 ± 100 1697 ± 446 1105 0.018* Dinner Station+ 721 ± 204 508 ± 120 -213 0.404 j. Shrub Paradise Valley 1 0 ± 0 0 ± 0 0 ± 0 -- Paradise Valley 2a 247 ± 44 65 ± 49 -182 0.080 Dinner Station+ 52 ± 21 26 ± 15 -26 0.307 + locations with variables that did not assume normality on log-transformed means. * Treatment has a significant effect on cover group within site, p < 0.05 a, b, c Significant variables containing the same letter are correlated within site (Spearman’s correlation coefficient > ± 0.75) Variables: GRSG = Greater sage-grouse 149
Table 6 Sites, elevations (m), treatment (trt), Shannon-Weaver diversity index values (H’), Effective species (Effective S), richness (S), evenness (J’), Hutcheson’s t-statistic (t), and probability values (P) for ungrazed (UG) and grazed (G) plots in aboveground and belowground for overall site diversity. Aboveground Belowground Site Trt H’ Effective S J’ t P H’ Effective S J’ t P S S Paradise Valley 1 UG 1.4 4.2 6 0.8 1.9 0.67 1.5 4.6 11 0.6 0.2 0.85 G 1.6 5.2 8 0.8 1.5 4.7 14 0.6 Paradise Valley 2 UG 2.1 7.8 16 0.7 0.9 0.06 1.8 6.1 12 0.7 2.3 0.02* G 2.1 8.4 15 0.8 1.9 6.7 12 0.8 Dinner Station UG 2.1 8.1 17 0.7 0.4 0.38 1.2 3.2 10 0.5 1.2 0.24 G 2.0 7.7 11 0.9 1.0 2.7 9 0.5
*p < .05 150
Figure 1 Mean relative percent (%) aboveground cover of all species recorded across all six sites. Proportion of species identified as forb, grass, or shrub are indicated by shaded bar.
Plant codes: AGCR = Agropyron cristatum (crested wheatgrass), ALAC4 = Allium acuminatum (tapertip onion), ALDE = Allysum desertorum (desert madwort), ARARA = Artemisia arbuscula (low sagebrush), ARLU = Artemisia ludoviciana (white sagebrush), ARTRW8 = Artemisia tridentata ssp. wyomingensis (Wyoming big sagebrush), BAHO = Balsamorhiza hookeri (Hooker’s balsamroot), BLSC = Blepharipappus scaber (rough eyelashweed), BRASS2 = Brassica ssp. (mustard), BRRA2 = Bromus racemosus (bald brome), BRTE = Bromus tectorum (cheatgrass), CETE5 = Ceratocephala testiculata (bur buttercup), CHVI8 = Chrysothamnus viscidiflorus (yellow rabbitbrush), CYIB = Cymopterus ibapensis (Ibapah springparsely), DEPAN = Descurainia paradisa (paradise tansymustard), ELEL5 = Elymus elymoides (squirreltail), ELLA3 = Elymus lanceolatus (thickspike wheatgrass), EPBR3 = Epilobium brachycarpum (annual willowherb), ERNA10 = Ericameria nauseosa (rubber rabbitbrush), GRSP = Grayia spinosa (spiny hopsage), LECI4 = Leymus cinerius (basin wildrye), LETR5 = Leymus triticoides (creeping wildrye), MEOB = Mertensia oblongifolia (oblong bluebells), PEKI = Penstemon kingii (King’s lambstongue), PHHO = Phlox hoodii (cushion phlox), PHLO2 = Phlox longifolia (longleaf phlox), POSE = Poa secunda (sandberg bluegrass), TACA8 = Taeniatherum caput-medusae (medusahead), TRGY = Trifolium gymnocarpon (hollyleaf clover), VUBR = Vulpia bromoides (six-weeks fescue) 151
Figure 2 Mean relative density (individuals/m2) of all species recorded across all sites in the seed bank. Proportion of species identified as forb, grass, or shrub are indicated by shaded bar.
Plant code: ARTEM = Artemisia ssp. (sagebrush), ARTRW8 = Artemisia tridentata ssp. wyomingensis (Wyoming big sagebrush), BRTE = Bromus tectorum (cheatgrass), CETE5 = Ceratocephala testiculata (bur buttercup), COPA3 = Collinsia parvifolia (blue-eyed Mary), CRYPT = Cryptantha ssp. (fiddleneck), DEPI = Descuraina pinnata (paradise tansymustard), DRVE2 = Draba verna (spring draba), ELEL5 = Elymus elymoides (squirreltail), GADE2 = Gayophytum decipiens (deceptive groundsmoke), GILIA = Gilia ssp. (gilia), HOUM = Holosteum umbellatum (jagged chickweed), JUBU = Juncus bufonius (toad rush), LITHO2 = Lithophragma ssp. (woodland star), LOMAT = Lomatium ssp. (desert parsley), MIGR = Microsteris gracillis (slender phlox), MYAP = Myosurus apetalus (bristly mousetail), NEMOP = Nemophila ssp. (baby blue eyes), PG1 = perennial grass unidentified, POSE = Poa secunda (sandberg bluegrass), RORIP = Rorippa ssp. (yellowcress), SIAL2 = Sysimbrium altissimum (tall tumblemustard), VUBR = Vulpia bromoides (six-weeks fescue) 152
Figure 3 Six plant species shared between the seed bank and aboveground cover reported as mean relative percent (%) cover for aboveground and mean relative density (seeds•m2) for belowground.
Plant codes: BRTE = Bromus tectorum (cheatgrass), VUBR = Vulpia bromoides (six-weeks fescue), ARTRW8 = Artemisia tridentata ssp. wyomingensis (Wyoming big sagebrush), CETE5 = Ceratocephala testiculata (bur buttercup), ELEL5 = Elymus elymoides (squirreltail), POSE = Poa secunda (sandberg bluegrass) 153
Figure 4 a.) Mean percent (%) aboveground cover of observed species categorized into 9 functional groups and compared between ungrazed (UG) and grazed (G) areas across ecological sites ascending from lowest (1339 m) to highest elevation (1799 m). b) Mean density (seeds•m2) of species that germinated in the soil seed bank categorized into 9 functional groups and compared between UG and G areas. Means and standard error (SE) (±) bars reported.
a)
**
**
b)
* * *
*
* * * *
* Treatment has a significant effect on cover group within site, p < 0.05 ** Treatment has a significant effect on cover group within site, p < 0.01 Variables: GNS = Gravelly North Slope, AF = annual forb, GRSG = Greater sage-grouse , NAF = native annual forb, NF = native forb, PG = perennial grass 154
Figure 5 a) Mean (%) aboveground cover of cheatgrass between ungrazed (UG) and grazed (G) areas in each site in order from lowest (1339 m) to highest elevation (1799 m). Dinner Station was the only site without observed cheatgrass cover in the canopy for both treatments. b) Mean density (seeds•m2) of cheatgrass that germinated in the seed bank between UG and G areas in each site. Means and standard error (SE) (±) bars reported. a)
b)
155
Figure 6 Plant species identified as indicators in aboveground communities across Paradise Valley 1 (PV1), Paradise Valley 2 (PV2), and Dinner Station (DS). Pearson’s correlation residuals extracted from chi-squared analysis were significant at ±2. Species with positive residuals show there is a preference for associating in the canopy within site. Species with negative residuals indicate a preference of disassociation in the canopy within site.
Plant codes: ALAC4 = Allium acuminatum, ARARA = Artemisia arbuscula, ARTRW8 = Artemisia tridentata ssp. wyomingensis, BLSC = Blepharipappus scaber, ERNA10 = Ericameria nauseosa, PHHO = Phlox hoodii, VUBR = Vulpia bromoides 156
Figure 7 Plant species identified as indicators in belowground communities across Paradise Valley 1 (PV1), Paradise Valley 2 (PV2), and Dinner Station (DS). Pearson’s correlation residuals extracted from chi-squared analysis were significant at ±2. Species with positive residuals indicate a preference for associating with the seed bank within the site. Species with negative residuals indicate a preference of disassociation with the seed bank within site.
Plant codes: BRTE = Bromus tectorum, COPA3 = Collinsia parvifolia, DRVE2 = Draba verna, HOUM = Holosteum umbellatum, MYAP = Myosurus apetalus, PG1 = perennial grass unidentified, VUBR = Vulpia bromoides 157
Figure 8 a) Comparison of Shannon-Weaver diversity index (H’) measuring aboveground diversity between ungrazed (UG) and grazed (G) areas across Paradise Valley 1 (PV1), Paradise Valley 2 (PV2), and Dinner Station (DS) in order by ascending elevation. b) Comparison of Shannon-Weaver diversity index (H’) measuring belowground diversity between ungrazed (UG) and grazed (G) areas across Paradise Valley 1 (PV1), Paradise Valley 2 (PV2), and Dinner Station (DS) in order by ascending elevation. Error bars indicate 95% confidence intervals. Higher values indicate greater diversity (H’) while lower values indicate lower diversity by site.
a) b)
*
* Treatment has a significant effect on diversity within site p < 0.05