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Article

Understory Vegetation Change Following Woodland Reduction Varies by Plant Community Type and Seeding Status: A Region-Wide Assessment of Ecological Benefits and Risks

1
U.S. Department of Agriculture, Agricultural Research Service, Forage and Range Research Laboratory, Utah State University, Logan, UT 84322-6300, USA
2
Great Basin Research Center, Utah Division of Wildlife Resources, Ephraim, UT 84627, USA
*
Author to whom correspondence should be addressed.
Plants 2020, 9(9), 1113; https://doi.org/10.3390/plants9091113
Received: 7 August 2020 / Revised: 19 August 2020 / Accepted: 27 August 2020 / Published: 28 August 2020
(This article belongs to the Section Plant Ecology)

Abstract

Woodland encroachment is a global issue linked to diminished ecosystem services, prompting the need for restoration efforts. However, restoration outcomes can be highly variable, making it difficult to interpret the ecological benefits and risks associated with woodland-reduction treatments within semiarid ecosystems. We addressed this uncertainty by assessing the magnitude and direction of vegetation change over a 15-year period at 129 sagebrush (Artemisia spp.) sites following pinyon (Pinus spp.) and juniper (Juniperus spp.) (P–J) reduction. Pretreatment vegetation indicated strong negative relationships between P–J cover and the abundance of understory plants (i.e., perennial grass and sagebrush cover) in most situations and all three components differed significantly among planned treatment types. Thus, to avoid confounding pretreatment vegetation and treatment type, we quantified overall treatment effects and tested whether distinct response patterns would be present among three dominant plant community types that vary in edaphic properties and occur within distinct temperature/precipitation regimes using meta-analysis (effect size = lnRR = ln[posttreatment cover/pretreatment cover]). We also quantified how restoration seedings contributed to overall changes in key understory vegetation components. Meta-analyses indicated that while P–J reduction caused significant positive overall effects on all shrub and herbaceous components (including invasive cheatgrass [Bromus tectorum] and exotic annual forbs), responses were contingent on treatment- and plant community-type combinations. Restoration seedings also had strong positive effects on understory vegetation by augmenting changes in perennial grass and perennial forb components, which similarly varied by plant community type. Collectively, our results identified specific situations where broad-scale efforts to reverse woodland encroachment substantially met short-term management goals of restoring valuable ecosystem services and where P–J reduction disposed certain plant community types to ecological risks, such as increasing the probability of native species displacement and stimulating an annual grass-fire cycle. Resource managers should carefully weigh these benefits and risks and incorporate additional, appropriate treatments and/or conservation measures for the unique preconditions of a given plant community in order to minimize exotic species responses and/or enhance desirable outcomes.
Keywords: conifer encroachment; large-scale restoration; seeding; ecological site potential; woodland expansion; effect-size analysis; regional assessment; vegetation analysis; functional group conifer encroachment; large-scale restoration; seeding; ecological site potential; woodland expansion; effect-size analysis; regional assessment; vegetation analysis; functional group

1. Introduction

Semiarid ecosystems are currently threatened by woody plant dominance due to encroachment (i.e., spreading) and infilling (i.e., densification), heightening the need to understand how these changes impact ecosystem functioning and critical ecosystem services [1,2]. Numerous factors, including elevated atmospheric CO2, increased N deposition, climate shifts, reductions in fire, and changes in grazing/browsing regimes are believed to play an important role in woody plant encroachment [2]. Similar to global trends, multiple interacting factors have been attributed to coniferous tree expansion (e.g., single-leaf piñon pine (Pinus monophyla Torr. and Frém.), Colorado piñon pine (P. edulis Engelm.), Utah juniper (Juniperus osteosperma Torr.), and western juniper (Juniperus occidentalis Hook); hereafter P–J (pinyon–juniper)) into semiarid shrub-steppe ecosystems in the Intermountain Region of western North America [3,4,5]. These factors include (1) natural range expansion [6], (2) decreased fire frequency due to cessation of periodic fires after European arrival, active fire control, and the creation of fire barriers [3,7,8,9,10,11], (3) introduction of livestock grazing and heavy grazing following the arrival of Europeans that reduced fuels and altered competitive interactions between herbaceous species and trees [4,8,10,12,13,14,15], (4) favorable climatic conditions, especially during wetter and cooler conditions between 1900 and 1950 [16,17,18,19,20], (5) afforestation following prior woodland reduction and harvesting [21], and (6) woodland recovery from megadroughts in late 1500s [6,10]. Prior to European settlement of the western U.S., P–J woodlands occurred on fewer than 3 million ha, but estimates indicate distribution across foothills, mesas and plateaus, and low mountain woodlands now occupy an area as high as 50 million ha in western U.S. [22,23], more than 18 million ha in the Intermountain West [17,22,24,25], and between 4 and 6 million ha in Utah alone (i.e., more than 25% of its land area) [23,25,26,27]. Current estimates also indicate that P–J woodlands have increased within the range of 125–625% since 1860 due to encroachment into shrub-steppe ecosystems that did not previously support trees and infilling within shrub-steppe woodlands [12,14,16,19,20,28,29,30]; yet patterns have not been uniform throughout the Intermountain Region and the relative importance of potential factors causing P–J expansion is largely unknown for most locations in this region [3].
While woody plant encroachment does not universally degrade ecosystems [31], in the Intermountain Region of western North America it is a primary conservation concern due to negative impacts on sagebrush (e.g., black sagebrush, Artemisia nova A. Nelson and big sagebrush, A. tridentata spp. Nutt.)-dominated semiarid shrubland and shrub-steppe ecosystems [4,7,8,14,16,20,32,33,34,35]. Encroachment of P–J in this region has been linked specifically to sharp reductions in herbaceous understory vegetation and species diversity [30,36,37,38,39,40,41,42,43,44,45], increases in flammable exotic annual grasses and the risk of creating an annual grass-fire cycle [32,33,46,47], soil instability, soil erosion, and reduced hydrological functioning [48,49,50,51,52,53]. The reduction and/or absence of desirable herbaceous vegetation on encroached sites has also led to inadequate seed banks to allow natural regeneration after the application of P–J reduction treatments [54,55,56], but see [57]. Furthermore, degraded understory vegetation associated with increasing tree density diminishes habitat suitability for wildlife species, including regionally important mule deer (Odocoileus hemionus) [42,58] and greater sage-grouse (Centrocercus urophasianus) [59,60,61,62,63,64]. Hazardous woody fuel build-up is also threatening sagebrush communities due to the extreme risk of intensive wildfires [12,19,65]. Given these conservation concerns and ecological impacts of woodland encroachment, proactive P–J reduction and restoration seeding are viewed as ways to improve the capacity of shrublands and shrub-steppe plant communities to support greater ecosystem services, including expediting the recovery of understory native shrub and herbaceous species [66,67], and creating suitable habitats for imperiled avian species [61,68]. Furthermore, proactive management associated with mechanical P–J reduction increases water accumulation in winter, infiltration rates following precipitation events, and soil water availability in spring [15,69,70,71,72,73,74], thus altering key ecological processes necessary to enhance understory vegetation [67,75,76]. However, the conservation benefits of such treatments have not been consistently realized [1,77,78,79,80], the efficacy of this management strategy has been highly variable, and the longevity of removal/reduction treatments often do not exceed 10 years [2,78]. Consequently, there is critical need for empirical assessments of large-scale restoration projects at the regional scale to uncover patterns in posttreatment vegetation dynamics, enhance our ability to choose the most appropriate site-specific treatments for future restoration [80,81,82,83], and inform the public on how restoration activities are achieving management goals [84,85,86,87,88].
Although the factors responsible for idiosyncratic restoration outcomes in woodland ecosystems are not fully understood [67,78,79,89,90], evidence indicates that pretreatment vegetation and unique biophysical conditions prevalent within distinct plant communities are important determinants of understory recovery [91,92,93,94] and habitat suitability for ground-nesting birds [95,96,97]. For example, the recovery of understory vegetation following P–J removal depends on both pretreatment levels of woodland encroachment [4,12,49,98] and abundance of native vegetation [39,79,91,99,100,101,102], which are often inversely related. This inverse relationship indicates that strong competitive interactions for soil resources are responsible for the contingence between pretreatment tree canopy cover and/or density in P–J woodlands and posttreatment herbage production [30,39,45,56,57,103,104,105]. For example, because rooting zones of trees can overlap substantially [106] and strongly compete with understory vegetation for limiting resources [107,108], P–J reduction is expected to liberate soil resources necessary for the understory recovery [76,94]. Thus, sites with greater pretreatment understory abundances of perennial grasses and native shrubs are expected to have higher recovery potential compared to sites with advanced phases of woodland development and severely degraded understory vegetation [94,109]. Sites with advanced woodland development are also more prone to invasive grass increases after P–J removal than less-developed woodland sites [39,91,110], but when pretreatment native vegetation contains a high abundance of perennial grasses, understory recovery can preclude posttreatment dominance of exotic annual species by competing for resources made available after P–J removal [111,112,113,114,115] as well as influence posttreatment native shrub abundance [90,113].
The paramount influence of pretreatment vegetation on posttreatment understory recovery may in fact muddle the interpretation of restoration outcomes in woodland ecosystems because the choice of treatment type is usually based on pragmatic and/or workable features of treatment applications, which can confound a clear interpretation of the influences of treatment type and pretreatment vegetation on posttreatment responses. For example, in the absence of fire to naturally regulate woodland encroachment [12,19,116], numerous mechanical treatments (i.e., chaining, mastication; e.g., shredding and dispersing mulch, and cutting) have been developed to function as fire surrogates for fuel reduction, watershed improvement, and to restore understory vegetation components [67,76,90,117,118,119]. However, the suitability of each treatment depends on pretreatment vegetation and ecological site characteristics [5,32], including the amount of understory herbaceous and shrub species and the severity of P–J encroachment [117,120]. Chaining is typically applied to sites with larger-diameter trees, higher tree cover, and degraded understory vegetation [121,122] and creates greater ground disturbance than mastication and cutting, which can in turn increase the density of invasive annual species [89,115,123,124]. Thus, chained sites are nearly always seeded and inherent soil disturbance associated with chaining is considered necessary to alter seed bed conditions and increase establishment of seeded species [95,125,126,127,128,129,130,131,132,133,134,135]. Mastication is also suitable for sites characterized by later stages of woodland development but is followed by seeding only if pretreatment understory conditions are degraded [40,60,89]. Mastication is also unique compared to the other treatments due to its production and dispersal of mulched residue that reduces bare ground, erosion and runoff [136,137,138,139,140], increases water infiltration rates, and reduces sediment yields relative to areas lacking the masticated residue [137]. Studies also indicate that this residue can potentially reduce seedling emergence of seeded species [132,141] and the tracked vehicles used to apply this treatment can decrease soil aggregate stability [137,142]. In contrast to chaining and mastication, cutting maintains understory shrub and herbaceous cover with minimal ground disturbance [91,102,124,139] and is most appropriate for sites with low tree density that do not require seeding [67,69,139,143]. Consequently, due to pretreatment vegetation conditions and disturbance regimes intrinsic to each treatment type, treatment types are expected to yield variable restoration outcomes [144,145], but it remains difficult to extrapolate restoration outcomes from disparate studies to other situations because treatment efficacy and pretreatment conditions are not mutually independent.
As P–J encroachment has occurred over a diverse range of topographic, climatic, and edaphic conditions in the Intermountain West [24,26,146,147,148], categorical plant community classifications that incorporate biophysical-site properties and vegetation-recovery potentials offer a practical platform to make impartial comparisons among treatment alternatives and decipher restoration outcomes [90,149]. Thus, constraining the direct comparison of treatment alternatives within plant community types can partially address the uncertainties created by idiosyncratic restoration outcomes in woodland ecosystems [78] and reveal the site-specific conditions where restoration may be most successful [94,109,150,151,152]. For example, plant communities encroached by P–J species are commonly classified by sagebrush taxa [147,153,154,155,156] that dominate within distinct soil temperature/soil moisture regimes. At one extreme, higher elevation mountain big sagebrush communities are characterized by cool/moist (i.e., frigid/xeric) precipitation/temperature regimes, receive higher average annual precipitation, and have higher soil water holding capacity compared to Wyoming big sagebrush and black communities that dominate warm/dry (mesic/aridic) regimes at lower elevations and on soils with lower water holding capacity [157,158,159,160]. Black sagebrush communities also dominate on gravelly soils with clay-textured or calcified subsoil horizons (i.e., caliche) that create shallow rooting depths and poor water infiltration compared to communities dominated by big sagebrush species [161,162]. Lower elevation black- and Wyoming-big sagebrush communities experiencing P–J encroachment are also more susceptible to annual grass invasion compared to mountain big sagebrush communities in the Intermountain Region [163,164], which leads to key differences in resistance to exotic plant invasion and resilience following disturbance and environmental stress [109,165,166,167]. These differences underpin the capacity of plant community types to serve as effective environmental surrogates [109,165] and emphasize the need to assess restoration outcomes of alternative P–J reduction treatments within plant community types in order to enhance site-specific management recommendations [165,168].
In this study, we assessed vegetation change over a 15-year period following landscape-scale P–J reduction applied at 129 woodland sites in Utah, USA. Our overarching objective was to assess both restoration benefits (i.e., positive effects on understory herbaceous-perennial and shrub components) and ecological risks (i.e., positive changes in exotic annual species) associated with this effort to reverse woody plant encroachment. To do this, we evaluated the relationships among pretreatment vegetation components and calculated effect-sizes for eight vegetation and soil surface variables to uncover the contingence of posttreatment vegetation change on plant community type for three common mechanical P–J reduction treatments using meta-analysis; an analytical approach particularly suitable to study different disturbance types applied across highly variable conditions [85,144]. Meta-analysis was also used to assess how restoration seedings contributed to overall changes in understory vegetation components within plant community types. Due to strong competitive interactions between P–J species and understory vegetation, we expected that increased soil resources following P–J reduction treatments would ubiquitously lead to significant positive changes in understory vegetation. However, because treatment types with higher disturbance intensity are routinely applied to sites with more advanced pretreatment woodland development, we expected that the magnitude of change for herbaceous and shrub components would be greater for chaining and mastication compared to the lower intensity cutting treatment. We also expected positive changes would be larger for understory vegetation in mountain big sagebrush communities that typically exhibit greater resilience to disturbance than other sagebrush communities (e.g., [94,157,169]). Reconciling these expectations will better inform ongoing conservation efforts to offset woody plant encroachment in semiarid ecosystems and refine the development of management guidelines that incorporate site-specific criteria when planning and executing restoration projects.

2. Results

2.1. Pre-Treatment Vegetation

Pretreatment differences among mechanical treatments and plant community types were found for six of the eight response variables (Table S1). Significant differences included higher P–J cover for chaining and mastication compared to cutting and variation in the amounts of both sagebrush and perennial grasses ranging from low values for chaining, intermediate values for mastication, and high values for cutting (Figure 1A). In contrast, differences among plant communities included lower annual grass cover and higher cryptogam cover, in black sagebrush communities compared to big sagebrush communities (Figure 1B). Perennial forb cover was significantly greater for mountain big sagebrush than Wyoming big sagebrush communities, while values were intermediate for black sagebrush communities. Inverse relationships between pretreatment P–J cover and both sagebrush and perennial grass cover were significant for all plant community types with the exception of sagebrush cover in black sagebrush communities (Figure 2A,B). Lastly, pretreatment cover values (mean ± SE) were significantly greater on unseeded sites than seeded sites for sagebrush (5.6 ± 0.9 vs. 2.7 ± 0.8, respectively; t = 3.1; p = 0.002), perennial grass (9.8 ± 1.0 vs. 1.8 ± 0.5, respectively; t = 7.3; p < 0.001), and perennial forb (0.9 ± 0.2 vs. 0.4 ± 0.1, respectively; t = 2.7; p = 0.008).

2.2. Plant Community Responses to P–J Reduction

The overall effects of P–J reduction were significant for all eight response variables; effects were positive for sagebrush, all herbaceous components, and soil surface variables but negative for P–J and bare ground (Figure 3, Figure 4 and Figure 5). Treatment effects (i.e., communities pooled) were also more positive for chaining and mastication compared to cutting for herbaceous vegetation, but plant communities did not generally respond differently within a treatment type except for annual grass (Figure 4C) and cryptogam (Figure 5B), which were more positive for black sagebrush communities within the chaining treatment.
Although plant community types differed for only two of the eight variables, the significance of effect sizes (i.e., Ho: μ = 0; α = 0.05) varied among plant communities in many instances. For example, greater resilience of mountain big sagebrush communities was evident from significant positive effect sizes for mountain big sagebrush communities but not the other community types for sagebrush cover (i.e., chaining treatment; Figure 2B), perennial grass cover (i.e., cutting treatment; Figure 3A), and a significant negative effect size for bare ground (i.e., cutting treatment; Figure 5A). In contrast, a non-significant effect size for bare ground within the chaining treatment for Wyoming big sagebrush communities indicated that vegetation recovery did not result in parallel changes in bare ground as was observed in the other community types (Figure 5A). Non-significant effects for perennial and annual forb cover within black sagebrush communities in the chaining treatment, where positive effects on annual grasses were most pronounced, also differs from responses seen in the two big sagebrush communities (Figure 4B,D). However, unlike chaining, the effect size for annual grass cover was not significantly different than zero for black sagebrush communities in the mastication treatment yet chaining had a significant effect on annual grass cover in both big sagebrush community types (Figure 4C).

2.3. Influence of Seeding on Post-Treatment Understory Recovery

Overall effect sizes for seeding were significant for perennial grass and perennial forb, but not sagebrush (Figure 6). Seeding also had significant, positive effects on perennial grass and perennial forb in many plant community–treatment combinations, but significant differences between seeded and unseeded sites were observed only at mountain big sagebrush sites for perennial forb (regardless of treatment) and for perennial grass in the mastication treatment. Lastly, overall annual grass responses were not significantly different between seeded and unseeded sites, nor were differences found for any of the treatment–plant community combinations (data not shown).

3. Discussion

3.1. Overall Effects of P–J Reduction and Seeding were Dependent on Disturbance Intensity and Pre-Treatment Vegetation

The overall effects of reversing woody plant encroachment through mechanical P–J reduction in our assessment revealed positive changes in all understory vegetation components and supported the expectation that higher disturbance associated with chaining and mastication would lead to more dramatic changes in herbaceous components relative to the cutting treatment. Thus, although high disturbance intensity can greatly alter soil surface attributes, injure non-target plants, and disrupt species recruitment [111,168,170,171], our results illustrated that sites treated with chaining and mastication stood to gain the most from P–J reduction (i.e., based on pretreatment conditions; Figure 1A) and experienced changes in understory conditions yielding both beneficial ecological services and undesirable ecological risks. In contrast, it was surprising that comprehensive positive changes in understory vegetation were not more prominent in mountain big sagebrush communities that typically exhibit greater resilience to disturbance than other sagebrush communities [94,109]. Instead, differences among plant communities were rarely observed within treatment types (i.e., annual grass and cryptogam under chaining; Figure 4C and Figure 5B, respectively); however, both significant and non-significant effect sizes were observed for all response variables except pinyon–juniper when comparing plant communities within treatment types. These nuanced results, thus, partially support the expected resilience of plant community types and highlight specific situations where caution is warranted in order to avoid unintended consequences associated with P–J reduction. In addition, while we observed dramatic overall effects of seeding on perennial grass and forb components in support of the widely held expectation that seeding contributes to understory recovery [66,142,165,172], these increases must be examined from the perspective of pretreatment conditions that dictated whether seeding was necessary as well as the treatment–plant community backdrop in order to fully understand and appraise situations where seeding can yield the most benefits to understory recovery.

3.2. Why Were Understory Responses More Pronounced for Chaining and Mastication Than Cutting?

While our results agree with previous studies indicating that pretreatment levels of encroachment and the abundance of residual perennial vegetation are key factors associated with variation in treatment outcomes [4,12,39,49,91,98,173], we stress that less-pronounced overall changes in herbaceous perennial vegetation within the cutting treatment should not be viewed as an unfavorable treatment outcome [89,103]. Instead, this pattern reflects the fact that cutting sites were in earlier phases of woodland development that had inherently greater pretreatment residual vegetation to support understory recovery (i.e., lower pretreatment P–J cover and higher pretreatment sagebrush perennial grass cover; Figure 1A; [67,91,174]). Our results also emphasize that although higher initial cover of perennial vegetation equated to lower overall posttreatment change for cut sites, positive responses in perennial grass and cryptogam suggest that increases in soil water and nutrients that typically accompany P–J reduction likely triggered these significant responses [39,72]. These changes in understory vegetation and soil surface cover probably contributed to the competitive effects of perennial grasses and the absence of positive changes to undesirable annual grass and forb components (i.e., Figure 4C,D) that typically increase following P–J reduction [125,175] as was the case for chaining and mastication. It is also feasible that greater residual perennial grass plants might have suppressed perennial forbs on cutting sites as reported in other studies when perennial grasses begin to dominate over time [71,110,126,176]. This interpretation is based on the understanding that herbaceous species typically exhibit overlapping resource use, perennial grasses are effective competititors, and that cryptogams can potentially prevent annual grass establishment [177,178,179].
In contrast to cutting, inherently low pretreatment perennial cover grass in the chaining and mastication treatments magnified the capacity of annual grasses to proliferate following P–J reduction [89,98]. For example, chaining directly disturbs the soil surface, creating favorable safe-site microenvironments for seedling emergence and establishment [180,181]. While this mechanism explains the robust changes associated with chaining in understory vegetation, including the only significant positive change in sagebrush (i.e., at mountain big sagebrush sites), it had a similar effect on annual grasses, particularly at black and mountain big sagebrush sites [79,182,183,184]. Mastication is also known to promote annual grass establishment, more so than perennial grasses ([98,132,141,185,186,187], but see [188]) because the production and distribution of mulch favors annual grass growth by reducing soil temperature, increasing soil moisture, and elevating inorganic nitrogen supply to plants [39,73,132,189,190]. These results indicate that annual grasses will likely proliferate in the short-term even when perennial grasses increase following P–J reduction [116,142,187]. However, because perennial grasses are known to effectively suppress annual grasses (i.e., [110,175,191]), the expectation is that steady increases in perennial grass cover will diminish this threat over time with proper posttreatment management [67,118,167,191,192]. Consequently, extra effort during the posttreatment period (i.e., [5,167,192]) will be essential to enhance the capacity of understory herbaceous vegetation to recover and mitigate the risk of stimulating an annual grass-fire cycle (i.e., [39,116,142,193]). In addition, because livestock grazing is a key factor in the expansion of P–J through its direct influence on both perennial grass and shrub cover, it may be necessary to adjust management plans to reduce the speed of P–J recovery on treated sites [5,194,195].

3.3. What Ecological Processes Were Responsible for the Differences in Understory Resilience and Annual Grass Response among Plant Community Types?

Although sagebrush plant community types are known to vary widely in environmental and topo-edaphic characteristics [123,126,154,167,192], we found only marginal support for the expectation of more pronounced understory vegetation responses in mountain big sagebrush (i.e., cool/moist temperature/moisture regimes; moderately deep, loamy to clay loam soils; >300 mm annual precipitation; [94,157,169]) compared to Wyoming big sagebrush (i.e., warm/dry; moderately deep, loamy soils; 200–300 mm annual precipitation) or black sagebrush plant communities (i.e., warm/dry; shallow, stony, calcareous soils; <300 mm annual precipitation) [164,192,196]. Instead, our results indicated strong positive perennial grass and forb recovery after P–J reduction in the chaining and mastication treatments and similar levels of resilience among plant communities (Figure 4A,B). It was also surprising that the positive changes in perennial grass and forb seen in black and Wyoming big sagebrush sites matched those of mountain big sagebrush sites within the chaining and mastication treatments. We speculate that the unexpected positive responses for black and Wyoming big sagebrush communities might have stemmed from posttreatment increases in soil moisture (i.e., [69,72,182]) causing greater net increases in resource availability on these warm/dry sites than on cool/moist mountain big sagebrush communities that typically receive greater annual precipitation because of their distribution at higher elevations. This interpretation is also supported by the fact that lifeforms known to be effective indicators of resource availability (i.e., cryptogams and annual grasses [197,198,199,200]), experienced greater changes from chaining in the black sagebrush compared to mountain big sagebrush sites. Similarly, the positive changes in cryptogams on drier black and Wyoming sites became statistically significant in response to mastication and cutting, whereas changes were not significant on mountain big sagebrush sites. Our explanation for the unexpected resilience on the drier sites supports broadscale studies illustrating that favorable changes in site ecohydrology are more pronounced in P–J woodlands that experience greater increases in understory vegetation, which in turn reduces runoff and increases water infiltration [76,201,202].
Variation in annual grass response among plant community types in the chaining treatment were also an unexpected result of our assessment. Surprisingly, greater positive annual grass response for black sagebrush sites where pretreatment annual grass cover was lowest was opposite to the response observed in Wyoming big sagebrush communities. As significant positive changes in perennial grasses were observed during the posttreatment period for all plant community types, this pattern was probably a consequence of neither perennial nor annual forbs capitalizing on posttreatment increases in soil water and nutrients that accompany P–J reduction [44,72,203], which allowed greater positive changes for annual grasses in black compared to Wyoming big sagebrush plant communities. In contrast, competitive interactions among understory vegetation for soil resources cannot explain the positive annual grass responses we observed at higher elevation in mountain big sagebrush plant communities that are expected to have higher perennial plant productivity (i.e., cool/moist temperature/precipitation regimes) and resistance to exotic plant invasion than warm/dry sagebrush sites [44,94,169,172]. These results indicate that biotic resistance to annual grass expansion can be overridden when annual grasses are present prior to applying P–J reduction treatments, even when all perennial understory components simultaneously experience strong positive changes, and potentially provide competition for soil resources [136,187,204,205]. This conclusion also applies to mastication sites where annual grass responses were generally positive even though the other herbaceous understory components exhibited strong positive change. Thus, analogous to our previous recommendations to mitigate annual grasses after P–J reduction (i.e., Section 3.2), posttreatment management plans should routinely include interventions that target areas most susceptible to annual grass expansion. For example, while successional models recognize the risk of annual grass and annual forb expansion in the first few years after P–J reduction [5,110,206,207], broadcast herbicide applications of hotter/drier areas and/or spot herbicide applications of smaller microsites previously occupied by trees could be used during this critical period to target areas where annual species flourish due to the absence of perennial grass competition [208,209,210].

3.4. When Is Seeding Essential for Understory Recovery Following P–J Reduction?

Seeding after juniper removal is generally considered essential to provide effective suppression of exotic annual grasses through competition for spring soil moisture [66,72,126,130] and reestablish understory herbaceous vegetation [118,125,126,131,211], especially when native species seed reserves are depauperate [66,115,126,212,213]. However, these benefits are often assumed (i.e., the majority of mechanically treated P–J woodlands were also seeded [25,126,214,215]), and few studies have specifically isolated the benefits of P–J reduction alone versus the combined application of P–J reduction and seeding [40,89,193]. While responses between unseeded and seeded sites did not vary for annual grass as illustrated in a recent study [89], the variable herbaceous perennial responses we observed among plant communities and treatments indicate situations where seeding is most essential to achieve restoration goals.
As pretreatment abundance of perennial understory vegetation typically dictates whether sites are seeded (e.g., Figure 1; [67]), we were not surprised that the overall changes in perennial grass and forb components were more pronounced for seeded compared to unseeded sites and for mastication compared to cutting treatments. It was also not surprising that these benefits were more evident at cool, high elevation mountain big sagebrush sites that receive greater precipitation compared to warm, low elevation Wyoming big sagebrush sites ([151,172], but see [165]). In fact, only mountain big sagebrush sites treated with mastication showed consistently greater positive changes to both herbaceous components for seeded compared to unseeded sites. These results agree with previous studies reporting greater increases for mastication than cutting [142] and enhanced seedling establishment in mastication treatments where mulch deposition ameliorates seedbed temperature and increases surface soil water availability and inorganic soil nitrogen content [73,98,132,189,190]. Our results also indicated that the majority of responses for perennial grass and forb components became significant only for seeded sites, suggesting that seeding might be necessary more often than typically prescribed [89]. In contrast, the absence of perennial forb response at both seeded and unseeded Wyoming big sagebrush sites treated with cutting illustrates that seeding may not be necessary in this situation because cutting sites typically have higher pretreatment abundance of perennial understory vegetation and lower exotic annual weeds compared to sites receiving mastication [72,132,187,189].
Although sagebrush recovery was not deterred by P–J treatments, there were no instances across community and/or treatment types where seeding enhanced sagebrush recovery. These results suggest that factors other than seed limitation (e.g., [66,115,126]) may be responsible for poor shrub recovery. For example, sagebrush establishment is known to be both episodic [216,217] and strongly sensitive to low spring precipitation, which could have hampered seed germination and seedling establishment, even when seeds were available (e.g., [172,218]). In addition, high utilization of treated areas by big game ungulates can limit sagebrush recovery [89,219,220]. For example, our capacity to detect changes associated with seeding [165] was likely reduced by mule deer, which preferentially browse sagebrush on thinned areas to meet their winter dietary needs [221]. Lastly, perennial and annual grass seedlings strongly compete with sagebrush seedlings and can greatly reduce nutrient acquisition, growth, and survival of sagebrush seedlings [178,222,223]. In particular, rapid establishment of seeded perennial grasses can hinder the establishment of new sagebrush seedlings and slow sagebrush recovery on restoration sites [151,165,224,225].

3.5. Conclusions

Our results illustrated strong associations among pretreatment vegetation components, a critical factor underscoring the pressing need to reverse woody plant encroachment across this semiarid ecosystem. Furthermore, we showed that recognizing intrinsic differences in pretreatment vegetation among treatment and community types is critical to disentangle confounding factors and interpret posttreatment response to P–J reduction. We found strong and consistent changes in bare ground as well as robust changes in understory vegetation components. As expected, variation among plant community types was larger for higher-disturbance chaining and mastication treatments compared to lower-impact cutting treatments, but understory recovery was surprisingly similar among plant community types that experience highly variable temperature/precipitation regimes. However, positive changes in perennial understory components did not effectively diminish annual grass responses, which appeared to be an inherent ecological risk of modifying soil surface conditions. We also found strong positive effects of seeding on desirable herbaceous recovery for the majority of plant community–treatment combinations, suggesting that seeding is an essential determinant of whether the management objectives of reversing woodland encroachment are met or not. Finally, our results confirm that P–J reduction does not negatively impact shrub abundance [226]. Collectively, this assessment contributes to an improved understanding of restoration outcomes following landscape-level treatments to reduce woody plant encroachment and provides empirical evidence regarding how such treatments provision ecosystem services (i.e., livestock forage and wildlife habitat) via increases in understory herbaceous cover [2,227,228] and regulate additional ecosystem services (i.e., erosion control) through reductions in bare ground and increases in cryptogam cover [227,229,230].

4. Materials and Methods

4.1. Project Sites

To evaluate changes in vegetation and ground surface variables, we used data obtained from Utah Watershed Restoration Initiative (UWRI) project sites that were treated for P–J reduction between 1999 and 2016 (UWRI; https://wri.utah.gov/wri/). Project sites were primarily distributed across Utah within three Ecoregions, namely, Central Basin and Range, Wasatch and Uinta Mountain, and Colorado Plateau [231,232] and ranged in elevation between 1600 and 2454 m, with a mean elevation of 1995 m (Figure 7). Soil textures were generally classified as loam with derivations of sandy, clay, sandy clay, and silt loam. After considering all possible project sites, we selected a subset of 129 sites that met the following criteria: (1) mechanical treatments were applied to increase the abundance of big sagebrush and herbaceous species and enhance the wildlife habitat for mule deer (Odocoileus hemionus), Rocky Mountain elk (Cervus canadensis), and/or greater sage-grouse (Centrocercus urophasianus); (2) pretreatment plant communities could be defined by dominant shrub species (i.e., black sagebrush (Artemisia nova), Wyoming big sagebrush (A. tridentata ssp. wyomingensis), or mountain big sagebrush (A. tridentata ssp. vaseyana)); (3) one of three prevalent P–J reduction treatments were applied (i.e., chaining, mastication, cutting); (4) both pre- and posttreatment data were available for analysis.
Plant community types within this region differ in many factors, including soils, elevation, and biophysical indicators [233] that have been related to post-disturbance ecosystem resilience and resistance to exotic plant invasions [109,164,167,169]. For example, A. tridentata ssp. vaseyana dominates higher elevation montane areas with cold–moist temperature/precipitation regimes, greater primary productivity, and higher overall resistance and resilience capacity compared to the other two plant community types [167,169]. In contrast A. tridentata ssp. wyomingensis and A. nova typically occur at lower elevations with warm–dry temperature/precipitation regimes and notably lower resistance and resilience tendencies [109,158,159,167,169], with the former being more common at higher elevation moister soils and the latter dominating where shallow soils are underlain by a distinct petrocalcic (caliche) layer [157,159].

4.2. Treatments and Seeding

All treatments to reduce P–J abundance on project sites were applied with mechanical devices as part of the Utah Watershed Restoration Initiative (UWRI) (https://wri.utah.gov/wri/). The choice of treatment applied to each site was primarily determined by pretreatment woodland developmental phases that broadly differ in P–J cover (i.e., Phase I < 10%, Phase II = 10–30%, and Phase III > 30%; [4,12,49]); chaining treatments were accomplished by pulling a segment of a naval surplus anchor chain between two bulldozers. The chains were fashioned into an Ely chain by welding short lengths of iron across each link [180]. This treatment uproots trees, scarifies the soil surface, and creates a seedbed for broadcast seeding. To increase efficacy of uprooting trees on sites dominated by live trees (e.g., not killed by previous wildfire), the chain was pulled in two opposite directions [129]. Mastication involved using a rubber-tired or tracked industrial tractor affixed with a rotary cutter that shredded tree biomass to ground level and distributed debris in patches without creating much ground disturbance, other than soil compaction due to tractor treads [137]. Cutting was accomplished with chainsaws by felling individual trees and distributing slash haphazardly on the landscape. In general, UWRI projects apply the cutting (i.e., lop-and-scatter) treatments for plant communities classified as late Phase I to early Phase II, mastication for Phase I to Phase III communities, and chaining for late Phase II to Phase III communities. As cutting treatments typically occur in the early encroachment stages (Phase I), the understory shrub and herbaceous components of the communities are often still relatively intact; whereas tree mastication and chaining treatments typically occur in later encroachment stages (Phase II and Phase III), where the understory shrub and herbaceous components of the plant communities have been depleted to varying degrees. As such, cutting treatments seldom receive supplemental seeding, while mastication treatments receive supplemental seeding when deemed necessary, and chaining treatments receive supplemental seeding the majority of the time.
Supplemental aerial or broadcast seeding of grasses and forbs typically occurs between the first and second pass of the chain to cover seed [126]. In contrast, seeding of shrub and smaller seeded species typically occurs following the second pass. Some larger seeded shrub species such as bitterbrush (Purshia tridentata [Pursh] DC) and fourwing saltbush (Atriplex canescens [Pursh] Nutt.) may be applied using dribbler units attached to the dozer or tractor performing the treatment and pressed into the seedbed with the tire or track action or seeded with small drills following treatment. Supplemental seeding was selectively applied to certain treatment sites when land managers determined there was not an adequate seed source present to reestablish desirable understory species. Seed mixes for each study site varied, but generally consisted of native and introduced grass, forb, and shrub species with emphasis placed on establishing understory vegetation to rapidly stabilize the soil surface from erosion and provide a competitive matrix to minimize exotic annual grass invasion. Species mixes were generally broadcast-seeded aerially using either a fixed-wing airplane or helicopter, with some variation based on species and treatment method (e.g., shrub species applied by tractors doing the reduction treatments).

4.3. Vegetation and Ground Surface Sampling

Data collection occurred from 1997 to 2016, both prior to applying P–J reduction treatments and in subsequent years (typically every three to five years). In addition, roughly half of the project sites were monitored two- or three-times posttreatment. Sampling was conducted by establishing a 152.4-m baseline transect within treatment areas. Along the baseline transect, five 30.5 m belts were placed perpendicular on the 15.2 m mark of each belt at predetermined meter marks (3.4 m, 40.8 m, 78.9 m, 113.0 m, and 150.9 m). A steel stake was placed at the beginning of each belt to ensure consistent placement of future sampling locations. Along each of the five belts, 20, 25 × 25 cm quadrats were placed at 1.5 m intervals to measure canopy cover for herbaceous species and ground surface variables.
Canopy cover for herbaceous species was determined using an ocular cover estimation procedure using seven Daubenmire cover classes within the quadrats [234,235]. The seven cover classes were (1) 0.01–1%, (2) 1.1–5%, (3) 5.1–25%, (4) 25.1–50%, (5) 50.1–75%, (6) 75.1–95%, and (7) 95.1–100%. Similarly, with the quadrat frame on the soil surface, basal cover was also estimated for cryptogam, litter, and bare ground. It is important to note that while cryptogams encompass a broad range of lifeforms in woodland ecosystems (i.e., mosses, algae, lichens, and liverworts [236]), this category primarily consisted of bryophytes (mosses) as opposed to biological cryptogamic or microphytic crusts [237]. To determine basal and canopy cover for each belt, the midpoint for each cover class value was summed and divided by the number of sampling quadrats (i.e., 20). The five belts were used to determine mean and standard error cover percentage and for a given site. Cover of mature big sagebrush and P–J were estimated using the canopy line-intercept method [234,238]. Cover percentages were calculated by dividing the total length along each belt covered by a particular species of tree or shrub by the total length of the belt. Using individual species data, vegetation was grouped into the different categories according to functional group designation: P–J, sagebrush (i.e., Artemisia spp.), perennial grass, perennial forb, annual grass, and annual forb.

4.4. Data Analysis

To assess pretreatment vegetation, the Shapiro–Wilk Goodness-of-Fit Test was applied to all response variables and transformations to improve normality were applied based on Akaike Information Criterion values. Analysis of variance (ANOVA) was used to assess the effects of P–J reduction treatment and plant community type on pretreatment vegetation and ground cover variables (α = 0.05). For significant factors, differences among means were determined with Tukey (HSD) tests (α = 0.05). Variables that could not be transformed to meet ANOVA assumptions were analyzed for main-, but not interaction-effects with non-parametric Kruskal–Wallis tests. For significant factors, differences among means were determined with non-parametric Wilcoxon tests of each pair (α = 0.05). In addition, the associations between P–J and sagebrush and between P–J and perennial grasses were assessed using linear regression analysis in JMP ver. 14 (SAS Institute Corp. Cary, IN, USA). The significance of relationships was determined separately for each plant community type (α = 0.05). Lastly, pretreatment differences between unseeded and seeded sites for perennial grass, perennial forb, and shrub cover were analyzed using unpaired (i.e., independent samples) 2-sample Student’s t-test (α = 0.05).
Due to the limitations of our study design (i.e., variable treatment years, monitoring years and inequality of sites monitored each year) we quantified changes in vegetation and surface variables with standardized effect-size metrics. Effect sizes were calculated using the metafor package [239] for R (www.r-project.org) as the natural log of the ratio between post- and pretreatment (ln[post/pre] = lnRR) for each project site (n = 5) [240,241]. This approach was necessary because control areas were not available for analysis, as is typical in woody plant reduction studies. Depending on the elapsed time since treatments were applied, more than one effect size was calculated for each study site in most cases. Meta-analysis is considered an ideal method to analyze effect sizes and synthesize the outcomes of different treatment types across multi-site, long-term experiments [242]. In R, we used multi-level meta-analysis to test null hypotheses that mean effect sizes are equal to zero (z-test; Ho: μ = 0; α = 0.05). Mixed-effect models accounted for variances associated with sampling error, the random effects of study site (i.e., between-study error), and repeated sampling over time at study sites (i.e., within study error) [85,243]. In addition, fixed effects of treatment and plant community type were coded as moderators. For each variable, we conducted three analyses: (1) a comparison of the three treatment types (plant communities pooled), (2) separate comparisons of plant community type for each treatment type, and (3) an overall analysis that included the entire dataset. Effect sizes were considered significant if 95% confidence intervals did not overlap zero [227,244,245]. Similarly, two effect sizes were considered significantly different if 95% confidence intervals did not overlap each other.

Supplementary Materials

The following are available online at https://www.mdpi.com/2223-7747/9/9/1113/s1, Table S1: Analysis of variance (ANOVA) results for the effects of pinyon-juniper reduction treatment and plant community type on pretreatment vegetation and ground cover variables. Note: variables that could not be transformed to meet ANOVA assumptions were analyzed for main-, but not interaction-effects (i.e., indicated by dashes) with non-parametric Kruskal–Wallis tests. Asterisks indicate significance (*** p < 0.001, ** p < 0.01, * p < 0.05).

Author Contributions

K.L.G. conceived the idea for this study, oversaw data compilation, and co-wrote the paper. T.A.M. analyzed the dataset and co-wrote the paper. Both authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by Federal Aid to Wildlife Grant W-82-R.

Acknowledgments

The authors wish to thank Melissa Landeen and Joe Robins for providing helpful suggestions to improve this manuscript and four anonymous reviewers of an earlier version of this manuscript.

Conflicts of Interest

The authors declare no conflict of interest.

References

  1. Wilcox, B.P.; Birt, A.; Fuhlendorf, S.D.; Archer, S.R. Emerging frameworks for understanding and mitigating woody plant encroachment in grassy biomes. Curr. Opin. Environ. Sustain. 2018, 32, 46–52. [Google Scholar] [CrossRef]
  2. Archer, S.R.; Andersen, E.M.; Predick, K.I.; Schwinning, S.; Steidl, R.J.; Woods, S.R. Woody plant encroachment: Causes and consequences. In Rangeland Systems: Processes, Management and Challenges; Briske, D.D., Ed.; Springer Nature: Cham, Switzerland, 2017; pp. 25–83. [Google Scholar]
  3. Romme, W.H.; Allen, C.D.; Balley, J.D.; Baker, W.L.; Bestelmeyer, B.T.; Brown, P.M.; Eisenhart, K.S.; Floyd, M.L.; Huffman, D.W.; Jacobs, B.F.; et al. Historical and modern disturbance regimes, stand structures, and landscape dynamics in pinon-juniper vegetation of the western United States. Rangel. Ecol. Manag. 2009, 62, 203–222. [Google Scholar] [CrossRef]
  4. Miller, R.; Svejcar, T.; Rose, J. Conversion of shrub steppe to juniper woodland. In Proceedings: Ecology and Mangement of Pinyon-Juniper Communities within the Interior West, 15–18 September 1997, Provo, UT; Proc. RMRS-P-9; Monsen, S.B., Stevens, R., Eds.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Ogden, UT, USA, 1999; pp. 385–390. [Google Scholar]
  5. Miller, R.F.; Chambers, J.C.; Evers, L.; Williams, C.J.; Snyder, K.A.; Roundy, B.A.; Pierson, F.B. The Ecology, History, Ecohydrology, and Management of Pinyon Juniper Woodlands in the Great Basin and Northern Colorodao Plateau of the Western United States; Gen. Tech. Rep. RMRS-GTR-403; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Fort Collins, CO, USA, 2019; p. 284.
  6. Swetnam, T.W.; Allen, C.D.; Betancourt, J.L. Applied historical ecology: Using the past to manage the future. Ecol. Appl. 1999, 9, 1189–1206. [Google Scholar] [CrossRef]
  7. Baker, W.L.; Shinneman, D.J. Fire and restoration of pinon-juniper woodlands in the western United States: A review. For. Ecol. Manag. 2004, 189, 1–21. [Google Scholar] [CrossRef]
  8. Burkhardt, J.W.; Tisdale, E.W. Causes of juniper invasion in southwestern Idaho. Ecology 1976, 57, 472–484. [Google Scholar] [CrossRef]
  9. Bauer, J.M.; Weisberg, P.J. Fire history of a central Nevada pinyon-juniper woodland. Can. J. For. Res. 2009, 39, 1589–1599. [Google Scholar] [CrossRef]
  10. Shinneman, D.J.; Baker, W.L. Historical fire and multidecadal drought as context for pinon-juniper woodland restoration in western Colorado. Ecol. Appl. 2009, 19, 1231–1245. [Google Scholar] [CrossRef] [PubMed]
  11. Miller, R.F.; Rose, J.A. Fire history and western juniper encroachment in sagebrush steppe. J. Range Manag. 1999, 52, 550–559. [Google Scholar] [CrossRef]
  12. Miller, R.F.; Tausch, R.J.; McArthur, E.D.; Johnson, D.D.; Sanderson, S.C. Age Structure and Expansion of Piñon-Juniper Woodlands: A Regional Perspective in the Intermountain West; RMRS-RP-69; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Ed.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Fort Collins, CO, USA, 2008; p. 17.
  13. Miller, R.E.; Rose, J.A. Historic expansion of Juniperus occidentalis (western juniper) in Southeastern Oregon. Great Basin Nat. 1995, 55, 37–45. [Google Scholar]
  14. Blackburn, W.H.; Tueller, P.T. Pinyon and juniper invasion in black sagebrush communities in east-central Nevada. Ecology 1970, 51, 841–848. [Google Scholar] [CrossRef]
  15. Mollnau, C.; Newton, M.; Stringham, T. Soil water dynamics and water use in a western juniper (Juniperus occidentalis) woodland. J. Arid Environ. 2014, 102, 117–126. [Google Scholar] [CrossRef]
  16. Miller, R.F.; Wigand, P.E. Holocene changes in semiarid pinyon-juniper woodlands. BioScience 1994, 44, 465–474. [Google Scholar] [CrossRef]
  17. Miller, R.F.; Tausch, R.J. The role of fire in juniper and pinyon woodlands: A descriptive analysis. In Fire Conference 2000: The first National Congress on Fire Ecology, Prevention, and Management, Proceedings of the Invasive Species Workshop: The Role of Fire in the Control and Spread of Invasive Species, 27 November–1 December 2000, San Diego, CA; Misc. Pub. No. 11; Galley, K.E.M., Wilson, T.P., Eds.; Tall Timbers Research Station: Tallahassee, FL, USA, 2001; pp. 15–30. [Google Scholar]
  18. Fritts, H.C. Relationships of ring widths in arid- site conifers to variations in monthly temperature and precipitation. Ecol. Monogr. 1974, 44, 411–440. [Google Scholar] [CrossRef]
  19. Miller, R.F.; Knick, S.T.; Pyke, D.A.; Meinke, C.W.; Hanser, S.E.; Wisdom, M.J.; Hild, A.L. Characteristics of sagebrush habitats and limitations to long-term conservation. In Greater Sage-Grouse: Ecology and Conservation of a Landscape Species and its Habitats; Knick, S.T., Connelly, J.W., Eds.; University of California Press: Berkeley, CA, USA, 2011; Volume 38, pp. 145–184. [Google Scholar]
  20. Cottam, W.P.; Stewart, G. Plant succession as a result of grazing and of meadow dessication by erosion since settlement in 1862. J. For. 1940, 38, 613–626. [Google Scholar]
  21. Ko, D.W.; Sparrow, A.D.; Weisberg, P.J. Land-use legacy of historical tree harvesting for charcoal production in a semi-arid woodland. For. Ecol. Manag. 2011, 261, 1283–1292. [Google Scholar] [CrossRef]
  22. West, N.E. Intermountain deserts, shrub steppes, and woodlands. In North American Terrestrial Vegetation; Barbour, M.G., Billings, W.D., Eds.; Cambridge University Press: New York, NY, USA, 1988; pp. 209–230. [Google Scholar]
  23. Mitchell, J.E.; Roberts, T.C. Distribution of pinyon-juniper in the western United States. In Proceedings: Ecology and Management of Pinyon-Juniper Communities within the Interior West, 15–18 September 1997, Provo, UT; RMRS-P-9; Monsen, S.B., Stevens, R., Eds.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Ogden, UT, USA, 1999; pp. 146–154. [Google Scholar]
  24. Tueller, P.T.; Beeson, C.D.; Tausch, R.J.; West, N.E.; Rea, K.H. Pinyon-Juniper Woodlands of the Great Basin: Distribution, Flora, Vegetatal Cover; Res. Rep. INT-229; U.S. Department of Agriculture, Forest Service, Intermountain Forest and Range Experiment Station: Ogden, UT, USA, 1979; p. 22.
  25. Evans, R.A. Management of Pinyon-Juniper Woodlands; Gen. Tech. Rep. INT-249; U.S. Department of Agriculture, Forest Service, Intermountain Research Station: Ogden, UT, USA, 1988; p. 34.
  26. (GLTI). Inventorying, Classifying, and Correlating Juniper and Pinyon Communities to Soils in Western United States; U.S. Department of Agriculture, Natural Resources Conservation Service, Grazing Lands Technology Institute: Fort Worth, TX, USA, 1997; p. 39.
  27. Krueger, W.C.; Winward, A.H. Influence of cattle and big game grazing on understory structure of a douglas fir-ponderosa pine-kentucky bluegrass community. J. Range Manag. 1974, 27, 450–453. [Google Scholar] [CrossRef]
  28. Weisberg, P.J.; Lingua, E.; Pillai, R.B. Spatial patterns of pinyon-juniper woodland expansion in central Nevada. Rangel. Ecol. Manag. 2007, 60, 115–124. [Google Scholar] [CrossRef]
  29. Page, D.; Gottfried, G.; Tausch, R.; Lanner, R.; Ritter, S. Management of Pinyon-Juniper “Woodland” Ecosystems. A Position of the Intermountain Society of American Foresters. 2013. Available online: http://www.usu.edu/saf/PJWoodlandsPositionStatement.pdf (accessed on 27 August 2020).
  30. Tausch, R.J.; West, N.E.; Nabi, A.A. Tree age and dominance patterns in Great-Basin pinyon-juniper woodlands. J. Range Manag. 1981, 34, 259–264. [Google Scholar] [CrossRef]
  31. Eldridge, D.J.; Bowker, M.A.; Maestre, F.T.; Roger, E.; Reynolds, J.F.; Whitford, W.G. Impacts of shrub encroachment on ecosystem structure and functioning: Towards a global synthesis. Ecol. Lett. 2011, 14, 709–722. [Google Scholar] [CrossRef]
  32. Davies, K.W.; Boyd, C.S.; Beck, J.L.; Bates, J.D.; Svejcar, T.J.; Gregg, M.A. Saving the sagebrush sea: An ecosystem conservation plan for big sagebrush plant communities. Biol. Conserv. 2011, 144, 2573–2584. [Google Scholar] [CrossRef]
  33. Miller, R.F.; Naugle, D.E.; Maestas, J.D.; Hagen, C.A.; Hall, G. Special issue: Targeted woodland removal to recover at-risk grouse and their sagebrush-steppe and prairie ecosystems. Rangel. Ecol. Manag. 2017, 70, 1–8. [Google Scholar] [CrossRef]
  34. Petersen, S.L.; Stringham, T.K. Intercanopy community structure across a heterogeneous landscape in a western juniper-encroached ecosystem. J. Veg. Sci. 2009, 20, 1163–1175. [Google Scholar] [CrossRef]
  35. Burkhardt, J.W.; Tisdale, E.W. Nature and successional status of western juniper in Idaho. J. Range Manag. 1969, 22, 264–270. [Google Scholar] [CrossRef]
  36. West, N.E. Basic synecological relationshiops of sagebrush-dominated lands in the Great Basin and the Colorado Plateau. In The Sagebrush Ecosystem: A Symposium; Utah State University, College of Natural Resources: Logan, UT, USA, 1979; pp. 33–41. [Google Scholar]
  37. Tausch, R.J.; Tueller, P.T. Foliage biomass and cover relationships between tree-dominated and shrub-dominated communities in pinyon-juniper woodlands. Great Basin Nat. 1990, 50, 121–134. [Google Scholar]
  38. Tausch, R.J.; West, N.E. Plant species composition patterns with differnces in tree dominance on a southwestern Utah pinyon-juniper site. In Desired Future Conditions for Pinyon-Juniper Ecosystems, 8–12 August 1994, Flagstaff, AZ; Gen. Tech. Rep. RM-258; Shaw, D.W., Aldon, E.F., LoSapio, C., Eds.; U.S. Department of Agriculture, Rocky Mountain Forest and Range Experimental Station: Fort Collins, CO, USA, 1995; pp. 16–23. [Google Scholar]
  39. Roundy, B.A.; Miller, R.F.; Tausch, R.J.; Young, K.; Hulet, A.; Rau, B.; Jessop, B.; Chambers, J.C.; Eggett, D. Understory cover responses to piñon–juniper treatments across tree dominance gradients in the Great Basin. Rangel. Ecol. Manag. 2014, 67, 482–494. [Google Scholar] [CrossRef]
  40. O’Meara, T.E.; Haufler, J.B.; Stelter, L.H.; Nagy, J.G. Nongame wildlife responses to chaining of pinyon-juniper woodlands. J. Wildl. Manag. 1981, 45, 381–389. [Google Scholar] [CrossRef]
  41. Rau, B.M.; Tausch, R.; Reiner, A.; Johnson, D.W.; Chambers, J.C.; Blank, R.R. Developing a model framework for predicting effects of woody expansion and fire on ecosystem carbon and nitrogen in a pinyon-juniper woodland. J. Arid Environ. 2012, 76, 97–104. [Google Scholar] [CrossRef]
  42. Short, H.L.; Evans, W.; Boeker, E.L. The use of natural and modified pinyon pine Juniper woodlands by deer and elk. J. Wildl. Manag. 1977, 41, 543–559. [Google Scholar] [CrossRef]
  43. Bristow, N.A.; Weisberg, P.J.; Tausch, R.J. A 40-year record of tree establishment following chaining and prescribed fire treatments in singleleaf pinyon (Pinus monophylla) and Utah juniper (Juniperus osteosperma) Woodlands. Rangel. Ecol. Manag. 2014, 67, 389–396. [Google Scholar] [CrossRef]
  44. Bates, J.D.; Davies, K.W. Effects of conifer treatments on soil nutrient availability and plant composition in sagebrush steppe. For. Ecol. Manag. 2017, 400, 631–644. [Google Scholar] [CrossRef]
  45. Jameson, D.A. The relationship of tree overstory and herbaceous understory vegetation. J. Range Manag. 1967, 20, 247–249. [Google Scholar] [CrossRef]
  46. Barger, N.N.; Archer, S.R.; Campbell, J.L.; Huang, C.Y.; Morton, J.A.; Knapp, A.K. Woody plant proliferation in North American drylands: A synthesis of impacts on ecosystem carbon balance. J. Geophys. Res. Biogeosci. 2011, 116. [Google Scholar] [CrossRef]
  47. D’Antonio, C.M.; Vitousek, P.M. Biological invasions by exotic grasses, the grass/fire cycle, and global change. Ann. Rev. Ecol. Syst. 1992, 23, 63–87. [Google Scholar] [CrossRef]
  48. Roundy, B.A.; Farmer, M.; Olson, J.; Petersen, S.; Nelson, D.R.; Davis, J.; Vernon, J. Runoff and sediment response to tree control and seeding on a high soil erosion potential site in Utah: Evidence for reversal of an abiotic threshold. Ecohydrology 2017, 10, 1–9. [Google Scholar] [CrossRef]
  49. Miller, R.F.; Svejcar, T.J.; Rose, J.A. Impacts of western juniper on plant community composition and structure. J. Range Manag. 2000, 53, 574–585. [Google Scholar] [CrossRef]
  50. Pierson, F.B.; Williams, C.J.; Kormos, P.R.; Hardegree, S.P.; Clark, P.E.; Rau, B.M. Hydrologic vulnerability of sagebrush steppe following pinyon and juniper encroachment. Rangel. Ecol. Manag. 2010, 63, 614–629. [Google Scholar] [CrossRef]
  51. Huxman, T.E.; Wilcox, B.P.; Breshears, D.D.; Scott, R.L.; Snyder, K.A.; Small, E.E.; Hultine, K.; Pockman, W.T.; Jackson, R.B. Ecohydrological implications of woody plant encroachment. Ecology 2005, 86, 308–319. [Google Scholar] [CrossRef]
  52. Williams, C.J.; Pierson, F.B.; Al-Hamdan, O.Z.; Kormos, P.R.; Hardegree, S.P.; Clark, P.E. Can wildfire serve as an ecohydrologic threshold-reversal mechanism on juniper-encroached shrublands? Ecohydrology 2014, 7, 453–477. [Google Scholar] [CrossRef]
  53. Madsen, M.D.; Zvirzdin, D.L.; Petersen, S.L.; Hopkins, B.G.; Roundy, B.A.; Chandler, D.G. Soil water repellency within a burned pinon-juniper woodland: Spatial distribution, severity, and ecohydrologic implications. Soil Sci. Soc. Am. J. 2011, 75, 1543–1553. [Google Scholar] [CrossRef]
  54. Poulsen, C.L.; Walker, S.C.; Stevens, R. Soil seed banking in pinyon-juniper areas with differing levels of tree cover, understory density and composition. In Proceedings: Ecology and Management of Pinyon-Juniper Communities within the Interior West, 15–18 September 1997, Provo, UT; Proc. RMRS-P-9; Monsen, S.B., Stevens, R., Eds.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Ogden, UT, USA, 1999; pp. 141–145. [Google Scholar]
  55. Naillon, D.; Memmott, K.; Monsen, S.B. A comparison of understory species at three densities in a pinyon-juniper woodland. In Proceedings: Ecology and Management of Pinyon-Juniper Communities within the Interior West, 15–18 September 1997, Provo, UT; Proc. RMRS-P-9; Monsen, S.B., Stevens, R., Eds.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Ogden, Utah, USA, 1999; pp. 72–75. [Google Scholar]
  56. Koniak, S.; Everett, R.L. Seed reserves in soils of successional stages of pinyon woodlands. Am. Midl. Nat. 1982, 108, 295–303. [Google Scholar] [CrossRef]
  57. Allen, E.A.; Nowak, R.S. Effect of pinyon–juniper tree cover on the soil seed bank. Rangel. Ecol. Manag. 2008, 61, 63–73. [Google Scholar] [CrossRef]
  58. Suminski, R.R. Management implications for mule dear winter range in northern pinyon-juniper. In Managing Pinyon-Juniper Ecosystems for Sustainability and Social Need, 26–30 April 1993, Sante Fe, NM; Gen Tech. Rep. RM-236; Aldon, E.F., Shaw, D.W., Eds.; U.S. Department of Agriculture, Forest Service, Rocky Moutain Forest and Range Experiment Staion: Fort Collins, CO, USA, 1993; pp. 133–139. [Google Scholar]
  59. Commons, M.L.; Baydack, R.K.; Braun, C.E. Sage grouse response to pinyon-juniper management. In Proceedings: Ecology and Management of Pinyon-Juniper Communities in the Interior West, 15–18 September 1997, Provo, UT; RMRS-P-9; Monsen, S.B., Stevens, R., Eds.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Ogden, UT, USA, 1999; pp. 238–239. [Google Scholar]
  60. Frey, S.N.; Curtis, R.; Heaton, K. Response of a small population of greater sage-grouse to tree removal: Implications of limiting factors. Hum. Wildl. Interact. 2013, 7, 260–272. [Google Scholar]
  61. Baruch-Mordo, S.; Evans, J.S.; Severson, J.P.; Naugle, D.E.; Maestas, J.D.; Kiesecker, J.M.; Falkowski, M.J.; Hagen, C.A.; Reese, K.P. Saving sage-grouse from the trees: A proactive solution to reducing a key threat to a candidate species. Biol. Conserv. 2013, 167, 233–241. [Google Scholar] [CrossRef]
  62. Hagen, C.A. Greater Sage-Grouse Conservation Assessment and Strategy for Oregon: A Plan To Maintain and Enhance Populations of Habitat; Oregon Department of Fish and Wildlife: Salem, OR, USA, 2011; p. 207. [Google Scholar]
  63. Knick, S.T.; Hanser, S.E.; Leu, M. Ecological scale of bird community response to piñon-juniper removal. Rangel. Ecol. Manag. 2014, 67, 553–562. [Google Scholar] [CrossRef]
  64. Crawford, J.A.; Olson, R.A.; West, N.E.; Mosley, J.C.; Schroeder, M.A.; Whitson, T.D.; Miller, R.F.; Gregg, M.A.; Boyd, C.S. Ecology and management of sage-grouse and sage-grouse habitat. J. Range Manag. 2004, 57, 2–19. [Google Scholar] [CrossRef]
  65. Tausch, R.J. Transitions and thresholds: Influences and implications for management in pinyon and juniper woodlands. In Proceedings: Ecology and Management of Pinyon-Juniper Communities within the Interior West, 15–18 September 1997, Provo, UT; RMRS-P-9; Monsen, S.B., Stevens, R., Eds.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Ogden, UT, USA, 1999; pp. 361–365. [Google Scholar]
  66. Davies, K.W.; Bates, J.D.; Boyd, C.S. Postwildfire seeding to restore native vegetation and limit exotic annuals: An evaluation in juniper-dominated sagebrush steppe. Restor. Ecol. 2019, 27, 120–127. [Google Scholar] [CrossRef]
  67. Miller, R.F.; Ratchford, J.; Roundy, B.A.; Tausch, R.J.; Hulet, A.; Chambers, J. Response of conifer-encroached shrublands in the Great Basin to prescribed fire and mechanical treatments. Rangel. Ecol. Manag. 2014, 67, 468–481. [Google Scholar] [CrossRef]
  68. Farzan, S.; Young, D.J.N.; Dedrick, A.G.; Hamilton, M.; Porse, E.C.; Coates, P.S.; Sampson, G. Western juniper management: Assessing strategies for improving greater sage-grouse habitat and rangeland productivity. Environ. Manag. 2015, 56, 675–683. [Google Scholar] [CrossRef]
  69. Gifford, G.F.; Shaw, C.B. Soil moisture patterns on two chained pinyon-juniper sites in Utah. J. Range Manag. 1973, 26, 436–440. [Google Scholar] [CrossRef]
  70. Everett, R.L.; Sharrow, S.H. Soil Water and Temperature in Harvested and Nonharvested Pinyon-Juniper Stands; Res. Paper INT-342; U.S. Department of Agriculture, Forest Service, Intermountain Research Station: Ogden, UT, USA, 1985; p. 5.
  71. Bates, J.D.; Miller, R.F.; Svejcar, T.J. Understory dynamics in cut and uncut western juniper woodlands. J. Range Manag. 2000, 53, 119–126. [Google Scholar] [CrossRef]
  72. Roundy, B.A.; Young, K.; Cline, N.; Hulet, A.; Miller, R.F.; Tausch, R.J.; Chambers, J.C.; Rau, B. Piñon-juniper reduction increases soil water availability of the resource growth pool. Rangel. Ecol. Manag. 2014, 67, 495–505. [Google Scholar] [CrossRef]
  73. Young, K.R.; Roundy, B.A.; Eggett, D.L. Tree reduction and debris from mastication of Utah juniper alter the soil climate in sagebrush steppe. For. Ecol. Manag. 2013, 310, 777–785. [Google Scholar] [CrossRef]
  74. Tennesen, M. When juniper and woody plants invade, water may retreat. Science 2008, 322, 1630–1631. [Google Scholar] [CrossRef] [PubMed]
  75. Pyke, D.A.; Chambers, J.C.; Pellant, M.; Knick, S.T.; Miller, R.F.; Beck, J.L.; Doescher, P.S.; Schupp, E.W.; Roundy, B.A.; Brunson, M.; et al. Restoration Handbook for Sagebrush Steppe Ecosystems with Emphasis on Greater Sage-Grouse Habitat—Part. 1. Concepts for Understanding and Applying Restoration; Circular 1416; U.S. Department of the Interior, U.S. Geological Survey: Reston, VA, USA, 2015; p. 44. [CrossRef]
  76. Williams, C.J.; Pierson, F.B.; Kormos, P.R.; Al-Hamdan, O.Z.; Nouwakpo, S.K.; Weltz, M.A. Vegetation, hydrologic, and erosion responses of sagebrush steppe 9 yr following mechanical tree removal. Rangel. Ecol. Manag. 2019, 72, 47–68. [Google Scholar] [CrossRef]
  77. Fulbright, T.E.; Davies, K.W.; Archer, S.R. Wildlife responses to brush management: A contemporary evaluation. Rangel. Ecol. Manag. 2018, 71, 35–44. [Google Scholar] [CrossRef]
  78. Archer, S.R.; Davies, K.W.; Fulbright, T.E.; McDaniel, K.C.; Wilcox, B.P.; Predick, K.I. Brush management as a rangeland conservation strategy: A critical evaluation. In Conservation Benefits of Rangeland Practices: Assessment, Recommendations, and Knowledge Gaps; Briske, D.D., Ed.; Allen Press: Lawrence, KS, USA, 2011; pp. 105–170. [Google Scholar]
  79. Provencher, L.; Thompson, J. Vegetation responses to pinyon-juniper treatments in eastern Nevada. Rangel. Ecol. Manag. 2014, 67, 195–205. [Google Scholar] [CrossRef]
  80. Gottfried, G.J.; Severson, K.E. Managing pinyon-juniper woodlands. Rangelands 1994, 16, 234–236. [Google Scholar]
  81. Roccaforte, J.P.; Fulé, P.Z.; Covington, W.W. Monitoring landscape-scale ponderosa pine restoration treatment implementation and effectiveness. Restor. Ecol. 2010, 18, 820–833. [Google Scholar] [CrossRef]
  82. McIver, J.; Brunson, M. Multidisciplinary, multisite evaluation of alternative sagebrush steppe restoration treatments: The SageSTEP Project. Rangel. Ecol. Manag. 2014, 67, 435–439. [Google Scholar] [CrossRef]
  83. Copeland, S.M.; Munson, S.M.; Bradford, J.B.; Butterfield, B.J.; Morgan, J. Influence of climate, post-treatment weather extremes, and soil factors on vegetation recovery after restoration treatments in the southwestern US. Appl. Veg. Sci. 2019, 22, 85–95. [Google Scholar] [CrossRef]
  84. Benayas, J.M.R.; Newton, A.C.; Diaz, A.; Bullock, J.M. Enhancement of biodiversity and ecosystem services by ecological restoration: A meta-analysis. Science 2009, 325, 1121–1124. [Google Scholar] [CrossRef] [PubMed]
  85. Gurevitch, J.; Koricheva, J.; Nakagawa, S.; Stewart, G. Meta-analysis and the science of research synthesis. Nature 2018, 555, 175–182. [Google Scholar] [CrossRef] [PubMed]
  86. Gomez-Aparicio, L.; Lortie, C.J. Advancing plant ecology through meta-analyses. J. Ecol. 2014, 102, 823–827. [Google Scholar] [CrossRef]
  87. Willms, J.; Bartuszevige, A.; Schwilk, D.W.; Kennedy, P.L. The effects of thinning and burning on understory vegetation in North America: A meta-analysis. For. Ecol. Manag. 2017, 392, 184–194. [Google Scholar] [CrossRef]
  88. Shindler, B.; Gordon, R.; Brunson, M.W.; Olsen, C. Public perceptions of sagebrush ecosystem management in the Great Basin. Rangel. Ecol. Manag. 2011, 64, 335–343. [Google Scholar] [CrossRef]
  89. Redmond, M.D.; Zelikova, T.J.; Barger, N.N. Limits to understory plant restoration following fuel-reduction treatments in a pinon-juniper woodland. Environ. Manag. 2014, 54, 1139–1152. [Google Scholar] [CrossRef]
  90. West, N.E. Factors affecting treatment success in the pinyon-juniper type. In Proceedings: Second Utah Shrub Ecology Workshop; Johnson, K.L., Ed.; Utah State University: Logan, UT, USA, 1984; pp. 21–33. [Google Scholar]
  91. Williams, R.E.; Roundy, B.A.; Hulet, A.; Miller, R.F.; Tausch, R.J.; Chambers, J.C.; Matthews, J.; Schooley, R.; Eggett, D. Pretreatment tree dominance and conifer removal treatments affect plant succession in sagebrush communities. Rangel. Ecol. Manag. 2017, 70, 759–773. [Google Scholar] [CrossRef]
  92. Miller, R.F.; Chambers, J.C.; Pellant, M. A Field Guide for Selecting the Most Appropriate Treatment in Sagebrush and Piñon-Juniper Ecosystems in the Great Basin; Gen. Tech. Rep. RMRS-GTR-322; Station, R.M.R., Ed.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Fort Collins, CO, USA, 2014; p. 72.
  93. House, J.I.; Archer, S.; Breshears, D.D.; Scholes, R.J. Conundrums in mixed woody-herbaceous plant systems. J. Biogeogr. 2003, 30, 1763–1777. [Google Scholar] [CrossRef]
  94. Roundy, B.A.; Chambers, J.C.; Pyke, D.A.; Miller, R.F.; Tausch, F.J.; Schupp, E.W.; Rau, B.; Gruell, T. Resilience and resistance in sagebrush ecosystems are associated with seasonal soil temperature and water availability. Ecosphere 2018, 9, e02417. [Google Scholar] [CrossRef]
  95. Stevens, R. Restoration of native communities by chaining and seeding. In Proceedings: Ecology and Managment of Pinyon-Juniper Communities within the Interior West, 15–18 September 1997, Provo, UT; RMRS-P-9; Monsen, S.B., Stevens, R., Eds.; U.S. Department of Agriculture, Rocky Mountain Research Station: Ogden, UT, USA, 1999; pp. 285–289. [Google Scholar]
  96. Baughman, C.; Forbis, T.A.; Provencher, L. Response of two sagebrush sites to low-disturbance, mechanical removal of pinyon and juniper. Invasive Plant Sci. Manag. 2010, 3, 122–129. [Google Scholar] [CrossRef]
  97. Rosenstock, S.S.; Van Riper, C. Breeding bird responses to juniper woodland expansion. J. Range Manag. 2001, 54, 226–232. [Google Scholar] [CrossRef]
  98. Bybee, J.; Roundy, B.A.; Young, K.R.; Hulet, A.; Roundy, D.B.; Crook, L.; Aanderud, Z.; Eggett, D.L.; Cline, N.L. Vegetation response to piñon and juniper tree shredding. Rangel. Ecol. Manag. 2016, 69, 224–234. [Google Scholar] [CrossRef]
  99. Hessing, M.B.; Johnson, C.D.; Balda, R.P. Early secondary succession of a pinyon-juniper woodland in a northern Arizona powerline corridor. Southwest Nat. 1982, 27, 1–9. [Google Scholar] [CrossRef]
  100. Schott, M.R.; Pieper, R.D. Succssion in pinyon-juniper vegetation in New Mexico. Rangelands 1986, 8, 126–128. [Google Scholar]
  101. Urza, A.K.; Weisberg, P.J.; Chambers, J.C.; Dhaemers, J.M.; Board, D. Post-fire vegetation response at the woodland-shrubland interface is mediated by the pre-fire community. Ecosphere 2017, 8, e01851. [Google Scholar] [CrossRef]
  102. Bernau, C.R.; Strand, E.K.; Bunting, S.C. Fuel bed response to vegetation treatments in juniper-invaded sagebrush steppe. Fire Ecol. 2018, 14. [Google Scholar] [CrossRef]
  103. Huffman, D.W.; Stoddard, M.T.; Springer, J.D.; Crouse, J.E. Understory responses to tree thinning and seeding indicate stability of degraded pinyon-juniper woodlands. Rangel. Ecol. Manag. 2017, 70, 484–492. [Google Scholar] [CrossRef]
  104. Davenport, D.W.; Breshears, D.D.; Wilcox, B.P.; Allen, C.D. Viewpoint: Sustainability of pinon-juniper ecosystems—A unifying perspective of soil erosion thresholds. J. Range Manag. 1998, 51, 231–240. [Google Scholar] [CrossRef]
  105. Clary, W.P.; Jameson, D.A. Herbage production following tree and shrub removal in the pinyon-juniper type of Arizona. J. Range Manag. 1981, 34, 109–113. [Google Scholar] [CrossRef]
  106. Knoop, W.T.; Walker, B.H. Interactions of woody and herbaceous vegetation in a southern African savanna. J. Ecol. 1985, 73, 235–253. [Google Scholar] [CrossRef]
  107. Riginos, C. Grass competition suppresses tree growth across multiple demographic stages. Ecology 2009, 90, 335–340. [Google Scholar] [CrossRef] [PubMed]
  108. Riginos, C.; Grace, J.B.; Augustine, D.J.; Young, T.P. Local versus landscape-scale effects of savanna trees on grasses. J. Ecol. 2009, 97, 1337–1345. [Google Scholar] [CrossRef]
  109. Chambers, J.C.; Maestas, J.D.; Pyke, D.A.; Boyd, C.S.; Pellant, M.; Wuenschel, A. Using resilience and resistance concepts to manage persistent threats to sagebrush ecosystems and greater sage-grouse. Rangel. Ecol. Manag. 2017, 70, 149–164. [Google Scholar] [CrossRef]
  110. Barney, M.A.; Frischknecht, N.C. Vegetation changes following fire in the pinyon-juniper type of west-central Utah. J. Range Manag. 1974, 27, 91–96. [Google Scholar] [CrossRef]
  111. Kerns, B.K.; Day, M.A. The importance of disturbance by fire and other abiotic and biotic factors in driving cheatgrass invasion varies based on invasion stage. Biol. Invas. 2017, 19, 1853–1862. [Google Scholar] [CrossRef]
  112. Chambers, J.C.; Roundy, B.A.; Blank, R.R.; Meyer, S.E.; Whittaker, A. What makes Great Basin sagebrush ecosystems invasible by Bromus tectorum? Ecol. Monogr. 2007, 77, 117–145. [Google Scholar] [CrossRef]
  113. Everett, R.L.; Sharrow, S.H. Understory response to tree harvesting of singleleaf pinyon and Utah juniper. Great Basin Nat. 1985, 45, 105–112. [Google Scholar]
  114. Bates, J.D.; Sharp, R.N.; Davies, K.W. Sagebrush steppe recovery after fire varies by development phase of Juniperus occidentalis woodland. Int. J. Wildl. Fire 2014, 23, 117. [Google Scholar] [CrossRef]
  115. Kerns, B.K.; Day, M.A. Fuel reduction, seeding, and vegetation in a juniper woodland. Rangel. Ecol. Manag. 2014, 67, 667–679. [Google Scholar] [CrossRef]
  116. Davies, K.W.; Rios, R.C.; Bates, J.D.; Johnson, D.D.; Kerby, J.; Boyd, C.S. To burn or not to burn: Comparing reintroducing fire with cutting an encroaching conifer for conservation of an imperiled shrub-steppe. Ecol. Evol. 2019, 9, 9137–9148. [Google Scholar] [CrossRef]
  117. Miller, R.F.; Bates, J.D.; Svejcar, T.J.; Pierson, F.B.; Eddleman, L.E. Biology, Ecology, and Management of Western Juniper; Tech. Bull. 152; Oregon State University, Agricultural Experiment Station, Ed.; Oregon State University, Agricultural Experiment Station: Corvallis, OR, USA, 2005; p. 7. [Google Scholar]
  118. Skousen, J.G.; Davis, J.N.; Brotherson, J.D. Pinyon-juniper chaining and seeding for big game in central Utah. J. Range Manag. 1989, 42, 98–104. [Google Scholar] [CrossRef]
  119. Little, E.L. Managing southwestern piñon-juniper woodlands: The past half centrury and the future. In Managing Piñon-Juniper Ecosystems for Sustainability and Social Needs, 26–30 April 1993, Sante Fe, NM; Gen. Tech. Rep. RM-236; Aldon, E.F., Shaw, D.W., Eds.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station: Fort Collins, CO, USA, 1993; pp. 105–107. [Google Scholar]
  120. Tausch, R.J.; Miller, R.F.; Roundy, B.A.; Chambers, J.C. Piñon and Juniper Field Guide: Asking the Right Questions to Select Appropriate Management Actions; Circular 1335; U.S. Department of the Interior, U.S. Geological Survey: Reston, VA, USA, 2009; p. 96.
  121. Fairchild, J.A. Pinyon-juniper chaining design guidelines for big game wither range enhancement projects. In Proceedings: Ecology and Management of Pinyon-Juniper Communities within the Interior West, 15–18 September 1997, Provo, UT; RMRS-P-9; Monsen, S.B., Stevens, R., Eds.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Ogden, UT, USA, 1999; pp. 278–280. [Google Scholar]
  122. Crow, C.; Riper, C.V. Avian community responses to mechanical thinning of a pinyon-juniper woodland: Specialist sensitivity to tree reduction. Nat. Areas J. 2010, 30, 191–201. [Google Scholar] [CrossRef]
  123. Koniak, S. Broadcast Seeding Success in Eight Pinyon-Juniper Stands after Wildfire; Res. Note INT-334; U.S. Department of Agriculture, Forest Service, Intermountain Forest and Range Experiment Station: Ogden, UT, USA, 1983; p. 4.
  124. Huffman, D.W.; Stoddard, M.T.; Springer, J.D.; Crouse, J.E.; Chancellor, W.W. Understory plant community responses to hazardous fuels reduction treatments in pinyon-juniper woodlands of Arizona, USA. For. Ecol. Manag. 2013, 289, 478–488. [Google Scholar] [CrossRef]
  125. Ott, J.E.; McArthur, E.D.; Roundy, B.A. Vegetation of chained and non-chained seedings after wildfire in Utah. J. Range Manag. 2003, 56, 81–91. [Google Scholar] [CrossRef]
  126. Stevens, R. Thirty years of pinyon-juniper big game habitat improvement projects: What have we learned? In Proceedings—Pinyon-Juniper Conference; Gen. Tech. Rep. INT-215; Everett, R.L., Ed.; U.S. Department of Agriculture, Forest Service, Intermountain Research Station: Ogden, UT, USA, 1987; pp. 558–571. [Google Scholar]
  127. Stevens, R. Mechanical chaining and seeding. In Proceedings of the Ecology and Managment of Pinyon-Juniper Communities within the Interior West, Provo, UT, USA, 15–18 September 1997; Proc. RMRS-P-9. pp. 281–284. [Google Scholar]
  128. Daniel, T.W.; Rivers, R.J.; Isaacson, H.E.; Eberhard, E.J.; LeBaron, A.D. Management Alternatives for Pinyon-Juniper Woodlands. Part. A: The Ecology of the Pinyon-Juniper Type of the Colorado Plateau and the Basin and Range Provinces; Bureau of Land Management and Utah Agriculutural Experiment Station: Logan, UT, USA, 1966; p. 242. [Google Scholar]
  129. Tausch, R.J.; Tueller, P.T. Plant succession following chaining of pinyon-juniper woodlands in eastern Nevada. J. Range Manag. 1977, 30, 44–49. [Google Scholar] [CrossRef]
  130. Madsen, M.D.; Zvirzdin, D.L.; Petersen, S.L.; Hopkins, B.G.; Roundy, B.A. Anchor chaining’s influence on soil hydrology and seeding success in burned Pinon-Juniper woodlands. Rangel. Ecol. Manag. 2015, 68, 231–240. [Google Scholar] [CrossRef]
  131. Skousen, J.; Davis, J.N.; Brotherson, J.D. Comparison of vegetation patterns resulting from bulldozing and 2-way chaining on a Utah pinyon-juniper big game range. Great Basin Nat. 1986, 46, 508–512. [Google Scholar]
  132. Young, K.R.; Roundy, B.A.; Eggett, D.L. Plant establishment in masticated Utah juniper woodlands. Rangel. Ecol. Manag. 2013, 66, 597–607. [Google Scholar] [CrossRef]
  133. Juran, C.; Roundy, B.A.; Davis, J.N. Wildfire rehabilitation success with and without chaining on the Henry Mountains, Utah. In Proceedings—Shrublands under Fire: Disturbance and Recovery in a Changing World, 6–8 June 2006, Cedar City, UT; RMRS-P-52; Kitchen, S.G., Pendleton, R.L., Monaco, T.A., Vernon, J., Eds.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Fort Collins, CO, USA, 2008; pp. 91–106. [Google Scholar]
  134. Thompson, T.W.; Roundy, B.A.; McArthur, E.D.; Jessop, B.D.; Waldron, B.; Davis, J.N. Fire rehabilitation using native and introduced species: A landscape trial. Rangel Ecol. Manag. 2006, 59, 237–248. [Google Scholar] [CrossRef]
  135. Clary, W.P. Plant Density and Cover Responses to Several Seeding Techniques Following Wildfire; Res. Note INT-384; U.S. Department of Agriculture, Forest Service, Intermountain Research Station: Ogden, UT, USA, 1988; p. 6.
  136. Stoddard, M.T.; Huffman, D.W.; Alcoze, T.M.; Fule, P.Z. Effects of slash on herbaceous communities in pinyon-juniper woodlands of northern Arizona. Rangel. Ecol. Manag. 2008, 61, 485–495. [Google Scholar] [CrossRef]
  137. Cline, N.L.; Roundy, B.A.; Pierson, F.B.; Kormos, P.; Williams, C.J. Hydrologic response to mechanical shredding in a juniper woodland. Rangel. Ecol. Manag. 2010, 63, 467–477. [Google Scholar] [CrossRef]
  138. Pierson, F.B.; Williams, C.J.; Kormos, P.R.; Al-Hamdan, O.Z. Short-term effects of tree removal on infiltration, runoff, and erosion in woodland-encroached sagebrush steppe. Rangel. Ecol. Manag. 2014, 67, 522–538. [Google Scholar] [CrossRef]
  139. Brockway, D.G.; Gatewood, R.G.; Paris, R.B. Restoring grassland savannas from degraded pinyon-juniper woodlands: Effects of mechanical overstory reduction and slash treatment alternatives. J. Environ. Manag. 2002, 64, 179–197. [Google Scholar] [CrossRef]
  140. Reiner, A.L.; Vaillant, N.M.; Fites-Kaufman, J.; Dailey, S.N. Mastication and prescribed fire impacts on fuels in a 25-year old ponderosa pine plantation, southern Sierra Nevada. For. Ecol. Manag. 2009, 258, 2365–2372. [Google Scholar] [CrossRef]
  141. Faist, A.; Stone, H.; Tripp, E. Impacts of mastication: Soil seed bank responses to a forest thinning treatment in three Colorado (USA) conifer forest types. Forests 2015, 6, 3060–3074. [Google Scholar] [CrossRef]
  142. Ross, M.R.; Castle, S.C.; Barger, N.N. Effects of fuels reductions on plant communities and soils in a piñon-juniper woodland. J. Arid Environ. 2012, 79, 84–92. [Google Scholar] [CrossRef]
  143. Miller, R.F.; Chambers, J.C.; Pyke, D.A.; Pierson, F.B.; Williams, J.C. 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; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Fort Collins, CO, USA, 2013; p. 126.
  144. Murphy, G.E.P.; Romanuk, T.N. A meta-analysis of community response predictability to anthropogenic disturbances. Am. Nat. 2012, 180, 316–327. [Google Scholar] [CrossRef]
  145. Lavin, M.; Brummer, T.J.; Quire, R.; Maxwell, B.D.; Rew, L.J. Physical disturbance shapes vascular plant diversity more profoundly than fire in the sagebrush steppe of southeastern Idaho, USA. Ecol. Evol. 2013, 3, 1626–1641. [Google Scholar] [CrossRef]
  146. Pieper, R.D. Spatial variation of piñon-juniper woodlands in New Mexico. In Managing Pinyon-Juniper Ecosystems for Sustainability and Social Needs, 26–30 April 1993, Sante Fe, NM; Gen Tech. Rep. RM-236; Aldon, E.F., Shaw, D.W., Eds.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station: Fort Collins, CO, USA, 1993; pp. 89–92. [Google Scholar]
  147. West, N.E.; Tausch, R.J.; Tueller, P.T. A Management-Oriented Classification of Pinyon-Juniper Woodlands of the Great Basin; RMRS-GTR-12; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Ogden, UT, USA, 1998; p. 42.
  148. Leonard, S.G.; Miles, R.L.; Summerfield, H.A. Soils of the pinyon-juniper woodlands. In Proceedings—Pinyon-Juniper Conference, 13–16 January 1986, Reno, NV; Gen. Tech. Rep. INT-215; Everett, R.L., Ed.; U.S. Department of Agricuture, Forest Service, Intermountain Research Station: Ogden, UT, USA, 1987; pp. 227–230. [Google Scholar]
  149. Evangelista, P.; Stohlgren, T.J.; Guenther, D.; Stewart, S. Vegetation response to fire and postburn seeding treatments in Juniper woodlands of the Grand Staircase-Escalante National Monument, Utah. West. N. Am. Nat. 2004, 64, 293–305. [Google Scholar]
  150. Pyke, D.A.; Shaff, S.E.; Lindgren, A.I.; Schupp, E.W.; Doescher, P.S.; Chambers, J.C.; Burnham, J.S.; Huso, M.M. Region-wide ecological responses of arid Wyoming big sagebrush communities to fuel treatments. Rangel. Ecol. Manag. 2014, 67, 455–467. [Google Scholar] [CrossRef]
  151. Knutson, K.C.; Pyke, D.A.; Wirth, T.A.; Arkle, R.S.; Pilliod, D.S.; Brooks, M.L.; Chambers, J.C.; Grace, J.B. Long-term effects of seeding after wildfire on vegetation in Great Basin shrubland ecosystems. J. Appl Ecol. 2014, 51, 1414–1424. [Google Scholar] [CrossRef]
  152. Germino, M.J.; Barnard, D.M.; Davidson, B.E.; Arkle, R.S.; Pilliod, D.S.; Fisk, M.R.; Applestein, C. Thresholds and hotspots for shrub restoration following a heterogeneous megafire. Landsc. Ecol. 2018, 33, 1177–1194. [Google Scholar] [CrossRef]
  153. West, N.E.; Tausch, R.J.; Rea, K.H.; Tueller, P.T. Taxonomic determination, distribution, and ecological indicator values of sagebrush within pinyon-juniper woodlands of Great Basin. J. Range Manag. 1978, 31, 87–92. [Google Scholar] [CrossRef]
  154. Stevens, R. Species adapted for seeding mountain brush, big, black, and low sagebrush, and pinyon-juniper communities. In Managing Intermountain Rangelands—Improvement of Range and Wildlife Habitats: Proceedings, 15–17 September 1981, Twin Falls, ID, 22–24 June 1982, Elko, NV; Gen. Tech. Rep. INT-157; Monsen, S.B., Shaw, N., Eds.; U.S. Department of Agriculture, Forest Service, Intermountain Forest and Range Experimental Station: Ogden, UT, USA, 1983; pp. 78–82. [Google Scholar]
  155. Moir, W.H.; Carleton, J.O. Classification of pinyon-juniper (P-J) sites on national forests in the southwest. In Proceedings—Pinyon-Juniper Conference, 13–16 January 1986, Reno, NV; Gen. Tech. Rep. INT-215; Everett, R.L., Ed.; U.S. Department of Agricuture, Forest Service, Intermountain Research Station: Ogden, UT, USA, 1987; pp. 216–226. [Google Scholar]
  156. West, N.E.; Rea, K.H.; Tausch, R.J. Basic synecological relationships in juniper-pinyon woodlands. In The Pinyon-juniper Ecosystem: A Symposium; Utah Agricultural Experiment Station: Logan, UT, USA, 1975; pp. 41–54. [Google Scholar]
  157. McArthur, E.D. Ecology, distribution, and values of sagebrush within the Intermountain Region. In Proceedings: Ecology And Management of Annual Rangelands, 18–21 May 1992, Boise, ID; Gen. Tech. Rep. INT-GTR-313; Monsen, S.B., Kitchen, S.G., Eds.; U.S. Department of Agriculture, Forest Service, Intermountain Research Station: Ogden, UT, USA, 1994; pp. 347–351. [Google Scholar]
  158. Winward, A.H. Using sagebrush ecology in wildland management. In Proceedings of the First Utah Shrub Ecology Workshop; Johnson, K.L., Ed.; Utah State University, College of Natural Resources: Logan, UT, USA, 1983; pp. 15–19. [Google Scholar]
  159. West, N.E. (Ed.) Great Basin-Colorado Plateau Sagebrush Semi-Desert. In Temperate Deserts and Semi-Deserts; Elsevier: Amsterdam, The Netherlands, 1983; pp. 331–349. [Google Scholar]
  160. Wilder, L.E.; Veblen, K.E.; Schupp, E.W.; Monaco, T.A. Seedling emergence patterns of six restoration species in soils from two big sagebrush plant communities. West. N. Am. Nat. 2019, 79, 233–246. [Google Scholar] [CrossRef]
  161. Jensen, M.E. Interpretation of environmental gradients which influence sagebrush community distribution in northeastern Nevada. J. Range Manag. 1990, 43, 161–167. [Google Scholar] [CrossRef]
  162. Thatcher, A.P. Distribution of sagebrush as related to site differences in Albany County, Wyoming. J. Range Manag. 1959, 12, 55–61. [Google Scholar] [CrossRef]
  163. Goodrich, S. Classification and capabilities of woody sagebrush communities of western North America with emphasis on sage-grouse habitat. In Sage-Grouse Habitat Restoration Symposium Proceedings, 4–7 June, Boise, ID; RMRS-P-38; Shaw, N.L., Pellant, M., Monsen, S.B., Eds.; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Fort Collins, CO, USA, 2005; pp. 17–37. [Google Scholar]
  164. Shultz, L.M. Monograph of Artemisia subgenus Tridentatae (Asteraceae-Anthemideae). Bot. Monogr. 2009, 89, 1–131. [Google Scholar]
  165. Wilder, L.E.; Veblen, K.E.; Gunnell, K.L.; Monaco, T.A. Influence of fire and mechanical sagebrush reduction treatments on restoration seedings in Utah, United States. Restor. Ecol. 2019, 27, 308–319. [Google Scholar] [CrossRef]
  166. Maestas, J.D.; Campbell, S.B. Mapping Potential Ecosystem Resilience and Resistance Across Sage-Grouse Range Using Soil Temperature and Moisture Regimes. 2016. Available online: http://www.usu.edu/saf/PJWoodlandsPositionStatement.pdf (accessed on 27 August 2020).
  167. Chambers, J.C.; Miller, R.F.; Board, D.I.; Pyke, D.A.; Roundy, B.A.; Grace, J.B.; Schupp, E.W.; Tausch, R.J. Resilience and resistance of sagebrush ecosystems: Implications for state and transition models and management treatments. Rangel. Ecol. Manag. 2014, 67, 440–454. [Google Scholar] [CrossRef]
  168. Riginos, C.; Veblen, K.E.; Thacker, E.T.; Gunnell, K.L.; Monaco, T.A. Disturbance type and sagebrush community type affect plant community structure after shrub reduction. Rangel. Ecol. Manag. 2019, 72, 619–631. [Google Scholar] [CrossRef]
  169. Chambers, J.C.; Bradley, B.A.; Brown, C.S.; D’Antonio, C.; Germino, M.J.; Grace, J.B.; Hardegree, S.P.; Miller, R.F.; Pyke, D.A. Resilience to stress and disturbance, and resistance to Bromus tectorum L. invasion in cold desert shrublands of western North America. Ecosystems 2014, 17, 360–375. [Google Scholar] [CrossRef]
  170. Lembrechts, J.J.; Pauchard, A.; Lenoir, J.; Nunez, M.A.; Geron, C.; Ven, A.; Bravo-Monasterio, P.; Teneb, E.; Nijs, I.; Milbau, A. Disturbance is the key to plant invasions in cold environments. Proc. Natl. Acad. Sci. USA 2016, 113, 14061–14066. [Google Scholar] [CrossRef] [PubMed]
  171. Mitchell, R.M.; Bakker, J.D.; Vincent, J.B.; Davies, G.M. Relative importance of abiotic, biotic, and disturbance drivers of plant community structure in the sagebrush steppe. Ecol. Appl. 2017, 27, 756–768. [Google Scholar] [CrossRef] [PubMed]
  172. Urza, A.K.; Weisberg, P.J.; Chambers, J.C.; Board, D.; Flake, S.W. Seeding native species increases resistance to annual grass invasion following prescribed burning of semiarid woodlands. Biol. Invas. 2019, 21, 1993–2007. [Google Scholar] [CrossRef]
  173. Johnson, D.D.; Miller, R.F. Structure and development of expanding western juniper woodlands as influenced by two topographic variables. For. Ecol. Manag. 2006, 229, 7–15. [Google Scholar] [CrossRef]
  174. Jacobs, B.F.; Romme, W.H.; Allen, C.D. Mapping “old” vs. “young” pinon-juniper stands with a predictive topo-climatic model. Ecol. Appl. 2008, 18, 1627–1641. [Google Scholar] [CrossRef]
  175. Davis, J.N.; Harper, K.T. Weedy annuals and establishment of seeded species on a chained juniper-pinyon woodland in central Utah. In Wildland Shrub Dieoffs Following Excessivey Wet Periods: A Synthesis; McArthur, E.D., Romney, E.M., Smith, S.D., Tueller, P.T., Eds.; U.S. Department of Agriculture, Forest Service, Intermountain Research Station: Ogden, UT, USA, 1990; pp. 72–79. [Google Scholar]
  176. Everett, R.L.; Ward, K. Early plant succession on pinyon-juniper controlled burns. Northwest Sci. 1984, 58, 57–68. [Google Scholar]
  177. James, J.J.; Davies, K.W.; Sheley, R.L.; Aanderud, Z.T. Linking nitrogen partitioning and species abundance to invasion resistance in the Great Basin. Oecologia 2008, 156, 637–648. [Google Scholar] [CrossRef]
  178. Leonard, E.D.; Monaco, T.A.; Stark, J.M.; Ryel, R.J. Invasive forb, annual grass, and exotic shrub competition with three sagebrush-steppe growth forms: Acquisition of a spring 15N tracer. Invasive Plant Sci. Manag. 2008, 1, 168–177. [Google Scholar] [CrossRef]
  179. Deines, L.; Rosentreter, R.; Eldridge, D.J.; Serpe, M.D. Germination and seedling establishment of two annual grasses on lichen-dominated biological soil crusts. Plant Soil 2007, 295, 23–35. [Google Scholar] [CrossRef]
  180. Cain, D. The Ely Chain: A Practical Handbook of Principles and Practices of Chaining and Vegetative Manipulation; U.S. Department of the Interior, Bureau of Land Management: Ely, NV, USA, 1971; p. 32.
  181. McKenzie, D.; Jensen, F.R.; Johnsen, T.N.; Young, J.A. Chains for mechanical brush control. Rangelands 1984, 6, 122–127. [Google Scholar]
  182. Evans, R.A.; Young, J.A. Plant succession following control of western juniper (Juniperus occidentalis) with picloram. Weed Sci. 1985, 33, 63–68. [Google Scholar] [CrossRef]
  183. Havrilla, C.A.; Faist, A.M.; Barger, N.N. Understory plant community responses to fuel-reduction treatments and seeding in an upland piñon-juniper woodland. Rangel. Ecol. Manag. 2017, 70, 609–620. [Google Scholar] [CrossRef]
  184. Stephens, G.J.; Johnston, D.B.; Jonas, J.L.; Paschke, M.W. Understory responses to mechanical treatment of pinyon-juniper in northwestern Colorado. Rangel. Ecol. Manag. 2016, 69, 351–359. [Google Scholar] [CrossRef]
  185. Owen, S.M.; Sieg, C.H.; Gehring, C.A.; Bowker, M.A. Above- and belowground responses to tree thinning depend on the treatment of tree debris. For. Ecol. Manag. 2009, 259, 71–80. [Google Scholar] [CrossRef]
  186. Potts, J.B.; Stephens, S.L. Invasive and native plant responses to shrubland fuel reduction: Comparing prescribed fire, mastication, and treatment season. Biol. Conserv. 2009, 142, 1657–1664. [Google Scholar] [CrossRef]
  187. Coop, J.D.; Grant, T.A.; Magee, P.A.; Moore, E.A. Mastication treatment effects on vegetation and fuels in piñon-juniper woodlands of central Colorado, USA. For. Ecol. Manag. 2017, 396, 68–84. [Google Scholar] [CrossRef]
  188. Rubin, R.L.; Roybal, C.M. Plant community responses to mastication and mulching of one-seed juniper (Juniperus monosperma). Rangel. Ecol. Manag. 2018, 71, 753–756. [Google Scholar] [CrossRef]
  189. Young, K.R.; Roundy, B.R.; Eggett, D.L. Mechanical mastication of Utah juniper encroaching sagebrush steppe increases inorganic soil N. Appl. Environ. Soil Sci. 2014, 2014. [Google Scholar] [CrossRef]
  190. Aanderud, Z.T.; Schoolmaster, D.R.; Rigby, D.; Bybee, J.; Campbell, T.; Roundy, B.A. Soils mediate the impact of fine woody debris on invasive and native grasses as whole trees are mechanically shredded into firebreaks in piñon-juniper woodlands. J. Arid Environ. 2017, 137, 60–68. [Google Scholar] [CrossRef]
  191. Koniak, S. Succession in pinyon-juniper woodlands following wildfire in the Great Basin. Great Basin Nat. 1985, 45, 556–566. [Google Scholar]
  192. Miller, R.F.; Chambers, J.C.; Pellant, M. A Field Guide for Rapid Assessment of Post-Wildfire Recovery Potential in Sagebrush and Pinon-Juniper Ecosystems in the Great Basin; Gen. Tech. Rep. RMRS-GTR-338; U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station: Fort Collins, CO, USA, 2015; p. 70.
  193. Redmond, M.D.; Cobb, N.S.; Miller, M.E.; Barger, N.N. Long-term effects of chaining treatments on vegetation structure in pinon-juniper woodlands of the Colorado Plateau. For. Ecol. Manag. 2013, 305, 120–128. [Google Scholar] [CrossRef]
  194. Van Auken, O.W. Causes and consequences of woody plant encroachment into western North American grasslands. J. Environ. Manag. 2009, 90, 2931–2942. [Google Scholar] [CrossRef] [PubMed]
  195. Caracciolo, D.; Istanbulluoglu, E.; Noto, L.V. An ecohydrological cellular automata model investigation of juniper tree encroachment in a western North American landscape. Ecosystems 2017, 20, 1104–1123. [Google Scholar] [CrossRef]
  196. West, N.E. (Ed.) Overview of North American temperate deserts and semi-deserts. In Temperate Deserts and Semi-Deserts; Elsevier Scientific Publishing Company: Amsterdam, The Netherlands, 1983; pp. 321–330. [Google Scholar]
  197. Proctor, M.C.F.; Tuba, Z. Poikilohydry and homoihydry: Antithesis or spectrum of possibilities? New Phytol. 2002, 156, 327–349. [Google Scholar] [CrossRef]
  198. Hirsch-Schantz, M.C.; Monaco, T.A.; Call, C.A.; Sheley, R.L. Large-scale downy brome treatments alter plant-soil relationships and promote perennial grasses in salt desert shrublands. Rangel. Ecol. Manag. 2014, 67, 255–265. [Google Scholar] [CrossRef]
  199. Compagnoni, A.; Adler, P.B. Warming, soil moisture, and loss of snow increase Bromus tectorum’s population growth rate. Elementa 2014, 2, 20. [Google Scholar] [CrossRef]
  200. Condon, L.A.; Pyke, D.A. Filling the interspace-restoring arid land mosses: Source populations, organic matter, and overwintering govern success. Ecol. Evol. 2016, 6, 7623–7632. [Google Scholar] [CrossRef]
  201. Williams, C.J.; Snyder, K.A.; Pierson, F.B. Spatial and temporal variability of the impacts of pinyon and juniper reduction on hydrologic and rrosion processes across climatic gradients in the western US: A regional synthesis. Water 2018, 10, 1607. [Google Scholar] [CrossRef]
  202. Jacobs, B.F.; Gatewood, R.G. Reintroduction of fire maintains structure of mechanically restored pinyon-juniper savanna (New Mexico). Ecol. Restor. 2002, 20, 207–208. [Google Scholar]
  203. Monaco, T.A.; Mangold, J.M.; Mealor, B.A.; Mealor, R.D.; Brown, C.S. Downy brome control and impacts on perennial grass abundance: A systematic review spanning 64 years. Rangel. Ecol. Manag. 2017, 70, 396–404. [Google Scholar] [CrossRef]
  204. Jacobs, B.F. Restoration of degraded transitional (piñon-juniper) woodland sites improves ecohydrologic condition and primes understory resilience to subsequent disturbance. Ecohydrology 2015, 8, 1417–1428. [Google Scholar] [CrossRef]
  205. Ashcroft, N.K.; Fernald, A.G.; VanLeeuwen, D.M.; Baker, T.T.; Cibils, A.F.; Boren, J.C. The effects of thinning trees and scattering slash on runoff and sediment yield within dense piñon-juniper woodlands in New Mexico, United States. J. Soil Water Conserv. 2017, 72, 122–130. [Google Scholar] [CrossRef]
  206. West, N.E.; Van Pelt, N.S. Successional patterns in pinyon-juniper woodlands. In Proceedings: Pinyon-Juniper Conference. USDA Forest Service, 13–16 January, Reno, NV; Gen. Tech. Rep. INT-215; Everett, R.L., Ed.; U.S. Department of Agriculture, Forest Service, Intermountain Research Station: Ogden, UT, USA, 1987; pp. 43–52. [Google Scholar]
  207. Bates, J.D.; Boyd, C.S.; Davies, K.W. Longer-term post-fire succession on Wyoming big sagebrush steppe. Int. J. Wildl. Fire 2020, 29, 229–239. [Google Scholar] [CrossRef]
  208. Bates, J.D.; Davies, K.W. Seasonal burning of juniper woodlands and spatial recovery of herbaceous vegetation. For. Ecol. Manag. 2016, 361, 117–130. [Google Scholar] [CrossRef]
  209. Kane, J.M.; Meinhardt, K.A.; Chang, T.; Cardall, B.L.; Michalet, R.; Whitham, T.G. Drought-induced mortality of a foundation species (Juniperus monosperma) promotes positive afterlife effects in understory vegetation. Plant Ecol. 2011, 212, 733–741. [Google Scholar] [CrossRef]
  210. Coultrap, D.E.; Fulgham, K.E.; Lancaster, D.L.; Gustafson, J.; Lile, D.F.; George, M.R. Relationship between western Juniper (Juniperus occidentalis) and understory vegetation. Invasive Plant Sci. Manag. 2008, 1, 3–11. [Google Scholar] [CrossRef]
  211. Everett, R.L.; Sharrow, S.H. Response of Grass Species to Tree Harvesting in Singleleaf Pinyon-Utah Juniper Stands; Res. Paper INT-334; U.S. Department of Agriculture, Forest Service, Intermountain Research Station: Ogden, UT, USA, 1985; p. 7.
  212. Sheley, R.L.; Bates, J.D. Restoring western juniper- (Juniperus occidentalis) infested rangeland after prescribed fire. Weed Sci. 2008, 56, 469–476. [Google Scholar] [CrossRef]
  213. Davies, K.W.; Bates, J.D.; Madsen, M.D.; Nafus, A.M. Restoration of mountain big sagebrush steppe following prescribed burning to control western juniper. Environ. Manag. 2014, 53, 1015–1022. [Google Scholar] [CrossRef]
  214. Plummer, A.P.; Christensen, D.R.; Monsen, S.B. Restoring Big Game Range in Utah; Publ. 69-3.; Utah Division of Fish and Game: Salt Lake, UT, USA, 1969; p. 183. [Google Scholar]
  215. Redmond, M.D.; Golden, E.S.; Cobb, N.S.; Barger, N.N. Vegetation management across Colorado Plateau BLM Lands: 1950–2003. Rangel. Ecol. Manag. 2014, 67, 636–640. [Google Scholar] [CrossRef]
  216. Hourihan, E.; Schultz, B.W.; Perryman, B.L. Climatic influences on establishment pulses of four Artemisia species in Nevada. Rangel. Ecol. Manag. 2018, 71, 77–86. [Google Scholar] [CrossRef]
  217. Nelson, Z.J.; Weisberg, P.J.; Kitchen, S.G. Influence of climate and environment on post-fire recovery of mountain big sagebrush. Int. J. Wildl. Fire 2014, 23, 131–142. [Google Scholar] [CrossRef]
  218. Ziegenhagen, L.L.; Miller, R.F. Postfire recovery of two shrubs in the interiors of large burns in the Intermountain West, USA. West. N. Am. Nat. 2009, 69, 195–205. [Google Scholar] [CrossRef]
  219. Rosenstock, S.S.; Monsen, S.B.; Stevens, R.; Jorgensen, K.R. Mule Deer Diets on a Chained and Seeded Central Utah Pinyon-Juniper Range; Res. Paper INT-410; U.S. Department of Agriculture, Forest Service, Intermountain Research Station: Ogden, UT, USA, 1989; p. 4.
  220. Rosenstock, S.S.; Stevens, R. Herbivore effects on seeded alfalfa at four pinyon-juniper sites in central Utah. J. Range Manag. 1989, 42, 483–490. [Google Scholar] [CrossRef]
  221. Sorensen, G.E.; Kramer, D.W.; Cain, J.W.; Taylor, C.A.; Gipson, P.S.; Wallace, M.C.; Cox, R.D.; Ballard, W.B. Mule deer habitat selection following vegetation thinning treatments in New Mexico. Wildl. Soc. Bull. 2020, 44. [Google Scholar] [CrossRef]
  222. Gunnell, K.L.; Monaco, T.A.; Call, C.A.; Ransom, C.V. Seedling interference and niche differentiation between crested wheatgrass and contrasting native Great Basin species. Rangel. Ecol. Manag. 2010, 63, 443–449. [Google Scholar] [CrossRef]
  223. McAdoo, J.K.; Boyd, C.S.; Sheley, R.L. Site, competition, and plant stock influence transplant success of Wyoming big sagebrush. Rangel. Ecol. Manag. 2013, 66, 305–312. [Google Scholar] [CrossRef]
  224. Williams, J.R.; Morris, L.R.; Gunnell, K.L.; Johanson, J.K.; Monaco, T.A. Variation in sagebrush communities historically seeded with crested wheatgrass in the eastern Great Basin. Rangel. Ecol. Manag. 2017, 70, 683–690. [Google Scholar] [CrossRef]
  225. Rottler, C.M.; Burke, I.C.; Palmquist, K.A.; Bradford, J.B.; Lauenroth, W.K. Current reclamation practices after oil and gas development do not speed up succession or plant community recovery in big sagebrush ecosystems in Wyoming. Restor. Ecol. 2017, 26. [Google Scholar] [CrossRef]
  226. McIver, J.; Brunson, M.; Bunting, S.; Chambers, J.; Doescher, P.; Grace, J.; Hulet, A.; Johnson, D.; Knick, S.; Miller, R.; et al. A synopsis of short-term response to alternative restoration treatments in sagebrush-steppe: The SageSTEP Project. Rangel. Ecol. Manag. 2014, 67, 584–598. [Google Scholar] [CrossRef]
  227. Daryanto, S.; Wang, L.; Fu, B.; Zhao, W.; Wang, S. Vegetation responses and trade-offs with soil-related ecosystem services after shrub removal: A meta-analysis. Land Degrad. Dev. 2019, 30, 1219–1228. [Google Scholar] [CrossRef]
  228. Anadon, J.D.; Sala, O.E.; Turner, B.L., 2nd; Bennett, E.M. Effect of woody-plant encroachment on livestock production in North and South America. Proc. Natl. Acad. Sci. USA 2014, 111, 12948–12953. [Google Scholar] [CrossRef] [PubMed]
  229. Brunson, M. Unwanted no more: Land use, ecosystem services, and opportunities for resilience in human-influenced shrublands. Rangelands 2014, 36, 5–11. [Google Scholar] [CrossRef]
  230. Archer, S.R.; Predick, K.I. An ecosystem services perspective on brush management: Research priorities for competing land-use objectives. J. Ecol. 2014, 102, 1394–1407. [Google Scholar] [CrossRef]
  231. Omernik, J.M.; Griffith, G.E. Ecoregions of the conterminous United States: Evolution of a hierarchical spatial framework. Environ. Manag. 2014, 54, 1249–1266. [Google Scholar] [CrossRef] [PubMed]
  232. Bailey, R.G. Ecoregions—The Ecosystem Geography of the Oceans and Continents; Springer: New York, NY, USA, 1995; p. 176. [Google Scholar]
  233. Benson, B. Technical Note: Pinyon and Utah Juniper Site Evaluation Procedure for Utah; U.S. Department of Agriculture, Natural Resources Conservation Service: Salt Lake, UT, USA, 2014; p. 7.
  234. Bonham, C.D. Measurements of Terrestrial Vegetation; Wiley-Blackwell: Chichester, UK, 2013. [Google Scholar]
  235. Bonham, C.C.; Mergen, D.E.; Montoya, S. Plant cover estimation: A contiguous Daubenmire frame. Rangelands 2004, 26, 17–22. [Google Scholar] [CrossRef]
  236. Eldridge, D.J. Cryptogams, vascular plants, and soil hydrological relations: Some preliminary results from the semiarid woodlands of eastern Australia. Great Basin Nat. 1993, 53, 48–58. [Google Scholar]
  237. West, N.E. Structure and function of microphytic soil crusts in wildland ecosystems of arid and semi-arid regions. Adv. Ecol. Res. 1990, 20, 179–223. [Google Scholar]
  238. Canfield, R. Application of the line interception method in sampling range vegetation. J. For. 1941, 39, 388–394. [Google Scholar]
  239. Viechtbauer, W. Conducting meta-analyses in R with the metafor package. J. Stat. Softw. 2010, 36, 1–48. [Google Scholar] [CrossRef]
  240. Schmidt, F.L.; Oh, I.S.; Hayes, T.L. Fixed-versus random-effects models in meta-analysis: Model properties and an empirical comparison of differences in results. Br. J. Math. Stat. Psychol. 2009, 62, 97–128. [Google Scholar] [CrossRef] [PubMed]
  241. Hedges, L.V.; Gurevitch, J.; Curtis, P.S. The meta-analysis of response ratios in experimental ecology. Ecology 1999, 80, 1150–1156. [Google Scholar] [CrossRef]
  242. Koricheva, J.; Gurevitch, J. Uses and misuses of meta-analysis in plant ecology. J. Ecol. 2014, 102, 828–844. [Google Scholar] [CrossRef]
  243. Nakagawa, S.; Noble, D.W.; Senior, A.M.; Lagisz, M. Meta-evaluation of meta-analysis: Ten appraisal questions for biologists. BMC Biol. 2017, 15, 18. [Google Scholar] [CrossRef] [PubMed]
  244. Nakagawa, S.; Cuthill, I.C. Effect size, confidence interval and statistical significance: A practical guide for biologists. Biol. Rev. 2007, 82, 591–605. [Google Scholar] [CrossRef]
  245. Curtis, P.S.; Wang, X. A meta-analysis of elevated CO2 effects on woody plant mass, form, and physiology. Oecologia 1998, 113, 299–313. [Google Scholar] [CrossRef]
Figure 1. Mean (± SE) pretreatment cover showing differences among treatments (A) and plant community types (i.e., black sagebrush (Artemisia nova), Wyoming big sagebrush (A. tridentata ssp. wyomingensis), and mountain big sagebrush (A. tridentata ssp. vaseyana]) (B).
Figure 1. Mean (± SE) pretreatment cover showing differences among treatments (A) and plant community types (i.e., black sagebrush (Artemisia nova), Wyoming big sagebrush (A. tridentata ssp. wyomingensis), and mountain big sagebrush (A. tridentata ssp. vaseyana]) (B).
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Figure 2. Prediction lines for linear relationships between pretreatment cover of pinyon–juniper and sagebrush (A) and between pinyon–juniper and perennial grass (B) in three sagebrush community types (i.e., black sagebrush (Artemisia nova; n = 43), Wyoming big sagebrush (A. tridentata ssp. wyomingensis; n = 63), and mountain big sagebrush (A. tridentata ssp. vaseyana; n = 128). Shaded regions are 95% confidence intervals. Asterisks indicate significance (*** p < 0.001, ** p < 0.01).
Figure 2. Prediction lines for linear relationships between pretreatment cover of pinyon–juniper and sagebrush (A) and between pinyon–juniper and perennial grass (B) in three sagebrush community types (i.e., black sagebrush (Artemisia nova; n = 43), Wyoming big sagebrush (A. tridentata ssp. wyomingensis; n = 63), and mountain big sagebrush (A. tridentata ssp. vaseyana; n = 128). Shaded regions are 95% confidence intervals. Asterisks indicate significance (*** p < 0.001, ** p < 0.01).
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Figure 3. Mean effect sizes and 95% confidence intervals (lnRR; pre vs. posttreatment cover) for pinyon-juniper (A) and sagebrush (B) responses following pinyon–juniper reduction in three sagebrush community types (i.e., black sagebrush (Artemisia nova), Wyoming big sagebrush (A. tridentata ssp. wyomingensis), and mountain big sagebrush (A. tridentata ssp. vaseyana) when treatments are both pooled and separated (solid black circles). Overall effects (blue diamonds) indicate effect sizes for all observations; values in parentheses are the number of observations used to calculate effect sizes. Effect sizes are considered significantly different from one another if confidence intervals do not overlap.
Figure 3. Mean effect sizes and 95% confidence intervals (lnRR; pre vs. posttreatment cover) for pinyon-juniper (A) and sagebrush (B) responses following pinyon–juniper reduction in three sagebrush community types (i.e., black sagebrush (Artemisia nova), Wyoming big sagebrush (A. tridentata ssp. wyomingensis), and mountain big sagebrush (A. tridentata ssp. vaseyana) when treatments are both pooled and separated (solid black circles). Overall effects (blue diamonds) indicate effect sizes for all observations; values in parentheses are the number of observations used to calculate effect sizes. Effect sizes are considered significantly different from one another if confidence intervals do not overlap.
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Figure 4. Mean effect sizes and 95% confidence intervals (lnRR; pre vs. posttreatment cover) for perennial grass (A), perennial forb (B), annual grass (C), and annual forb (D) vegetation responses following pinyon–juniper reduction in three sagebrush community types (i.e., black sagebrush (Artemisia nova), Wyoming big sagebrush (A. tridentata ssp. wyomingensis), and mountain big sagebrush (A. tridentata ssp. vaseyana)) when treatments are both pooled or separated (solid black circles). Overall effects (blue diamonds) indicate effect sizes for all observations; values in parentheses are the number of observations used to calculate effect sizes. Effect sizes are considered significantly different from one another if confidence intervals do not overlap.
Figure 4. Mean effect sizes and 95% confidence intervals (lnRR; pre vs. posttreatment cover) for perennial grass (A), perennial forb (B), annual grass (C), and annual forb (D) vegetation responses following pinyon–juniper reduction in three sagebrush community types (i.e., black sagebrush (Artemisia nova), Wyoming big sagebrush (A. tridentata ssp. wyomingensis), and mountain big sagebrush (A. tridentata ssp. vaseyana)) when treatments are both pooled or separated (solid black circles). Overall effects (blue diamonds) indicate effect sizes for all observations; values in parentheses are the number of observations used to calculate effect sizes. Effect sizes are considered significantly different from one another if confidence intervals do not overlap.
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Figure 5. Mean effect sizes and 95% confidence intervals (lnRR; pre vs. posttreatment cover) for bare ground (A) and cryptogam (B) following pinyon–juniper reduction in three sagebrush community types (i.e., black sagebrush (Artemisia nova), Wyoming big sagebrush (A. tridentata ssp. wyomingensis), and mountain big sagebrush (A. tridentata ssp. vaseyana)) when treatments are both pooled or separated (solid black circles). Overall effects (blue diamonds) indicate effect sizes for all observations; values in parentheses are the number of observations used to calculate effect sizes. Effect sizes are considered significantly different from one another if confidence intervals do not overlap.
Figure 5. Mean effect sizes and 95% confidence intervals (lnRR; pre vs. posttreatment cover) for bare ground (A) and cryptogam (B) following pinyon–juniper reduction in three sagebrush community types (i.e., black sagebrush (Artemisia nova), Wyoming big sagebrush (A. tridentata ssp. wyomingensis), and mountain big sagebrush (A. tridentata ssp. vaseyana)) when treatments are both pooled or separated (solid black circles). Overall effects (blue diamonds) indicate effect sizes for all observations; values in parentheses are the number of observations used to calculate effect sizes. Effect sizes are considered significantly different from one another if confidence intervals do not overlap.
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Figure 6. Mean effect sizes and 95% confidence intervals (lnRR; pre vs. posttreatment cover) for sagebrush (A), perennial grass (B), and perennial forb (C) responses in three sagebrush community types following pinyon–juniper reduction by mastication and cutting treatments (solid black circles) for unseeded and seeded sites. Overall effects (blue diamonds) indicate effect sizes for all observations; values in parentheses are the number of observations used to calculate effect sizes. Effect sizes are considered significantly different from one another if confidence intervals do not overlap.
Figure 6. Mean effect sizes and 95% confidence intervals (lnRR; pre vs. posttreatment cover) for sagebrush (A), perennial grass (B), and perennial forb (C) responses in three sagebrush community types following pinyon–juniper reduction by mastication and cutting treatments (solid black circles) for unseeded and seeded sites. Overall effects (blue diamonds) indicate effect sizes for all observations; values in parentheses are the number of observations used to calculate effect sizes. Effect sizes are considered significantly different from one another if confidence intervals do not overlap.
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Figure 7. Map of pinyon–juniper reduction study sites in Utah.
Figure 7. Map of pinyon–juniper reduction study sites in Utah.
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