INTRODUCTION

Sagebrush ecosystems have long been considered among the most endangered in North America (Noss et al. 1995; Knick et al. 2003), with perhaps a third of presettlement area of sagebrush already converted to other land uses or highly degraded. Over the past 100 yr, fire suppression, livestock grazing, urban expansion, oil and gas extraction, expansion of native conifers such as juniper and piñon pine (Juniperus occidentalis, J. osteosperma; Pinus monophylla, P. edulis), and invasion of exotic weeds such as cheatgrass (Bromus tectorum) have contributed most to the decline of sagebrush communities in the intermountain region (Pellant 1994; Miller and Tausch 2001; Ingelfinger and Anderson 2004). At higher elevations, conifer expansion and depletion of fine fuels due to heavy livestock grazing has shifted fire regimes from relatively frequent and low (< 50 yr mean fire return interval) to more infrequent and high (> 50 yr mean fire interval) severity (Miller and Rose 1999; Miller and Tausch 2001; Miller and Heyerdahl 2008). At lower-elevation treeless sagebrush ecosystems, cheatgrass has invaded at the expense of native perennial species, and mean fire return intervals have shifted from > 50 yr to < 10 yr in some places (Whisenant 1990; D'Antonio and Vitousek 1992). Under current climatic conditions, both piñon and juniper woodlands and exotic annual grasses have the potential to dominate an even greater area (Wisdom et. al 2002), and global warming is likely to exacerbate this trend (Pyke and Knick 2003; Tausch and Nowak 2000; Neilson et al. 2005; Balch et al. 2013; Bradley 2010).

For several years now, land managers have attempted to arrest the conversion of sagebrush steppe lands into woodland and cheatgrass systems, restore a desirable herbaceous understory, and reduce fuel loads by applying treatments such as prescribed fire, mowing, chaining, cutting, masticating, and/or herbicides. Although site-specific information exists on the effectiveness and ecological effects of some treatments, there is scant multivariate scientific information available on treatment outcomes over the range of environmental and ecological conditions that occur across sagebrush ecosystems. The Sagebrush Steppe Treatment Evaluation Project (SageSTEP) evaluates the ecological effects of prescribed fire and its surrogates (mechanical and herbicide treatments) at 21 sagebrush steppe sites in the Great Basin and surrounding areas (McIver et al. 2010). The multisite design of SageSTEP is intended to provide information on how different site conditions influence treatment response, while the multivariate design is intended to understand how treatments influence relationships within systems and to identify potential trade-offs among variables.

Butterflies have long been considered indicators of ecosystem condition, thus allowing insights about the likely responses of a larger set of fauna of conservation concern (Thomas 1983; Swengel 1998; Fleishman 2000). Furthermore, the decline of several species of threatened and endangered butterflies has been linked to habitat loss due to invasive plant invasion (Russell and Schultz 2010). This is primarily because native butterflies are closely linked to native plants (Ehrlich and Raven 1965). Since sagebrush steppe restoration is keenly concerned with the control of invasive species, it makes sense to monitor faunal components that would likely be sensitive to changes in the balance between native and exotic plant species. More generally, butterflies are good indicators of ecosystem condition due to their sensitivity to changes in the distribution and abundance of native host plants (Ehrlich and Raven 1965) and to native and exotic nectar sources (Holl 1995).

Butterflies are also easy to count and identify on the wing (Pollard 1977) and so can be sampled with relatively little impact to their populations. Further, butterfly larvae are intimately linked to native host plants, particularly perennial forbs and grasses, so assessing the effects on them will tell us something about effects of treatment on the plant community and linkages between flora and fauna (Ehrlich and Raven 1965). Finally, testing the effects of land management treatments on the fauna can give us more insight on the extent to which management practices, especially those with which flora and fauna have no evolutionary history (mechanical and herbicide treatments), result in unintended or undesirable consequences. Although some butterfly species can adapt to sudden loss of host plants or nectar sources (Singer et al. 1994; Boughton 1999), mechanical or herbicide treatments may have other structural or functional effects that are unique enough to cause problems for native species.

In this article, we describe butterfly species composition across a network of 16 of the 21 SageSTEP sites and relate this to plant species composition, habitat structure, and site characteristics. We then report on the response of butterfly species, species groups, and communities to prescribed fire and fire surrogate treatments. We expected that butterfly community composition would vary in accordance with known species distributions in the Great Basin and that it would correspond roughly to native plant community composition. We also expected that prescribed fire would have somewhat different effects on butterflies when compared to its “fire surrogates,” such as herbicides and mechanical treatments, and that effects would decrease with time after treatment.

METHODS

Study Sites and Treatment Plots

Butterflies were sampled between 2006 and 2012 at 16 sites within the SageSTEP Network on sagebrush steppe lands in the Great Basin and surrounding areas. Nine sites make up the SageSTEP woodland experiment, representing sagebrush systems that are relatively mesic (259–462 mm annual precipitation; Table 1) and characterized by expansion of piñon and juniper into areas that were historically sagebrush steppe. The nine sites are divided into three regions, each dominated by a different woodland overstory: 1) western juniper region: four sites in Oregon and northern California, dominated by western juniper (Juniperus occidentalis Hook.); 2) piñon-juniper region: three sites in Nevada, with overstory shared by singleleaf piñon (Pinus monophylla Torr.& Frém.) and Utah juniper (Juniperus osteosperma [Torr.] Little); and 3) juniper-piñon region: two sites in Utah, with overstory dominated by Utah juniper, with minor representation of Colorado piñon (Pinus edulis Engelm.; McIver et al. 2010). Seven sites make up the sage-cheat experiment, representing sagebrush systems that are treeless, lower elevation, more xeric (214–364 mm annual precipitation), and characterized by cheatgrass invasion of sagebrush steppe. The sage-cheat experiment is composed of three sites in Utah, Nevada, and western Idaho; two sites in Oregon; and two sites in Washington (Table 1). Although all 16 sites are classified as cool desert and have similar vegetation and land use patterns (Bestelmeyer et al. 2009), weather patterns differ markedly across this geographic range. Sites in California, Oregon, Washington, and southwestern Idaho have a Pacific maritime climate, with nearly all precipitation originating in the Pacific Ocean and falling between November and June. Sites in Nevada, Utah, and eastern Idaho have a more continental climate, with less precipitation falling from November to June and relatively more summer rains originating from the Gulf of Mexico, usually in July and August.

Table 1.

SageSTEP site information, including site acronym and name, state, year treated, percent plot area burned in prescribed fire, elevation, percent slope, aspect, dominant tree species, current native vegetation, plot position within site, mean plot area, and minimum intertransect distance.

For the woodland experiment, each site comprised three or four 10–20-ha plots, with each plot receiving one distinct treatment, randomly assigned (Table 1). We selected one plot as unmanipulated control, applied prescribed fire to a second plot, and clear-cut all trees on a third plot. At both Utah juniper-piñon woodland sites, we masticated all trees within a fourth plot with a Bullhog rotary mower (McIver and Brunson 2014). Prescribed fire was applied first between August and November 2006, 2007, or 2008. The goal was to accomplish 100% tree mortality by fire within each prescribed fire plot in an effort to release the residual understory; due to variation in weather conditions, prescribed fires burned between 38% and 85% of each plot area (Table 1). Clear-cut and mastication treatments were implemented within 6 mo of fire treatments. For the clear-cut treatment, all trees > 2 m tall were cut down and left on the ground across the contour. For the mastication treatment, all trees > 2 m tall were shredded with the rotary mower, and residue was left where initially deposited.

For the sage-cheat experiment, each site comprised four 25–75-ha plots, with each plot receiving one distinct treatment, randomly assigned (Table 1). We selected one plot as unmanipulated control and applied prescribed fire, a mowing treatment, and a broadleaf herbicide treatment to the remaining three plots. Prescribed fire was applied first, from May to October 2006, 2007, or 2008, and was intended to blacken 100% of each plot area. For six of the seven sites, prescribed fires burned between 40% and 79% of each plot area (Table 1); at Roberts, only 8% of the plot area burned, so the prescribed fire treatment was not evaluated for this site. Once fire was implemented for each site, both herbicide and mowing treatments were applied to two other plots within the following 8 mo. Both treatments were designed to remove about 50% of sagebrush cover to reduce woody fuels and release the understory herbaceous species. The herbicide tebuthiuron (N-[5-1,1-dimethylethyl-1,3,4-thiadiazol-2-yl]-N,N′-dimethylurea) was applied over the entire plot at a rate dictated by prior testing to remove 50% of the shrub overstory. Rotary mowers were set at a predetermined height to remove and distribute roughly 50% of sagebrush biomass over each entire plot. It should be noted that the Roberts sage-cheat site experienced a severe wildfire (Jefferson Fire) on July 13, 2010, that killed nearly all vegetation in two of the four plots. Since treatments were applied in 2007 at Roberts, we present only 3 yr of posttreatment data for this site (2008–2010), with the 2010 butterfly sample collected just 3 wk prior to the wildfire.

RESULTS

A total of 5 933 butterflies were observed at the 16 sites during the 7-yr study period, comprising five families and 52 species (Appendix 1; available online at  http://dx.doi.org/10.2111/REM-D-13-00127.s1). Over 72% of the total count was represented by the 10 most commonly observed species; 10 species were observed fewer than four times. The average number of butterflies counted per 1 000-m survey across all years at all sites was 13.52 (± 1.57 SE) and was reasonably consistent over the 7 yr, except in 2007 (35% of average) and 2009 (153% of average). Woodland sites had about three times the average count per 1 000-m survey (woodland = 17.16 ± 2.23 SE individuals; sage-cheat = 5.46 ± 0.53 SE individuals), and nearly twice the average survey richness (woodland = 3.09 ± 0.13 SE species; sage-cheat = 1.79 ± 0.08 SE species) compared to sage-cheat sites. Butterfly species richness was correlated with overall plant species richness at the plot scale (r² = 0.45; P < 0.01; y = 0.3x − 0.9), with average plot-level plant species richness per year nearly twice as high at woodland sites (43.5 ± 1.66 SE spp.) compared to the relatively lower-elevation sage-cheat sites (25.6 ± 1.56 SE spp.).

At woodland sites, NMS ordination distinguished the three woodland regions along axis 1 and sites within each region along axis 2 (Fig. 1a). In the western juniper region, the principal indicator taxa for the Blue Mountain site include common blues (PLIC), juniper hairstreaks (CAGR), and Edith's checkerspot (EUED), with ochre ringlets (COTU) indicating the other three western juniper sites. The piñon-juniper sites ordinated toward the center and include several indicator taxa, principally the pine elfin (INER) and large whites (POIA) for Seven Mile; skippers (HEIA), Melissa blues (PLME), and fritillaries (SPIA) for South Ruby; and sulfurs (COAS), Riding's satyr (NERI), and Anicia Checkerspots (EUAN) for Marking Corral. In the juniper-piñon region, the principal indicator taxa for Greenville Bench include checkered skippers (PYCO) and, for Onaqui, desert marbles (EULO). Principal environmental correlates (r² > 0.50) include higher cover of duff, embedded litter, and Idaho fescue (Festuca idahoensis) toward the western juniper region vs. higher mean gap sizes and bluebunch wheatgrass cover (Pseudoregneria spicata) toward the juniper-piñon region. Overall butterfly abundance was higher toward the sites ordinating toward the bottom of the graph (BM, SV). Analysis with the MRPP demonstrated highly significant separation of each region in ordination space, with all pairwise P-values < 0.001. Moreover, when MRPPs were run for site comparisons, the majority of pairwise P-values (84%) were < 0.03; the remainder were < 0.05. Likewise, MRPP analyses for pairwise interannual comparisons were all significant (P < 0.02) with the exception of the comparison between 2009 and 2010; thus, community structure of butterflies varied markedly not only among woodland sites but also among survey years. On the other hand, the MRPP yielded no significant community structure differences among woodland treatments for any pairwise treatment comparison (P > 0.10).

Figure 1.

See Accompanying Table.

At sage-cheat sites, NMS ordination yielded similar results as observed for the woodlands, with four more or fewer distinct groups of sites recognizable (Fig. 1b). The most compositionally diverse of the sage-cheat sites was Moses Coulee, which ordinated by itself as a distinct group of plots, with four key indicator taxa (gray hairstreak [STME], common blue [PLIC], ochre ringlet [COTU], and wood nymphs [CEPE]). The two geographically close Hart Mountain Refuge sites (Gray Butte and Rock Creek) clustered together, with both sites featuring a dominance of desert marbles (EULO). Interestingly, despite their greater geographic separation, Saddle Mountain and Owyhee had very similar compositions of butterfly genera, with both sites featuring an abundance of skippers (HEIA) and large whites (POIA). Finally, the Onaqui and Roberts sites (the two most eastern sage-cheat sites) were also quite similar in generic composition, with each site featuring an abundance of Melissa blues (PLME), ladies (VACA), sulfurs (COAS), and checkered skippers (PYCO). The principal environmental correlates (r² > 0.40) of axis 1 were shallow-rooted native bunchgrasses (PSG), particularly Sandberg's bluegrass (POSE) in the northwest and squirreltail (ELEL5) in the east, and weather factors at the time of survey (higher wind in the northwest [Wind] and higher temperature in the east [TEMP]). Higher axis 2 scores were correlated with plant species richness (Prich) and cover of perennial forbs (PFb), both of which were attributes of the sites ordinating toward the top of the graph. Analysis with the MRPP indicated that most site-level pairwise comparisons were significantly different (all < 0.03), with the exception of the two Hart Mountain Refuge sites, Gray Butte vs. Rock Creek (P = 0.12), and the two most easterly sites, Onaqui vs. Roberts (P = 0.07). Like the woodland sites, interannual variation was also marked, with each year different from every other year, with the exception of 2009 and 2010 (P < 0.03 for all pairwise comparisons except 2009 and 2010). However, MRPP analysis of treated sites yielded no significant differences in community structure among sage-cheat treatments for any pairwise treatment comparison (P > 0.10 for all pairwise comparisons). Finally, when the woodland and sage-cheat butterfly main matrices were each compared statistically to their floral matrix counterparts (Mantel test), the null hypothesis of no relationship between each pair of main matrices was rejected (P < 0.000001), indicating distinct among-site similarity in the ordination of butterfly and floral communities.

High spatial variation in butterfly community structure, together with marked interannual variation in counts at most sites, made determination of treatment effects challenging. Within the context of substantial spatial and temporal variation, however, certain patterns of treatment response were observed. When all sites were analyzed as a whole (N = 16 sites), treated plots had higher transient abundance and richness compared to untreated controls (Table 2) starting in the second year after treatment and lasting through year 4 (Fig. 2). No treatment effects were observed for local butterflies at the network level, although both abundance and richness increased with time after treatment in most plots regardless of treatment.

Table 2.

Posttreatment means and standard errors for local and transient butterfly richness and abundance, and indication of interannual variation (*) analyzed for the network as a whole (N = 16) with two-factor general linear modeling (treatment × time since treatment). Different letters indicate significant pairwise difference between treatment and control (P < 0.05).

Figure 2.

Mean (± SE) a, transient richness and b, abundance at the Network Level (N = 16 sites) for untreated control plots and combined fire and mechanical plots, 1–4 yr after treatment. *Above fire/mechanical error bar indicates significant difference (P < 0.05) between treatment and control for comparison at each year after treatment.

In the woodland experiment, two of the eight functional groups and two of the eight common and widespread species exhibited significant treatment response. Among transients, the number of legume-feeding sulfurs (SU-Fab) and the number of transient whites (WT-Bra) were higher in plots treated either with fire or by mechanical means (Table 3). Sulfurs were consistently more abundant in treated plots throughout the 4-yr posttreatment time period (Fig. 3a), while transient whites were more abundant in treated plots only in posttreatment years 2 and 3 (Fig. 3b). Higher numbers of transients (both sulfurs and whites) in treated plots were mirrored by vegetation data, which showed that both annual and perennial percent forb cover increased with treatment of any kind relative to untreated controls (Table 4). In particular, annual forb cover increased markedly in burn plots, with mean posttreatment cover averaging nearly fourfold that of untreated controls (13.76% in burn plots vs. 3.53% in control plots). Among local butterflies, numbers of Melissa blues increased in burned and clear-cut plots, and the effect size of its plot-level response was correlated with the effect size of the cover of its Astragalus host plants (Fig. 4; r² = 0.30; y = 0.64x − 0.03; P < 0.05). Although the mean multisite effect size correlation for fire and mechanical treatment plots was very similar (two symbols labeled with “TOT” in Fig. 4), individual sites typically varied markedly in effect size correlations for fire vs. mechanical treatments (Fig. 4). For example, effect sizes for both Astragalus and Melissa blue were high for the Blue Mountain (BM) prescribed fire plot but low for the mechanical plot there, while the Walker Butte (WB) site showed the opposite pattern. We observed no other effect size correlation between local butterflies and their principal larval host plants. The only observed decreases in butterfly numbers observed in the woodland experiment were for legume-feeding blues in bullhog plots and for the juniper hairstreak (Table 3). The difference in blues was due entirely to a region effect, in which numbers were lower for all plots in the juniper-piñon region. Since the bullhog treatment was applied only to the two juniper-piñon sites, this led to the apparent bullhog plot effect. The juniper hairstreak, on the other hand, declined in abundance after treatment at all sites where it was initially common, primarily the western juniper and the piñon-juniper sites Marking Corral and South Ruby (Table 3). Having a larva that feeds on juniper, removal of its host plant had clear effects on abundance of this species, and this effect persisted through 4 yr of posttreatment time. Finally, significant interannual variation was observed for nearly every analyzed taxon in the woodland experiment, with numbers generally increasing with time after treatment, due to relatively low counts in 2007 and generally high counts in 2009 and 2011. The only taxon that did not exhibit interannual variation was the juniper hairstreak, which had consistent survey counts relative to treatment throughout the study period (Table 3).

Table 3.

Posttreatment means and standard errors for variables in the woodland experiment for which significant treatment effects or interannual variation (*) was observed with two-factor general linear model (treatment × time since treatment). Different letters indicate significant differences between treatment and control (P < 0.05). WJ indicates western juniper; PJ, piñon-juniper; JP, juniper-piñon.

Figure 3.

Mean (± SE) abundance of a, sulfurs and b, transient whites for the woodland experiment (N = 9 sites) for untreated control plots and combined fire and mechanical plots, 1–4 yr after treatment. *Above fire/mechanical error bar indicates significant difference (P < 0.05) between treatment and control for comparison at each year after treatment.

Table 4.

Posttreatment means and standard errors (years 1–3) for annual and perennial forb cover in sage-cheat and woodland experiments. Different letters indicate significant difference in pairwise comparisons with two-factor general linear model (treatment × time since treatment). * indicates significant interannual variation.

Figure 4.

Effect size of Melissa blues (Plebejus melissa) vs. the effect size of one of its primary larval host plants Astragalus spp. for pooled posttreatment samples taken in prescribed burn and mechanically treated plots at those woodland sites at which Melissa blues were present. Effect size metric used was Hedge's D = (mean count in control plot − mean count in treatment plot)/ pooled standard deviation. Woodland site abbreviations: BM indicates Blue Mountain; GR, Greenville Bench; MC, Marking Corral; ON, Onaqui; SR, South Ruby; SV, Seven Mile; WB, Walker Butte; TOT, all-site average.

In the sage-cheat experiment, we observed persistently higher local species abundance and richness in burn plots at five of the six sites at which our prescribed burn blackened at least 40% of the plot area (Gray Butte, Moses Coulee, Rock Creek, Owyhee, Saddle Mt); local butterfly abundance and richness in mow plots or plots treated with the broadleaf herbicide tebuthirion were no different than controls (Table 5). The burn effect on the abundance of local butterflies persisted through the 4th yr posttreatment, with control and burn plot abundance similar only in year 2 (Fig. 5). We also observed persistently higher numbers of grass-feeding skippers (SK-Poa) and local mustard-feeding whites (WL-Bra) after burning, but mowing or herbicide application had no apparent effect on these taxa (Table 5). Local butterfly abundance declined with time since treatment in most plots, with relatively higher counts in 2008 and 2009 and lower counts in 2010 and 2011. Much of this effect was due to decreases over time in the numbers of western branded skippers and in local whites (primarily marbles; see Appendix 1; available online at  http://dx.doi.org/10.2111/REM-D-13-00127.s1). Local species richness also varied through time, but variability was not clearly or consistently linked to year effects. Among transients, numbers of Becker's white (Pontia beckerii) were lower in mowed plots relative to control or burn plots, with this effect persisting through 4 yr of posttreatment (Table 5). Neither transient abundance nor richness varied markedly at sage-cheat sites over time.

Table 5.

Posttreatment means and standard errors for variables in the sage-cheat experiment for which significant treatment effects or interannual variation (*) was observed in analysis with two-factor general linear modeling (treatment × time since treatment). Different letters indicate significant pairwise difference between treatment and control (P < 0.05).

Figure 5.

Mean (± SE) local abundance for the sage-cheat experiment (N± =± 7 sites) for untreated control plots and combined fire and mechanical plots, 1–4 yr after treatment. *Above fire/mechanical error bar indicates significant difference (P± <± 0.05) between treatment and control for comparison at each year after treatment.

DISCUSSION

Observed butterfly community structure generally conformed to known patterns of species distribution in the Great Basin and showed a close relationship to native plant communities across the SageSTEP network of sites. However, both spatial (among-site) and temporal (among-year) variation in butterfly community structure was very high and tended to overwhelm patterns of treatment response. When species and species groups did respond to treatment, response was generally positive regardless of treatment type, with response to prescribed fire vs. its mechanical surrogates (clear-cutting, mastication) more similar than expected. Similarity in response among treatments was likely due to the fact that woody vegetation removal, whether by fire or by machine, tended to increase soil water availability (Roundy 2014), which enhanced grass and forb production (Table 4) and in turn provided more resources for butterfly larvae (host plants) and adults (nectar). Finally, observed treatment responses were persistent, with most variables showing divergent trajectories between control and treated plots through 4 yr of posttreatment time.

A total of 52 species of butterflies were observed at the 16 SageSTEP study sites over a 7-yr period, a relatively low number compared to other butterfly studies of comparable scope conducted in the Great Basin. For example, in a 3-yr montane canyon study examining the principal factors that explain patterns of butterfly species richness, Fleishman et al. (2000) observed 33 and 40 butterfly species from only two mountain ranges in central Nevada (Toiyabe and Toquima, respectively), nearly double the maximum richness we found at our most diverse woodland sites, after 7 yr of observations (Blue Mountain, 18 species; Marking Corral, 17 species). Lack of available water (Murphy and Wilcox 1986), great distance to water (Fleishman et al. 1997), and restriction of sampling to an early phenological window (May through mid-July) all probably contributed to the relatively low species richness observed in the current study, especially at the sage-cheat sites. In addition, the higher species richness we observed at the higher-elevation woodland sites was likely due in part to the positive correlation with plant species richness, which has been reported in other studies (Hogsden and Hutchinson 2004).

The pronounced differences in butterfly community structure among sites, at the species, generic, and group levels, is one of the most striking results of the current study. The broad geographic extent of the SageSTEP study might explain some among-site differences in species composition due to geographic range limits of individual species. But nearly 64% of Great Basin butterfly species are widespread in distribution, occurring in their preferred habitats not only in the Great Basin but also in the Sierra Nevada to the west and the Rocky Mountains to the east (Austin and Murphy 1987). More likely, among-site differences are due to several factors, including availability of host plants, landscape context, and topographic features as well as site history. Certainly, when ordinations of butterflies and plants are compared within each experiment (woodland and sage-cheat), patterns of among-site distances in ordination space are remarkably similar (Mantel test), reflecting the strong relationship between butterflies and the native flora. In any case, the magnitude of among-site variability observed in the current study is not unprecedented. For example, working at a number of sites within the Toquima Range, Fleishman (2000) observed substantial spatial and temporal variability in butterfly species composition and richness. Her data also indicated that butterfly community similarity decreased with the distance between inventoried units, with the most distant units tending to be markedly dissimilar. Furthermore, Fleishman et al. (2000) also reported considerable among-site differences in the relationship between butterfly communities and environmental gradients, with surveys in the Toquima and Toiyabe ranges indicating opposite correlations between species richness and elevation. Although we do not yet have the sample sizes necessary to quantify patterns of interannual variation in butterfly communities, it is also clear from other work that temporal variation tends to be considerable as well, with year-to-year surveys producing distinctively different results at the same sites (Pollard et al. 1998; Fleishman 2000; Fleishman et al. 2000; Ross and Miller 2000; Kleintjes et al. 2004).

At the level of the butterfly community, treatments designed to restore degraded sagebrush steppe habitat produced measurable impact only on transient richness and abundance, both of which increased after treatment (Table 2; Fig. 2). However, when community response was measured by the combination of relative abundance and species composition (community structure), no measurable effects were observed. Part of the reason for this is that marked spatial (among-site) and temporal (among-year) variability in butterfly numbers and species composition created so much “noise” in the data that treatment-induced “signals” were difficult to pick out of community-level data. Indeed, variation in butterfly communities among sites and through the years often produced a much stronger signal in community data than did treatments, as demonstrated by the significant interannual variation observed for eight of the 20 variables analyzed. Neither Fleishman (2000) nor Ross and Miller (2000) reported marked effects of prescribed fire on butterflies when effects were evaluated at the community level (total richness or abundance). Rather, both studies identified among-site, among-plot, or among-year variability as a major contributing factor in their determination of no effect. In a study on prairie restoration, however, Vogel et al. (2007) were able to detect a compositional effect of treatment, with burning and grazing treatments generating similar richness but somewhat different community structures. They suggest that no one practice will benefit all species or even all species within habitat-specialist or habitat-generalist guilds.

At the species and species-group level, however, a few notable treatment effects were observed. The most obvious was the decline in the number of juniper hairstreaks in woodlands after the removal of trees by either prescribed fire or mechanical treatments. The reason for hairstreak decline is obvious: larvae feed on juniper vegetation, and treatments thus decreased the availability of larval host plants. But nearly every other species or species group that was measurably affected increased in numbers after treatment in both sage-cheat and woodland experiments, indicating that butterfly habitat generally improved as a result of treatment. Moreover, these effects generally persisted through 4 yr posttreatment, indicating that the mechanisms behind treatment response are long lasting.

At sage-cheat sites, prescribed fire had the most obvious effect, with local butterfly abundance and richness consistently higher in fire-treated plots (Table 5). These effects were largely due to higher abundance of grass-feeding skippers (SK-Poa) and mustard-feeding local whites (WL-Bra) in fire plots relative to controls. Skippers are relatively sedentary as adults, so it is possible that these modest differences were due to improved larval feeding habitat, which included a variety of native bunchgrasses. It is also possible that larval host plant resources for desert marbles (the most common representative of the local white group) improved with burning, although at no site at which it was common did any of its known mustard host plants (Arabis, Descuriana, Lepidium, Sisymbrium, Streptanthus) increase in cover in burned plots. The fire effect was also persistent through 4 yr of posttreatment time, and there was no evidence that numbers of these groups were converging over time in fire vs. controls or other treated plots. The only species for which negative treatment effects was observed was the transient Becker's white (Pontia beckerii), which declined in mowed plots relative to controls or fire-treated plots. The mechanism for this decline is unclear, as annual forb nectar resources were generally higher in mowed plots (Table 4), and there was no evidence that potential larval host plants (mustards) declined after mowing.

At woodland sites, mechanical treatments, including both clear-cutting and mastication, caused increases in the abundance of legume-feeding sulfurs (SU-Fab) and mustard-feeding transient whites (WT-Bra; Table 3; Fig. 3). Similarly, numbers of Melissa blues were higher after both prescribed fire and clear-cutting (Table 3; Fig. 4). Positive responses to treatment are most likely due to the fact that removal of trees by any means begins a cascade of effects that has the ultimate result of improving both larval and adult feeding habitat for most sagebrush steppe butterfly species. In particular, water is the most important limiting resource in sagebrush steppe systems (Chambers et al. 2007, 2014), and piñon and juniper trees are the most effective competitors for it. When trees are removed, soil water availability markedly increases (Roundy 2014), and these increases are accompanied by shifts in resource utilization toward shrubs (mechanical treatments only) and herbaceous vegetation (both mechanical and burning treatments). Since many sagebrush steppe butterfly species, as well as prairie species, are linked to native herbaceous vegetation (grasses and forbs) for larval feeding (Ehrlich and Raven 1965; Boggs and Freeman 2005; Moranz et al. 2012) and since many adults depend on forb flowers for adult feeding (Murphy 1983; Boggs and Freeman 2005), increases in the production of particular larval host plant species (e.g., Astragalus; Fig. 4) and forb cover in general (e.g., Table 4) will tend to improve butterfly foraging habitat. In any case, the fact that increases in soil water availability have, like observed butterfly effects, persisted through 4 yr of treatment (Roundy 2014) suggests that enhanced soil water availability is the root mechanism behind increases in butterfly abundance at most sites.

Enhancement of larval food plant availability by both fire and mechanical treatments is the most likely mechanism behind observed increases in Melissa blues. This interpretation is supported by the positive correlation between the plot-level effect size of Melissa blues and that of one of its primary host plants Astragalus spp. (Fig. 4). Certainly, larval food resources can have significant impacts on adult life history features of holometabolous insects, including body size, which can in turn influence population growth (Boggs 2003). In our study, while Melissa blues clearly responded positively to restoration treatments, juniper hairstreaks responded negatively because of the removal of their larval host plants. This underlines the fact that any significant habitat alteration is likely to benefit some species and impact others (Vogel et al. 2007). One would expect, however, that as long as restoration practices are implemented on sufficiently small scales, positive and negative effects on species will tend to balance out at the landscape level.

Ross and Miller (2000) also suggested that increases in specific larval host plants (e.g., lupine) were linked to increases in the abundance of butterflies that feed on them (common blue: Plebejus icariodes) but also identified improved nectar resources as the primary mechanism behind increased butterfly abundance 1 yr after burning in western juniper woodlands in eastern Oregon. Most likely, improvement of adult nectar habitat is the most likely mechanism behind treatment-induced increases in the number of transients such as sulfurs and large whites. Since many nectar species are annual forbs, which generally increased in cover after treatment (Table 4; see also Miller et al. 2014), tree-removal treatments essentially created “bull's-eyes” of nectar resources at the plot scale that could have attracted strong-flying adult species of butterflies from outside the plots, such as large whites and sulfurs. Similar results were found by Kleintjes et al. (2004), who reported increases in butterfly abundance and richness after mechanical treatments to remove trees in piñon-woodlands in northern New Mexico. They also reported increases in herbaceous cover overall and increases in 5 of the 10 most common nectar and larval host plants after treatment and suggested that the treated watershed became an “oasis” that attracted nectaring adults from adjacent areas. In prairie habitat, Vogel et al. (2007) reported similar linkages between butterfly response and vegetation, with butterfly abundance and diversity responding positively to burning or mowing treatments and best explained by a negative association to bare ground and by a positive association with percent forb cover. It is also possible, however, that increases in the number of sulfurs was due in part to the creation of more “open” habitat that some of these species are known to prefer (e.g., Colias eurytheme; Scott 1986; Meyer and Sisk 2001) or to increased insolation of treated stands (Waltz and Covington 2004). Whatever the mechanism, the negative correlation between woody cover and butterfly abundance and richness has been noted elsewhere (Erhardt 1985), reinforcing the close linkage between butterflies and herbaceous vegetation (Pollard et al 1998; Grill et al. 2005; Vogel et al. 2007). Certainly, for most butterfly studies in which investigators have evaluated treatments designed to remove or reduce woody vegetation in semiarid systems, the linkages between butterflies and herbaceous vegetation have been emphasized. This suggests that treatment effects on the herbaceous flora and the butterfly fauna will likely move in parallel for the most part, even though it will always be necessary to monitor both components to be certain that no unintended consequences arise from management treatments.

MANAGEMENT IMPLICATIONS

Management activities, especially those that replace stands, are very likely to change species composition of invertebrates due to habitat changes that favor some species and impact others. With a juniper-feeding larva, juniper hairstreaks exhibited a decline in numbers, short of local extirpation, at all sites at which they were common. This result was expected and is no cause for alarm but does emphasize the importance of maintaining a balance across the landscape in the spatial extent of management activities that replace stands. While most other butterfly species and species-group variables did not change with treatment, most of those that did change increased in numbers. This is most likely due to the fact that removal of woody vegetation by any means (fire or fire surrogate treatments) increased water availability for herbaceous vegetation, which increased its cover in the short term and led to improvement in both larval food and adult nectar resources. Most of the significant effects observed in this study persisted for 4 yr after treatment. That trend would be expected to continue for some time, until enhanced soil water resources are captured by regrowing vegetation.

Strong ties to the native plant community favor butterflies as a monitoring tool to assess environmental change in the Great Basin. Yet high temporal and spatial variability in numbers suggests that monitoring would have to be long term and of considerable spatial extent in order to yield meaningful information.