A total of eight steppe and forest study sites were chosen that span the meadow steppe-xerophytic forest ecotone in eastern Washington (See Suppl. material 1: Table S1). The co-dominance of Symphoricarpos albus with Festuca idahoensis and Pseudoroegneria spicata characterize the mature vegetation in the Festuca idahoensis/Symphoricarpos albus habitat type (sensu Daubenmire 1970) in the four eastern Washington meadow-steppe sites (1250 m² each). The four forest sites (1250 m² each) are dominated by Pinus ponderosa with co-dominate Symphoricarpos albus in the understory (hereafter termed the P. ponderosa/S. albus habitat type, sensu Daubenmire and Daubenmire 1968). Sites were 40.9 ± 6.1 km apart; the adjacent sites were at least > 0.5 km apart.

A seed mixture of three grasses and three forbs (a native, naturalized, and invasive species of each taxonomic category) was used in seed addition plots in this study. The native perennials Pseudoroegneria spicata and Geum triflorum are prevalent in meadow-steppe (Daubenmire 1970); these species are less prominent in P. ponderosa forests (Daubenmire and Daubenmire 1968). Secale cereale, a naturalized annual, is a Washington Class C noxious weed that appears as a volunteer in many cultivated crops (Gaines and Swan 1972, Washington Noxious Weed Control Board [WNWCB]: http://www.nwcb.wa.gov) and establishes, albeit rarely, in meadow steppe and P. ponderosa forest (Connolly et al. 2014, USDA PLANTS database: http://plants.usda.gov). Centaurea cyanus, a naturalized annual, is also registered on the WNWCB monitor list and is widely established at low density throughout the meadow steppe and P. ponderosa forest in eastern Washington (Roche and Talbot 1986, USDA PLANTS database: http://plants.usda.gov). The invasive annual Bromus tectorum is abundant, even dominant, in the meadow steppe (Daubenmire 1970, Mack 1981) but infrequent in P. ponderosa forest (Daubenmire and Daubenmire 1968, Pierson and Mack 1990). Cirsium arvense, an invasive perennial, commonly occurs in anthropogenically disturbed sites and is present in both habitat types (http://www.nwcb.wa.gov, Connolly 2013). Seeds of G. triflorum, Ce. cyanus, B. tectorum, and Ci. arvense were collected in bulk from our meadow steppe sites from May – September 2010 and 2011; seeds of P. spicata and S. cereale were obtained from a local vendor (Rainer Seed Company, Davenport, WA, USA) to insure we had adequate numbers of locally produced seeds for all treatments (described below).

We substantiate the immigrant class (naturalized vs. invasive) of each non-native test species based on 1) a preliminary vegetation analysis conducted at all 8 study sites (Connolly et al. 2014) and 2) state and regional published accounts habitats (e.g., Gaines and Sawn 1972, Roche and Talbot 1986) of the relative abundance of these species. Importantly, some work has evaluated the mechanisms driving competition dynamics between these specific native perennials and introduced annuals (e.g., Madsen et al. 2012), but outcomes remain unclear and suggest evaluation of their respective establishment potential and relative performance across environmental and disturbance gradients may help identify the drivers of introduced plant colonization and persistence in natural sites.

The effects of seed addition and disturbance were assessed in late July-early August 2011 in six experimental blocks arranged in a 2 × 3 grid at each site (25 × 50 m); blocks were 25-m apart. Each block was comprised of four hardware cloth exclosures (aboveground dimensions were 45 × 45 × 45 cm tall, 1 cm² openings); exclosures in each experimental block were arranged 2-m apart in a square (24 exclosures per site, 192 exclosures total across all sites). Before its installation each exclosure was sprayed with enamel paint (Krylon®) to prevent leachate from the hardware cloth affecting plant growth within the exclosure. Exclosures were embedded 15-cm deep into the mineral soil to exclude the treatment being confounded by vertebrate seed predators.

Each block contained a complete 2 × 2 factorial cross with seed addition and disturbance as factors. To generate disturbance treatments, we removed all vegetation and litter from the soil surface and churned the top 3 cm of mineral soil without removing any soil. Disturbed soil was then leveled within each exclosure to minimize differences in soil microtopography among these exclosures (Harper 1977). We extended disturbance treatments in a 0.5-m buffer zone around each disturbance treatment exclosure to minimize shading by neighboring understory plants. Vegetation was left intact within and around undisturbed control exclosures. Exclosures were embedded carefully around each replicate assigned to the undisturbed treatment and produced no detectable changes to plant cover within or around the exclosures. Importantly, this disturbance treatment did not necessarily release experimentally sown plants from competition but rather increased the availability of some resources (e.g., light, Suppl. material 2: Figure S1) that are known to influence plant competition in understory environments.

In early August 2011, 96 exclosures amongst the sites were sown with an admixture of seeds containing three grasses (P. spicata, S. cereale, B. tectorum) and the three forbs (G. triflorum, Ce. cyanus, Ci. arvense). Seeds were sown evenly across a 30 × 30 cm square at the center of each exclosure (0.09 m² sampling area, 50 seeds of each species, 300 seeds sown total per exclosure). Seeds were pressed firmly onto the soil surface to minimize post-dispersal seed movement. In the remaining 96 exclosures amongst the sites no seeds were added in order to measure natural recruitment of study species and evaluate the contribution of seed addition to plant establishment counts.

Exclosures were monitored monthly for damage and other extraneous events; plants were counted in early July 2012 to estimate establishment. Following July counts, all above ground plant biomass was harvested within each exclosure, separated by species, dried (48 hours at 70°C) and weighed. Plant establishment was quantified early in the growing season and before the production of reproductive structures in order to minimize the possibility of introducing non-native species. Natural recruitment by species other than our six test species was rare within these exclosures; nonetheless these recruits were excluded from the analysis. Average individual seedling biomass was estimated by dividing total biomass for each species in each exclosure by the number of that species in the exclosure. Plots that received seed addition were treated with glyphosate herbicide (Roundup®, Monsanto Company) at the cessation of the study. Additionally, the immediate area in a 15-m radius surrounding each exclosure was monitored throughout 2012 and 2013 to detect and remove extraneous introductions.

We used general linear mixed models to evaluate whether seed addition, disturbance, and plant immigration class (Native vs. Naturalized vs. Invasive) influenced the number of individuals that established within each community (Steppe vs. Forest). July individual counts of each species were averaged across all blocks at a site to generate site-level averages for each treatment combination and for each species. Ten exclosures were damaged in March 2011. These units were excluded from analysis as vertebrate seed predators and grazers can strongly influence plant establishment in these habitats (Connolly et al. 2014) and may generate undetectable variation in seedling recruitment. Five of eight sites, however, had no damage to exclosures, and no site with damaged exclosures had fewer than three replicates of each treatment combination with which to generate site-level averages for each species. Site-level averages for plant counts for each species were used as model response variables and all fixed effects (habitat, seed addition, disturbance, plant immigration class) and their possible interactions were included in analysis of the response variables. Site identification and the interaction between site, seed addition, and disturbance were included in this model as random effects to account for the nested structure of the design. Average individual counts for each species were log (x+1) transformed prior to analysis. We used post hoc tests to evaluate pairwise contrasts using the Tukey-Kramer method to control for multiple comparisons (Littell et al. 2006).

Analysis of average individual biomass followed a similar model structure but was limited to seed addition plots to insure the analysis was conducted between individuals with similar durations of residence time within each plot. Individual biomass estimates of each species were averaged across all blocks at a site to generate site-level averages for each treatment combination and for each species. Individual biomasses were square-root transformed before analysis. Average Ci. arvense biomass at one steppe site (Smoot Hill - Summit) was a significant outlier differing from the species’ other mean values by over three standard errors and was driven by the rapid second year growth of an adult Ci. arvense already residing in the plot. Omitting this observation permits the analysis to satisfy assumptions of normality; consequently, final model analysis for average individual biomass did not include this observation. Models evaluating plant establishment and biomass employed the Kenward-Roger approximation to estimate appropriate degrees of freedom (Littell et al. 2006). All analyses were conducted in SAS (Proc GLIMMIX, SAS 9.3; Cary, North Carolina, USA).

Our experimental design incorporated the effect of plant immigration class (Native, Naturalized, or Invasive) by evaluating two representative species from each class (one grass, one forb). Although the species selected represent common or dominant plants in these forest and steppe communities (See Study species section) and site-level quality can be assessed by the relative abundance of these native and non-native species (Daubenmire 1970, Mack 1981, Pierson and Mack 1990), we were only able to accommodate two species of each plant immigration status within each plot in our experimental design. Given the limited number of species within each immigration class, we must tentatively interpret conclusions drawn from the main effect immigration class or interactions including immigration class. In order to accommodate interpretation at the species level, we include supplemental results and figures that evaluate plant species as a main effect instead of plant immigration class in the same general linear mixed model framework (Suppl. material 3: Tables S2–S3, Figs S2–S3). Importantly, given the early experimental harvest date and relatively large plot size we assume species sown in our seed mixtures demonstrated independent responses to treatments and had negligible effects on the overall emergence and growth of other species occurring in the same plot. Ancillary analysis using statistical models that helps account for that lack of independence with multivariate responses (i.e., MANOVA general linear models evaluating the response of multiple species sown in the same plot) indicate similar results for main fixed effects to those derived from mixed models (Suppl. material 4: Tables S4–S7).

Seed limitation influences recruitment of many native (Turnbull et al. 2000, Clark et al. 2007) and non-native species (Jongejans et al. 2007, Swope and Parker 2010, Connolly et al. 2014). Seed limitation can be the product of 1) a paucity of reproducing plants, 2) poor seed dispersal, 3) biotic agents that directly reduce seed number, 4) poor propagule viability, or 5) some combination thereof (Harper 1977, Seabloom et al. 2003, Davis 2009). Non-native plants are unlikely to be dispersal-limited between communities in our study region as the propagules of non-native species can readily traverse the PNW steppe-forest ecotone and establish (albeit rarely and for short durations) in disturbed coniferous forest sites (e.g., Pierson and Mack 1990, Dodson and Felder 2006). Unlike native and invasive species, adult S. cereale and C. cyanus are however rare at both forest and steppe sites (Connolly et al. 2014), implicating the lack of reproducing plants, poor seed dispersal, or both as a major limiting factor for naturalized species within these communities. Moreover, preferential attack by granviores and consistent losses caused by pathogenic soil fungi in both habitats also contribute substantially to seed limitation, occasionally eliminating entire experimentally-introduced populations (Connolly 2013, Connolly et al. 2014).

Differences in species’ biomass production between habitat types and with or without disturbance may also influence non-native propagule pressure and contribute to seed limitation for non-native species. For example, individual B. tectorum biomass correlates strongly with total seed mass produced per individual plant (R²_adj = 0.861; P < 0.001, Almquist 2013) and our study shows average B. tectorum biomass was quantitatively greater in undisturbed steppe (43.4 ± 8.2 mg [mean ± SE]) than in undisturbed forest (10.0 ± 4.0 mg) at the time of July 2012 harvest (Suppl. material 3: Fig. S3) suggesting that average annual seed production per B. tectorum individual is likely greater in the PNW meadow steppe than in the adjacent, undisturbed ponderosa pine understory. Plants were harvested from exclosures before the generation of reproductive tillers to eliminate unintentional plant introductions at these sites, but previous estimates of B. tectorum fitness within each of these communities corroborate this hypothesis (steppe: 16–20 seeds per adult plant, Pyke 1986; forest: 0.7–0.9 seeds per adult plant, Pierson and Mack 1990). Additionally, disturbance of plant canopy cover in our study resulted in 87.5% and 31.7% greater individual B. tectorum biomass in the forest and steppe, respectively (Suppl. material 3: Fig. S3). These disturbance-mediated effects on productivity may also increase individual seed production for non-native plants. By limiting productivity, plant cover likely limits non-native plant seed production, influences seed dispersal dynamics, lowers propagule pressure, and facilitates community resistance to the establishment of light-requiring non-native plants.

Disturbance can facilitate a species’ transition from naturalization to invasion (Crooks and Soulé 1999, Groves 2006, Chakraborty and Li 2010, Richardson and Pyšek 2012) by increasing resource availability and eliminating competitors (Davis et al. 2000, Davis and Pelsor 2001, Myers and Harms 2009, Richardson and Pyšek 2012, Leffler et al. 2016). In our study, naturalized species’ establishment in disturbed plots in the forest and steppe were equivalent to or exceeded the establishment of co-occurring invasive and native species in identical treatments (Suppl. material 3: Fig. S2), suggesting that resources provided by the removal of understory (< 1.5m high) canopies (e.g., light [Fig. S1], soil nutrients, water) helped meet a major requirement for recruitment for these naturalized species. Seedlings of invasive species may have higher relative growth rates and net assimilation rates than introduced, non-invasive congeners (Grotkopp et al. 2010) and, consequently, invaders may be more robust in resource scarce (e.g., undisturbed) sites than co-occurring naturalized species. Disturbance has a strong, positive effect on the growth of these two naturalized species, suggesting resource limitation, and in particular light limitation, may be a consistent, effective biotic barrier against some members of this class of plant immigrants in PNW forests. Residence time, however, can also influence the potential for naturalized species to invade (Groves 2006) and, while the two naturalized species examined in this study have likely occupied PNW natural habitats for over 100 years (e.g., Gaines and Sawn 1972, Roche and Talbot 1986), it is possible that sufficient time has not elapsed to permit the expansion of these species within these habitats. Further research is needed to determine the extent to which the interaction of resource availability, disturbance regimes, and species residence time in a novel habitat affects the differential establishment of invasive and naturalized species (Grotkopp et al 2002, Groves 2006, Moravcová et al. 2010).

Competition in the PNW coniferous forest understory is a strong biotic barrier to invasive species that are abundant in the adjacent steppe, particularly B. tectorum (Pierson and Mack 1990). For example, low light availability at the soil surface in the P. ponderosa forest understory may cause low non-native species recruitment and individual seedling biomass. P. ponderosa forest understory lowered light transmittance at the soil surface to 20% of ambient conditions in June 2012, whereas shading in undisturbed steppe only lowered light transmittance to 60% of ambient conditions at the soil surface (see Suppl. material 2: Methods S1 and Fig. S1). Shading may directly influence the survival of some invasive species; for example, Bakker (1960) reported large Ci. arvense seedling mortality if light intensities fall below 20% of full sunlight – a threshold similar to that measured beneath the understory at our Ponderosa Pine forest sites. Additionally, shade lowers the probability that non-native seeds receive essential light-related germination cues (Pons 2000, Jensen and Gutekunst 2003) and may slow non-native seedling growth rate and result in lower fecundity through modifications of seedling microclimate (e.g. low temperatures, increased snow cover, Mack and Pyke 1984, Pierson and Mack 1990).

The environmental tolerances of introduced species interact with a novel habitat to determine a species’ potential for naturalization (Richardson and Pyšek 2012), and climatic mismatch between an invader and a novel habitat may preclude non-native plant establishment (Alpert et al. 2000). Consequently, pre-adaptation to forest understories will raise the likelihood that an introduced species will naturalize in the interior of these temperate North American forests. Shade-tolerant non-native perennials (e.g. Berberis thunbergii, Celastrus orbiculatus, Lonicera spp.) readily establish in eastern North American forests (Zheng et al. 2006) and could plausibly be introduced as horticultural escapes and even naturalized in western coniferous forests (Smith and Mack 2013). Some non-native grasses may also tolerate low light levels in North American forest understories and may be candidates for future naturalizations and potential invasions (e.g. Miscanthus sinensis, Horton et al. 2010; leptomorphic bamboos, Smith and Mack 2013). Understanding the interactions between the physical tolerances of introduced species and the severity of competition in novel habitats would improve predictions of non-native plant naturalization or invasion potential on a habitat-specific level (Chytrý et al. 2008, Richardson and Pyšek 2012).

Few studies directly evaluate the relationship between biotic resistance and the relative abundance of introduced species (van Kleunen et al. 2010, Richardson and Pyšek 2012). However, here we report the results of one part of a three-experiment series evaluating how functionally different components of biotic resistance (i.e., seed predation [Connolly et al. 2014], seed parasitism [Connolly 2013], competition [reported here]) relate to the prevalence of non-native plants between habitats differing in susceptibility to invasion. Invasive plants are conspicuous by their tolerance or avoidance, or both, of most biotic barriers in the extensively invaded PNW steppe (Mack 1986), whereas naturalized species are significantly restricted, and occasionally eliminated, by the joint action of biotic interactions in the same habitat and at the same time. In the examples investigated here, community resistance to invasions is substantial in adjacent low-elevation PNW coniferous forest. For the species we evaluated, limitations to recruitment and performance imposed by a dense canopy and seed limitation imposed by granivores and, to a lesser extent, seed pathogens ensure that undisturbed forests interiors are likely to be well defended against the encroachment of many non-native species, particularly annual grasses. Collectively, our work demonstrates that biotic resistance likely plays a role both in determining 1) the distribution of some non-native species amongst a region’s communities and 2) the position of a non-native species along the introduced-naturalized-invasive species continuum in a community. Further work evaluating the potential synergistic interactions between multiple biotic barriers with a larger suite of representatives from each immigration class (e.g., Suwa and Louda 2011, Maron et al. 2012, Maron et al. 2013) will help elucidate how biotic interactions ultimately influence demography of non-native plants and the distribution of non-native plants within Pacific Northwest steppe and forest communities.