We conducted the study in three parks managed by the U.S. National Park Service: Death Valley National Park, Mojave National Preserve, and Lake Mead National Recreation Area, in the U.S. states of California, Nevada, and Arizona (Fig. 1). Each park exceeds 0.5 million ha and includes the largest national park in the lower 48 states (Death Valley), the third largest (Mojave), and the fourth largest (Lake Mead). These parks are in the Mojave Desert, where landforms include canyons, alluvial fans, cinder cones, low hills, mountains, dry lake beds, and intermittently flowing stream channels (Fig. 2). Predominant vegetation types include desert holly (Atriplex hymenelytra) and other shrub communities in the lowest-elevation basins, creosote bush-bursage (Larrea tridentata-Ambrosia dumosa) shrubland to elevations of 1200 m, blackbrush (Coleogyne ramosissima) shrubland at middle elevations, single-leaf pinyon-Utah juniper (Pinus monophylla-Juniperus osteosperma) woodland starting around 1,600 m on mountain slopes, and conifer forest such as white fir (Abies concolor) or bristlecone pine (Pinus longaeva) on the highest peaks (Keeler-Wolf 2007). Most annual plants are winter annuals, germinating in fall/winter (beginning in November) and growing until April (Beatley 1974).

Figure 1.

Location of three parks managed by the National Park Service in which we measured non-native plant species on 1,662 plots, Mojave Desert, southwestern USA.

Figure 2.

Views of national parks showing the variety of contexts in which non-native plants occur. Death Valley National Park: top: Death Valley floor where non-natives were generally sparse; middle: an area previously dominated by native shrubland and converted largely to non-native Bromus annual grassland following wildfire; bottom: Panamint Mountains where Bromus tectorum was the major non-native species. Mojave National Preserve: top: developed area with a history of human occupation and disturbance (Zzyzx, California); middle: Yucca brevifolia-Coleogyne ramosissima mature native shrubland, among the most susceptible communities to wildfire spread facilitated by non-native grasses; bottom: this community type following wildfire. Lake Mead National Recreation Area: top: Tamarix spp. (tall, green, leafy trees) infesting riparian areas around the Lake Mead shoreline; middle: shoreline activities can distribute non-native plants, making treating non-natives along the shoreline a priority for park managers; bottom: natural washes can serve as vectors for dispersal of non-natives.

Among the parks, Death Valley contains the lowest (along the Death Valley floor) and highest elevations (Telescope Peak in the Panamint Mountains; Table 1). Mojave Preserve contains low-lying basins and its highest elevations in the Clark Mountains, with much of the park of intermediate elevation (800–1500 m). Spirit Mountain, at 1,720 m in the Newberry Mountains, is the highest peak in Lake Mead National Recreation Area. Climate varies across the region and with elevation. The Death Valley, California, weather station at 58 m below sea level receives only 6 cm/yr of precipitation and has an average January daily low temperature of 4 °C and July high of 47 °C (1961–2012 records; Western Regional Climate Center, Reno, Nevada). In contrast, a station 1,326 m in elevation receives 27 cm/yr of precipitation, with a January average daily low of 3 °C and July high of 34 °C (1958–2011; Mitchell Caverns, California, in south-central Mojave Preserve).

Before they were designated, the parks incurred anthropogenic disturbance including clearing for townsites, agriculture, or ranches in the 1800s and early 1900s; localized mining; alteration to springs and seeps (e.g., piping water elsewhere); road and trail building; and ranching operations with cattle and sheep (Lovich and Bainbridge 1999). Non-native burros were kept as work animals by miners and continue to inhabit these parks as feral animals (Beever and Pyke 2005). Livestock grazing allotments were decommissioned in the late 1990s in Lake Mead National Recreation Area and partially decommissioned from 1998–2002 in Mojave National Preserve. Extensive roadless areas exist, but the parks do contain widespread road networks, such as 3,700 km of roads within Mojave National Preserve (Vogel and Hughson 2009). Combined human visitation to the parks was 7.8 million visitors in 2012, including 1 million in Death Valley, 0.5 million in Mojave, and 6.3 million in Lake Mead (National Park Service, Public Use Statistics Office, Denver, Colorado).

We sampled all three parks using similar stratified-random designs. We divided Death Valley National Park into 16 zones corresponding to major mountain ranges (e.g., Panamint Mountains) or valleys (e.g., Death Valley floor). Using an existing vegetation map of the park (5-ha minimum mapping unit; Thomas et al. 2004) and a Geographic Information System random point generator (ArcGIS 9.3, Esri Corp., Redlands, California), we generated 5 potential points for sampling within each vegetation type. For instance, the largest zone (Last Chance Mountain Range) contained 43 vegetation types and had 215 potential sample points. Sample points were then evaluated and field visited in random order within zones, with the goal of sampling 2 (the first 2, if possible) of 5 potential sites. Potential points were rejected because of safety concerns (e.g., cliff faces were not sampled) or unsuitability (e.g., developed areas such as campgrounds), and the next potential point was evaluated.

We divided Mojave National Preserve into 31 zones according to broad landforms (e.g., Cima Volcanic Field) in a 1:100,000-scale geologic map (Miller et al. 1991). Then, to capture elevational variation, each zone was stratified by 250-m elevation bands (e.g., 750–1,000 m). Finally, these elevation bands were stratified by predicted land cover based on the Thomas et al. (2004) vegetation map. Again using a random point generator, we selected three points for potential sampling within each vegetation type × elevation × landform zone stratum. Generation of three potential sample points within a stratum provided field crews the ability to reject unsuitable sites, as at Death Valley.

The Thomas et al. (2004) vegetation map did not extend to Lake Mead National Recreation Area, so the park was stratified into 1-km² pixels based on climate (derived from PRISM), topography (digital elevation models), and soil parent material (Lato 2006). Climate layers included July average maximum temperature (> 41 °C, ≤ 41 °C), January average minimum temperature (> 2.5 °C, ≤ 2.5 °C), average annual May through October precipitation (> 8 cm, ≤ 8 cm), and average annual November through April precipitation (6–9 cm, 9–11 cm, and 11–16 cm). Topography was categorized as drainage, flat (< 1% slope gradient), gently sloping (1 to ≤ 10% gradient), and for slope gradients > 10%, by slope aspect as northeast (0–89°), southwest (180–269°), or neutral (90–179° or 270–359°). There were 188 unique combinations of these variables extant on the landscape, and 2–3 points were randomly sampled per combination.

We used the same procedures for field data collection in all three parks. At each sample point, we surveyed a square plot of 0.1 ha for areal cover of non-native plant species (including annual, biennial, and perennial plants) using the following cover classes: present but < 1%, 1–5%, > 5–15%, > 15–25%, >25–50%, > 50–75%, and > 75%. We recorded both live and dead annual plants as a measure of cumulative presence for two reasons: 1) live annual plants are ephemeral, absent many years and when present, for only a short time in winter/spring; and 2) fuel provided by dead annual plants poses a fire hazard to mature Mojave Desert plant communities (Brisbin et al. 2013). The length of time that dead annuals persist as upright stalks varies, but Beatley (1966) noted that red brome (Bromus rubens), a major non-native in the Mojave Desert, can stand approximately two years. We summed cover of live and dead stalks of annual plants into a single cover estimate by species for each plot. Nomenclature and classification of species by longevity/growth forms (e.g., perennial forb) and native/exotic followed Natural Resources Conservation Service (2013).

We established a total of 1,662 plots that encompassed 99% of elevation ranges within parks (Table 1). We sampled high elevations in warmer months and low elevations in cooler months. We worked in Death Valley National Park between May 17 and July 2, 2010, and between January 4 and May 22, 2011. Between September 30, 2010 and June 7, 2011, we sampled Mojave National Preserve. We sampled Lake Mead National Recreation Area between February 18 and May 13, 2010, and between September 9 and October 22, 2010.

For all plots combined and each park separately, we calculated the total number of non-native species, percentage of plots containing one or more non-native species, mean non-native richness (species/0.1-ha plot), and frequency of each species. We used Pearson correlation coefficients to examine relationships between elevation and non-native species richness and cover. We compared species prioritization rankings from two systems: NatureServe’s I-rank (Randall et al. 2008) and the California Invasive Plant Council system (Cal-IPC; Warner et al. 2003). To compare species recorded on plots with all known records of non-native species within each park, we obtained species lists maintained by each park. Using PC-ORD v. 6 (McCune and Mefford 1999), we calculated Sørensen similarities between parks of species composition (presence/absence data) recorded on plots and from park species lists.

Eighty-two percent of plots contained at least one non-native plant species (Table 1, 2). Non-native richness ranged from 0–10 species/0.1 ha, with a median of 2 species and mean of 1.60 ± 0.03 (± standard error of mean). Thirty-one percent of plots contained one non-native species, 30% two, 17% three, and 4% four to ten species. For example, considering two species of Bromus, 36% of plots contained neither species, 3% only cheatgrass (Bromus tectorum), 50% only Bromus rubens, and 10% both (Fig. 3). Total non-native cover ranged from 0-81%, with a median of 0.4% and mean of 2.5 ± 0.1%. Most plots had low cover and few had high cover: 60% had < 1% cover, 26% had 1–5%, 8% had 5–10%, and 6% had > 10% cover.

Figure 3.

Presence or absence of the non-native annuals Bromus rubens and Bromus tectorum in national park units of the Mojave Desert, southwestern USA.

Of 29 total non-native species on plots, 59% were annuals, 10% annuals/biennials, 14% annual to perennials, and 17% perennials (Table 2). By growth form, 55% were forbs, 35% grasses, 7% shrubs, and 3% trees. Annual forbs (31%) and annual grasses (28%) were the most prevalent groups. The most frequent species included: Bromus rubens (60% of plots), redstem filaree (Erodium cicutarium; 39%), Schismus spp. (28%), Bromus tectorum (13%), prickly Russian thistle (Salsola tragus; 4%), Sahara mustard (Brassica tournefortii; 4%), and saltcedar (Tamarix ramosissima; 3%).

Results were mixed regarding availability of species prioritization rankings, consistency between ranking systems, and relationship between a species’ rank and its frequency (Table 2). Nineteen of 28 taxa (68%, with Schismus spp. grouped to genus) had rankings available from the Cal-IPC system and 17 were ranked by the NatureServe I-rank system. There were six species (21%) not ranked by either. Of the 9 taxa ranked by both systems, consistency varied. Ranking was consistent for Bromus tectorum and Tamarix ramossisma, with both systems ranking the species as high priority and capable of pervasive impacts. However, the two systems returned opposite rankings for Schismus spp. Cal-IPC ranked the taxon as ‘low’ priority, while NatureServe ranked it as ‘high’ priority.

We did not detect an overall correlation between elevation and non-native richness or cover (Fig. 4). The only trend apparent was that all plots containing > 2 non-native species occurred at elevations < 2,000 m. Individual species displayed stronger relationships with elevation than did total non-native measures (Fig. 5). Brassica tournefortii and Schismus spp. were most frequent at elevations below 1,200 m. Although infesting a broad elevation range, Erodium cicutarium was most frequent at middle elevations between 400 and 1,600 m. Distribution of the most frequent species, Bromus rubens, also was centered on middle elevations. Bromus tectorum exhibited a different pattern: it was most frequent at elevations above 1,600 m in Death Valley National Park and Mojave National Preserve, and at the highest elevations present in Lake Mead National Recreation Area.

Figure 4.

Scatterplot of elevation and non-native plant species richness and cover derived from 1,662 plots in Death Valley National Park, Mojave National Preserve, and Lake Mead National Recreation Area in the Mojave Desert, southwestern USA. There was no relationship between elevation and non-native richness or cover (Pearson r = 0.00). The inset graph in (b) shows percentages of plots infested by either non-native grasses or forbs, or both (‘neither’ signifies 18% of plots not invaded and 2% invaded only by woody species).

Figure 5.

Relationship between elevation and frequency of major non-native plant species in three Mojave Desert parks, southwestern USA. High elevations were absent in Lake Mead.

The total number of non-native species detected on plots within parks was similar, ranging from 17–22 species/park (Table 1). However, mean non-native richness/plot was twice as high in Mojave National Preserve and Lake Mead National Recreation Area as in Death Valley National Park. Mojave Preserve had the fewest un-invaded plots and 3× as many plots containing at least two non-native species as did Death Valley.

Bromus rubens was the most frequent species in all three parks, and its highest frequency was in Mojave National Preserve (Table 2). Notable differences in species frequencies included low frequency of Schismus spp. and Erodium cicutarium in Death Valley relative to the other parks, and higher frequency of Bromus tectorum in Death Valley and Mojave compared to Lake Mead. Additionally, Brassica tournefortii was not detected on any plots in Death Valley, whereas the species was the fourth most frequent at Lake Mead.

Lists maintained by each park contained similar numbers of non-native species, ranging from 73–83 species (Table 1). The percentage of a park’s non-native flora detected on plots was also similar among parks at 23–30%. Most species on plots were on these lists except for some new records that the plots produced: Chilean chess (Bromus berteroanus) and alfalfa (Medicago sativa) in Death Valley, and common mallow (Malva neglecta) in Mojave Preserve. Compositional similarity of non-native species lists was 51–60% among parks, slightly lower than the 57–76% similarity for plots among parks (Table 3).

Present invasion status of these parks could be interpreted from different viewpoints. On one hand, the fact that 82% of plots were invaded by at least one non-native species is alarming. Moreover, the ecosystem engineer, Bromus rubens, occurred in 60% of plots. By providing copious and persistent fuel, this species promotes spread of wildfire, a novel disturbance requiring centuries for recovery of mature perennial communities in this desert (Abella 2010, Steers and Allen 2010). Some other frequent species, such as Tamarix ramosissima, also can dramatically impact indigenous ecosystems, including riparian areas which are hotspots of native biodiversity (Shafroth et al. 2005). On the other hand, 60% of plots had < 1% cover, indicating extensive minimally infested area.

Plot-based surveys of landscapes like ours provide information on species distribution and abundance and are not exhaustive botanical inventories (Barnett et al. 2007). Our plots contained 21% of the 139 non-native species on inventory lists of these parks. Our study years were near average for detecting annual plants, based on receiving 132% of long-term average (74 years) precipitation for the 2010 growing season (October 2009 through April 2010) and 104% for the 2011 growing season (October 2010 through April 2011; Las Vegas, Nevada airport station). Many of the undetected species are uncommon in the backcountry and inhabit only specific sites, such as campgrounds, roadsides, and cultural areas (e.g., historical cabins including non-native landscaping vegetation, or orchards).

Although non-native species measures such as total species and species composition were generally similar among parks, some notable differences existed. Mojave National Preserve had the fewest un-invaded plots, and Death Valley National Park had the lowest non-native richness/plot and fewest plots containing multiple species. Mojave Preserve has the most extensive history of disturbance and was most recently placed under National Park Service protection in 1994 (Beever and Pyke 2005). In the Czech Republic, the later reserves were created, the more non-native plants they contained (Pyšek et al. 2003). Mojave Preserve also has extensive middle elevations most susceptible to non-native plant fuel production and wildfire spread (Van Linn et al. 2013). Less invasion in Death Valley might relate to the park containing elevation extremes, which were least invaded, and the lowest frequencies of Schismus spp. and Erodium cicutarium. Another difference was that Lake Mead contained the lowest frequency of Bromus tectorum, likely because high elevations were absent (Abella and Tendick 2013).

Although correlations between elevation and non-native richness and cover were not detected, individual species were most frequent within particular elevation ranges. Additionally, elevation extremes (below sea level and > 2,000 m) were least invaded in terms of non-native species richness. If climate becomes warmer and drier in the region as some projections suggest (Barrows and Murphy-Mariscal 2012), lower elevations might become even less invasible, and higher elevations more so. Forecasting how invasibility might change at high elevations is difficult, because high elevations may already be invasible and simply have not received seed pressure (Keeley et al. 2003). In Rocky Mountain National Park, Colorado, Bromus tectorum frequency increased by 50% across high-elevation ecotonal plots over a 12-year period from 1996–2007 that was relatively dry (Bromberg et al. 2011). Given minimal invasion at the highest elevations in our study parks, early detection and treatment of new invaders and newly invaded sites might be particularly appropriate.

Unfortunately, the most frequently detected species, such as Bromus rubens, are not simply ‘innocuous’ inhabitants of the parks, but rather the most damaging type of non-native species (i.e. ecosystem engineers; Crooks 2002). These non-natives can disrupt critical ecological functions, fundamentally conflicting with the national park goal of promoting native species and processes. One such example warranting further attention is food availability to the desert tortoise (Gopherus agassizii). This long-lived (~ 50 years) reptile inhabits all three parks and is listed as threatened under the U.S. Endangered Species Act. Two studies in the Mojave Desert reported that despite being among the most abundant plants, the non-native Schismus spp. and Bromus rubens were avoided by foraging tortoises (Jennings 1997, Oftedal et al. 2002). Although Schismus spp. represented 98% of the plants encountered, tortoises ate < 0.1% of them (Oftedal et al. 2002). Compared to native annual forbs, Schismus has lower water and protein content and high potassium toxicity. Moreover, the pointed florets of Bromus can injure tortoises directly when ingested (Medica and Eckert 2007). Non-native annuals compete with native annual forage plants, and natives have increased when Schismus and Bromus were removed (Brooks 2000). Declining populations of desert tortoises face pervasive dominance of non-preferred food plants, which further influence habitat conditions by providing fuel facilitating spread of desert wildfires (Brooks and Berry 2006).

Our results showing how widespread invading species are on these landscapes exemplify a broad issue of biological invasions being a driver of contemporary species evolution (Leger and Espeland 2010). Two examples from our set of invading species illustrate this. Within 7 years of being exposed to elevated CO₂, Bromus rubens plants evolved lower rates of leaf stomatal conductance, a physiological adaptation linked with improved water-use efficiency (Grossman and Rice 2014). Most of the invading species in southwestern deserts are annuals, such as Bromus, able to rapidly evolve (Table 2). This underscores the importance of management actions to reduce population sizes of non-natives and their capacity for evolving traits that make them even more competitive (Leger and Espeland 2010). Genetics of native species may also be shifting in response to environments altered by non-native plants. In the Great Basin Desert, growth of native perennial grasses responded most rapidly to watering on sites that were most heavily invaded by Bromus tectorum (Leger 2008). This implied that native plants were adapting to become more competitive with Bromus (Leger 2008). Adaptation to the presence of a non-native species may be beneficial for persistence of some native plants. However, it is undesirable from a national park perspective, where native species are supposed to evolve through natural processes, not through anthropogenic species introductions.

These assessment data are an initial step towards non-native plant distribution mapping, which needs to consider extreme spatio-temporal variability in desert ephemeral plants. Distribution and abundance of annual plants varies both with inherent site productivity and weather in any particular year (Wallace and Thomas 2008, Casady et al. 2013). In addition to being a practical strategy given difficulty in sampling numerous desert sites in a short spring growing season, we included live and dead annual plants to both represent cumulative recent ‘presence’ and importance of live and dead biomass as fuel. Thus, the survey data could facilitate spatial modeling of site productivity for these species across the landscape, and serve as baseline data for modeling temporal variation.

Our findings revealed several considerations regarding species prioritization as a management tool. Not all species of management interest had ‘off the shelf’ rankings available, necessitating that managers develop their own rankings, a difficult task for little-studied species. Even if a species has no ranking available, existing ranking systems may still offer a useful framework for developing customized rankings. Results also suggested that comparing different ranking systems, when available, is useful to assess consistency of rankings (Andreu and Vilà 2010). Rankings can differ for numerous reasons, such as different emphases (e.g., inclusion or exclusion of management difficulty as an evaluation factor), and date of the ranking which affects information available. As one example, our comparisons illustrated that Schismus spp. were ranked oppositely (‘low’ and ‘high’ priority) by two ranking systems (Randall et al. 2008, Warner et al. 2003). In addition to providing fuel for wildfire and competing with native plants (Abella and Smith 2013), the observation that Schismus is non-preferred forage for the desert tortoise suggests that the ‘low’ ranking for Schismus warrants re-evaluation (Oftedal et al. 2002).

Prioritizing species currently at the extremes – those that are infrequent (but capable of impacts) and those that are widespread and capable of major impacts – may maximize use of limited treatment resources. For example, early treatment of currently infrequent species such as ripgut brome (Bromus diandrus) and spiny sowthistle (Sonchus asper) would follow a principle that early detection and treatment is the most cost-effective and successful strategy (Lodge et al. 2006). This also helps reduce risk that these species become future problems.

Rather than viewing pervasive, high-impact invaders like Bromus rubens as ‘hopeless’, treating these species at priority sites is likely important to avoid negating other management efforts and indeed protecting core values of parks. Over 28,700 ha (4.5%) of Mojave National Preserve burned in fires partly fueled by Bromus rubens between 2005 and 2011, destroying mature desert vegetation, as well as cultural resources (Hegeman et al. 2014). Strategically treating priority sites can reduce landscape fuel connectivity (Brisbin et al. 2013). Treatments suitable for broad areas such as those infested by Bromus rubens require further experimentation, but early timed herbicide application has reduced Bromus while promoting natives (Allen et al. 2005). Competitive native species also can reduce Bromus (Abella et al. 2012), and biocontrol agents are under evaluation (Baughman and Meyer 2013).

We identified sites containing multiple non-native plant species, which can affect candidate treatment strategies and their effectiveness. The potential influences of multiple species are numerous, such as: (i) herbicide effectiveness can vary with plant growth form, (ii) treatment timing can be difficult when species’ phenologies differ, (iii) required treatment duration can fluctuate among species varying in soil seed bank longevity, (iv) more complicated treatment regimes can increase costs and potential for negatively impacting native species, and (v) chances increase that other non-native species replace a focal treated species (Abella 2014).

What evidence exists for relationships of multiple species with treatment difficulty in the Mojave Desert? Brooks (2000) found that hand pulling Bromus rubens or Schismus spp. increased native annuals but also increased the non-native forb Erodium cicutarium. Similarly, hand weeding Brassica tournefortii increased Erodium (Marushia et al. 2010). In a post-fire environment dominated by non-native annuals, Steers and Allen (2010) had more encouraging results where herbicide not only reduced the grasses Bromus and Schismus, it also reduced Erodium. Native annual forbs increased, and the native grass sixweeks fescue (Vulpia octoflora) was not damaged (Steers and Allen 2010). The treatments exploited the accelerated early season phenology of non-native annuals (compared to native annuals) by applying herbicide early in the growing season. Refining knowledge of the earliest possible time for treating non-natives – which might vary among years – is warranted to manage single- and multi-species infestations while promoting natives (Marushia et al. 2010, Abella et al. 2013).