By the end of the twenty first century, climate warming is predicted to increase average annual temperatures by 2°C to 5°C across North America, resulting in a 500-800 km northward shift in comparable average temperatures (IPCC 2014). As such, ecologists are grappling with how climate change could alter species distributions, including those of invasive species, and the potential consequences for ecosystem structure, function, and services (e.g., Walther et al. 2002; Parmesan and Yohe 2003; Pearson and Dawson 2003; Chen et al. 2011). The present-day distributions of species, and their potential responses to climate change, depend on numerous attributes, including historical ranges, modes of dispersal, environmental conditions, and species interactions (Urban et al. 2016). Thus, forecasting potential climate change effects on invasive species requires an understanding of multiple invasion processes, including dispersal and colonization abilities and responses to abiotic and biotic factors within the new range (Dukes and Mooney 1999; Hobbs et al. 2006; Broennimann et al. 2007; Hellmann et al. 2008; Walther et al. 2009; Bradley et al. 2010; Gallagher et al. 2010; Václavík and Meentemeyer 2012; Parker et al. 2013). For example, climate change may benefit non-native species through changes in abiotic barriers to colonization and/or changes in the biotic resistance of communities, both of which could produce range shifts in species invasion (Bradley et al. 2009; Allen and Bradley 2016; Merow et al. 2017). Depending on their nature, species interactions may also contract or expand non-native species ranges, thus accelerating or slowing climate tracking, either of which could significantly influence management efforts (HilleRisLambers et al. 2013; Urban et al. 2013).
Here, using the U.S. Pacific Northwest (PNW) coastal dune ecosystem as a case study, we examined how rising temperatures alter the abundance and competitive interactions of two invasive beachgrasses with implications for range shifts and dune functions and services. Foredunes (i.e., the seaward-most sand ridges running parallel to the shoreline) in the PNW are dominated by two congeneric, non-native invasive beachgrasses, Ammophila arenaria and Ammophila breviligulata. Their introduction and spread throughout the PNW has caused significant changes to the geomorphology, ecology, and ecosystem services of these systems (Zarnetske et al. 2010, 2012, 2015; Hacker et al. 2012; Seabloom et al. 2013; Biel et al. 2017, 2019a). The two beachgrasses have transformed the PNW dune system from a hummocky system of open sand and sparse vegetation cover to one dominated by tall, well–vegetated foredunes dominated by near Ammophila monocultures (Wiedemann and Pickart 1996). The bioengineering of foredunes has displaced numerous endemic plants and animals and has resulted in significant population declines of numerous plants, insects, and shore-nesting birds (Slobodchikoff and Doyen 1977; Wiedemann 1984; USFWS 1993, 2013). Moreover, the two congeners differ in the changes they create to coastal dune ecosystems. Ammophila arenaria creates taller and narrower foredunes than A. breviligulata (Seabloom and Wiedemann 1994; Hacker et al. 2012; Zarnetske et al. 2012; Biel et al. 2019a), providing superior protection against winter coastal storms (Seabloom et al. 2013) and promoting greater endemic plant diversity (Hacker et al. 2012; David et al. 2015).
The present-day range of the two non-native Ammophila species, and forecasts for the future, hinge on knowing how a combination of historical introduction events, dispersal abilities, abiotic tolerances, and biotic interactions affect their distribution and abundance. Ammophila arenaria and A. breviligulata were introduced to the U.S. Pacific coast for sand stabilization in the early 20th century, but the spatial extent of plantings differed between species. Between the 1870s and 1960s, A. arenaria plantings occurred extensively throughout the U.S. Pacific coast and it subsequently colonized dunes between Los Angeles, CA, USA (34°N) and the Queen Charlotte Islands, BC, Canada (54°N) (Green 1965; Breckon and Barbour 1974; Wiedemann and Pickart 2008). Ammophila arenaria was pervasive throughout this range but is presently restricted in its dominance on dunes from central Oregon to southern California (Seabloom and Wiedemann 1994; Buell et al. 1995; Hacker et al. 2012). In contrast, A. breviligulata plantings were primarily relegated to northern Oregon and southern Washington in the 1930s. Since its introduction, A. breviligulata has expanded predominantly northward to Vancouver Island, BC, Canada (49°N), but also southward into northern Oregon, displacing A. arenaria as the dominant beachgrass as far south as Seaside (46°N) (Fig. 1) (Seabloom and Wiedemann 1994; Page 2001; Hacker et al. 2012). There are also small extant A. breviligulata populations on beaches, including Seaside to Cascade Head (45°N), Siltcoos (44°N), Coos Bay (43°N), OR, and San Francisco, CA (38°N) (Hickman 1993; Hacker et al. 2012). Although the Ammophila congeners have spread since their initial introductions (Seabloom and Wiedemann 1994; Hacker et al. 2012; David et al. 2015), their present-day distributions in part simply reflect their respective planting histories.
However, the mostly segregated distribution of the two congeners is also a consequence of several interacting factors including species interactions, dispersal limitation, and temperature. Based on multidecadal field surveys, A. breviligulata has gradually replaced A. arenaria as the dominant vegetation on foredunes in regions where the two species co-occur (Fig. 1) (Seabloom and Wiedemann 1994; Hacker et al. 2012; David et al. 2015; Zarnetske et al. 2015). Manipulative experiments show that A. breviligulata is competitively dominant to A. arenaria, especially under lower sand deposition regimes characteristic of the northern coast (Baye 1990; Zarnetske et al. 2013). Thus, though A. arenaria is present in northern Oregon and southwest Washington, A. breviligulata has largely displaced A. arenaria, effectively limiting its northern extent (Fig. 1). Moreover, the present-day southerly range of A. breviligulata is partly a consequence of dispersal limitation. Ammophila primarily invades new territory via establishment of rhizome fragments, and to a lesser extent via seed recruitment (Maun 1984; van der Putten 1990), which can be carried offshore and transported hundreds of kilometers via ocean currents (Baye 1990; Aptekar and Rajmanek 2000). The PNW coastal climate is conducive to this mode of long-distance dispersal, with harsh winter storms that erode foredunes (Ruggiero et al. 2010), leading to rhizome fragmentation and spreading when sprouting is at its peak (Pavlik 1983; Konlechner et al. 2016). However, prevailing winds and currents during winter months are directed northerly (Komar 1998), limiting beach grass dispersal southward. Additionally, the Pacific Northwest is divided into many discrete littoral cells separated by large, rocky headlands (Komar 1985). These headlands redirect alongshore currents offshore, limiting transportation of sediment and propagules between adjacent littoral cells (Bray et al. 1995). Consequently, while A. breviligulata has rapidly expanded its range northward, geographic barriers and ocean currents have likely slowed its southern expansion and limited its potential range (Fig. 1) (Hacker et al. 2012).
Finally, evidence suggests that the two congeners differ in their temperature tolerances with A. arenaria showing better tolerance to warmer temperatures. In its native range, the optimal temperature for A. breviligulata varies between 15°C and 30°C, depending on environmental as well as population genetic variability, with reductions in primary production as temperatures approach 20°C (Seneca and Cooper 1971; Yuan et al. 1993; Emery and Rudgers 2013). In contrast, A. arenaria grows in regions where temperatures regularly exceed 40°C and is known to survive and grow when exposed to temperatures of 50°C or more (Huiskes 1979). Along the PNW coast where both species co-occur, mean maximum daily temperatures range from 14°C to 18°C and average monthly rainfall ranges from 15 mm to 55 mm between June and September, depending upon the month and location (Table S1). At the southern range limit of A. arenaria (34°N at Los Angeles, CA; Breckon and Barbour 1974), temperatures range from 18°C to 21°C and average monthly rainfall ranges from 0.25 mm to 5.3 mm (Table S1). With climate change, PNW annual air temperatures are expected to increase by 2.5-3.4°C by 2080 relative to a 1980s baseline, depending upon the climate scenario (Mote and Salathé 2010; Salathé et al. 2010), creating temperature conditions similar to the present-day climate at the southern distribution of A. arenaria in California.
Therefore, although multidecadal observational surveys and manipulative experiments demonstrate that A. breviligulata effectively constrains the northern limit of A. arenaria by competitive displacement, and dispersal limitation constrains the southern limit of A. breviligulata, it is unknown whether these range limits are sensitive to temperature. As a result, forecasting the response of the two Ammophila species to climate warming is essential because changes to their distribution and abundance could influence foredune ecology, morphology, and ecosystem services (Hacker et al. 2012; Seabloom et al. 2013; Biel et al. 2017, 2019a).
In this study, we report the effects of warming and sand burial on the abundance, morphology, and competitive interactions of the two Ammophila congeners. We used large sand-filled planters to manipulated temperature, sand burial, and the relative abundances of A. arenaria and A. breviligulata to address the following questions. (1) Does warming and sand burial alter plant growth and morphological traits for each of the Ammophila species? We hypothesized that A. breviligulata would exhibit higher sensitivity to warming via reductions in biomass, while A. arenaria would show little response. (2) Does warming and sand burial alter the direction and/or magnitude of the competitive interaction between the congeners? If so, do they exert density-independent or density-dependent effects on Ammophila growth? Given that competitive displacement of A. arenaria by A. breviligulata presently limits the distribution and abundance of A. arenaria, if A. breviligulata exhibits a negative response to warming and/or sand burial, then we hypothesize that such changes might foster coexistence or even a reversal of competitive dominance. By extension, latitudinal gradients in temperature and rising temperatures from climate change would impede southerly range expansion of A. breviligulata. However, if neither species is sensitive to rising temperatures, then climate is unlikely to prevent further A. breviligulata invasion and displacement of A. arenaria.