Mediterranean, invasive, woody species grow larger ...

RESEARCH ARTICLE A M E R I C A N J O U R N A L O F B O TA N Y

Mediterranean, invasive, woody species grow larger than their less-invasive counterparts under potential global environmental change1 Jennifer Erskine-Ogden2,3, Eva Grotkopp2, and Marcel Rejmánek

PREMISE OF THE STUDY: Revealing biological differences between invasive and noninvasive species is essential for predicting species’ distribution changes with global environmental change. While most research has focused on differences between invasive and noninvasive species under favorable conditions using herbaceous species, invasive woody angiosperms are also of great ecological concern. Our study focused on how growth and allocation may change for invasive and noninvasive, mediterranean, woody angiosperms under future conditions caused by global change, specifically increased nitrogen deposition and drought. METHODS: We tested how seedling functional traits differed between invasive and noninvasive woody angiosperms under different experimental conditions in a greenhouse setting. We compared growth rates and allocation patterns using two levels of soil nitrogen and three levels of watering. We also examined trait log response ratios to increases in nitrogen and increases in water. Our study sampled angiosperm trees and shrubs, incorporating congeneric/ confamilial relationships through 13 phylogenetically controlled contrasts. KEY RESULTS: Three functional traits were highly and positively associated with plant invasiveness for most conditions studied: seedling plant mass, leaf area, and height. Invasive species also had significantly higher root mass ratios at low water regardless of nitrogen input. Invasive and noninvasive species had similar log response ratios to increases in nitrogen and watering for studied traits. CONCLUSIONS: Mediterranean, woody, invasive species’ larger mass, leaf area, and early height advantage under elevated nitrogen input and increased root production in drought conditions may lead to increased invasion of these species with expected global climate change. KEY WORDS climate change; drought; functional traits; growth rate; leaf area; log response ratio; nitrogen deposition; root mass ratio

Global climate change and anthropogenic resource additions are altering many habitats worldwide (Sala et al., 2000; IPCC, 2014). Increasing atmospheric concentration of CO2 as well as rising overall temperatures are leading to more extreme weather events, changes in rainfall patterns and both increases and decreases in total precipitation (Walther et al., 2009; Diez et al., 2012; IPCC, 2014). Nitrogen (N) deposition is increasing worldwide due to changes in agricultural methods (chemical fertilization and a proliferation of large animal feed lots), industry, transportation, and urbanization (Fowler et al., 1998; Fenn et al., 2010; Ma et al., 2015). Available N enrichment can have considerable impacts by changing community 1

Manuscript received 20 November 2015; revision accepted 25 January 2016. Department of Evolution and Ecology, University of California, Davis, California 95616 USA 2 Authors contributed equally to the work. 3 Author for correspondence (e-mail: [email protected]) doi:10.3732/ajb.1500494

composition and promoting the decline of species richness across ecosystems worldwide (Chapin et al., 2000; Fenn et al., 2010; Ochoa-Hueso et al., 2011; Field et al., 2014). Species’ abundances, ranges, and overall biodiversity are being altered by these increases in available N and C, coinciding with global climate change and anthropogenic disturbances (Walther et al., 2009; Diez et al., 2012; Meier et al., 2012; Porter et al., 2013). Global climate change and increased available N may favor the successful introduction, naturalization, and spread of invasive plant species into new regions (Walther et al., 2009; Pardo et al., 2011; Porter et al., 2013) and can confer many advantages to these invasive species (Bradley et al., 2010; Dukes, 2011; Diez et al., 2012; González-Muñoz et al., 2014). Invasive species are able to capitalize on available N in naturally nutrient poor communities such as the South African fynbos (Sharma et al., 2010) and serpentine soils (Huenneke et al., 1990) by efficiently utilizing nutrient pulses (Davis et al., 2000; Dawson et al., 2012; Vallano et al., 2012) and thus

A M E R I C A N J O U R N A L O F B OTA N Y 103(4): 613–624, 2016; http://www.amjbot.org/ © 2016 Botanical Society of America • 613

614

•

A M E R I C A N J O U R N A L O F B OTA N Y

outcompete low-nutrient-adapted plants. They also have been shown to take advantage of early precipitation events through phenological priority effects (Wolkovich and Cleland, 2011). Both capitalizing on nutrient pulses and phenological priority effects can lead to higher recruitment rates by invasive species into disturbed environments (Hobbs and Huenneke, 1992). Species with the ability to take advantage of increased resource availability and/or pulse events have the capacity to become more invasive (Richards et al., 2006; Davidson et al., 2011; Godoy et al., 2011; Morris et al., 2011; Dostál et al., 2013). Possible mechanisms for opportunistic invasion by plants include greater trait plasticity that allows species to invade new habitats (Drenovsky et al., 2012a) and rapid evolution after naturalization leading to invasion into a wider breadth of habitats (Bossdorf et al., 2005; Whitney and Gabler, 2008; Schlaepfer et al., 2010). On the other hand, invasive species can be at a disadvantage to native species in the novel environment if native species are better able to tolerate increasingly stressful conditions such as prolonged drought (review by Daehler, 2003; Bradley et al., 2010; Diez et al., 2012; but see Burns et al., 2007; Funk and Vitousek, 2007). Understanding the potential invasiveness of species with continued global anthropogenic change is a priority for ecologists and land managers (Tulloss and Cadenasso, 2015). The effects of climate change on plant community composition are very complex because temperature increases and water availability depend greatly on local geomorphology, soil properties, and the reactions of species to these changes (Rapacciuolo et al., 2014; Tulloss and Cadenasso, 2015). Therefore, experiments controlling environmental inputs such as nitrogen and water are needed. One approach that should be considered when predicting a species’ invasive potential is to uncover underlying biological trait differences among invasive and noninvasive species. Research has shown that invasive species grow faster and have greater biomass as seedlings than noninvasive species (Grotkopp et al., 2010; van Kleunen et al., 2010b; Dawson et al., 2011; Morris et al., 2011). However, most of these analyses were conducted under favorable growing conditions (high nutrient levels with sufficient moisture for optimal growth), and plants were harvested soon after germination and near their maximal relative growth rate. While these high-resource conditions are often used to differentiate invasive vs noninvasive species based on their functional traits, these conditions usually are not encountered in natural areas. In wildlands, N levels vary greatly due to geology and topography, elevated levels of fertilizer run-off and atmospheric N deposition (Fenn et al., 2010; Tulloss and Cadenasso, 2015). Likewise, precipitation patterns also vary. The importance of comparative studies representing different but realistic potential environmental conditions has been recognized (Richards et al., 2006; Standish et al., 2012), yet few experiments have examined invasive and noninvasive species under low-resource conditions (Leishman and Thomson, 2005; Funk and Vitousek, 2007; Muth and Pigliucci, 2007). Using a functional-trait-based tactic for impact assessment, prediction, and management of invasive species’ under changing environmental conditions is an apt approach (Drenovsky et al., 2012a). Our design was chosen to test for potential competitive and physiological advantages invasive species may have as they grow during a finite rainy season in a mediterranean climate. Because our questions were related to growth over a determined length of time and not how functional traits compare at a given size, our study did not account for functional trait variation due to ontogeny (Coleman et al., 1994). The main goal of our study was to understand how functional traits differ among woody, mediterranean, invasive and phylogenetically

related, less-invasive species under varying environmental conditions during early establishment. Woody invasive species are studied much less frequently than herbaceous invasives yet they are as great of an ecological concern (Richardson and Rejmánek, 2011; fig. 1b in Hulme et al., 2013; Rejmánek, 2014). We quantified whether differences in growth and allocation patterns present under low nitrogen levels typically found in mediterranean wildlands and different watering conditions also remained under increased nitrogen and decreased water availability representing potential future environmental change. Under high-resource conditions, we expected invasive species to have greater height, leaf area, plant mass, and relative growth rate (RGR) than less-invasive species. We expected these differences to diminish under less-favorable conditions. We hypothesized that our mediterranean invasive species would have higher root mass ratios (RMR) than phylogenetically related, less-invasive species under the more stressful low-water conditions. We also hypothesized that invasive species would have greater log response ratios (ln RR) to increases in N and water (W) than less-invasive species and thus greater trait plasticity. Including increases in carbon (C) availability to the experiment would have given a more complete picture of the potential effects of global climate change on functional traits, but we did not have the space or capacity to include this aspect in our study. We quantified chlorophylls (chl) a and b and total carotenoids (Car) contents on a leaf area basis. We also examined the chl a/b and total Car/total chl ratios for a subset of our mediterranean contrasts to gain some insight on how pigments might contribute to growth under favorable conditions and to photoprotection under drought and/or nutrient stress (Demmig-Adams and Adams, 1996; Kitajima and Hogan, 2003). Additionally, we examined root growth and architecture for several contrasts.

MATERIALS AND METHODS Species selection and status—Species chosen (Table 1) were all woody horticultural species native to mediterranean climates outside of California. We did not use native California plants as noninvasive species within our contrasts because native does not necessarily equate to noninvasive elsewhere and we did not want to confound our study (Muth and Pigliucci, 2006; van Kleunen et al., 2010a). All species chosen were horticultural species with readily available commercial or wild-growing seeds and had been naturalized in California for at least 50 years. We used 50 years since naturalization to minimize any differences in time since introduction and to avoid the potential of a lag-time in invasiveness (Crooks and Soule, 1999; van Kleunen et al., 2010a). Our study focused on angiosperm trees and shrubs and incorporated congeneric or confamilial relationships into phylogenetically controlled contrasts (hereafter called “contrasts”). In this way, we examined differences between invasive and less-invasive species within clades (Felsenstein, 1985; Harvey and Pagel, 1991; Funk et al., 2015). Analyses did not take into account the magnitude of the differences in invasiveness within contrasts, and phylogenetic distances were not standardized. The contrasts used were a subset of those from Grotkopp et al. (2010) but included an additional invasive species, Caesalpinia pulcherrima, in the Caesalpinia contrast and an additional contrast formed by Pittosporum undulatum and Pittosporum tobira. Acacia, Acer, Brachychiton, Crataegus, and broom contrasts as well as Eucalyptus pauciflora were grown but had insufficient

A P R I L 2016 , V O LU M E 103

• E R S K I N E - O G D E N E T A L. — F U N C T I O N A L T R A I T S A N D W O O DY I N VA S I V E S P E C I E S

• 615

TABLE 1. Contrasts, species, invasive status (based on number of invasive reports; see footnote), and year grown.

Contrast Buddleja Eucalyptus 1

Morus Pittosporum Anacardiaceae

Berberis Caesalpinia

Eriobotrya Erythrina Eucalyptus 2

Eucalyptus 3 Lavandula Leptospermum

Species

Invasive statusa

Year

Buddleja davidii Franch. Buddleja globosa Hope Eucalyptus camaldulensis Dehnh. Eucalyptus pulverulenta Sims Eucalyptus nicholii Maiden & Blakely Morus alba L. Morus rubra L. Pittosporum undulatum Vent. Pittosporum tobira (Thunb.) W.T.Aiton Schinus molle L. Schinus terebinthifolia Raddi Searsia lancea (L. f.) F.A.Barkley Berberis thunbergii DC. Berberis koreana Palib. Caesalpinia gilliesii (Hook.) D.Dietr. Caesalpinia pulcherrima (L.) Sw. Caesalpinia cacalaco Humb. & Bonpl. Eriobotrya japonica (Thunb.) Lindl. Eriobotrya deflexa (Hemsl.) Nakai Erythrina crista-galli L. Erythrina coralloides DC. Eucalyptus camaldulensis Dehnh. Eucalyptus globulus Labill. Eucalyptus macrocarpa Hook. Eucalyptus conferruminata D.Carr & S.Carrab Eucalyptus baueriana Schauer Lavandula stoechas L. Lavandula angustifolia Mill. Leptospermum laevigatum (Gaertn.) F.Muell. Leptospermum lanigerum (Sol. ex Aiton) Sm.

2 1 2 0 0 2+ 1 2 2 2 2+ 1 2 0 2 1 0 2 0 2 0 2 2 0 2 0 2 1 2 0

2006 2006 2006 2006 2006 2006 2006 2006 2006 2007 2007 2007 2007 2007 2007 2007 2007 2007 2007 2007 2007 2007 2007 2007 2007 2007 2007 2007 2007 2007

Notes: Invasive species are indicated in bold text. a Invasive status was determined using Randall’s (2007) compendium and supplemented with expert opinion. Species with a 0 ranking are noninvasive, invasive potential increases to 2+, the most highly invasive species. b Commonly sold in California and South Africa as Eucalyptus lehmannii (Schauer) Benth.

germination to be included in all treatments and therefore have been excluded from the data analyses. Since all species could not be studied simultaneously due to space and labor limitations, we performed one experiment in 2006 with four contrasts of nine species and a similar experiment in 2007 with nine contrasts of 21 species. Richardson et al. (2000, p. 98) define invasive species as “naturalized plants that produce reproductive offspring, often in very large numbers, at considerable distances from parent plants…and thus have the potential to spread over a considerable area.” For our analyses, invasive woody species were those that are invasive in California or have been reported as clearly invasive in other states or regions of the world (Randall, 2007). While most species cannot be categorized as purely invasive or noninvasive, extremely invasive and noninvasive species have traits included in a suite of continuous characteristics at opposing ends of life-history spectra (Baker, 1965; Westoby, 1998; Reich et al., 2003; Wright et al., 2004; fig. 13.12 in Rejmánek et al., 2013; Stahl et al., 2013). For example, invasive species tend to have faster growth, shorter generation times, high propagule pressure, vegetative reproduction, and leaves with lower construction costs, while noninvasive species can have slower growth, longer juvenile periods, only sexual reproduction and at low rates, and more expensive leaves with greater longevity. Because these traits are numerous and mostly continuous, we will refer to our “noninvasive” species as “less-invasive”. The less-invasive species were those that, despite widespread planting, have not been reported as invasive anywhere, or only have limited, local establishment (Richardson et al., 2000).

We used Randall’s (2007) A Global Compendium of Weeds, a compilation of reports, both peer-reviewed and gray literature, to distinguish our invasive and less-invasive species. It includes the location of a species’ invasion and the level/type of invasiveness reported within each article. We ranked each species’ invasiveness based on the number of cited reports of invasiveness and the severity of its weed classification. Species were ranked from 0 to 2+, with 0 being noninvasive and 2+ being highly invasive. The species rankings were defined as follows: 0 (0–7 citations and not classified as “environmental weeds”), 1 (8–15 citations or fewer citations but classified as “environmental weeds’), 2 (16–24 citations) and 2+ (25 or more citations) (Table 1). This classification scheme allowed us to compare less-invasive species with more-invasive species within a contrast, but was not meant to give absolute categories of invasiveness. We recognize that A Global Compendium of Weeds (Randall, 2007) has, at times, uneven species coverage, redundant citations within references, and classifications that may not be uniformly applied to all reports. It also was considered the most comprehensive source on invasive species at the time of species selection. In the case of the Pittosporum contrast, in which both species have a ranking of “2” for invasiveness, P. tobira is considered much less invasive than P. undulatum (Rejmánek, 2013 and references therein). Status of all used species was also compared with records in the database of invasive trees and shrubs (Rejmánek and Richardson, 2013). Environmental conditions and experimental design—Seeds were

surface-sterilized with a diluted bleach solution (0.30% sodium

616

•


hypochlorite), stratified and planted in March 2006 or 2007 in Tall One Treepots (36 cm tall, 2.83 L; Stuewe and Sons, Corvallis, Oregon, USA) filled with sterilized UC Mix (see Grotkopp et al., 2010) with the following added per cubic meter (approximately 26.7 kg dry): 1.8 kg oyster lime, 1.8 kg dolomite 65, 1.8 kg single super phosphate. Timing of seedling emergence was recorded for all plants, and pots were thinned within a week to one seedling per pot. We used 8 ppm as our low N level because it was typical for low-resource natural areas isolated from some of the most direct sources of anthropogenic N deposition, such as agricultural run-off and artificial fertilization. Our high N level of 160 ppm approached that of the fertilized water that is routinely used in greenhouses (202 ppm) and was within the range of elevated N levels reported in California (Fenn et al., 2010). Pots were watered with either high N (160 ppm) or low N (8 ppm) solutions. Diluted GrowMore 4-18-38 No Boron (0.5% NH4+, 3.5% NO3−; National Research and Chemical, Gardena, California, USA) was used in conjunction with a Ca(NO3)2 solution to make the high N fertilizer solution, while even further-diluted GrowMore 4-18-38 No Boron and CaSO4 solutions constituted the low N fertilizer to achieve equivalent final mineral concentrations of Ca++ (238 ppm), P (36 ppm) and K (76 ppm) for both the high N (NH) and low N (NL) treatments. An N electrode was used to monitor the N concentrations in the irrigation water so that the N injector rates into the irrigation lines could be adjusted if needed. In mediterranean climates, seeds typically germinate during a cool and wet winter that is followed by a hot and dry summer. To simulate the rainy mild mediterranean winters and to ensure establishment, we hand-watered our seedlings to soil saturation as needed (when the top layer of soil was dry; typically every 2–3 d) until the experimental watering (W) treatments began. For the 2006 experiment, watering treatments began 60 d after emergence. After data collection for the first year, we were concerned that this timeframe was too late to detect differences in relative growth rates among species since RGR decreases considerably after maximal RGR, i.e., initial exponential growth (Hunt, 1982). For the 2007 experiment, we decreased the time of full watering before the W treatments commenced from 60 d to 30 d after emergence. Pots of each species at each N level were randomly distributed among benches in two adjacent greenhouses at the University of California, Davis. We randomly assigned each N/W treatment to two benches per greenhouse, with a total of four benches for each of the six N/W treatments. Water-level treatments were applied by hand-watering the pots to soil saturation in the high-watering (WH), moderate-watering (WM) and low-watering (WL) treatment blocks every 2 d, 5 d, and 10 d, respectively, until the plants were harvested at 90 d old (30 d after W treatment commenced) in 2006 (T2). We did not observe the intended levels of water stress and die back under the lowwatering/high-drought (WL) conditions in 2006, so we increased the differences among watering treatments for the 2007 experiment to watering every 3 d, 6 d, and 12 d, and we increased the duration of watering treatments to 45 d so that plants were harvested at 75 d after emergence in 2007 (T2). While our experimental conditions were realistic with respect to the extremes of N levels found in California, we were unable to withhold water from our seedlings for an extended time to mimic mediterranean summer drought without potentially killing most of the potted plants. With the roots unable to access a natural soil profile as they would in the field, we had to adapt the levels of drought to be appropriate for a greenhouse experiment. We had little to no plant mortality due to drought.

The 2006 greenhouses had min/max day temperatures of 19.4°/42.2°C and min/max night temperatures of 11.1°/23.3°C averaging 29.5° ± 0.23°C day and 18.3° ± 0.13°C night. The 2007 greenhouses had min/max day temperatures of 21.7°/41.7°C and min/max night temperatures of 13.9°/25.6°C averaging 31.9° ± 0.18°C day and 19.2° ± 0.09°C night temperatures. Growth analysis—Plants were harvested at two time points in each experiment. In 2006, plants were first harvested at 60 d, and for 2007 the initial harvest was at 30 d after emergence (here forth referred to as T1 for both experiments). For the first harvest (T1), 10–15 plants per species at each N level from both greenhouses were harvested without water stress to mimic the end of the mediterranean winter. The second harvest (T2) was done 30 d after watering treatments commenced in 2006 (90 d after emergence) and 45 d after watering treatments begin in 2007 (75 d after emergence). Typically, 5–6 plants/treatment/species were randomly harvested from each of the greenhouses for a total of 10–12 plants/treatment/ species harvested for the second time point. For both T1 and T2 harvests, plant height was measured (except Pittosporum), leaves were separated from each plant and leaf area (LA) was measured with either a LI-COR LI-3100 Area Meter (Lincoln, Nebraska, USA) or, if too small for accurate LI-COR measurements, through digital photos of leaves using ImageJ (Rasband, 1997–2014). Roots were cut at the root–shoot junction and carefully washed to remove soil. All roots, stems, and leaves were dried in a 65°C oven for a minimum of 48 h and then weighed. We did not include the cotyledon mass for the Eriobotrya contrast because these large storage cotyledons inconsistently fell from the plants. Due to poor germination, the Leptospermum contrast (2007) was only harvested at the first time interval (before W treatments commenced). We used data collected to calculate leaf area ratio (LAR: cm2 · g−1plant), specific leaf area (SLA: cm2 · g−1leaf), leaf mass ratio (LMR: gleaf · g−1plant), root mass ratio (RMR: groot · g−1plant), total plant mass (g), net assimilation rate (NAR: mgplant · cm−2leaf · d−1), relative growth rate (RGR: mgplant · g−1plant · d−1) and absolute growth rate (AGR: mgplant · day−1). For the first harvest (T1), 10–15 plants per species per nitrogen level from each greenhouse were used to estimate the initial biomass for calculating NAR, RGR, and AGR at each nitrogen level (James et al., 2009). RGR is generally considered most important at the earliest stages of plant growth (Hunt, 1982). As a plant’s total biomass increases, its rate of growth relative to its initially measured plant mass declines. On the other hand, absolute growth rate does not take existing biomass into account, just the amount of plant mass gained over a time period. While we recognize that variation in functional traits (e.g., RMR, LAR) can be due to ontogeny rather than plasticity (Coleman et al., 1994), our study focused on functional traits conferring invasiveness in nonnative species invading mediterranean climates where timing of growth is important for survival. Because of this, as well as space logistics within our greenhouses, we did not include an ontogenic component (functional growth curves with multiple harvests) to our data collection or analyses. Root study of Anacardiaceae, Berberis, and Lavandula—In 2007, five contrasts were randomly chosen for a separate root experiment. Because of poor germination and extreme root brittleness, three contrasts, Berberis, Lavandula, and Anacardiaceae were used for maximum root length, and only the Lavandula and Anacardiaceae contrasts produced reliable root morphology data. Plants were

A P R I L 2016 , V O LU M E 103


grown in TP616 Tree Pots (41 cm tall, 6.23 L; Stuewe and Sons) under the same NL and NH treatments as the main experiment and under WH conditions. Thirty days after emergence, 10 plants per N treatment had their pots cut open vertically and maximum root length was measured. Roots were washed carefully to avoid breaking the root system and were scanned with WinRHIZO 2003 (Instruments Régent, Quebec City, Quebec, Canada) to analyze root morphology and topology. We measured root length (cm), specific root length (m · g−1), average diameter (mm), volume (mm3), length per unit volume (mm−2), and numbers of tips and forks. After scanning, roots were dried at 65°C for a minimum of 48 h and weighed. Pigment analysis—In 2006, pigments were extracted from leaf tis-

sue with ammonical acetone and quantified using a spectrophotometer (Bausch & Lomb Spectronic 21, Rochester, New York, USA) reading absorbances at 480, 645, 663, and 710 nm to estimate amounts of chl a, chl b, total chl, and total Car (Hendry and Price, 1993) on a leaf area basis (pigment content). A hole-punch was used to cut the samples (two to four per plant) from leaves located at midheight on each plant so that leaves were all of approximately similar ages for pigment analysis. Leaves of eight plants per species for each N/W treatment were sampled at 0 and 10 d after W treatments began (combined into “early” time point due to small sample sizes) and at 20 and 30 d after W treatments commenced (combined into “later” time point). Statistical analyses—Most statistical analyses were performed separately for 2006 and 2007 because the W treatments and harvesting times differed substantially between the two years. To compare general patterns in growth and allocation traits between invasive and less-invasive species, we performed mixed-model restricted maximum-likelihood (REML) ANOVA with invasiveness (Inv; two levels), nitrogen (N; two levels), water (W; three levels; only at T2 as T1 had no water treatments), and all interactions as fixed factors. Our random factors were contrast, species nested within contrast (following Funk et al., 2015), greenhouse, and block (nested within N and greenhouse for T1; nested within N, W, and greenhouse for T2). We used contrast as a random factor because our 12–13 contrasts sample across the angiosperm phylogeny. Residuals of variables were tested for normality and homogeneity of variance. Data were transformed when necessary and the ANOVAs weighted by the inverse of the variance of the residuals to fit the assumptions of ANOVA (Neter et al., 1990). We used the REML method of ANOVA because our data were not completely balanced. Because we were not interested in all possible comparisons of means, we used a sequential Bonferroni correction (Sokal and Rohlf, 1995) to test for significance among the means of interest (e.g., between invasive and noninvasive species at each significant treatment level). For those interested, uncorrected within-contrast comparisons of trait responses between invasive and less-invasive species are available in Appendix S1 (see Supplemental Data with the online version of this article). For the root study, plants only were harvested at T1 and followed the method described above; fixed factors included invasiveness, N, and contrast in a full factorial design, and greenhouse as a random factor. For the pigment analyses, fixed factors included invasiveness, N, W, and contrast in a full factorial design; random factors were greenhouse and block nested within N, W, and greenhouse. JMP 10.0.0 (SAS Institute, Cary, North Carolina, USA) software was used for all statistical analyses.

• 617

We calculated natural log response ratios (ln RR) of contrasts to analyze the proportional changes in traits of invasive and lessinvasive species in response to increases in N (ln RRN) and in W (ln RRW) at T2. These ln RRs were calculated as: ln (mean at highresource level) – ln (mean at low-resource level) (Hedges et al., 1999; Liancourt et al., 2005) by contrasting paired high- and lowresource blocks from each greenhouse (two pairs per greenhouse) consistently for all analyses (n = 4 block sets). We examined species’ responses for plant mass, plant height, LA, LAR, SLA, LMR, and RMR to increased N and to increased W (using only the data for WH and WL). We then performed REML full factorial ANOVA as described previously to detect whether the invasive and the phylogenetically related, less-invasive species respond differently to N and W. Because contrast is a random factor and ln-transformed ratios were used, we combined both years of data into single analyses for each trait, for a total of two separate ANOVAs (variable = ln RRN; fixed factors = Inv, W, Inv × W) and (variable = ln RRw; fixed factors = Inv, N, Inv × N) per trait. Random factors were contrast, species nested within contrast, greenhouse, and block set nested within greenhouse. When an interaction was significant, we used a t test with a sequential Bonferroni correction (Sokal and Rohlf, 1995) because we were specifically looking for differences in responses to a single factor (either N or W) between invasive and less-invasive species.

RESULTS Height, plant mass, and leaf area T1—Overall, at T1 (before W

treatments began), invasive plants were significantly taller, heavier, and had greater total LA than less-invasive plants regardless of N level (height: 2006 F1, 3.5 = 10.03, P < 0.05, 2007 F1, 10.8 = 4.94, P < 0.05; mass: 2006 see below, 2007 F1, 11.0 = 8.22, P < 0.05; LA: 2006 F1, 101 = 6.89, P < 0.05, 2007 F1, 11.5 = 11.16, P < 0.01; Fig. 1; Appendix S2 for ANOVA table and Appendix S3 for means and SDs for all T1 data [see online Supplemental Data]) with the exception of mass in the 2006 contrasts. For the 2006 contrasts, there was a significant Inv × N interaction for plant mass (Fig. 1A; F1, 167.0 = 5.93, P < 0.05) because these invasive species had a greater proportional increase in mass for the NH treatment than did phylogenetically related, lessinvasive species (NL: P = NS; NH: P < 0.05; Bonferroni-corrected post hoc tests). Height T2—There were no differences between invasive and less-

invasive species at the most stressful WL treatment of the 2006 species, but as water increased, invasive species were significantly taller than less-invasive species regardless of N (Inv × W: F2, 547.6 = 4.23, P < 0.05) (Fig. 2A). For the 2007 species, invasive plants were taller than less-invasive species under NL conditions for all water treatments (Inv × N 2007: F1, 1058.0 = 6.37; P < 0.05, Bonferroni-corrected post hoc tests) (Fig. 2B). Plant mass and leaf area T2—At the second harvest (T2), the general trends observed were similar among the 2006 and 2007 contrasts (Fig. 2C–F; online Appendix S4 for ANOVA table and online Appendix S5 for means and SDs for all T2 data). Under the stressful NL conditions, plant mass and leaf area were low, with W having little or no effect regardless of invasive status. Under NH conditions, as W level increased to more favorable conditions, plant mass and LA were significantly higher for invasive species (at WM and WH

618

•


Root mass ratio—In general, NL

conditions resulted in 40–70% greater RMR than NH conditions at both harvests (N: T1 2006 F1, 21.9 = 523.46, T1 2007 F1, 42.4 = 111.37, T2 2006 F1, 585.0 = 1356.68; 2007 F1, 44.5 = 553.95; all P < 0.001) (Figs. 1G, 1H, 2G, 2H; online Appendices S2, S4). During early growth and establishment (T1 harvest), RMR was not significantly different between invasive and less-invasive species (Fig. 1G, H). However, at T2, invasive species had 20% greater RMR than their less-invasive counterparts, regardless of N condition (Inv: 2006 F1, 4.0 = 10.36, P < 0.05; Inv: 2007 F1, 10.2 = 10.65, P < 0.01) (Fig. 2G, H). For the 2007 contrasts, there was an Inv × W interaction (Inv × W: F2, 1067.0 = 3.92, P < 0.05); RMR was significantly higher for the invasive species under WL conditions (P < 0.01, Bonferroni-corrected post hoc tests) and WM conditions (P < 0.05, Bonferroni-corrected post hoc tests) regardless of N level (Fig. 2H). Growth rates and components—

For the 2006 and 2007 contrasts, there were no significant differences in RGR or NAR between invasive and less-invasive species, regardless of N or W (2006 RGR calculated from 60 to 90 d; 2007 RGR calculated from 30 to 75 d); 2007 interactions for RGR and NAR were no longer significant with Bonferroni-corrected post hoc tests; online Appendix S4). There were significant Inv × FIGURE 1 Overall (A, B) plant height (square-root transformed), (C, D) plant mass (square-root transformed), (E, N × W interactions for AGR in F) leaf area (square-root transformed) and (G, H) root mass ratio (RMR; angular transformation) for invasive and less-invasive plants harvested at time 1 (T1) under low nitrogen (Low N) and high nitrogen (High N) conditions both years (2006: F2, 564.1 = 3.66, in 2006 (A, C, E, G) and 2007 (B, D, F, H), respectively. Closed squares, 2006 invasive species; opened squares, 2007: F2, 945.7 = 2.99, P ≤ 0.05 for 2006 less-invasive species; closed circles, 2007 invasive species; opened circles, 2007 less-invasive species. All both; online Appendix S4). All data for figures were back-transformed for graphical purposes. Back transformations of standard errors (SEs) species had similar AGR at NL redo not necessarily visually reflect significant differences among data points for all figures. For summary ANOVA gardless of water and invasive status. At NH, as W increased, instatistics, see online Appendix S2; for means and SDs of traits, see online Appendix S3. vasive species had greater AGRs than did less-invasive species (WL: P = NS; WH: P < 0.05; Bonferroni-corrected post hoc tests, Fig. 2C–F; online Appendix S4) (N × W 2006 (mass: F 2, 25.8 = 65.36; both years). LA: F 2, 32.8 = 80.83; all P < 0.001); Inv × N 2006 (mass: F1, 569.8 = 19.97; For LAR, LMR, and SLA at T1, there were no significant differLA: F1, 569.0 = 33.78; both P < 0.001); Inv × N × W 2007 (mass: F 2, 1052.0 = ences between invasive and less-invasive species (online Appendix 5.07; LA: F2, 1043.0 = 6.57; both P < 0.01). Invasive species, in fact, S2; Bonferroni-corrected post hoc tests for 2007 SLA, Inv × N interhad 30–45% higher trait values than less-invasive species at WH action NS). At T2 less-invasive species had significantly greater (Fig. 2C–F).

A P R I L 2016 , V O LU M E 103


• 619

(Inv × N: NS, Bonferroni-corrected post hoc tests). SLA was not significantly different between invasive and less-invasive species in 2006 or 2007 (2006 Inv × N: NS, Bonferroni-corrected post hoc tests; 2007 Inv: all analyses NS). Log response ratios to increases in N and increases in W—Inva-

sive and less-invasive species responded similarly to increased N or increased W for plant mass, plant height, LA, LAR, SLA, LMR, and RMR. Pigment analyses—There were significant Contrast × Inv × N interactions at both time points for pigment contents (chl a, chl b, total chl, and total Car) measured on a leaf area basis (mmol · m−2leaf area) indicating that individual contrasts behaved differently with regard to pigment contents (F3, 358–410 = 9.91–37.76, all P < 0.001). Within the Buddleja contrast, at both time points, invasive and less-invasive species had similar chl a, chl b, and total chl contents at both N levels (NS), but lessinvasives had greater Car contents at NL only (Inv × N: early, F1, 96 = 6.79, P < 0.01; late, F1, 76 = 21.62, P < 0.001). For the Eucalyptus 1 and Morus contrasts, invasive species had more chl a, chl b, total chl, and Car at NH for both times (Inv × N: F1, 66–131 = 9.98–83.04, all P < 0.001). The invasive Pittosporum species had more chl a, chl b and total chl on a leaf area basis than the lessFIGURE 2 Overall (A, B) plant height (square-root transformed), (C, D) plant mass (square-root transformed), (E, F) leaf area (square-root transformed) and (G, H) root mass ratio (RMR; angular transformation) for invasive and less- invasive species for the early time invasive plants harvested at T2 under all environmental treatment conditions in 2006 (A, C, E, G) and 2007 (B, D, F, at NL (Inv × N: F1, 55.4–95 = 7.66– H), respectively. Closed squares, 2006 low nitrogen, invasive species; opened squares, 2006 low nitrogen, less- 36.03, P < 0.01). On the other invasive species; closed circles, 2007 low nitrogen, invasive species; opened circles, 2007 low nitrogen, less-invasive hand, at the later time point, the species; closed triangles 2006 and 2007 high nitrogen, invasive species; open triangles, 2006 and 2007 high nitro- invasive species had greater piggen, less-invasive species. All data for figures back-transformed for graphical purposes. Back transformations of ments at NH (Inv × N: early, F1, 96 = SEs do not necessarily visually reflect significant differences among data points for all figures. For summary 6.67–10.39, P < 0.05; late, F1, 83 = 29.92–30.29, all P < 0.001). The ANOVA statistics, see online Appendix S4; for means and SDs of traits, see online Appendix S5. invasive Pittosporum species had LAR than invasive species did regardless of N or W for the 2007 greater Car content than did the less-invasive species under all concontrasts (Inv: F1, 17.1 = 6.57, P < 0.05; online Appendix S4). In 2006, ditions at both time points (Inv: F1, 55.4–66.0 = 15.21–36.05, P < 0.001). under NL, less-invasive species also had significantly higher LAR We also analyzed chl a/b and total Car/total chl to gain an unthan did invasive species, while at NH, trait values were similar derstanding of the photoprotective aspects of pigments (Demmig(Inv × N: F1, 585.0 = 36.94, P < 0.001). Less-invasive species had higher Adams and Adams, 1996). Both chl a/b and chl/Car had significant LMR overall than invasive species did in 2007 (Inv: F1, 10.0 = 7.71, Contrast × Inv × N (or W) interactions at both the early and late P < 0.05), but there were no differences within the 2006 contrasts time points (data not shown). Contrasts showed different patterns:

620

•


sometimes the invasive species having greater ratios of pigments, sometimes the less-invasive species having the greater ratio, and sometimes no significant difference at all (data not shown). Root analyses—Invasive species in the two contrasts analyzed for

root morphology characteristics, Anacardiaceae and Lavandula, had more potential foraging ability with significantly more total forks (R2 = 0.770, F1, 57 = 7.55, P < 0.01) and tips (R2 = 0.628, F1, 56.1 = 22.15, P < 0.001) than their less-invasive counterparts. There was a significant Inv × N × Contrast interaction for average root diameter (R2 = 0.910, F1, 57 = 12.78, P < 0.001) The less-invasive Anacardiaceae species had a significantly lower average root diameter than its invasive counterparts, but only under NL (Tukey post hoc: P < 0.001). The less-invasive Lavandula species had a significantly greater average root diameter than its invasive counterpart regardless of N treatment (Tukey post hoc: P < 0.01). Invasive and less-invasive species did not differ in specific root length regardless of N condition or contrast, but the less-invasive species in the Anacardiaceae, Lavandula, and Berberis had greater maximum root length (R2 = 0.694, F1, 128.3 = 17.36, P < 0.001) than their invasive counterparts did. Total root length was significantly higher under NH treatments regardless of contrast or invasive status (R2 = 0.685, F1, 57 = 14.329, P < 0.001). Both contrasts had significantly greater length per unit volume under NH regardless of species invasiveness (R2 = 0.614, F1, 57 = 18.3429, P < 0.001). DISCUSSION Changing climate, weather patterns, and anthropogenic resource additions such as N, CO2, and methane, are presently altering the landscape and the makeup of plant communities worldwide. Species that are able to exploit increases in nitrogen input can maintain or achieve superior functional trait values, have higher survival rates and are more prolific than their counterparts (Dawson et al., 2012; Vallano et al., 2012; Porter et al., 2013; Field et al., 2014). We found that with increased nitrogen input, the invasive, mediterranean, woody species had greater biomass and total leaf area than did their less-invasive counterparts. During early growth, invasive species were taller regardless of N availability. While other studies also found that invasive species had greater size (i.e., plant mass, total leaf area, height) than less-invasive species in both favorable and stressful environments, these studies focused on a mixture of life forms, did not always correct for phylogeny and sometimes included comparisons with native species (van Kleunen et al., 2010b; Schlaepfer et al., 2010; Godoy et al., 2012). Advantages are only biologically meaningful if they are related to increases in fitness. While we did not directly measure fitness, biomass is often used as a strong indicator or proxy for fitness (Richards et al., 2006; Davidson et al., 2011). Greater biomass, LA, and height can give plants a competitive edge by shading out neighbors and increasing light capturing abilities for further growth (Weiner and Thomas, 1986; Falster and Westoby, 2003; Morris et al., 2011), especially during intermittent favorable wet winters. As we expected, differences in biomass, leaf area, and height between invasive and closely related less-invasive species became smaller as conditions became more stressful (decreasing W and especially under NL), but under no conditions did the less-invasive species have significantly greater biomass, LA, or height than their invasive counterparts.

More stressful environmental conditions are expected over the next century, as the typical mediterranean climate of a wet winter season followed by a long dry summer is predicted to become more extreme (Sala et al., 2000; Hayhoe et al., 2004; Walther et al., 2009; Diffenbaugh et al., 2015). Root allocation, specifically RMR, along with other traits such as a variety of hydraulic characteristics, fog moisture utilization, and canopy dormancy patterns, are especially important functional traits for perennial species’ establishment and survival in these mediterranean climates (Morris et al., 2011; West et al., 2012). In general, we found that RMR increased significantly under NL and with decreasing W. The older invasive seedlings (T2) had greater RMR than their less-invasive counterparts did under WL conditions, regardless of N, a trait that may make them more stress tolerant and able to survive over extended periods of drought. These results are typical for invasive mediterranean species as well as drought-tolerant species (Drenovsky et al., 2008), but may not apply to invasive species in general (Hastwell and Panetta, 2005; van Kleunen et al., 2011; Dawson et al., 2012). While the invasive species studied did not have greater RMR than their less-invasive counterparts as very young seedlings (T1 harvest), this result could be because the seedlings were still quite small. For the two contrasts included in the root morphology study (Anacardiaceae and Lavandula), the invasive species had significantly more root tips and forks than their less-invasive counterparts and could indicate greater foraging capabilities for these invasive species (Drenovsky et al., 2008; Keser et al., 2014). While we found significant size and root allocation differences between our invasive and less-invasive species, we did not find this to be true with a number of performance traits tested. Our results for leaf allocation traits did not agree with a meta-analysis, which showed that invasive species had greater values for traits related to performance such as physiology, leaf and shoot allocation, growth rate, size, and fitness (van Kleunen et al., 2010b). These studies included nonwoody species with different growth forms from a wide range of climates and various experimental conditions. We found, in our study of mediterranean woody species, no difference in SLA between invasive and less-invasive species. In fact the less-invasive species sometimes had higher leaf allocation traits such as LMR and LAR than their invasive counterparts. Lower SLA is often associated with more arid conditions because the leaves are often thicker, denser, and more drought tolerant (Stahl et al., 2013). Invasive species in harsher climates, such as those in mediterranean regions, may not have the same traits as invasive species from more mesic habitats. Lower leaf allocation is also consistent with the increases we found in RMR, a trait associated with more arid conditions or extended droughts. For invasive species in mediterranean climates, establishing an extensive root system before the end of the mild, rainy winter is critical for survival in the dry seasons ahead (Morris et al., 2011). This trade-off between leaf and root allocation may explain why invasive and less-invasive species did not differ significantly in SLA. It also may explain why we found no consistent trends for LMR and LAR in our mediterranean species (Grotkopp et al., 2010; Godoy et al., 2011, 2012). Physiological advantages involving pigments were exhibited by mediterranean invasive species relative to their less-invasive counterparts in the subset of contrasts studied. While both the invasive and less-invasive species acclimated to stressful conditions by altering chl a/b and Car/chl ratios, the invasive species in three of the four contrasts studied (Eucalyptus I, Morus, and Pittosporum) contained more carotenoids and chlorophyll on a leaf area basis than

A P R I L 2016 , V O LU M E 103


did less-invasive species at high N. More carotenoids may aid these invasive species with photoprotection under stressful conditions such as high light and drought (Zinnert, et al., 2013; Oliveira et al., 2014). In environments with elevated N, increased chlorophyll content may benefit invasive species by increasing photosynthetic rates per leaf area (though not necessarily on a per mole chlorophyll basis) (Pattison et al., 1998; McDowell, 2002; Kitajima and Hogan, 2003). The role of greater pigment amounts in aiding invasive species as a mechanism of increased growth warrants further investigation as our study was limited to only four contrasts. While we found no difference in NAR between invasive and less-invasive species for all of our contrasts studied, this is not uncommon when calculated over a long time interval. Direct gas exchange and other biophysical measurements would have been much more sensitive to differences in photosynthesis (Shen et al., 2011). Relative growth rate (RGR) has been associated with invasiveness in previous studies of ours (Grotkopp and Rejmánek, 2007; Grotkopp et al., 2010), as well as a meta-analysis by van Kleunen (2010b). Contrary to expectations, we found no difference in RGR between invasive, mediterranean, woody species and their lessinvasive counterparts for the times tested. This result is not surprising, considering maximal RGR (when plants are growing at their fastest, exponential rate) generally occurs shortly after germination (Hunt, 1982), after which, RGR slows down considerably. Our plants were harvested as late as 90 d after emergence, long past maximal RGR. Another factor that likely played a major role in our nonsignificant results is the trade-off between water-use efficiency (WUE) and RGR. When mediterranean species germinate in mild, wet winters, a high RGR can be important for the competitive advantages of greater biomass, height, and leaf area in these early stages of growth. However, as conditions become drier during the spring and summer months, increases in WUE become essential at the expense of RGR. This trade-off between WUE and RGR has been shown in woody species (Stahl et al., 2013), forbs (Drenovsky et al., 2012b), and desert annuals (Huxman et al., 2013). What does remain, however, is that invasive species are significantly larger, in terms of plant mass, total leaf area, and early on, height. We, therefore, infer that invasive species had a significantly higher maximal RGR during the earliest stages of growth (Grotkopp et al., 2010; Dawson et al., 2011) before the watering treatments began, resulting in greater overall mass and leaf area. Larger size confers competitive advantages to invasive species, especially during intermittent, favorable, wet winters. Larger size can augment solar interception, shade out neighbors, increase reproductive rates, and increase root-mining capabilities for water and nutrients (Weiner and Thomas, 1986; Falster and Westoby, 2003; Morris et al., 2011). Invasive species consistently had higher AGR than phylogenetically related less-invasive species across all experimental treatments. Because invasive species are already larger to begin with, and they are gaining more biomass on an absolute scale (AGR), they will remain bigger even if their relative growth rates at that time do not differ from their less-invasive counterparts. Future studies that include multiple harvests to produce functional growth curves would elucidate differences, if any, in growth trajectories or mechanisms between invasive and lessinvasive species while controlling for biomass and ontogeny (Coleman et al., 1994). For our study, we were not concerned with using biomass as a covariable when examining traits such as RMR and LAR even though they may be correlated (either inversely or positively) with biomass. We were quantifying changes in trait values

• 621

based on time since germination, an important factor when comparing species germinating in mediterranean climates with very dry summer seasons. Clearly, species that germinate first will have a head start on accumulating biomass (getting big), but at the risk of germinating during an isolated early rain. Research has found, in fact, that some mediterranean invasive species germinate before the native species, giving them the added advantage of a phenological priority effect as well (Wolkovich and Cleland, 2011). During extended drought, large size and leaf area can be a disadvantage depending on what other drought tolerance strategies are present (isohydry, fog moisture utilization, rooting depth). Extensive field studies would be needed to elucidate the other potential drought tolerance strategies these invasive species could use under more severe drought conditions beyond increased RMR (see West et al. (2012). The question then remains whether these differences in growth and allocation patterns between non-native invasive and less-invasive mediterranean woody species are related to trait plasticity within a species. Researchers have quantified plasticity of functional traits such as biomass, survival, flowering traits, WUE, and root mass ratio in response to changes in environmental factors. Several studies have found plasticity of these traits to be greater for invasive species than for less-invasive species or their native counterparts (reviewed by Richards et al., 2006; meta-analysis by Davidson et al., 2011). Other studies have found no difference or inconsistent differences in plasticity of traits between invasive and presumably less-invasive species (Hastwell and Panetta, 2005; Godoy et al., 2011; van Kleunen et al., 2011; Drenovsky et al., 2012b). We found that across the evaluated resource gradients, trait plasticity, measured as log response ratios, was similar between invasive and less-invasive species for all traits. With environmental change, greater responses are not necessarily advantageous; as conditions become more stressful, maintaining performance and fitnessrelated trait values may be more important for survival, i.e., fitness homeostasis (Rejmánek, 2000; Richards et al., 2006). Species native to mediterranean climates are generally adapted to more stressful, nutrient-poor (low N) conditions, with reliable winter rains and a prolonged summer drought. These weather patterns are no longer consistent because of global climate change. Mediterranean climates are becoming more stochastic, consisting of sporadic El Niño years with heavy rainfall scattered among more frequent and severe drought years. Inconsistencies in weather patterns and variability of N deposition may be the key to invasive species’ success in these regions; they have the ability to take advantage of additional resources such as anthropogenic N inputs, resulting in increases in biomass, leaf area, height, and RMR during favorable wet years (Richards et al., 2006; Dawson et al., 2012). They can then maintain growth and survive under more stressful conditions and have significantly greater root mass ratios than their less-invasive counterparts. Most plants in arid regions are N limited when moderate and favorable precipitation patterns occur, but are water limited at the lowest precipitation levels, supporting this conclusion (Yahdjian et al., 2011). Our data show that mediterranean, woody, invasive species clearly have higher trait values for plant mass and leaf area under higher N and more favorable water conditions than their less-invasive counterparts. They have greater absolute growth, higher RMR and, at times, height under both high and low nitrogen conditions. This combination of traits (plant mass, leaf area, RMR, height, and AGR) forms a fundamental difference between invasive and less-invasive, mediterranean, woody

622

•


species under many of our experimental conditions and potential future wildland conditions. Our results emphasize the importance of examining multiple traits in a variety of environments and testing different life forms when distinguishing invasive and less-invasive species. The functional traits important for mediterranean, invasive, woody species were different from those for invasive herbaceous species as well as for species invading mesic environments. Because of the breadth of families covered by our experiments, information gained from our research could be included in screening procedures of potential horticultural woody species in mediterranean environments as global environmental change modifies species distributions (Daehler et al., 2004; Leung et al., 2012; Meier et al., 2012). ACKNOWLEDGEMENTS The authors thank all the undergraduates who helped in the greenhouse. Special thanks go to P. Riley, T. Metcalfe, and M. Bower for logistical help; R. Drenovsky and P. Ludwig for valuable comments on earlier drafts; and R. Brandon Pratt, Amy McPherson, and three anonymous reviewers for helpful comments. This work was funded by a grant from the USDA-CSREES UC-IPM Exotic/Invasive Pests and Diseases Research Program (Project #05XN027) and supported by the University of California Agricultural Experiment Station. LITERATURE CITED Baker, H. G. 1965. Characteristic and modes of origin of weeds. In H. G. Baker and G. L. Stebbins [eds.], The genetics of colonizing species, 147–168. Academic Press, New York, New York, USA. Bossdorf, O., H. Auge, L. Lafuma, W. E. Rogers, E. Siemann, and D. Prati. 2005. Phenotypic and genetic differentiation between native and introduced plant populations. Oecologia 144: 1–11. Bradley, B. A., D. M. Blumenthal, D. S. Wilcove, and L. H. Ziska. 2010. Predicting plant invasions in an era of global change. Trends in Ecology & Evolution 25: 310–318. Burns, J. H., S. L. Halpern, and A. A. Winn. 2007. A test for a cost of opportunism in invasive species in the Commelinaceae. Biological Invasions 9: 213–225. Chapin, F. S. III, E. S. Zavaleta, V. T. Erviner, R. L. Naylor, P. M. Vitousek, H. L. Reynolds, D. U. Hooper, et al. 2000. Consequences of changing biodiversity. Nature 405: 234–242. Coleman, J. S., K. D. M. McConnaughay, and D. D. Ackerly. 1994. Interpreting phenotypic variation in plants. Trends in Evolution and Ecology 9: 187–191. Crooks, J. A., and M. E. Soule. 1999. Lag times in population explosions of invasive species: Causes and implications. In O. T. Sandlund, P. J. Schei, and A. Viken [eds.], Invasive species and biodiversity management, 103–125. Kluwer, Dordrecht, Netherlands. Daehler, C. C. 2003. Performance comparisons of co-occurring native and alien invasive plants: Implications for conservation and restoration. Annual Review of Ecology, Evolution, and Systematics 34: 183–211. Daehler, C. C., J. S. Denslow, S. Ansari, and H. C. Kuo. 2004. A risk-assessment system for screening out invasive pest plants from Hawaii and other Pacific Islands. Conservation Biology 18: 360–368. Dawson, W., M. Fischer, and M. van Kleunen. 2011. The maximum relative growth rate of common UK plant species is positively associated with their global invasiveness. Global Ecology and Biogeography 20: 299–306. Dawson, W., R. P. Rohr, M. van Kleunen, and M. Fischer. 2012. Alien plant species with a wider global distribution are better able to capitalize on increased resource availability. New Phytologist 194: 859–867. Davidson, A. M., M. Jennions, and A. B. Nicotra. 2011. Do invasive species show higher phenotypic plasticity than native species and, if so, is it adaptive? A meta-analysis. Ecology Letters 14: 419–431.

Davis, M. A., J. P. Grime, and K. Thompson. 2000. Fluctuating resources in plant communities: A general theory of invisibility. Journal of Ecology 88: 528–534. Demmig-Adams, B., and W. W. Adams III. 1996. The role of xanthophyll cycle carotenoids in the protection of photosynthesis. Trends in Plant Science 1: 21–26. Diez, J. M., C. M. D’Antonio, J. S. Dukes, E. D. Grosholz, J. D. Olden, C. J. B. Sorte, D. M. Blumenthal, et al. 2012. Will extreme climatic events facilitate biological invasions? Frontiers in Ecology and the Environment 10: 249–257. Diffenbaugh, N. S., D. L. Swain, and D. Touma. 2015. Anthropogenic warming has increased drought risk in California. Proceedings of the National Academy of Sciences, USA 112: 3931–3936. Dostál, P., W. Dawson, M. van Kleunen, L. H. Keser, and M. Fischer. 2013. Central European plant species from more productive habitats are more invasive at a global scale. Global Ecology and Biogeography 22: 64–72. Drenovsky, R. E., B. J. Grewell, C. M. D’Antonio, J. L. Funk, J. J. James, N. Molinari, I. M. Parker, and C. L. Richards. 2012a. A functional trait perspective on plant invasion. Annals of Botany 110: 141–153. Drenovsky, R. E., A. Khasanova, and J. J. James. 2012b. Trait convergence and plasticity among native and invasive species in resource-poor environments. American Journal of Botany 99: 629–639. Drenovsky, R. E., C. E. Martin, M. R. Falasco, and J. J. James. 2008. Variation in resource acquisition and utilization traits between native and invasive perennial forbs. American Journal of Botany 95: 681–687. Dukes, J. S. 2011. Climate change. In D. Simberloff and M. Rejmánek [eds.], Encyclopedia of biological invasions, 113–117. University of California Press, Berkeley, California, USA. Falster, D. S., and M. Westoby. 2003. Plant height and evolutionary games. Trends in Ecology & Evolution 18: 337–343. Felsenstein, J. 1985. Phylogenies and the comparative method. American Naturalist 125: 1–15. Fenn, M. E., E. B. Allen, S. B. Weiss, S. Jovan, L. H. Geiser, G. S. Tonnesen, R. F. Johnson, et al. 2010. Nitrogen critical loads and management alternatives for N-impacted ecosystems in California. Journal of Environmental Management 91: 2404–2423. Field, C. D., N. B. Dise, R. J. Payne, A. J. Britton, B. A. Emmett, R. C. Helliwell, S. Hughes, et al. 2014. The role of nitrogen deposition in widespread plant community change across semi-natural habitats. Ecosystems (New York, N.Y.) 17: 864–877. Fowler, D., C. E. R. Pitcairn, M. A. Sutton, C. Flechard, B. Loubet, M. Coyle, and R. C. Munro. 1998. The mass budget of atmospheric ammonia in woodland within 1 km of livestock buildings. Environmental Pollution 102 (Supplement 1): 343–348. Funk, J. L., C. S. Rakovski, and J. M. Macpherson. 2015. On the analysis of phylogenetically paired designs. Ecology and Evolution 5: 940–947. Funk, J. L., and P. M. Vitousek. 2007. Resource-use efficiency and plant invasion in low-resource systems. Nature 446: 1079–1081. Godoy, O., F. Valladares, and P. Castro-Díez. 2011. Multispecies comparison reveals that invasive and native plants differ in their traits but not in their plasticity. Functional Ecology 25: 1248–1259. Godoy, O., F. Valladares, and P. Castro-Díez. 2012. The relative importance for plant invasiveness of trait means, and their plasticity and integration in a multivariate framework. The New Phytologist 195: 912–922. González-Muñoz, N., J. C. Linares, P. Castro-Díez, and U. Sass-Klassen. 2014. Predicting climate change impacts on native and invasive tree species using radial growth and twenty-first century climate scenarios. European Journal of Forest Research 133: 1073–1086. Grotkopp, E., J. Erskine-Ogden, and M. Rejmánek. 2010. Assessing potential invasiveness of woody horticultural plant species using seedling growth rate traits. Journal of Applied Ecology 47: 1320–1328. Grotkopp, E., and M. Rejmánek. 2007. High seedling relative growth rate and specific leaf area are traits of invasive species: phylogenetically independent contrasts of woody angiosperms. American Journal of Botany 94: 526–532. Harvey, P. H., and M. D. Pagel. 1991. The comparative method in evolutionary biology. Oxford University Press, Oxford, UK. Hastwell, G. T., and F. D. Panetta. 2005. Can differential responses to nutrients explain the success of environmental weeds? Journal of Vegetation Science 16: 77–84.

A P R I L 2016 , V O LU M E 103


Hayhoe, K., D. Cayan, C. B. Field, P. C. Frumhoff, E. P. Maurer, N. L. Miller, S. C. Moser, et al. 2004. Emissions pathways, climate change, and impacts on California. Proceedings of the National Academy of Sciences, USA 101: 12422–12427. Hedges, L. V., J. Gurevitch, and P. S. Curtis. 1999. The meta-analysis of response ratios in experimental ecology. Ecology 80: 1150–1156. Hendry, G. A. F., and A. H. Price. 1993. Stress indicators: Chlorophylls and carotenoids. In G. A. F. Hendry and J. P. Grime [eds.], Methods in comparative ecology: A laboratory manual, 148–152. Chapman and Hall, London, UK. Hobbs, R. J., and L. F. Huenneke. 1992. Disturbance, diversity and invasion— Implications for conservation. Conservation Biology 6: 324–337. Huenneke, L. F., S. P. Hamburg, R. Koide, H. A. Mooney, and P. M. Vitousek. 1990. Effects of soil resources on plant invasion and community structure in California serpentine grassland. Ecology 71: 478–491. Hulme, P. E., P. Pyšek, V. Jarošik, J. Pergl, U. Schaffner, and M. Vilà. 2013. Bias and error in understanding plant invasion impacts. Trends in Ecology & Evolution 28: 212–218. Hunt, R. 1982. Plant growth curves: The functional approach to plant growth analysis. Thomson Litho, East Kilbride, UK. Huxman, T. E., S. Kimball, A. L. Angert, J. R. Gremer, G. A. Barron-Gafford, and D. L. Venable. 2013. Understanding past, contemporary, and future dynamics of plants, populations, and communities using Sonoran Desert winter annuals. American Journal of Botany 100: 1369–1380. IPCC. 2014. Summary for policymakers. In C. B. Field, V. R. Barros, D. J. Dokken, K. J. Mach, M. D. Mastrandrea, T. E. Bilir, M. Chatterjee, et al. [eds.], Climate change 2014: Impacts, adaptation, and vulnerability. Part A: Global and sectoral aspects. Contribution of Working Group II to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change, 1–32. Cambridge University Press, Cambridge, UK. James, J. J., J. M. Mangold, R. L. Sheley, and T. Svejcar. 2009. Root plasticity of native and invasive Great Basin species in response to soil nitrogen heterogeneity. Plant Ecology 202: 211–220. Keser, L. H., W. Dawson, Y.-B. Song, F.-H. Yu, M. Fischer, M. Dong, and M. van Kluenen. 2014. Invasive clonal plant species have a greater root-foraging plasticity than non-invasive ones. Oecologia 174: 1055–1064. Kitajima, K., and K. P. Hogan. 2003. Increases of chlorophyll a/b ratios during acclimation of tropical woody seedlings to nitrogen limitation and high light. Plant, Cell & Environment 26: 857–865. Leishman, M. R., and V. P. Thomson. 2005. Experimental evidence for the effects of additional water, nutrients and physical disturbance on invasive plants in low fertility Hawkesbury Sandstone soils, Sydney, Australia. Journal of Ecology 93: 38–49. Leung, B., N. Roura-Pascual, S. Bacher, J. Heikkilä, L. Brotons, M. A. Burgman, K. Dehnen-Schmutz, et al. 2012. TEASIng apart alien species risk assessments: A framework for best practices. Ecology Letters 15: 1475–1493. Liancourt, P., R. M. Callaway, and R. Michalet. 2005. Stress tolerance and competitive-response ability determine the outcome of biotic interactions. Ecology 86: 1611–1618. Ma, G., Y. Wang, X. Bao, Y. Hu, Y. Liu, L. He, T. Wang, and F. Meng. 2015. Nitrogen pollution characteristics and source analysis using the stable isotope tracing method in Ashi River, northeast China. Environmental Earth Sciences 73: 4831–4839. McDowell, S. C. L. 2002. Photosynthetic characteristics of invasive and noninvasive species of Rubus (Rosaceae). American Journal of Botany 89: 1431–1438. Meier, E. S., H. Lischke, D. R. Schmatz, and N. E. Zimmermann. 2012. Climate, competition and connectivity affect future migration and ranges of European trees. Global Ecology and Biogeography 21: 164–178. Morris, T. L., K. J. Esler, N. N. Barger, S. M. Jacobs, and M. D. Cramer. 2011. Ecophysiological traits associated with the competitive ability of invasive Australian acacias. Diversity & Distributions 17: 898–910. Muth, N. Z., and M. Pigliucci. 2006. Traits of invasives reconsidered: Phenotypic comparisons of introduced invasive and introduced noninvasive plant species within two closely related clades. American Journal of Botany 93: 188–196.

• 623

Muth, N. Z., and M. Pigliucci. 2007. Implementation of a novel framework for assessing species plasticity in biological invasions: Responses of Centaurea and Crepis to phosphorus and water availability. Journal of Ecology 95: 1001–1013. Neter, J., W. Wasserman, and M. H. Kutner. 1990. Applied linear statistical models: Regression, analysis of variance and experimental design, 3rd ed. Irwin, Homewood, Illinois, USA. Ochoa-Hueso, R., E. B. Allen, C. Branquinho, C. Cruz, T. Dias, M. E. Fenn, E. Manrique, et al. 2011. Nitrogen deposition effects on Mediterraneantype ecosystems: An ecological assessment. Environmental Pollution 159: 2265–2279. Oliveira, T. M., V. Matzek, C. D. Medeiros, R. Rivas, H. M. Falcão, and M. G. Santos. 2014. Stress tolerance and ecophysiological ability of an invader and a native species in a seasonally dry tropical forest. PLoS One 9: e105514. Pardo, L. H., M. E. Fenn, C. L. Goodale, L. H. Geiser, C. T. Driscoll, E. B. Allen, J. S. Baron, et al. 2011. Effects of nitrogen deposition and empirical nitrogen critical loads for ecoregions of the United States. Ecological Applications 21: 3049–3082. Pattison, R. R., G. Goldstein, and A. Ares. 1998. Growth, biomass allocation and photosynthesis of invasive and native Hawaiian rainforest species. Oecologia 117: 449–459. Porter, E. M., W. D. Bowman, C. M. Clark, J. E. Compton, L. H. Pardo, and J. L. Soong. 2013. Interactive effects of anthropogenic nitrogen enrichment and climate change on terrestrial and aquatic biodiversity. Biogeochemistry 114: 93–120. Randall, R. P. 2007. A global compendium of weeds [online]. Website http:// www.hear.org/gcw/ [accessed 16 June 2009]. Rapacciuolo, G., S. P. Maher, A. C. Schneider, T. H. Talisin, M. D. Jabis, R. E. Walsh, K. J. Iknayan, et al. 2014. Beyond a warming fingerprint: Individualistic biogeographic responses to heterogeneous climate change in California. Global Change Biology 20: 2841–2855. Rasband, W. S. 1997–2014. ImageJ. Computer program and documentation distributed by the author, website http://imagej.nih.gov/ij/ [accessed 04 April 2006]. Reich, P. B., I. J. Wright, J. Cavendare-Bares, J. M. Craine, J. Oleksyn, M. Westoby, and M. B. Walters. 2003. The evolution of plant functional variation: Traits, spectra, and strategies. international Journal of Plant Sciences 164 (Supplement): S143–S164. Rejmánek, M. 2000. Invasive plants: Approaches and predictions. Austral Ecology 25: 497–506. Rejmánek, M. 2013. Victorian box: An invasive tree spreading in California. Cal-IPC News 20 (4) & 21 (1): 8–9 [online]. Website http://www.cal-ipc.org/ resources/news/pdf/Cal-IPC_News_Winter2013.pdf [accessed 09 September 2015]. Rejmánek, M. 2014. Invasive trees and shrubs: Where do they come from and what we should expect in the future? Biological Invasions 16: 483–498. Rejmánek, M., P. Pyšek, and D. Richardson. 2013. Plant invasions and invasibility of plant communities. In E. van der Maarel and J. Franklin [eds.], Vegetation ecology, 2nd ed., 387–424. John Wiley, Oxford, UK. Rejmánek, M., and D. Richardson. 2013. Trees and shrubs as invasive alien species—2013 update of the global database. Diversity & Distributions 19: 1093–1094. Richards, C. L., O. Bossdorf, N. Z. Muth, J. Gurevitch, and M. Pigliucci. 2006. Jack of all trades, master of none? On the role of phenotypic plasticity in plant invasions. Ecology Letters 9: 981–993. Richardson, D. M., P. Pyšek, M. Rejmánek, M. G. Barbour, F. D. Panetta, and C. J. West. 2000. Naturalization and invasion of alien plants: Concepts and definitions. Diversity & Distributions 6: 93–107. Richardson, D. M., and M. Rejmánek. 2011. Trees and shrubs as invasive alien species—A global review. Diversity & Distributions 17: 788–809. Sala, O. E., F. S. Chapin III, J. J. Armesto, E. Berlow, J. Bloomfield, R. Dirzo, E. Huber-Sanwald, et al. 2000. Global biodiversity scenarios for the year 2100. Science 287: 1770–1774. Schlaepfer, D. R., M. Glättli, M. Fischer, and M. van Kleunen. 2010. A multispecies experiment in their native range indicates pre-adaptation of invasive alien plant species. New Phytologist 185: 1087–1099. Sharma, G. P., S. A. Muhl, K. J. Esler, and S. J. Milton. 2010. Competitive interactions between the alien invasive annual grass Avena fatua and indigenous herbaceous plants in South African Renosterveld: The role of nitrogen enrichment. Biological Invasions 12: 3371–3378.

624

•


Shen, X. Y., S. L. Peng, B. M. Chen, J. X. Pang, L. Y. Chen, H. M. Xu, and Y. P. Hou. 2011. Do higher resource capture ability and utilization efficiency facilitate the successful invasion of native plants? Biological Invasions 13: 869–881. Sokal, R. R., and F. J. Rohlf. 1995. Biometry, 3rd ed. W. H. Freeman, New York, New York, USA. Stahl, U., J. Kattge, B. Reu, W. Voigt, K. Ogle, J. Dickie, and C. Wirth. 2013. Whole-plant trait spectra of North American woody plant species reflect fundamental ecological strategies. Ecosphere 4: 128. Standish, R. J., J. B. Fontaine, R. J. Harris, W. D. Stock, and R. J. Hobbs. 2012. Interactive effects of altered rainfall and simulated nitrogen deposition on seedling establishment in a global hotspot. Oikos 121: 2014–2025. Tulloss, E. M., and M. L. Cadenasso. 2015. Nitrogen deposition across scales: Hotspots and gradients in a California savanna landscape. Ecosphere 6: 167. Vallano, D. H., P. C. Selmants, and E. S. Zavaleta. 2012. Simulated nitrogen deposition enhances the performance of an exotic grass relative to native serpentine grassland competitors. Plant Ecology 213: 1015–1026. van Kleunen, M., W. Dawson, D. Schlaeper, J. M. Jeschke, and M. Fischer. 2010a. Are invaders different? A conceptual framework of comparative approaches for assessing determinants of invasiveness. Ecology Letters 13: 947–958. van Kleunen, M., D. R. Schlaepfer, M. Glaettli, and M. Fischer. 2011. Preadapted for invasiveness: Do species traits or their plastic response to shading differ between invasive and non-invasive plant species in their native range? Journal of Biogeography 38: 1294–1304. van Kleunen, M., E. Weber, and M. Fischer. 2010b. A meta-analysis of trait differences between invasive and non-invasive plant species. Ecology Letters 13: 235–245.

Walther, G.-R., A. Roques, P. E. Hulme, M. T. Sykes, P. Pyšek, I. Kühn, M. Zobel, et al. 2009. Alien species in a warmer world: Risks and opportunities. Trends in Ecology & Evolution 24: 686–693. Weiner, J., and S. C. Thomas. 1986. Size variability and competition in plant monocultures. Oikos 47: 211–222. West, A. G., T. E. Dawson, E. C. February, G. F. Midgley, W. J. Bond, and T. L. Alston. 2012. Diverse functional responses to drought in a Mediterranean-type shrubland in South Africa. New Phytologist 195: 396–407. Westoby, M. 1998. A leaf–height–seed (LHS) plant ecology strategy scheme. Plant and Soil 199: 213–227. Whitney, K. D., and C. A. Gabler. 2008. Rapid evolution in introduced species, “invasive traits” and recipient communities: Challenges for predicting invasive potential. Diversity & Distributions 14: 569–580. Wright, I. J., P. B. Reich, M. Westoby, D. D. Ackerly, Z. Baruch, F. Bongers, J. Cavender-Bares, et al. 2004. The worldwide leaf economics spectrum. Nature 428: 821–827. Wolkovich, E. M., and E. E. Cleland. 2011. The phenology of plant invasions: A community ecology perspective. Frontiers in Ecology and the Environment 9: 287–294. Yahdjian, L., L. Gherardi, and O. E. Sala. 2011. Nitrogen limitation in aridsubhumid ecosystems: A meta-analysis of fertilization studies. Journal of Arid Environments 75: 675–680. Zinnert, J. C., S. A. Shiflett, J. K. Vick, and D. R. Young. 2013. Plant functional traits of a shrub invader relative to sympatric native shrubs. Ecosphere 4: 119.