Limits and Effects of Invasion by the Nonindigenous Wetland Plant

Biological Invasions (2005) 7: 687–698 Ó Springer 2005 DOI 10.1007/s10530-004-5858-y

Limits and eﬀects of invasion by the nonindigenous wetland plant Lythrum salicaria (purple loosestrife): a seed bank analysis

Sarah B. Yakimowski1, Heather A. Hager2,3,* & Christopher G. Eckert1 1Department of Biology, Queen’s University, Kingston, ON, Canada K7L 3N6; 2Department of Biology, University of Regina, Regina, SK, Canada S4S 0A2; 3Present address: 409 Salt Springs Church Road, Brantford, ON, Canada N3T 5L9; *Author for correspondence (e-mail: [email protected])

Received 5 March 2004; accepted in revised form 18 June 2004

Key words: competition, Lythrum salicaria, propagule pressure, seed bank richness, seed dispersal, seedling emergence, seedling interactions, wetland marshes

Abstract

We used seed bank analyses to investigate the role of dispersal in limiting invasion by Eurasian Lythrum salicaria within and among North American wetlands, and the changes in seed bank diversity associated with this invader. We compared the number and species composition of seedlings emerging from soil sampled in 11 uninvaded wetlands and paired uninvaded and invaded sites within 10 invaded wetlands under both seedling competition and noncompetitive conditions. Almost no L. salicaria emerged in samples from uninvaded wetlands, indicating dispersal limitation despite prodigious seed production in nearby wetlands. However L. salicaria emerged in all samples from uninvaded sites in invaded wetlands, suggesting environmental limits on establishment within invaded wetlands. Conditions that provided opportunities for seedlings to compete reduced survival of Typha spp. but not L. salicaria seedlings. However, this was due to species-specific differences in post-emergence mortality rather than response to competition. Competition did reduce seedling mass, but this effect did not differ among species. Species richness of emerging seedlings was lower for invaded than uninvaded wetlands. Lower seed bank richness may be a cause or consequence of L. salicaria invasion. Efforts to reduce seed dispersal to uninvaded wetlands would likely slow the spread of this invader.

Introduction that plant distributions can be limited at a local scale by propagule availability (Tilman 1997; Determining what limits the abundance and dis- Levine 2000; Turnbull et al. 2000; Foster and Til- tribution of invasive species is crucial for control- man 2003). Dispersal will not lead to range ling them, yet the factors affecting spread in expansion unless propagules arrive at ‘safe sites’ native ecosystems are known for few invaders suitable for establishment (Harper 1977; John- (Parker et al. 1999). In general, a species may be stone 1986). Establishment may be inhibited by absent from a particular area because it has abiotic factors such as temperature, moisture, or failed to disperse there (Turnbull et al. 2000) or light availability, and biotic factors such as com- ecological conditions are unsuitable for establish- petitors, herbivores, or parasites. ment (Johnstone 1986; Williamson and Fitter During biological invasion, the relative impor- 1996). Seed addition experiments have shown tance of propagule dispersal vs environmental 688 suitability in limiting a species’ distribution may cumstantial evidence suggests that its regional vary with spatial scale. The leading edge of inva- distribution is limited primarily by environmental sion within a climatically appropriate region may factors. Environmental limitation is also sup- be limited by propagule dispersal (Higgins and ported by observations (Day et al. 1988) and Richardson 1999). Behind the leading edge, the experiments showing that L. salicaria invasion is distribution of the species is likely to be limited facilitated by the removal of plant competitors by a combination of dispersal and environmental and litter (Rachich and Reader 1999; Shadel and factors such as in stable metapopulations. How- Molofsky 2002; Hager 2004). The role of seed ever, the role of dispersal in limiting invasion at dispersal in limiting spread at any spatial scale any spatial scale remains unstudied for most has never been investigated. high-profile invasive species (but see Vila` and The second goal of this study is to use seed D’Antonio 1998; Wiser et al. 1998; Levine 2000; bank analyses to better understand the ecologi- D’Antonio et al. 2001). Determining the extent cal effects of L. salicaria invasion. Whether to which dispersal limits distribution is most invasion by nonindigenous plants negatively directly assessed using long-term seed addition affects the diversity and functioning of native experiments (e.g. Levine 2000). However, seed ecosystems is a contentious issue (Robinson et addition to uninvaded areas is difficult to justify al. 1995; Stohlgren et al. 1999; Levine and for species thought to have severe ecological or D’Antonio 1999; Symstad 2000), and L. salicar- economic effects. For invasive species with seed ia is a case in point (Anderson 1995; Hager and dormancy, an alternative approach is to evaluate McCoy 1998). This species is widely thought to the occurrence and viability of seeds in substrate form large monospecific stands (Rawinski and samples from areas where the species occurs Malecki 1984; Thompson et al. 1987; Mal et al. compared with where it is absent. Yet, compara- 1992; Weihe and Neely 1997; Mullin 1998) and tive seed bank analysis has been conducted for to displace dominant Typha spp. that define the few species (e.g. D’Antonio et al. 2001). physical structure (Dennison and Berry 1993), The main goal of this study is to use seed bank hydrology and nutrient cycling (Mitsch and analyses to investigate the role of propagule dis- Gosselink 1993) of marshlands, thereby altering persal in limiting the distribution of one of the these processes. However, recent field surveys of most high-profile and putatively aggressive invad- established plants have failed to detect the ers of natural habitat, Lythrum salicaria L. (Lyt- expected negative correlation between L. salicar- hraceae, purple loosestrife). This is a perennial ia abundance and native species richness plant introduced to North America from Eurasia (Treberg and Husband 1999; Farnsworth and in the early 1800s (Thompson et al. 1987). Ellis 2001; Hager and Vinebrooke 2004). Exami- Within invaded regions, L. salicaria exhibits a nation of the seed bank in invaded and unin- patchy distribution, with established populations vaded habitat can provide insight into the in some wetlands but not others. Even within relation of invasion with native species diversity, invaded wetlands, L. salicaria occurs in some as the representation of species in the seed bank patches but is absent from similar adjacent will largely indicate long-term recruitment patches (H.A. Hager, S.B. Yakimowski and C.G. potential (D’Antonio et al. 2001). Eckert, personal observation). Propagule avail- Here, we use comparative seed bank analyses ability seems unlikely to be limiting, as the spe- and seed-addition experiments to address two cies is renowned for its prodigious sexual main questions: (1) has seed dispersal limited the fertility; individual plants can produce hundreds spread of L. salicaria; and (2) is there an associa- of thousands of small seeds annually (Shamsi tion between invasion by L. salicaria and reduced and Whitehead 1974; A˚ gren 1996), which are recruitment potential of native species? To thought to be widely dispersed by wind, water address the first question, we quantify the occur- and animals (Thompson et al. 1987; Mal et al. rence of viable L. salicaria seeds in soil sampled 1992; Skinner et al. 1994). In addition, L. salicaria from invaded and uninvaded sites, both within forms large seed banks ( 8.2 seeds mL)1 in one and between wetlands. We also quantify the stand, Welling and Becker 1990). Therefore, cir- emergence of viable seeds added to soil samples 689 to determine whether the emergence of L. salicaria L. salicaria stems (L+) and ‡400 contiguous m2 seedlings is affected by soil characteristics among of Typha spp. stems without L. salicaria (L)). sites. To address the second question, we com- We determined the location of the L+ and L) pare the emergence of all species among samples sites within each invaded wetland arbitrarily, from invaded and uninvaded sites, under compet- except that L+ sites were located in dense itive and noncompetitive conditions, because patches of mature L. salicaria that contained recruitment is likely to be affected by competition many dead L. salicaria stems, indicating that the among seedlings. We specifically examine recruit- species had been established for several years. ment of Typha spp., as they are the dominant L) sites were ‡25 m from L+ sites and exhibited structural species in these wetlands, and widely no evidence of past or present L. salicaria. At thought to suffer from competition with L. sali- each of the 31 sample sites (11 U, 10 L) and 10 caria (e.g. Thompson et al. 1987; Mal et al. L+), we collected nine 5-cm diameter · 10-cm 1997). deep sediment cores at 1-m intervals from areas with <20cm standing water. Three cores were taken in late June and used in outdoor garden Materials and methods experiments, and six were taken in August and used in greenhouse experiments. Sampling Study sites and sampling design occurred before dispersal of L. salicaria or Typha spp. seed from the current year’s reproduction. In 2000, we sampled 21 Typha angustifolia-domi- Depletion of the seed bank by germination was nated marshes within a 400-km2 area surround- probably minor between sampling dates, because ing the west side of the Lake Minnetonka area in light penetration through the vegetation canopy Hennepin and Carver counties, southeastern is low. Cores from each site and time were Minnesota, USA. Lythrum salicaria was first pooled and stored at 4 °C. recorded near the study area in Ramsey county in 1924, and was well established in 68 of 87 Garden and greenhouse germination environments counties in Minnesota by 1994 (Skinner et al. 1994). We sampled 11 uninvaded (U) and Two outdoor garden experiments were conducted 10 invaded (L) wetlands. All wetlands are at the Carlos Avery Wildlife Management Area, classified as palustrine, emergent, seasonally Minnesota during July–August 2000. For each or semi-permanently flooded (Cowardin et al. site, pooled core samples were mixed thoroughly 1979; US Fish and Wildlife Service, National and a 100 mL subsample was spread as a 5 mm Wetlands Inventory, http://www.nwi.fws.gov/ layer onto potting soil in each of three 16.5-cm spec_note.htm). There were no obvious physical, diameter pots (31 sites · 3 replicate samples 93 biological or landscape-level features that differ- pots per experiment). Pots were placed in plastic¼ entiated U and L wetlands, which were of com- wading pools and randomly repositioned every parable size (mean ± SD: U = 17.7 ± 5.3 ha, 3–4 days. Water levels were maintained using min 0.5, max 57.2; L 10.3 ± 2.5 ha, well water and natural rainfall so that the soil min ¼ 0.9, max 26.3).¼ U wetlands¼ were located was saturated but not flooded. within¼ 1.2–23.8¼ km of each other, L wetlands Two more experiments were conducted in a within 0.1–13.8 km, and U and L wetlands greenhouse at Queen’s University in Kingston, within 6.0–20.3 km. Some invaded wetlands Ontario, Canada from November 2000-January occurred as close as 1 km to the U wetlands we 2001. For each site, a 100 ml subsample of mixed sampled, but were not included in the study. U core was spread onto potting soil as above, in wetlands had no evidence of L. salicaria (i.e. each of three 15.2 · 10.2 cm pots arranged six seedlings, adults or dead stems). per flat, three flats per tray. Pots were randomly Within each invaded wetland, we sampled a positioned and rotated weekly. Soil was always site with L. salicaria (L+) and a site where L. sal- saturated but not flooded. Germinating seeds icaria was absent (L)). Each invaded wetland were subjected to 16 h d)1 of 413 lmol m)2 s)1 had ‡400 contiguous m2 of closely situated supplementary lighting and 20–25 °C to simulate 690 early summer when germination occurs in natural competitive conditions. We compared L+ with populations. L) sites using a 3-way, mixed model analysis of variance (ANOVA) with wetland (W) as a ran- Competitive and noncompetitive conditions dom effect and site type (T: L+ vs L)) and environment (E: garden vs greenhouse, or com- We quantified seedling emergence from seed petitive vs noncompetitive) as fixed effects. The bank samples under both noncompetitive and denominators for the F-tests were as follows: T competitive conditions. We obtained noncompeti- used MSW·T; E used MSW·E;T·E, W·E and tive conditions in both garden and greenhouse W·T used MSW·E·T;W·E·T used MSResidual; experiments by removing seedlings as soon as and W used a synthetic denominator they could be identified. Unidentifiable species (MSW·E + MSW·T)MSW·E·T). We compared U were transplanted and grown for later identifica- with either L+ or L) sites using a partially tion. We recorded emergence until no new seed- nested 3-way ANOVA with site type as a main lings appeared. In the garden, we obtained fixed effect, wetland nested within site type competitive conditions by allowing seedlings to (W[T]) as a random effect and environment as a remain in the pots for 6 weeks, and then count- fully crossed fixed effect. The denominators for ing them and harvesting aboveground biomass, the F-tests were as follows: T used MSW[T]; W[T], which was sorted by species, dried to constant E and T·E used MSE·W[T]; and E·W[T] used mass and weighed to 0.001 g. MSResidual. Emergence of L. salicaria after seed addition was compared with similar models, Seed addition without environment as a factor. Data were transformed to normalize residuals and eliminate We added viable L. salicaria seed to soil samples heteroscedasticity using Box–Cox power transfor- under greenhouse conditions to determine mations (Y¢ Yk; Neter et al. 1990). We com- whether differences in soil characteristics among pared the¼ proportional change in average U, L) and L+ sites might have contributed to individual seedling mass between site types using differences in the emergence of L. salicaria. We t-tests performed on site means of log10-trans- sowed an average (±1 SE) of 354 ± 26 seeds into formed data. Because we contrasted U, L) and each of three replicate soil subsamples from each L+ sites using three simultaneous analyses, we site that were autoclaved to kill naturally occur- held the type I error (a) at 0.05 using sequential ring seed (no seedlings emerged from autoclaved Bonferroni to reduce the per test a accordingly soil without seed added) and monitored emer- (Rice 1989). Means are presented ± 1 SE unless gence under non-competitive greenhouse condi- otherwise indicated. tions. We also sowed the same amount of seed into three replicate samples of untreated soil from each U and L) site to verify that autoclav- Results ing did not make soil less suitable for germination. These seeds were collected in fall 2000 from Comparison of L. salicaria emergence between four L+ sites, and exhibited 96% emergence in invaded and uninvaded sites potting soil in the greenhouse. Many more L. salicaria seedlings emerged in soil Statistical analyses sampled from invaded than uninvaded sites in both garden and greenhouse environments (Fig- For all experiments, we quantified the emergence ure 1). Seedlings emerged from all invaded wet- of L. salicaria, Typha spp. and all other species land samples (both L+ and L) sites) but only pooled (species list available from the authors). two uninvaded (U) wetland samples (Fisher’s From these data, we calculated total seedling exact test comparing U with L: P 0.0002). emergence and species richness excluding L. sali- Only five L. salicaria seedlings emerged¼ from two caria and Typha spp. We also calculated individ- of 33 U samples under garden conditions, and a ual seedling dry mass for each species under single L. salicaria from one U sample under 691

Lythrum salicaria emergence tended to be in the same wetlands as the L+ 1000 sites with the highest densities of L. salicaria seeds (Figure 1). Differences in soil characteristics between invaded and uninvaded sites did not contribute 100 to variation in L. salicaria emergence. Overall, 92% of seeds sown on autoclaved soil yielded seedlings, and there was no difference in seedling number between L+ (306 ± 6), L) (318 ± 6) 10 and U (317 ± 5) samples (all P > 0.2). Emer- gence of L. salicaria seed added to untreated soils Number of seedlings + 1 was also very high for both L) (338 ± 17) and U sites (313 ± 14). 1 9 UL–L+ sites Site type Comparison of seed bank species diversity between invaded and uninvaded sites Figure 1. Emergence of Lythrum salicaria seedlings from soil samples collected from 11 uninvaded wetlands (U) and sites in 10 invaded wetlands with (L+) and without (L)) standing The number of Typha seedlings emerging varied populations of L. salicaria. Each point is the number of seed- more than 100-fold among 100 mL soil samples lings emerging from a 100 ml soil sample, averaged across (range 3–411), but there was no difference three replicate samples per site subjected to each of two between¼ L+, L) and U samples (Figure 2; growth environments (garden and greenhouse, n 6 samples ¼ Table 1). Environment strongly affected emer- per point; Log10 scale for display only). There was no difference in seedling emergence between environments, nor did the gence, as about 2 times more Typha emerged in environment affect differences between site types. Lines join garden (103 ± 11) than greenhouse (49 ± 6), L) and L+ sites within each invaded wetland. but the effects of environment and site type did not interact. Among invaded wetlands, the number of Typha emerging correlated positively greenhouse conditions. We verified the near rather than negatively with the number of L. sali- absence of L. salicaria seeds in U samples by caria emerging (L+ garden: r +0.36, P 0.31; examining three 1-ml aliquots of soil from each L+ greenhouse: r +0.67, P¼ 0.03; L) garden:¼ U sample under a dissecting microscope. r +0.65, P 0.04;¼ L) greenhouse:¼ r +0.90, Although L. salicaria seedlings emerged in all P¼ 0.0004). ¼ ¼ samples from invaded wetlands, an average of ¼Emergence of other species aside from Typha about 10-times more seedlings emerged from L+ spp. and L. salicaria also varied widely among than L) samples in both environments (gar- samples (range 0–82 seedlings/sample) and, in den 267 ± 72 and 22 ± 13 seedlings per sam- contrast to Typha¼ , was 2-times higher in green- ¼ ple, respectively; greenhouse 319 ± 111 and house (13.8 ± 1.8) than garden (6.8 ± 1.6; ¼ 36 ± 19). Seedling numbers differed significantly Table 1). More seedlings emerged from U than between site types (F1,9 87.7, P < 0.0001), but L+ or L) samples, but there was no difference ¼ not environments (F1,9 1.3, P 0.28), and the between L+ and L) samples within invaded wet- ¼ ¼ difference between L+ and L) did not vary lands (Figure 2; Table 1). Effects of site type and between environments (E·T interaction environment did not interact (E·T: all P > 0.7). F1,9 1.4, P 0.27). There was significant varia- Among invaded wetlands, the number of seed- ¼ ¼ tion in emergence among the 10 invaded wet- lings of other native species did not correlate lands (F9,8 7.8, P 0.0052), and the number of with the number of L. salicaria seedlings (L+ ¼ ¼ seedlings emerging from L) sites correlated posi- garden: r +0.18, P 0.6; L+ greenhouse: tively with the number emerging from L+ sites r )0.34, ¼P 0.3; L¼) garden: r )0.37, (garden: r +0.73, P 0.02; greenhouse: P¼ 0.3; L) greenhouse:¼ r )0.26, P 0.5).¼ ¼ ¼ r +0.78, P 0.008). Thus, L) sites containing ¼Higher seedling numbers¼ were reflected¼ in ¼ ¼ considerable numbers of viable L. salicaria seeds higher richness of other species emerging from U 692

Seedlings of Typha spp. Effect of competitive conditions on emergence 300 Competitive conditions reduced total seedling 200 number by 24% (Figure 3; Table 2) but did not alter differences in emergence among site types 100 (no site type · competition interactions: F1,9 < 0.01, P 0.9 for L. salicaria from L+ vs 0 L) samples; see¼ Table 2 for Typha and other spe- Seedlings of other species cies). However, the effect of competitive condi- 60 tions (compared to noncompetitive conditions)

40 a diﬀered among species. The number of Typha seedlings declined by 42%, whereas L. salicaria 20 seedlings declined by only 9% (F1,9 5.23, b b P 0.05 for the L+ vs L) analysis) and¼ seed- ¼

Per 100 mL sample 0 lings of other species increased by 27%. However, Species richness 10 this did not result from competitive effects per se. The density of seedlings was >3-times higher in 8 L+ (301 ± 44) than L) (95 ± 20) or U a 6 (76 ± 12) samples, yet the difference in Typha 4 b b seedlings between competitive vs noncompetitive 2 conditions did not vary between L+, L) or U 0 sites (Table 2). U L– L+ Seedlings surviving to harvest in L) or U sam- Site type ples had greater mean individual dry mass than Figure 2. Emergence of native wetland species seedlings from those in L+ samples, where seedling density was 100 mL soil samples collected from 11 uninvaded wetlands >3 times higher (Figure 4). The proportional (U) and sites in 10 invaded wetlands with (L+) and without decline in seedling mass between L) and L+ (L)) standing populations of L. salicaria. Box plots show the pots did not differ between seedling types (differ- distribution of site means (based on three replicate samples subjected to each of two growth environments, n 6 samples ence between L) and L+ in log10[dry mass] with ¼ 95% confidence limits: L. salicaria 0.54 [0.18– per point) for number of Typha seedlings per sample (top), ¼ number of seedlings of other species (middle; excluding L. sal- 0.91], Typha 0.47 [0.12–0.82], others 0.39 icaria and Typha spp.), and the richness of other species [0.08–0.69]). ¼ ¼ emerging from each sample (bottom). Boxes for U wetlands are shaded. The bottom line in each box plot is the 25th per- centile, the midline is the median and the top is the 75th per- centile. The whiskers extend 1.5 interquartile distances from Discussion the median or to the more extreme points (circles), whichever is closest. Means are shown by the hatched lines within the Limitations to the regional and local distribution boxes. Contrasts between sites types are shown with letters: of L. salicaria those not sharing a letter are significantly different. Although emergence differed between environments, the environment generally did not affect differences between site types (see Our results suggest that the distribution of L. sal- Table 1 and text). icaria in Minnesota marshes is limited by both seed dispersal and environmental conditions, but than L+ and L) samples (Figure 2; Table 1). the importance of these factors varies with spatial Again, L+ and L) sites did not differ. Species scale. Within invaded wetlands, the distribution richness was 60% higher in greenhouse seems limited primarily by environmental condi- (3.3 ± 0.3 other species/sample) than garden tions, because a viable L. salicaria seed bank was (2.0 ± 0.2). Of 30 species identified, four were always present at L) sites. Most L) samples not native to North America, but these made up yielded a considerable number of L. salicaria only 4.5% of all seedlings, and removing them seedlings (>10/100 ml), and two produced seed- from the analysis did not change the results. ling on par with most L+ sites, although the 693

Table 1. Analysis of variation in seedling emergence for species from soil samples collected from 11 uninvaded wetlands (U) and sites in 10 invaded wetlands with (L+) and without (L)) standing populations of L. salicaria.a Variable/eﬀect L+ vs L) L+ vs U L) vs U FPFPFP Seedlings of Typha spp. k = 0.4 k = 0.2 k = 0.2 Site type (T) 0.27 0.61 0.08 0.77 0.00 0.99 Environment (E) 26.86 0.0006 12.33 0.0023 16.40 0.0007 E · T 0.02 0.87 0.42 0.52 0.86 0.37

Seedlings of other species k = –0.2 k = 0.2 k = 0.2 Site type (T) 0.04 0.85 13.11 0.0018 12.54 0.0022 Environment (E) 13.78 0.0048 9.54 0.0060 12.03 0.0026 E · T 0.00 0.99 0.04 0.85 0.01 0.91

Richness of other species k = 0.4 k = 0.6 k = 0.5 Site type (T) 1.60 0.24 7.92 0.011 11.77 0.0028 Environment (E) 2.06 0.18 16.48 0.0007 18.35 0.0004 E · T 0.20 0.67 3.44 0.079 7.41 0.0135 aComparisons between L+ and L) sites use 3-way factorial ANOVA. Comparisons between U and both L+ and L) sites use a partially-nested 3-way ANOVA (see Materials and methods for details). For all analyses, the whole ANOVA model was significant (all P < 0.008, all r2 > 0.52). Only selected effects from the full models are shown here (all df = 1). All variables were power transformed (Y¢ =Yk), and k is presented for each analysis. P-values remaining significant after sequential Bonferroni correction are in bold. average density (29 seedlings/100 ml soil) was At a larger scale, the distribution of L. salicar- much lower than that of L+ samples (293/ ia among Minnesota wetlands is currently limited 100 ml). In addition, the occurrence of viable by seed dispersal. Virtually no seed was found in seeds was greater at L) sites in the same wet- soil samples from uninvaded wetlands that were lands as L+ sites with a dense L. salicaria seed within 1–20 km of wetlands with large L. salicaria bank, and a large and dense stand of reproduc- populations. Although this suggests that the tive plants. Yet, there was no evidence of these propagule pressure by L. salicaria on uninvaded L) sites ever having had established L. salicaria. wetlands is very low, it does not rule out some These results are similar to those of other seed role for environmental factors for seeds that do bank studies (e.g., Drake 1998). reach these wetlands. We ruled out the possibility Many biotic and abiotic factors may poten- that soil characteristics might prevent germina- tially inhibit seedling establishment. For exam- tion of L. salicaria for the U sites we sampled. ple, low light caused by dense Schizachyrium Experimental sowing of seed into a variety of condensatum prevented establishment of Melinis invaded and uninvaded wetlands in the field is minutiflora in Hawaiian forests (D’Antonio required to test other environmental effects. For et al. 2001). Because seedling recruitment in example, sowing experiments confirmed that seed Typha-dominated marshes is often limited by supply rather than environmental factors limits light (Grace and Harrison 1986), L. salicaria the spread of non-native species in riparian habi- establishment can be negatively affected by the tat (Levine 2000). presence of adult Typha spp. and plant litter Although the interaction between seed dis- (Hager 2004). Soil characteristics do not pre- persal and environmental variation in controlling vent L. salicaria recruitment at L) sites, as we the regional distribution of an invasive species observed uniformly high emergence of L. sali- may be complex, our results indicate that L. salicaria seed added to both autoclaved and caria propagules are not widely dispersed among untreated soils. Experimental sowing of L. sali- wetlands in regions where it has been established caria seed into L) sites shows that the presence for many years. Efforts to prevent dispersal of of plants and plant litter reduces establishment seed into uninvaded wetlands would seem likely (Hager 2003). to slow the spread of this invader. 694

correlated positively with L. salicaria emergence Competitive Noncompetitive in invaded wetlands, rather than negatively. Lythrum seedlings Typha seedlings Competitive conditions reduced the emergence 350 160 and survival of Typha seedlings by 59%, com- 300 a pared to non-competitive conditions, while having 250 120 little or no effect on L. salicaria. However, this 200 large decrease in Typha occurred across all site 150 80 types, despite 3-fold variation in total seedling 100 density and species composition. These results 40 50 b indicate that the lower success of Typha, relative c 0 0 to L. salicaria (and other species) after six weeks of growth probably reflects a difference among Other seedlings Species richness 25 4 a a species in post-emergence survival rather than response to competition. A major increase in the 20 3 density of L. salicaria seedlings does not decrease 15 ab post-emergence survival of Typha seedlings across b 2 the densities observed here. Overall, 88% of Ty- 10 b pha seedlings emerged before the first Lythrum b 1 5 seedling, which would reduce opportunities for L. salicaria to competitively suppress Typha. 0 0 UL–L+UL–L+ We observed a negative effect of competition on the mean mass of the seedlings that survived, Site type but the relation between total seedling density Figure 3.Effect of competition on emergence of Lythrum sali- and seedling mass did not differ among Typha caria and other wetland species seedlings from 100 mL soil spp., L. salicaria and other species. This suggests samples collected from 11 uninvaded wetlands (U) and sites that individual seedlings suffer from competition in 10 invaded wetlands with (L+) and without (L)) standing populations of L. salicaria. Under noncompetitive conditions to similar extents at this stage of development, (open circles) seedlings were removed from pots as they regardless of species identity or the composition emerged, whereas under competitive conditions (solid circles) of competitors. In contrast, experiments suggest seedlings were not removed and emergence was scored six that mature L. salicaria specifically outcompetes weeks after sowing. Each point is the mean ± 1 SE across mature Typha spp. (Mal et al. 1997; Weihe and sites. Analysis of these data is in Table 2. Contrasts between sites types are shown with letters: those not sharing a letter Neely 1997). are significantly different. There were no interactions between In addition to lower post-emergence survival the effects of site type and competition. of Typha than L. salicaria and other wetland species, lower emergence of Typha in the greenhouse than garden also suggests that Typha recruitment Lythrum–Typha interactions is more sensitive to environmental conditions. Our greenhouse was maintained at 20–25 °C, but It is widely expected that invasion by nonindige- the preference by germinating Typha for temper- nous species reduces the abundance and diversity ature fluctuations of 7 °C (Thompson and Grime of native species, and this claim has been made 1983) could account for lower emergence in the repeatedly with respect to L. salicaria (e.g. Ra- greenhouse. Typha seedlings were more prone to winski and Malecki 1984; Thompson et al. 1987; browning and drying in the greenhouse than Weihe and Neely 1997; Mullin 1998; Blossey were other species. This sensitivity might disad- et al. 2001). However, our results provide no evi- vantage Typha in competition with L. salicaria dence for a negative effect of L. salicaria on the during early life-history stages. Experiments that recruitment potential of Typha. There was no dif- examine interactions over a broader range of life ference in Typha emergence among site types (U, history stages and environments are required to L) and L+) under noncompetitive or competi- better understand how L. salicaria invasion tive conditions. Moreover, Typha spp. emergence affects Typha abundance. 695

Table 2.Eﬀect of competition on seedling emergence for species from soil samples collected from 11 uninvaded wetlands (U) and sites in 10 invaded wetlands with (L+) and without (L)) standing populations of L. salicaria (see Figure 3).a Variable/eﬀect L+ vs L) L+ vs U L) vs U FPFPFP Seedlings of Typha spp. k = )0.2 k = 0.2 k = 0.2 Site type (T) 0.01 0.92 0.01 0.90 0.15 0.70 Competition (C) 72.90 0.0001 23.89 0.0001 21.83 0.0002 C · T 0.00 0.99 0.23 0.63 0.09 0.76

Seedlings of other species k = 0.2 k = 0.2 k = 0.2 Site type (T) 1.10 0.32 10.59 0.0042 13.04 0.0019 Competition (C) 10.20 0.011 5.37 0.032 3.37 0.082 C · T 0.42 0.53 0.61 0.44 0.12 0.73

Richness of other species k = 0.8 k = 0.6 k = 0.6 Site type (T) 0.01 0.91 5.84 0.025 6.27 0.022 Competition (C) 0.32 0.58 0.15 0.70 1.20 0.29 C · T 1.17 0.31 0.31 0.59 0.10 0.75 aComparisons between L+ and L) sites use 3-way factorial ANOVA. Comparisons between U and both L+ and L) sites use a partially-nested 3-way ANOVA (see methods). For all analyses, the whole ANOVA model was significant (all P < 0.03, all r2 > 0.45). Only selected effects from the full models are shown here (all df = 1). All variables were power transformed (Y¢ =Yk), and k is presented for each analysis. P-values remaining significant after sequential Bonferroni correction are in bold.

Lythrum salicaria Eﬀect of Lythrum salicaria on the recruitment Typha spp. Other native species potential of other wetland plant species 1.3 Surveys of established plants in invaded versus 1.1 uninvaded habitats have failed to detect the expected negative association between L. salicaria 0.9 abundance and native species diversity (Jean and Bouchard 1993; Treberg and Husband 1999; 0.7 Farnsworth and Ellis 2001; Keller 2000; Hager

0.5 and Vinebrooke 2004). Likewise, we did not ﬁnd a diﬀerence in species richness of seedlings (average seedling mass) seedling (average

10 0.3 emerging from the seed bank (excluding L. salicaria and Typha) between L+ and L) sites in Log UL – L+ 0.1 invaded wetlands. However, species richness of 1.6 1.8 2.0 2.2 2.4 seedlings was higher for uninvaded wetlands than Log (total seedlings) 10 either L) or L+ sites in invaded wetlands. The abundance of other species was also 4-fold Figure 4.Effect of competitive conditions on individual seedling mass. For each of the three classes of emerging seedlings, higher in U than L) or L+ samples. Higher mean ± 1 SE seedling dry mass is compared between U, L) abundance was greater than could be accounted and L+ samples that differed greatly in seedling density. for by the additional species emerging (only Seedling mass differed significantly between L) and L+ sam- 2-fold greater than L- or L+). These results ples for all three seedling types (paired t-tests on site means: L. salicaria t 3.44, P 0.0088; Typha spp. t 3.03, suggest a negative association between L. salicar- P 0.014; other¼ species t ¼2.87, P 0.018). There were¼ no ia and the potential richness and abundance of ¼ ¼ ¼ differences in average seedling mass for each seedling type other species among, but not within wetlands. between L) and U samples (2-sample t-tests on site means: Interpreting this result is difficult. One possibil- Typha spp. t 0.46, P 0.65; other species t 0.65, ity is that L. salicaria reduces native plant diver- P 0.52). No L.¼ salicaria seedlings¼ emerged from U samples¼ ¼ under competitive conditions, so seedling dry mass could not sity of whole wetlands, including patches where it be quantified for this site type. is absent. Alternatively, wetlands with low species 696 diversity might be more susceptible to invasion, Blossey B, Skinner LC and Taylor J (2001) Impact and man- perhaps because of low resource competition agement of purple loosestrife (Lythrum salicaria) in North (Robinson et al. 1995; Symstad 2000). However, America. Biodiversity and Conservation 10: 1787–1807 Brown JH, Stevens GC and Kaufman DM (1996) The geo- few studies have shown a negative association graphic range: size, shape, boundaries and internal structure. between species richness and invasibility (Tilman Annual Review of Ecology and Systematics 27: 597–623 1997; Costello et al. 2000). Instead, seed addition Costello DA, Lunt ID and Williams JE (2000) Effects of inva- experiments find positive correlations between sion by the indigenous shrub Acacia sophorae on plant species richness and establishment of new species, composition of coastal grasslands in south-eastern Austra- lia. Biological Conservation 96: 113–121 and the occurrence of invaders is often positively Cowardin LM, Carter V, Golet F and LaRoe E (1979) Classi- correlated with species richness in unmanipulated fication of Wetlands and Deepwater Habitats of the United communities (Robinson et al. 1995; Stohlgren States. US Fish and Wildlife Service, 103 pp et al. 1999; Levine and D’Antonio 1999; Symstad D’Antonio CM, Hughes RF and Vitousek PM (2001) Factors 2000). Species richness explains very little of the influencing dynamics of two invasive C4 grasses in seasonally dry Hawaiian woodlands. Ecology 82: 89–104 variation in invasibility, especially in early stages Day R, Keddy P, McNeill J and Carleton T (1988) Fertility of invasion (Wiser et al. 1998). Rather, factors and disturbance gradients: a summary model for riverine correlated with richness determine community in- marsh vegetation. Ecology 69: 1044–1054 vasibility (Robinson et al. 1995; Symstad 2000; Dennison MS and Berry JF (1993) Wetlands: Guide to Science, Levine and D’Antonio 1999; Levine 2000). For Law, and Technology. Noyes Publications, New Jersey Drake DR (1998) Relationships among the seed rain seed instance, disturbance by humans probably bank and vegetation of a Hawaiian forest. Journal of Veg- reduces wetland diversity while facilitating intro- etation Science 9: 103–112 duction of nonindigenous species (Findlay and Emery SL and Perry JA (1996) Decomposition rates and Houlahan 1997). Our finding of different associa- phosphorous concentrations of purple loosestrife (Lythrum tions between L. salicaria and species richness salicaria) and cattail (Typha spp.) in fourteen Minnesota wetlands. Hydrobiologia 323: 129–138 within wetlands than among wetlands emphasizes Farnsworth EJ and Ellis DR (2001) Is purple loosestrife (Ly- that empirical patterns of association between thrum salicaria) an invasive threat to freshwater wetlands? invasive and native species may vary with spatial Conflicting evidence from several ecological metrics. Wet- scale. lands 21: 199–209 Findlay CS and Houlahan J (1997) Anthropogenic correlates of species richness in southeastern Ontario wetlands. Con- servation Biology 11: 1000–1009 Acknowledgements Foster BL and Tilman D (2003) Seed limitation and the regu- lation of community structure in oak savanna grassland. 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