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Ecology |
State University of New York at Stony Brook, Department of Ecology and Evolution, 650 Life Sciences Bldg., SUNY–Stony Brook, Stony Brook, New York 11794 USA
Received for publication May 25, 2005. Accepted for publication October 25, 2005.
ABSTRACT
In attempting to determine the traits associated with invasive plant species, ecologists have often used species native to the invaded range as "control species." Because many native species themselves are aggressive colonizers, comparisons using this type of control do not necessarily yield information relevant to distinctions between invasive and noninvasive species. Here we implement an alternative study design that compares phenological, architectural, size, and fitness traits of several introduced invasive species to introduced noninvasive species within two genera of Asteraceae (Crepis and Centaurea). While there were many significant differences between the genera, there were few shared attributes among invasive or noninvasive congeners, even for traits as seemingly important as the number of inflorescences produced and the size of seed heads. Instead, the results suggest that differences in invasiveness between closely related species is better explained as the result of complex trait interactions and specific introduction histories.
Key Words: Asteraceae • Centaurea • comparative method • Crepis • exotic species • invasion success • invasive species • nonindigenous species
Although the introduction of some species into novel geographical areas can be viewed as beneficial (e.g., crops, horticultural varieties, biological control agents) and by most measures the majority of introductions are in fact benign (Williamson, 1996
), there is a great deal of concern regarding the detrimental effects that certain introduced species may cause (Pimentel et al., 2000
). Unfortunately, to date there are few identifiable indicators that allow us to proactively weed out the occasional problematic invaders from the crowd of benign and beneficial prospects (Enserink, 1999
; Mack et al., 2000
; but see Rejmanek and Richardson, 1996
; Reichard and Hamilton, 1997
; Rejmanek, 2000
; Kolar and Lodge, 2001
, 2002
).
With respect to the traits common to introduced invasive species, Kolar and Lodge (2002)
describe a potential paradox in invasion biology, noting that some researchers are content to view this problem as identified and solved, while others believe that invasive traits are largely idiosyncratic and, therefore, predicting which species will be invasive is impossible. Taking the position that both of these claims are too extreme and in want of support, Kolar and Lodge (2002)
suggest, we think correctly, that there may be a more realistic and fruitful middle ground where the prediction of invasiveness is feasible. Finding this middle ground requires the recognition that "invasive species" is not a homogeneous class of natural objects and that progress in this field may be better achieved by focusing our efforts on meaningfully circumscribed groups, be they taxonomic, ecological, biogeographical, or otherwise. The successful identification of local trends regarding invasive species (in lieu of general laws or an endless collection of anecdotes) could be quite valuable to both scientists and public policymakers.
With this mindset of tackling invasion biology in tractable allotments, the importance of delineating appropriate groups for comparison becomes paramount. If identifying appropriate groups of organisms for recognizing local trends among invasive species is the first step, then identifying the proper groups for comparison is the second. With some exceptions (e.g., Rejmanek and Richardson, 1996
; Grotkopp et al., 2002
; Kolar and Lodge, 2002
; Gerlach and Rice, 2003
; Mandak, 2003
; Bellingham et al., 2004
; Burns, 2004
; Sutherland, 2004
; Lloret et al., 2005
), controls for studies of introduced invasives have been selected from among the closely related species (often congeners) native to the region of invasion. The use of closely related species is helpful on two scores: first, it allows one to directly address the issue of the lack of phylogenetic independence (because species are all related to each other at some level, they have a shared history and therefore are not statistically independent data points; Harvey and Pagel, 1991
); second, it insures that the differences between study species are not so great as to be essentially uninformative.
On the other hand, the common practice of using native species as a control can lead to misleading comparisons, considering that there are many aggressively colonizing native species. A logical alternative can be arrived at when one appreciates that one of the primary goals of invasion biology is to sort out what makes introduced invasive species different from the host of noninvasive introductions. To the extent that this is the case, a well-informed study of the issue requires representatives of both groups of interest, that is, introduced invasives and introduced noninvasives. Fortunately, there is often quite a large number of introduced noninvasives from which to choose, though this does not make the task trivial when one begins to focus on particular geographic regions, ecosystems, or taxonomic groups.
In this study we examine phenological, architectural, and fitness trait differences between populations of introduced invasive and introduced noninvasive plant species within two closely related genera of Asteraceae, hypothesizing that variable modes of invasiveness may exist between different evolutionary lineages. Our premises led to the following questions and expectations: (1) Are there differences in traits between introduced invasive and introduced noninvasive species that are consistent across evolutionary lineages? We predicted that differences between introduced invasives and introduced noninvasives that hold up across genera would be largely restricted to fitness or fitness proxy traits (e.g., onset of reproductive maturity, quantity of offspring), due to the likely existence of numerous different multivariate phenotypes that may yield similar fitness values. (2) Are there differences in traits between introduced invasive and introduced noninvasive species that are found only within one clade? We predicted that genus-specific differences between introduced invasives and introduced noninvasives would most likely be found among the phenological (e.g., time to bolting) and architectural (e.g., branching pattern) traits because fitness trait patterns are expected to be similar between clades. (3) Are there major differences in traits among introduced invasive species within each clade? We predicted that there would be a detectable amount of variability among introduced invasive species within each genus—even in the presence of clade-specific patterns, as the result of idiosyncratic ecological and evolutionary histories.
MATERIALS AND METHODS
Plant material
We chose species of the genera Centaurea and Crepis (both Asteraceae) for our study based, in part, on our ability to collect multiple invasive and noninvasive introductions to North America within both genera (Table 1). To assess differential rates of a species' range expansion, which we feel is ecologically consistent with the concept of invasiveness (Richardson et al., 2000
; Colautti and MacIsaac, 2004
), we characterized each species based on the number of U.S. states with a record of its presence in The PLANTS Database (USDA, NRCS, 2002
) (invasive species, present in 
24 of the 48 continental U.S. states; noninvasive species, in 
16 states). Intrastate level (county by county) presence/absence records confirm our characterization of the invasive status at this smaller spatial scale (i.e., for the species used in this study, invasives are largely widespread both across and within states, while noninvasives are largely restricted at both scales; Appendix S1, see Supplemental Data accompanying online version of this article). Estimates of the dates of introduction to North America from native Eurasian ranges are given as the earliest found herbarium record compiled from various North American herbaria and available floras (Table 1). Some of these dates no doubt underestimate the actual age of the introduction, and it is also possible that some dates represent populations that did not naturalize and, in some sense, overestimate the relevant age of the introduction.
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), the same is true of noxious weed lists where, depending on the specific case, a species can be listed without any evidence of it occurring in a given state or be absent from a list despite being quite common. Additional problems that hamper attempts to characterize invasiveness (including the current study and the vast majority of previous studies) include (1) the reality that invasiveness is more accurately a continuous variable, by no means intrinsically categorical; (2) the fact that "invasiveness" may change through time, as indicated by the "lag time" that is often noted for invasive species; and (3) the difficulty in estimating the rate of spread due to the absence of information on the frequency and localities of a species' introduction. These problems are worth noting and deserve further investigation when possible. However, the inability of most invasion studies to address these problems does not invalidate the utility of such studies, nor does it place invasion biology outside the norm of biology in the degree to which we must logistically treat various aspects of our systems as "all else being equal."
While the intraspecific phylogenetic relationships within Crepis and Centaurea are poorly known (Babcock, 1947
; Susanna et al., 1995
; Whitton et al., 1995
; Garcia-Jacas et al., 2001
), there is little doubt that the genera themselves represent separate clades (at the level of tribe or subfamily; Bremer, 1994
; Bayer and Starr, 1998
; Panero and Funk, 2002
).
All study species are short-lived, predominantly annual, herbaceous plants occurring in a variety of disturbed habitats. The degree of phenotypic similarity between genera allowed us to directly compare a large number of traits potentially relevant to invasion. Twenty-six different seed accessions of the nine study species were obtained from wild-collected populations, European botanical gardens, and North American collaborators (see Table 1 for details).
Plant handling and experimental setup
For all 26 populations, 2–4 seeds (based on availability) were planted into each of ten 4 x 4 x 4.5 cm starter pots with standard autoclaved Pro-mix potting soil (Premier Horticulture, Red Hill, Pennsylvania, USA) and placed in the University of Tennessee White Avenue greenhouse. Ambient lighting and photoperiod were augmented with greenhouse lights for 16 h per day. Once established, individual seedlings were transplanted into 13 x 13 x 13.5 cm pots and distributed in a randomized block design comprising 10 complete blocks. Planting was carried out in October and November 2001.
All traits were measured at the individual level on plants transplanted to larger pots. These traits included (1) days to bolting, a measure of time spent in the vegetative stage; (2) days from bolting to initial flowering (anthesis), a measure of time until reproductive maturity; (3) days from flowering to initial dehiscence, an estimate of time until seed dispersal; (4) number of rosette leaves at bolting, an estimate of investment in the vegetative phase of growth; (5) number of basal stems, a component of plant architecture; (6) stem diameter (size); (7) stem length (size); (8) branch order (number of branch nodes encountered in tracing the longest stem backwards to the rosette), another characterization of plant architecture during the reproductive phase; (9) above ground biomass (dry mass), a measure of overall growth; (10) involucre diameter, a component of reproductive fitness; and (11) number of inflorescences, another component of reproductive fitness.
Data analysis
Data were checked for violations of assumptions of normality and homoscedasticity (days to initial flowering and number of basal stems were log transformed, and a power transformation was applied to number of inflorescences) and analyzed with a two-way analysis of variance (ANOVA). Calculation of sums of squares and significance tests were carried out using JMP (SAS Institute, Cary, North Carolina, USA), version 5.1. For each trait, the full model included the following main effects and interactions: genus (overall differences between Centaurea and Crepis, treated as a fixed effect), provenance (to account for overall differences between European and North American collected populations, fixed effect), invasive status (overall differences between invasive and noninvasive groups, fixed effect), block (microenvironmental effects, random), genus x invasive status (i.e., differences between invasive and noninvasive species that are specific to individual genera, fixed), species nested within genus x invasive status (i.e., differences between species within each genus/invasive status combination, random), population nested within species (i.e., differences between populations of the same species, random), and error (residual variance).
To reduce the likelihood of our not detecting significant effects (Type II error), we have chosen to highlight all tests where the associated P value is less than the typical 0.05. After Moran (2003)
, in lieu of using the Bonferroni correction (normally applied to maintain the overall probability of committing a Type I error), we report the likelihood (p) of finding a particular number of significant test results below our 
value (K) , given the total number of tests performed (N), by the following formula:
) and discuss the results from a statistically conservative perspective.
Power analyses were carried out using G·Power (Buchner et al., 1997
) on all ANOVA tests. We followed the conventions of Cohen (1992)
and evaluated whether we had the statistical power to detect "medium" and "small" effects caused by our treatments. "Medium" effect size (ES) means that treatment differences are "visible to the naked eye of a careful observer" (e.g., the invasive plants have visibly more fruits than the noninvasive plants), whereas "small" ES means that the treatment differences are "noticeably smaller than the medium but not so small as to be trivial" (e.g., the invasive plants have more fruits than the noninvasive plants, but it is not as obvious) (Cohen, 1992
). This analysis accounts for the possibility that an effect may be statistically nonsignificant, not because the effect is actually biologically insignificant, but rather is due to limited sample size. Power values range from 0 to 1 and are calculated for each effect in the model, based upon the degrees of freedom and the ES (small, medium, or large) of interest. Power values of 0.8 and higher are considered to be sufficient to conclude that there was enough statistical power to detect an effect of the size of interest (Cohen, 1992
).
To examine the multivariate relationships among traits, we performed a discriminant analysis on the characters for which we had data for all species (days to bolting, days from bolting to flowering, number of leaves, number of basal stems, stem length, stem diameter, branch order, biomass, involucre diameter, and number of inflorescences). Different plots in canonical components space allowed us to identify which characters most contributed to identifying plants based on the combined criteria of genus and invasive status. These analyses were also carried out using JMP, version 5.1.
RESULTS
Phenological traits
An analysis of variance showed that the main effect of genus was significant for one of three phenological traits (days from bolting to flowering; Table 2). While there were no significant differences in time to bolting, upon bolting Crepis development was, in general, more rapid (Fig. 1).
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The main effect of provenance explained relatively little of the variance in the phenological traits (Table 2), although it was significant for days to flowering (on average introduced plants flowered 2 days earlier than those collected from the range of origin) and days from flowering to seed set (on average introduced plants set seed 6 days later than those collected from the range of origin). The nested population effect was significant for days to bolting, indicating a moderate degree of variation for this trait within some of the study species (Table 2). There were no significant invasive status, genus x invasive status, and block effects for any of the phenological traits.
Architectural and size traits
The main effect of genus was significant for stem diameter and number of rosette leaves and marginally significant for number of basal stems (Table 3). Crepis species had more rosette leaves, and more, but thinner basal stems than Centaurea species (Fig. 2).
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The main effect of provenance was significant for the number of leaves at bolting (on average, introduced plants had 11 fewer rosette leaves than those collected from the range of origin, Table 3). The nested population effect was significant for number and length of basal stems, number of rosette leaves, and biomass, indicating a moderate degree of variation for this trait within some of the study species (Table 3).
The significant block effect on biomass (Table 3) was most likely the result of mortality of some of the smallest species in some blocks coupled with a loss of some of the larger species in others, although it may also have been due to microenvironmental heterogeneity in the greenhouse setting. There were no significant invasive status or genus x invasive status effects for any of the architectural or size traits.
Fitness traits
Analyses of variance of the fitness traits showed that the main effect of genus was marginally significant for involucre diameter (Table 4). Centaurea species generally had larger seed heads than Crepis species (Fig. 3).
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The nested population effect also was significant for involucre diameter, indicating a moderate degree of variation for this trait within some of the study species (Table 4). There were no significant provenance, invasive status, genus x invasive status, or block effects for any of the fitness proxy traits.
Given that for the least consistently significant of the model effects (excepting the block and the provenance effects for which we had no a priori expectations), we found three significant tests (K = 3) for a specific effect (as for the genus main effect) of 11 total tests (N = 11) (Tables 2–4), the probability that this would happen by chance is p = 0.01 (Moran, 2003
). Because we found more than three significant results for all other effects showing significance, we feel confident that our analyses are robust against Type I errors.
Multivariate analysis
A discriminant analysis performed on the genus x invasive status combinations (i.e., invasive Centaurea, noninvasive Centaurea, invasive Crepis, and noninvasive Crepis) yielded a first discriminant function that primarily separated the groups according to genus (accounting for 88.7% of the variance; Fig. 4a, b). A second discriminant function largely separated invasive Crepis from noninvasive congeners (6.8% of explained variance; Fig. 4a), and a third discriminant function partially distinguished invasive Centaurea from noninvasive congeners (4.5% of total explained variance; Fig. 4b).
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DISCUSSION
Overall, our results suggest that while there are different characteristic phenotypes within Centaurea and Crepis, there are few phenotypic traits that distinguish invasive and noninvasive introduced species in these genera. This lack of a distinguishable invasive phenotype was apparent across genera (i.e., phenotypes shared by invasive species at the subfamily or tribe level) and within them (i.e., genus-specific invasive species patterns). Despite the fact that much of the variation in traits occurred between the two genera, there was substantial variation among congeners for many traits. Given this, any causal explanation of variation in invasion success related to "traits of invasives" must either reside in traits not examined in this study or as more complex interactions of the traits assessed here. Of course, these possibilities are not mutually exclusive.
Phenological traits
In his list of the traits of the hypothetical ideal weed, Baker (1965)
suggests that, all else being equal, invasive species might be expected to have rapid seedling growth and spend less time in the vegetative condition. This is quite reasonable because shortened development times might increase the number of generations per year (in annuals), reduce the likelihood of mortality before reproductive maturity, or both. While it has been claimed that a reduced juvenile period is widespread among invasive species (Rejmanek and Reichard, 2001
), appropriately controlled and replicated studies have been largely restricted to long-lived species (Richardson et al., 1990
; Rejmanek and Richardson, 1996
; Grotkopp et al., 2002
; Bellingham et al., 2004
) or measures of relative growth rate (Burns, 2004
).
In the current study we found no detectable differences in the phenologies of introduced invasive species when compared to introduced noninvasives, across or within the two genera examined. On average, the development of an invasive Crepis or Centaurea was no more rapid than that of a noninvasive congener (Fig. 1).
Architectural and size traits
Despite the aforementioned lack of shorter developmental schedules of invasive species in our study system, it is possible to imagine that invasives could still grow larger, more robust plants than noninvasives in the same growing season. With respect to the height of invasive species, a review of empirical studies by Kolar and Lodge (2001)
provides limited evidence that, in some cases, invasive species are taller than native or noninvasive species. However, in the recent analysis of Lloret et al. (2005)
there is no observed relationship between invasiveness and growth form or stem height within the alien flora of the Mediterranean islands.
Our results are similar to those of Lloret et al. (2005)
; we found no detectable correlations with introduced species invasiveness in any of the size or architecture traits examined. There was very little variation among congeners with respect to stem length or diameter, and where there was appreciable variation among congeners for architectural traits (number of basal stems and branches), no apparent trends distinguished invasive species from noninvasive ones (Fig. 2).
Fitness traits
It seems logical that invasive species would differ significantly from noninvasive species with respect to fitness traits, particularly if they are similar in many other respects, as is largely the case among our congeners here. Indeed, Baker (1965)
included this on the list of traits of the ideal weed (partitioned into high seed output, seed output under a wide range of environmental conditions, and seed and vegetative components of reproduction). While some researchers have examined seed mass in similarly controlled studies (Rejmanek and Richardson, 1996
; Grotkopp et al., 2002
; Lloret et al., 2005
) and others have considered survival (Gerlach and Rice, 2003
; Bellingham et al., 2004
) or germination rate (Gerlach and Rice, 2003
; Mandak, 2003
), only Gerlach and Rice (2003)
were able to measure lifetime reproductive output.
Gerlach and Rice (2003)
found appreciable differences in the number and mass of inflorescences in their field study (including both C. solstitialis and C. sulphurea, also used here), and these differences corresponded, to a degree, with the invasive status of the species in question (the most invasive species, given as C. solstitialis, had both more numerous and massive inflorescences than less or noninvasive species; however, the moderately invasive C. melitensis did not differ from the noninvasive C. sulphurea with respect to the number of inflorescences and had less massive seed heads). Our results, while indicating significant variability among congeners in involucre diameter and number of inflorescences, revealed no general pattern with respect to invasive status (Fig. 3). Although the species in each genus that produced the most inflorescences was an invasive, there were also invasive species that did not differ significantly from the congener with the lowest average reproductive output. In the particular case of C. solstitialis and C. sulphurea, we did not detect a significant difference in the number of inflorescences (Fig. 3), as Gerlach and Rice (2003)
observed in their study of these species. The substantial differences in growing conditions between our study (largely benign greenhouse conditions) and that of Gerlach and Rice
(multiple field environments) may explain these discrepancies.
Limitations
Our common garden study differs significantly from most previous assessments of the traits of invasive species in that the inclusion and replication of introduced noninvasive species allowed us to explicitly address the differences between successful and unsuccessful invasions. While this is important, our methods were not without drawbacks. The common garden design and directed taxonomic focus left us with a suite of species that largely had similar life histories, necessarily reducing our ability to comment on the likely importance of these traits. We recognize that variation in life history traits not examined here (e.g., perennation, reproductive biology, vegetative habit, dispersal mechanisms) may often trump variation in quantitative traits when trying to explain range or habitat expansion (see, for example, Gerlach and Rice, 2003
; Mandak, 2003
; Sutherland, 2004
; Lloret et al., 2005
). However, that the suite of species used here largely shares a common reproductive biology, growth form, and habitat, yet includes both invasives and noninvasives, suggests that more than life history variation contributes to differential invasiveness.
Another consequence of our common garden approach is that we necessarily included fewer species than broader comparative analyses of data compiled from multiple field, garden, greenhouse, herbaria, floras, or other accounts (e.g., Sutherland, 2004
; Lloret at al., 2005
). While this limits our ability to generalize, it helps us avoid the confounding effects resulting from variation in the environments, localities, or studies from which more comprehensive species accounts may be assembled.
The current study is also limited in its characterization and comparison of species because all plants were grown in a single, largely benign, greenhouse environment. Our understanding of the role of phenotypic plasticity in the differential success of plant invasions, addressed as early as 40 years ago by Baker (1965)
, has recently been expanded by various empirical studies (Williams and Black, 1994
; Pattison et al., 1998
; Milberg et al., 1999
; Schweitzer and Larson, 1999
; Kaufman and Smouse, 2001
; Gerlach and Rice, 2003
; Burns, 2004
; DeWalt et al., 2004
; Gleason and Ares, 2004
; Suding et al., 2004
; Wilson et al., 2004
; Brock et al., 2005
; Hastwell and Panetta, 2005
; Leishman and Thomson, 2005
). We plan to broaden our own efforts by exploring the sensitivities of the current findings to relevant environmental variability.
Conclusions
That we did not detect any general relationships between the invasive status and individual quantitative traits (fitness or otherwise) is surprising given how similar (and hence comparable) the study species are with respect to life history and growth form. Insofar as this study successfully maximized the degree of "all else being equal" among our invasive and noninvasive introductions, these results suggest the following: (1) The differences between closely related invasive and noninvasive introductions may not be properties of the species, as much as they are the results of specific and contingent introduction histories (Colautti
and MacIsaac, 2004; Puth and Post, 2005
). (2) To the extent that causally relevant phenotypic differences exist between closely related invasive and noninvasive, introduced species, these differences are still likely to be the result of multiple trait interactions and do not necessarily result from straightforward fitness differences (Rejmanek, 2000
; Grotkopp et al., 2002
).
While the study of species invasions will continue to benefit from more in-depth case studies, the appropriateness of "control species" will be a critical factor determining how much any analysis contributes to our understanding of specific cases and general patterns.
FOOTNOTES
We thank K. Kennard, L. Smith, and A. Muth for assistance in the greenhouse and field, K. McFarland for technical assistance, and J. Hodges and J. D. Rule for site access. We also thank A. Liston and J. DiTomaso for assistance in acquiring various seed accessions as well as valuable population distribution information. J. Banta, O. Bossdorf, C. Richards and several anonymous reviewers provided valuable feedback on the analyses and manuscript. The Kew Royal Botanical Gardens, Marburg University Botanical Garden, Botanical Garden of the University of Göttingen, Civico Orto Botanico, Trieste, National Botanical Garden of Belgium, Friedrich-Schiller-Universität, and the USDA National Seed Storage Laboratory all provided various seed accessions. This work was financially supported by The University of Tennessee Department of Ecology and Evolutionary Biology and Department of Botany as well as the U.S. National Science Foundation (grants DEB0089493 and IBN0321466). 
2 Author for correspondence (e-mail: nmuth{at}life.bio.sunysb.edu
)
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