Journal of Vegetation Science && (2012)

Limited effects of dominant species population sourceon community composition during communityassembly

David J. Gibson, Sara G. Baer, Ryan P. Klopf, Lewis K. Reed, Ben R. Wodika & Jason E. Willand

Keywords

Community assembly; Dominant species;

Extended phenotype; Grassland; Restoration;

Seed pool; Tallgrass prairie

Nomenclature

Mohlenbrock (2002)

Received 6 March 2012

Accepted 3 August 2012

Co-ordinating Editor: Sandor Bartha

Gibson, D.J. (corresponding author,

djgibson@siu.edu), Baer, S.G. (sgbaer@siu.

edu), Klopf, R.P. (rklopf@gmail.com),

Reed, L.K. (lewiskreed@gmail.com),

Wodika, B.R. (BenWodika@gmail.com) &

Willand, J.E. (jwilland@siu.edu):

Department of Plant Biology and Center for

Ecology, Southern Illinois University

Carbondale, Carbondale, IL, 62901-6509, USA

Abstract

Question: To what extent do dominant species population sources and subordi-

nate species pools affect diversity and composition of an assembling grassland

community?

Location: Illinois, USA.

Methods: Percentage cover of all species were recorded annually in 36 1-m2

quadrats assigned to a factorial combination of dominant species population

source (functionally distinct cultivar or non-cultivar seed source) and designed

species pool (three levels varying in species identity, but with equal functional

group representation and richness) during the first 4 yr of community assembly

in an experimental grassland restoration.

Results: Univariate and multivariate analyses showed that individual species

abundance, life form and community composition differed significantly among

designed species pools, but were not strongly affected by population source of

the dominant species (cultivar or non-cultivar). There were fewer C4 species in

cultivar plots but only in one of three designed species pools during one of 4 yr

of community assembly. The number of legume and forb species was higher in

cultivar plots, but also only in one of the 4 yr of study. Other changes in species

richness and abundance were solely related to successional change.

Conclusions: Non-dominant species introduced to restore plant communities

strongly affects plant community composition, and composition can show fidel-

ity to designed species pools. Only marginal or temporary effects of dominant

species seed source were observed in the assembling plant community. Thus, we

found no strong evidence that the source of dominant species, in this case culti-

vars compared to local ecotypes, has consequences for community assembly in

the early stages of restoration (1–4 yr). The absence of a strong dominant spe-

cies source effectmaybe exacerbatedby the assembly of diverse plant communities,

resulting in a stronger effect of subordinate species seedmixture in restoration.

Introduction

There is increasing evidence that community assembly is

related to abiotic and biotic filters that constrain success of

species that are able to colonize from the regional species

pool (Baer et al. 2004; Hobbs & Norton 2004; Chase et al.

2005; Gibson et al. 2011). Niche-based models recognize

environmental (abiotic) filters and processes of limiting

similarity (i.e. biotic filters such as competition, facilitation

and mutualism) as being important determinants of com-

munity assembly (Abrams 1983). Within a site, where

environmental conditions are fairly homogeneous, com-

munity heterogeneity can arise from local species sorting,

partly determined by functional traits and phylogeny

(Houseman & Gross 2011; Pavoine et al. 2011; Weiher

et al. 2011).

The importance of intra-specific variation within the

dominant, matrix species on community assembly repre-

sents a poorly understood aspect of limiting similarity.

Dominant species are generally abundant, relatively large

and, as individual species, make a substantial contribution

to primary production. Subordinate species are generally

Journal of Vegetation ScienceDoi: 10.1111/j.1654-1103.2012.01475.x© 2012 International Association for Vegetation Science 1

relatively smaller in stature, less abundant and occupy

more restricted microhabitats (Grime 1998), but can con-

tribute most to floristic diversity, particularly in grassland

(Gibson 2009). The mass ratio hypothesis (Grime 1998)

suggests that the traits and functional diversity of domi-

nant species largely determine ecosystem function. These

dominant species can constrain subordinate species success

and diversity (Baer et al. 2004; McCain et al. 2010), with

consequences for ecosystem processes such as primary pro-

duction (Polley et al. 2007; Sasaki & Lauenroth 2011).

Intra-specific variation in dominant species can scale to

affect associated communities, and has been referred to as

the ‘extended phenotype hypothesis’ (Dawkins 1982).

This concept suggests that high levels of genotypic and

functional phenotypic diversity of dominant species can

have considerable ecological influence on assembling com-

munities (Vellend 2006; Fridley & Grime 2010; Vellend

et al. 2010; Gibson et al. 2012) at multiple scales (Bangert

et al. 2008). For example, distinct genotypes of Populus,

raised in a common garden, differentially influenced asso-

ciated foliar insect communities, and the litter of different

genotypes supported different communities of decomposer

microorganisms (Whitham et al. 2003). However,

whether variation in dominant plant species affects

within-trophic level community structure during commu-

nity assembly offers a novel context to test whether intra-

specific variation can lead to an extended phenotype.

Grassland habitats offer a dynamic, non-equilibrium

model system in which to test hypotheses of community

assembly (Collins et al. 1998). However, much of this eco-

system has been lost due to anthropogenic influence on

the environment (largely for agriculture) in most parts of

the world (Ellis & Ramankutty 2008), necessitating resto-

ration programmes to increase the extent of grassland and

improve the quality of degraded remnants (White et al.

2000; Gibson 2009). Addition of propagules (e.g. seeds) is

the most common approach to restoring and steering the

recovery of grassland communities degraded through agri-

culture. The species mix should influence composition of

the assembling community, and dominant species (partic-

ularly matrix grasses) are likely to modulate resultant

community diversity (Baer et al. 2004). However, the role

of dominant species in community assembly, especially

aspects of limiting similarity (Abrams 1983) and the exclu-

sion of subordinate species, is not well understood. Two of

these aspects that are likely to be of particular importance

are the population source of the dominant species (Bangert

& Whitham 2007), and inter-specific interactions arising

from propagules in the regional species pool (Zobel 1997)

introduced to restore a community. In the context of

restoration, this means that the seed source of dominant

species and composition of the sown species mix could sig-

nificantly influence restoration trajectory, diversity and

ecosystem function (Lesica & Allendorf 1999; Zedler 2005;

Wilsey 2010; Larson et al. 2011). Through competitive

and facilitative interactions, both the genetic identity (pop-

ulation source) of dominant species and the local species

pool can act as biotic filters constraining community

assembly (Whitlock et al. 2007; Gibson et al. 2012).

The objective of this study was to conduct an experi-

mental test of the applicability of the ‘extended pheno-

type hypothesis’ in assembling grassland communities

sown with different population sources of commonly

dominant grass species. Preliminary evidence demon-

strates variation in genetic structure and functional traits

between population sources of the dominant grasses.

Specifically, we investigated whether dominant species

sources differentially affect diversity and composition of

an assembling grassland community. We have demon-

strated that cultivars of the dominant grasses used in this

experiment were genetically distinct from local popula-

tion sources (Gustafson et al. 2004b). Further, we have

documented that cultivar and non-cultivar sources of the

dominant grasses varied in relevant functional traits,

including enhanced leaf-level photosynthetic rates (Lam-

bert et al. 2011) and increased root length and root

surface area (Klopf & Baer 2011). The cultivar and non-

cultivar population sources affect inter-specific competi-

tive interactions (Gustafson et al. 2004a). Specifically, we

contrasted the effect of sowing cultivars of the dominant

grasses selected and recommended for use in the region

of our study site with locally collected, non-cultivar dom-

inant grass sources on the developing structure of plant

communities using three distinct pools of subordinate

species. The three species pools sown into the source of

dominant grasses was intended to elucidate whether

dominant grass source effects were a general phenome-

non (consistent across species pool) or contingent upon

inter-specific interactions. We measured species composi-

tion during the first 4 yr of community assembly in

response to dominant grass source and designed species

pool (DSP) to test three hypotheses: Our first hypothesis

(H1) was that prairie seeded with cultivars of the domi-

nant grasses would obtain a different (i.e. more grass

dominated and less diverse) community composition

over time compared with prairie seeded with locally col-

lected, non-cultivar seed sources. We expected a differ-

ence to arise because the cultivars have been selected for

agronomic traits of high productivity, seed viability, ger-

mination, as well as drought tolerance and disease resis-

tance (Alderson & Sharp 1995). We expected cultivars to

be more competitive than wild-collected sources (Gustaf-

son et al. 2004a), allowing them to attain greater domi-

nance than local ecotypes of the native grasses. Second,

we hypothesized (H2) the effect of dominant species

source would be consistent across DSPs of subordinate

Journal of Vegetation Science2 Doi: 10.1111/j.1654-1103.2012.01475.x© 2012 International Association for Vegetation Science

Grassland assembly D.J. Gibson et al.

species if the effect predicted in H1 is a general phenome-

non. Lastly, we hypothesized (H3) that the effect predicted

in H1 (greater grass dominance in cultivar- relative to

non-cultivar-sown plots) would reduce divergence in

community composition among the DSPs seeded with cul-

tivars relative to DSPs seeded with non-cultivars. More

specifically, we predicted differences in community com-

position would diminish among DSPs over time as the

contrasting effect of cultivar vs native population sources

of the dominant species increased over time and seeded

species became more prominent in the community.

Methods

In March 2006, we established an experimental tallgrass

prairie restoration in a former agricultural field at the

Southern Illinois University Agronomy Research Center in

Jackson County, Illinois, USA (37°41′ N, 89°14′ W). Soil

at the field site was a silty clay loam (Herman 1979). The

climate has had an average annual temperature of 13.4 °C(average minimum 7.4 °C, average maximum 19.4 °C),with average annual rainfall of 1212 mm, of which an

average of 52% fell during the growing season from 1 April

through 30 September (1990–2009 record, Carbondale,

IL). Total precipitation received each year of this study

(2006, 2007, 2008 and 2009) was 1475, 1084, 1492 and

1545 mm, of which 45, 42, 47 and 61% of total precipita-

tion was received during the growing season, respectively.

The restoration experiment was a split-plot design

(Appendix S1), with dominant grass population source

(cultivar or non-cultivar) assigned as the whole-plot factor

addressing H1 according to a randomized complete block

design (six whole plots per source population; n = 12).

We blocked according to former agricultural field use;

each block contained six 7m 9 23 m whole plots. The

source population treatment comprised three dominant

tallgrass prairie species: Andropogon gerardii Vitman,

Sorghastrum nutans (L.) Nash. and Schizachyrium scoparium

(Michx.) Nash. These three grasses were chosen because

they are the most frequent and dominant species in the

tallgrass prairie region (Weaver 1954; Dodd 1983).

Hypothesis 2 was addressed through a subplot factor of

DSP consisting of three pools containing unique species

randomly assigned to one of three 5m 9 5 m subplots

within each whole plot (n = 36, 12 subplots per DSP).

Each DSP consisted of 15 species and the same number of

species represented in general functional groups: C4 grass,

C3 grass, legumes and non-legumine herbaceous forbs

(Table 1). All of the sown species were native to the

region and occur in tallgrass prairie (Mohlenbrock 2002).

Each of the three DSPs was equivalent in richness and

coarsely defined functional groups, but varied in terms of

species identity, and hence any emergent functional group

traits. The experiment was monitored for the first 4 yr of

community assembly, allowing a test of H3.

Non-cultivar (local ecotype) seeds of the dominant

grasses for the whole plot treatment were collected from

nearby prairie remnants within 100 km of Carbondale, IL.

A small portion (7%) non-cultivar ‘Missouri ecotype’ seed

ofA. gerardiiwas included in the non-cultivar seedmix due

to limited availability of viable non-cultivar A. gerardii seed

from local prairie remnants. We used cultivars recom-

mended and widely sown in the region: A. gerardii ‘Roun-

tree’, S. nutans ‘Rumsey’ and S. scoparium ‘Aldous’. We

seeded 100 live seeds of each dominant grass species per

square meter. Cultivar percentage live seed (PLS) was pro-

vided by a native seed supplier (Hamilton Seed Co., Elk

Creek, MO, US). Non-cultivar seed was tested for PLS at a

seed testing lab (Hulsey Seed Laboratory, Inc., Decatur, GA,

US) according to the guidelines of theAssociation ofOfficial

Seed Analysts (AOSA). Seeds of subordinate species for the

subplot DSP treatment were obtained from the same native

Table 1. Designed species pools assigned to subplots within each whole

plot seeded with either cultivars or non-cultivars of the dominant prairie

grasses.

Designed species

pool 1

Designed species

pool 2

Designed species

pool 3

Non-legume Forbs

Asclepias tuberosa Achillea millefolium Brickellia

eupatorioides

Asclepias verticillata Asclepias syriaca Aster

oolentangiensis

Aster oblongifolius Echinacea purpurea Heliopsis

helianthoides

Callirhoe involucrata Eupatorium

altissimum

Monarda fistulosa

Delphinium

carolinianum

subsp. virescens

Liatris pycnostachya Penstemon digitalis

Oenothera

macrocarpa

Oenothera biennis Rudbeckia hirta

Ratibida pinnata Oligoneuron rigidum Silphium laciniatum

Rosa arkansana Ruellia humilis Solidago speciosa

Senecio plattensis Silphium integrifolium Vernonia fasciculata

Legumes

Baptisia alba var.

macrophylla

Amorpha canescens Astragalus

canadensis

Chamaecrista

fasciculata

Dalea candida Baptisia bracteata

Psoralidium

tenuiflorum

Desmanthus illinoensis Lespedeza capitata

Schrankia nuttallii Desmodium illinoense Dalea purpurea

C3 Grass

Elymus canadensis Koeleria macrantha Agrostis hyemalis

Non-dominant C4 Grass Non-dominant C4 Grass Non-dominant C4

Grass

Sporobolus heterolepis Bouteloua curtipendula Panicum virgatum

Journal of Vegetation ScienceDoi: 10.1111/j.1654-1103.2012.01475.x© 2012 International Association for Vegetation Science 3

D.J. Gibson et al. Grassland assembly

plant supplier (Hamilton Seed Co.). Subordinate species

unique to eachDSPwere seeded at a rate of 20 seed�m�2 per

species (total = 300 seed�m�2of subordinate species) follow-

ing recommendations for mesic prairie restorations (Diboll

2005). Live seedwas not determined for subordinate spe-

cies, butwas consistent amongsubplots assigned toeachDSP.

Prior to seeding, the field site was tilled using a tandem

disk and field cultivator. Seeds of the dominant grasses

and subordinate species were mixed with wet sand and

hand broadcast into plots on 1 Mar 2006. Following seed-

ing, the plots were manually compacted to ensure seed

contact with the soil. The 6-m buffers around the whole

plots were seeded with two native grasses, Elymus canaden-

sis L. and Bouteloua curtipendula (Michx.) Torr, using a

tractor-pulled Brillion Alfalfa seeder. In subsequent years,

2-m wide pathways were mowed in the buffers to main-

tain open walkways. Following standard practice for

early-stage prairie restorations (Packard & Mutel 2005),

the entire site was burned annually in the dormant season

to control volunteers and enhance establishment of the

native prairie species.

Data collection

From 2006 through 2009, canopy cover (%) and species

richness (number per m2) were recorded from early and

late summer surveys in permanently marked 1-m2 quad-

rats (n = 36) located in the centre of each subplot. Maxi-

mum cover of each species from the two surveys was used

for all community analyses. Nomenclature and exotic spe-

cies status followsMohlenbrock (2002).

In addition to examining total cover and total richness,

individual species were categorized into the following spe-

cies groups: sown species, sown dominant grass species

(sum cover of A. gerardii, S. nutans and S. scoparium),

design pool species (i.e. species in DSPs 1, 2 and 3), volun-

teers (i.e. unsown species regardless of nativity), grasses,

forbs, legumes, exotics (i.e. non-native volunteers), C3

species and C4 species. Shannon’s diversity (H′) and Simp-

son’s evenness indices were calculated in DECODA.

Data analysis

We analysed the effects of seed source (cultivars or non-

cultivar seed), DSP (three pools), year (2006–2009) and

possible interactions among these three factors according

to a split-plot design, with the whole-plot factor assigned

according to a randomized complete block design and year

as the repeated measure using the mixed model procedure

in SAS (SAS Institute Inc., Cary, NC, US; 2002–2008).

Data were log-transformed prior to analyses. Degrees of

freedom were calculated using the Kenward–Roger

estimation recommended for repeated measures using

mixed-model analyses (Littell et al. 2006). The most

appropriate covariance structure for each model was

determined by comparing AIC, AICC and BIC model fit

statistics for unstructured, compound symmetry and first-

order autoregressive covariance structures. Least-squares

mean comparisons were used to test for differences among

sources, species design pools, year and interactions

between these factors (a = 0.05).

Non-metric multidimensional scaling (NMDS) ordina-

tion analysis was conducted on the species cover by plot

matrix using DECODA software. NMDS is considered to be

the most robust method of ordination (Minchin 1987).

Species cover values were standardized to unit maxima,

and species occurring in <15 plots (i.e. 10% of the plots

over all years) were excluded from the analysis. NMDS

was run using 100 random starting configurations on the

Bray–Curtis dissimilarity measure. The relationship of the

abundance of species to the retained ordination solution

was assessed by calculating species scores for each species

in the NMDS space. The species scores were calculated as

the weighted average of the abundance scores of the sam-

ples in which the species occurred in each dimension.

These weighted averages were used to plot species as

points in the NMDS ordination and are referred to as spe-

cies centroids because they show the centre of the species’

distribution with respect to the ordination axes. The rela-

tionship between species functional groups (Table 2) and

the ordination solution was tested by fitting vectors of

maximum correlation (Faith & Norris 1989; Kantvilas &

Minchin 1989). Vector significance was assessed following

permutation tests to generate correlation values. Vectors

significantly correlated with the ordination were retained

for plotting in ordination space relative to the ordination

centroid. Analysis of similarity (ANOSIM) was used to test

for the relationship between seed source, DSPs and block

and the ordination solution retained for interpretation.

Results

Cover, richness and diversity

Over the 4 yr of study, 126 species of vascular plant were

recorded in the plots, with a decline from 85 species in

2006 to 53 in 2009. In the first year of establishment, the

plots were dominated by agricultural weeds, including Pa

nicum dichotomum (frequency = 75% of plots, mean cover

= 16%), Bromus inermis (frequency = 69%, cover = 8.8%)

and Digitaria sanguinalis (frequency = 100%, cover =7.0%). The dominant grasses established at a high fre-

quency (occurrence in sampling plots) in 2006, albeit with

lower cover than expected (Andropogon gerardii:

frequency = 78%, cover = 1.1 ± 0.2%; Sorghastrum nutans:

frequency = 86%, cover = 0.9 ± 0.1%; Schizachyrium scop

arium: frequency = 92%, cover = 1.2 ± 0.1%). By 2009,

Journal of Vegetation Science4 Doi: 10.1111/j.1654-1103.2012.01475.x© 2012 International Association for Vegetation Science

Grassland assembly D.J. Gibson et al.

the plots were dominated by sown prairie species, includ-

ing the dominant grasses (S. nutans: frequency = 89%,

cover = 29.3 ± 3.0%; A. gerardii: frequency = 83.3%,

cover = 16.9 ± 2.2%; and S. scoparium: frequency = 42%,

cover = 3.1 ± 0.8%) and sown forbs (e.g., Rudbeckia hirta:

frequency = 50% despite being seeded into only 33% of

the plots, cover = 5.6 ± 1.4%; Solidago canadensis: fre-

quency = 64%, cover = 4.5 ± 1.0%; and Silphium inte-

grifolium: frequency = 36%, cover = 7.4 ± 2.0%). Some

agricultural weeds persisted in the plots into 2009, e.g. Stel

laria media (frequency = 75%, cover = 1.3 ± 0.2%) and

Trifolium repens (frequency = 61%, cover = 12.8 ± 3.8%).

There was no effect of seed source on the cover of any

plant groups examined. Change in cover was significant

over time (i.e. main effect of year) for most plant categories

(Table 2). Cover of the dominant grasses, with the excep-

tion of S. scoparium, increased from 2006 through 2009 in

all of the DSPs and as a group, as did total cover (Fig. 1a,b),

and cover of sown species, forbs, legumes, all C4 and all C3

species (Fig. 1c). In contrast, cover of volunteers decreased

by more than 50% between 2006 and 2009 (57.4 ± 5.4%

to 25.7 ± 5.1%) and the cover of exotics decreased from

18.2 ± 2.5% in 2006 to 13.4 ± 3.7% in 2007 (with no

significant change thereafter; Fig. 1c).

Table 2. Repeatedmeasures mixed-model analysis of species functional groups based on (a) cover, (b) richness and (c) diversity and evenness.

Functional group Covariance

structure

Source (S) Designed

Species

Pool (DSP)

Year (Y) S*DSP S*Y DSP*Y S*DSP*Y R-value

(a)

df 1,9 2,20 3,90 2,20 3,90 6,90 6,90

Total cover cs 0.00 0.06 42.52*** 0.02 0.70 0.42 0.78 0.62***

Sown spp. cs 0.76 1.59 480.63*** 0.50 1.83 0.60 1.24 0.86***

Grasses un 0.4930 3.6730* 18.2128*** 0.4030 2.2128 0.8735.9 0.4335.9 0.48***

Forbs un 0.0211.7 1.9917.7 41.1228*** 0.5617.7 1.1228 1.8335.9 0.6835.9 0.67***

Legumes un 0.5527.6 2.2327.6 30.4228*** 0.0927.6 0.4528 1.4535.9 0.4535.9 0.42***

C4 un 1.0727.8 0.2727.8 19.6228*** 0.4227.8 1.9928 1.5735.9 0.6335.9 0.72***

C3 ar(1) 0.079.11 0.1621 8.1881.1*** 0.6721 0.6581.1 1.2184.1 0.4584.1 0.50***

Pool species ar(1) 3.78† 0.329.5 123.0588.2*** 0.0525 2.5388.2† 4.9590.3** 0.7390.3 0.62***

Volunteers un 0.4611.9 1.1215.1 17.1928*** 1.0415.1 0.6828 0.5835.9 0.2435.9 0.59***

Annuals un 0.2430 3.2830† 49.8828*** 0.3330 0.6228 3.0635.9* 0.9335.9 0.49***

Exotics un 2.210.1 2.378.5 6.5428** 1.408.5 1.7928 0.3435.9 0.5735.9 0.51***

Andropogon gerardii un 0.9518.3 1.2925.7 39.4328*** 1.1225.7 0.3728 1.0435.9 1.2435.9 –

Sorghastrum nutans un 0.2030 0.4330 56.2628*** 0.2430 1.7628 0.6135.9 0.7035.9 –

Schizachyrium scoparium un 0.6324.2 0.1821.5 7.5628** 1.2921.5 0.1528 0.7135.9 0.4135.9 –

Total dominant grasses un 0.6019.6 0.2719.7 67.7528*** 0.7019.7 2.4728† 1.1435.9 1.3635.9 0.83***

(b)

df 1,9 2,20 3,90 2,20 3,90 6,90 6,90

Total richness cs 0.17 7.17** 121.33*** 1.73 0.48 0.62 0.33 0.80***

Number sown species ar(1) 1.8836.3 5.5236.3** 4.2987.1** 1.7536.3 0.8187.1 0.4589.9 1.0889.9 0.27**

Grasses un 3.7720.2† 2.2922.7 88.028*** 0.9022.7 2.0228 0.5535.9 0.4235.9 0.72***

Forbs cs 1.91 14.95*** 91.04*** 0.64 2.72* 1.02 0.41 0.75***

Legumes ar(1) 0.4636.4 5.7536.4** 3.4590.3** 1.0536.4 3.5890.3* 0.7191.9 0.3791.9 0.41***

C4 cs 2.71 3.91* 34.57*** 4.75* 3.75* 0.69 0.69 0.65***

C3 ar(1) 1.109.01 12.3636.6*** 94.1683.9*** 0.4936.6 2.0883.9 1.0088.1 0.4388.1 0.77***

Pool species ar(1) 3.1432.7† 5.532.7** 2.6887.4* 0.6232.7 1.5087.4 0.8889.2 0.7589.2 0.30**

Volunteers un 0.479.51 1.1625.6 274.2928*** 0.9525.6 0.4128 1.6735.9 0.7435.9 0.87***

Annuals cs 0.63 9.17** 160.13*** 1.98 0.19 1.74 0.91 0.87***

Exotics un 4.1215.3† 0.0917.5 113.5628*** 1.6517.5 0.1025 0.5035.9 2.2735.9

† 0.70***

(c)

df 1,9 2.20 3,90 2,20 3,90 6,90 6,90

H′ un 0.0310 1.2220.3 34.6528*** 0.3120.3 0.4428 0.3835.9 0.8635.9 0.55***

Evenness cs 0.00 0.22 0.76 0.12 0.25 0.82 0.12 0.06

Values are F-statistics for species functional groups, and R correlation values for vector analyses of species functional groups with the NMDS ordination.†P � 0.1, *P � 0.05, **P � 0.01, ***P � 0.001. Covariance structure, cs = compound symmetry, un = unstructured, and ar(1) = first-order autore-

gressive. Degrees of freedom (df) calculated according to the Kenwood–Roger (KR) correction, which uses an invariant denominator df for each effect with

a compound symmetry covariance structure, but is variable for un and ar(1) covariance structures (shown as subscript following F-statistics).

Journal of Vegetation ScienceDoi: 10.1111/j.1654-1103.2012.01475.x© 2012 International Association for Vegetation Science 5

D.J. Gibson et al. Grassland assembly

Designed species pools differentially affected total grass

cover across all years (Table 2a). There was significantly

higher cover of grasses in DSP 1 (50.2 ± 3.9%) than DSP 2

(38.7 ± 3.4%) over all years resulting from the establish-

ment success of S. nutans and Elymus canadensis seeded in

DSP 1 (Cover in 2009, DSP 1 = 32.3 ± 5.8% and

3.3 ± 1.5%, respectively, DSP 2 = 26.3 ± 5.2 and

0.2 ± 0.2%, DSP 3 = 27.1 ± 5.4% and 0.3 ± 0.3%). Cover

of designed pool species and annuals exhibited an interac-

tion with year resulting from differences in cover of

designed pool species in 2008 and 2009, respectively,

which did not occur in other years. More specifically, in

2008 the cover of designed pool species was significantly

higher in DSP 3 (54.4 ± 5.6%) compared to DSP 1

(36.7 ± 4.4%); DSP 2 was intermediate at 40.4 ± 4.9%. In

2009, the cover of annual species was highest in DSP 1

(17.1 ± 7.6%) compared with both DSP 2 (2.8 ± 0.7%)

and DSP 3 (1.9 ± 0.4%).

Source of dominant grasses did not influence total rich-

ness, but weakly (0.10 > P > 0.05) affected the richness of

grasses, designed pool species and exotics over all years

(Table 2b) and richness of forbs, legumes and C4 grasses in

some years. Total grass richness was higher with the non-

cultivar seed source (5.3 ± 0.3 grass species�m�2) com-

pared to growing with the cultivar seed source

(4.7 ± 0.3 spp.�m�2), whereas richness of designed pool

species was lower in the presence of non-cultivar grasses

(5.9 ± 0.2 spp. m�2) than cultivars (6.7 ± 0.2 spp. m�2).

The number of exotic species was also marginally affected

by seed source, with fewer (2.9 ± 0.3 spp. m�2) exotic

species in plots with cultivars than the non-cultivar seed

sources (3.4 ± 0.3 species m�2). The number of legume,

forb and C4 species exhibited an interaction between

source and year resulting from different levels of richness

in association with cultivar sources in only 2007, the sec-

ond year of community assembly. In 2007, legume rich-

ness and forb richness were higher in the presence of grass

cultivars than non-cultivars (Fig. 3a,b). In contrast, the

number of C4 species was higher in plots seeded with a

non-cultivar source compared with a cultivar seed source

in this year (Fig. 2c).

Richness of C3 species, exotic and volunteer species

was strongly and only affected by year (Table 2b). Total

species richness, number of volunteers and annuals

decreased each year (Fig. 2d,e,g), and the number of

exotics decreased from 6.3 ± 0.3 species m�2 in 2006 to

2.2 ± 0.3 in 2007 (with no significant change thereaf-

ter; Fig. 2f). Richness of sown species, grasses, C3

species and designed pool species were affected by year

(C3 species and grasses decreasing; Fig. 3i,k), but also

by either seed source of dominant grasses or DSP (see

below).

Designed species pool strongly affected total richness

and richness of all species groups except the number of

grasses, exotics and volunteers (Table 2b). The total

number of species, number of sown species, forbs,

legumes, pool species and C3 species was highest in DSP 2

(Fig. 3). The number of annuals was highest in DSP 1.

There was an interaction between DSP and seed source

on the number of C4 species, with the lowest number of

Cov

er (%

)

0

10

20

30

40

50

Andropogon gerardiiSorghastrum nutansSchizachyrium scopariumTotal Dominant grasses

Cov

er (%

)

0

20

40

60

80

100

120

140Total cover Sown Designed Pool Volunteers

A

AB

C

DCB

D C

BA

B

A

A

B

A

B

AA

A

A

B B

B

C

C

D

D

C

Year2006 2007 2008 2009

Cov

er (%

)

0

20

40

60

80

ForbsLegumes C3 C4 Annuals Exotics

B

BA

A

C

BB

A

A A

BB

A

B BB

C B

AA

(a)

(b)

(c)

Fig. 1. Temporal changes in percentage cover (mean ± SE) of (a) all

species, sown, designed species pool (pool) and volunteer species, (b) the

dominant grasses collectively and by species, and (c) species groups from

2006 to 2009, corresponding to the first through fourth year of community

establishment. Means within a cover group accompanied by the same

letter were not significantly different (a = 0.05; letters not shown for cover

of designed pool species and annuals because these species groups

exhibited an interaction between time and designed pool (Table 2a).

Journal of Vegetation Science6 Doi: 10.1111/j.1654-1103.2012.01475.x© 2012 International Association for Vegetation Science

Grassland assembly D.J. Gibson et al.

these species in DSP 2 seeded with cultivar compared

with non-cultivar seed (Fig. 3h).

Diversity (H′) decreased through time but not differen-

tially among DSPs or seed sources of dominant grasses

(Table 2c). In 2006, 2007, 2008 and 2009 H′ was

2.62 ± 0.09, 2.21 ± 0.06, 2.11 ± 0.05 and 1.84 ± 0.04,

respectively, and significantly different between all years

except 2007 and 2008. Evenness was unrelated to DSP,

source or time. Average evenness of the restored commu-

nity was 0.37 ± 0.01.

Community composition

A three-dimensional NMDS ordination solution was

retained for interpretation of community change over time

and in the response to the biotic factors that were manipu-

lated (stress = 0.18). The distribution of plots in the 3-D

solution was related to year of sampling (ANOSIM

R = 0.63, P < 0.0001; Fig. 4a), DSP (R = 0.45, P < 0.0001)

and weakly to sample block (R = 0.08, P < 0.0001), but

not seed source of the dominant species (R = 0.01,

P = 0.25). Sample plots showed a trajectory through ordi-

nation space with a clear separation of DSPs (Fig. 4b). The

change in composition from 1 yr to the next decreased

from 2006 through 2009 (decreasing distance between

plots from 1 yr to the next in multivariate space; Fig. 4a).

Species centroids (data not shown) showed that the plots

were initially dominated by a high frequency of volunteer

species, including Poa annua (33% of plots in 2006, 2% in

2009) and Veronica peregrina (34% and 0%, respectively),

Year

2

4

6

8

10

0

4

8

12

16

20

24

4

8

12

16

20

24

0

2

4

6

8

10

(a) Legumes

0

2

4

6

8

10CultivarNon-cultivar

*

(b) Forbs

4

8

12

16

20

24 CultivarNon-cultivar*

(c) C4 species

Year2006 2007 2008 2009

No

spec

ies

. m–2

No

spec

ies

. m–2

No

spec

ies

. m–2

No species . m

–2N

o species . m–2

No species . m

–2N

o species . m–2

0

2

4

6

8

10 CultivarNon-cultivar

*

A

B

CD

(g) Annuals

(f) Exotics

(e) Volunteers

(d) Total richness

A

A

A

B BB

B

B

C

C

D

D

0

4

8

12

16

20

24(h) Sown species

AA BBA

4

8

12

16

20

24(i) C3 speciesA

BC

D

0

2

4

6

8

10(j) DSP species

AB A A B

Year2006 2007 2008 2009 2006 2007 2008 2009

2

4

6

8

10(k) Grasses

B B

C

A

A B AB AB

AB

CD

A

BB C

Fig. 2. Richness (mean ± SE) of (a) legumes, (b) forbs and (c) all C4 species from 2006 to 2009 by dominant grass seed source (to illustrate the source x

time interaction for these groups), (d) total richness and richness of (e) volunteers (i.e. unsown plant species), (f) exotics (i.e. non-native volunteer species),

(g) annuals, (h) sown species, (i) C3 species, (j) design pool species, and (k) grasses by year. Cultivar vs non-cultivar patterns for each year were the same

for (d–k) (i.e. non-significant interaction between seed source and year; year effect only; Table 2b). Asterisks indicate significant differences between

cultivar and non-cultivar seed sources within a year (a = 0.05) (a–c). Means accompanied by the same letter were not significantly different (a = 0.05).

Journal of Vegetation ScienceDoi: 10.1111/j.1654-1103.2012.01475.x© 2012 International Association for Vegetation Science 7

D.J. Gibson et al. Grassland assembly

but through time became increasingly dominated by the

sown species, including the three dominant grasses

(Fig. 1b). The most frequently occurring species character-

izing the plots in each designed pool were increasingly

those included in the DSP mix. By 2009, the designed pool

species with the highest cover were Baptisia alba var macro-

phylla (19.5 ± 6.6% cover), Chamaecrista fasciculata (14.0 ±7.8%) and Ratibida pinnata (14.0 ± 4.5%) in DSP 1, Silphi-

um integrifolium (19.8 ± 3.6%) and Desmanthus illinoensis

(7.9 ± 3.5%) in DSP 2, and Rudbeckia hirta (14.4 ± 2.7%)

and Silphium laciniatum (12.3 ± 1.3%) in DSP 3. Vector

analysis showed that species functional groups were signif-

icantly related to the ordination solution (Table 2; all

P < 0.01). A plot of these vectors (not shown) reflects a

contrast in composition from plots with the highest num-

ber of species overall, and in most species functional

groups, including volunteer species, to plots with fewer

species but higher cover in all groups. The vectors indicated

that while sown species increased in both richness and

cover through time, volunteers decreased in richness and

cover. Evenness was unrelated to the distribution of plots

within the ordination (vector correlation R = 0.06,

P = 0.91).

Discussion

Counter to other studies that have supported the ‘extended

phenotype hypothesis’ (Whitham et al. 2003; Bailey et al.

2009), but consistent with Wilsey (2010), we found no

strong evidence in support of our hypotheses that popula-

tion source of dominant species consistently affected plant

community composition (H1 and H2, this study) or ecosys-

tem processes (S.G. Baer, D.J. Gibson, A.M. Lambert, L.K.

Reed, R.E. Campbell, R.P. Klopf, J. Willand & B.Wodika in

(a) Total richness

No

spec

ies

. m–2

0

5

10

15

20

25

B A B(b) Pool species

0

5

10

15

20

25

BA B

(d) Sown species

No species . m

–2

5

10

15

20

BA

B

(c) Forb species

0

5

10

15

20

BA

B

(e) C3 species

0

5

10

15

20

(f) Legume species

0

5

10

15

20

BA

B

A A B

(g) Annual species

0

2

4

6

8

BBA

Designed species pool1 2 3 1 2 3

0

2

4

6

8CultivarNon-cultivar

AA AAAB B

(h) C4 species

Fig. 3. Richness (mean ± SE) of species life-form groups significantly affected by designed species pool (a–g), and (h) the interaction between designed

species pool and seed source of dominant species. Mean values with the same letter were not significantly different (a = 0.05).

Journal of Vegetation Science8 Doi: 10.1111/j.1654-1103.2012.01475.x© 2012 International Association for Vegetation Science

Grassland assembly D.J. Gibson et al.

review) during development of a diverse grassland assem-

blage. The use of DSPs reinforced the generality of the role

of this anticipated biotic filter in structuring assembling

plant communities. Dominant species population source

had only a weak effect across all DSPs on the richness of

some species groups. An overwhelmingly consistent effect

of DSP with no divergence through time in response to

population source on developing community composition

through 4 yr refuted our third hypothesis. This finding

supports models of community assembly in which individ-

ual species or a particular mix of species are given primacy,

allowing species niche partitioning or facilitative effects to

drive the development of species associations (Fridley

2001).

As expected, we observed rapid changes in community

composition during the early stages of community assem-

bly of a prairie restoration (Cottam & Wilson 1966; Camill

et al. 2004). In contrast to some restorations (Kindscher &

Tieszen 1998), we observed a decrease in diversity as the

sown, native prairie species came to dominate the commu-

nity at the expense of the early volunteer colonists. Most

importantly, the weak effect of dominant species seed

source that we observed during these first 4 yr of commu-

nity assembly does not support the commonly held dogma

regarding the importance of the seed source on commu-

nity assembly in grassland restorations, discouraging the

use of cultivars (Lesica & Allendorf 1999; Gustafson et al.

2005). Concern regarding appropriate germplasm for rein-

troduction into communities is largely based on the

genetic/population principles that founder effects deter-

mine genetic variability among colonizing species, particu-

larly dominant matrix species (Montalvo et al. 1997;

Hufford & Mazer 2003). However, this concern has not

been tested, as far as we know, with species in a restoration

and in a community common garden setting with the pres-

ence of inter-specific interactions. Further, the hierarchical

effects of different genetic structure on community assem-

bly and ecosystem processes in multi-species settings is

poorly understood. We contrasted the effects of seeding

with locally adapted cultivars vs locally collected non-culti-

var seed, which we have documented to contain a differ-

ent genetic structure (Gustafson et al. 2004b). Cultivars

have been selected for plant traits associated with ‘vigour’

and we have documented enhanced traits in cultivars

(i.e. physiological performance, root architecture and

below-ground net primary productivity) relative to non-

cultivars of the dominant prairie grasses in a common gar-

den setting (Klopf & Baer 2011; Lambert et al. 2011).

Despite trait differences documented between the popula-

tion sources, the outcome of any inter-specific interactions

between the dominant grasses and subordinate species did

not lead to any consistent differences in species composition

during the first 4 yr of community assembly. The occur-

rence of such an effectwas perhaps unlikely during the first

2 yr of community assembly because the dominant grasses

were not the most abundant species (i.e. dominant grasses

cover was <20%). However, in the third and fourth year of

restoration, native grass cover exceeded 40%, principally

due to high cover of S. nutans. Moreover, in an additional

field experiment in Illinois, seed source effects of dominant

grasses did not scale to affect species diversity (R.P. Klopf, S.

G. Baer, & D.J. Gibson unpubl.). Dynamic patterns of com-

munity change during the early phases of community

assembly are important to characterize as they may persist

much later in succession (Ejrnæs et al. 2006; Roscher et al.

2009) in the absence of disturbance (Bartha et al. 2003).

These early successional transient prairie communities are

important in the context of landscape diversity.

The different DSPs led to differences in overall commu-

nity composition (indicated by the multivariate analysis)

demonstrating that sown species influence local species

–1.5

–1.0

–0.5

0.0

0.5

1.0

1.5

–1.0

–0.50.0

0.51.0

1.5

–1.5–1.0

–0.50.00.51.0

NM

DS

3

NMDS 1

NMDS 2

2006200720082009

(a) Year of sampling

–1.5

–1.0

–0.5

0.0

0.5

1.0

1.5

–1.0

–0.50.0

0.51.0

1.5

–1.5–1.0

–0.50.00.51.0

NM

DS

3

NMDS 1

NMDS 2

123

(b) Designed species pools

Fig. 4. Three-dimensional NMDS plots (stress = 0.18) with sample plots

labelled by (a) year of sampling and (b) designed species pool.

Journal of Vegetation ScienceDoi: 10.1111/j.1654-1103.2012.01475.x© 2012 International Association for Vegetation Science 9

D.J. Gibson et al. Grassland assembly

composition, which has important implications for restor-

ing plant community heterogeneity across a landscape that

can be fraught with difficulty in restoration (Baer et al.

2005). Although the dominant grasses characterize the

system generally as tallgrass prairie, subordinate species

mixtures represented in our DSPs characterize local inter-

actions and local heterogeneity. The DSPs also differen-

tially influenced richness of several functional groups

(i.e. forbs, legumes, annuals, C3 and C4 species), despite

that the cover of each group was not affected by the DSPs.

The implications are that ‘species matter’ in this system, or

at least in the mixtures that we sowed, supporting argu-

ments about species function (Lawton 1994; Giller &

O’Donovan 2007). Even though species mixtures may

contain the same initial representation of functional

groups, as they did in our DSPs, other trait differences

among species can affect community composition (Wein-

berg 1984; Helsen et al. 2012) and drive ecosystem func-

tion (Isbell et al. 2011). In our study, four legume species,

for example, were seeded equally into each DSP, but after

4 yr came to be abundant in only two of the three pools,

likely resulting from non-equivalency in establishment

among this group. In contrast, one or more composite spe-

cies characterized each of the three DSPs after 4 yr, reflect-

ing similar patterns of colonization and perhaps increased

functional equivalency among these species (Ejrnæs et al.

2006; Harpole & Tilman 2006).

Although DSP effects on richness were predominant, a

small number of these effects did occur in the context of

the seed source of the dominant grasses (e.g. the DSP by

seed source interaction), albeit not every year. These inter-

actions are consistent with the operation of a hierarchical

species filter (Hobbs & Norton 2004; Gibson et al. 2012),

where species sorting occurs through multiple abiotic and

biotic filters. In this case, our DSPs were representative of

initial abiotic filtering via local seed pool limitations, with

later biotic filtering imposed in a weak manner by the seed

source (cultivar or non-cultivar source) of the dominant

grasses. Longer-term study of species composition is

needed to determine if weak seed source effects become

more or even less prevalent as the communities mature.

These results have relevance for practitioners and land

managers because there is an on-going debate regarding

the source of seed to use in restoration (Hufford & Mazer

2003; Gustafson et al. 2005). General concerns are that

cultivar seed sources may introduce non-local germplasm

into native communities and cause genetic swamping. We

suggest that these concerns may needmore context depen-

dency, particularly if native communities are nearly non-

existent on the landscape (e.g. grassland widely converted

to agriculture). In many instances, we do not know ‘how

local is local’ or how well suited (locally adapted) remain-

ing populations are for use in restoration (McKay et al.

2005). Less is understood about the hierarchical ecological

consequences of the genetic structure of sources used in

restoration. If seed sources of dominant species are not well

adapted to the local environment, then increased estab-

lishment success and growth of subordinate species should

promote diversity (Mijnsbrugee et al. 2010). This study

suggests that during initial community assembly, the com-

position of the subordinate species seed mixture is more

important than the seed source of the dominant species,

particularly in diverse restored communities.

Acknowledgements

Funding for this research was provided by the National Sci-

ence Foundation (DEB 0516429). We thank Ryan E.

Campbell, Allison M. Lambert, Elizabeth M. Bach and

many undergraduate student workers for their assistance

with data collection, and the Southern Illinois University

Carbondale Agronomy Research Center for site prepara-

tion andmaintenance.

References

Abrams, P. 1983. The theory of limiting similarity. Annual Review

of Ecology and Systematics 14: 359–376.

Alderson, J. & Sharp, C.W. 1995.Grass varieties in the United States.

United States Department of Agriculture, CRC Press, Inc.,

Boca Raton, FL, US.

Baer, S.G., Blair, J.M., Collins, S.L. & Knapp, A.K. 2004. Plant

community responses to resource availability and heteroge-

neity during restoration. Oecologia 139: 617–639.

Baer, S.G., Collins, S.L., Blair, J.M., Knapp, A.K. & Fiedler, A.K.

2005. Soil heterogeneity effects on tallgrass prairie commu-

nity heterogeneity: an application of ecological theory to res-

toration ecology. Restoration Ecology 13: 413–424.

Bailey, J.K., Schweitzer, J.A., Asbeda, F., Koricheva, J., LeRoy,

C.J., Madritch, M.D., Rehill, B.J., Bangert, R.K., Fischer, D.

G., Allan, G.J. & Whitham, T.G. 2009. From genes to ecosys-

tems: a synthesis of the effects of plant genetic factors across

levels of organization. Philosophical Transactions of the Royal

Society of London, Series B: Biological Sciences 364: 1607–1616.

Bangert, R.K. & Whitham, T.G. 2007. Genetic assembly rules

and community phenotypes. Evolutionary Ecology 21: 549–

560.

Bangert, R.K., Lonsdorf, E.V., Wimp, G.M., Shuster, S.M.,

Fischer, D., Schweitzer, J.A., Allan, G.J., Bailey, J.K. &

Whitham, T.G. 2008. Genetic structure of a foundation spe-

cies: scaling community phenotypes from the individual to

the region.Heredity 100: 121–131.

Bartha, S., Meiners, S.J., Pickett, S.T.A. & Cadenasso, M.L. 2003.

Plant colonization windows in a mesic old field succession.

Applied Vegetation Science 6: 205–212.

Camill, P., McKone, M.J., Sturges, S.T., Severud, W.J., Ellis, E.

C., Limmer, J., Martin, C.B., Navratil, R.T., Purdie, A.J.,

Journal of Vegetation Science10 Doi: 10.1111/j.1654-1103.2012.01475.x© 2012 International Association for Vegetation Science

Grassland assembly D.J. Gibson et al.

Sandel, B.S., Talukder, S. & Trout, A. 2004. Community-

and ecosystem-level changes in a species-rich tallgrass prairie

restoration. Ecological Applications 14: 1680–1694.

Chase, J., Amarasekare, P., Cottenie, K., Gonzalez, A., Holt,

R.D., Holyoak, M., Hoopes, M.F., Leibold, M.A., Loreau, M.,

Mouguet, N., Shurin, J.B. & Tilman, D. 2005. Competing

theories for competitive metacommunities. In: Holyoak, M.,

Leibold, M.A. & Holt, R.D. (eds.) Metacommunities: spatial

dynamics and ecological communities. pp. 335–354. University

of Chicago Press, Chicago, IL, US.

Collins, S.L., Knapp, A.K., Hartnett, D.C. & Briggs, J.M. 1998.

The dynamic tallgrass prairie: synthesis and research oppor-

tunities. In: Knapp, A.K., Briggs, J.M., Hartnett, D.C. & Col-

lins, S.L. (eds.) Grassland dynamics: long-term ecological research

in tallgrass prairie. pp. 301–315. Oxford University Press,

New York, NY, US.

Cottam, G. & Wilson, H.C. 1966. Community dynamics on an

artificial prairie. Ecology 47: 88–96.

Dawkins, R. 1982. The extended phenotype. Oxford University

Press, Oxford, UK.

Diboll, N. 2005. Designing seed mixes. In: Packard, S. & Mutel,

C.F. (eds.) The tallgrass restoration handbook. pp. 135–150.

Island Press, Washington, DC, US.

Dodd, J.D. 1983. Grassland associations in North America. In:

Gould, F.W. & Shaw, M.R. (eds.) Grass systematics, 2nd ed.

pp. 343–357. Texas A&MPress, College Station, TX, US.

Ejrnæs, R., Bruun, H.H. & Graae, B.J. 2006. Community assem-

bly in experimental grasslands: suitable environment or

timely arrival? Ecology 87: 1225–1233.

Ellis, E.C. & Ramankutty, N. 2008. Putting people in the map:

anthropogenic biomes of the world. Frontiers in Ecology and

the Environment 6: 439–447.

Faith, D.P. & Norris, R.H. 1989. Correlation of environmental

variables with patterns of distribution and abundance of

common and rare freshwater macroinvertebrates. Biological

Conservation 50: 77–98.

Fridley, J.D. 2001. The influence of species diversity on ecosys-

tem productivity: how, where, andwhy?Oikos 93: 514–526.

Fridley, J.D. & Grime, J.P. 2010. Community and ecosystem

effects of intraspecific genetic diversity in grassland micro-

cosms of varying species diversity. Ecology 91: 2272–2283.

Gibson, D.J. 2009. Grasses and grassland ecology. Oxford University

Press, Oxford, UK.

Gibson, D.J., Urban, J.W. & Baer, S.G. 2011. Mowing and fertil-

izer effects on seedling establishment in a successional old-

field. Journal of Plant Ecology 4: 157–168.

Gibson, D.J., Allstadt, A., Baer, S.G. & Geisler, M. 2012. Incorpo-

rating the genetic diversity of foundation species into com-

munity assembly and diversity. Oikos 121: 497–507.

Giller, P.S. & O’Donovan, G. 2007. Biodiversity and ecosystem

function: do species matter? Biology and Environment: Proceed-

ings of the Royal Irish Academy 102: 129–139.

Grime, J.P. 1998. Benefits of plant diversity to ecosystems:

immediate, filter and founder effects. Journal of Ecology 86:

902–910.

Gustafson, D.J., Gibson, D.J. & Nickrent, D.L. 2004a. Competi-

tive relationships of Andropogon gerardii (Big Bluestem) from

remnant and restored native populations and select culti-

vated varieties. Functional Ecology 18: 451–457.

Gustafson, D.J., Gibson, D.J. & Nickrent, D.L. 2004b. Conserva-

tion genetics of two co-dominant grass species in an

endangered grassland ecosystem. Journal of Applied Ecology

41: 389–397.

Gustafson, D.J., Gibson, D.J. & Nickrent, D.L. 2005. Empirical

support for the use of local seed sources in prairie restoration.

Native Plants Journal 6: 25–28.

Harpole,W.S. & Tilman, D. 2006. Non-neutral patterns of species

abundance in grassland communities. Ecology Letters 9:

15–23.

Helsen, K., Hermy, M. & Honnay, O. 2012. Trait but not species

convergence during plant community assembly in restored

semi-natural grasslands. Oikos in press; doi: 10.1111/j/1600-

0706.2012.20499.x.

Herman, R.J. 1979. Soil survey of Jackson County, Illinois. National

Cooperative Soil Survey,Washington, DC, US.

Hobbs, R.J. & Norton, D.A. 2004. Ecological filters, thresholds,

and gradients in resistance to ecosystem reassembly. In:

Temperton, V.M., Hobbs, R.J., Nuttle, T. & Halle, S. (eds.)

Assembly rules and restoration ecology: bridging the gap between

theory and practice. pp. 72–95. Island Press, Washington, DC,

US.

Houseman, G.R. & Gross, K.L. 2011. Linking grassland plant

diversity to species pools, sorting, and plant traits. Journal of

Ecology 99: 464–472.

Hufford, K.M. & Mazer, S.J. 2003. Plant ecotypes: genetic differ-

entiation in an age of ecological restoration. Trends in Ecology

and Evolution 18: 147–155.

Isbell, F., Calcagno, V., Hector, A., Connolly, J., Harpole, W.S.,

Reich, P.B., Scherer-Lorenzen, M., Schmid, B., Tilman, D.,

van Ruijven, J., Weigelt, A., Wilsey, B.J., Zavaleta, E.S. &

Loreau, M. 2011. High plant diversity is needed to maintain

ecosystem services.Nature 477: 199–202.

Kantvilas, G. & Minchin, P.R. 1989. An analysis of epiphytic

lichen communities in Tasmanian cool temperate rain forest.

Plant Ecology 84: 91–112.

Kindscher, K. & Tieszen, L.L. 1998. Floristic and soil organicmat-

ter changes after five and thirty-five years of native tallgrass

prairie restoration. Restoration Ecology 6: 181–196.

Klopf, R.P. & Baer, S.G. 2011. Root dynamics of cultivar and

non-cultivar population sources of three dominant grasses

during intial establishment of tallgrass prairie. Restoration

Ecology 19: 112–117.

Lambert, A.M., Baer, S.G. & Gibson, D.J. 2011. Intraspecific vari-

ation in ecophysiology of three dominant tallgrass prairie

grasses used in restoration: cultivar versus non-cultivar pop-

ulation sources. Restoration Ecology 19: 43–52.

Larson, D.L., Bright, J.B., Drobney, P., Larson, J.L., Palaia, N.,

Rabie, P.A., Vacek, S. & Wells, D. 2011. Effects of planting

method and seed mix richness on the early stages of tallgrass

prairie restoration. Biological Conservation 144: 3127–3139.

Journal of Vegetation ScienceDoi: 10.1111/j.1654-1103.2012.01475.x© 2012 International Association for Vegetation Science 11

D.J. Gibson et al. Grassland assembly

Lawton, J.H. 1994. What do species do in ecosystem? Oikos 71:

367–374.

Lesica, P. & Allendorf, F.W. 1999. Ecological genetics and the

restoration of plant communities: mix or match? Restoration

Ecology 7: 42–50.

Littell, R.C., Milliken, G.A., Stroup, W.W., Wolfinger, R.D. &

Schabenberger, O. 2006. SAS® for mixed models, 2nd ed. SAS

Institute, Cary, NC, US.

McCain, K.N.S., Baer, S.G., Blair, J.M. & Wilson, G.W.T. 2010.

Dominant grasses suppress local diversity in restored tallgrass

prairie. Restoration Ecology 18: 40–49.

McKay, J.K., Christian, C.E., Harrison, S. & Rice, K.J. 2005.

“How local is local?” – A review of practical and conceptual

issues in the genetics of restoration. Restoration Ecology 13:

342–440.

Mijnsbrugee, K.V., Bischoff, A. & Smith, B. 2010. A question of

origin: where and how to collect seed for ecological restora-

tion. Basic and Applied Ecology 11: 300–311.

Minchin, P.R. 1987. An evaluation of the relative robustness

of techniques for ecological ordination. Plant Ecology 69:

89–107.

Mohlenbrock, R.H. 2002. Vascular flora of Illinois. Southern Illi-

nois University Press, Carbondale & Edwardsville, IL, US.

Montalvo, A.M., Williams, S.L., Rice, K.J., Buchmann, S.L.,

Cory, C., Handel, S.N., Nabhan, G.P., Primack, R. & Robich-

aux, R.H. 1997. Restoration Biology: a Population Biology

Perspective. Restoration Ecology 5: 277–290.

Packard, S. & Mutel, C.F. (eds.) 2005. The tallgrass restoration

handbook, 2nd ed. Island Press, Washington, DC, US.

Pavoine, S., Vela, E., Gachet, S., de Belair, G. & Bonsall, M.B.

2011. Linking patterns in phylogeny, traits, abiotic variables

and space: a novel approach to linking environmental filter-

ing and plant community assembly. Journal of Ecology 99:

165–175.

Polley, H.W.,Wilsey, B.J. & Derner, J.D. 2007. Dominant species

constrain effects of species diversity on temporal variability

in biomass production of tallgrass prairie. Oikos 116: 2044–

2052.

Roscher, C., Temperton, V.M., Buchmann, N. & Schulze, E.-D.

2009. Community assembly and biomass production in

regularly and never weeded experimental grasslands. Acta

Oecologica 35: 206–217.

Sasaki, T. & Lauenroth, W. 2011. Dominant species, rather than

diversity, regulates temporal stability of plant communities.

Oecologia 166: 761–768.

Vellend, M. 2006. The consequences of genetic diversity in com-

petitive communities. Ecology 87: 304–311.

Vellend, M., Drummond, E.B.M. & Tomimatsu, H. 2010. Effects

of genotype identity and diversity on the invasiveness and

invasibility of plant populations. Oecologia 162: 371–381.

Weaver, J.E. 1954. North American prairie. Johnsen Co., Lincoln,

NE, US.

Weiher, E., Freund, D., Bunton, T., Stefanski, A., Lee, T. &

Bentivenga, S. 2011. Advances, challenges and a developing

synthesis of ecological community assembly theory. Philo-

sophical Transactions of the Royal Society of London, Series B:

Biological Sciences 366: 2403–2413.

Weinberg, J.R. 1984. Interactions between functional groups in

soft-substrata: do species differences matter? Journal of Exper-

imental Marine Biology and Ecology 80: 11–28.

White, R., Murray, S. & Rohweder, M. 2000. Pilot analysis of glo-

bal ecosystems: grassland ecosystems technical report. World

Resources Institute, Washington, DC, US.

Whitham, T.G., Young, W.P., Martinsen, G.D., Gehring, C.A.,

Schwietzer, J.A., Shuster, S.M., Wimp, G.M., Fischer, D.G.,

Bailey, J.K., Lindroth, R.L., Woolbright, S. & Kuske, C.R.

2003. Community and ecosystem genetics: a consequence of

the extended phenotype. Ecology 84: 559–573.

Whitlock, R., Grime, J.P., Booth, R.E. & Burke, T. 2007. The role

of genotypic diversity in determining grassland community

structure under constant environmental conditions. Journal

of Ecology 95: 895–907.

Wilsey, B.J. 2010. Productivity and subordinate species response

to dominant grass species and seed source during restoration.

Restoration Ecology 18: 628–637.

Zedler, J. 2005. Ecological restoration: guidance from theory.

San Francisco Estuary andWatershed Science 3: 1–31.

Zobel, M. 1997. The relative role of species pools in determining

plant species richness: an alternative explanation of species

coexistence? Trends in Ecology and Evolution 12: 266–269.

Supporting Information

Additional supporting information may be found in the

online version of this article:

Appendix S1. The split-plot design used to test

whether intraspecific functional variation of dominant

species affects community composition and structure over

the first 4 yr of community assembly. Six whole plots were

randomly assigned to cultivar (n = 3) and non-cultivar

(n = 3) source of three dominant grasses within each of

two adjacent agricultural fields (blocks). Three unique

designed species pools (A–C) of subordinate prairie species

were randomly assigned to 5 9 5 m subplots within each

whole-plot. This figure also appears in Baer et al. (in

review).

Please note: Wiley-Blackwell are not responsible for

the content or functionality of any supporting materials

supplied by the authors. Any queries (other than missing

material) should be directed to the corresponding author

for the article.

Journal of Vegetation Science12 Doi: 10.1111/j.1654-1103.2012.01475.x© 2012 International Association for Vegetation Science

Grassland assembly D.J. Gibson et al.