BIODIVERSITY RESEARCH: Experimental introduction of the alien plant Hieracium lepidulum reveals no significant impact on montane plant communities in New Zealand

BIODIVERSITYRESEARCH

Experimental introduction of the alienplant Hieracium lepidulum reveals nosignificant impact on montane plantcommunities in New Zealand

Ross Meffin*, Alice L. Miller�, Philip E. Hulme and Richard P. Duncan

INTRODUCTION

It is clear that some alien plants can have significant impacts

on the plant communities they invade; there are well-

documented examples of plant invasions leading to declines

in abundance and the local extinction of native species (Pysek

& Pysek, 1995; Vila et al., 2006; Gaertner et al., 2009; Hejda

et al., 2009) and to substantial changes in ecosystem structure

and function (Vitousek & Walker, 1989; D’Antonio &

Vitousek, 1992; Ogle et al., 2003). While these problems

warrant considerable concern, there is some evidence that alien

plant species may have significant impacts in only a minority

of cases (Levine et al., 2003; Sax & Gaines, 2008). Many alien

plants appear to invade and coexist in native communities

without significant impact (Stohlgren et al., 2006). Such

contrasting results, however, may also reflect the scale

Bio-Protection Research Centre, PO Box 84,

Lincoln University, Lincoln 7647, New

Zealand

*Correspondence: Ross Meffin, Bio-Protection

Research Centre, PO Box 84, Lincoln

University, Lincoln 7647, New Zealand.

E-mail: [email protected]

�Present address: National Park Service, Joshua

Tree National Park, 74485 National Park Drive,

Twentynine Palms, California 92277, USA.

ABSTRACT

Aim There is debate over whether alien plants necessarily alter the communities

they invade or can coexist with native species without discernable impacts. We

followed the fate of montane plant communities in response to the experimental

sowing of the alien weed Hieracium lepidulum, looking for changes in plant

community composition and structure over 6 years.

Location Craigieburn Range, New Zealand.

Methods We used a replicated randomised block design, with 30 · 30 cm plots

(n = 756) subdivided into 5 · 5 cm cells to examine and compare the effects of

H. lepidulum at 0.09 m2 (plot) and 0.0025 m2 (cell) scales. Plots were sown with

between 0 and 15,625 H. lepidulum seeds in 2003, forming gradients of invader

density and cover. Measurements comprised community richness, evenness and

diversity along with H. lepidulum density and cover at both scales. The

relationships between the invader and local community attributes were

modelled using hierarchical mixed-effect models.

Results Plant communities differed in the extent to which they became invaded,

with H. lepidulum cover in the plots ranging from 0% to 52%, with a mean of

only 1.89%. Plot species richness increased from 2003 to 2009, with a component

of this increase (+0.002 species per year) associated with increasing H. lepidulum

density. Other relationships between the plant community and H. lepidulum were

generally non-significant.

Main conclusions In these montane plant communities, it appears H. lepidulum

coexists with the native community with no measurable negative effects after

6 years on species richness, evenness or diversity, even where density and cover of

the invader are highest. We suggest H. lepidulum has persisted preferentially at

those sites with abiotic conditions sufficient to support a species-rich assemblage.

Keywords

Biodiversity, biological invasions, facilitation, hierarchical mixed model, invasive

species.

Diversity and Distributions, (Diversity Distrib.) (2010) 16, 804–815

DOI:10.1111/j.1472-4642.2010.00684.x804 www.blackwellpublishing.com/ddi ª 2010 Blackwell Publishing Ltd

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dependence of alien plant impacts (Gaertner et al., 2009).

Negative interactions between alien and native species are most

likely to occur at fine spatial scales (< 1 m2), while at coarser

scales native plant diversity and alien abundance may covary

with favourable environmental factors (e.g. resources, distur-

bance) leading to positive associations between the two (Levine

& D’Antonio, 1999). Indeed, the frequency at which alien

plants impact native communities, and the magnitude of those

impacts, is unclear because of a lack of studies representative of

the range of alien plant life-forms, invaded native communities

and spatial scales examined (Mack, 1996; Parker et al., 1999;

Byers et al., 2002; Mills et al., 2009).

Furthermore, several difficulties arise in trying to quantify

invader impacts. First, most studies to date have been

correlative, comparing already invaded communities to neigh-

bouring uninvaded reference communities. In such studies, it

is difficult to disentangle cause and effect, since species-poor

communities may be more vulnerable to invasion (Levine &

D’Antonio, 1999), or both native and alien species richness

may be driven by similar environmental factors, potentially

resulting in already species-rich communities becoming even

richer (Stohlgren et al., 2003; Hulme, 2006). Second, correl-

ative studies have tended to focus on highly invaded commu-

nities where, because of the dominance of the alien, impacts

are most likely to be found. But such dominance may not be

representative of the average invader density across habitats,

and such studies may tend to overstate the effects of alien

plants (Truscott et al., 2008). Consequently, few studies have

examined impacts across a sufficiently broad gradient of alien

abundance and native species richness or diversity, meaning

that situations where alien species increase rather than decrease

species richness may be missed (Vila et al., 2006).

Alternatively, the impact of plant invaders can be assessed

experimentally via species removals (e.g. Ogle et al., 2003;

Hulme & Bremner, 2006; Stinson et al., 2007; Truscott et al.,

2008). While removals can overcome the confounding effects

of environmental variation, there are two further issues. First,

removing the invader may disturb the native community, and

the effects of the invader’s absence may then be confounded

with those of the disturbance. Second, removal studies measure

recovery of the native community following release from the

invader, which may differ from the initial impact of the

invader on that community (Diaz et al., 2003).

A further approach is to intentionally introduce the invader

into native communities and to compare the response of

experimentally invaded plots with uninvaded control plots.

This technique is infrequently used and can pose ethical

problems relating to deliberately introducing alien species into

uninvaded communities (Truscott et al., 2008; Vila et al.,

2008), but has two advantages over other methods. First, it is

experimental, so confounding factors can be dealt with

through stratification, replication and randomisation; and

second, it allows direct measurement of invader impacts, as

opposed to the effects of invader removal.

Our aim is to examine the long-term impacts of an invasive

alien plant on the richness, diversity and evenness of native plant

communities using intentional introductions to establish an

experimental gradient of invader density and abundance. We use

the alien plant Hieracium lepidulum (botanical nomenclature

throughout this paper follows Webb et al., 1988) as a model,

since it is widely regarded as posing a significant threat to the

diversity of montane and subalpine native forest and grassland

communities in the South Island, New Zealand (Wiser et al.,

1998; Rose & Frampton, 1999; Wiser & Allen, 2000; Radford

et al., 2007). By establishing an experimental gradient of alien

plant densities and assessing impacts over 6 years and at two

spatial scales, we undertake one of the most thorough assess-

ments of alien plant impacts on vegetation to date.

We hypothesise that H. lepidulum primarily impacts the

native plant community through competition for resources

(Rose & Frampton, 1999; Radford et al., 2007), and conse-

quently that its effects should be most apparent at the fine

spatial scale of plant neighbourhoods. However, even if impacts

result from other putative mechanisms, they should still be

apparent at these scales. Such impacts would lead to population

declines and local extinctions, and thus changes in the local

species richness, evenness and diversity of native plant commu-

nities. We therefore examined the effect of H. lepidulum

invasion on native community composition and structure, by

following changes in the richness, evenness and diversity of

native plants at two spatial scales over a period of 6 years.

METHODS

Study site

The study was conducted in the Craigieburn Stream catch-

ment, Craigieburn Forest Park, Canterbury, New Zealand

(43�10¢ S, 172�45¢ E). The terrain is mountainous, with

elevations ranging from 800 to 2000 m. The dominant

vegetation is mountain beech forest (Nothofagus solandri var.

cliffortiodes) to an elevation of approximately 1400 m, above

which it gives way to subalpine scrub, tussock and alpine

herbfields. The mean annual temperature is 8.2 �C at 914 m,

and mean annual precipitation is 1533 mm (Miller, 2006).

Study species

Hieracium lepidulum is a broad-leaved, taprooted, rosette-

forming perennial herb growing to a height of approximately

30–40 cm and is apomictic, relying entirely on wind-dispersed

seed for establishment and spread. Flowering and seed

dispersal occur from November to May, after which the leaves

die back and the plant overwinters as a rhizome (Chapman

et al., 2004; Miller, 2006). It was introduced from Europe to

New Zealand, where it was first recorded in 1941 (Miller,

2006). It has become invasive in the South Island and has

steadily increased in distribution and abundance over the last

60 years (Rose et al., 1995; Duncan et al., 1997, 2001; Wiser

et al., 1998; Mark et al., 1999; Wiser & Allen, 2000). This

invasion success is thought to be because of competitive

advantages which allow H. lepidulum to displace native species

Multi-scale experimental test of alien plant impacts

Diversity and Distributions, 16, 804–815, ª 2010 Blackwell Publishing Ltd 805

and become dominant over time, sometimes forming dense

mats to the exclusion of other species (Rose & Frampton, 1999;

Radford et al., 2007). Along with other members of the genus

invasive in New Zealand, notably Hieracium pilosella and

Hieracium praeltum, it is thought to compete strongly with

native species (Fan & Harris, 1996; Moen & Meurk, 2001;

Weigelt et al., 2002; Winkler & Stocklin, 2002; Lamoureaux

et al., 2003), leading to declines in the species diversity of plant

communities it has invaded (Scott et al., 1990; Wiser & Allen,

2000; Espie, 2001). Unlike those congeners, H. lepidulum is

able to establish in shade under forest, shrub or tussock

canopies and is able to invade higher rainfall and higher

elevation areas (Wiser et al., 1998; Rose & Frampton, 1999).

Consequently, this species is of particular conservation concern

because of its widespread distribution and potential to impact

the extensive and largely intact native communities in montane

and subalpine regions throughout the South Island (Wiser &

Allen, 2000; Miller, 2006; Radford et al., 2007).

Experimental set-up

We experimentally added H. lepidulum seeds at different

densities to plots located in six habitat types: forest creek,

intact forest, forest canopy gap (formed by treefall), subalpine

creek, subalpine tussock grassland and subalpine scrub. We

used a randomized block design, with each habitat having six

replicate blocks. Subalpine creek and forest creek blocks were

adjacent to the main watercourse draining the catchment,

Craigieburn Stream. Forest creek and subalpine creek blocks

were located at a random distance along the stream below tree

line and above tree line, respectively. For intact forest and

tussock grassland habitat, blocks were located at a random

distance along Craigieburn Stream and a random distance

perpendicular to the watercourse into the forest or tussock

habitat. The minimum distance from the watercourse was no

less than 20 m (to ensure the habitat was not strongly

influenced by the creek) and no more than 200 m (for logistical

reasons). Canopy gap blocks were located at the nearest treefall

gap greater than 25 m2, a random distance along the creek and

a random distance into the forest, as above. Scrub habitat

blocks were located by mapping the scrub patches above tree

line in the study catchment using aerial photographs and then

randomly selecting six representative patches.

Each block comprised three replicate plots of each of seven

H. lepidulum seed density treatments (21 plots per block):

control with no seed addition and no initial wetting, proce-

dural control with no seed addition but with wetting and seed

addition at rates of 25, 125, 625, 3125 and 15625 seeds per

30 · 30 cm plot with wetting (or 278, 1389, 6944, 34722 and

173611 seeds m)2). Lower sowing treatment levels represent

seed densities in the range expected in areas invaded by

H. lepidulum, based on its observed densities and rates of seed

production (Miller, 2006). Plots with higher densities con-

tained more seeds than pilot studies indicated were likely to

occur naturally in the study site, but allowed us to quantify the

effects of H. lepidulum across a broad gradient of seed densities

and gave us the best chance of establishing H. lepidulum at the

high densities and covers reported from other sites. Wetting

was used to prevent the seeds from being blown away during

seed addition. The plots in each block were randomly

positioned in a 16-m2 area such that there was at least a 30-

cm buffer between each plot, and seed density treatments were

randomly assigned to each plot following plot placement. Plots

were permanently marked with two labelled corner pegs.

Naturally occurring H. lepidulum plants were uncommon in

the study blocks at the start of the experiment, but any plants

occurring in or within 20 m of the edge of the block were

removed at the start of the study.

Data collection

Plot abiotic characteristics

Data on environmental conditions in each block were collected

prior to seed addition: soil attributes were measured at the

block level, and light measurements were taken for each plot.

Soil depth to bedrock was measured in the centre of each block

using a metal probe. Soil samples were collected from the

corners of each block using a 10-cm corer and then bulked.

Five grams wet weight of soil from each block was dried and

reweighed to calculate gravimetric soil moisture content. The

remaining soil was air-dried, sieved and analysed for pH,

Olsen-soluble phosphorus, carbon, nitrogen and potassium.

The altitude and aspect of each block were recorded. For each

plot, available light was measured as the percentage of

photosynthetic photon flux density (%PPFD) under overcast

skies. To do this, 15 instantaneous light measurements (Qi)

were taken 1 cm above the centre of each plot using point

quantum sensors (LI-190SA, LICOR, Lincoln, NE, USA), while

readings under open light conditions (Qo) were recorded

simultaneously. %PPFD was then calculated as Qi/Qo · 100.

Quantifying Hieracium lepidulum abundance and native

community structure

Hieracium lepidulum seeds were added to the plots in March

2003 (autumn), and the number of H. lepidulum plants in each

plot was recorded in the following summer (December 2003)

and then recorded annually in February or March from 2004 to

2009. In plots where high densities of H. lepidulum plants

made a complete census impractical, a 1 cm · 1 cm grid of

cells was overlaid on the plot, and the count from 20 randomly

selected cells were multiplied by 45 to estimate the total plot

count.

Prior to seed addition, the plant community in each plot was

characterized by recording the presence of all vascular plant

species. Changes in composition were tracked by remeasuring

each plot in 2006, 2007 and 2009, again recording all vascular

plant species present.

In 2009, additional measurements were made to characterize

invader impacts. To quantify the impact of H. lepidulum at

small spatial scales more accurately, where we hypothesise

R. Meffin et al.

806 Diversity and Distributions, 16, 804–815, ª 2010 Blackwell Publishing Ltd

competitive effects to be most apparent, each plot was divided

into a regular grid of 36 5 · 5 cm cells. In each cell, we

recorded the number and cover of H. lepidulum plants, and the

presence and cover of all other vascular plant species. Cell

values were averaged to obtain cover estimates for each species

at the plot level.

Ten plots (1%) could not be relocated during the study;

most likely due to disturbances such as treefalls. In addition,

four plots contained no native species throughout the study

and were thus uninformative when considering H. lepidulum

impacts on native communities. We excluded these 14 plots

from the analyses.

Analysis

We conducted our analysis in three parts: (1) a longitudinal

analysis of changes in native species richness in the plots over

6 years as a function of H. lepidulum density, and analyses of

variation in native community richness, diversity and evenness

in relation to H. lepidulum density and cover in 2009 at both

the (2) plot and (3) cell scales.

Longitudinal analysis

We first examined how plot species richness changed through

time by fitting linear trends to the data using a hierarchical

regression model (Gelman & Hill, 2006). Time (years since

addition of H. lepidulum seed), habitat and their interaction

were included as fixed effects, with the interaction term

allowing for a different trend in species richness through time

in each habitat. There were repeat measurements per plot, and

the plots were nested within blocks, so we included plot nested

within block as random effects to account for this non-

independence.

If invasion by H. lepidulum caused local extinctions through

competitive exclusion, we would expect species richness to

have declined more in heavily invaded plots relative to

uninvaded ones. To test this, we fitted a second regression

model with the same random effects, but with 2009 density of

H. lepidulum, time, the interaction between density and time,

and habitat as fixed effects. The density by time interaction

tests whether more or less invaded plots (as measured by

H. lepidulum density in 2009) show different trends in native

species richness through time. In addition, we included soil

characteristics (carbon, nitrogen, phosphorus and potassium

concentrations, pH, soil moisture and soil depth) and physical

attributes (plot altitude, aspect and available light) as addi-

tional fixed effects. These could influence both native species

richness and H. lepidulum density so were included as

covariates to account for this.

This model was compared with another candidate to allow

for the possibility that H. lepidulum could have different

impacts in different habitats. To do this, we fitted a second

model with an additional 2009 H. lepidulum density by habitat

interaction. We selected the better model using Akaike

Information Criterion (AIC), a measure of the explanatory

power of a model which penalises models requiring more

parameters; a lower AIC value indicates the ‘better’ model

(Akaike, 1974). As the alternative model did not show

significant reduction in AIC, it was rejected in favour of the

original model.

Coefficient estimates for the final model were tested for

significance using anova in the package languageR (Baayen,

2008). When conducting this test on a mixed-effect model,

there is uncertainty in calculating the number of denominator

degrees of freedom (Bates, 2006; Baayen et al., 2008). To

overcome this, these were set to the number of observations

minus the number of fixed-effects coefficients. This produces

anti-conservative P values for small data sets, but here

(n = 742) provides a good indication of significance (Baayen

et al., 2008).

Plot scale

Using the data from 2009, we related the density and cover of

H. lepidulum in the experimental plots to three measures of

native community structure: plot species richness, Shannon

diversity and Shannon evenness (Magurran, 2003; Spellerberg

& Fedor, 2003). For each response variable (2009 plot richness,

evenness and diversity), we fitted two hierarchical regression

models: one with H. lepidulum density and the other with

H. lepidulum cover as fixed effects, and including block as a

random effect to account for the nesting of plots within blocks.

The influence of habitat on these relationships was then tested

by fitting a second set of regression models, this time including

habitat and its interactions with H. lepidulum density and

cover as additional fixed effects. To test coefficient estimates

for significance, we used Markov Chain Monte Carlo sampling

with 10,000 iterations to generate posterior distributions using

the package languageR (Baayen, 2008). Multiple tests of

significance inflate the risk of type I errors, to account for

this we applied a Holm–BonFerroni correction.

Finally, we may expect competitive interactions to be

stronger within morphologically and functionally similar sets

of species, as a result of competition for the same resources

(Ortega & Pearson, 2005). To test this, we related H. lepidulum

density and cover in 2009 to the richness, diversity and

evenness of species in each of 10 functional groups (basal herb,

clumped herb, creeping herb, shrub, erect leafy, grass, mat/

cushion, prostrate shrub, tussock and ‘other’) based on

morphological and life history characteristics (Cornelissen

et al., 2003), and informed by field observations. Of these, the

basal herb, creeping herb and erect leafy groups are expected to

have stronger competitive interactions with the morphologi-

cally and functionally similar H. lepidulum (Ortega & Pearson,

2005).

Cell scale

Because the outcome of competitive interactions may be more

apparent at small spatial scales, we repeated the plot-scale

analyses at the cell scale. We fitted the same models but



included plots nested within blocks as random effects to

account for non-independence because of the nesting of cells

within plots within blocks.

All hierarchical models were fitted using the lme4 package

(Bates & Maechler, 2009) in R version 2.9.2 (R_Develop-

ment_Core_Team, 2009).

RESULTS

Hieracium lepidulum introduction

Average H. lepidulum densities peaked in December 2003, in

the summer immediately following seed addition, when plots

contained from 0 to 4435 individuals. Average densities then

declined in all seed addition treatments until around 2007,

where densities appeared to stabilize (Fig. 1). The seed

addition treatments created a gradient of H. lepidulum density

across the plots, which was maintained for the duration of the

study (Fig. 1). Density ranged from 0 to 383 individuals per

plot, with an overall mean of 12.65 ± 1.5 in all treatment plots

and 0.24 ± 0.19 in all control plots (Table 1, Fig. 2, showing

combined treatment and control plots). Percent cover of the

invader ranged from 0 to 52% with an overall mean of

1.89 ± 0.017% in treatment plots and 0.09 ± 0.04% in control

plots (Table 1, Fig. 2, showing combined control and treat-

ment plots).

Longitudinal analysis

Fitting a hierarchical regression model with time, habitat and

their interaction as fixed effects, plot species richness increased

with time in all habitats, but particularly in those above tree

Year

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2003 2004 2005 2006 2007 2008 2009

Seeds sown per plot:15,625312562512525

Figure 1 Mean density (number of individuals per 30 · 30 cm

plot and their 95% confidence intervals, 108 plots per treatment)

of Hieracium lepidulum plants through time in plots sown at 2003

at five-treatment seed sowing densities (controls not shown). Tab

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R. Meffin et al.


line (Fig. 3). Two models incorporating the influence of

H. lepidulum were considered; the first, with time, 2009

H. lepidulum density, their interaction, habitat and covariates

(AIC = 11,316) was selected over the second, which had an

additional interaction term between H. lepidulum density and

habitat (AIC = 11,320). The selected model indicated an

overall mean increase of 0.2 species per plot per year over

the duration of the study (Table 2). There was a highly

significant H. lepidulum density by time interaction, but this

was positive, implying that plots with more H. lepidulum

showed a greater increase in species richness through time, the

opposite to what we would expect if invasion by H. lepidulum

resulted in competitive exclusion. Plots above tree line had

significantly more species than those below (Table 2, see also

Fig. 3). Habitats varied considerably in their mean abiotic

characteristics (see Appendix S1 in supporting Information).

Having accounted for differences among habitats, lower

elevation plots and those with higher available light had more

species on average, but other covariates were non-significant.

Plot scale

Overall plot species richness was positively related to H. lepi-

dulum density in 2009 (Fig. 4). As density and cover of the

invader were reasonably well correlated, a similar positive

relationship was found between plot species richness and

H. lepidulum cover (not shown). While these relationships

were not significant, they are opposite in direction to what we

would expect if H. lepidulum were outcompeting native species

for limited resources. When habitat and its interaction with

2009 H. lepidulum density and cover were added to these

models as additional fixed effects, allowing for different trends

in each habitat, relationships with H. lepidulum density and

cover were mostly non-significant and weakly positive (see

Appendix S2 in Supporting Information); similar to those

found at the cell scale (see below). The only significant

relationship was in the forest creek habitat, where mean plant

species diversity was greater in plots with higher H. lepidulum

cover.

Cell scale

Overall, the relationship between species richness and H. lepi-

dulum density was positive at the cell scale (Fig. 4), and the

relationship with cover of the invader was significantly positive

(Fig. 5). Mean species evenness was also significantly greater

overall in plots with higher H. lepidulum densities. In models

allowing for impacts to vary by habitat, most relationships

0 100 200 300 400

010

100

2009 Hieracium density

Fre

quen

cy

0 10 20 30 40 50 60

010

100

2009 % Hieracium cover

Fre

quen

cy

Figure 2 Frequency histograms of Hieracium lepidulum density

(number if individuals per 30 · 30 cm plot, top) and cover

(percent cover per 30 · 30 cm plot, bottom) in 2009, showing

combined treatment and control plots.

02

46

810

1214

Year

Plo

t spe

cies

ric

hnes

s

Subalpine creekTussockOverallScrub

2003 2004 2005 2006 2007 2008 2009

2003 2004 2005 2006 2007 2008 2009

02

46

810

1214

Year

Plo

t spe

cies

ric

hnes

s

OverallForest creekForest gapForest

Figure 3 Change in mean species richness per 30 · 30 cm plot

(circles with 95% confidence intervals, n = 742), overall and in

habitats above (top) and below tree line (bottom). Lines show the

best-fit from a hierarchical regression model with time (years since

2003), habitat and their interaction as fixed effects, and plots

nested within blocks as random effects.



were non-significant, although evenness was significantly

greater in plots with higher H. lepidulum cover in the tussock

habitat.

Functional groups

Relationships between community attributes and H. lepidulum

density and cover were mostly weak and non-significant when

the data were partitioned by functional group at the plot scale

(see Appendix S1). However, species belonging to the clumped

herb group had a significantly less even distribution of

abundances in plots where H. lepidulum cover was lower.

That is, one or more clumped herb species were dispropor-

tionately more abundant in plots with lower H. lepidulum

cover.

At the cell scale, functional group diversity and evenness also

showed non-significant relationships with H. lepidulum

(Fig. 6). By contrast, functional group richness showed multi-

ple small, but significant, interactions with H. lepidulum

density and cover at the cell scale. Competitive effects are

hypothesised to be strongest at the cell scale, so this contrast

may be expected, although cover of the invader was signifi-

cantly positively associated with basal herb, creeping herb and

grass functional group richness. There were significant negative

associations with tussock and shrub functional group richness.

DISCUSSION

Six years after intentionally introducing H. lepidulum plants at

a wide range of densities, we found no significant impacts on

the resident plant communities. Rather, H. lepidulum appears

to have simply entered these communities and to co-exist as an

additional member. Indeed, plot species richness increased

over the 6 years of our study, with a higher rate of increase in

those plots with greater H. lepidulum density. Furthermore, in

2009 plant community richness, evenness and diversity showed

no evidence of significant decline at the plot or cell scales in

response to H. lepidulum density or cover. Wiser et al. (1998)

also reported a positive association between H. lepidulum

presence and native richness in a long-term study of invasion

Table 2 Coefficient estimates (± SE) for the longitudinal

hierarchical regression model describing change in plot species

richness (n = 742). Habitat coefficients are all relative to the

reference habitat forest creek. Also shown are F values and

associated P values.

Estimate SE F P

Intercept )3.988 8.801

Year 0.208 0.012 412.164 0.000

2009 Hieracium lepidulum

Density

)0.005 0.002 0.003 0.957

Year : 2009 H. lepidulum

Density

0.002 0.000 31.935 0.000

Habitat

Forest Creek 0 – 10.636 0.000

Forest )1.571 1.986

Forest Gap )0.119 2.168

Subalpine Creek 13.049 4.621

Scrub 9.346 4.400

Tussock 9.562 4.040

Altitude )0.024 0.009 5.937 0.015

Aspect 0.015 0.013 2.301 0.129

Light (%PPFD) 0.017 0.005 10.338 0.001

Soil

Moisture 0.003 0.036 0.042 0.839

Depth 0.012 0.009 1.575 0.210

pH 0.229 1.912 0.327 0.568

Carbon 0.039 0.797 0.467 0.494

Nitrogen )2.347 12.496 0.045 0.831

Phosphorus -0.045 0.142 0.108 0.742

Potassium 0.444 0.891 0.220 0.639

100101# Hieracium individuals per cell

Cel

l spe

cies

ric

hnes

s

02

46

810

12

338100100

05

1015

2025

# Hieracium individuals per plot

Plo

t spe

cies

ric

hnes

s

Figure 4 Hierarchical regression models of species richness, with

2009 Hieracium lepidulum density as the fixed effect at the

30 · 30 cm plot (top, n = 747) and 5 · 5 cm cell (bottom,

n = 26,892) scales. Data points have been jittered for clarity. Note

the log scales on the x axes.

R. Meffin et al.


into a similar N. solandri var. cliffortiodes forest. However, the

observational nature of this study means it is unclear whether

this represents a lack of impact, as reported here, or

preferential invasion of species-rich sites.

The idea that invaders may often be accepted into the

resident community without discernable impact is not new

(Williamson & Fitter, 1996), but has seen a recent revival of

interest with the biotic acceptance theory of Stohlgren et al.

Coefficient estimate: density

OverallForest creek

ForestForest gap

Subalpine creekScrub

Tussock

OverallForest creek

ForestForest gap


Tussock

OverallForest creek

ForestForest gap


Tussock

Ric

hnes

sD

iver

sity

Eve

nnes

s

Coefficient estimate: cover

–0.15 –0.10 –0.05 0.00 0.05 0.10 0.15 –0.15 –0.10 –0.05 0.00 0.05 0.10 0.15

Figure 5 Coefficient estimates for hierarchical regression models of plant community richness, diversity and evenness in response to

Hieracium lepidulum density and cover, overall and by habitat at the cell (5 · 5 cm) scale (n = 26,892). Significant results (after

Holm–Bonferroni correction) indicated in bold with solid circles.

Coefficient estimate: density

Basal herbClumped herbCreeping herb

ShrubErect leafy

GrassMat/cushion

OtherProstrate shrub

Tussock


ShrubErect leafy

GrassMat/cushion


Tussock


ShrubErect leafy

GrassMat/cushion


Tussock

Ric

hnes

sD

iver

sity

Eve

nnes

s

Coefficient estimate: cover

–0.015 –0.005 0.000 0.005 0.010 0.015 –0.004 –0.002 0.000 0.002 0.004

Figure 6 Coefficient estimates for hierarchical regression models of plant community richness, diversity and evenness in response to

Hieracium lepidulum density and cover at the cell (5 · 5 cm) scale (n = 26,892) and partitioned by functional group. Significant results

(after Holm–Bonferroni correction) indicated in bold with solid circles.



(2006) and support from studies such as Mills et al.(2009).

Grice (2006), in a review of invasive weed impacts in Australia,

also suggested that research is dominated by a few high profile

invasive species with clear impacts, while the remaining

majority may well have negligible effects on the invaded

communities. However, these studies have often been per-

formed at larger spatial scales than those studied here and thus

may also reflect covariation in alien abundance and native

plant richness across environmental suitability gradients. Our

results showed a positive relationship between invader abun-

dance and native richness, indicating no impact, even at a scale

of 5 · 5 cm. It is difficult to believe species are not interacting

at this spatial scale.

Why are impacts on native diversity, richness or evenness

negligible in this system? First, H. lepidulum does not appear

to be competitively superior to native species; it is smaller

statured than herbaceous community dominants in all hab-

itats, and observations over 6 years indicate that it does not

grow more rapidly than native species. Thus, its ability to

invade may be more a reflection of its opportunistic nature

and ability to colonize small-scale disturbances rapidly.

Second, plant growth rates and belowground resource levels

tend to be low at montane sites near tree line (Richardson &

Friedland, 2009), and this may limit plant competition – with

most species more akin to stress tolerators (sensu Grime,

1979) than competitors. Under such circumstances, positive

rather than negative plant–plant interactions may be more

likely (Callaway et al., 2002; Brooker et al., 2008). In an

examination of putative factors influencing H. lepidulum

establishment, Radford et al. (2010) reported that H. lepidu-

lum cover in invaded communities decreased significantly after

nutrient addition, suggesting nutrient limitation suppressed

competition between the invader and resident community.

The combination of an opportunistic colonizing strategy and

facilitation by native species could explain the positive increase

in richness when H. lepidulum invades these communities.

Alternatively, these communities may not be saturated, and

thus opportunities exist for additional species to enter,

especially where the environment is conducive to growth. If

species richness increased most in plots with conditions

conducive to plant growth, then it might be expected that

H. lepidulum would also perform better in those plots,

persisting preferentially because of the same favourable

conditions.

Teasing apart these two possible explanations for the

positive relationship is challenging, although evidence for the

latter explanation stems from the analysis of functional groups.

We expected groups functionally similar to H. lepidulum, such

as basal herb, creeping herb and erect leafy, to be dispropor-

tionately impacted. In fact, the model parameter estimates are

positive, suggesting some plots are favourable sites for both the

native species in these functional groups and H. lepidulum.

Significant negative parameter estimates were obtained for

clumped herb, mat/cushion, shrub and tussock groups. The

shrub and tussock groups contain robust, comparatively

large-statured species, with the negative parameter estimates

suggesting the presence of these species provides some

resistance to invasion. Similarly, the clumped herb group

consists primarily of vegetatively spreading Celmisia species,

which form dense clumps, and the mat/cushion group tend to

also form a dense ground cover, which may restrict the

availability of sites suitable for H. lepidulum to establish.

Nevertheless, previous studies have reported apparent

impacts at some locations, with H. lepidulum appearing to

form dense mats to the exclusion of native species (Rose &

Frampton, 1999), and other members of the Hieracium genus

appear to impact montane plant communities (Scott et al.,

1990). Given the unbiased distribution of study blocks across

the major habitat types found in the area, it is likely that the

results of our experimental sowing describe the average impact

of H. lepidulum across the site and within each community. It

seems that under certain conditions, such as at lower

elevations, or perhaps given more time, H. lepidulum can

reach higher densities/cover than we were able to experimen-

tally establish and may have an impact at sites where this

occurs.

Our results show that invasion by H. lepidulum into

montane forest and grassland habitats has no discernable

negative impacts on key plant community attributes across

habitats, scales and functional groups. The unexpected nature

of this result highlights the need for rigorous quantitative

assessment of invader impacts. As mechanism influences both

the type and spatial scale of impact, it needs to be considered

when planning research to ensure the appropriate attributes of

the native community are measured at the appropriate scales.

The lag time between invasion and impact means that studies

must also be conducted over appropriate temporal scales. To

better understand the frequency and magnitude of invader

impacts, we need to investigate the impacts of a representative

range of invasions, rather than only those where impacts are

clearest. Experimental introductions represent a valuable and

under-utilised tool for assessing the impacts of invasive plants.

Further studies of this nature are required if we are to

formulate a general understanding of the interactions between

invader, invaded community and impact.

ACKNOWLEDGEMENTS

We gratefully acknowledge the Miss E.L. Hellaby Indigenous

Grasslands Research Trust for funding. The Bio-Protection

research centre provided financial support for the preparation

of this article. Ross Meffin was funded by the New Zealand

Plant Protection Society Dan Watkins Scholarship in Weed

Science. Thanks to Robyn Damaryhoman, Laura Spence and

others for valuable field assistance.

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SUPPORTING INFORMATION

Additional Supporting Information may be found in the online

version of this article:

Appendix S1 Mean abiotic characteristics of plots by habitat.

Appendix S2 Results of hierarchical mixed models of richness,

evenness and diversity of native plant communities in response

to experimental gradients of cover and density of Hieracium

lepidulum.

R. Meffin et al.


As a service to our authors and readers, this journal provides

supporting information supplied by the authors. Such mate-

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BIOSKETCHES

Ross Meffin is a PhD student supervised by Richard P.

Duncan and Philip E. Hulme at the Bio-Protection Research

Centre, Lincoln University. He is broadly interested in ecology,

with a current focus on plant invasions.

Author contributions: R.M., R.P.D. and P.E.H. conceived the

research. A.L.M. set up the experiment. R.M., A.L.M. and

others collected the data. R.M. carried out the research with

advice from R.P.D. and P.E.H. All authors contributed to

writing the manuscript.

Editor: David Richardson



BIODIVERSITY RESEARCH: Experimental introduction of the alien plant Hieracium lepidulum reveals no significant impact on montane plant communities in New Zealand

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