Invasion success and impacts of
Hieracium lepidulum in a
New Zealand tussock grassland and
montane forest
A thesis
submitted in partial fulfilment
of the requirements for the Degree of
Master of Science
at
Lincoln University
by
Ross Meffin
Lincoln University
2010
A version of chapter three of this thesis, with the title
“Experimental introduction of the alien weed Hieracium
lepidulum reveals no significant impact on montane plant
communities in New Zealand”, authored by Ross Meffin, Alice
L. Miller, Philip E. Hulme and Richard P. Duncan is currently
under review for publication in the journal Diversity and
Distributions.
ii
Abstract of a thesis submitted in partial fulfilment of the
requirements for the Degree of M.Sc.
Invasion success and impacts of Hieracium lepidulum
in a New Zealand tussock grassland and montane forest
by
R. Meffin
Invasive species represent a major concern; they can result in serious ecological and
economic losses and are recognised as one of the most serious threats to global species
diversity. Plant invasions are of particular concern in New Zealand, which has high
proportions of both naturalised and endemic plant species.
In this thesis I focussed on the invasive plant Hieracium lepidulum, an exotic weed
introduced from Europe to New Zealand prior to 1941. It is invasive in a variety of
habitats in the South Island, where it has steadily increased in distribution and abundance
over the last 50 years, and is thought to have detrimental impacts on native plant
communities. I investigated factors influencing its invasion success and tested for impacts
on native plant communities, making extensive use of existing plots into which H.
lepidulum was experimentally introduced in 2003.
I examined how community richness, turnover, resource availability and propagule
pressure of the invader interacted to determine the invasion success of H. lepidulum.
Results differed markedly above and below treeline. Above treeline, plots with higher
richness and turnover were more invaded; below treeline, plots with higher available light
were more invaded. In both habitats, these findings were modified by the influence of
propagule pressure; at low propagule pressure, site characteristics were non-significant in
explaining invasion success, while at higher propagule pressure these effects became
significant.
To test for impacts resulting in altered community composition and structure, I looked for
changes in community richness, diversity and evenness subsequent to H. lepidulum
introduction. As impacts may be more apparent at fine spatial scales, I made measurements
at a 5 x 5 cm cell scale in addition to the established 30 x 30 cm plot scale. Plot species
richness increased from 2003 to 2009 and a component of this increase was associated
iii
with H. lepidulum density. Other relationships between the plant community and H.
lepidulum were generally non-significant. Results showed that H. lepidulum has had no
negative effects on community richness, evenness or diversity.
Despite being able to opportunistically colonise grassland sites with high turnover, and
forest sites subject to canopy disturbance, dependant on propagule pressure, it appears H.
lepidulum has not impacted community composition or structure.
Keywords: diversity, Hieracium lepidulum, exotic weeds, hierarchical data, biological
invasions, invasive species, mixed model, invasibility, turnover, carousel model, propagule
pressure, invasion resistance, resource fluctuation, impact.
iv
Table of contents
Abstract.................................................................................................................................iii
Table of contents...................................................................................................................v
List of figures......................................................................................................................vii
List of tables..........................................................................................................................x
Chapter 1:
Introduction...............................................................................................11
1.
Overview......................................................................................................11
2.
Study Species...............................................................................................12
3.
Methodological approach.............................................................................14
Chapter 2:
3.1.
Experimental introduction...............................................................14
3.2.
Spatial scale.....................................................................................15
3.3.
Analysis...........................................................................................16
Above and below treeline processes interact with
propagule pressure to determine invasion success..................................18
Abstract........................................................................................................18
1.
Introduction..................................................................................................18
2.
Methods........................................................................................................22
2.1. Study site...............................................................................................22
2.2. Experimental setup................................................................................22
2.3. Data collection.......................................................................................23
2.3.1. Plot abiotic characteristics........................................................23
2.3.2. Quantifying H. lepidulum abundance and
native community structure..................................................23
3.
Results..........................................................................................................25
3.1. Individual models..................................................................................25
3.2. Below treeline.......................................................................................33
3.3. Above treeline.......................................................................................33
4.
Discussion....................................................................................................36
v
Chapter 3: No significant impacts on native community composition or structure....40
Abstract....................................................................................................................40
1.
Introduction..................................................................................................40
2.
Methods.......................................................................................................42
2.1. Study site...............................................................................................42
2.2. Experimental setup................................................................................43
2.3. Data collection......................................................................................43
2.3.1. Plot abiotic characteristics.....................................................43
2.3.2. Quantifying H. lepidulum abundance and native community
structure................................................................................43
2.4. Analysis.................................................................................................44
2.4.1. Longitudinal analysis.............................................................45
2.4.2. Plot scale................................................................................46
2.4.3. Cell scale................................................................................47
3.
Results..........................................................................................................47
3.1. Hieracium lepidulum............................................................................47
3.2. Longitudinal analysis............................................................................47
3.3. Plot scale...............................................................................................53
3.4. Cell scale...............................................................................................53
3.5. Functional groups..................................................................................53
4. Discussion............................................................................................................60
Chapter 4: General discussion..........................................................................................63
1. Key results and implications................................................................................63
2. Further research and recommendations...............................................................64
Acknowledgements............................................................................................................66
Bibliography.......................................................................................................................67
vi
List of figures
Figure 1 Histogram of mean Hieracium lepidulum density in 2009 and 95% confidence
intervals, six years after addition of H. lepidulum seeds at five treatment densities (25, 125,
625, 3125, 15625 seeds respectively per 30 x 30 cm plot, columns T1 – T5) and control
and procedural controls (no seeds added, no seeds added with wetting, columns C and
PC).......................................................................................................................................25
Figure 2 Fitted hierarchical regression model of 2009 Hieracium lepidulum density per 30
x 30 cm plot, with species richness, seed addition treatment and their interaction as fixed
effects (n = 529)...................................................................................................................26
Figure 3 Fitted hierarchical regression model of 2009 Hieracium lepidulum density per 30
x 30 cm plot, with extinction rate, seed addition treatment and their interaction as fixed
effects (n = 529)...................................................................................................................27
Figure 4 Fitted hierarchical regression model of 2009 Hieracium lepidulum density per 30
x 30 cm plot, with available light, seed addition treatment and their interaction as fixed
effects (n = 529)...................................................................................................................28
Figure 5 Fitted hierarchical regression model of log transformed 2009 Hieracium
lepidulum density per 30 x 30 cm plot, with soil moisture, seed addition treatment and
their interaction as fixed effects (n = 529)...........................................................................29
Figure 6 Fitted hierarchical regression model of log transformed 2009 Hieracium
lepidulum density per 30 x 30 cm plot, with N:C ratio, seed addition treatment and their
interaction as fixed effects (n = 529)...................................................................................30
Figure 7 Coefficient estimates from fitted hierarchical regression models of below treeline
2009 Hieracium lepidulum density per 30 x 30 cm plot, with 2003 plot species richness,
extinction rate, available light, soil moisture, N:C ratio and their interactions with seed
addition treatment as fixed effects (n = 529). Significant results (p < 0.05) indicated with a
bold line and solid circle......................................................................................................32
vii
Figure 8 Coefficient estimates from fitted hierarchical regression models of below treeline
2009 Hieracium lepidulum density per 30 x 30 cm plot, with 2003 plot species richness,
extinction rate, available light, soil moisture, N:C ratio and their interactions with seed
addition treatment as fixed effects (n = 529). Significant results (p < 0.05) indicated with a
bold line and solid circle......................................................................................................33
Figure 9 Mean density (number of individuals per 30 x 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)
..............................................................................................................................................46
Figure 10 Frequency histograms of Hieracium lepidulum density (number of individuals
per 30 x 30 cm plot) and percent cover (total foliar cover per plot) in 2009 across all seed
addition and control plots....................................................................................................47
Figure 11 Change in mean species richness per 30 x 30cm plot (circles with 95%
confidence intervals, n = 742), overall and in habitats above (top) and below treeline
(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......................................................................................................................49
Figure 12 Hierarchical regression models of species richness, with 2009 Hieracium
lepidulum density and cover as the fixed effects at the 30 x 30 cm plot scale (n = 747).
Data points have been jittered for clarity. Note the log scales on the x axes.......................52
Figure 13 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 plot (30 x 30 cm) scale (n = 742). Significant results (p < 0.05
after
Holm-Bonferroni
correction)
indicated
in
bold
with
solid
circles...................................................................................................................................53
Figure 14 Hierarchical regression models of species richness, with 2009 Hieracium
lepidulum density and cover as the fixed effects at the 5 x 5 cm cell scale (n = 26,892).
viii
Data points have been jittered for clarity. Note the log scales on the x
axes......................................................................................................................................54
Figure 15 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 x 5 cm) scale (n = 26,892). Significant results (p < 0.05
after
Holm-Bonferroni
correction)
indicated
in
bold
with
solid
circles...................................................................................................................................55
Figure 16 Coefficient estimates for hierarchical regression models of plant community
richness, diversity and evenness in response to Hieracium lepidulum density and cover at
the plot (30 x 30 cm) scale (n = 747) and partitioned by functional group. Significant
results (p < 0.05 after Holm-Bonferroni correction) indicated in bold with solid
circles...................................................................................................................................56
Figure 17 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 x 5 cm) scale (n = 26,892) and partitioned by functional group. Significant
results (p < 0.05 after Holm-Bonferroni correction) indicated in bold with solid
circles...................................................................................................................................57
ix
List of tables
Table 1 Mean (± s.e) community species richness, diversity, evenness, and Hieracium
lepidulum density and % foliar cover per 30 x 30 cm plot for each of six habitat types in
2009. Treatment plots had five densities (25, 125, 625, 3125 and 15625 seeds/plot) of H.
lepidulum seeds added in 2003 (n = 532), control plots had no seed added (n
=215)....................................................................................................................................48
Table 2 Coefficient estimates (± s.e.) 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...................................................................................................................................50
x
Chapter 1: Introduction
1.
Overview
Invasive species represent a major concern; they can result in serious ecological and
economic losses (1995; Pimentel et al. 2000; Levine et al. 2003; Dukes and Mooney 2004)
and are recognised as one of the most serious threats to species diversity, second only to
habitat destruction (Diamond 1989; Burhenne-Guilmin et al. 1994; Walker and Steffen
1997; Williamson 1999). Examples of the impacts of invasive plants are well documented;
they can considerably alter the composition and structure of native communities, causing
declines in abundance and local extinctions of native plant species (e.g. Pysek and Pysek
1995; Vila et al. 2006; Hejda et al. 2009) and modify ecosystem structure and function
(e.g. Vitousek and Walker 1989; D'Antonio and Vitousek 1992; Ogle et al. 2003). Weed
invasions are of particular concern in New Zealand, which has 2,190 naturalised exotic
plant species, of which 325 are recognized as invasive weeds by the Department of
Conservation, compared with just 1,896 indigenous plant species (Timmins 2004), many
of which are endemic. Understanding factors controlling the spatial distribution of these
invasive species and how they may alter the communities they invade is crucial to
management and conservation efforts, in addition to providing insights into the ecological
processes at work.
In this thesis I focus on the invasive herbaceous weed Hieracium lepidulum, widespread
and considered to be of conservation concern in the South Island of New Zealand. The
thesis consists primarily of two chapters presented as stand-alone scientific papers; chapter
two looks at the spatial distribution of H. lepidulum, investigating factors influencing its
invasion success, while chapter three looks at its subsequent effects, testing for detrimental
impacts on native plant community structure and composition. Chapter one provides a
concise review of current knowledge concerning the study species and invasion success
and impacts of invasive plants in general, along with an outline of the methodological
approach taken in the studies presented. Chapter four presents a summary and synthesis of
the main results, and places them in the context of current research.
The outcome of any particular invasion is thought to be determined by invader propensity
to invade new habitats (invasiveness), the inherent susceptibility of a community to
invasion (invasibility) and the size and frequency of introductions (propagule pressure);
and is likely to be the result of the complex interplay among all of these (Lonsdale 1999).
11
Here I focus on one species, H. Lepidulum, taking its characteristics as given, and examine
the effects of community invasibility and propagule pressure. Propagule pressure is
thought to be a major driver of invasion success (Tilman 1997; Cassey et al. 2004; Von
Holle and Simberloff 2005; Colautti et al. 2006) and I investigate how it interacts with
factors influencing invasibility: community richness, turnover and resource availability.
Community richness is used as a measure to test the diversity-resistance hypothesis, which
suggests more diverse communities can monopolise resources more effectively and are
inherently more stable, and thus have greater ‘biotic resistance’ to invasion (Elton 1958;
MacArthur 1972; Kennedy et al. 2002). Turnover may provide opportunities for invaders
to establish as sites become vacant in a shifting mosaic of species composition, as
suggested in the ‘carousel model’ of turnover (VanderMaarel and Sykes 1993). The
influence of resource availability is likely to be complex; the shape of the relationship is
likely to depend on where on the resource gradient a site sits, and the relative
competitiveness of the invader and community dominants at that position. Furthermore
both richness and turnover may covary with resource availability.
Factors determining invader impact on resident communities are not well understood, and
it appears many invasive weeds may coexist as additional community members without
significant impact (Stohlgren et al. 2006). Evidence suggests impacts may be scale
dependant (Levine and D'Antonio 1999) and there is still much debate over the relative
importance of traits of the alien invader and how these interact with a those of a given
invaded community. There is a need for further studies across a representative of the range
of alien plant life-forms, invaded native communities and spatial scales (Mack 1996; Byers
et al. 2002; Mills et al. 2009). By quantifying the impacts of H. lepidulum in a variety of
habitats in the study site at two spatial scales, I undertake one of the most thorough
investigations of impact of an invader to date.
2.
Study species
Hieracium lepidulum is a broad leaved, tap-rooted, 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 is an exotic weed introduced from Europe to New
Zealand, where it was first recorded in 1941 (Miller 2006). It has become invasive in the
South Island, where it has steadily increased in distribution and abundance over the last 50
12
years (Rose et al. 1995; Duncan et al. 1997; Wiser et al. 1998; Mark et al. 1999; Wiser and
Allen 2000; Duncan et al. 2001).
Studies of factors influencing H. lepidulum success have produced a variety of
conclusions; propagule pressure, reduced competition due to disturbance, and competitive
advantages at both high and low resource availabilities have all been cited as determining
factors. Furthermore, many of these findings appear to be contradictory.
One of the more comprehensive studies of factors influencing the invasion success of H.
lepidulum, notable for tracking the invasion as it occurred, is that of Wiser et al. (1998).
This study investigated invasion into mountain beech forest (Nothofagus solandri var.
Cliffortiodes) over 23 years. They found that initially, when propagule pressure was low,
this limited dispersal and invasion success; but as the invasion progressed and propagule
pressure increased, plots with more species and resources were more invaded - in
contradiction to the expectations of the diversity resistance hypothesis. They speculated
that diversity may actually promote invasion, suggesting potential dominance of one or
more competitively superior species in low richness communities, and increased temporal
and spatial heterogeneity in species rich communities as possible mechanisms. They also
reported that disturbance in the form of canopy gaps, with an attendant increase in
resources in the form of light, initially promoted invasion, but this effect weakened with
time.
In a pot study investigating the effects of resource availability and disturbance on the
competitiveness of H. lepidulum relative to species co-occurring in the field, Radford et al.
(2006) concluded that H. lepidulum was not a strong competitor in either high or low
resource environments, and that reduced competition due to disturbance was likely to
facilitate its success. In another study of invasion success, Radford et al. (2009)
investigated the possibility that a high leaf area to total biomass ratio may provide H.
lepidulum with a growth advantage over native species. Their findings showed this was the
case in low resource environments, but that the advantage was diminished in high resource
environments, contradicting the expectations of the resource enrichment hypothesis (Davis
et al. 2000). They suggested that this growth advantage at low resource levels enabled H.
lepidulum to successfully invade environments where native species were resource limited,
although perhaps this advantage may not extend to sites with extremely low resource
levels.
Other studies have suggested invasion success may be due to high propagule pressure and
an ability to colonise sheltered inter-tussock micro-sites (Rose and Frampton 1999; Rose et
13
al. 2004) in conjunction with competitive advantages which then allow H. lepidulum to
displace native species and become dominant over time (Rose and Frampton 1999; Rose
and Frampton 2007).
At present it difficult to reconcile these varying findings, but they suggest there may be
multiple mechanisms operating, and that these may differ between environments above
and below treeline, with high and low resource availability, and between disturbed and
undisturbed sites.
It is believed that H. lepidulum may have significant negative impacts on native plant
communities, although little work has been done to quantify any such effect. It is thought
H. lepidulum may compete strongly with native species (Fan and Harris 1996; Moen and
Meurk 2001; Weigelt et al. 2002; Winkler and Stocklin 2002; Lamoureaux et al. 2003)
leading to declines in the species diversity of plant communities it has invaded (Scott et al.
1990; Rose and Frampton 1999; Wiser and Allen 2000; Espie 2001; Rose et al. 2004; Rose
and Frampton 2007).
In contrast Wiser et al. (1998) found a positive correlation between H. lepidulum
occurrence and community richness; this could be interpreted as suggesting that rather
than negatively impacting native species, H. lepidulum actually promotes native richness.
However, no data was collected quantifying the richness of these sites prior to invasion; as
such the possibility that this result instead reflects preferential invasion of species rich sites
cannot be discounted.
As evidence is limited to observational, correlative studies it is difficult to determine
whether observed relationships between the invader and native communities reflect
impact, or preferential invasion of species poor or rich sites. The experimental approach
described below provides the first explicit test of the effects of H. lepidulum invasion on
community structure and composition, able to distinguish between invasibility and impact.
3.
Methodological approach
3.1. Experimental introduction
Many studies of both invasion success and impact have been limited by the correlative
approach taken, comparing already invaded and uninvaded habitats (Levine et al. 2003).
Tests of the diversity resistance hypothesis which take such an approach run the risk of
attributing invasion to low species diversity, when this may in fact be the result of invasion
and its subsequent impact. Studies of impact may be similarly confounded, attributing low
diversity in invaded sites to impacts of the invader, when in fact they have been
14
preferentially invaded due their low diversity, as suggested in the diversity resistance
hypothesis.
One alternative for impact studies is to perform removal experiments (e.g. Ogle et al.
2003; Stinson et al. 2007), in which the invasive plant in question is removed from
experimental plots, and the effects of this removal examined. The act of removal itself, and
the associated disturbance may obscure any invader effects (Diaz et al. 2003), particularly
in communities not naturally subject to high levels of disturbance.
An alternative is experimental sowing or introduction of an invasive species in a natural or
semi-natural habitat. While some caution is warranted, appropriately applied introduction
experiments have great potential to bypass the difficulties encountered by other
methodologies. In this study I was able to utilise such an introduction experiment
established in 2003 by Alice Miller (2006) as part of a study of the spatial distribution of
H. lepidulum. These plots have subsequently been remeasured by a variety of individuals,
including myself (in 2009), to maintain a long term data set. This provided me with long
term, replicated data, tracking the dynamics of the community and invader from the time
of introduction, and did not necessitate species removals and the associated disturbance.
The layout and establishment of the plots, and measurements made prior to 2009 were thus
the work of other researchers, but the studies presented in this thesis represent novel
analyses and applications of the data set, to answer questions not previously addressed.
3.2. Spatial scale
In studies of invasibility, the relationship between native and exotic diversity is thought to
be scale dependant (Shea and Chesson 2002; Davies et al. 2005). At fine scales (< 1 m2, in
the range of plot sizes typically used in experimental studies) competitive effects are
expected to result in a negative relationship between the two; at coarser scales (often used
in observational studies) competitive effects are less apparent, while both native and
exotic diversity are likely to covary with resource availability, disturbance and spatial
heterogeneity leading to a positive relationship (Drake et al. 1989; Levine and D'Antonio
1999).
Spatial scale is also important in studies of invasive species impact (Pauchard and Shea
2006). Impacts due to resource competition occur in the neighbourhood of the invader, at
small scales - competition for light occurs when two individuals overlap in horizontal
extent above ground, allowing one plant to overtop and shade out another, competition for
water and nutrients occurs when two individuals overlap in extent below ground and their
15
root systems are intermingled in the same patch of soil (Keddy and Shipley 1989). When
impacts operate through ecosystem processes, for example by altering nutrient cycling or
fire regimes (Dantonio and Vitousek 1992; Dukes and Mooney 2004), the effects can
extend far beyond the immediate proximity of the plant and operate at the landscape scale.
To account for these effects, I conducted my study of impacts at two spatial scales, a 30 x
30 cm plot scale and a 5 x 5 cm cell scale. While impacts due to competitive exclusion
should be apparent at the plot scale, at such a fine scale as 5 x 5 cm it is difficult to believe
any impacts would be missed. Furthermore, this gave me the opportunity to compare any
impacts found at the two scales, potentially giving me the ability to distinguish competitive
exclusion and ecosystem process impacts.
I conducted my study of invasibility at the plot scale, allowing me to test the diversity
resistance hypothesis rigorously. At this scale competitive effects should be apparent and
those of spatial heterogeneity should be minimised; thus any negative relationship between
diversity and invasibility should be apparent at this scale. As there were no data available
to quantify native community diversity at the cell scale prior to H. lepidulum introduction,
I was unable to test the diversity resistance hypothesis at this scale.
3.3. Analysis
I used the statistical software R (R_Development_Core_Team 2009) to formulate
hierarchical mixed-effect models using the package lme4 (Bates and Maechler 2009). This
technique fits a linear model to data with fixed and random effects; fixed effects are those
affecting the entire population, or repeatable manipulations of interest, such as
experimental treatments, covariates and their interactions; random effects are those
operating on experimental units within the data set, such blocks or repeated measurements,
and represent stochastic variation in a larger population. For each effect a coefficient
describing the magnitude and direction of its relationship with the response variable is
estimated using Restricted Maximum Likelihood (REML). These coefficients can be
interpreted as the effect of that parameter on the intercept (for single terms) or slope (for
interaction terms) of relationship being modelled.
My experimental design incorporated cells nested within plots, plots nested within blocks,
and repeated measures on the same plots. By modelling these grouped experimental units
as random effects, I was able to into account their non-independence through time, due to
repeated measures, and spatially, due to the nested blocks design. Other approaches, such
as fitting a linear model, risk pseudoreplication and violating assumptions of
16
independence. Treating these variables as random effects also maximises statistical power,
as it uses fewer degrees of freedom than treating them as fixed effects and estimating a
parameter for each level. Instead, for each effect a mean and standard deviation are
calculated, describing a normal distribution to which the observed values of each level
belong. Separating out these random effects thus allows better estimates to be made of the
fixed effects of experimental interest.
When fitting a linear mixed effect model two sets of assumptions need to be satisfied.
Firstly the within group errors must be independent, normally distributed about zero with
constant variance, and be independent of the random effects. Secondly, the random effects
must be normally distributed with a constant specified covariance matrix, and must be
independent among groups. To validate these assumptions I followed the methods in
Pinheiro and Bates (2000); firstly I checked the distribution of within group residuals, and
plots of the fitted versus observed values, then examined normal plots of the random
effects and scatter plots of the random effects. Assumptions were satisfactorily met for all
models presented.
17
Chapter 2: Above and below treeline
processes interact with propagule pressure
to determine invasion success
Abstract
Explaining the variable outcomes documented across a range of invasions is a major area
of research interest, giving rise to a variety of hypotheses. As yet, unifying these
hypotheses has proved difficult, and attempting to do so may indeed be fruitless. Most
explanations of invasion success enjoy at least some theoretical and empirical support, and
it seems possible that explaining the variety of outcomes, or indeed some specific
outcomes, may require a combination of hypotheses. I examine the invasion success of
Hieracium lepidulum above treeline and below treeline; experimentally adding seeds of the
invader at different densities (0 - 15625 seeds per 30 x 30 cm plot, n=529) using a
randomised, replicated block design. I examine how community richness, turnover,
resource availability and propagule pressure interact to determine density of the invader
six years after introduction, using hierarchical regression models. Results differed
markedly above and below treeline. Above treeline, plots with higher richness and
turnover were more invaded; below treeline, plots with higher available light were more
invaded, suggesting different mechanisms facilitated invasion. In both habitats, these
findings were modified by the influence of propagule pressure; at low propagule pressure,
measured site characteristics were non-significant in explaining invasion success, while at
high propagule pressure these effects became significant. These results highlight the
complex nature of plant invasions, and the importance of propagule pressure. In addition,
findings contradict the expectations of the diversity-resistance hypothesis, and are one of
the only fine-scale, experimental studies to do so.
1.
Introduction
As global travel and trade have increased, plants have being introduced outside their native
ranges; many have become invasive in these new habitats, and can result in serious
ecological and economic losses (Hobbs and Humphries 1995; Pimentel et al. 2000; Levine
et al. 2003; Dukes and Mooney 2004). However, many introduced plants fail to become
invasive, and amongst those which do, there is considerable variation in the magnitude of
invasion. Explaining this variability has become a major area of research interest; there are
18
many hypotheses to explain differences in invasion success (see Theoharides and Dukes
2007; Catford et al. 2009 for recent reviews), and available evidence provides at least
some support for all (Catford et al. 2009). As such a general theory of invasion success is
elusive, and may remain so.
It is recognised that the outcome of a particular invasion is most likely the result of the
complex interplay among invader identity, community susceptibility
to invasion
(invasibility) and the size and frequency of introductions (propagule pressure) (Lonsdale
1999). Thus, for a given invader, the degree of invasion success is likely to depend on
community invasibility and its interaction with propagule pressure; however, these
interactions are not often studied (but see Thomsen et al. 2006; Edward et al. 2009;
Eschtruth and Battles 2009).
Propagules of the invader are a prerequisite for any invasion to take place, and may be the
prime determinant of success (Tilman 1997; Cassey et al. 2004; Von Holle and Simberloff
2005; Colautti et al. 2006); increased propagule supply is thought to lead to greater
invasion success (Williamson and Fitter 1996; Cassey et al. 2004). Each propagule can be
visualised as passing through a process where variability in propagule genetic makeup,
introduction site, introduction time and other factors combine to determine the chances of
each propagule successfully establishing. It follows from this that more introductions
result in a greater probability of some being successful, and hence propagule pressure will
influence invasion success. Furthermore, at sufficiently high propagule pressures, despite
this stochasticity, enough propagules will be introduced to opportune sites at opportune
times that the effects of community resistance become overwhelmed, and previously
uninvasible communities become invasible (Von Holle and Simberloff 2005). Despite this,
few studies have quantified the effects of propagule pressure and how they may interact
with community invasibility to determine invasion success (but see Colautti et al. 2006;
Edward et al. 2009; Eschtruth and Battles 2009).
Community invasibility is likely to depend on a number of factors. The diversity-resistance
hypothesis suggests that more diverse communities can monopolise resources more
effectively and are inherently more stable, and thus have greater ‘biotic resistance’ to
invasion (Elton 1958; MacArthur 1972; Kennedy et al. 2002). We may thus expect more
species rich communities to be less invasible, particularly finer spatial scales where
competitive effects are strongest, and the confounding effects of spatial heterogeneity and
covariation between richness, resource availability and other variables are weakest (Drake
et al. 1989; Levine and D'Antonio 1999; Shea and Chesson 2002; Davies et al. 2005).
19
Species turnover may also influence community invasibility. In the ‘carousel model’ of
turnover (VanderMaarel and Sykes 1993), at fine spatial scales, species are thought to
appear and disappear from year to year, forming a shifting mosaic as individuals die and
colonise new sites; while at coarser scales community composition remains unchanged –
this effect has been documented in particular in grasslands (VanderMaarel and Sykes 1993;
VanderMaarel 1996). It seems likely that invaders may behave similarly to native species
(Huston 2004), and ‘jump aboard’ this carousel, colonising vacant sites, and thus more
readily invade high turnover communities where such sites become available more
frequently.
Background resource availability is also often thought to influence invasibility, although
for any given community the both positive and negative relationships are possible. Species
richness tends to show a humped relationship along resource gradients (Mittelbach et al.
2001; Stein et al. 2008). As we may expect the same factors to drive both native and exotic
(Huston 2004), we would expect exotic richness, and thus invasibility, to show a similar
humped relationship. At low resource levels we may expect a positive relationship between
resource availability and invasibility, while at high resource levels we may expect the
inverse to be true, and at intermediate levels there may be little influence. Just such a
humped relationship was reported along a moisture gradient by Rejamanek (1988);
although mostly relationships cited in the literature are positive (e.g. Huenneke et al. 1990;
Burke and Grime 1996),
non-significant (e.g. Sandler et al. 2007), or negative
(e.g.Tanentzap and Bazely 2009). These studies may well have been conducted across
gradients of a magnitude that revealed only part of the overall humped relationship. Most
often positive relationships are reported (Funk and Vitousek 2007), and this may reflect the
majority of studies taking place in resource limited communities. Furthermore, invasion
outcomes along a resource gradient will depend on the relative competitive strengths of the
invader and community dominants at low, high and intermediate resource levels.
Studies of factors influencing H. lepidulum invasion success have produced a variety of
conclusions; propagule pressure, reduced competition due to disturbance, and competitive
advantages at both high and low resource availabilities have all been cited as determining
factors, and many of these findings appear to be contradictory. In a study tracking invasion
of H. lepidulum into mountain beech forest (Nothofagus solandri var. Cliffortiodes),
similar to that found in the study site, Wiser et al. (1998) found that at low propagule
pressures this factor limited invasion success; but at higher propagule pressure, sites with
higher species richness and more resources - notably in the form of light resulting from
20
treefall gaps – were more invaded. Studies of the competitive abilities of H. lepidulum
relative to co-occurring species have been somewhat inconclusive, suggesting both that it
is not a strong competitor at high or low resource availability, but colonised disturbed sites
where competition was reduced (Radford et al. 2006), and that it has a competitive growth
advantage over co-occurring species at low resource levels (Radford et al. 2009). Other
studies have suggested invasion success may be due to high propagule pressure and an
ability to colonise sheltered inter-tussock micro-sites (Rose and Frampton 1999; Rose et al.
2004), in conjunction with competitive advantages which then allow H. lepidulum to
displace native species and become dominant over time (Rose and Frampton 1999; Rose
and Frampton 2007). At present it is difficult to reconcile these varying findings, but they
suggest there may be multiple mechanisms operating - above and below treeline, with high
and low resource availability, and in disturbed and undisturbed sites.
In this study I investigate four variables thought to influence invasion success, or
implicated in existing studies of H. lepidulum invasion; community richness and turnover,
resource availability, and propagule pressure, focusing in particular on interactions
between invasibility and propagule pressure. Both biotic (richness and turnover) and
abiotic (resource availability) characteristics of communities are likely to differ
significantly above and below treeline; for example, in the study area below treeline
communities tend to be dominated by relatively few, long lived species, and are often
limited by light availability, whereas above treeline communities may be more diverse,
with higher turnover rates and limited by nutrient availability. We may thus expect
invasion success to be dependent on propagule pressure and light availability below
treeline, as suggested by Wiser et al. (1998); but above treeline, in grasslands comparable
to those studied by VanderMaarel and Sykes (1993), we may expect the interaction of
propagule pressure and turnover to be of primary importance.
Hieracium lepidulum seeds were experimentally added at different densities to 30 x 30 cm
plots in three above treeline and three below treeline communities, and the subsequent
fates of the invader and native communities followed over six years. At this plot size,
competitive effects as proposed by the diversity-resistance hypothesis should be apparent,
as should any effects of turnover (sensu VanderMaarel and Sykes 1993). This design
allowed me to examine the effects on H. lepidulum invasion success of 1) propagule
pressure, which was manipulated via seed addition treatments; 2) native community
richness; 3) species turnover; and 4) background resource availability.
21
2.
Methods
2.1. Study site
The study was conducted in the Craigieburn Stream catchment, Craigieburn Forest Park,
Canterbury, New Zealand (43o10’S, 172°45'E). The terrain is mountainous, with elevations
ranging from 800m to 2000m. The dominant vegetation is mountain beech forest
(Nothofagus solandri var. cliffortiodes) to an elevation of approximately 1400m, above
which it gives way to subalpine scrub, tussock grasslands, and alpine herbfields. The mean
annual temperature is 8.2°C, and mean annual precipitation is 1533 mm (Miller 2006).
2.2. Experimental setup
Hieracium lepidulum seeds were experimentally added at different densities to plots
located in six habitat types: three below treeline - forest creek, intact forest, forest canopy
gap (formed by treefall); and three above treeline - alpine creek, tussock grassland and
subalpine scrub. A randomised block design was used, 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 a random distance along the stream below treeline and above treeline
respectively. For intact forest and tussock grassland habitat, blocks were located 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
treeline 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 addition
treatments (21 plots per block): control with no seed addition and no initial wetting,
procedural control with no seed addition but with wetting, and seed addition at rates of 25,
125, 625, 3125 and 15625 seeds per 30 x 30 cm plot with wetting (or 278, 1389, 6944,
34722, and 173611 seeds/m2). Lower 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 contained more seeds than
expected to occur naturally, but allowed examination of a ‘worst case’ scenario. Wetting
22
was used to prevent the seeds from blowing 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.
2.3. Data collection
2.3.1. 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 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 re-weighed to
calculate gravimetric soil moisture content. The remaining soil was air dried, sieved and
analysed for carbon and nitrogen. For each plot, available light was measured as the
percentage of photosynthetic photon flux density (%PPFD) under overcast skies. To do
this, fifteen 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 x 100.
2.3.2. Quantifying H. lepidulum abundance and native community structure
Hieracium lepidulum seeds were added to the plots in March 2003 (autumn), and the
numbers of H. lepidulum plants in each plot was recorded in the following summer
(December 2003) and then recorded annually in February or March from 2004-2009. One
measurement was taken in December 2003 to capture peak densities which occurred at that
time, subsequent measurements were taken in February and March each year for
phenological consistency. In plots where high densities of H. lepidulum plants made a
complete census impractical, a 1 cm x 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 characterised by recording the
presence of all vascular plant species. Changes in composition were tracked by
23
remeasuring each plot in 2006, 2007 and 2009, again recording all vascular plant species
present. Identification was to species level in nearly all cases; in cases where there was a
lack of morphological features allowing identification to species level in the field,
identification was to genus level, and such groupings were kept consistent between years.
Ten plots (1%) which could not be relocated during the study were excluded from the
analyses.
2.4. Analysis
Propagule pressure was quantified as one five H. lepidulum seed addition treatments;
treatments one to five corresponding to additions of 25, 125, 625, 3125 and 15625 seeds
per plot (or 278, 1389, 6944, 34722, and 173611 seeds/m2).
I used native community richness in 2003 to quantify the influence of native community
richness on invasibility, as this reflects the state of the community prior to experimental
introduction of H. lepidulum.
As measures of turnover, I calculated the rates at which species colonised and became
locally extinct in each plot over the six years of the study. Using the presence and absence
of species at each of the four measurement times, I calculated the numbers of colonising
and locally extinct species over each measurement interval (Rusch and Vandermaarel
1992), and used these to calculate the number of species to become locally extinct and
colonisation per year. I then took the mean of these three rates; this approach minimised
the possibility that a species may enter and leave the plot between measurements.
Extinction rate was well correlated with colonisation rate (see results below), and as it is
local extinctions which provide an opportunity for H. lepidulum to establish, it was chosen
as the measure of species turnover.
As measures of the amount of resources available for plant growth I used light (%PPFD),
soil moisture content and soil nitrogen as measured by nitrogen to carbon ratio (N:C ratio).
The relative importance of these three resources is unknown, so to weight them equally, I
transformed each by subtracting the mean and dividing by one standard deviation.
To investigate the nature of the relationships between these variables and 2009 H.
lepidulum density, I fitted five hierarchical regression models to the data (Gelman and Hill
2006), one for each putative variable thought to influence invasibility (2003 richness,
extinction rate, available light, N:C ratio and soil moisture); these and their interactions
with propagule pressure were the fixed effects of the five models, and each model had
block as a random effect to account for the non independence of plots nested within
24
blocks. For the soil moisture and N:C ratio models, H. lepidulum density was log
transformed to improve model fit.
To test the relative importance of species richness, extinction rate and each of the three
resources, I fitted two further models - to data above and below treeline. Both incorporated
those five variables and the interaction of each with seed treatment as fixed effects, and
plot as a random effect. Contrasting results from previous studies indicate that there may
be significant differences in factors promoting H. lepidulum invasion between
communities above and below treeline.
Coefficient estimates from each model were tested for significance using t-tests. There is
uncertainty in estimating the denominator number of degrees of freedom for coefficients of
mixed effects models (Bates 2006; Baayen et al. 2008). To overcome this, I set these to be
the number of observations minus the number of fixed effects coefficients. While for small
data sets this produces anti-conservative p values, for data sets of this size (n = 529) it
provides a good measure of significance (Baayen et al. 2008).
3.
Results
Seed addition treatments resulted in differential invasion of plots by H. lepidulum (Fig. 1).
Plots subject to treatment one were not significantly more invaded than control plots or
procedural control plots (no seeds added and no seeds added with wetting). Plots subject to
treatments two, three, four and five all had significantly more H. lepidulum individuals in
2009 than controls. All treatments also differed significantly from each other, except
treatments four and five, where there was no significant difference in H. lepidulum
invasion success in 2009. Local extinction rate was well correlated with colonisation rate
(Pearson’s product-moment correlation = 0.88). As local extinctions provide opportunities
for H. lepidulum to establish, I used extinction rate as my measure of species turnover.
3.1. Individual models
Hierarchical regression models of 2009 H. lepidulum density fitted with 2003 richness,
extinction rate, light, soil N:C ratio and soil moisture as fixed effects all showed positive
interactions with seed addition treatment, but these were significant only at higher seed
addition densities. In the species richness model, plots with higher species richness in
2003 had significantly higher densities of H. lepidulum in 2009 for treatments three, four
and five (additions of 625, 3125 and 15625 seeds per 30 x 30 cm plot; Fig. 2), the opposite
of what is predicted by the diversity-resistance hypothesis. At lower seed addition densities
25
(treatments one and two) 2003 richness had no significant influence on H. lepidulum
invasion. This same pattern across treatments was also found for the extinction rate model
(Fig. 3); for treatments three, four and five plots with higher extinction rates contained
significantly more H. lepidulum individuals in 2009. Models of resource availability
showed qualitatively similar patterns, with plots with more available resources being more
heavily invaded; here the interactions were significant only for treatments four and five for
the light model (Fig. 4); three four and five for the soil moisture model (Fig. 5); and four
and five for the N:C ratio model (Fig. 6).
26
60
50
40
30
0
10
20
Individuals per plot
C
PC
T1
T2
T3
T4
T5
Figure 1
Histogram of mean Hieracium lepidulum density in 2009 and 95% confidence intervals,
six years after addition of H. lepidulum seeds at five treatment densities (25, 125, 625,
3125, 15625 seeds respectively per 30 x 30 cm plot, columns T1 – T5) and control and
procedural controls (no seeds added, no seeds added with wetting, columns C and PC)
27
400
treatment seeds sown
0
100
2009 H. lepidulum density
200
300
15625
3125
625
125
25
0
5
10
2003 plot richness
15
20
Figure 2
Fitted hierarchical regression model of 2009 Hieracium lepidulum density per 30 x 30 cm
plot, with species richness, seed addition treatment and their interaction as fixed effects (n
= 529).
28
400
treatment seeds sown
0
100
2009 H. lepidulum density
200
300
15625
3125
625
125
25
0
1
2
3
plot species extinction rate
4
5
Figure 3
Fitted hierarchical regression model of 2009 Hieracium lepidulum density per 30 x 30 cm
plot, with extinction rate, seed addition treatment and their interaction as fixed effects (n =
529).
29
400
treatment seeds sown
0
100
2009 H. lepidulum density
200
300
15625
3125
625
125
25
-1
0
1
light index
2
3
Figure 4
Fitted hierarchical regression model of 2009 Hieracium lepidulum density per 30 x 30 cm
plot, with available light, seed addition treatment and their interaction as fixed effects (n =
529).
30
1000
treatment seeds sown
0
2009 H. lepidulum density
10
100
15625
3125
625
125
25
-3
-2
-1
0
1
soil moisture index
2
3
Figure 5
Fitted hierarchical regression model of log transformed 2009 Hieracium lepidulum density
per 30 x 30 cm plot, with soil moisture, seed addition treatment and their interaction as
fixed effects (n = 529).
31
1000
treatment seeds sown
0
2009 H. lepidulum density
10
100
15625
3125
625
125
25
-3
-2
-1
0
N:C Ratio
1
2
3
Figure 6
Fitted hierarchical regression model of log transformed 2009 Hieracium lepidulum density
per 30 x 30 cm plot, with N:C ratio, seed addition treatment and their interaction as fixed
effects (n = 529).
32
3.2. Below treeline
In the model fitted to data below treeline with 2003 richness, extinction rate, light, N:C
ratio, soil moisture and their interactions with seed addition treatment as fixed effects,
richness and extinction rate were not significant in explaining invasion success. Of the
resources examined, only available light was significant in explaining H. lepidulum
invasion (Fig 7). Again, the effect was significant only at higher propagule pressures
(treatments four and five), where plots with more available light were significantly more
invaded by H. lepidulum. Below treeline habitats varied in available light; plots in forest
gap and forest creek habitats had more available light (8.3 ± 4.8 and 10.5 ± 7.8 %PPFD
respectively) than forest plots (3.6 ± 2.1 %PPFD). Controlling for other variables,
treatments four and five had more H. lepidulum individuals in 2009, but this was
significant only for treatment four.
3.3. Above treeline
In the equivalent overall model fitted to data above treeline, plots with higher 2003
richness had more H. lepidulum individuals in 2009 for treatments four and five, although
significantly so only for treatment four. Plots with higher extinction rates were
significantly more invaded by H. lepidulum for treatment five (Fig. 8). Resource
availability was non-significant in explaining H. lepidulum invasion success.
33
40
coefficient estimate
20
0
-20
-40
N:C ratio:seed treatment 5
4
3
2
soil moisture:seed treatment 5
4
3
2
light:seed treatment 5
4
3
2
extinction rate:seed treatment 5
4
3
2
2003 richness:seed treatment 5
4
3
2
seed treatment 5
4
3
2
N:C ratio
soil moisture
light
extinction rate
2003 richness
(intercept)
Figure 7
Coefficient estimates from fitted hierarchical regression models of below treeline 2009
Hieracium lepidulum density per 30 x 30 cm plot, with 2003 plot species richness,
extinction rate, available light, soil moisture, N:C ratio and their interactions with seed
addition treatment as fixed effects (n = 529). Significant results (p < 0.05) indicated with a
bold line and solid circle.
34
100
coefficient estimate
50
0
-50
-100
N:C ratio:seed treatment 5
4
3
2
soil moisture:seed treatment 5
4
3
2
light:seed treatment 5
4
3
2
extinction rate:seed treatment 5
4
3
2
2003 richness:seed treatment 5
4
3
2
seed treatment 5
4
3
2
N:C ratio
soil moisture
light
extinction rate
2003 richness
(intercept)
Figure 8
Coefficient estimates from fitted hierarchical regression models of above treeline 2009
Hieracium lepidulum density per 30 x 30 cm plot, with 2003 plot species richness,
extinction rate, available light, soil moisture, N:C ratio and their interactions with seed
addition treatment as fixed effects (n = 529). Significant results (p < 0.05) indicated with a
bold line and solid circle.
35
4.
Discussion
Results showed marked differences in variables influencing invasion success above and
below treeline, and significant interactions between these variables and propagule
pressure. Above treeline, plots with higher turnover (as measured by extinction rates) and
species richness were more invaded. It seems here that plots with higher rates of species
turnover provided more opportunities for H. lepidulum (along with others species in the
regional pool) to establish; H. lepidulum may act in a similar way to existing community
members, and hitch a ride on the ‘carousel’ of species turnover (VanderMaarel and Sykes
1993). An alternative is that local extinctions occurred as a result of disturbance, herbivory,
or disease; factors cited in the nutrient enrichment hypothesis, which suggests that
resultant pulses of resources provide windows of opportunity for invaders to establish
(Davis et al. 2000; Davis and Pelsor 2001). Disturbance by spring avalanches, human
activities or herbivory by introduced hares may have destroyed all individuals of a species
within a plot, and the resulting increase in resources (due to higher availability and reduced
consumption) may have facilitated invasion, although this remains speculative.
Expectations of the resource enrichment hypothesis were more convincingly matched
below treeline. Here invasion success was primarily controlled by light availability and
propagule pressure, echoing the findings of Wiser et al. (1998). Higher light levels below
treeline occurred in plots situated in treefall gaps and in gallery forest along stream
margins. Such treefall gaps are a form of disturbance; stream margins are also subject to
relatively frequent treefalls due to the effects of intermittent flooding and steep banks. The
resultant increase in available light can be thought of as a resource pulse, albeit on a
relatively long temporal scale.
Perhaps the clearest result from this study is the overriding role of propagule supply in
determining invasion success. Seed addition treatment showed significant interactions with
species richness, extinction rate and resource availability (although in overall models light
was the only resource significant in explaining H. lepidulum density after controlling for
other variables). Of particular interest, these interactions were only significant at higher
propagule pressures. Thus it appears that at low propagule pressures, species richness,
species turnover and resource availability had no significant influence on H. lepidulum
invasion; while above a certain threshold propagule supply, H. lepidulum could overcome
barriers to establishment and preferentially invade plots with higher turnover and richness
(above treeline), disturbance meditated resource pulses (below treeline).
Below this
threshold it appears the effects of stochastic processes may dominate and relatively few
36
plots are invaded, while above the threshold there are sufficient propagules that abiotic and
biotic site effects become apparent. In this context, it is worth noting that Wiser et al.
(1998) reported propagule pressure to increase in the forest interior as the invasion in their
study progressed, and as a result switched from being dispersal limited to resource (light)
limited.
The diversity-resistance hypothesis was not supported by these results; instead of the
predicted negative association between native richness and invader density, I found
positive or non-significant relationships: species rich plots were in general more invasible.
Evidence in the literature for the diversity-resistance hypothesis is mixed, and the
apparently contradictory results have been the source of considerable debate (Levine and
D'Antonio 1999; Wardle 1999; Naeem et al. 2000; Kennedy et al. 2002; Fridley et al.
2007). In general, observational studies have not supported the hypothesis, showing a
positive relationship between native richness and invasibility (PlantyTabacchi et al. 1996;
Sax 2002; Stohlgren et al. 2003; Stohlgren et al. 2006), while experimental studies in the
main have supported the hypothesis (Tilman 1997; Levine 2000; Kennedy et al. 2002).
This apparent conflict is most likely due to the scale dependence of the relationship (Shea
and Chesson 2002; Davies et al. 2005). At fine scales (< 1 m2, in the range of plot sizes
typically used in experimental studies) competitive effects are expected to result in a
negative relationship between the two; at coarser scales often used in observational studies
competitive effects are less apparent, while both native and exotic richness are likely to
covary with resource availability, disturbance and spatial heterogeneity (Drake et al. 1989).
Importantly, this study was experimental, and conducted at a relatively fine (30 x 30 cm)
spatial resolution. At this scale the effects of interspecific competition should be apparent
and those of spatial heterogeneity should be minimal; we would thus expect to see any
putative effect whereby community richness confers resistance to invasions. Two
experimental addition studies have found similar positive relationships (Peart and Foin
1985; Robinson et al. 1995), but in both cases the result was attributed to the presence of a
heavily dominant species which suppressed both native and exotic species, resulting in
low richness plots being relatively uninvaded. This does not appear to be the case here as
relatively few plots were dominated by one species in such manner.
How else could such a positive relationship come about? Native and exotic species may
well respond to the same abiotic variables (Huston 2004), and thus sites with high native
richness are also likely to be favourable sites for H. lepidulum. Alternatively, biotic factors
may play a role, namely via facilitation. In high stress environments such as those found
37
near treeline (Richardson and Friedland 2009), relationships between species are more
likely to be facilitative than competitive (Callaway et al. 2002; Brooker et al. 2008) - for
example sites with vegetation are likely to provide more shelter and better moisture
retention than bare ground – and thus species rich sites may favour, rather than inhibit, H.
lepidulum establishment.
Given this lack of support for the diversity-resistance hypothesis, what could be the
explanation in terms of the underlying theory? The hypothesis has its origins in the work
of Elton (1958) and MacArthur (MacArthur 1955; 1972). Elton suggested that more
diverse communities are inherently more stable in the face of perturbations from outside
influences, but there is some debate whether this implies less susceptibility to invasion or
impact (Levine and D'Antonio 1999). Similarly, MacArthur (1955) hypothesised that as
more diverse communities have more links between species, breaking one of these will
have a smaller effect on community stability than in relatively in species poor
communities, thus rendering the more diverse community more stable. However, from the
standpoint of the carousel model of turnover, what is meant by ‘stability’ and how this may
prevent invasive species entering a community seems less clear. The diversity-resistance
hypothesis has also been criticised on the grounds that there may be a confounding
sampling effect: more diverse communities are more likely to contain a native species
which can strongly compete with and resist the invader, rather than diversity per se
conferring resistance (Wardle 2001).
The results presented here illustrate the complex nature of invasions, and that several
different factors may be responsible for the success of any given invasion. Here, turnover,
resource pulses and propagule pressure all play a role, with different processes operating
above and below treeline. However, the effects of turnover and resource pulses were only
apparent when propagule pressure was sufficiently high, and below this threshold all sites
appeared equally invasible. As such it appears propagule pressure may be the overriding
determinant of invasion success, allowing invaders to take advantage of natural
community turnover and fluctuating resource levels.
The existence of a threshold propagule pressure above which H. lepidulum can invade
disturbed habitats more successfully is significant from a conservation perspective, in that
weed control aimed at limiting propagule pressure may be an effective and efficient way to
limit spread and establishment in new areas, in particular forest interiors. Controls on
human activity to limit disturbance and propagule supply may also benefit any
management programme.
38
Further experimental studies are required, and could fruitfully investigate two factors
thought to influence invasion success, not incorporated here. I considered community
diversity to be species richness, however
other measures are possible, and there is
evidence (Levine and D'Antonio 1999) that trophic diversity is a more relevant measure of
a community’s ability to effectively sequester resources and thus resist invasion. Invader
identity is also key, yet there are surprisingly few studies which specifically control for
differences between invaders. Such studies would be invaluable in furthering progress
towards been able to accurately predict invasion success across a range of communities
and invaders.
39
Chapter 3: No significant impacts on plant
community composition or structure
Abstract
There is debate over whether invasive plants necessarily impact the communities they
invade, or can coexist without discernable impacts. I followed the fate of alpine grassland,
scrub and montane forest communities in response to the experimental sowing of the alien
weed Hieracium lepidulum, looking for changes in community composition and structure
over six years. I used a replicated randomised block design, with 30 x 30 cm plots (n=756)
subdivided into 5 x 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 0 - 15625 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. Plot species richness increased from
2003 to 2009; furthermore, some of this increase (+0.002 species/year) was associated
with H. lepidulum density. Other relationships between the plant community and H.
lepidulum were generally non-significant. It appears H. lepidulum coexists with the
community in the study site, and has had no negative effects on richness, evenness or
diversity, even where density and cover of the alien are highest.
1.
Introduction
In chapter one I found that after experimental introduction, H. lepidulum preferentially
established in species rich and high turnover sites above treeline, and sites with canopy
disturbance below treeline. In this chapter I investigate the impacts on resident plant
community composition and structure that may result.
It is well known that some introduced plant species can impact native plant communities,
considerably altering their composition and structure, causing declines in abundance and
local extinctions of native plant species (e.g. Pysek and Pysek 1995; Vila et al. 2006;
Hejda et al. 2009) and modifying ecosystem structure and function (e.g. Vitousek and
Walker 1989; D'Antonio and Vitousek 1992; Ogle et al. 2003). While these problems
warrant considerable concern, evidence suggests such impacts due to alien plant species
may be the exception rather than the rule (Williamson 1996; Levine et al. 2003; Sax and
40
Gaines 2008), and it appears many invasive weeds may coexist as additional community
members without significant impact (Stohlgren et al. 2006). Resolution of these apparently
conflicting results can be found by considering the scale-dependency of alien plant
impacts; negative interactions between alien and native species generally operate at fine
spatial scales (< 1m2), while at coarser scales heterogeneity may mean native plant
diversity and alien abundance covary with favourable environmental factors (e.g. resource
availability, disturbance regime) leading to positive associations between the two (Levine
and D'Antonio 1999). In addition, there is still much debate over the relative importance of
identity and traits of the alien invader and how these interact with a given invaded
community.
Indeed, the frequency and extent to which alien plants impact native
communities are at present unclear due to a lack of studies representative of the range of
alien plant life-forms, invaded native communities and spatial scales examined (Mack
1996; Byers et al. 2002; Mills et al. 2009). More case studies and a move towards
generality are needed if we are to understand the nature and scale of the problem invasive
weeds pose, and target limited management resources effectively (Parker et al. 1999;
Levine et al. 2003).
Several difficulties arise in trying to quantify invader impacts. The majority of studies
take an observational approach, comparing native community diversity and composition
between already invaded and uninvaded habitats (Levine et al. 2003). However, such a
correlative approach reveals little about the causality of any differences detected, since
species poor communities may be more readily invaded (Levine and D'Antonio 1999), and
both native and alien species richness may covary with similar environmental drivers
(Hulme 2006). One alternative is experimentally removing the invader from plots, and
examining the response of the invaded community to its absence (e.g. Ogle et al. 2003;
Hulme and Bremner 2006; Stinson et al. 2007; Truscott et al. 2008). While overcoming the
confounding effects faced by observational studies, two further difficulties arise (Diaz et
al. 2003). First, the act of removal itself, and the associated disturbance may obscure any
invader effects, particularly in communities not naturally subject to high levels of
disturbance. Second, the true impact of the invader may not be measured, as the native
community which establishes following release from the invader may differ from the preinvasion community.
A third alternative is to experimentally introduce the invader into the native community
and compare its response between invaded and uninvaded control plots. While care needs
to be taken in selecting appropriate study species and sites to avoid potential ethical issues
41
involved (Truscott et al. 2008; Vila et al. 2008), such introductions have three advantages
over other methodologies. First, they allow questions of causality to be investigated due to
their experimental nature; second, they keep anthropogenic disturbance to a minimum; and
third, allow direct measurement of invader impacts.
Another consideration is that of temporal scale. When the state of a plant community is
perturbed by an invader, it may be some time before it reaches a new stable state or returns
to the original one; extinctions or declines in abundance of native plant species may occur
over timescales of years or even decades (Gaertner et al. 2009), and studies of invasive
plant impacts need to span similar time periods to properly assess the long term impact of
an invader.
My aim is to examine the long-term impacts of an invasive alien plant on native
community composition and structure, using intentional introductions to establish
experimental gradients of invader density and abundance. I use the alien plant Hieracium
lepidulum 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 (Rose and Frampton 1999; Radford et al. 2007).
I hypothesise that H. lepidulum primarily impacts the native plant community through
competitive displacement (Rose and Frampton 1999; Moen and Meurk 2001; Rose et al.
2004; Radford et al. 2007; Rose and Frampton 2007; Radford et al. 2009), and
consequently 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 communities. I therefore examined the impact of H. lepidulum invasion by
following changes in these three key traits of the native plant community at two spatial
scales over a period of six years. By establishing an experimental gradient of alien plant
densities and assessing impacts over six years and at two spatial scales I undertake one of
the most through assessments of alien plant impacts to date.
2.
Methods
2.1. Study site
The study was conducted in the Craigieburn Stream catchment, Craigieburn Forest Park,
Canterbury, New Zealand (43o10’S, 172°45'E). The terrain is mountainous, with elevations
42
ranging from 800m to 2000m. The dominant vegetation is mountain beech forest
(Nothofagus solandri var. cliffortiodes) to an elevation of approximately 1400m, above
which it gives way to subalpine scrub, tussock and alpine herbfields. The mean annual
temperature is 8.2°C, and mean annual precipitation is 1533 mm (Miller 2006).
2.2. Experimental setup
This study used the same experimental introduction experiment described in chapter two,
established by Miller (2006).
Hieracium lepidulum seeds were experimentally added at different densities to plots
located in six habitat types: forest creek, intact forest, forest canopy gap (formed by
treefall), alpine creek, tussock grassland and subalpine scrub. A randomised block design
was used, with each habitat having six replicate blocks. 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, procedural control with no seed
addition but with wetting, and seed addition at rates of 25, 125, 625, 3125 and 15625 seeds
per 30 x 30 cm plot with wetting (or 278, 1389, 6944, 34722, and 173611 seeds/m2).
2.3. Data collection
2.3.1. Plot abiotic characteristics
Data on environmental conditions in each block were collected prior to seed addition. For
each block, altitude and aspect were recorded, soil depth in the centre was measured with a
metal probe, and soil samples were collected from the corners using a 10 cm corer and
then bulked. Five grams wet weight of soil from each block was dried and re-weighed to
calculate gravimetric soil moisture content. The remaining soil was air dried, sieved and
analysed for pH, Olsen-soluble phosphorous, carbon, nitrogen and potassium. Light
measurements were taken for each plot as described in chapter two.
2.3.2. Quantifying H. lepidulum abundance and native community structure
Data were collected in two phases; H. lepidulum density and community richness from
2003 to 2009 (see chapter two for details), plus additional measurements made only in
2009 which provided the basis to calculate indices of richness, diversity and evenness at
the plot and cell scales (see below). I adopted this approach to take advantage of preexisting measurements taken in the course of associated experiments.
43
In 2009 additional measurements were made to characterise invader impacts. To better
quantify the impact of H. lepidulum at fine spatial scales, where I hypothesise competitive
effects to be most apparent, each plot was divided into a regular grid of 36 5 cm x 5 cm
cells. In each cell I 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. I excluded these 14 plots from the analyses.
I calculated three indices of community attributes expected to show impacts due to H.
lepidulum; richness, Shannon diversity and Shannon evenness (Magurran 2003;
Spellerberg and Fedor 2003). H. lepidulum is hypothesised to compete with members of
the plant community for resources, leading to reduced abundance or local extinction of
some. If so, local extinction will result in lower richness and diversity indices, while
reduced abundance of some species will result in lower diversity and evenness indices.
Shannon Diversity was utilised due to its relative sensitivity to rare species, which may be
more likely to manifest impacts. All three indices were calculated at both the plot and cell
scales in 2009. Richness is the number of species present per plot or cell excluding H.
lepidulum. Shannon diversity ( ) and evenness ( ) were calculated as follows:
Where S is species richness and
is the proportional abundance (cover) of the th species.
can take any positive value, with higher values indicating greater richness and/or more
evenly spread abundances among species. Values of
range between 0 and 1; where lower
values indicate dominance by one or a few species and higher values indicate an
increasingly even spread of abundances among species.
2.4. Analysis
The analysis was conducted in three parts: a longitudinal analysis of changes in plot
species richness over six years as a function of H. lepidulum density, and two analyses that
examined variation in native community richness, diversity and evenness in relation to H.
lepidulum density and cover in 2009; one at the plot and the other at the cell scale. I used
44
the statistical software R (R_Development_Core_Team 2009) to formulate hierarchical
mixed-effect models using the package lme4 (Bates and Maechler 2009).
When fitting a linear mixed effect model two sets of assumptions need to be satisfied.
Firstly the within group errors must be independent, normally distributed about zero with
constant variance, and be independent of the random effects. Secondly, the random effects
must be normally distributed with a constant specified covariance matrix, and must be
independent among groups. To validate these assumptions I followed the methods in
Pinheiro and Bates (2000); firstly I checked the distribution of within group residuals, and
plots of the fitted versus observed values, then examined normal plots of the random
effects and scatter plots of the random effects. Assumptions were satisfactorily met.
2.4.1. Longitudinal analysis
I first examined how plot species richness changed through time by fitting linear trends to
the data using a hierarchical regression model (Gelman and 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 I included plot nested within block as random effects to account for this nonindependence.
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, I 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, I included soil characteristics (carbon,
nitrogen, phosphorous and potassium levels, pH, soil moisture and soil depth) and physical
attributes (plot altitude, aspect and available light) as additional 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. I fitted a second model with
an additional 2009 H. lepidulum density by habitat interaction. I selected the better model
using Akaike Information Criterion (AIC), a measure of the explanatory power of a model
45
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).
2.4.2. Plot scale
I obtained a more detailed picture of the effects of H. lepidulum on community structure
and composition by investigating the relationships between plot richness, diversity and
evenness and H. lepidulum density and cover in 2009. For each response variable (2009
plot richness, evenness and diversity) I fitted two hierarchical linear regression models:
one with H. lepidulum density and the other with H. lepidulum cover as fixed effects,
including block as a random effect to account for plots being nested within blocks. The
influence of habitat on these relationships was then tested by fitting a second set of six
regressions, this time including habitat and its interactions with H. lepidulum density and
cover as additional fixed effects. To test coefficient estimates for significance, I 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 I applied a Holm-BonFerroni correction.
Finally, competitive interactions are expected to be stronger within functionally similar
sets of species (Ortega and Pearson 2005). To test this, I 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 functionally similar H. lepidulum.
46
2.4.3. Cell scale
To test whether stronger competitive interactions led to greater impacts at fine spatial
scales, I repeated the plot scale analyses at the cell scale. Models were the same as those
fitted at the plot scale, but with plot nested within block as random effects to account for
cells being nested within plots within blocks. Response variables were cell richness,
diversity and evenness. Fixed effects were cell H. lepidulum density and cover; then cell
H. lepidulum density and cover, habitat and their interaction. Coefficient estimates were
tested for significance using MCMC sampling. Analyses were again conducted with the
data partitioned by functional group. Holm-BonFerroni corrections were applied.
3.
Results
3.1. Hieracium lepidulum
Peak H. lepidulum densities occurred in December 2003, in the summer immediately
following sowing, when plots contained from 0 to 4435 individuals. Average densities then
declined until around 2007, where they appeared to stabilise (Fig. 9). The seed addition
treatments created a gradient of H. lepidulum density across the plots, which was
maintained for the duration of the study (Figs. 9 & 10). In 2009, the density and cover of
H. lepidulum in plots were reasonably well correlated (Pearson’s product moment
correlation = 0.52, p < 0.01) and a there was a similar H. lepidulum cover gradient (Fig.
10). Density ranged from 0 to 383 individuals per plot, with an overall mean of 12.65 ± 1.5
in all treatment plots and 0.34 ± 0.19 in all control plots (Table 1). 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.
3.2. 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 treeline (Fig. 11). Two models incorporating the influence of H. lepidulum were
considered; the first, with time, 2009 H. lepidulum density, their interaction, habitat and
covariates (AIC = 11316) was selected over the second, which had an additional
interaction term between H. lepidulum density and habitat (AIC = 11320). The selected
model indicated an overall mean increase of 0.2 species per plot per year over the duration
of the study (Table 2). Furthermore, there was a highly significant H. lepidulum density by
47
time interaction. This was also positive, implying that plots with more H. lepidulum
showed a greater increase in species richness through time, the opposite of what we would
expect if invasion by H. lepidulum resulted in competitive exclusion. Plots above treeline
had significantly more species than those below (Table 2, see also Fig. 11). Lower
elevation plots and those with higher available light had more species on average, but other
10000
potential covariates were non-significant.
Seeds Sown per Plot:
100
10
0
1
# Hieracium Individulals per Plot
1000
15625
3125
625
125
25
2003
2004
2005
2006
2007
2008
2009
year
Figure 9
Mean density (number of individuals per 30 x 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)
48
100
10
0
Frequency
0
100
200
300
400
100
10
0
Frequency
2009 Hieracium Density
0
10
20
30
40
50
60
2009 % Hieracium Cover
Figure 10
Frequency histograms of Hieracium lepidulum density (number of individuals per 30 x 30
cm plot) and percent cover (total foliar cover per plot) in 2009 across all seed addition and
control plots.
49
Table 1
Mean (± s.e) species richness, diversity, evenness, and Hieracium lepidulum density and %
foliar cover per 30 x 30 cm plot for each of six habitat types in 2009. Table 1
Treatment plots had five densities (25, 125, 625, 3125 and 15625 seeds/plot) of H.
lepidulum seeds added in 2003 (n = 532), control plots had no seed added (n = 215).
Treatment plots
Habitat
Richness
Diversity
Evenness
H. lepidulum H. lepidulum
density
cover
Forest Creek
4.34 ± 0.44 1.08 ± 0.09 0.66 ± 0.04 5.77 ± 1.26
1.83 ± 0.37
Forest
1.41 ± 0.09 0.39 ± 0.04 0.51 ± 0.05 0.18 ± 0.14
0.03 ± 0.04
Forest Gap
2.60 ± 0.17 0.78 ± 0.05 0.69 ± 0.04 1.99 ± 0.65
1.45 ± 0.51
Alpine Creek 11.31 ± 0.45 2.20 ± 0.05 0.86 ± 0.01 33.08 ± 6.56
2.04 ± 0.39
Scrub
4.65 ± 0.24 1.34 ± 0.05 0.80 ± 0.01 10.09 ± 2.32
3.99 ± 0.77
Tussock
7.70 ± 0.42 1.78 ± 0.06 0.83 ± 0.01 24.73 ± 4.85
1.96 ± 0.34
Overall
5.34 ± 0.17 1.27 ± 0.03 0.72 ± 0.01
12.65 ± 1.50
1.39 ± 0.17
Richness
H. lepidulum H. lepidulum
Control plots
Habitat
Diversity
Evenness
density
cover
Forest Creek
4.23 ± 0.64 1.11 ± 0.09 0.67 ± 0.06 1.31 ± 1.1
0.36 ± 0.19
Forest
1.47 ± 0.13 0.38 ± 0.04 0.54 ± 0.08 0
0
Forest Gap
3.22 ± 0.29 0.92 ± 0.05 0.62 ± 0.07 0
0
Alpine Creek 12.16± 0.78 2.25 ± 0.05 0.85 ± 0.01 0
0
Scrub
5.11 ± 0.41 1.49 ± 0.05 0.73 ± 0.04 0.44 ± 0.18
0.14 ± 0.08
Tussock
7.50 ± 0.67 1.73 ± 0.06 0.86 ± 0.01 0.31 ± 0.18
0.03 ± 0.02
Overall
5.26 ± 0.31 1.25 ± 0.06 0.71 ± 0.02 0.24 ± 0.19
0.09 ± 0.04
50
14
10
8
6
4
0
2
Plot Species Richness
12
Subalpine Creek
Tussock
Overall
Scrub
2003
2004
2005
2006
2007
2008
2009
14
Year
10
8
6
4
0
2
Plot Species Richness
12
Overall
Forest Creek
Forest Gap
Forest
2003
2004
2005
2006
2007
2008
2009
Year
Figure 11
Change in mean species richness per 30 x 30cm plot (circles with 95% confidence
intervals, n = 742), overall and in habitats above (top) and below treeline (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.
51
Table 2
Coefficient estimates (± s.e.) 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 s.e.
F
p
(Intercept)
-3.988
8.801
Time
0.208
0.012
412.164 0.000
2009 H. lepidulum Density
-0.005
0.002
0.003
0.957
Year : 2009 H. lepidulum Density
0.002
0.000
31.935
0.000
10.636
0.000
Habitat
Forest Creek
0
-
Forest
-1.571
1.986
Forest Gap
-0.119
2.168
Subalpine
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
Phosphorous
-0.045
0.142
0.108
0.742
Potassium
0.444
0.891
0.220
0.639
Creek
52
3.3. Plot scale
Overall plot species richness was positively related with H. lepidulum density and cover in
2009 (Fig. 12). While these relationships were not significant, they are the opposite of
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, to allow for different trends in each
habitat, relationships with H. lepidulum density and cover were mostly non-significant and
weakly positive (Fig. 13); similar to those found at the cell scale (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.
3.4. Cell scale
Overall the relationship between species richness and H. lepidulum density was positive at
the cell scale (Fig. 14), and the relationship with cover of the invader was significantly
positive (Figs. 14 & 15). 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 were non-significant, although evenness was significantly
greater in plots with higher H. lepidulum cover in the tussock habitat.
3.5. 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 (Fig. 16). 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 disproportionately 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. 17). By contrast, functional group richness showed
multiple 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 significantly positively associated
with basal herb, creeping herb and grass functional group richness. There were significant
negative associations with tussock and shrub functional group richness.
53
25
0
5
10
15
20
25
20
15
10
0
5
Plot Species Richness
0
10
100
383
0
# Hieracium Individuals per Plot
10
100
Plot % Hieracium Cover
Figure 12
Hierarchical regression models of species richness, with 2009 Hieracium lepidulum
density and cover as the fixed effects at the 30 x 30 cm plot scale (n = 747). Data points
have been jittered for clarity. Note the log scales on the x axes.
54
0.6
Coefficient Estimate: Cover
0.4
0.2
0.0
-0.2
-0.4
-0.6
Tussock
Scrub
Subalpine Creek
Forest Gap
Forest
Forest Creek
Overall
Coefficient Estimate: Density
0.10
0.05
0.00
Tussock
Scrub
Subalpine Creek
Forest Gap
Forest
Forest Creek
Overall
-0.05
Tussock
Scrub
Subalpine Creek
Forest Gap
Forest
Forest Creek
Overall
-0.10
E venness
D iv e rs ity
R ic h n e s s
Figure 13
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 plot (30 x 30 cm) scale (n = 742). Significant results (p < 0.05 after HolmBonferroni correction) indicated in bold with solid circles.
55
12
0
2
4
6
8
10
12
10
8
6
4
Cell Species Richness
2
0
1
10
100
0
# Hieracium Individuals per Cell
10
100
Cell % Hieracium Cover
Figure 14
Hierarchical regression models of species richness, with 2009 Hieracium lepidulum
density and cover as the fixed effects at the 5 x 5 cm cell scale (n = 26,892). Data points
have been jittered for clarity. Note the log scales on the x axes.
56
Figure 15
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 x 5 cm) scale (n = 26,892). Significant results (p < 0.05 after
Holm-Bonferroni correction) indicated in bold with solid circles.
57
Tussock
Prostrate Shrub
Other
Mat/Cushion
Grass
Erect Leafy
Shrub
Creeping Herb
Clumped Herb
Basal Herb
Tussock
Prostrate Shrub
Other
Mat/Cushion
Grass
Erect Leafy
Shrub
Creeping Herb
Clumped Herb
Basal Herb
Tussock
Prostrate Shrub
Other
Mat/Cushion
Grass
Erect Leafy
Shrub
Creeping Herb
Clumped Herb
Basal Herb
E ve n e s s
D iv e rsity
0.000
0.002
0.004
coefficient estimate: density
-0.006 -0.004 -0.002
R ich n e s s
0.006
-0.10
0.00
0.05
coefficient estimate: cover
-0.05
0.10
coefficient estimate: cover
0.15
0.10
0.05
0.00
-0.05
-0.10
-0.15
Tussock
Scrub
Subalpine Creek
Forest Gap
Forest
Forest Creek
Overall
coefficient estimate: density
0.15
0.10
0.05
0.00
Tussock
Scrub
Subalpine Creek
Forest Gap
Forest
Forest Creek
Overall
-0.05
Tussock
Scrub
Subalpine Creek
Forest Gap
Forest
Forest Creek
Overall
-0.15 -0.10
E venness
D iv e rsity
R ic h n e ss
Figure 16
Coefficient estimates for hierarchical regression models of plant community richness,
diversity and evenness in response to Hieracium lepidulum density and cover at the plot
(30 x 30 cm) scale (n = 747) and partitioned by functional group. Significant results (p <
0.05 after Holm-Bonferroni correction) indicated in bold with solid circles.
58
Figure 17
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
x 5 cm) scale (n = 26,892) and partitioned by functional group. Significant results (p <
0.05 after Holm-Bonferroni correction) indicated in bold with solid circles.
59
Tussock
Prostrate Shrub
Other
Mat/Cushion
Grass
Erect Leafy
Shrub
Creeping Herb
Clumped Herb
Basal Herb
Tussock
Prostrate Shrub
Other
Mat/Cushion
Grass
Erect Leafy
Shrub
Creeping Herb
Clumped Herb
Basal Herb
Tussock
Prostrate Shrub
Other
Mat/Cushion
Grass
Erect Leafy
Shrub
Creeping Herb
Clumped Herb
Basal Herb
E venness
D iv e rs ity
-0.015
R ic h n e ss
coefficient estimate: density
-0.005 0.000 0.005 0.010 0.015
-0.004
0.000
0.002
coefficient estimate: cover
-0.002
0.004
4.
Discussion
It appears H. lepidulum has not had any significant impacts on the resident plant
community, even after six years across a wide range of invader densities. Instead, it seems
to have slotted into the community where it co-exists as an additional member. Plot species
richness increased over the six years of our study, and this increase was significantly
greater in plots with higher H. lepidulum densities. Furthermore, in 2009 plant community
richness, evenness and diversity showed no significant declines at the plot or cell scales in
response to H. lepidulum density or cover.
The idea that invaders may often be accepted into the resident community without
discernable impact is not new (Williamson and Fitter 1996), but has seen a recent revival
of interest with the biotic acceptance theory of Stohlgren et al. (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 a few high profile invasive species with clear impacts
dominate research, while the majority may well have negligible effects on invaded
communities. However, the coarser spatial scales at which these studies have been
conducted may mean such results reflect covariation in alien and native plant richness
across environmental suitability gradients, rather than the outcomes of interaction between
the invader and resident community. My results showed a positive relationship between
invader abundance and native richness, indicating no impact, even at a scale of 5cm x 5cm.
At such a fine spatial scale we certainly expect to see the outcomes of any interspecies
interactions.
The observed lack of impacts on native plant richness, diversity and evenness is surprising
given concerns over the species. How has H. lepidulum entered the system without
discernable impact? H. lepidulum appears not to be competitively superior to native
species (Radford et al. 2006); it is smaller statured than community dominants in all
habitats, and observations over six years indicate that it does not grow more rapidly than
native species. This suggests two mechanisms by which it could establish without impact.
First, these communities may be following a succession trajectory and not be saturated. If
so, opportunities exist for additional species to enter, especially in plots where the
environment is conducive to growth due to unutilised resources such as moisture, nutrients
or light. Species richness would increase most in plots with conditions conducive to plant
growth, and it might be expected that H. lepidulum would also perform better in those
plots, persisting preferentially due to the same favourable conditions. Conversely, after
60
initial establishment at high densities due to propagule pressure, H. lepidulum numbers
may then have dropped disproportionately in those plots with less favourable conditions.
Second, its ability to invade may be more a reflection of its opportunistic nature: H.
lepidulum may adventitiously colonise plots where turnover or a favourable disturbance
regime intermittently frees up sites and resources.
Evidence for the former explanation can be found in the analyses of functional groups.
Due to increased competition between similar life forms, I expected groups functionally
similar to H. lepidulum, such as basal herb, creeping herb and erect leafy, to be
disproportionately impacted. Contrary to this, they have positive parameter estimates,
suggesting some plots are favourable sites for both these groups and H. lepidulum.
Evidence for the latter can be found in chapter two of this thesis.
Hieracium lepidulum may not have been able to enter all plots with equal ease. 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. The negative parameter estimates for these groups suggest they resist invasion.
The clumped herb group consists primarily of vegetatively spreading Celmisia species,
which form dense clumps of vegetation. In the case of the mat/cushion group, the dense
ground cover exhibited by these species may well restrict the availability of sites suitable
for H. lepidulum to establish
Nevertheless, previous studies have reported apparent impacts; H. lepidulum forms dense
mats to the exclusion of native species in some localities (Rose and Frampton 1999), and
other members of the Hieracium genus appear to impact montane plant communities
(Scott et al. 1990). These results are not mutually exclusive with those presented here. It is
possible that under certain conditions H. lepidulum may reach higher densities/cover than I
was able to experimentally establish, and may have localised impacts at sites where this
occurs. However, given the randomised block design I used incorporating the major habitat
types found in the area, my results should describe the average impact of H. lepidulum
across these communities.
My results show that after six years, invasion by H. lepidulum has no discernable negative
impacts on key plant community attributes across habitats, spatial 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
61
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.
62
Chapter 4: General discussion
The aim of this thesis was to investigate factors allowing Hieracium lepidulum to
successfully invade montane forest, mixed subalpine tussock grasslands and herbfields in
the Craigieburn Catchment, Craigieburn Forest Park, New Zealand; and to determine the
nature and magnitude of any impacts this invasion has had on native plant communities.
Hieracium lepidulum seeds were added to plots at different densities, and the fate of the
invader and native community tracked over six years. Measurements were made at two
spatial scales; this feature, in combination with the long term, experimental approach,
makes this one of the most thorough investigations of plant invasion and subsequent
impact made to date.
1.
Key results and implications
Above treeline, plots with higher turnover and species richness were more invaded. It
seems H. lepidulum may act in a similar way to existing community members, and hitch a
ride on the ‘carousel’ of species turnover (VanderMaarel and Sykes 1993). Below treeline,
invasion success was primarily controlled by light availability and propagule pressure.
Increased light availability was associated with disturbance in the form of treefall gaps and
along stream margins. This increased light can be interpreted as a long term resource pulse,
thus providing support for the nutrient enrichment hypothesis (Davis et al. 2000; Davis and
Pelsor 2001).
Propagule supply was central in determining invasion success; the variables discussed
above were only significant at higher propagule pressures. Thus it appears that above a
certain threshold propagule supply, H. lepidulum could overcome stochastic effects to
preferentially invade plots with higher turnover (above treeline), disturbance meditated
resource pulses (below treeline) and higher richness.
Results did not support the diversity resistance hypothesis, but instead showed the opposite
of what it predicts - six years after introduction, sites with higher initial species richness
had higher densities of H. lepidulum. These sites were also the most resource rich. While
previous studies have both supported (e.g. Tilman 1997; Levine 2000; Kennedy et al.
2002) and contradicted (e.g. PlantyTabacchi et al. 1996; Lonsdale 1999; Sax 2002)the
diversity resistance hypothesis, the latter have generally been coarse scale, observational
studies. This study was experimental and conducted at a relatively fine spatial scale, and as
such this result contrary to the hypothesis represents a significant finding.
63
Despite H. lepidulum successfully invading and establishing at a wide range of densities, I
found no evidence of impacts on the native plant community. Species richness increased
over the duration of the study, significantly more so in plots with higher H. lepidulum
densities. Furthermore, H. lepidulum density and cover had positive relationships with
species richness, even at the cell scale where competitive effects are most likely to result in
impacts. Evenness and diversity mostly showed non-significant relationships with the
invader, and there were few differences between habitats.
How can this ability to successfully invade, without discernable impact be explained? It
seems that H. lepidulum has joined the community as an extra member, adventitiously
colonising sites with temporarily elevated resource availability, rather than by
outcompeting native species. Utilising the framework for understanding the interplay
between invasion and impact developed by MacDougall et al. (2009), building on the
ideas of Chesson (2000), we can visualise H. lepidulum as having a niche difference rather
than a fitness difference with the native community. In this framework, niche differences
offer an advantage only when a species is rare, resulting in negative feedback loops which
promote coexistence - for example a taller species will receive more light than its
neighbours, but this advantage will lessen as it becomes more common and in effect
competes with itself; on the other hand fitness differences confer an advantage regardless
of rarity, and therefore promote dominance – for example greater fecundity will always
benefit a species, or a species which promotes and tolerates fire will experience increasing
advantage over fire sensitive species as its numbers increase. Either a niche or a fitness
difference is sufficient for an invader to establish, but a fitness difference is required for
the invader to become dominant and impact the resident community. I suggest the niche
difference in this system may be the ability to colonise recently disturbed sites more
rapidly than native species - as H. lepidulum density increases in these sites, the advantage
diminishes and becomes self limiting. Conversely it seems H. lepidulum has no
competitive advantage that persists at high invader densities, and thus is unable to become
dominant and negatively impact the native community.
2.
Further research and recommendations
While H. lepidulum has successfully invaded the study area it has had no discernable
negative impacts on key plant community attributes. Despite this, the possibility remains
that particular ruderal native species with similar strategies may be impacted by H.
lepidulum invasion. Furthermore, there is no doubt that H. lepidulum alters the qualitative
64
nature of invaded habitats, and when flowering can have a considerable visual impact,
which may be of concern in conservation areas valued for their aesthetic as well as
ecological values. As such, further research into impacts on potentially vulnerable native
species along with management strategies may be warranted. Controls on human activity
to limit disturbance and propagule supply may benefit any management programme
These results show that to better understand invasion success and invader impacts we need
to study a representative range of invasions rather than only those where invasion and
impact are clearest. Future studies need to account for invader identity, community
composition, abiotic conditions, and propagule pressure; and be conducted at appropriate
spatial and temporal scales. Furthermore, the complex nature of invasions means several
factors may be important in explaining the success and impacts of a particular invader.
Experimental introductions represent a valuable and under-utilised tool for such studies,
providing opportunities for rigorous, quantitative investigations of plant invasions.
65
Acknowledgements
A big thank you to:
My supervisors, Richard Duncan and Phil Hulme, for their guidance, knowledge, insight
and enthusiasm.
Alice Miller, whose hard work and foresight in establishing the experimental plots at
Craigieburn made this thesis possible.
The Miss E.L. Hellaby Indigenous Grasslands Research Trust,
The New Zealand Plant Protection Society Dan Watkins Scholarship in Weed Science and
The Bio-Protection Research Centre writing scholarship
for research funding and financial support.
Robyn Damaryhoman for a huge effort in the field and entering endless data.
Laura Spence, Myles Mackintosh and others for valuable field assistance.
Yukari Izumo, Hamish Meffin and Elaine Meffin for being there, and much more.
66
Bibliography
Akaike, H. (1974). A New Look at the Statistical Model Identification. {IEEE}
Transactions on Automatic Control 19(6): 716-723.
Baayen, R. H. (2008). languageR: Data sets and functions with "Analyzing Linguistic
Data: A practical introduction to statistics".
Baayen, R. H., D. J. Davidson and D. M. Bates (2008). Mixed-effects modeling with
crossed random effects for subjects and items. Journal of Memory and Language
59(4): 390-412.
Bates, D. (2006). "lmer, p-values and all that." Retrieved 22-0ct-2009, 2009.
Bates, D. and M. Maechler (2009). lme4: Linear mixed-effects models using S4 classes.
Brooker, R. W., F. T. Maestre, R. M. Callaway, C. L. Lortie, L. A. Cavieres, G. Kunstler, P.
Liancourt, K. Tielborger, J. M. J. Travis, F. Anthelme, C. Armas, L. Coll, E.
Corcket, S. Delzon, E. Forey, Z. Kikvidze, J. Olofsson, F. Pugnaire, C. L. Quiroz,
P. Saccone, K. Schiffers, M. Seifan, B. Touzard and R. Michalet (2008).
Facilitation in plant communities: the past, the present, and the future. Journal of
Ecology 96(1): 18-34.
Burhenne-Guilmin, F., L. Glowka, A. F. Krattiger, J. A. McNeely, W. H. Lesser, K. R.
Miller, Y. St. Hill and R. Senanayake (1994). An introduction to the Convention on
Biological Diversity. Widening perspectives on biodiversity.: 15-18.
Burke, M. J. W. and J. P. Grime (1996). An experimental study of plant community
invasibility. Ecology 77(3): 776-790.
Byers, J. E., S. Reichard, J. M. Randall, I. M. Parker, C. S. Smith, W. M. Lonsdale, I. A. E.
Atkinson, T. R. Seastedt, M. Williamson, E. Chornesky and D. Hayes (2002).
Directing research to reduce the impacts of nonindigenous species. Conservation
Biology 16(3): 630-640.
Callaway, R. M., R. W. Brooker, P. Choler, Z. Kikvidze, C. J. Lortie, R. Michalet, L.
Paolini, F. I. Pugnaire, B. Newingham, E. T. Aschehoug, C. Armas, D. Kikodze and
B. J. Cook (2002). Positive interactions among alpine plants increase with stress.
Nature 417(6891): 844-848.
Cassey, P., T. M. Blackburn, S. Sol, R. P. Duncan and J. L. Lockwood (2004). Global
patterns of introduction effort and establishment success in birds. Proceedings of
the Royal Society of London Series B-Biological Sciences 271: S405-S408.
Catford, J. A., R. Jansson and C. Nilsson (2009). Reducing redundancy in invasion
ecology by integrating hypotheses into a single theoretical framework. Diversity
and Distributions 15(1): 22-40.
Chapman, H., B. Robson and M. L. Pearson (2004). Population genetic structure of a
colonising, triploid weed, Hieracium lepidulum. Heredity 92(3): 182-188.
Chesson, P. (2000). Mechanisms of maintenance of species diversity. Annual Review of
Ecology and Systematics 31: 343-+.
Colautti, R. I., I. A. Grigorovich and H. J. MacIsaac (2006). Propagule pressure: A null
model for biological invasions. Biological Invasions 8(5): 1023-1037.
Cornelissen, J. H. C., S. Lavorel, E. Garnier, S. Diaz, N. Buchmann, D. E. Gurvich, P. B.
Reich, H. ter Steege, H. D. Morgan, M. G. A. van der Heijden, J. G. Pausas and H.
Poorter (2003). A handbook of protocols for standardised and easy measurement of
plant functional traits worldwide. Australian Journal of Botany 51(4): 335-380.
67
D'Antonio, C. M. and P. M. Vitousek (1992). Biological invasions by exotic grasses, the
grass fire cycle, and global change. Annual Review of Ecology and Systematics 23:
63-87.
Dantonio, C. M. and P. M. Vitousek (1992). BIOLOGICAL INVASIONS BY EXOTIC
GRASSES, THE GRASS FIRE CYCLE, AND GLOBAL CHANGE. Annual
Review of Ecology and Systematics 23: 63-87.
Davies, K. F., P. Chesson, S. Harrison, B. D. Inouye, B. A. Melbourne and K. J. Rice
(2005). Spatial heterogeneity explains the scale dependence of the native-exotic
diversity relationship. Ecology 86(6): 1602-1610.
Davis, M. A., J. P. Grime and K. Thompson (2000). Fluctuating resources in plant
communities: a general theory of invasibility. Journal of Ecology 88(3): 528-534.
Davis, M. A. and M. Pelsor (2001). Experimental support for a resource-based mechanistic
model of invasibility. Ecology Letters 4(5): 421-428.
Diamond, J. M. (1989). The present, past and future of human-caused extinctions.
Philosophical Transactions of the Royal Society of London Series B-Biological
Sciences 325(1228): 469-477.
Diaz, S., A. J. Symstad, F. S. Chapin, D. A. Wardle and L. F. Huenneke (2003). Functional
diversity revealed by removal experiments. Trends in Ecology & Evolution 18(3):
140-146.
Drake, J. A., H. A. Mooney and F. DiCastri, Eds. (1989). Invasibilty of plant communities.
Biologcal invasions: a global perspective. Chichester, John Wiley.
Dukes, J. S. and H. A. Mooney (2004). Disruption of ecosystem processes in western
North America by invasive species. Revista Chilena De Historia Natural 77(3):
411-437.
Duncan, R. P., K. M. Colhoun and B. D. Foran (1997). The distribution and abundance of
Hieracium species (Hawkweeds) in the dry grasslands of Canterbury and Otago.
New Zealand Journal of Ecology 21: 51-62.
Duncan, R. P., R. J. Webster and C. A. Jensen (2001). Declining plant species richness in
the tussock grasslands of Canterbury and Otago, South Island, New Zealand. New
Zealand Journal of Ecology 25: 35-47.
Edward, E., P. K. T. Munishi and P. E. Hulme (2009). Relative Roles of Disturbance and
Propagule Pressure on the Invasion of Humid Tropical Forest by Cordia alliodora
(Boraginaceae) in Tanzania. Biotropica 41(2): 171-178.
Elton, C. S. (1958). The ecology of invasions by animals and plants. The ecology of
invasions by animals and plants.: 181 pp.
Eschtruth, A. K. and J. J. Battles (2009). Assessing the relative importance of disturbance,
herbivory, diversity, and propagule pressure in exotic plant invasion. Ecological
Monographs 79(2): 265-280.
Espie, P. (2001). Hieracium in New Zealand: Ecology and Management. Invermay,
AgResearch Ltd.
Fan, J. W. and W. Harris (1996). Effects of soil fertility level and cutting frequency on
interference among Hieracium pilosella, H-praealtum, Rumex acetosella, and
Festuca novae-zelandiae. New Zealand Journal of Agricultural Research 39(1): 132.
Fridley, J. D., J. J. Stachowicz, S. Naeem, D. F. Sax, E. W. Seabloom, M. D. Smith, T. J.
Stohlgren, D. Tilman and B. Von Holle (2007). The invasion paradox: Reconciling
pattern and process in species invasions. Ecology 88(1): 3-17.
Funk, J. L. and P. M. Vitousek (2007). Resource-use efficiency and plant invasion in lowresource systems. Nature 446(7139): 1079-1081.
68
Gaertner, M., A. Den Breeyen, C. Hui and D. M. Richardson (2009). Impacts of alien plant
invasions on species richness in Mediterranean-type ecosystems: a meta-analysis.
Progress in Physical Geography 33(3): 319-338.
Gelman, A. and J. Hill (2006). Data analysis using regression and multilevel/hierachical
models. Cambridge, Cambridge University Press.
Grice, A. C. (2006). The impacts of invasive plant species on the biodiversity of Australian
rangelands. Rangeland Journal 28(1): 27-35.
Hejda, M., P. Pysek and V. Jarosik (2009). Impact of invasive plants on the species
richness, diversity and composition of invaded communities. Journal of Ecology
97(3): 393-403.
Hobbs, R. J. and S. E. Humphries (1995). An integrated approach to the ecology and
management of plant invasions. Conservation Biology 9(4): 761-770.
Huenneke, L. F., S. P. Hamburg, R. Koide, H. A. Mooney and P. M. Vitousek (1990).
Effects of soil resources on plant invasion and community structure in Californian
serpentine grassland. Ecology 71(2): 478-491.
Hulme, P. E. (2006). Beyond control: wider implications for the management of biological
invasions. Journal of Applied Ecology 43(5): 835-847.
Hulme, P. E. and E. T. Bremner (2006). Assessing the impact of Impatiens glandulifera on
riparian habitats: partitioning diversity components following species removal.
Journal of Applied Ecology 43(1): 43-50.
Huston, M. A. (2004). Management strategies for plant invasions: manipulating
productivity, disturbance, and competition. Diversity and Distributions 10(3): 167178.
Keddy, P. A. and B. Shipley (1989). Competitive hierachies in herbaceous plant
communities. Oikos 54(2): 234-241.
Kennedy, T. A., S. Naeem, K. M. Howe, J. M. H. Knops, D. Tilman and P. Reich (2002).
Biodiversity as a barrier to ecological invasion. Nature 417(6889): 636-638.
Lamoureaux, S. L., D. Kelly and N. D. Barlow (2003). Population dynamics in mature
stands of Hieracium pilosella in New Zealand. Plant Ecology 166(2): 263-273.
Levine, J. M. (2000). Species diversity and biological invasions: Relating local process to
community pattern. Science 288(5467): 852-854.
Levine, J. M. and C. M. D'Antonio (1999). Elton revisited: a review of evidence linking
diversity and invasibility. Oikos 87(1): 15-26.
Levine, J. M., M. Vila, C. M. D'Antonio, J. S. Dukes, K. Grigulis and S. Lavorel (2003).
Mechanisms underlying the impacts of exotic plant invasions. Proceedings of the
Royal Society of London Series B-Biological Sciences 270(1517): 775-781.
Lonsdale, W. M. (1999). Global patterns of plant invasions and the concept of invasibility.
Ecology 80(5): 1522-1536.
MacArthur, R. (1955). Fluctuations of Animal Populations and a Measure of Community
Stability. Ecology 36(3): 533-536.
MacArthur, R. H. (1972). Geographical ecology; patterns in the distribution of species.
New York, Harper and Row.
MacDougall, A. S., B. Gilbert and J. M. Levine (2009). Plant invasions and the niche.
Journal of Ecology 97(4): 609-615.
Mack, R. N. (1996). Predicting the identity and fate of plant invaders: Emergent and
emerging approaches. Biological Conservation 78(1-2): 107-121.
Magurran, A. E. (2003). Measuring biological diversity. Measuring biological diversity:
viii + 256 pp.
Mark, A., G. Rogers and J. Barkla (1999). General observations on vegetation condition
and trends based on resampling permanent photographic points in Mt Aspiring
69
National Park, Feb-March, 1999. Dunedin, University of Otago/Department of
Conservation: 2.
Miller, A. L. (2006). Untangling spatial distribution patterns of the invasive herb
Hieracium lepidulum Stenstr. (Asteracea) in a New Zealand mountain landscape.
Bio-protection and Ecology. Christchurch, Lincoln University. PhD: 146.
Mills, J. E., J. A. Reinartz, G. A. Meyer and E. B. Young (2009). Exotic shrub invasion in
an undisturbed wetland has little community-level effect over a 15-year period.
Biological Invasions 11(8): 1803-1820.
Mittelbach, G. G., C. F. Steiner, S. M. Scheiner, K. L. Gross, H. L. Reynolds, R. B. Waide,
M. R. Willig, S. I. Dodson and L. Gough (2001). What is the observed relationship
between species richness and productivity? Ecology 82(9): 2381-2396.
Moen, J. and C. D. Meurk (2001). Competitive abilities of three indigenous New Zealand
plant species in relation to the introduced plant Hieracium pilosella. Basic and
Applied Ecology 2(3): 243-250.
Naeem, S., J. M. H. Knops, D. Tilman, K. M. Howe, T. Kennedy and S. Gale (2000). Plant
diversity increases resistance to invasion in the absence of covarying extrinsic
factors. Oikos 91(1): 97-108.
Ogle, S. M., W. A. Reiners and K. G. Gerow (2003). Impacts of exotic annual brome
grasses (Bromus spp.) on ecosystem properties of northern mixed grass prairie.
American Midland Naturalist 149(1): 46-58.
Ortega, Y. K. and D. E. Pearson (2005). Weak vs. strong invaders of natural plant
communities: Assessing invasibility and impact. Ecological Applications 15(2):
651-661.
Parker, I. M., D. Simberloff, W. M. Lonsdale, K. Goodell, M. Wonham, P. M. Kareiva, M.
H. Williamson, B. Von Holle, P. B. Moyle, J. E. Byers and L. Goldwasser (1999).
Impact: toward a framework for understanding the ecological effects of invaders.
Biological Invasions 1(1): 3-19.
Pauchard, A. and K. Shea (2006). Integrating the study of non-native plant invasions
across spatial scales. Biological Invasions 8(3): 399-413.
Peart, D. R. and T. C. Foin (1985). Analysis and predictionof population and community
change: a grassland case study. The population study of vegetation. J. White. Junk,
Dordecht.
Pimentel, D., L. Lach, R. Zuniga and D. Morrison (2000). Environmental and economic
costs of nonindigenous species in the United States. Bioscience 50(1): 53-65.
Pinheiro, J. and D. Bates (2000). Mixed effects models in S and S-Plus. New York,
Springer.
PlantyTabacchi, A. M., E. Tabacchi, R. J. Naiman, C. Deferrari and H. Decamps (1996).
Invasibility of species rich communities in riparian zones. Conservation Biology
10(2): 598-607.
Pysek, P. and A. Pysek (1995). Invasion by Heracleum mantegazzianum in different
habitats in the Czech Republic. Journal of Vegetation Science 6(5): 711-718.
R_Development_Core_Team (2009). R: A Language and Environment for Statistical
Computing.
Radford, I. J., K. J. M. Dickinson and J. M. Lord (2006). Nutrient stress and performance
of invasive Hieracium lepidulum and co-occurring species in New Zealand. Basic
and Applied Ecology 7(4): 320-333.
Radford, I. J., K. J. M. Dickinson and J. M. Lord (2007). Functional and performance
comparisons of invasive Hieracium lepidulum and co-occurring species in New
Zealand. Austral Ecology 32(3): 338-354.
70
Radford, I. J., K. J. M. Dickinson and J. M. Lord (2009). Does the invader Hieracium
lepidulum have a comparative growth advantage over co-occurring plants? High
leaf area and low metabolic costs as invasive traits. New Zealand Journal of Botany
47(4): 395-403.
Rejmanek, M. (1988). Invasibilty of plant communities. Biological Invasions. D. R. Drake,
H. A. Mooney, F. DiCastriet al. Chichester, Wiley and sons. 37: 369-388.
Richardson, A. D. and A. J. Friedland (2009). A review of the theories to explain Arctic
and alpine treelines around the world. Journal of Sustainable Forestry 28(1/2): 218242.
Robinson, G. R., J. F. Quinn and M. L. Stanton (1995). INVASIBILITY OF
EXPERIMENTAL HABITAT ISLANDS IN A CALIFORNIA WINTER ANNUAL
GRASSLAND. Ecology 76(3): 786-794.
Rose, A. B. and C. M. Frampton (1999). Effects of microsite characteristics on Hieracium
seedling establishment in tall- and short-tussock grasslands, Marlborough, New
Zealand. New Zealand Journal of Botany 37(1): 107-118.
Rose, A. B. and C. M. Frampton (2007). Rapid short-tussock grassland decline with and
without grazing, Marlborough, New Zealand. New Zealand Journal of Ecology
31(2): 232-244.
Rose, A. B., K. H. Platt and C. M. Frampton (1995). Vegetation change over 25 years in a
New Zealand short-tussock grassland: effects of sheep grazing and exotic
invasions. New Zealand Journal of Ecology 19: 163-174.
Rose, A. B., P. A. Suisted and C. M. Frampton (2004). Recovery, invasion, and decline
over 37 years in a Marlborough short-tussock grassland, New Zealand. New
Zealand Journal of Botany 42(1): 77-87.
Rusch, G. and E. Vandermaarel (1992). Species turnover and seedling recruitment in
limestone grasslands. Oikos 63(1): 139-146.
Sandler, H. A., P. Alpert and D. Shumaker (2007). Invasion of natural and agricultural
cranberry bogs by introduced and native plants. Plant Ecology 190(2): 219-231.
Sax, D. F. (2002). Native and naturalized plant diversity are positively correlated in scrub
communities of California and Chile. Diversity and Distributions 8(4): 193-210.
Sax, D. F. and S. D. Gaines (2008). Species invasions and extinction: The future of native
biodiversity on islands. Proceedings of the National Academy of Sciences of the
United States of America 105: 11490-11497.
Scott, D., J. S. Robertson and W. J. Archie (1990). Plant dynamics of New Zealand tussock
grassland infested with Hieracium pilosella. 2. Transition matrices of vegetational
changes. Journal of Applied Ecology 27(1): 235-241.
Shea, K. and P. Chesson (2002). Community ecology theory as a framework for biological
invasions. Trends in Ecology & Evolution 17(4): 170-176.
Spellerberg, I. F. and P. J. Fedor (2003). A tribute to Claude Shannon (1916-2001) and a
plea for more rigorous use of species richness, species diversity and the 'ShannonWiener' Index. Global Ecology and Biogeography 12(3): 177-179.
Stein, C., H. Auge, M. Fischer, W. W. Weisser and D. Prati (2008). Dispersal and seed
limitation affect diversity and productivity of montane grasslands. Oikos 117(10):
1469-1478.
Stinson, K., S. Kaufman, L. Durbin and F. Lowenstein (2007). Impacts of garlic mustard
invasion on a forest understory community. Northeastern Naturalist 14: 73-88.
Stohlgren, T. J., D. T. Barnett and J. Kartesz (2003). The rich get richer: patterns of plant
invasions in the United States. Frontiers in Ecology and the Environment 1(1): 1114.
71
Stohlgren, T. J., C. Jarnevich, G. W. Chong and P. H. Evangelista (2006). Scale and plant
invasions: a theory of biotic acceptance. Preslia 78(4): 405-426.
Tanentzap, A. J. and D. R. Bazely (2009). Propagule pressure and resource availability
determine plant community invasibility in a temperate forest understorey. Oikos
118(2): 300-308.
Theoharides, K. A. and J. S. Dukes (2007). Plant invasion across space and time: factors
affecting nonindigenous species success during four stages of invasion. New
Phytologist 176(2): 256-273.
Thomsen, M. A., C. M. D'Antonio, K. B. Suttle and W. P. Sousa (2006). Ecological
resistance, seed density and their interactions determine patterns of invasion in a
California coastal grassland. Ecology Letters 9(2): 160-170.
Tilman, D. (1997). Community invasibility, recruitment limitation, and grassland
biodiversity. Ecology 78(1): 81-92.
Timmins, S. M. (2004). How weed lists help protect native biodiversity in New Zealand.
Weed Technology 18: 1292-1300.
Truscott, A. M., S. C. Palmer, C. Soulsby, S. Westaway and P. E. Hulme (2008).
Consequences of invasion by the alien plant Mimulus guttatus on the-species
composition and soil properties of riparian plant communities in Scotland.
Perspectives in Plant Ecology Evolution and Systematics 10(4): 231-240.
Truscott, A. M., S. C. Palmer, C. Soulsby, S. Westaway and P. E. Hulme (2008).
Consequences of invasion by the alien plant Mimulus guttatus on the species
composition and soil properties of riparian plant communities in Scotland.
Perspectives in Plant Ecology, Evolution and Systematics 10(4): 232-240.
VanderMaarel, E. (1996). Pattern and process in the plant community: Fifty years after
A.S. Watt. Journal of Vegetation Science 7(1): 19-28.
VanderMaarel, E. and M. T. Sykes (1993). Small-scale plant-species turnover in a
limestone grassland - the carousel model and some comments on the niche concept.
Journal of Vegetation Science 4(2): 179-188.
Vila, M., A. S. D. Siamantziouras, G. Brundu, I. Camarda, P. Lambdon, F. Medail, E.
Moragues, C. M. Suehs, A. Traveset, A. Y. Troumbis and P. E. Hulme (2008).
Widespread resistance of Mediterranean island ecosystems to the establishment of
three alien species. Diversity and Distributions 14(5): 839-851.
Vila, M., M. Tessier, C. M. Suehs, G. Brundu, L. Carta, A. Galanidis, P. Lambdon, M.
Manca, F. Medail, E. Moragues, A. Traveset, A. Y. Troumbis and P. E. Hulme
(2006). Local and regional assessments of the impacts of plant invaders on
vegetation structure and soil properties of Mediterranean islands. Journal of
Biogeography 33(5): 853-861.
Vitousek, P. M. and L. R. Walker (1989). Biological invasion by Myrcia faya in Hawaii plant demography, nitrogen fixation, ecosystem effects. Ecological Monographs
59(3): 247-265.
Von Holle, B. and D. Simberloff (2005). Ecological resistance to biological invasion
overwhelmed by propagule pressure. Ecology 86(12): 3212-3218.
Walker, B. H. and W. Steffen (1997). An overview of the implications of global change for
natural and managed terrestrial ecosystems. . Conservation Ecology. 1(2).
Wardle, D. A. (1999). Is "sampling effect" a problem for experiments investigating
biodiversity-ecosystem function relationships? Oikos 87(2): 403-407.
Wardle, D. A. (2001). Experimental demonstration that plant diversity reduces invasibility
- evidence of a biological mechanism or a consequence of sampling effect? Oikos
95(1): 161-170.
72
Weigelt, A., T. Steinlein and W. Beyschlag (2002). Does plant competition intensity rather
depend on biomass or on species identity? Basic and Applied Ecology 3(1): 85-94.
Williamson, M. (1996). Biological Invasions. London, Chapman Hall.
Williamson, M. (1999). Invasions. Ecography 22(1): 5-12.
Williamson, M. and A. Fitter (1996). The varying success of invaders. Ecology 77(6):
1661-1666.
Winkler, E. and J. Stocklin (2002). Sexual and vegetative reproduction of Hieracium
pilosella L. under competition and disturbance: a grid-based simulation model.
Annals of Botany 89(5): 525-536.
Wiser, S. and R. Allen (2000). Hieracium lepidulum invasion of indigenous ecosystems.
Conservation Advisory Science Notes 278: 1-9.
Wiser, S. K. and R. B. Allen (2000). Hieracium lepidulum invasion of indigenous
ecosystems. (Conservation Advisory Science Notes. Wellington, Department of
Conservation Head Office: 1-9.
Wiser, S. K., R. B. Allen, P. W. Clinton and K. H. Platt (1998). Community structure and
forest invasion by an exotic herb over 23 years. Ecology 79: 2071-2081.
Wiser, S. K., R. B. Allen, P. W. Clinton and K. H. Platt (1998). Community structure and
forest invasion by an exotic herb over 23 years. Ecology 79(6): 2071-2081.
73