Download 1 Research paper Native grass establishment in grassy woodlands

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Research paper
Native grass establishment in grassy woodlands with nutrient enriched soil and exotic
grass invasion
Elizabeth A. Lindsay and Saul A. Cunningham
Commonwealth Scientific and Industrial Research Organization
GPO Box 1700, Canberra 2601 ACT, Australia
*corresponding author
[email protected], Telephone +612 6246 4121, Fax +612 6246 4362
Running title: Native grass addition in disturbed grassy woodlands
Keywords: woodland restoration, exotic annual grass, plant-soil relationships, seed addition,
C3 grass, south-eastern Australia
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Abstract
Successful re-introduction of species to the ground-layer of disturbed woodlands may
require management of competition with adult plants and addressing the potential influence of
soil nutrients. We investigated multiple factors that could potentially limit ground-layer
restoration in Eucalyptus melliodora-E. blakelyi grassy woodlands with modified ground
layers. We established plots over ground-layers varying in exotic annual and native perennial
grass cover. First we investigated relationships between the existing vegetation and soil.
Second, we evaluated whether the existing vegetation, soil fertility and habitat factors, such as
distance to trees, were barriers to native grass re-establishment by adding Poa labrillardieri
and Bothriochloa macra seeds to plots with an intact ground-layer, clipped ground-layer and
plots cleared of all vegetation.
Soil phosphorous had a positive relationship with exotic cover and a negative
relationship with native cover. Clipping vegetation prior to seeding had a negative effect on
Poa labrillardieri emergence and no effect on Bothriochloa macra, but survival of both
species was greater when seeds were sown close to trees. Both species established better in
cleared plots, indicating competition can inhibit restoration. Bothriochloa macra
establishment was low when soil carbon and C3 native grass cover were high. In contrast,
Bothriochloa macra established successfully in areas with high exotic cover and nutrient
enriched soil. Poa labrillardieri seedling survival was low in all circumstances. We conclude
that under comparable soil conditions it could be more difficult to re-establish additional
native grass species in woodland ground-layers dominated by simplified native grass
assemblages than those dominated by exotic annual grasses.
Keywords: woodland restoration, exotic annual grass, plant-soil relationships, seed addition,
C3 grass, south-eastern Australia
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Introduction
There can be many barriers to the restoration of vegetation communities with a history of
disturbance by livestock grazing and exotic plant invasion. Reintroduction of plant species by
seeding can be impeded by factors including elevated soil nutrients, and availability of microsites for germination (Musil 1993; Corbin & D'Antonio 2004). In exotic plant dominated
systems there can also be strong competition for light, water and space (Fogarty & Facelli
1999; Cione et al. 2002; D'Antonio & Meyerson 2002).
Grassy woodlands dominated by Yellow box (Eucalyptus melliodora) and Red Gum (E.
blakelyi) occur on the tablelands and slopes of south-eastern Australia, and are listed as an
endangered ecological community (DEWHA 2006). This community has an open tree canopy
and a diverse ground-layer of native grasses and herbs. These woodlands now occur as
remnants in an agricultural landscape, and many of the patches have been grazed by livestock
over the past 150 years (Benson 1991).
The grassy woodland understory is normally dominated by tall cool season (e.g. Poa
sieberiana) or warm season tussock forming perennial grasses (e.g. Themeda australis).
Heavy livestock grazing leads to increases in cool season species such as Microlaena stipodes
and Austrodanthonia spp (Lodge & Whalley 1985; Prober & Thiele 1995). Further grazing
pressure and elevated soil phosphorus or nitrate can favor invasion by annual exotic grasses
(Garden et al. 2001; Prober & Thiele 2005). Once established annual grasses may outcompete
native perennial grasses (Brown & Rice 2000; Groves et al. 2003). A similar successional
sequence from tall indigenous perennial grasses to short introduced annual species has also
been described for disturbed Californian grasslands (Seabloom et al. 2003).
Even after livestock grazing is removed weeds can persist and native plant diversity
often remains low in grassy ecosystems (Meissner & Facelli 1999; Spooner et al. 2002).
Weed persistence has been related to elevated soil nutrients (Prober et al. 2002b; Lenz &
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Facelli 2006), and while fertility remains high annual grasses may be more competitive than
native grasses (Groves et al. 2003). Replacement of tall C4 perennial grasses with short C3
perennial and annual grasses is likely to alter the spatial structure of the grassy layer. Canopy
gaps in grasslands can be important for the germination and survival of some, but not all, forb
species (Morgan 1998; Clarke & Davison 2004) and it remains to be seen if canopy gaps
assist with the reintroduction of native perennial grasses (Clarke & Davison 2004). Biomass
removal could be a simple way to create microsites more suitable for native seeds to
germinate.
Commonly used grassy woodland restoration techniques are limited to planting of trees
and shrubs as tube stock and fencing of the remnant to prevent further grazing by livestock
(Briggs et al. 2008), but these techniques do not directly address that element that has been
most changed; the ground-layer. There have been several studies conducted on plant
reintroduction by seeding into grassy ecosystems in Australia (e.g. Gibson-Roy et al. 2007;
Prober & Lunt 2009). However, most of these techniques have been limited to small scale
experiments and no guidelines exist on the regeneration requirements of many ground layer
species (Clarke & Davison 2004).
In this study we investigated multiple factors that could potentially limit ground-layer
restoration in Eucalyptus melliodora-E. blakelyi woodlands with modified ground layers by looking at
plant-soil relationships and conducting a seed addition experiment. First we investigated relationships
between the existing vegetation and soil properties in plots established over a gradient of exotic annual
grass and native perennial grass cover. We wanted to determine if there were strong links between the
vegetation composition and soil properties, such that vegetation could be used as an indicator of soil
nutrient status for selecting areas suitable for ground-layer restoration activities such as seed addition.
We hypothesized that elevated annual exotic grass cover would be associated with elevated soil
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nitrogen and phosphorous, and high native plant cover and diversity would be associated with low soil
phosphorous.
Second, we conducted a seed-addition experiment on the same plots using the cool season
tussock forming Poa labrillardieri (common tussock grass) and the warm season basal tuft forming
Bothriochloa macra (red grass). We aimed to reintroduce native grass species into weed dominated
areas, in particular by adding functional groups (warm season grasses and tall cool season tussock
grasses), that historically dominated the ground-layer of grassy woodlands, but were absent from the
field sites at the time of study. The seed-addition experiment aimed to provide a better understanding
of how restoration would work under different soil and vegetation conditions. We hypothesized that
elevated nitrate and exotic grass cover would be the main barriers to native grass addition. We also
postulated that large amounts of grass biomass could impede native grass emergence and survival, and
that clipping and removal of the ground-layer vegetation could facilitate re-establishment of native
grasses. Mature trees and coarse woody debris can influence soil nutrient levels and soil moisture
(Wilson 2002; Evans et al. 2003), so in addition we measured distance to tress and logs to determine if
these factors influenced emergence and establishment of red grass and common tussock grass.
Methods
We selected four patches of grassy woodland with both native and exotic plant species
in the understorey at the CSIRO Ginninderra Experiment Station Canberra, ACT, Australia
(149.080°E 35.193°S). The overstory was dominated by Eucalyptus melliodora (Yellow box),
with less cover of E. blakelyi (Blakely’s Red Gum) and E. mannifera (Brittle Gum). Sites
were flat to gently undulating and were separated by 0.8 to 2.8km. Livestock grazing has
occurred in the area for more than 100 years and all sites were actively grazed by merino
sheep (Ovis aries) and eastern grey kangaroos (Macropus giganteus). The soils were similar
across the four sites, being at least 25 cm deep and dominated by red earths and yellow
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podzolic soils (Sleeman 1979). Fertilizer had not been applied near the study areas for at least
six years.
Ten plant species were commonly encountered in the ground-layer (Table 1). The
native vegetation was dominated by perennial grasses (Mean cover 79 ± 33% 1SD), with C3
perennial grasses (69 ± 38%) more common than C4 grasses (9.7 ± 20%), and forbs rare. The
exotic component was dominated by annual grasses (79 ± 26%) and no shrubs were present in
the plots. Native plant richness was low (0 to 10 species per plot) with an average of 3.6 ± 1.7
species per 1 m2 plot. Exotic plant richness was similar (0 to 9 species), with an average of
4.5 ± 2 species per m2. Average total plant richness was 8.1 ± 3 species per m2.
Experimental design
We established 24 1 m2 plots at each of the four sites in October 2007 (Total n= 96).
Each plot was delineated within one of three vegetation types, 1) native dominated, 2) exotic
dominated or 3) co-dominated by native and exotic species. There were eight replicate plots
of each vegetation type at each site. Vegetation cover in the native dominated plots was at
least 75% native, and vegetation in the exotic dominated plots was at least 75% exotic. The
mixed vegetation plots had a similar cover of native and exotic vegetation (Approximately
50% native and 50% exotic). The average total plant cover across all vegetation types was 58
± 14% (1SD). Vertebrate herbivores (sheep, kangaroos, rabbits) were effectively excluded by
1m high fences (4 cm mesh).
We surveyed percent vegetation cover of all plant species in each plot (nine 33 x 33
cm grids per 1 m2 plot. Plots were initially surveyed in November 2007, and resurveyed in
November 2008 to determine impacts of seed addition and herbivore exclusion on the existing
vegetation (data not presented). Plant nomenclature follows Harden (1992-2007).
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We measured several plot level variables that could affect the plant assemblage and
grass establishment including distance to the nearest mature tree (Diameter breast height >
20cm), distance to the nearest log (coarse woody debris >10 cm diameter) and noted the
presence or absence of tree canopy cover above each plot. We also collected a soil sample
(4.5 cm x 5 cm depth) from the corner of each plot which was analyzed for total carbon, total
nitrogen, ammonium, nitrate, pH (1:5 CaCl2) and plant available phosphorus (bicarbonate
extractable), using the methods of Rayment and Higginson (1992).
Seed-addition experiment
Warm season and tall cool season tussock grasses historically dominated the groundlayer of grassy woodlands, but they were absent or in low cover at the study sites at the time
of study. We chose a C3 tussock (common tussock, Poa labrillardieri) and a C4 tuft grass
(red grass, Bothriochloa macra) as focal species for the seed-addition experiment. Common
tussock grass is a tall C3 that can tolerate wet and dry conditions. Red grass is a low-growing
basal tuft forming C4 grass that is dormant in winter and persists well through drought
(Milthorpe & Wynne 2001). Red grass is thought to act as an initial colonizer (G. Robertson
2008, Friends of the Grasslands, Canberra, ACT, personal communication). No botanical
surveys had been conducted in the study sites prior to disturbance, but both species are known
to be part of the E. melliodora- E. blakelyi box-gum grassy woodland community (DEWHA,
2006). Common tussock grass was present in undisturbed areas within 5 km of the experiment
station and red grass was present on the experiment station, but not in our sites.
The main seed-addition experiment was conducted in the 96 plots initially surveyed to
examine plant community-soil relationships. Two treatments, vegetation clipping (clipped or
unclipped) and grass seed-addition (seed added or no seed), were randomly applied to the
plots in a replicated orthogonal design with two replicates of each treatment at each site (3
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vegetation types x 2 seed treatments x 2 clipping treatments x 2 replicates = 24) (Table 2).
Clipped plots had vegetation trimmed to 3 cm above the soil. Plots were raked to remove
clipped material and raking was applied to unclipped plots as a procedural control. Raking
created furrows ~0.5cm deep in non-vegetated areas.
Two additional fully cleared plots were established at each site (2 x 4 sites = 8) on
mixed vegetation (i.e. an even cover of native and exotic vegetation). All vegetation was
removed from these plots by hand (no herbicide), and all cleared plots had gras seed added as
described below. These were established to determine if the existing vegetation provided
competition and as a partial control if there was no emergence in the vegetated plots. These
results were analyzed separately.
Seeds were purchased from Clean Seeds Pty Ltd (Bungendore, N.S.W.). The red grass
seeds were harvested from a commercial native grass seed farm, and the common tussock
seeds were harvested 25 km west of the study sites from wild populations. Red grass seeds
were clay coated to prevent predation and common tussock grass seeds were refined to
remove awnlets. We obtained 57% germination for red grass and 39% for common tussock
after 35 days at 25ºC in ambient light conditions in a glasshouse.
All seeds were sown in late spring (November 2007) and adequate rainfall was
received at this time. Seed addition plots received a mix of red grass and common tussock
grass seed in November 2007 (in excess of ~3000 and ~6000 seeds respectively). Seeds were
applied evenly by hand in the furrows from raking. Washed river sand (~2kg/plot) was then
added to prevent seeds from blowing away and to reduce predation and desiccation.
Red grass seeds have an optimal germination at 25 to 30ºC (Waters et al. 2000), which
was experienced in the months after sowing. Each plot was watered (2L) two to three times
per week, unless it rained, until the first frost occurred, in May 2008. The region has an
average annual rainfall of 622 mm (Australian Bureau of Meteorology), but only 529 mm was
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recorded from November 2007 to November 2008 at the weather station (Environdata,
Australia) located on the Station. Rainfall was highest during summer (November 106 mm),
and lowest in winter (May 7.6 mm). The air temperature ranged from 35.6 ºC (max, January)
to -6.3 ºC (min, August).
Plant counts
We counted emergent seedlings seven weeks after sowing (Time 1), followed by a count of
seedlings surviving at 20 weeks post sowing (Time 2). The final survey was conducted 12
months after sowing, in November 2008 (Time 3). Initial seedling counts were conducted
along fifty 1 m long transects spaced 2 cm apart. Subsequent surveys were done with transects
spaced 4 cm apart. As an indicator of plant size for red grass, we also determined the
proportion of plants that had at least one leaf greater or equal to 4.5 cm in length at 12
months.
Data analysis
Plant and soil relationships
To investigate relationships between the existing plant composition and cover and plot
properties (soil nutrient levels, distance to log, distance to tree, presence of canopy cover) we
performed multiple linear regressions using the All-subsets Regression procedure in Genstat
11 (Payne et al. 2008). Model fit comparison was done with the Akaike Information Criterion
and the adjusted r2. Cool and warm season native grasses may respond differently to grazing
and fertility (Garden et al. 2001), so we investigated separate models to explain their cover.
We did not explore models with highly negatively correlated predictor variables (e.g. native
and exotic plant covers), and we used a ratio of carbon and nitrogen when both variables were
important. We investigated relationships between the soil nutrients and whether soil nutrient
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levels were correlated with distance to the nearest tree or log by calculating the Pearson
correlation coefficient.
Seed addition
This experiment was designed as a fully factorial treatment structure; however there was more
variation in the plant cover of the vegetation types (native dominated, 50:50 mix, exotic
dominated) than desired. As a result we choose to treat the original vegetation cover type as a
continuous variable and analyse most of the data using multiple regression.
No common tussock or red grass plants were detected in the plots without seed
addition, therefore in assessing determinants of emergence and establishment we have only
examined the seed-added plots (n=48), again using all-subsets regression. We investigated
models for common tussock emergence, red grass emergence, red grass seedling survival, red
grass establishment and proportion of red grass with leaf blades ≥4.5 cm long. There was
insufficient data to analyze the second and third common tussock counts. Plant counts were
square root or log(x+1) transformed prior to analysis. Predictor variables explored in the
models were soil nitrogen, nitrate, ammonium, phosphate, pH, carbon, average native plant
cover, C4 native grass cover, C3 native grass cover, average exotic cover, native plant
richness, exotic plant richness, total plant richness, total ground cover, area of bare ground,
distance to nearest tree, distance to nearest log and presence of canopy cover. Soil variables
were log transformed and plant cover variables square root transformed. Clipping (clipped vs.
unclipped) was included as a grouping factor. We did not investigate models for the
difference between times 1, 2 and 3 because our observations suggested that both emergence
and death occurred between surveys, making population differences over time difficult to
interpret.
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To assess the effect of vegetation removal on seedling establishment we compared
seedling densities in cleared and seeded plots with those in the unclipped but seeded plots,
matched for similar vegetation cover (even mix of native and exotic cover). We used a two
factor ANOVA with site as a random factor and vegetation presence as a fixed factor (α =
0.05).
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Results
Plant and soil relationships
The plots occurred over a soil nutrient gradient, with values ranging from 16 to 143 g/kg total
carbon, 1.17 to 8.5 g/kg total nitrogen and 6.2 to 57 m/kg plant available phosphorus. Plant
and soil variables explained 17 to 39.6% of the variation in the initial plant cover variables
(Table 3). Exotic plant cover was positively correlated with soil phosphorus and distance to
the nearest tree (Table 3). Native plant cover was best explained by exotic richness (correlation), soil nitrogen (+ correlation) and soil phosphorus (- correlation) C3 native grass
cover was best explained by total soil nitrogen (+ correlation), nitrate (- correlation) and
exotic species richness (- correlation) (Table 3). C4 grass cover was best explained by
distance to the nearest log (+ correlation). We found that soil carbon, nitrogen, phosphorus
and nitrate were all significantly negatively correlated with distance to the nearest log
(Appendix 1). Exotic richness was partially explained by native cover. Whereas native plant
richness was best explained by soil nitrogen (+ correlation) and phosphorus (- correlation).
There were positive correlations between soil carbon, nitrogen, nitrate and phosphorus
(Appendix 1). Carbon was the only soil parameter that varied with distance to the nearest tree,
being higher close to the base of trees (Appendix 1), where the distance from based ranged
from 0.1 to 13.1m, average 4.3 ± 3.2 m.
Seed addition
Seven weeks following sowing, there were 62.1± 67 (1 SD) red grass seedlings/m2. Clipping
vegetation prior to seed addition did not enhance the red grass emergence or establishment,
and this factor was not important in any of the models. Red grass emergence was negatively
correlated with soil carbon and distance to tree and positively correlated with soil nitrate and
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total plant species richness (Table 4). Emergence was greatest in plots with low soil carbon,
high soil nitrate and plant richness and in close proximity to a tree.
Twenty weeks after sowing there was an average of 20.2 ± 20 red grass plants per m2.
At this time seedling number was negatively correlated with C3 native grass cover, soil
carbon and distance to nearest tree (Table 4).
Red grass plants established at an average of 11.6 ± 19 (1SD) plants per m2 in the
vegetated plots, with establishment occurring in 81% of the exotic grass dominated plots and
44% of the native dominated plots. One red grass plant was flowering by the final survey. Red
grass establishment was also negatively correlated with soil carbon and native C3 grass cover
(Table 4), such that the greatest establishment generally occurred when the plot had low soil
carbon and low native grass cover (Fig. 1). There were seven plots where more plants were
counted in the final survey than in the 20 week survey, indicating emergence was continuing
to take place. There was high rainfall and temperatures above 25ºC in this period.
The proportion of red grass plants with a leaf longer or equal to 4.5 cm after 1 year was
negatively correlated with the soil C:N ratio and positively correlated with total plant richness
(Table 4). In other words there were more large red grass plants in soil with a greater
proportion of nitrogen relative to carbon and higher plant richness, a similar result to that for
emergence.
Common tussock had a lower initial emergence count than the red grass, with 8.04±15
(1 SD) seedlings detected per m2. This decreased to a total of 12 seedlings after 20 weeks, and
only two after 12 months across all sown plots. Emergence was generally highest in plots
close to a tree and where vegetation was not clipped, with a multiple regression model
explaining 29% of the variation in common tussock emergence (Table 3). Emergence was
negatively correlated with distance to nearest tree (non-linear relationship) and the clipping
treatment was significant.
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Common tussock and red grass emergence was often greatest in the same plots, with a
positive correlation between the two (r2=0.48, p>0.001). Exotic plant cover and richness were
not significant explanatory variables for the common tussock grass or red grass counts. There
was no relationship between exotic vegetation and seedling counts.
There were more red grass and common tussock grass seedlings in the plots cleared of
all vegetation compared to the plots with a mixture of native and exotic vegetation (Table 5).
There was also greater red grass establishment in the cleared plots. Nevertheless there was
great variation among the cleared plots, with the standard deviations of the plant densities
comparable to the means (Table 5).
We examined relationships between the native and exotic plant cover and soil
parameters important for explaining red grass emergence and establishment and the existing
vegetation (Fig. 2). Most of the soil nutrient concentrations were positively correlated with
each other. There were no explanatory variables in common between the models for common
tussock emergence and native or exotic plant cover.
Discussion
Red grass plants established in Yellow Box-Red gum woodlands with high exotic plant cover
and nutrient enriched soil, but tussock grass plants only established in plots cleared of all
vegetation. We hypothesized that elevated nitrate and annual exotic grass cover would be the
main barriers to native grass addition, but we found no evidence for this. Exotic plant cover
and richness were not important predictors of emergence or establishment in any of the
regression models. When plant cover and richness were examined individually there was
either no relationship or a non-significant positive trend with emergence and establishment
counts. Clipping the potentially competing ground-layer vegetation prior to seed addition
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offered no benefit to either species. Clarke and Davison (2004) also found little benefit of
clipping or burning grass on seedling emergence in woodlands.
High soil carbon and C3 native grass cover appeared to be the main barriers to red
grass establishment in our experiment. We hypothesized that elevated nitrate or nitrogen
would be one of the main abiotic factors limiting establishment, and not carbon. Elevated soil
nitrate was positively associated with red grass emergence, while elevated nitrogen was
associated with a greater proportion of large red grass plants. Even though elevated nitrate
was positively associated with red grass emergence it should be noted that the conditions
suitable for germination are not always suitable for long term establishment (Turnbull 2000).
There was a negative correlation between C3 perennial grass cover and nitrate in this study.
Exotic plant cover in woodlands has been associated with elevated nitrate (Prober et al.
2002b), and red grass cover in pastures tends to decrease with increasing fertilizer use
(Garden et al. 2001). Our results suggest that red grass could be established from direct
seeding in relatively nutrient rich conditions when livestock grazing is excluded.
Red grass emergence was higher when total plant diversity was greater. Plots with
very low plant diversity could be more exposed to sun, wind or soil pathogens. Where as plots
with higher diversity could be more structurally complex and provide an array of
microenvironments (Silvertown 1981), some of which are suitable for red grass. C3 native
grass cover was a better indicator of red grass failure than total native plant cover or exotic
plant cover. It is possible that the small native forbs and sedges present offered less
competition than grasses that respond favorably to increased fertility and grazing pressure.
Emergence and survival of both seeded species was higher when sown close to a
mature eucalypt tree. Common tussock grass emergence was also greater when the preexisting grass was not clipped. Existing grass and trees could influence the seedling
microhabitat, providing shade, protection from wind or increased moisture availability.
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Young common tussock plants have fine leaves (<2 mm diameter), and we observed leaf
damage on these after hot sunny days, which did not appear on the larger red grass leaves.
Common tussock seedlings may require protection from the existing vegetation until they are
established.
Soil properties have been shown to change with distance from the base of eucalypt
trees, with nitrogen and carbon in particular higher within the canopy zone (Prober et al.
2002a; Wilson 2002). We found soil carbon to vary with distance to the nearest tree, but no
other parameters. Even though the plots were all at different distances from mature eucalypt
trees (0.1 to 13.1 m), the entire study area would be under some influence from the canopy or
roots of one or more trees.
The low number of surviving common tussock grass plants meant we could not
examine the factors influencing establishment success of this species. Other studies have
similarly recorded low emergence for this species in the field (Gibson-Roy et al. 2007).
Larger seeded species tend to have greater establishment success than smaller seeded species
in seed sowing experiments (Moles & Westoby 2002). This effect could explain part of the
success red grass had over common tussock grass. We do not know if the red grass success
was to the detriment of the common tussock grass, because we only considered one seed
mixing ratio and seeding rate. There was a positive correlation between the emergence rates
of both species, suggesting their microhabitat requirements are similar.
Like trees, some grass populations self thin to more stable densities (Lonsdale &
Watkinson 1983). In some plots, red grass plants established in higher density than would
naturally occur in grassy woodlands, and so death of some individuals might be desirable in
terms of a transition to more natural conditions. Common tussock grows to form a large plant
(base width ~30 cm and height ~1.2 m), and less than one adult plant per m2 would be
common in these woodland communities. Therefore it is likely that the establishment rate
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obtained on the cleared plots would be satisfactory in the longer term if there was little further
mortality.
The order in which species are sown or emerge can effect the success of other species
(Abraham et al. 2009). Red grass is thought to be good colonizer, which is supported by our
success in re-introducing it to disturbed woodlands. It is unknown if red grass is capable of
transforming the system in such a way as to facilitate establishment of other species, with
nursing effects only examined for a few species (Padilla & Pugnaire 2006). We suggest future
experiments could trial sowing red grass first, followed by sowing of species such as Poa spp
or Themeda australis after red grass establishment.
Native plant cover and richness were strongly negatively related to soil phosphorus,
whereas elevated soil nitrogen had a weak relationship with native richness and cover. C4
perennial grass cover increased with increasing distance to nearest log, with many nutrients
higher in plots closer to logs. The soil nutrient status of some of the native dominated plots
was comparable to those found for reference condition woodlands (Prober et al. 2002a).
However, some of the mixed vegetation and exotic dominated plots had elevated carbon,
ammonia, nitrate and plant available phosphorus up to 40 times higher than reference
conditions. Nutrients added to remnant vegetation through agricultural activities are often
very persistent (McLauchlan 2006; Lindsay et al. In review), with phosphorus often in stable
or immobile forms (Holford 1997).
The most common C3 native grasses were Austrodanthonia spp and Microlaena
stipoides, with many plots dominated by these. Their cover was positively correlated with
nitrogen, but negatively correlated with nitrate, with Australian native plants reported to have
a preference for ammonium over nitrate (Bidwell et al. 2006). Areas dominated by
Austrodanthonia spp and Microlaena stipoides can be very stable (Garden et al. 2000), with
their cover positively correlated with fertilizer application (Garden et al. 2001). These native
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perennial C3 grasses appeared to offer more competition to red grass than the annual exotic
C3 grasses. Cole et al. (2005) also thought perennial grasses offered competition to Themeda
triandra in a seed addition experiment. Grassy woodland communities can have multiple
stable states (Spooner & Allcock 2006), and it could be difficult to increase the diversity in
woodlands with a simplified native grass understory, and more intervention could be required
than seed addition into furrows, such as burning or greater soil disturbance (Clarke & Davison
2004).
While some glasshouse experiments have shown perennial native grasses to compete
poorly with annual exotic grasses in pots (Allcock 2002; Groves et al. 2003), we suggest that
that field experiments like ours provide a more realistic context for predicting restoration
outcomes. Under field conditions variation in recruitment becomes a more important factor,
and might be the crucial factor for establishment of native perennial grasses. Native perennial
grasses in an exotic grass invaded grassland in California were found to be recruitment
limited and not poor competitors, with the perennial grasses able to grow on lower resources
than those left by annual grasses (Seabloom et al. 2003). Perennial species could be better
competitors for below ground resources due to their deep root systems and greater allocation
to roots than annual grasses (Holmes & Rice 1996).
This work has implications for eucalypt woodlands and other grassy ecosystems that
are or have been part of agricultural landscapes. In making recommendations (see
Implications for Practice) we have looked at factors that influenced emergence, establishment
and the composition of the current native vegetation. Microhabitat, soil carbon and C3 native
grass cover all appeared to be important factors determining the success of the sown grass
species. These results suggest that it may be easier to reintroduce native grasses to an annual
exotic grass dominated grassy woodland, than to a woodland with a simplified understorey
dominated by C3 perennial grasses that respond positively to increased soil fertility and
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grazing pressure (Microlanea stipodes and Austrodanthonia spp), under comparable soil
conditions.
Implications for Practice
• Re-introduction of native grasses by seed addition can be more successful in woodlands
dominated by annual exotic grasses than simplified assemblages of C3 perennial grasses.
• Soil nutrient enrichment can be a barrier to reintroduction in low nutrient systems, but it is
possible to reintroduce some species, such as red grass, under such conditions.
• Biomass removal, such as clipping or mowing prior to seed addition may not benefit all
grass species
• Removal of ground-layer vegetation can lead to greater establishment of native species, but
may be limited by proximity to undesirable propagule supply.
Acknowledgments
We would like to thank staff at the Ginninderra Experiment Station, Juliana Skinner and Ben
Durant for field work assistance. This work was funded by Land and Water Australia.
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Benson, J. 1991. The effect of 200 years of European settlement on the vegetation and flora of
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Table 1 Common name, scientific names and life form (perennial (P) or annual (A)), for the
most common native (N) and exotic (E) plant species in the understory of the grassy eucalypt
woodland study sites. Two distinct species of Austrodanthonia were present, but one showed
variable traits and could not be identified to species level.
Scientific name
Common name
Life form
Origin
Austrodanthonia sp.
Wallaby grass
P
N
Austrodanthonia racemosa
Slender wallaby grass
P
N
Austrostipa scabra
Corkscrew spear grass
P
N
Lomandra filiformis
Wattle mat-rush
P
N
Microlaena stipoides
Weeping grass
P
N
Bromus diandrus
Giant brome
A
E
Bromus molliformis
Soft brome
A
E
Hordeum leporinum
Barley grass
A
E
Lolium rigidum
Annual rye grass
A
E
Vulpia bromoides
Rats tail fescue
A
E
26
Table 2 Factors and resultant treatments combinations for the main seed addition experiment.
All treatment combinations were replicated twice in each of the four woodland sites (Total n=
2 x 12 x 4 = 96). Clipped refers to the vegetation being cut to 3 cm height. The plots with
grass seed added had a mixture of red grass and common tussock seeds added.
Ground-layer vegetation type
Grass seed added
No seed added
Clipped
Clipped
Unclipped
Unclipped
Clipped
Clipped
Unclipped
Unclipped
Clipped
Clipped
Unclipped
Unclipped
Native dominated
Mixed (native & exotic)
Exotic dominated
27
Table 3 Multiple linear regression models that best explained the average exotic and native
cover and richness, and total plant richness in each plot. The adjusted r2 and the A.I.C were
used to compare model fit. All soil concentrations were log(x+1) transformed and plant cover
was square root transformed prior to analysis. The last column indicates the direction of the
regression slope between the response and predictor variable.
Response variable
Exotic cover
Total native cover
r2 (%)
Predictor
adjusted
variables
22.3
Overall model
32.8
F
p
line
14.7
<0.001
Phosphorus
<0.001
+
Distance to tree
0.006
+
Overall model
16.5
<0.001
Nitrogen
<0.001
+
Phosphorus
<0.001
-
Exotic richness
0.008
-
C4 native grass cover
23.6
Distance to log
30.9
<0001
C3 native grass cover
22.9
Overall model
10.43
<.001
Nitrogen
0.013
+
Nitrate
<0001
-
Exotic richness
0.014
-
Exotic richness
15.9
Native cover
9.70
0.003
Native richness
39.6
Overall model
31.7
<0.001
Total plant richness
Total plant richness
(exotic cover
excluded)
24.6
17.0
Slope of
Nitrogen
0.126
+
Phosphorus
<0.001
-
Overall model
8.66
<0.001
Phosphorus
<0.001
-
Exotic cover
0.014
+
0.002
-
Phosphorus
10.4
28
Table 4 Multiple regression models that best explained red grass germination (7 weeks, T1),
seedling survival (20 weeks, T2), establishment (12 months, T3), the proportion with the
longest leaf ≥4.5 cm long as well as poa tussock germination. Model fit comparisons were
done with the adjusted r2 and the A.I.C. The last column indicates the direction of the
regression slope between the response and predictor variable.
Category
Germination
Response
r2 (%)
variable
adjusted
Red grass
36.6
T1 (log)
Survival
Red grass
21.6
T2 (sqrt)
Establishment
Red grass
13.6
T3 (sqrt)
Size
Germination
Proportion
21.5
Predictor variables
F
p
Slope
of line
Overall model
7.63
<.001
Total plant richness
<.001
+
Distance to tree (log)
0.003
-
Carbon (log)
0.058
-
Nitrate (log)
0.005
+
Overall model
5.22
0.004
C3 grass cover
0.004
-
Distance to tree (sqrt)
0.028
-
Carbon (log)
0.056
-
Overall model
4.71
0.014
C3 grass cover
0.02
-
Carbon (log)
0.06
-
Overall model
7.18
0.002
red grass
C:N ratio
0.008
-
leaf ≥4.5 cm
Total plant richness
0.012
+
Poa T1 (log) 29.1
Overall model
7.17
0.001
Distance to tree (log)
0.001
-
Clipped
0.007
-
Tree x Clipped
0.074
+
29
Table 5. Summary of the two factor ANOVA results testing the difference for red grass and
common tussock counts per m2 (mean ± standard deviation) in the plots cleared of all
vegetation (cleared) and the plots with a mixture of native and exotic plants (midpoint on the
cover gradient). Grass counts were performed at three time periods, germination (T1, 7
weeks), seedling survival (T2, 20 weeks) and establishment (T3, 12 months) and data was
square root transformed.
Mean
S.D
per m2
Cleared vs. mix
F1,11
p
Mean
S.D
per m2
Cleared
Mixed vegetation
Poa T1
17.37
0.002
58.6
49
0.63
0.96
Poa count T2
21.29
<0.001
51.3
50
0
0
Poa count T3
NA
0.75
1.5
0
0
Red grass count T1
39.5
<0001
274
190
25.6
18.5
Red grass count T2
17.69
0.002
112
98.8
10.1
8.19
Red grass count T3
12.29
0.005
114
119
5.65
5.78
NA There was insufficient data to analyze the common tussock counts at T3.
30
Figure captions
Figure 1 Fitted regression lines and 95% confidence intervals for the multiple linear
regression model for red grass establishment per m2 (square root transformed) and (a) soil
carbon (g/kg) (log transformed) and (b) C3 native grass cover.
Figure 2 Diagram showing the variables that potentially influenced a) native and exotic plant
cover and red grass germination (T1), and b) red grass establishment (T3). Positive (solid
line) and negative (dashed line) correlations are indicated next to each arrow. Arrows indicate
possible pathways of how parameters other than those in the multiple regression models could
influence emergence and establishment. The variables were determined by model fit
comparison and multiple linear regression as indicated in tables 2 and 3. The correlations
between the soil nutrients are shown in Appendix 1.
31
Appendix 1 Correlations (r) between soil parameters that were part of significant multiple
regression models in Table 3 and 5, and between soil parameters and the sample distance to
the nearest mature eucalypt tree and log.
Variable 1
Variable 2
r
p
Carbon
Nitrate
0.525
<0.001
Carbon
Nitrogen
0.980
<0.001
Carbon
Phosphorus
0.777
<0.001
Nitrogen
Nitrate
0.600
<0.001
Nitrogen
Phosphorus
0.799
<0.001
Ammonia
Tree distance
0.04
0.562
Carbon
Tree distance
-0.393
<0.001
Nitrate
Tree distance
0.138
0.175
Nitrogen
Tree distance
0.170
0.130
Phosphorus
Tree distance
0.09
0.244
Ammonia
Log distance
0.265
0.995
Carbon
Log distance
-0.634
<0.001
Nitrate
Log distance
-0.370
0.002
Nitrogen
Log distance
-0.689
<0.001
Phosphorus
Log distance
-0.692
<0.001