The Effect of Fire and Grazing on the Cumberland Plain Woodlands Samantha Clarke University of Wollongong

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2004 The effect of fire and grazing on the Cumberland Plain Woodlands Samantha Clarke University of Wollongong

Recommended Citation Clarke, Samantha, The effect of fire and grazing on the Cumberland Plain Woodlands, Master of Science - Research thesis, School of Biological Sciences, University of Wollongong, 2004. http://ro.uow.edu.au/theses/2700

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The Effect of Fire and Grazing on the Cumberland Plain Woodlands

A thesis submitted in partial fulfillment of the requirements for the award of the degree

Master of Science (Research)

from

THE UNIVERSITY OF WOLLONGONG

By

SAMANTHA CLARKE

Bachelor of Science (Biology)

DEPARTMENT OF BIOLOGICAL SCIENCES 2004 CERTIFICATION

I, Samantha Clarke, declare that this thesis, submitted in partial fulfillment of the requirements for the award of Master of Science (Research), in the Department of Biological Sciences, University of Wollongong, is wholly my own work unless otherwise referenced or acknowledged. The document has not been submitted for qualifications at any other academic institution.

Samantha Clarke

20 June 2004 ABSTRACT

Temperate grassy woodlands throughout the world have suffered the effects of changed disturbance regimes, in particular, fire and grazing, due to human activities. Since European settlement fire and tree clearing has been used to modify grassy woodland vegetation for livestock grazing and agriculture. As a consequence some species, particularly shrubs and trees, have been reduced or eliminated and both native and introduced grasses have become more dominant. It is important to understand how disturbance regimes are affecting grassy woodlands to ensure long-term survival and diversity of the community.

The Cumberland Plain Woodlands are a good example of the magnitude of destruction endured by a typical grassy w~odland community. They are found in small remnants within the western Sydney region and are severely fragmented and disturbed with approximately only 6 % remaining. The Cumberland Plain Woodlands are listed as an endangered ecological community under the Threatened Species ConseNation Act, 1995 and the Environment Protection and Biodiversity ConseNation Act, 1999. This listing is a consequence of widespread destruction and fragmentation of the community due to tree clearing, agriculture and urbanisation. Little is known of the effect of changed disturbance regimes on the Cumberland Plain Woodland remnants and other vulnerable communities growing upon shale in western Sydney with most research focussing on the vegetation communities growing upon Hawkesbury Sandstone, which initiated this research.

The study area was located within one of the largest remnants of Cumberland Plain Woodlands, which is found at Holsworthy Military Area, western Sydney. Holsworthy has been part of the army training area for the last 100 years and therefore is isolated from growing urbanisation, agriculture and livestock grazing. The impact of frequent fire is of concern in this area as past management practices were to burn the area regularly to remove vegetation and create an open understorey able to be used for army training activities. Also, small firearms training activities often ignite fires in the woodland due to bullets missing targets and igniting the vegetation. Fires in the area are generally left to burn out without any intervention due to the danger of explosion of unexploded ordnances throughout the training area. High fire frequency and the cessation of livestock grazing at Holsworthy highlighted the necessity to answer the following questions: (i) How do plant species in the Cumberland Plain Woodlands respond to different fire regimes? ('ii) How do these plant species respond to grazing by animals other than domestic livestock, such as macropods and rabbits? And (iii) what effect does the heat and smoke from fire have on grass germination?

An exclosure experiment using various fire and grazing treatments was used to answer these questions and determine if the richness and abundance of species changed over a five-year period. Also, a laboratory germination experiment was used to determine how 22 grass species from the Cumberland Plain Woodlands respond to heat and smoke treatments designed to simulate the process of fire.

This study found that (i) past fires affected plant community composition in the Cumberland Plain Woodlands at Holsworthy Military Area, generally reducing the abundance of shrubs relative to grasses, although some particular shrub species were not reduced in abundance, (ii) community composition did not appear to be strongly affected by grazing, and (iii) germination of some grass species was stimulated by heat and/or smoke but germination of others was reduced.

This research indicates that fire is an important and frequent disturbance in the Cumberland Plain Woodlands at Holsworthy Military Area. Understanding the current fire regime is necessary to determine appropriate management strategies to ensure the survival of the Cumberland Plain Woodlands and other grassy woodland communities. This project enhances the research and monitoring component of the Plan of Management for the recovery of the Cumberland Plain Woodlands, Holsworthy Military Area. It highlights the need for landscape heterogeneity to maintain survival and diversity of the

ii community and increases the knowledge of the effect of heat and smoke from fire on the germination of grasses in grassy woodlands.

iii Acknowledgments

I would like to thank many people for their help and guidance during the course of this project, I cannot name them all here but their support and encouragement has been gratefully appreciated. I would like to make a special thank you to some of them:

I would like to thank Marina Peterson from the Department of Defence for her willingness to support the ongoing process of this research, financial support and advice on army procedures.

A big thank you to Kris French, my supervisor, for her encouragement to pursue my ambitions and for providing the opportunity to undertake this project. Kris's advice, support and friend~hip have been fantastic.

Warrant Officer Mick Blacksland, Holsworthy Range Control and security staff of Total Asset Management Services for help in accessing my study sites.

Sarah Hill for help in the initial location of the experimental sites and for sharing valuable information of the study area. Sarah's help has been invaluable and much appreciated.

Jean Clarke, Sarah Hill, Maria Adams, Marie Turner and Kris French who were all excellent field assistants.

Lotte von Richter, Cathy Offord and staff of the Royal Botanic Gardens, Mt Annan for advice on procedures of the germination experiment and help with species selection and location and for use of the research laboratory. Thanks to Lotte for technical assistance in the field and laboratory.

Belinda Pellow has provided help with plant identification, map production and much more. I am very grateful to Belinda for help and advice on so many issues.

iv The Janet Cosh Herbarium girls, Belinda Pellow and Jean Clarke, thanks to you both for being so supportive and for your friendship.

Thanks to Jack Baker for his encouragement, advice and comments on earlier drafts of this thesis.

David Keith for advice on experimental design and analysis and for comments on earlier drafts of the germination chapter.

Jane Wasley for assistance with the construction of the smoke machine.

Ken Russell for help and advice on statistical analysis.

Jocelyn Howell and Doug Benson for advice on the grasses of the Cumberland Plain Woodlands.

Lastly, to my family Ginge, Laura and Stacey. Thanks to Ginge for his constant encouragement, patience and tolerance. Thanks to Laura and Stacey, for their patience and acceptance. Without the love and support of my family this project would have been impossible and I am forever grateful.

v Table of Contents

ABSTRACT ACKNOWLEDGEMENTS TABLE OF CONTENTS LIST OF TABLES LIST OF FIGURES LIST OF APPENDIXES

Chapter 1 - Introduction 1 1.1 Background 1 1.2 Disturbance in grassy woodlands 1 1.3 Distribution of grassy woodlands 2 1.4 The effect of fire on grassy woodlands_ 4 1.4.1 Fire frequency 4 1.4.2 Fire intensity 5 1.4.3 Season of fire 5 1.5 The effect of grazing on grassy woodlands 6 1.5.1 Domestic livestock 7 1.5.2 Rabbits 7 1.5.3 Native Herbivores 8 1.6 Fire and grazing interactions 8 1. 7 Management of grassy woodlands in Australia 10 1.8 Cumberland Plain Woodlands 11 1.8.1 Fire and grazing at Holsworthy 12 1.9 Aims of the study 12

Chapter 2 - The effect of fire and grazing on plant species richness and abundance 14

2.1 Introduction 14 2.2 Methods 17 2.2.1 Location 17 2.2.2 Experimental design 17

vi 2.2.3 Surveys 18 2.2.4 Total and mean species richness 19 2.2.5 Rainfall 19 2.2.6 Statistical analysis 19 2.3 Results - Part A - Changes in Species Richness 1997-2002 23 2.3.1 Overview of the yearly changes in total species richness 23 2.3.2 Rainfall 26 2.3.3 Differences among fence and fire treatments 28 2.3.4 Multivariate analysis of presence/absence data 29 2.3.4.1 Total species 29 2.3.4.2 Shrubs 31 2.3.4.3 Herbs and Poaceae 32 2.3.5 Grazing and fire interactions 36 2.4 Results - Part B - Changes in Cover_Abundance 2001-2002 37 2.4.1 Multivariate analysis of cover abundance data 37 2.4.1.1 Total species 37 2.4.1.2 Shrubs 38 2.4.1.3 Herbs 43 2.4.1.4 Poaceae 43 2.5 Cover abundance of Themeda australis 47 2.6 Discussion 48 2.6.1 The impact of fire on species richness and cover abundance 48 2.6.2 The impact of grazing on species richness and cover abundance 50 2.6.3 The interactive effect of fire and grazing on species richness and abundance 51

Chapter 3 - The germination response of 22 Poaceae species to heat and smoke 53

3.1 Introduction 53 3.2 Materials and methods 56 3.2.1 Seed collection 56 3.2.2 Laboratory germination studies 57 3.2.3 Heat treatments 58 3.2.4 Smoke water treatment 58

vii 3.2.5 Sowing 59 3.2 .6 Statistical analysis 59 3.3 Results 62 3.3.1 Heat treatment response 63 3.3.2 Smoke treatment response 66 3.3.3 Heat and smoke interaction 69 3.4 Discussion 71 3.4.1 Heat treatments 72 3.4.2 Smoke treatment 74 3.4.3 Heat and smoke interaction 75

Chapter 4 - Concluding discussion 77

4.1 The effect of fire and grazing on species richness and abundance 78 4.2 The effect of heat and smoke on grass germination 79 4.3 Conclusions 80

References 83

viii List of Tables

Table 2.1: Experimental design showing the number of replications for each fire treatment. 18 Table 2.2: Values used to record cover abundance (CA) of species in survey sites based on the Braun-Blanquet scale (Poore, 1955). 18 Table 2.3: SIMPER showing the average dissimilarity (Ave Diss) between the species richness of Total species for 2001 and 2002 showing the most discriminating species. Diss/SD = standard deviations between dissimilarity indices, Contrib (%)­ percent contribution to the dissimialrity index, cum %=cumulative percent. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt +summer burn 2000, CBS2=control burnt 1997 + summer burn 2000. Bold indicates a ·strong discriminator between treatments. 30

Table 2.4: SIMPER showing the average dissimilarity (Ave Diss) between the species richness of shrub species for 2001 and 2002 showing the top 30% of species. Diss/SD = standard deviations between dissimilarity indices, Contrib (%)-percent contribution to the dissimialrity index, cum %=cumulative percent. Fire treatment symbols are; SB=summer burn 1997, UBS2=unburnt +summer burn 2000, CBS2=control burnt 1997 + summer burn 2000. Bold indicates a strong discriminator between treatments. 32

Table 2.5: SIMPER table of the species richness of herb species (1997) showing significant differences between grazed and ungrazed sites. Species shown are the top 30% of the average dissimilarity (Ave Diss) between sites. CF = closed fence, OF = open fence. Diss/SD = standard deviations between dissimilarity indices, Contrib % = the percent contribution to the dissimilarity index, Cum % = cumulative %. Bold indicates a strong discriminator between treatments. 33

IX Table 2.6: SIMPER table of the species richness of Poaceae species (1997 & 2001) showing significant differences between ·grazed and ungrazed sites. Species shown are the top 30% & 50% of the average dissimilarity (Ave Diss) between sites. CF = closed fence, OF = open fence. Diss/SD = standard deviations between dissimilarity indices, Contrib % = the percent contribution to the dissimilarity index, Cum % = cumulative %. Bold indicates a strong discriminator between 34 treatments.

Table 2.7: SIMPER showing the average dissimilarity (Ave Diss) between the species richness of Poaceae species for 1999, 2001 and 2002 showing the top 30% of species. Diss/SD = standard deviations between dissimilarity indices, Contrib (%)-percent contribution to the dissimialrity index, cum %=cumulative percent. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 + summer burn 2000. Bold indicates a strong discriminator 35 between treatments.

Table 2.8: Analysis of similarities (ANOSIM) of the untransformed cover abundance data for 2001 and 2002 in the fire treated sites. Pairwise tests are presented for the sites that showed significant differences. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control 37 burnt 1997 +summer burn 2000.

Table 2.9: Similarity percentages (SIMPER) of untransformed 2001 data showing the average abundance and the standard deviation between dissimilarity values (Diss/SD) of shrub species between closed fence (CF) and open fence (OF) treatments. Contrib %=percentage contribution to the 39 dissimilarity index, Cum %=cumulative percent.

Table 2.10: Similarity percentages (SIMPER) of untransformed 2001 and 2002 data showing the average abundance and the standard deviation between dissimilarity values (Diss/SD) of shrub species between fire treatments. Contrib %=percentage contribution to the dissimilarity index, Cum %=cumulative percent. Fire treatment symbols are; SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 40 +summer burn 2000.

x Table 2.11: Similarity percentages (SIMPER) of untransformed 2001 data showing the average abundance and the standard deviation between dissimilarity values (Diss/SD) of Fabaceae ·species between closed fence (CF) and open fence (OF) treatments. Contrib %=percentage contribution to the 42 dissimilarity index, Cum %=cumulative percent.

Table 2.12: Similarity percentages (SIMPER) of untransformed 2001 data showing the average abundance and the standard deviation between dissimilarity values (Diss/SD) of Poaceae species between closed fence (CF) and open fence (OF) treatments. Contrib %=percentage contribution to the 43 dissimilarity index, Cum %=cumulative percent.

Table 2.13: Similarity percentages (SIMPER) of untransformed 2001 and 2002 data showing the average abundance and the standard deviation between dissimilarity values (Diss/SD) of Poaceae species between fire treatments. Contrib %=percentage contribution to the dissimilarity index, Cum %=cumulative percent. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 46 +summer burn 2000.

Table 3.1: Shows the species used, collection date, location of collection (MA = Mt Annan Botanic Garden (native annex of the Royal Botanic Gardens, Sydney), SADA = Small Arms Danger Area, Holsworthy Military Area), date of sowing, the age of the seed at time of sowing and whether all enclosing structures were removed. *indicates introduced species. 57

Table 3.2: Mean germination (%) and ± standard deviation of each species in order from highest to lowest germination. The age of seed (Age); time to reach total germination (Germ Time); the germination response of species that did not respond to the heat or smoke treatments but (Good=71-100%, Moderate=70- 31 %, Poor=0-30%); H indicates a response to heat, either to H=high or L=low temperatures; S indicates a response to smoke, Y=a positive response, HxS indicates a heat/smoke interaction, the letter X indicates an interaction. * indicates an 62 introduced species.

Xl Table 3.3: Phylogeny and classification of 22 Poaceae species (Mallet & Orchard, 2002). The table is grouped into 6 subfamilies, 8 tribes, 16 genera and 22 species. The response ·category shows the treatment response of each species; H=heat, S=smoke, HxS=heat/smoke interaction. 63

Table 3.4: Results of Two-way ANOVA table for 16 species that showed significant differences in germination following heating and smoke treatments. The F ratio (F), the probability (p)=0.05 are shown. Cochran's C test shows homoscedasity except numbers in bold where variances were heterogeneous for P<0.05 & <0.01. * indicates an introduced species. 64

Table 3.5: Tukey's HSD tests on species where significant ANOVA was found. Critical value for q; Heat =4.046, HxS =4.602. Smoke category shows whether the significant difference shows a decrease (0) or an increase (I) in 65 germination with the addition of smoke water.

Xll List of Figures

Figure 2.1 : Location of Holsworthy Military Area, western Sydney, Australia. 21

Figure 2.2: Study sites showing corresponding fire treatments. 22

Figure 2.3: Species richness of the community in the years (a) 1997, (b) 1999, (c) 2001, (d) 2002. 24

Figure 2.4: Total number of species of each major lifeform category present in each survey year. 25

Figure 2.5: Number of species recorded from the Asteraceae and Orchidaceae families for each survey year. 25

2 Figure 2.6: Mean (± s.d.) and total species richness (100 m ) in the 48 sites for each survey year. MSR = Mean Species Richness, TSR =Total Species Richness. 26

Figure 2.7: Annual rainfall (mm) for the years 1996-2002 in the Holsworthy Military Area taken from the Michael Wenden Centre, Liverpool. 27

Figure 2.8: Mean (± s.d.) number of species (MSR) per plot in 1997, 1999, 2001 and 2002 with corresponding annual rainfall (mm). 27

Figure 2.9: Mean (± s.d.) number of species in each fence treatment in 1997, 1999, 2001 and 2002. CF = closed fence treatment (no grazing) and OF = open fence (grazing) (n=16). 28

Figure 2.10: The mean (± s.d.) number of species recorded in each fire treatment group in 1999, 2001 and 2002. UB=unburnt, CB=control burn, SB=summer burn (1997 wildfire), UBS2=unburnt + summer burn2 (2000 wildfire), CBS2=control burn + summer burn2 (2000 wildfire), SBS2=summer burn + summer burn2 (2000 wildfire). 29

Figure 2.11: Change in species richness in fire and grazing treatments in 1999 compared to 1997. UB=unburnt 1997, CB=control burn 1997, SB=summer burn 1997. CF=closed fence, OF=open fence. Different letters represent a significant difference via Tukey's HSD test (p=0.05). 36

Xlll Figure 2.12: nMDS ordination sites of fire treatments for the cover abundance of total species in 2001 and 2002. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 +summer burn 2000, SBS2=summer burn 1997 + summer burn 2000. 38

Figure 2.13: nMDS ordination sites of fence treatments for the cover abundance of shrub species in 2001 and 2002. Fence treatment symbols are CF = closed fence, OF = open fence. 39

Figure 2.14: nMDS ordination sites of fire treatments for the cover abundance of shrub species in 2001 and 2002. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 +summer burn 2000, SBS2=summer burn 1997 + summer burn 2000. 41

Figure 2.15: nMDS ordination sites of fence treatments for the cover abundance of Fabaceae species in 2001. Fence treatment symbols are CF = closed fence, OF = open fence. 42

Figure 2.16: Average cover abundances (based on ordinal Braun Blanquet values) for Poaceae species in 2001 that contributed up to 30% of the average dissimilarities between significantly different fire treatment groups (based on ANOSIM). Fire treatment symbols are; UB=unburnt, CB=contro'I burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 + summer burn 2000, SBS2=summer burn 1997 + summer burn 2000. At=Austrodanthonia tenuior, Dm=Dichelachne micrantha, Es=Entolasia stricta, Ms=Microlaena stipoides, Ta=Themeda australis . 44

Figure 2.17: Average abundances (based on ordinal Braun Blanquet values) for Poaceae different fire treatment groups (based on ANOSIM). Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 +summer burn 2000, SBS2=summer burn 1997 + summer burn 2000. Dm=Oichelachne micrantha, Es=Entolasia stricta, Microlaena stipoides, Ps=Panicum simile, Ta=Themeda australis. 45

Figure 2.18: The mean percent cover of Themeda australis in the closed fence (CF) and open fence (OF) sites in 2001 and 2002. 47

XIV Figure 2.19: The mean cover abundance of Themeda australis in the fire treated sites in 2001 and 2002. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt +summer burn 2000, CBS2=control burnt 1997 + summer burn 2000, SBS2=summer burn 1997 + 47 summer burn 2000.

Figure 3.1: Smoke machine used to produce smoke water. This procedure was performed in a fume hood. 61

Figure 3.2: The mean germination data (± s.d.) of (a) Aristida ramosa and (b) Austrodanthonia racemosa var. racemosa to temperature (°C). Differences in letters above each temperature indicates a significant difference via Tukey's HSD. 64

Figure 3.3: The mean data (± s.d.) (back transformed to %) of species showing significant difference in the heat treatments (temperature °C). Differences in letters above each temperature indicates a significant difference via Tukf?y'S HSD. 66

Figure 3.4: The mean percentage germination (± s.d.) under smoke treatments in Aristida vagans. 67

Figure 3.5: The mean germination (± s.d.) (back transformed to %) of the large significant differences between no smoke/smoke treated seeds. 68

Figure 3.6: The mean germination (germ) data (%) (back transformed to %) of six species showing significant differences in heat/smoke interactions using Tukey's HSD. Tukey's HSD was performed at each temperature to investigate the impact of smoke. Temperature is in °C. Differences in letters above each temperature indicates a significant difference via Tukey's HSD. 70

xv List of Appendixes

Appendix 1: List of species identified at the Small Arms Danger ·Area, Holsworthy between 1997 and 2002. Reg Con = regional conservation status, C = conserved, V = vulnerable, * = introduced species. Species that are considered to be of particular significance in the western Sydney context are shown 94-98 in bold type (James et at., 1999).

Appendix 2: Australian Map Grid references taken from a Global Positioning System of the forty-eight study sites at the 99-100 Small Arms Danger Area, Holsworthy.

XVI Chapter 1 Introduction

1.1 Background Plant communities throughout the world experience many natural and human induced disturbances. A disturbance is described as 'any relatively discrete event in time that removes organisms and opens up space which can be colonised by individuals of the same or different species' (Bond and van Wilgen, 1996). It is generally accepted that disturbance can vary in intensity, frequency and time (Gill et al., 1981, Leck et al., 1989, Bond and van Wilgen, 1996) and may arise from such activities as grazing animals, pathogens, humans (trampling, mowing, ploughing, tree clearing) or natural phenomenon such as fire, cyclones, drought and soil erosion (Grime, 1974, Flannigan & Bergeron, 1998, Ishikawa et al., 1999). S_hort-term or pulse disturbances such as cyclones, wildfire etc. can be beneficial to increase diversity by opening up areas for colonisation of new individuals (Connell, 1978) whereas long-term or press disturbances such as intensive grazing, agriculture, tree clearing, and many other human activities can be detrimental to many species due to the constant destruction of individuals.

1.2 Disturbance in Grassy Woodlands Human activities have caused widespread disturbance in grassy woodlands throughout the world and have been influential in altering the disturbance regimes to the detriment of many organisms and indeed whole communities (Bond & van Wilgen, 1996). Apart from disturbances such as land clearing and utilisation, two major disturbances of natural habitats induced by humans are altered fire and grazing regimes (Mcintyre & Lavorel, 1994, Yates & Hobbs, 1997).

It is widely accepted that the occurrence of fire in vegetation communities plays an important role in maintaining species diversity (Withers & Ashton, 1977, Stuwe & Parsons, 1977, Shea et al., 1979, Benson, 1985, Clark, 1988, Williams et al., 1994, Morrison et al., 1995, Cary & Morrison, 1995, Abrahamson & Abrahamson, 1996, Morgan, 1996, Bradstock et al., 1997, Changxiang et al., 1997, Morrison & Renwick, 2000, Morgan, 1998b). When fire regimes are altered and grazing by introduced herbivores is present, there is a shift in vegetation composition, a reduction in the diversity of native species and an increase in the occurrence of exotics ('Hacker, 1984, Neave & Tanton, 1989, Mcintyre & Lavorel, 1994, Tremont, 1994, Pettit et al., 1995, Prober & Thiele, 1995, Miller & Halpern, 1998, Fensham & Skull, 1999). Changes to natural fire and grazing regimes has been an integral factor in determining the present day distribution of grassy woodlands throughout the world.

1.3 Distribution of Grassy Woodlands A grassy woodland is a vegetation community with a continuous grass/herbaceous stratum with trees and scattered shrubs (Cole, 1986).

Outside Australia, grassy woodlands ar~ known as savannas (Cole, 1986) covering approximately 23 million km 2 between the equatorial rain forests and the mid-latitude deserts and semi-deserts (Cole, 1986).

Climatic conditions of the grassy woodlands include a predictable rainfall regime with rainfall highest in spring/summer, when moisture is most needed for favourable plant growth, and lowest in the cooler months (Cole, 1986). There are considerable climatic variations between the grassy woodlands of the world. The annual rainfall generally decreases with increasing latitude from approximately 2000 mm at the margin of the tropical forests, to 250 mm at the desert margins (Cole, 1986). Also, the length of the dry season increases from three to four months to eight or nine months and becomes more clearly defined and severe (Cole, 1986).

Grassy woodlands are most extensive in South America which experiences relatively high rainfall of around 2000 mm (Groves, 1994). The African savannas are found in a 200-1800 mm rainfall belt and are similar to the Australian grassy woodlands. They occupy similar rainfall zones and have a large number of Poaceae species, which is the most speciose family in the community, with common genera such as Aristida and Eragrostis, with common shrub species such as Acacia. Due to the long period of human

2 settlement in Africa, it is hard to distinguish between natural and derived grassy woodlands as pre- and post-settlement research is limited (Groves, 1994). India and Southeast Asia have seen dramatic changes in vegetation structure due to human settlement. Remnant grassy woodlands now prevail where open forest and dense evergreen forest once existed (Groves, 1994).

The grassy woodlands of Australia cover approximately 25% of the continent, with the present distribution reflecting patterns of land use such as tree clearing and pastoralism (Groves, 1994). They are situated in south-eastern Western Australia, northern Australia (including northern WA, Northern Territory and northern Queensland) and extend inland along eastern Australia, west of the Great Dividing Range (Groves, 1994).

In Australia, the range of plant spec_ies within each grassy woodland community is influenced mainly by local rainfall conditions. The tropical areas are dominated by grass species such as Themeda australis. Eastern Australia or the semi-arid regions support mainly Themeda spp., with Dicanthium spp. and Botriochloa spp. as co-dominants. In the West, dominant species include Austrostipa, Eragrostis and Chrysopogon.

Grassy woodlands of the world have suffered more disturbances as a result of human activity than any other community. These vegetation communities are the most threatened communities in Australia (Yates & Hobbs, 1997) due to grazing by domestic livestock, agriculture and tree clearing after European settlement (Yates & Hobbs, 1997). Throughout Australia, most of these woodlands are restricted to relatively small remnants of various size and quality (Mcintyre & Lavorel, 1994, Prober & Thiele, 1995) and have little resemblance to their structure prior to European settlement (Groves, 1994, Tremont, 1994). Prober and Thiele (1995) found that tree clearing acts to reduce habitat diversity for herbaceous species and observed that vegetation under tree canopies has less grass (eg, Themeda) and more herbs. Tremont (1994) suggested that a temperate tall-grass grassland community deficient in shrubs, forbs and leguminous herbs has replaced many grassy woodland communities.

... .) Originally, the extent of grassy woodlands in eastern Australia was on the richest soil (Prober & Thiele, 1995) and provided an ideal environment to produce feed for livestock (Bryant & Litt, 1971, Bryant, 1973, Leigh & Holgate, 1979, McNaughton, 1985, Bennett, 1994, Noy-Meir, 1995, Fensham & Skull, 1999, Henderson & Keith, 2002). To increase productivity, graziers have manipulated the natural fire regime in grassy woodlands to increase the abundance of palatable grass species and remove trees and shrubs for easier access (Tothill, 1969, Orr et al., 1997). These changes have caused a shift from woodland to grasslands in some areas with the succession of many exotic species (Stuwe & Parsons, 1977, Hacker, 1984, Lunt, 1990, Grice & Barchia, 1992, Tremont, 1994, Prober & Thiele, 1995, Pettit et al., 1995, Ash & Mcivor, 1998, Fensham & Skull, 1999). These problems are also prevalent in the grassy woodlands of Northern A1,1stralia (Winter, 1991) and Western Australia (Pettit et al., 1995).

1.4 The Effect of Fire on Grassy Woodlands Grassy woodland communities have been affected by changed fire regimes as a consequence of human activities. Fire frequency, intensity and season have had important effects on grassy woodlands.

1.4.1 Fire Frequency Fire frequency, or time between fires, can have an enormous impact on some vegetation communities (Whelan, 1995). Generally, frequent burning (ie <5 years) in grassy woodlands favours resprouting perennials over non­ resprouting species, disadvantages species which rely solely on on-site storage of seed, promotes herbaceous cover over woody plants which in turn reduces subsequent fire intensity (Groves & Burdon, 1986, Bock et al., 1995). In eastern Queensland, Heteropogon contortus numbers increased following repeated spring burning over three years (Tothill, 1969). Similarly, Orr et al., (1997) found that fire-resistant species such as H. contortus increased after annual spring burning regimes for four years while other grasses such as Aristida spp. and Bothriochloa spp. reduced in numbers. In the south-eastern

4 Arizona grasslands frequent burning favoured some species, eg Eragrostis intermedia (Bock et al., 1995).

Other research has shown that frequent fire (approx. 3 years) in grasslands is required to maintain plant diversity by preventing dominant grass species such as Themeda austra/is from out-competing other herbaceous species (Stuwe & Parsons, 1977, Lunt 1994, Tremont, 1994, Morgan, 1997, Lunt & Morgan, 1999b, Morgan & Lunt, 2000). This occurs by destroying individual plants and creating opportunities for seedling regeneration while allowing established plants to persist (Morgan, 1998a).

1.4.2 Fire Intensity Frequent burning in grassy communities reduces the fuel load and, therefore, produces fires of low intensity (Good, 19?1). These low-intensity fires favour grass and herb species able to regenerate vegetatively (Christensen et al., 1981) and tend only to stimulate the germination of seeds from the upper level of soil (Auld & O'Connell, 1991 ). While low-intensity fire may be sufficient to encourage regeneration of grassy understorey species it is not favourable for communities supporting large numbers of woody shrubs and leguminous species (Christensen et al., 1981). Research indicates that these communities require high-intensity fires to break seed dormancy and stimulate germination in many hard-seeded species (Shea et al., 1979, Benson, 1985, Clark, 1988, Williams et al., 1994, Morrison et al., 1995, Bradstock et al., 1997).

1.4.3 Season of Fire The season in which a fire occurs is also an important aspect of the fire regime in many communities (Bond & van Wilgen, 1996). Fire season may cause more intense fires in late summer and autumn than at other times of year due to low moisture content of fuel (Christensen et al., 1981). In grassy woodlands, the effect of burning on species richness and abundance is more important than· the season of the burn (Lunt, 1990, Hill, 2000). However, timing of a fire when related to other factors such as timing of favourable rainfall and the· availability of mature seed may influence germination of some

5 Eucalyptus spp. and because these trees form the whole upper canopy, are particularly important in grassy woodland communities (Hill & French, in press).

1.5 The Effect of Grazing on Grassy Woodlands Large grazing mammals are an important factor in the dynamics of grassland ecosystems (McNaughton, 1985). As recently as 150 years ago, most grassy woodlands and grasslands of the world supported large migratory populations of hoofed herbivores, bison in North America, saiga antelope on the Eurasian Steppe, wildebeest and zebra on the African Savanna (Frank, 1998). Most of these grassy woodlands have been largely eliminated (Frank, 1998). Cultivation has replaced approximately 20%, and the rest has been transformed into pasture and open grassland supporting domesticated

ungulates (Frank, 1998). In natural gra~ing ecosystems, the vegetation has evolved to survive seasonal grazing by migratory ungulates. Australia is unique in its lack of naturally occurring hard hoofed ungulates and were grazed only by soft-footed marsupials and small mammals before the arrival of European settlers adding to the impact of changed disturbance regimes within grassy woodlands on the continent.

McNaughton (1985) suggested that nutrient cycling through the breakdown of dung and urine of large grazing mammals is a major contributor to the stimulatory effect of grazing on growth of the Serengeti grasslands in South Africa. This, coupled with grazing on an intermediate level, promotes the highest productivity in the grasslands. It is only when equilibrium (Connell, 1978) is lost through high-intensity grazing· that vegetation composition changes due to invasion by annual grasses and unpalatable herbs (McNaughton, 1985). In a natural grazing ecosystem, this change in composition will operate as a negative feedback and tend to diminish grazing intensity until equilibrium is restored (McNaughton, 1985). In an anthropogenic grassy woodland community grazing by large mammals has a detrimental effect on the vegetation and may only be restored by elimination of livestock and smaller herbivores such as rabbits (Wimbush & Costin, 1979a).

6 1.5.1 Domestic Livestock Large ungulates were introduced into Australia approximately 170 years ago (Yates & Hobbs, 1997). Their presence has had many detrimental effects on the landscape and has been one of the most widespread and long-standing disturbances to grassy woodlands in temperate Australia (Tremont, 1994). Grazing by domestic livestock can change the structure of grassy woodland vegetation by preventing recruitment of trees and shrubs (Pettit et al., 1995), particularly in small woodland remnants (Prober & Thiele, 1995), thus transforming woodlands into grasslands (Tremont, 1994). This may "increase the risk of soil erosion, disrupt nutrient cycling, reduce habitat for animals and reduce the food source for pollinators" (Pettit et al., 1995) thus leading to major changes in ecosystem functioning (Pettit et al., 1995).

Livestock grazing can contribute to the su~cess of annual grasses and annual herbs, including many exotic species (Pettit et al., 1995). These successful species possess characteristics that allow them to thrive in highly disturbed · sites, such as seed dormancy, short life histories and early flowering (Tremont, 1994). In contrast, native perennial grasses and herbs tend to be slow growing and usually require several years to reach reproductive maturity (Tremont, 1994). Since perennial grasses and herbs make up the majority of the grassy understorey in natural grassy woodlands (Groves, 1994) and are the target food source for domestic livestock (Grice & Barchia, 1992), then a change in the grass layer composition as well as a change in vegetation structure will result in a change from woodlands to grasslands and will encourage the dominance of exotic annual species. Other herbivores that contribute to the grazing regime in many communities in Australia are rabbits and native animals such as macropods.

1.5.2 Rabbits Rabbits have the capacity to cause environmental damage that is irreparable and have contributed to the extinction of many native plant and animal species in Australia (ANZRCDP, 2001). They are non-selective grazers and will easily destroy young trees and shrubs and also compete with livestock and native herbivores for available pasture (ANZRCDP, 2001). Their grazing

7 morphology enables them to graze plants to ground level and eat roots eliminating future regeneration . Leigh et al. (1987) found that rabbits in subalpine environments substantially reduced the species richness, cover and biomass of forbs, destroyed flower and seed heads and disturbed soil by digging and scratching. Since rabbits do not favour dense vegetation (Leigh et al., 1987) and depend on grazing livestock and frequent burning to convert the grassy woodlands to a short, shrub-free area with a large proportion of introduced grasses and forbs (Myers & Parker, 1965) they are generally abundant in grassy woodlands. This may have an enormous effect on vegetation structure and diversity within the woodland community (Leigh et al. , 1987, Bennett, 1994).

1.5.3 Native herbivores

There has been little research on the g~azing capacity of native animals in vegetation communities of Australia. Most research has assessed sites with a history of introduced herbivores. Increased numbers of macropods in some areas has the capacity to alter vegetation composition and abundance due to increased grazing pressure (Neave & Tanton, 1989, Bennett, 1994). Arid woodland communities. are particularly affected by this grazing especially by the eastern grey kangaroo (A.C.T. Advisory Committee).

1.6 Fire and Grazing Interactions The interaction of fire and grazing regimes may have an influence on plant species richness that compounds the effect of each factor alone. Research indicates that grazing herbivores, both native and introduced, utilise recently burnt, regenerating areas more than unburnt areas (Groves, 1965, Bryant & Litt, 1971, Bryant, 1973, Stuwe & Parsons, 1977, Leigh & Holgate, 1979, Leigh et al., 1987, Lunt, 1990, Bennett, 1994, Noy-Meir, 1995, Henderson & Keith, 2002). Southwell and Jarman (1987) concluded that native macropods and domestic livestock utilise recently burnt areas at different times due to different feeding morphologies favouring new grass growth at certain times after regeneration. Macropods are able to forage within days after fire when new grasses emerge (Southwell & Jarman, 1987). The mouth and incisor morphology and head posture flexibility of macropod species allows them to

8 reach new grass growth. In contrast, cattle can only utilise recently burnt areas when new growth is dense and more accessible, generally after a month (Southwell & Jarman, 1987). This variation in herbivore utilisation influences vegetation structure. For example, Fensham & Skull (1999) have suggested that cattle grazed areas are richer and more diverse in perennial grasses and areas grazed by native herbivores will encourage an abundance of annual grasses.

Rabbits are also known to graze regrowth in burnt areas. Leigh et al. (1987) found that rabbit populations survived and multiplied in burnt areas but decreased in areas left unburnt. Their study found three interacting effects of fire and rabbit grazing: a greater and more prolonged exposure of bare soil due to scratching, increased grazing effects in the short term because

resprouting plants are heavily grazed ~fter the removal by fire of most available forage and increased grazing effects in the long term because of the modification of the vegetation by fire to permit greater rabbit numbers.

Fire exclusion also influences interactions between herbivores and vegetation communities in some areas. Bennett (1994) found that fire exclusion plus increased grazing pressure by rabbits and eastern grey kangaroos caused the expansion of Leptospermum laevigatum into grasslands at Yanakie Isthmus, Southern Victoria, due to an increase in the exposure of bare ground and the restriction of the feeding range of cattle (known to graze L. laevigatum). This expansion is particularly detrimental to grassland ecology because once L. /aevigatum is established it forms dense canopies restricting the growth of grasses and herbs {Bennett, 1994).

Currently, there is limited research on the effect of native herbivores following fire. Most studies to date have been short-term, three to four years (Bryant & Litt, 1971, Bryant, 1973, Leigh & Holgate, 1979, Leigh et al., 1987, Bennett, 1994, Noy-Meir, 1995, Orr et al., 1997, Lunt & Morgan, 1999a).

There is a need for long-term studies to examine vegetation structure and . change after many years of fire and grazing interactions and the cessation of

9 grazing in vulnerable areas. Studies by Wimbush and Costin (1979a,b) in the Snowy Mountains were conducted over 20 years and showed that the removal of stock and the cessation of frequent burning resulted in improved water quality in the catchment and increased the richness and diversity of native forbs. It is necessary to determine how vegetation communities are affected by altered fire and grazing regimes to ensure the conservation of threatened communities such as grassy woodlands. One study currently addressing this issue is a long-term project that was established in 1997 by the University of Wollongong (see Hill, 2000) and the Department of Defence to determine the impact of grazing and fire on the regeneration of woody shrubs and Eucalyptus species on a remnant of the Cumberland Plain Woodlands. So far, this research has suggested that grazing was more influential than fire in modifying the cover of grass and non-grass species but the long-term trends are unknown and wa_rrant further investigation.

1.7 Management of Grassy Woodlands in Australia The Australian landscape has been significantly modified by humans since European settlement (Hobbs & Hopkins, 1990), particularly the areas of grassy woodlands (Benson & Howell, 1990a,b, Yates & Hobbs, 1997, James et al., 1999), which are the most threatened habitats in Australia (Yates & Hobbs, 1997). Grassy woodlands were used for domestic livestock grazing, agriculture and tree clearing for timber inflicting many threats on the vegetation which include habitat loss, fragmentation, degradation, increased introduced species, altered fire and grazing regimes, pollution and climate change (James et al., 1999).

The western Sydney ~rea is a good example of the fate of grassy woodlands after human destruction where logging and clearing has had a significant impact of the vegetation along with the loss of grasses, herbs and shrubs due to grazing by domestic stock and rabbits and replacement by introduced species (James et al., 1999). There are many distinctive plant communities that are currently not conserved or poorly conserved with approximately 60% of species inadequately protected within the region (James et al., 1999). Many of the small remnants in the western Sydney area are considered to be

10 endangered or vulnerable, in particular, the Cumberland Plain Woodlands (James et al., 1999).

1.8 Cumberland Plain Woodlands The Cumberland Plain Woodlands were the dominant vegetation community in the western Sydney area before European settlement (Hill, 2000). It has a long history of grazing, clearing and fragmentation, which has caused a loss of over 90% of the vegetation community (James et al., 1999). It is a community where remnants differ considerably and are classified according to the dominant canopy species (James et al., 1999). In the woodlands 450 species have been recorded with more than half (230) of them classed as regionally vulnerable. The community is now considered to be rare and endangered (James et al., 1999, French et al., 2000a,b, Hill, 2000) and is listed in the NSW Threatened Species Conservation Act (1995) and the Environment Protection and Biodiversity Conservation Act (1999) as an Endangered Ecological Community.

In order to conserve the Cumberland Plain Woodlands it is imperative that the few remaining remnants are protected from further degradation and are allowed to regenerate without competition from weeds. To provide protection to this community it is necessary to understand the impact of such threatening processes by performing long-term detailed studies of the area to determine the regeneration capacity of the community and understand the consequences of changed disturbance regimes.

The Cumberland Plain Woodlands study area is located in the Holsworthy Military Area (HMA), 35 km southwest of Sydney. The HMA is an area of 18 000 ha of bushland which occurs on the boundary of the Woronora Plateau and the Cumberland Plain (French et al., 2000a, Hill, 2000). The Cumberland Plain Woodland community of the HMA is characterised by canopy species such as Eucalyptus fibrosa, E. punctata, E. globoidea and E. crebra. Other species include E. tereticornis, E. eugenioides, E. molucanna, E. longifolia and occasionally Angophora bakeri. The understorey is predominantly Themeda australis along with other grasses and herbs with sparse patches of

11 Bursaria spinosa, Dillwynia parvifolia, Pultenaea vil/osa, Daviesia ulicifolia and Lissanthe strigosa (James et al., 1999, French et al., 2000b, Hill, 2000).

1.8.1 Fire and Grazing History at Holsworthy Holsworthy has a history of frequent fires. Apart from the occurrence of wildfire it is often burnt by the activities of army personnel with most of the fires in the Small Arms Danger Area caused by tracer bullets used in small arms training activities (Hill, 2000). Also, it has been suggested that past management practices by army personnel was to burn the area regular,ly and some fires were left to burn out without any intervention (Hill, 2000).

The deep clay soils of the Cumberland Plain Woodlands were ideal for the early settlers to begin agriculture and so; much of the land was used for grazing and wheat-growing, which contin~ed until the late nineteenth century (Benson & Howell, 1990a, b). The Holsworthy area has been used as army land for the last 100 years and for the most part has been isolated from agricultural activities and grazing by domestic livestock with most grazing occurring from native herbivores and rabbits (Hill, 2000).

1.9 Aims This study is part of the long-term experiment established by the University of Wollongong in 1997 (Hill, 2000) to determine changes in species richness and abundance in a range of lifeform groups due to changed fire and grazing regimes. The aims of my study were to:

• Determine changes in species richness over six years in areas subject to· differing grazing and fire regimes in a range of lifeform groups: total species, shrubs, Fabaceae species, herbs, Asteraceae species, and Poaceae species.

• Determine changes in the abundance of species over two years in areas subject to differing grazing and fire regimes in a range of lifeform groups:

12 total species, shrubs, Fabaceae species, herbs, Asteraceae species and Poaceae species.

• Determine the effect of heat and smoke from fire on the g,ermination of Poaceae species in grassy woodlands.

The survey and germination chapters are designed as stand-alone papers and therefore some repetition is evident within the background information of the introduction and method sections. At present the Poaceae germ.ination paper has been submitted to the Australian Journal of Botany.

13 Chapter 2 The effect of fire and grazing on plant species richness and abundance

2.1 Introduction Over the last few centuries, human activities such as land clearing and utilisation have altered the natural disturbance regimes in many vegetation communities, particularly grasslands and grassy woodlands. Fire and grazing regimes in these communities have been the most affected with changes in frequency, intensity and season of disturbance causing changes in species richness and abundance (Mcintyre & Lavorel, 1994, Bond & van Wilgen, 1996, Yates & Hobbs, 1997).

A change to the natural fire regime has important implications for the survival of grassy woodland communities (Southwell and Jarman, 1987, Mcintyre and Lavorel, 1994). Burning frequency will influence the species richness of the community. If frequent burning (ie < 5 years) occurs, there will be an increase in resprouting perennials and herbaceous species such as grasses and forbs. It will also eliminate non-resprouting species, which rely solely on on-site storage of seeds (Groves & Burdon, 1986, Bock et al., 1995), and increase dominance of some species (Tothill, 1969, Bock et al., 1995, Orr et al., 1997). In contrast some research suggests that grasslands which are naturally deficient in trees and woody shrubs require frequent fire (approximately every 3 years) to maintain plant diversity by preventing species such as Themeda australis from out-competing other herbaceous species (Stuwe and Parsons, 1977, Lunt, 1994, Tremont, 1994, Morgan 1997, Lunt & Morgan, 1999a,b). This occurs by destroying individual Themeda plants and aerial plant parts, creating opportunities for seedling regeneration of other species while allowing established plants to persist (Morgan, 1998a). To completely understand how the woody and herbaceous species are affected by fire and grazing it is necessary to separate them into groups for analysis, which this study aims to do. Woody components include shrub species and Fabaceae species (which are often heat stimulated). The herbaceous component is separated into herbs, Asteraceae species and Poaceae species.

14 In general, frequent burning in grassy communities reduces the fuel load and as a consequence produces fires of low intensity (Good, 1981). Low intensity fires favour species able to regenerate vegetatively such as grasses and herbs (Christensen et al., 1981) but may not be favourable to woody shrubs and leguminous species which require less frequent, high intensity fires to break seed dormancy and encourage the germination of hard seeded species (Shea et al., 1979, Benson, 1985, Glark, 1988, Williams et al., 1994, Morrison et al., 1995, Bradstock et al., 1997).

The intensity of a fire will also be determined by the season in which it occurs and will be more intense in late summer and autumn due to the low moisture content of fuel (Christensen et al., 1981, Bond and van Wilgen, 1996). In

grassy woodlands, the season of a burn i~ less important than the effect of the burn on species richness (Lunt, 1990, HiU, 2000) but the timing of favourable rainfall and the availability of mature seeds when related to the time of a fire may influence germination in some species (Hill & French, in press).

The grazing regime of an area is an important aspect of community survival and if maintained at an intermediate level will promote the highest productivity in natural grazing ecosystems (McNaughton, 1985). In Australia, grassy woodlands and grasslands where grazing levels are high have been detrimentally affected by introduced herbivor,es and may only be restored by the elimination of livestock and smaller herbivores such as rabbits (Wimbush & Costin, 1979a,b).

The study site for this research is in the Small Arms Danger Area, Holsworthy Military Area, western Sydney (Figure 2.1 ); a unique remnant of the Cumberland Plain Woodlands that has been isolated from domestic grazing for the last 100 years. Grazing in this area is commonly by eastern grey kangaroos, common wallaroos and swamp wallabies but surveys in this area found that abundance and activity was low (Hill, 2000). There has been little research on the grazing capacity of native animals in the vegetation communities of Australia with most research assessing sites with a history of

15 introduced herbivores. High population numbers of macropods will have a detrimental effect on some vegetation communities and have the capacity to alter vegetation composition and abundance as a result of grazing pressure (Neave & Tanton, 1989, Bennett, 1994). Rabbits are also a common herbivore of the Cumberland Plain Woodlands but, like the macropods at the Holsworthy site, are present in relatively low numbers (Hill, 2000). If rabbit numbers increase, young trees and shrubs may be destroyed as the regeneration capacity of these species is reduced by rabbit grazing (ANZRCDP, 2001).

The interaction of fire and grazing regimes may have an influence on plant species richness that compounds the effect of each factor alone. Research indicates that grazing herbivores, native or introduced, utilise recently burnt, regenerating areas more than unburnt areas (Groves, 1965, Bryant & Litt, 1971, Bryant, 1973, Stuwe and Parsons, 1977, Leigh & Holgate, 1979, Leigh et al., 1987, Lunt, 1990, Bennett, 1994, Noy-Meir, 1995, Henderson and Keith, 2002).

Given that grassy woodlands experience fire and grazing regularly, it is important to determine how these vegetation communities are affected by altered disturbance regimes to ensure the conservation of these threatened communities. The aim of this research is to:

• Determine changes in species richness over six years in areas subject to differing grazing and fire regimes in a range of lifeform groups: total species, shrubs, Fabaceae species, herbs, Asteraceae species, and Poaceae species (Part A).

• Determine changes in the abundance· of species over two years in areas subject to differing grazing and fire regimes in a range of lifeform groups: total species, shrubs, Fabaceae species, herbs, Asteraceae species and Poaceae species (Part B).

16 2.2 METHODS

2.2.1 Location The study site for this project was located in the Holsworthy Military Area (HMA), western Sydney, Australia (Figure 2.1). Experimental sites were located along Fire Road 16 in the Small Arms Danger Area (SADA) (33°59"S 150°55"E) (Figure 2.2). The study area is at the boundary between the Woronora Plateau and the Cumberland Plain with the vegetation community classed as Plateau Forest on Shale, which is considered to be part of the Cumberland Plain Woodlands community (French et al., 2000a,b). The Cumberland Plain Woodlands have been classified as an endangered ecological community under the Threatened Species Conservation Act (1995) and the Environment Protection and Biodiversity Conservation Act (1999). This vegetation community is one accommodating several Eucalyptus species, scattered woody shrubs and ·a dense ground cover of herbs and grasses, particularly Themeda australis.

2.2.2 Experimental Design Forty-eight random quadrats (10 m x 10 m) were established in 1997 to investigate the effect of fire and grazing (see Hill, 2000 for technical procedures and experimental design). Thirty-two of these quadrats were randomly allocated to two fence treatments, fully fenced (CF), open (OF). The fenced treatment excluded all vertebrate herbivores . (ungrazed), while the open fence treatments were open to all vertebrate herbivores (grazed). The other sixteen quadrats were allocated to a third treatment (trampling), which is not analysed further in the grazing experiment.

Fire treatments were applied in an orthogonal design to the sites. Originally the fire treatments were unburned (UB) and control burn (CB), which were evenly applied to the forty-eight fence treatments (where n = 8). Since the beginning of the experiment, however, two separate wildfire events in summer 1997 (SB) and summer 2000 (S2) changed the experimental design. As a consequence the treatments are no longer evenly distributed between sites

17 and the unburned sites now only represent 8.3% of the fire treatments (Table 2.1).

Table 2.1: Experimental design showing the number of replications for each fire treatment.

No. Sites Code Treatment

4 UB unburnt 1997 7 CB control burn 1997 7 SB summer burn 1997 3 UBS2 unburnt 1997 + summer burn 2000 9 CBS2 control burn 1997 +summer burn 2000 2 SBS2 summer burn 1997 + summer burn 2000

2.2.3 Surveys Surveys of all quadrats were conducted in 1997 and 1999 then during August to December 2001 and October to December 2002. Presence/absence data were recorded for all species in the experimental sites in all years (Part A), however, cover abundance was only recorded in 2001 and 2002 (Part B) therefore cannot be compared to previous years. Unknown species were collected and identified at the Janet Cosh Herbarium using the Flora of NSW botanical key to species (Harden, 1990-1993). The Braun-Blanquet scale of cover abundance (CA) (Table 2.2) was used to score the cover abundance of each species. Also, the dominant species for each vegetation layer was recorded for all quadrats. Table 2.2: Values used to record cover abundance (CA) of species in survey sites based on the Braun-Blanquet scale (Poore, 1955).

CA Value Description 1 one-a few individuals 2 uncommon and < 5% cover 3 common and < 5% cover 4 very abundant and < 5% or 5-20% cover 5 20-50% cover 6 50-75% cover 7 75-100% cover

18 2.2.4 Total and Mean Species Richness Part A, which aims to determine changes in species richness over six years, includes illustrations of how different components of the flora have changed over the six-year study period including the total and mean species richness. Also, changes in the number of Asteraceae and Orchidaceae species are shown. Orchidaceae species could not be analysis further due to limited numbers in some years. Asteraceae species are further analysed using PRIMER (Plymouth Routines in Multivariate Ecological Research).

2.2.5 Rainfall Average annual rainfall data from each survey year was compared to mean species richness to indicate how changes in rainfall may be associated with changes in species richness, this was not formally analysed but represented graphically.

2.2.6 Statistical Analysis Non-metric multidimensional scaling (nMDS) was used to visualise fence and fire treatment groupings of quadrats according to similarity in species richness, using Bray-Curtis indices of similarity through PRIMER (Carr, 1996). The ordination was performed on: • Total species • Shrubs • Fabaceae species • Herbs • Asteraceae species • Poaceae species

Analysis of similarity (ANOSIM) was used to determine differences in the composition (Part A) and cover abundance (Part B) of species within the fence treatment groups and the fire treatment groups with pairwise tests indicating significant differences between treatments. Similarity percentage analysis (SIMPER) of all floristic groups was used to determine the similarities within fence and fire treatments in each year and which species were characteristic of these groups using pairwise comparisons. SIMPER also

19 determines which species were the major contributors to the average dissimilarity between each treatment type.

Analysis of variances (ANOVA) was used in part A to determine the change in species richness between 1997 & 1999 due to grazing/fire interactions using SYSTAT statistical software. Tukey's HSD was used to determine where differences lay in significant results (Zar, 1999). This analysis could not be performed on data after 1999 due to the wildfire of 2000 changing the experimental design and causing the loss of degrees of freedom.

ANOVA was used to determine differences in the mean percentage cover of Themeda australis between the grazing and fire treatments in part B. Tukey's HSD was used to determine where differences lay in the significant fire treatments (Zar, 1999).

20 Figure 2.1: Location of Holsworthy Military Area, western Sydney, Australia. +N

Australia

21 Figure 2.2: Survey sites showing corresponding fire treatments. + Holsworthy Military Area ----r--t'7tl 0 0.4 - -- - Small Arms Danger Area

•• ••

• • .: ·• .. ••

• •

••• •••• • •• •••

Fire Treatments Holsworthy Vegetation Types D Untx.Jrnt aeared Cortrol Burnt •D Woodland/heath complex Aateaux forest Summer Burn D • GJHy forest • Untx.Jrnt/Summer Burn 2000 Heath co111>lex •D Cortrol Bum/Summer Burn 2000 • Sedge land Melaleuca thicket Summer Burn/Summer Burn 2000 •D I II I

22 2.3 RESULTS - Part A - Changes in Species Richness 1997 - 2002

2.3.1 Overview of the Yearly Changes in Total Species Richness A total of 190 species has been recorded in the 48 quadrats over the six-year period from the beginning of the experiment in 1997. Richness in these grassy woodlands occurs at ground level with the most speciose groups being the herbs (69), shrubs (43) and grasses (26) (Figure 2.4).

The proportion of species in different groups within the community has changed over the six-year study (Figure 2.3). In 1997 the shrub layer was the major community component (32%), which reduced to 24-26% after 1999. The main reason for the reduction , was not a loss of shrub species (Figure 2.4), but an increase in the diversity of the herbaceous layer, which increased from 25% in 1997 to 35-39% from 1999 (~igure 2.3 and Figure 2.4). The most obvious changes are an increase in species from the Orchidaceae and Asteraceae families up to 2001 followed by a decrease in numbers in 2002 (Figure 2.5). The grass layer has ranged between 15 and 19 % over the study period, but appeared to increase in species richness during this time (Figure 2.4). Tree composition has declined from 11% in 1997 to 9% in 1999, 6% in 2001 and 7% in 2002. The others category, which includes climbers, ferns and sedges has remained constant over the study period, although the number of species increased by 50% from 1997 to 2002. This includes four species of Lomandra recorded in 2001. The weed composition ranged between 8 and 9 % between 1997 and 2001 then declined to 5 % in 2002 (Figure 2.3), although a few more species were found (Figure 2.4).

23 Weeds Weeds 9 0,1o 8% l Others _:..-.--- Others 6 01c,

11 % Trees 35%

,,,,,,,,,.,h,,,,,,,,,,,,,,,,,,,,,,,,JJ;'J/Jl/J///Jh'l'/h'/J/,,,N'/ /J//J,//Jh'//JJ/J//h,,N/hl/.

Grasses 18% 16%

32% Shrubs 25%

a b

Weeds ~ Wee~ Others 5% \ ------

6% 39% 7%

Grasses 15% 19%

Shrubs 24% Shrubs 26%

c d

Figure 2.3: Species richness of the community in the years (a) 1997, (b) 1999, (c) 2001 , (d) 2002. en en 80 c:QI 70 .!::: u 60 • 1997 cc 50 en 01999 QI 40 ·2 30 02001 c. en 20 132002 m 10 -0 I- 0 Herbs Shrubs Trees Grasses Weeds Others Lifeform

Figure 2.4: Total number of species of each major lifeform category present in each survey year.

xi 20 ·c:; Cll 15 fli • Asteraceae D Orchidaceae ~ 10 .c E :::I z 5

0 1997 1999 2001 2002 Year

Fig.ure 2.5: Number of species recorded from the Asteraceae and Orchidaceae fam ilies for each survey year. Mean species richness was highest in 1999 and the total species richness was highest in 2001 compared to other years. Mean and total species richness were both reduced in 2002 compared to 1999 and 2001 (Figure 2.6).

i 180,--~~~~~~~~~~ 0 160 ~ 140 ~ 120 '(j 100 Q) •MSR ~ 80 OTSR 'O 60 ~ 40 ~ 20 z::I 0 1997 1999 2001 2002 Year

2 Figure 2.6: Mean (± s.d.) and total species richness (100 m ) in the 48 sites for each survey year. MSR = Mean Species Richness, TSR =Total Species Richness.

2.3.2 Rainfall Rainfall patterns between 1996 and 2002 have varied (Figure 2.7). Average annual rainfall in the Liverpool, western Sydney area over the last 100 years ranged from 600 mm - 900 mm. 1997 experienced a particularly low annual rainfall (562.1 mm) and a relatively dry period between May and August (120.7 mm) when the vegetation surveys were conducted. Conversely, in 1999 the same period received 214.4 mm rainfall and experienced an above average annual rainfall (1048.6 mm). Rainfall was average between 2000 and 2002 with 2002 experiencing a very dry period between June and November with a total rainfall of 51.6 mm in this period when sampling was conducted (Figure 2. 7).

26 1200

1000 e 800 .§. 600 ~ c: 'iU 0::: 400 200

0 1996 1997 1998 1999 2000 2001 2002 Year

Figure 2.7: Annual rainfall (mm) for the years 1996-2002 in the Holsworthy Military Area taken from the Michael Wenden Centre, Liverpool.

Changes in annual average rainfall appear to correspond with changes in species richness. High rainfall was associated with increased mean species richness and reduced rainfall in the following years was reflected in a decline in mean species richness (Figure 2.8).

60 1000 .!!!"' 900 (.) 50 800 -E G) E c. 700 ....U) 40 - 0 600 J! ... c -MSR G) 30 500 ·; .c 0::: ~Rainfall E 400 ::I 20 ns z 300 ::I c c ns 200 c G) 10 c( ::!: 100 0 0 1997 1999 2001 2002 Year

Figure 2.8: Mean (± s.d.) number of species (MSR) per plot in 1997, 1999, 2001 and 2002 with corresponding annual rainfall (mm).

27 2.3 .. 3 Differences Among Fence and Fire Treatments The mean species richness (MSR) in 1997 was the same for both fence treatments (F1 31 =0.0788, p=0.7809). The MSR was higher in the ungrazed ' sites in all other years but was only significant in 2001 (1999 F =3.3288, . 1' 31 p=0.0784, 2001 F1,31=4.1916, p=0.0495, 2002 F1,31=2.5013, p=0.1246) (Figure 2.9).

I/) 60 ..,..------~ Q) ·~ 50 Q. ~ 40 0 ~ 30 !oCFl J:I ~ zs 20 ; 10 Q)

:!: 0 -+---'"-- 1997 1999 2001 2002 Year

Figure 2.9: Mean(± s.d.) number of species in each fence treatment in 1997, 1999, 2001 and 2002. CF= closed fence treatment (no grazing) and OF= open fence (grazing) (n=16).

AVOVA could not be performed on the fire treatment data after 1999 due ·to increased fire treatments after a wildfire in 2000 and the subsequent loss of degrees of freedom. However, the mean species richness for fire treatments in 1999, 2001 and 2002 indicated that on average, the unburnt sites contained over 20% fewer species in 2001 and over 30% fewer in 2002 than burnt sites. In all sites species richness was higher in 1999 and declined in 2001 and 2002 (Figure 2.10).

28 60 Cl) cu ·c:; 50 cu Q. ! 40 0 •1999 ; 30 .Q •2001 20 02002 §c c ns cu 10 ::E 0 UB CB SB UBS2 CBS2 SBS2 Fire Treatment

Figure 2.10: The mean (± s.d.) number of species recorded in each fire treatment group in 1999, 2001 and 2002. UB=unburnt, CB=control burn, SB=summer burn (1997 wildfire), UBS2=unburnt + summer burn2 (2000 wildfire), CBS2=control burn + summer burn2 (2000 wildfire), SBS2=summer burn+ summer burn2 (2000 wildfire).

2.3.4 Multivariate Analysis of Presence/Absence Data

2.3.4.1 Total Species Although in the initial survey in 1997 there were significant differences in species richness between fence types (Global R =0.098, p =0.015), this was not evident in 1999 (Global R = 0.006, p = 0.386), 2001 (Global R = 0.026, p = 0.237) or 2002 (Global R = 0.022, p = 0.271 ). Average dissimilarity (Bray Curtis indices) between grazed and ungrazed sites fell only marginally from 54% in 1997 to 48.4-52.6% in subsequent years.

Differences between fire treatments were not evident until 2001 (Global R = 0.144, p = 0.006) and 2002 (Global R = 0.148, p = 0.038). In general, differences were largely due to unburnt areas differing from a range of burning regimes and sites that had experienced the initial summer burn differing from sites where two burns occurred (SB v CBS2) (Table 2.3).

29 Table 2.3: SIMPER showing the average dissimilarity (Ave Diss) between the species richness of Total species for 2001 and 2002 showing the most discriminating species. Diss/SD = standard deviations between dissimilarity indices, Contrib (%)-percent contribution to the dissimilarity index, cum %=cumulative percent. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 + summer burn 2000. Bold indicates a strong discriminator between treatments.

Group Ave Diss Species Average Abundance Diss/SD Contrib (%) Cum(%)

Total species 2001 UB CBS2 UB CBS2 54.99 Pultenaea villosa 0 1.75 2.15 2.08 2.08 Wahlenbergia gracilis 0 1.92 2.18 2.03 4.11 Eucalyptus moluccana 0.75 0 1.69 1.88 5.99 Microlaena stipoides 2.5 1.69 1.86 7.85 Goodenia hederacea 1.25 0.17 1.52 1.76 9.61 UB UBS2 UB UBS2 52.26 Pultenaea villosa 0 1.8 1.36 2.6 2.6 Eucalyptus fibrosa 0 1 1.36 2.6 5.19 Eucalyptus tereticomis 0.75 0 1.02 1.96 7.15 Eucalyptus moluccana 0.75 0 1.02 1.96 9.1 Poa labil/ardieri 1.25 0 1.01 1.93 11 .03 UB CB UB CB 54.65 Eucalyptus tereticomis 0.75 0.17 0.91 1.67 1.67 Eucalyptus moluccana 0.75 0.17 0.91 1.66 3.33 Microlaena stipoides 1 1.83 0.91 1.66 4.99 Eucalyptus crebra 0 0.67 0.90 1.66 6.64 Hibbertia obtusifolia f .75 0.42 0.85 1.56 8.21 Pratia purpurescens 0.25 1.42 0.85 1.56 9.76 UB SB UB SB 54.1 Persoonia linearis 0 0.92 0.99 1.83 1.83 Eucalyptus moluccana 0.75 0 0.98 1.81 3.65 Diane/la longifolia 0.25 3 0.96 1.78 5.43 Hibbertia obtusifolia 1.75 0.15 0.92 1.7 7.13 Goodenia hederacea 1.25 0.23 0.87 1.6 8.73 Exocarpos strictus 0 1.54 0.86 1.6 10.33 ______, CBS2 SB CBS2 SB 49.59 Pultenaea villosa 1.75 0.54 0.89 1.79 1.79 Astroloma humifusum 0 1.31 0.83 1.67 3.46 Persoonia linearis 0.42 0.92 0.72 1.45 4.91 Glycine microphy/la 0.42 1.54 0.72 1.45 6.36 Exocarpos strictus 0.25 1.54 0.71 1.44 7.8 Hibbertia aspera 0.75 1.85 0.70 1.4 9.2 Acacia falcata 0.83 0.38 0.67 1.35 10.55 Total species 2002 UB CBS2 UB CBS2 46.21 Pultenaea villosa 0 1.33 1.67 2.66 2.66 Goodenia hederacea 0.75 0 1.67 2.63 5.29 Eucalyptus eugenioides 0.75 0.17 1.37 2.36 7.65 Cymbopogon refractus 0.75 0.25 1.25 2.28 9.93 CBS2 SB CBS2 SB 49.14 Astroloma humifusum 0 1.46 1.74 2.3 2.3 Persoonia /inearis 0.25 1.15 1.53 2.13 4.43 Pultenaea villosa 1.33 0.62 1.39 2.02 6.45 Exocarpos strictus 0.25 1.46 1.28 1.91 8.36 Echinopogon caespitosus 0.17 1.15 1.26 1.88 10.24 UBS2 SB UBS2 SB 49.93 Pultenaea villosa 1.4 0.62 2.15 2.4 2.4 Exocarpos strictus 0 1.46 1.71 2.17 4.57 Astroloma humifusum 0 1.46 1.71 2.17 6.75 Persoonia /inearis 0.2 1.15 1.68 2.17 8.91 Echinopogon caespitosus 0.2 1.15 1.21 1.76 10.68

30 Various species were shown to be the main discriminators between significant pairs within the community but few were found to be strong discriminators (Table 2.3). Strong discriminators were defined as those species that have dissimilarity to standard deviation ratio greater than 1.5 (Clarke & Warwick, 1994). However, three species were good indicators of particular fire treatments. The occurrence of Pultenaea vil/osa is most commonly associated with sites that were burnt in summer 2000, while Hibbertia obtusifolia and Goodenia hederacea occurred more often in the unburnt sites.

2.3.4.2 Shrubs When shrubs were considered separately, there were no differences in species richness between grazed and ungrazed sites in any year (Global R = 0.041-0.068, p = 0.092-0.159). Similarly, the Fabaceae family, which are chiefly shrub species, showed no differe".lces between grazed and ungrazed sites throughout the experiment (Global R = 0.025-0.068, p = 0.063-0.225).

For comparisons amongst fire treatments, analysis showed that there were significant differences between the UBS2 and CBS2 sites when compared to the SB sites in 2001 (Global R = 0.159, p = 0.019) and 2002 (Global R = 0.173, p = 0.015). In a similar way to the analysis of total species, Pultenaea villosa is one of the main discriminators between fire regimes with this species occurring mostly in the S2 sites when compared to SB sites. Astroloma humifusum is also a strong discriminator and only occurs in the sites burnt in summer 1997 when compared to the UBS2 and CBS2 sites. Exocarpos strictus and Persoonia linearis are also strong discriminators between the UBS2 and SB sites and are mostly associated with the SB sites (Table 2.4).

31 Table 2.4: SIMPER showing the average dissimilarity (Ave Diss) between the species richness of shrub species for 2001 and 2002 showing the top 30% of species. Diss/SD = standard deviations between dissimilarity indices, Contrib (%)-percent contribution to the dissimilarity index, cum %=cumulative percent. Fire treatment symbols are; SB=summer burn 1997, UBS2=unburnt +summer burn 2000, CBS2=control burnt 1997 +summer burn 2000. Bold indicates a strong discriminator between treatments.

Group Ave Diss Species Average Abundance Diss/SD Contrib (%) Cum(%)

Shrubs 2001 CBS2 SB CBS2 SB 66.03 Pultenaea villosa 1.75 0.54 1.45 7.72 7.72 Astroloma humifusum 0 1.31 1.33 7.58 15.3 Exocarpos strictus 0.25 1.54 1.13 6.16 21.47 Hibbertia aspera 0.75 1.85 1.03 6.07 27.53 UBS2 SB UBS2 SB 63.26 Pultenaea villosa 1.8 0.54 1.96 9.23 9.23 Astroloma humifusum 0 1.31 1.35 7.81 17.04 Exocarpos strictus 0 1.54 1.40 7.33 24.37 Bursaria spinosa 1.4 0.31 1.10 6.40 30.77

Shrubs 2002 CBS2 SB CBS2 SB 68.87 Astroloma humifusum 0 1.46 1.56 7.95 7.95 Persoonia linearis 0.25 1.15 1.44 7.22 15.18 Pu/tenaea villosa 1.33 0.62 1.28 7 22.18 Exocarpos strictus 0.25 1.46 1.19 6.62 28.79 UBS2 SB UBS2 SB 70.65 Pultenaea villosa 1.4 0.62 1.99 8.19 8.19 Exocarpos strictus 0 1.46 1.60 7.38 15.56 Astroloma humifusum 0 1.46 1.60 7.38 22.94 Persoonia linearis 0.2 1.15 1.61 7.06 30

2.3.4.3 Herbs and Poaceae

As with total species, the composition of herbaceous species was initially different between grazed and ungrazed sites one year after the fences were erected in 1997 (Global R =0.09, p =0.032). This difference was not evident in 1999 (Global R = 0.008, p =0.325), 2001 (Global R =-0.003, p =0.487) or 2002 (Global R = 0.008, p = 0.366). The main contributing species to the initial differences were a greater occurrence of Goodenia hederacea in ungrazed sites and a greater occurrence of Dichondra repens, Pratia purpurescens and Senecio lautus in grazed sites but none were strong discriminators (Table 2.5).

32 Table 2.5: SIMPER table of the species richness of herb species (1997) showing significant differences between grazed and ungrazed sites. Species shown are the top 30% of the average dissimilarity (Ave Diss) between sites. CF= closed fence, OF= open fence. Diss/SD = standard deviations between dissimilarity indices, Contrib % = the percent contribution to the dissimilarity index, Cum % = cumulative %. Bold indicates a strong discriminator between ·treatments.

Ave Diss Species Average Abundance Diss/SD Contrib (%) Cum(%) CF OF

48.41 Goodenia hederacea 0.56 0.19 1.05 6.64 6.64 Dichondra repens 0.38 0.63 1.04 6.45 13.09 Pratia purpurescens 0.38 0.56 1.01 6.2 19.29 Senecio lautus 0.25 0.56 1.04 6.14 25.43 Opercularia diphylla 0.5 0.69 0.97 6.05 31.48

When Asteraceae species, which are c~iefly herbaceous, were considered separately no differences between grazed and ungrazed sites were found (Global R = -0.005-0.051, p = 0.111-0.503).

A difference in the grasses between grazed and ungrazed sites was evident in 1997 (Global R =0 .077, p = 0.048) and 2001 (Global R =0.106, p =0.02), but no difference was evident in 1999 (Global R = -0.019, p = 0.635) or 2002 (Global R = -0.003, p = 0.456). The species contributing to differences in 1997 and 2001 were not the same, suggesting that the community was dynamic in the occurrence of species on an annual basis. Initially, Panicum simile and Aristida vagans were recorded more often in the grazed sites while Entolasia stricta was recorded more often in the ungrazed sites. Four years later, the differences were a greater occurrence of P. simile, Paspalidium distans and Austrodanthonia tenuior in grazed sites and a greater occurrence of Diche/achne micrantha in ungrazed sites but again, no species were strong discriminators between grazed or ungrazed sites in 1997 or 2001 (Table 2.6).

33 Table 2.6: SIMPER table of the species richness of Poaceae species (1997 & 2001) showing significant differences between grazed and ungrazed sites. Species shown are the top 30% & 50% of the average dissimilarity (Ave Diss) between sites. CF = closed fence, OF = open fence. Diss/SD = standard deviations between dissimilarity indices, Contrib % = the percent contribution to the dissimilarity index, Cum % = cumulative %. Bold indicates a strong discriminator between treatments.

Group Ave Diss Species Average Abundance Diss/SD Contrib (%) Cum(%) CF OF

Poaceae 1997 52.62 Panicum simile 0.25 0.63 1.05 20.6 20.6 Aristida vagans 0.25 0.5 0.93 18.24 38.84 Entolasia stricta 0.25 0.13 0.65 10.19 49.03

Poaceae 2001 51 .69 Panicum simile 0.69 1.75 1.04 7.59 7.59 Paspalidium distans 0.69 0.88 0.94 6.44 14.03 Dichelachne micrantha 1.19 0.69 0.93 6.33 20.36 Austrodanthonia tenuior 0.38 0.94 0.94 6.01 26.37 Entolasia stricta 0.88 0.88 0.86 5.94 32.3

There were no significant differences between fire treatments in the Asteraceae or other herbaceous species in any year (Asteraceae: Global R = 0.001-0.01, p = 0.442-0.487, Herbs: Global R = 0.013-0.061, p = 0.197- 0.543). The grass species showed significant differences between various fire treatments in 1999 (Global R =0.13, p = 0.027), 2001 (Global R =0.087, p = 0.048) and 2002 (Global R = 0.104, p = 0.027). The species responsible for differences showed no clear pattern in response to the fire treatments (Table 2.7).

Micro/aena stipoides was shown to be a strong discriminator species in 2001 occurring more often in the UBS2 and CBS2 sites when compared to the UB sites. Poa /abil/ardieri was a strong discriminator species between the UB and UBS2 sites in 2001 and occurred more often in the UBS2 sites. All other species were poor discriminators between fire treatments (Table 2.7).

34 Table 2.7: SIMPER showing · the average dissimilarity (Ave Diss) between the species richness of Poaceae species for 1999, 2001 and 2002 showing the top 30% of species. Diss/SD = standard deviations between dissimilarity indices, Contrib (%)-percent contribution to the dissimilarity index, cum %=cumulative percent. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 + summer burn 2000. Bo'ld indicates a strong discriminator between treatments.

Group Ave Diss Species Average Abundance Diss/SO Contrib (%) Cum(%)

Poaceae 1999 CB SB CB SB 41 .61 Echinopogon caespitosus 0.21 0.73 1.26 8.32 8.32 Paspalidium distans 0.67 0.4 1.04 7.25 15.57 Poa /abillardieri 0.54 0.53 0.94 6.73 22.3 Digitaria ramularis 0.33 0.47 0.95 6.26 28.56 Poaceae 2001 UB CBS2 UB CBS2 56.89 Microlaena stipoides 2.5 1.52 10.31 10.31 Poa /abillardieri 1.25 0.08 1.37 9.71 20.02 Paspalidium distans 0.75 0.67 1.01 7.02 27 .04 Entolasia stricta 1.25 0.33 0.92 6.47 33.51 UB UBS2 UB UBS2 53.26 Microlaena stipoides 2.8 1.63 9.35 9.35 Poa Jabillardieri 1.25 0 1.63 9.35 18.69 Panicum simile 0.75 2 0.97 6.73 25.43 Digitaria ramularis 0.25 1.04 6.34 31.77 UB CB UB CB 57 .65 Microlaena stipoides 1.83 1.33 8.55 8.55 Poa /abillardieri 1.25 0.33 1.22 7.91 16.47 Dichelachne micrantha 0 1.33 1.11 6.95 23.42 Panicum simile 0.75 1.17 0.96 6.43 29.84 CB SB 0.98 CB SB 51 .3 Dichelachne micrantha 1.33 0.85 1.01 6.92 6.92 Paspalidium distans 0.83 0.62 0.80 6.87 13.79 Entolasia stricta 0.25 1.15 0.85 5.81 19.6 Poa /abillardieri 0.33 0.38 0.88 5.78 25.37 Ento/asia marginata 0.08 0.77 5.66 31 .04 Poaceae 2002 CBS2 UBS2 CBS2 UBS2 47.9 Entolasia stricta 0.5 1.6 1.13 7.73 7.73 Austrodanthonia racemosa 0.08 0.8 1.10 7.68 15.41 Entolasia marginata 0 1.13 7.53 22.94 Poa labillardieri 0.25 0.6 1.07 6.51 29.45 CBS2 SB CBS2 SB 43.87 Paspalidium distans 0.25 1.10 8.15 8.15 Ento/asia stricta 0.5 1.54 1.02 8.06 16.21 Echinopogon caespitosus 0.17 1.15 1.23 8.04 24.25 Cymbopogon refractus 0.25 1.08 1.15 7.79 32.04 UBS2 SB UBS2 SB 44.86 Paspalidium distans 0.2 1.19 6.67 6.67 Echinopogon caespitosus 0.2 1.15 1.22 6.45 13.12 Ento/asia marginata 0.54 1.02 6.14 19.26 Austrodanthonia tenuior 0.6 0.15 1.11 6.02 25 .28 Poa /abillardieri 0.6 0.38 1.08 5.93 31 .21

35 2.3.5 Grazing/Fire Interactions An analysis of the interaction between grazing and fire treatments could only be performed on the 1999 data due wildfire burning many fire treatment sites in 2000. This changed the experimental design by reducing the number of replications and the ability to analyse them statistically due to a subsequent loss of degrees of freedom. I investigated the interaction by testing the difference in the change of species richness from 1997 to 1999 across fire treatments using ANOVA. There was a greater change in the ungrazed compared to the grazed sites (F1 26=4.802, p=0.038) and an interaction ' between fire and grazing (F1 26=8.070, p=0.002). Tukey's HSD indicates that ' the change in species richness between fence treatments was significantly different in the summer burn (SB) sites (q6,26=4.373<6.20). There were no significant differences between fence treatments in the unburnt (UB) sites

(q6,26=4.373>0.46) or the control burn {CB) sites (q6,26=4.373>0.62) (Figure 2.11).

a a a a b a :(l N' 40 ·c:; E 8- g 30 (J) !:.. •CF C: UI 20 ·- UI Q) Q) OOF Clc: .cc: 10 CIS CJ .c ·- 0 0:: 0 UB CB SB Fire Treatment

Figure 2.11: Change in species richness in fire and grazing treatments in 1999 compared to 1997. UB=unburnt 1997, CB=control burn 1997, SB=summer burn 1997. CF=closed fence, OF=open fence. Different letters represent a significant difference via Tukey's HSD test (p=0.05).

36 2.4 RESULTS - Part B - Changes in Cover Abundance 2001 - 2002

2.4.1 Multivariate Analysis of Cover Abundance Data

2.4.1.1 Total species For both years there were no differences in species cover abundance between grazing treatments (2001: Global R=0.035, p=0.175, 2002: Global R=0.027, p=0.219). In contrast, there were significant differences in cover abundance between fire treatments in 2001 (Global R=0.15, p=0.003) and 2002 (Global R=0.187, p=0.013). In general, the sites burnt in summer 2000 are different to other sites (Table 2.8, Figure 2.12).

Table 2.8: Analysis of similarities (ANOSIM) of the untransformed cover abundance data for 2001 and 2002 in the fire treated sites. Pairwise tests are presented for the sites that showed significant differences. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 +summer burn 2000.

Total species Fire ANOSIM

Year Global Test Pairwise Tests Global R p Groups R Statistic p

2001 . 0.15 0.003 UBxCB 0.392 0.012 UBXSB 0.452 0.008 UB x CBS2 0.518 0.006 UB x UBS2 0.494 0.016 SB x CBS2 0.161 0.004

2002 0.187 0.013 SB x CBS2 0.386 0.008

37 Total Species 2001 ------, ~ : r~~; : O :?'< A ua

Y CBS2

II SBS2 t- • ussz • cs • • • • f< SB

Total Species 2002 .. • Stress:• 0.2 ... UB • ,. Tw- CBS2 • ,. • ... D 8882 ' y• •... ' ...... ,. . • UBS2 T . t" • • + .... "' .. + ...... • CB • + + ... • • • t+- • .. 88

Figure 2.12: nMDS ordination sites of fire treatments for the cover abundance of total species in 2001 and 2002. Fire treatment symbols are; us=unburnt, CB=control burnt 1997, SB=surnrner burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 +summer burn 2000, SBS2=summer burn 1997 + summer burn 2000.

2.4.1.2 Shrubs

In 2001, there were significant differences between grazing treatments (Global R = 0.091, p= 0.048). Daviesia u/icifolia, Bossiaea prostrata and Lissanthe strigosa were greater in cover abundance in ungrazed sites while Pultenaea vil/osa has typically higher cover abundance values in the grazed sites (Table 2.9). No significant differences occurred in 2002 (Global R = 0.061, p=0.111) (Figure 2.13).

38 Table 2.9: Similarity percentages (SIMPER) of untransformed 2001 data showing the average abundance and the standard deviation between dissimilarity values (Diss/SD) of shrub species between closed fence (CF) and open fence (OF) treatments. Contrib %=percentage contribution to the dissimilarity index, Cum %=cumulative percent.

Shrubs 2001 Cover Abundance

Species Average Abundance Diss/SD Contrib. % Cum.%

CF OF

Daviesia ulicifolia 2.31 0.94 1.38 10.32 10.32 Bossiaea prostrata 1.38 0.5 1.06 7.32 17.63 Lissanthe strigosa 1.38 0.69 1.1 7.29 24.92 Pultenaea villosa 0.75 1.19 0.96 7.17 32.1

Stress: D.17 Shrubs 2001 c • • c D • D CF D D c D • D c• c• D °a •• • D • D • • c • • • • OF •

Stres4 0.18 Shrubs 2002 D c • • o CF • D • • D c • D • • c• • D D c • D D D • • OF • • D a • c

Figure 2.13: nMDS ordination sites of fence treatments for the cover abundance of shrub species in 2001 and 2002. Fence treatment symbols are CF = closed fence, OF = open fence.

39 As with the analysis using total species, differences occurred in the species richness of shrubs between fire treatments in 2001 (Global R= 0.143, p=0.047) and 2002 (Global R=0.144, p=0.047 (Table 2.10). In general, sites that were only burnt in the fire in summer 1997 (SB) were different in shrub cover abundance to sites burnt in 2000 (Table 2.10, 'Figure 2.14). Sites that were burnt in 2000 were distinguished by a greater and more consistent cover of Pultenaea vi/losa. Astroloma humifusum was absent in sites burnt in 2000 where it had been present in 1999. Similarly, Exocarpos strictus was lower in abundance after the 2000 burn. Interestingly, these two species are vertebrate dispersed and predicted to be negatively affected by fire (French & Westaby, 1996).

Table 2.10: Similarity percentages (SIMPER) of untransformed 2001and 2002 data showing the average abundance and the standard deviation between dissimilarity values (Diss/SD) of shrub species between fire treatments. Contrib %=percentage contribution to the dissimilarity index, Cum %=cumulative percent. Fire treatment symbols are; SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 +summer burn 2000.

Group Ave Diss. Species Average Abundance Diss/SD Contib (%) Cum(%}

2001 CBS2 SB CBS2 SB 74.7 Pu/tenaea vi/Josa 1.75 0.54 1.38 9.87 9.87 Hibbertia aspera var. aspera 0.75 1.85 1.26 8.42 18.29 Daviesia ulicifolia 1.17 1.62 1.19 7.6 25.89 Exocarpos strictus 0.25 1.54 1.19 7.45 33.34 2002 CBS2 SB CBS2 SB 77.41 Pultenaea villosa 1.33 0.62 1.32 8.83 8.83 Hibbertia aspera var. aspera 0.33 1.62 1.46 8.27 17.1 Daviesia ulicifolia 0.92 1.77 1.23 8.19 25.29 Astroloma humifusum 0 1.46 1.35 8.09 33.38 UBS2 SB UBS2 SB 77.26 Pultenaea villosa 1.4 0.62 2.17 9.09 9.09 Exocarpos strictus 0 1.46 1.22 8.34 17.43 Daviesia ulicifolia 1 1.77 1.36 8.18 25.61 Astroloma humifusum 0 1.46 1.35 8.13 33.74

40 l!~!SS: IJ 1 Shrubs 2001 • .. UB " • • ., C6S1 t' + • · ~ • • + • + • t II 8951 t- A + I"+ • ~ i-• ., • • UliS2 T ... "' l' 't' I • • ,. • ... .. ; CB • • • • ., t t- sa

Sin· ~. ! ': 0. Hl Shrubs 2002 • ... l.JB • + ..,.,. .,. · ~ • ...... T + +.,. • y CB52 + .A...... II :S8S:2 • • t- • 'f T• • ., • Ul:S2 • • +• .... • • .. ' . CG A • • .. • • y T ' + sa

Figure 2.14: nMDS ordination sites of fire treatments for the cover abundance of shrub species in 2001 and 2002. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 +summer burn 2000, SBS2=summer burn 1997 + summer burn 2000.

Fabaceae were analysed separately to investigate changes within this speciose shrub family. In 2001 there were significant differences in the cover abundance of Fabaceae species between fence treatments (Global R = 0.114, P.= 0.023) (Table 2.11). Daviesia ulicifolia and Glycine microphyl/a

~ were more cover abundant in the ungrazed sites while Glycine tabacina was more cover abundant in the grazed sites, although there was variability among sites. Dissimilarity/SD values are relatively low indicating that these changes are not likely to be good indicators of differences between grazed and ungrazed sites (Table 2.11) (see Clarke & Warwick 1994). No significant differences occurred in 2002 (Global R = 0.074, p=0.064) (Figure 2.15).

41 Table 2.11: Similarity percentages (SIMPER) of untransformed 2001 data showing the average abundance and the standard deviation between dissimilarity values (Diss/SD) of Fabaceae species between closed fence (CF) and open fence (OF) treatments. Contrib %=percentage contribution to the dissimilarity index, Cum %=cumulative percent.

Fabaceae 2001 Untransformed SIMPER

Species Average Abundance Diss/SD Contrib. % Cum.% CF OF

Daviesia u/icifolia 2.31 0.94 1.34 11 .24 11.24 Glycine tabacina 1.19 2.25 1.22 9.65 20.89 Glycine microphylla 1.56 0.75 0.99 9.39 30.28

Stress: 0.19 • c c Fabaceae 2001 c CF [J • a c • • c a c a• . a •' • a • OF • c • • • c

Figure 2.15: nMDS ordination sites of fence treatments for the cover abundance of Fabaceae species in 2001. Fence treatment symbols are CF= closed fence, OF = open fence.

There were no significant differences in the species cover abundance of Fabaceae between fire treatments (2001: Global R=0.004, p=0.497; 2002: Global R=0.098, p=0.117. This result differs from the result for all shrub species, suggesting that fire is influencing species from other families rather than species within this fire-adapted family. Pultenaea vil/osa was still evident, however, as being the most different in cover between fire treatments.

42 2.4.1.3 Herbs There was no significant difference in the cover abundance of herbs between grazing treatments in 2001 (Global R=-0.001, p=0.455) and 2002 (Global R=0 .022, p=0.268). Similarly, for fire treatments there were no significant differences in the cover abundance of herbs between sites in 2001 (Global R=-0 .079, p=0 .843) or 2002 (Global R=0.085, p=0.122).

Asteraceae species, which are mainly herbaceous, were considered separately but only analysed for 2001 due to a lack of species of this family recorded in 2002. There was no difference in species richness between fence treatments (Global R=0 .005, p=0.518). Similarly, there was no difference in species richness between fire treatments (Global R=0.001, p=0.479).

2.4.1.4 Poaceae There were significant differences in species cover abundance between grazing treatments in 2001 (Global R=0.066, p=0.046). The results indicate that Themeda australis was more cover abundant in the ungrazed sites, and Micro/aena stipoides and Panicum simile were more cover abundant in the grazed sites (Table 2.12), although the Diss/SD is low, indicating that changes in abundance of these species are not good at distinguishing between fence types. There were no significant differences in cover abundance between fence treatments in 2002 (Global R=-0.018, p=0.709).

Table 2.12: Similarity percentages (SIMPER) of untransformed 2001 data showing the average abundance and the standard deviation between dissimilarity values (Diss/SD} of Poaceae species between closed fence (CF} and open fence (OF} treatments. Contrib %=percentage contribution to the dissimilarity index, Cum %=cumulative percent.

Poaceae SIMPER 2001 Species Average Abundance Diss/SD Contrib. % Cum. % CF OF Untransformed 2001 Themeda australis 5.38 4.5 0.91 9.97 9.97 Microlaena stipoides 1.63 2.25 1.24 8.27 18.23 Panicum simile 0.69 1.75 1.29 7.68 25.92 Entolasia stricta 0.88 0.88 0.95 6.95 32.87

43 There were significant differences in the cover abundance between fire treatments in 2001 (Global R=0.15, p=0.004) and 2002 (Global R=0.17, p=0.003). In general, the unburnt sites were significantly different to the control burn and summer burn sites in 2001. Themeda australis and Microlaena stipoides var. stipoides were present in all significant pairwise comparisons with T. australis more cover abundant in the sites that were control burnt and summer burnt. Entolasia stricta appeared to be more cover abundant in the unburnt sites (Figure 2.16 & Table 2.13).

11> w g7 g8 l1l ca "O 'g6 c: ..55 _56 <( <( !' a;4 ILB ..... ILB , ·>3 ~4 ., 0 CCB:l2 0 CCB (.)2 (.) 11> Q)2 Cl ~1 ca..... ~o ~o <( Ta M; Es Ta M; ~ ~

IUga g8 co ca "O "C

.D§5 ]6 <( <( ..... ll.ES2 ..... ILEE2 ~4 ~4 0 DCB 0 Offi (.) (.) Q)2 Q)2 Cl caCl ~ ..... ~o <(~o <( Ta Mi On Ta M; ~ ~ES

Figure 2.16: Average cover abundances (based on ordinal Braun Blanquet value.s) for Poaceae species in 2001 that contributed up to 30% of the average dissimilarities between significantly different fire treatment groups (based on ANOSIM). Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000 CBS2=control burnt 1997 + summer burn 2000, SBS2=summer burn 1997 + summer burn '2000. At=Austrodanthonia tenuior, Dm=Dichelachne micrantha, Es=Entolasia stricta, Ms=Micro/aena stipoides, Ta=Themeda australis.

44 In 2002, Themeda australis, the dominant grass, was most cover-abundant in sites that had been control burnt (CB) in 1997, often being lower in cover abundance in sites that were unburnt. Other species varied between sites (Figure 2.17 & Table 2.13). As Themeda australis was found in most sites, its occurrence (pres/absence data) did not influence differences in species richness between fire treatments. Overall this suggests that T. austra/is was influenced by fire events but was not destroyed by fire. In order to investigate this further, differences in cover abundance of T. australis were analysed using univariate statistics.

Q) ~ gs lij8 «I "O "O c: ..56 .c§5 < < .... ILB Q>4 ILB g?4 > 0 CCB 0 oc:Ee (.)2 ~2 Q) c: Cl «I ~o ~o < Ta M A; ~ Ta l\t ES ~ ~

Q) ~ (.) ;a lij8 "O "O c: §5 ..56 .c <.... < 11.HX 1Cffi2 li>4 g?4 > 0 Dl.ES2 0 offi 02 ~2 Q) c: Cl :uo"' e!Q) 0 > > < Ta ES < Ta ES ~ On ~ ~

Figure 2.17: Average abundances (based on ordinal Braun Blanquet values) for Poaceae different fire treatment groups (based on ANOSIM). Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 + summer burn 2000, SBS2=summer burn 1997 + summer bum 2000. Dm=Dichelachne micrantha, Es=Entolasia stricta, Microlaena stipoides, Ps=Panicum simile, Ta=Themeda australis.

45 Table 2.13: Similarity percentages (SIMPER) of untransformed 2001 and 2002 data showing the average abundance and the standard deviation between dissimilarity values (Diss/SD) of Poaceae species between fire treatments. Contrib %=percentage contribution to the dissimilarity index, Cum %=cumulative percent. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt +summer burn 2000, CBS2=control burnt 1997 +summer burn 2000.

Group Ave Diss. Species Average Abundance Diss/SD Contib (%) Cum(%)

Poaceae 2001 UB CBS2 UB CBS2 57.49 Themeda australis 3.75 5.25 0.92 12.49 12.49 Microlaena stipoides 2.5 1.93 12.15 24.64 Entolasia stricta 1.25 0.33 0.94 7.19 31.83 UB CB UB CB 57.75 Themeda australis 3.75 5.75 0.91 12.54 12.54 Microlaena stipoides 1.83 1.69 9.59 22.13 Entolasia stricta 1.25 0.25 0.92 7.28 29.41 UBS2 CB UBS2 CB 53.42 Themeda australis 2.8 5.75 1.41 15.38 15.38 Microlaena stipoides 2.8 1.83 1.26 7 22.38 Dichelachne micrantha 1.33 1.13 6.98 29.36 UBS2 SB UBS2 SB 54.34 Themeda australis 2.8 5.77 1.38 14.84 14.84 Microlaena stipoides 2.8 1.85 1.16 7.98 22.84 Austrodanthonia tenuior 1.4 0.15 1.12 6.56 29.38 Poaceae 2002 UB CB UB CB 43.6 Themeda australis 4.25 6.17 1.17 12.62 12.62 Microlaena stipoides 2.25 2.83 1.07 8.32 20.95 Panicum simile 2.25 1.42 1.2 7.68 28.63 Entolasia stricta 1.5 0.67 1.37 7.13 35.76 UB CBS2 UB CBS2 47.28 Themeda australis 4.25 5.83 1.08 12.97 12.97 Microlaena stipoides 2.25 1.75 1.29 9.52 22.49 Entolasia stricta 1.5 0.5 1.58 8.9 31 .39 CBS2 SB CBS2 SB 44.21 Microlaena stipoides 1.75 3.08 1.14 10.29 10.29 Entolasia stricta 0.5 1.54 ~ . 08 9.08 19.37 Themeda austra/is 5.83 5.54 0.84 8.84 28.2 Dichelachne micrantha 0.83 1.46 1.18 7.2 35.41 UBS2 CB UBS2 CB 47 Themeda australis 3 6.17 1.27 18.87 18.87 Entolasia stricta 1.6 0.67 1.2 7.57 26.44 Panicum simile 2.2 1.42 1.21 6.76 33.2 UBS2 SB UBS2 SB 49.59 Themeda australis 3 5.54 1.26 15.75 15.75 Entolasia stricta 1.6 1.54 1.14 6.73 22.48 Panicum simile 2.2 1.08 1.33 6.65 29.13 Oiche/achne micrantha 1.46 1.26 6.15 35.28 CBS2 UBS2 CBS2 UBS2 50.7 Themeda australis 5.83 3 1.18 19.44 19.44 Entolasia stricta 0.5 1.6 1.16 8.72 28.15 Panicum simile 1.58 2.2 1.32 7.43 35 .58

46 2.5 Cover Abundance of Themeda australis The mean cover abundance of Themeda australis did not differ between grazed and ungrazed sites in 2001 (F 1 30= 11 .88, p=O .1810) or 2002 ' (F 1 30=12.44, p=0.1289) (Figure 2.18). However, mean cover abundance . ' was significantly different between fire treatments in 2001 (F5 42=3.8962, ' p=0.0054) and 2002 (Fs,42=3.3946, p=0.0115). Themeda austra/is cover

abundance was significantly higher in the CB (Tukey's HSD: q5,42=4.232<5.58), SB (qa,42=4.323<5.18) and CBS2 (q5 42=4.232<4.39) in 1 2001 when compared to the UBS2 sites and was significantly higher in the CB (q5 42=4.232<4.87) and CBS2 (q5 42=4.232<4.50) in 2002 when compared ' ' to the UBS2 sites (Figure 2.19).

100 ...Q) > 0 80 0 c 60 -Q) •CF ...u Q) 40 OOF ll.. c ca 20 Q) ::!: 0 2001 2002 Year

Figure 2.18: The mean percent cover of Themeda australis in the closed fence (CF) and open fence (OF) sites in 2001 and 2002.

!;. Cl> 6 >- cCJ 0 CG •2001 0 ,, 4 c c CG ::S 02002 Cl> .Q 2 ::!: <( 0 UB CB SB UBS2 CBS2 SBS2 Fire Treatment

Figure 2.19: The mean cover abundance of Themeda australis in the fire treated sites in 2001 and 2002. Fire treatment symbols are; UB=unburnt, CB=control burnt 1997, SB=summer burn 1997, UBS2=unburnt + summer burn 2000, CBS2=control burnt 1997 + summer burn 2000, SBS2=summer burn 1997 + summer burn 2000.

47 2.6 Discussion

2.6.1 The Impact of Fire on Species Richness and Cover Abundance This study has shown that the grass and shrub species have been the most ·affected by past fire events in this area in the richness and cover abundance of species. The response of grasses to different fire events varies between species while shrubs were generally more abundant in the sites that were last burnt in summer 1997, than more recently burnt sites.

Fire in grassy woodlands encourages the regeneration of many grass species, which can withstand frequent, low intensity fire as an ongoing process but affects species differently (McGowen, 1997, Orr et al., 1997). Past research on grasslands in eastern Queensland found that frequent burning (annual) favoured one species in particular (ie Heteropogon contortus) (Tothill, 1969, Orr et al., 1997) but reduced the numbers of other grass species such as Aristida spp. and Bothriochloa spp. (Orr et al., 1997). Similarly, the present research shows that frequent burning favoured one species in particular i.e. Microlaena stipoides var. stipoides. Most species favoured the summer burn periods of 1997 and 2000 but some species (Entolasia stricta and Poa labil/ardien) were more common in sites left unburnt indicating that these species are fire sensitive. In general, the cover abundance of grasses was lower in the unburnt sites suggesting that the grasses of the Cumberland Plain Woodlands at Holsworthy have adapted to frequent fire events but in order to accommodate all species the time-between and the time-of fire events must remain variable.

Themeda australis is the major plant species in the area with cover abundance typically more than 50% in most sites. In general, the cover abundance of T. australis was higher in the control burn and summer burn sites compared to the sites left unburnt. T. australis is a dominant species of many grassland and grassy woodland communities (Groves, 1974, Lunt, 1990, Morgan, 1996, 1998b, Lunt & Morgan, 1999a). Grassland communities experience frequent burning (ie < 3 years) which is required to maintain plant diversity by preventing species such as T. australis from out-competing other

48 herbaceous species (Bryant, 1973, Stuwe and Parsons, 1977, Lunt, 1994, Morgan, 1998a) and maintain,ing Themeda sward health (Morgan & Lunt, 1999). My results suggest that T. australis requires periodic burning of various intensities to maintain dominance but frequent burning will be detrimental to the shrub species in the woodlands at Holsworthy.

Frequent, low intensity fires do not favour most woody shrubs and leguminous species in the Cumberland Plain Woodlands, Holsworthy Military Area. This is reflected in the occurrence and cover abundance of shrubs being higher in the summer burn 1997 sites. Also, significant changes in climatic conditions at the beginning of the project may account for the success of many shrub species in 1999. The year preceding the 1997 summer burn had a below average annual rainfall (562.1 mm), which may have produced a higher intensity fire than usual due to low moisture levels. Following this, 1998 experienced above average rainfall (1048.6 mm), which provided ideal opportunities for seed bank germination in many shrub species.

While the summer burn of 1997 has been beneficial to many shrub species, in particular Pultenaea villosa, which had a greater and more consistent cover in these sites, it has been detrimental to other species. Astroloma humifusum was absent in sites that were burnt in 2000 (S2) but was present in these sites before the fire. Similarly, Exocarpos strictus was lower in abundance in the S2 sites after the burn in 2000. The fact that these two species are vertebrate dispersed may account for a loss of these species, as seeds are typically left on the soil surface subjecting them to high temperatures (ie ;::: 300 °C) in the event of fire (French & Westaby, 1996). Alternatively, P. villosa is most likely an ant-dispersed species, which ensures that the seeds are buried beneath the soil surface and protected from fatal heating (French & Westaby, 1996). P. vil/osa is from the leguminous family Fabaceae, which generally requires heat (usually 80-100°C) or other dormancy breaking mechanisms to stimulate germination. Therefore, burial by ants ensures that seeds will not be heated excessively and killed, allowing this species to survive in fire prone environments such as the Cumberland Plain Woodlands unless frequent fire destroys adult plants before they have set seed.

49 The mean species richness declined continuously after 1999 in all fire treatment sites suggesting that factors other than the fire treatments have been influential in determining the condition of the community. Rainfall patterns indicate that species richness is sensitive to changes in rainfall with MSR highest in 1999 after an annual average rainfall was above average in 1998 (1048.6 mm). The herbaceous species seem to be one of the most affected groups by climatic conditions. However, the richness of grasses has increased each year and is more abundant in periods of high rainfall. Opportunities for seed germination were increased in 1999 by the fire treatments removing tussocks of Themeda australis and allowing seedling growth of other species without competition from T. australis. Fire treatments and an above average rainfall may have enhanced germination in some species. Similarly, climatic conditions influence species richness and abundance in south-eastern Arizona (~ock et al., 1995) and Breckland, England (Watt, 1970).

2.6.2 The Impact of Grazing on Species Richness and Cover Abundance Mean species richness was higher in the ungrazed sites compared to the grazed sites in 2001 but was only marginally higher indicating that grazing pressure did not influence species richness in the area over the six-year study period. Some differences were noted in the grass and shrub species between grazed and ungrazed sites but these were not strong discriminators between the treatments suggesting that the responses of these species to grazing was variable between sites and hence could not be used to determine specific impacts of grazing pressure.

_These results are consistent with Hill (2000) who found that herbivores at Holsworthy did not have a measurable impact due to a relatively large area and potential home range of the herbivores. Research indicates that the impact of macropod grazing depends upon population density (Lunt, 1991, Neave & Tanton, 1992) with high population levels severely degrading vegetation and restricting the regeneration of many rare and threatened plants (Neave & Tanton, 1992).

50 Evidence of rabbits (ie scats and warrens) at Holsworthy was more common than macropod evidence during my visits. The condition of the vegetation suggests that grazing by rabbits has had minimal impact on the community at present. However, it is common that problems caused by rabbits go unrecognised until it is too late but this is genera'l'ly more severe in arid areas (ANZRCDP, 2001).

Such variable results within the fence treatments indicates that grazing is variable and of low intensity at present. Further research is necessary to determine herbivore numbers and impacts to ensure grazing frequency and intensity does not become detrimental to the Cumberland Plain Woodlands at Holsworthy.

2.6.3 The Interactive Effect of Fire and ~razing on Species Richness and Abundance

The interactive effect of fire and grazing was only apparent in the summer burnt sites, where species richness was higher in ungrazed sites compared to the grazed sites. This interaction suggests that herbivore grazing does have a detrimental effect on some sites in the study area and this will be more intense after summer burning. Due to below average rainfall in 1997 the understorey was most likely very dry and of poor nutritional quality for grazing herbivores, but after the summer burn 1997 the newly resprouting shoots and/or exposed green shoots within tussocks of some grass species would have attracted grazing macropods. Studies of the effect of grazing by native herbivores after fire found that macropods utilise burnt vegetation within days of fire due to the presence of newly emerging shoots of grasses, other herbaceous species and shrubs (Denny, 1985, Noble et al., 1985, Southwell & Jarman, 1987). The emergence of new shoots is particularly important to macropods as their mouth and incisor morphology and the flexibility of the head posture enables them to graze on newly emerging shoots immediately (Southwell & Jarman, 1987). Shoots of resprouting or germinating species can be destroyed by grazing macropods due to the removal of photosynthetic material necessary for further growth (McGowen, 1997). This response is reflected in the reduced species richness in the grazed sites after the burn

51 compared to the ungrazed sites. In a large-scale burn the abundance of newly resprouting/germinating species would dilute the effect of grazing due to the apparently low numbers of herbivores in the area.

In conclusion, the response of the Cumberland Plain Woodlands at Holsworthy to fire and grazing depends on factors such as the time-between and the time-of fire and changes in climatic conditions. No set regime will benefit the community as a whole as the response of the grasses and shrubs vary between disturbance events. Frequent burning will increase the abundance of species such as Microlaena stipoides var. stipoides and Pultenaea villosa while lack of fire will increase Entolasia stricta and Poa labillardieri. Periodic burning will favour many grass species, particularly Themeda australis, regardless of the season of burning and will allow the germination of other herbaceous species due to reduced competition. The maintenance of the Cumberland Plain Woodlands at Holsworthy will depend on variable fire and grazing regimes to ensure species diversity by providing conditions favourable to all species periodically.

52 Chapter 3 Germination response of 22 Poaceae species to heat and smoke

3.1 Introduction Temperate and tropical grasslands and grassy woodlands worldwide experience frequent burning which removes vegetation, requiring the establishment of new individuals (Southwell & Jarman, 1987, Mcintyre & Lavorel, 1994, Noy-Meir, 1995). Many species in fire-prone areas have adapted to frequent burning by responding to fire-related cues such as heat, smoke and light to stimulate germination by breaking seed dormancy (Baskin & Baskin, 1998).

Much research has been undertaken on the responses of seeds to heat for

species with thick seed coats (eg Faba~eae species) where the seed coat cracks allowing imbibition of the embryo (Purdie, 1977, Warcup, 1980, Auld and O'Connell, 1991, Cocks and Stock, 1997, Enright et al., 1997, Clarke et al., 2000, Read et al., 2000, Hanley et al., 2001). Some species of the Proteaceae, Casuarinaceae and Myrtaceae families whose seeds are enclosed within follicles in woody fruits have also adapted to fire with the release of seed after the application of heat (Bell et al., 1987, Whelan and York, 1998). These species are typically found in communities that are often burnt through natural wildfires, prescribed burning and human induced fire (Bell et al., 1987, Whelan and York, 1998, Kenny, 2000, Morris, 2000). Few studies have been undertaken on other families eg. Poaceae, even though the communities with many Poaceae species also experience frequent burning due to the burning practices of- many pastoralists in the woodland/grassland areas of Australia (Tothill, 1969, Southwell & Jarman, 1987, McGowen, 1997, Orr et al., 1997).

Plant-derived smoke as a seed germination cue has been investigated in the last 18 years since Keeley and Pizzorno (1986) found that charred wood can stimulate germination in herbs in the California chaparral. The current research has focussed mainly on South African (de Lange & Boucher, 1999, Brown, 1993, Baxter et al., _1995, Brown et al., 1998, van Staden et al., 2000)

53 and southwest Australian species (Dixon et al., 1995, Roche et al., 1997a,b, 1997, Roche et al., 1998, Read and Bellairs, 1999, Tieu et al., 1999, Lloyd et al., 2000) where summer burning is a common phenomenon. Smoke products can be applied in a gas (aerosol smoking), liquid (smoke water) or solid (charcoal) state (Dixon, 1996) and it is generally accepted that it is the chemical compounds found in smoke by-products that break the dormancy of seeds allowing them to germinate (de Lange and Bouche, 1990, Brown, 1993, Dixon et al., 1995, Tieu et al., 1999).

In temperate grasslands in Australia, research has focussed of the effects of frequent burning. It is considered that in order to maintain species diversity, frequent (ie < 3 year intervals) burning is necessary to remove thick tussocks of Themeda australis and allow the establishment of other herbaceous

species (Stuwe & Parsons, 1977, Lunt, 1~90, Morgan, 1998a, Lunt & Morgan, 1999 a,b, Lunt & Morgan, 2000). However, grassy woodlands such as the Cumberland Plain Woodlands contain trees and scattered woody shrubs, which require longer fire intervals to reach reproductive maturity. More frequent fires may cause a reduction in the soil seed bank and over time eliminate some species (Bond & van Wilgen, 1996).

The Cumberland Plain Woodlands of western Sydney originally covered approximately 200 000 ha which has now been reduced to approximately 6 % of it's original distribution due to land clearing, pastoralism, agriculture, urbanisation and changed fire and grazing regimes (Benson & Howell, 1990a,b, James et al., 1999). The Cumberland Plain Woodlands are listed . under the Threatened Species Conservation Act (1995) and the Environment Protection and Biodiversity Conservation Act (1999) as an endangered ecological community. This community is one consisting of Eucalyptus trees, sparse patches of woody shrubs and a dense ground cover of grasses and herbs. The focus of this research is a Cumberland Plain Woodlands remnant in the Small Arms Danger Area, Holsworthy Military Area, western Sydney (33°59'S 150°55'E). This remnant has a history of frequent burning (Hill, 2000) and in the last six years ( 1997 - 2002) has experienced three wildfires (Hill, 2000 and pers. obs.). The most speciose family in the community is

54 Poaceae, which forms an important component of the dense ground cover. The effect of frequent fire on the ability of grasses to germinate is unknown. This research aims to improve an understanding of the effects of the fire related products of heat and smoke on Poaceae germination as a basis for management aimed at the conservation of the Cumberland Plain Woodlands community by determining:

• The germination responses of grasses to a range of temperatures experienced in surface soils during bushfires (40-120 °C).

• The germination responses of grasses to smoke from bushfires.

• Whether the response to heat treatments depends on the application of a smoke treatment.

55 3.2 Materials and Methods

3.2.1 Seed Collection Seeds of 22 Poaceae species, including two exotic species, were collected from two Cumberland Plain Woodlands remnants, the Small Arms Danger Area (33°59°15"S 150°55°00"E) within the Holsworthy Military Area and the Royal Botanic Gardens, Mount Annan (34°03°15"S 150°45°10"E). Two species were not found in the Small Arms Danger Area but found nearby ie Bothriochloa macra and Eragrostis curvula. E. curvula is a common weed in other Cumberland Plain Woodland remnants (pers. obs.) but is rare in the Small Arms Danger Area. Seed was usually sown within 1-3 months except for Themeda australis, whose seed was not found in sufficient numbers so seed collected from the Royal Botanic Gardens, Mt. Annan two years previously was used (Table 3.1 ).

Collection took place between March and May 2002 for most species (Table 3.1 ). Seed was picked from mature flower spikes, placed into paper bags and kept at room temperature until required. All seeds were sorted with all structures enclosing the caryopsis removed except where removal caused damage to the seed (Table 3.1). Each seed was chosen on quality and appearance to maximise the chance that viable seeds were used. Malformed, thin or soft seeds were discarded.

56 Table 3.1: Shows the species used, collection date, location of coUection (MA = Mt Annan Botanic Garden (native annex of the Royal Botanic Gardens, Sydney}, SADA = Small Arms Danger Area, Holsworthy Military Area), date of sowing, the age of the seed at time of sowing and whether all enclosing structures were removed.* indicates introduced species.

Date Date Seed Seed Species Collected Location Sown Age (days) Condition

Aristida ramosa 23/04/2002 MA 0710812002 106 lemma intact Aristida vagans 12/05/2002 SADA 0710812002 87 lemma intact Austrodanthonia racemosa var. racemosa 12/05/2002 SADA 0410912002 115 caryopsis naked Austrodanthonia tenuior 19/03/2002 MA 10104/2002 22 caryopsis naked Austrostipa rudis ssp. rudis 0710412002 SADA 2010512002 43 caryopsis naked Austrostipa setacea 29/03/2002 SADA 2010512002 52 caryopsis naked Bothriochloa macra 0810412002 MA 0510612002 58 caryopsis naked Cymbopogon refractus 12/05/2002 SADA 04/09/2002 115 caryopsis naked Dichanthium sericeum 23/04/2002 MA 05/06/2002 43 caryopsis naked Dichelachne micrantha 12/05/2002 SADA 0710812002 87 caryopsis naked Digitaria ramularis 2910312002 SADA 03/05/2002 35 lemma intact Echinopogon caespitosus var. caespitosus 12/05/2002 SADA 07/08/2002 87 caryopsis naked Entolasia stricta 29/03/2002 SADA 20/05/2002 22 lemma and patea intact Eragrostis benthamii 12/05/2002 -SADA 05101,512002 24 caryopsis naked Eragrostis curvula * 12/05/2002 SADA 04/09/2002 105 caryopsis naked Eragrostis leptostachya 07/04/2002 SADA 07/04/2002 26 caryopsis naked Microlaena stipoides var. stipoides 12/05/2002 SADA 05/06/2002 24 caryopsis naked Panicum effusum 23/04/2002 MA 03/05/2002 10 caryopsis naked Panicum simile 12/05/2002 SADA 20/05/2002 8 caryopsis naked Paspafidium distans 29/03/2002 SADA 03/05/2002 35 lemma and palea intact Sporobolus indicus var. capensis * 07/04/2002 SADA 04/09/2002 150 caryopsis naked Themeda australis 20/03/2000 MA 20/03/2002 730 lemma and 'lower gtume intact

3.2.2 Laboratory Germination Studies Seed age varied between species at the time of each experiment but was constant within species (Table 3.1). For each species 800 seeds were separated into four heat treatments (no heat, 40 °C, 80 °C and 120 °C) and two smoke water treatments (no smoke, smoke) in an orthogonal balanced experiment with 20 seeds in each of five replicates per treatment (total 100 seeds per treatment).

57 3.2.3 Heat Treatments Temperatures of 40°C, 80°C and 120°C were applied to seeds to determine the germination response to heat. The control treatments were the seeds unheated ie left at room temperature (approximately 20-25° C). These temperatures were chosen to simulate a hot day (40° C), a cool-bum (80° C), and a hot burn (120° C). Auld and O'Connell (1991) and Cocks and Stock (1997) found the temperatures of 80 - 100° C to be the most successful in stimulating germination in legumes. Heat was applied for two minutes in a gravity convection oven. This duration was chosen on the basis that previous studies on hard-seeded species had an optimal duration of around five minutes (Martin et al., 1975, Mott et al., 1982, Jeffrey et al., 1988, Trabaud & Oustric, 1989, Valbuena et al., 1992, Bossard, 1993, Cocks & Stock, 1997, Hanley et al., 2001) so two minutes was considered sufficient for small grass seeds. Also, a study of post-fire germination in 35 eastern Australian Fabaceae found that duration of heating was less important than temperature as a determinant of germination (Auld & O'Connell, 1991). Each species seeds were heat treated together during which time temperature was carefully monitored using a mercury thermometer; an electronic temperature button was also used for the temperatures below 120 °C. Austrodanthonia tenuior and Themeda australis seeds were only tested at 40, 60 and 80 °C due to a malfunction in the oven, which invalidated the 120 °C treatment. The oven was allowed to stabilise at each temperature for one hour before use. Previous testing had identified that oven temperatures needed to be 5°C warmer to compensate for loss of heat upon opening the door.

3.2.4 Smoke Water Treatment Smoke water was prepared according to the methods described in Dixon et al. (1995) and Read and Bellairs (1999). The vegetation used for smoke production was a variety of ground litter from the Cumberland Plain Woodlands. This included leaf litter and twigs of three Eucalyptus species ie E. moluccana, E. crebra and E. punctata and material from Themeda australis, Poa /abillardieri plus other herbs and ground cover. The smoke product was bubbled through a 20-litre drum of distilled water for one hour

58 (Figure 3.1 ). A 1O % solution of the smoke water was used for the smoke treatment. This dilution was chosen because of the success rate in germination of spedes tested by Tieu et al. (1999) and Clarke et al. (2000) at this concentration.

3.2.5 Sowing Sowing occurred within 48 hours of the heat treatments. Half of the heated seeds were soaked in 25 ml of distilled water, the other half were soaked in 25 ml of a 10 % smoke water solution for 24 hours prior to sowing. After soaking, 20 seeds of each treatment were evenly placed in 9 cm diameter disposable Petr 1i dishes with two sheets of Whatman No. 1 filter paper, moistened with 10 ml of distilled water. The Petri dishes were then sealed with parafilm to reduce desiccation. This meant that there were five replicate dishes of 20 seeds for each heat and smoke treatment combination

Dishes were placed randomly in a cooled incubator with a 12-hour light at 25 °C and 12-hour dark at 15 °C constant setting. Past research suggests that ideal temperatures for grass germination vary between 20 and 35 °C (Mott, 1978, Lodge & Whalley, 1981). Temperatures for this experiment were chosen partly from past research and partly on the wish to avoid fungal growth, as fungal growth will occur more rapidly in warmer temperatures.

Seeds were checked daily for desiccation and moistened when necessary. Germination was scored weekly and germinated seeds removed from the dishes then resealed. Seeds were classed as germinated with the emergence of the radicle. Recording continued for 8 weeks or until 100 % germination. Germination time was classed as the total time taken for germination.

3.2.6 Statistical Analysis Prior to parametric analysis, a Cochran's C was use to determine whether the variances were homogeneous (Zar, 1999). A two-factor analysis of variance (ANOVA) was used to test for significant differences in percentage germination between heat, smoke and heat x smoke interactions, following back transformation to percent. Two species had heterogeneous variances

59 and so a one-way ANOVA was used to test treatments separately using SYSTAT statistical software. The Tu key honestly significant difference test was used to test for differences between treatments that were shown to be significant. Where a significant interaction occurred the difference between smoke treatments was tested for each temperature (Zar, 1999).

60 Figure 3.1: Smoke machine used to produce smoke water. This procedure was performed in a fume hood.

61 3.3 RESULTS

The germination response of species to heat and smoke varied within and among species, with overall germination ranging from 2 to 97 % (Table 3.2). Sixteen of the species showed significant responses to heat and/or smoke. These responses were neither genera nor tribe specific (Table 3.3), except possibly within the genus Panicum which showed a positive response to smoke. Germination time also varied between species and ranged between 14 and 42 days (Table 3.2), with no obvious correlation between total

germination and germination time (F 1,19 =0.4879, p=0.4933), total germination

and seed age (F 1,19=2.7377, p=0.1144) or germination time and seed age

(F 1,19 =0.1363, p=0.7161).

Table 3.2: Mean germination (%) and ± standard deviation of each species in order from highest to lowest germination. The age of seed (Age); time to reach total germination (Germ Time); the germination response of species that did not respond to the heat or smoke treatments but (Good=71-100%, Moderate=70-31%, Poor=0-30%); H indicates a response to heat, either to H=high or l=low temperatures; S indicates a response to smoke, Y=a positive response, HxS indicates a heat/smoke interaction, the letter X indicates an interaction. * indicates an introduced species.

Treatment Response Species Germ without %Germ Age Germ Time treatment H S HxS (days) (days) response

Cymbopogon refractus 97 .13 :!. 5.42 115 21 x Dichanthium sericeum 97.13:!_3.56 43 14 Good Aristida ramosa 96.5 :!. 3.24 115 28 L Aristida vagans 94 .38:!_4.11 87 14 y Bothriochloa macra 92.63 :!. 9.40 58 14 x Austrodanthonia racemosa var. racemosa 91.25±.11.14 115 28 L Microlaena stipoides var. stipoides 88.63 :!: 7.59 24 21 H Echinopogon caespitosus var. caespitosus 86.63 :!: 16.23 87 21 x Austrodanthonia tenuior 69.88 :!: 14.21 22 42 y Dichelachne micrantha 59.88 :!: 10.65 87 14 Moderate Eragrostis benthamii 50 :!: 33.03 24 14 H y Themeda australis 39.63 :!: 14.43 730 21 Moderate Oigitaria ramularis 36±. 10.14 35 14 Moderate Austrostipa setacea 18.63 :!: 14.89 52 35 Poor Sporobofus indicus var. capensis * 12.5 :!: 10.62 150 21 x Austrostipa rudis ssp. rudis 11.5±.17.66 43 35 x Eragrostis cur.tu/a * 10.25 :!: 11.98 105 14 x y Panicum simile 10.25±.11.71 8 14 Eragrostis feptostachya 6.38 :!: 5.99 26 14 y Entolasia stricta 5.75 :!: 8.66 22 42 Poor y Panicum effusum 3:!,5.04 10 14 Paspalidium distans 2.13 :!: 3.18 35 35 H

62 Table 3.3: Phylogeny and classification of 22 Poaceae species (Mallet & Orchard, 2002). The table is grouped into 6 subfamilies, 8 tribes, 16 genera and 22 species. The response category shows the treatment response of each species; H=heat, S=smoke, HxS=heat/smoke interaction.

Synoptic Classification

Subfamily Tribe Genus Species Response

Aristidoideae Aristideae Aristida ramosa H vagans s Chloridoideae Cynodonteae Eragrostis benthamii H,S curvu/a HxS leptostachya s Sporobolus indicus var. capensis HxS Danthonioideae Danthonieae Austrodanthonia racemosa var. racemosa H tenuior s Ehrhartoideae Ehrharteae Microlaena stipoides var. stipoides H Panicoideae Andropogoneae Bothriochloa ma era HxS Cymbopogon refract us HxS Dichanthium sericeum Nil Themeda australis Nil Paniceae Digitaria ramularis Nil Entolasia stricta Nil Panicum effusum s simile s Paspalidium distans H Pooideae Aveneae Dichelachne micrantha Nil Echinopogon caespitosus var. caespitosus HxS Stipeae Austrostipa rudis ssp. rudis HxS setacea Nil

Austrodanthonia racemosa var. racemosa and Panicum simile showed heterogeneous variances (Table 3.4). Significant responses to treatments of these species must be interpreted with caution although McGuinness (2002) does suggest that if a formal test of homogeneity is required, Cochran's test, with a = 0.01, may be used, because it is less sensitive to heterogeneity caused by small variances, which indicates that A. racemosa var. racemosa results are probably valid (Table 3.4) .

3.3.1 Heat Treatment Response Five species were significantly affected by the heat treatments but unaffected by smoke (Table 3.4). Of these, Aristida ramosa (F3,32=4.402, p=0.01) and Austrodanthonia racemosa var. racemosa (F3 32=17.392, p=<0.001) were I negatively affected by temperatures above 40° C (Figure 3.2).

63 Table 3.4: Results of Two-way ANOVA table for 16 species that showed significant differences in germination foUowing heating and smoke treatments. The F ratio (F), the probability (p)=0.05 are shown. Cochran's C test shows homoscedasity except numbers in bold where variances were heterogeneous for P<0.05 & <0.01. * indicates an introduced species.

Species Heat Smoke HeatxSmoke Cochran'sC F3,32 p F1,32 p F3,32 p Co.os;s,s=0.3595; Co.01 ;s,s=0.4226

Aristida ramosa 4.402 0.01 0.2503 Aristida vagans 4.9 0.034 0.249 Austrodanthonia racemosa var. racemosa 17.392 <0.001 0.368 Austrodanthonia tenuior 9.88 0.004 0.3103 Austrostipa rudis ssp. rudis 30.19 <0.001 4.112 0.014 0.283 Bothrioch/oa macra 10.909 <0.001 4.401 0.044 5.94 0.002 0.2793 Cymbopogon refractus 4.823 0.007 0.1662 Echinopogon caespitosus var. caespitosus 43.884 <0.001 11.729 0.002 15.243 <0.001 0.2086 Eragrostis benthamii 60.559 <0.001 59.797 <0.001 0.2412 Eragrostis curvula* 27.616 <0.001 5.323 0.004 0.2129 Eragrostis leptostachya 14.25 0.001 0.2232 Microlaena stipoides var. stipoides 6.343 0.002 0.3158 Panicum effusum 17.001 <0.001 0.234 Panicum simile 93.524 <0.001 0.5653 Paspalidium distans 2.991 0.045 0.2159 Sporobolus indicus var. capensis* 5.151 0.03 3.238 0.035 0.3343

ab a b b - a a a b ~ 100 ~ 110 c: c: 0 0 ;i ;i 100 Ill Cllc: 90 c: ·e... ·e... 90 GI 80 GI C> C> 80 c: c: Cll Ill GI 70 GI 70 :!E :!E 0 40 80 120 Control 40 80 120 Temperature f"C) Temperature ("C\

a. Aristida ramosa b. Austrodanthonia racemosa var. racemosa

Figure 3.2: The mean germination data (± s.d.) of (a) Aristida ramosa and (b) Austrodanthonia racemosa var. racemosa to temperature (°C). Differences in letters above each temperature indicates a significant difference via Tukey's HSD.

64 Eragrostis benthamii showed the most outstanding increase in germination at 120° C (F3 3260.559, p=<0.001) (Figure 3.3a). Microlaena stipoides var. 1 stipoides showed a varied yet significant response to the heat treatments (F3 326.343, p=0.002). The Tukey's HSD (Table 3.5) showed that there was . ' a significant difference between 40° and 80° C. Germination was reduced at 80° C but increased when heated to 120° C (Figure 3.3b). Paspalidium distans germination was very low (ie 2.13 ± 3.18 %, fig 2c) but there was a significant difference between heat treatments (F3 32 2.991, p=0.045) I although the Tukey's HSD did not distinguish between means, this shows that some heat is required for germination (Table 3.4). Figure 3.3c shows minimal difference between temperatures. However, the results suggest that heating to 40-120° C is likely to stimulate germination. Table 3.5: Tukey's HSD tests on species where significant ANOVA was found. Critical value for q; Heat =4.046, HxS =4.602. Smoke category shows whether the significant difference shows a decrease (D) or an increase (I) in germination with the addition of smoke wate r.

Tukey's HSD Table

Heat Smoke Heatx Smoke Responsive Species (qcrit=4.046) P=0.05 (qcrit=4.602)

Aristida ramosa 40x80=4.51 40x120=4.11 Aristida vagans 0.034 (D) Austrodanthonia racemosa var. racemosa Cx120=5.14 40x120=5.74 80x120=4.83 Austrodanthonia tenuior 0.004 (I) Austrostipa rudis ssp. rudis C+S = 4.92 Echinopogon caespitosus var. caesJ?itosus C>S = 10.53 Eragrostis benthamii Cx80=6.99 <0.001 (I) Cx120=8.33 40x120=11.99 80x120=15.32 Eragrostis curvula 120°:-S<+S = 8.37 Eragrostis /eptostachya 0.001 (I) Microlaena stipoides var. stipoides 40x80=4.26 Panicum effusum <0.001 (I) Panicum simile <0.001 (I) Paspalidium distans P=0.045 Sporobolus indicus var. capensis C>S = 4.39

65 ab 80 a a b 100 ab a b ab ! 70 ! 90 8 so c: ; 50 :8 c: ..c: 80 ~ 40 -~ .. 70 ~ 30 (!) ; 20 c: .. 60 :::e.. 10 :::e.. 0 50 Control 40 BO 120 Control 40 BO 120 Temperature f"C\ Temperature ("C\

a: Eragrostis benthamii b: Micro/aena stipoides var. stipoides

15 a a a a ;; ~ 10 .2 7i c: - ~.. 5 (!) c: .. 0 :::e Control 40 BO 120 -5 Temperature ("C\

c: Paspalidium distans

Figure 3.3: The mean data (± s.d.) (back transformed to %) of species showing significant difference in the heat treatments (temperature °C). Differences in letters above each temperature indicates a significant difference via Tukey's HSD.

3.3.2 Smoke Treatment Response The application of smoke water affected germination in six. Five species showed a positive response to the smoke water treatments but did not respond to heat (Table 3.4). Of these, one species showed a negative response to smoke. Aristida vagans had 5.75 % less germination when treated with smoke (F =4.9, p=0.034) (Figure 3.4). 32 I 1

66 90

~ 85 c 0 :; 8 0 c 75 C)~ c : 70 :Iii 65 No Smoke Smoke Treatment

Figure 3.4: The mean percentage germination (± s.d.) under smoke treatments in Aristida vagans.

The application of smoke increased the germination in five species. Significant differences were found within Austrodanthonia tenuior

(F32,1=9.880, p=0.004), Eragrostis benthamii (F32,1=59.797, p=<0.001),

Eragrostis leptostachya (F32 1=14.25, p=0.001), Panicum effusum ' (F32,1 =17.001, p=<0.001 ), Panicum simile (F32,1 =93.5241, p<0.001). Austrodanthonia tenuior germination was marginally increased with the application of smoke, while the other four species showed a large increase in germination with the application of smoke (Figure 3.5). E. benthamii increased by approximately 41 % (Figure 3.5b) and E. leptostachya by 60% (Figure 3.Sc). The most significant difference in the smoke treatment was the increase in germination of P. effusum of 94 % (Figure 3.5d) and P. simile of 97.5 % (Figure 3.5e).

67 ~ - 70 !:... ~ c ; 60 0 ;; 60 ~ 50 nl ni c 40 c 40 ·~ ~ 30 GI GI (!) 20 (!) 20 c c nl 10 ni Cll GI 0 :E 0 :E No Smoke Smoke No Smoke Smoke Treatment Treatment

a: Austrodanthonia tenuior b: Eragrostis benthamii

~ 20 :.e- 14 ~ ~ 12 c c 0 15 0 ; 10 ftl ..nl c .E 8 10 e... E... 6 CD Cll C> (!) 4 c 5 c nl 2 .. Cll Cll :IE 0 :IE 0 No Smoke Smoke No Smoke Smoke Treatment Treatment

c: Eragrostis leptostachya d: Panicum effusum

No Smoke Smoke Treatment

e: Panicum simile

Figure 3.5: The mean germination (± s.d.) (back transformed to %) of the large significant differences between no smoke/smoke treated seeds.

68 3.3.3 Heat and Smoke Interaction

Six species were influenced by an interaction between smoke and heat (Table 3.4). Overall germination in Austrostipa rudis ssp. rudis (F3 32=4.112, 1 p=0.014) was higher in the smoke treated seeds for all but the control treatment (Figure 3.6a). Germination was also higher in Bothriochloa macra

(F3 ,32=5.94, p=0.002) when treated with smoke without any heat (i.e. control) but for all other heat treatments, smoke water did not influence germination (Figure 3.6b). Echinopogon caespitosus var. caespitosus (F3 32=11.73, ' p=0.002) had lower germination at 40°C than at other temperatures but smoke decreased germination only in the control treatment (Figure 3.6c). Sporobolus indicus var. capensis (F3 32=3.238, p=0.035) increased in the ' smoke treatments with increasing temperature (Figure 3.6d), but was only different at the control temperature (when smoke was negatively influencing germination) and at 120°C (when it was positively influencing germination). For Eragrostis curvu/a germination increased in the smoke treated seeds at each temperature but was only significant at 120°C (F3,32=5.323, p=0.004 Figure 3.6e). Finally, in Cymbopogon refractus (F3 32=4.823, p=0.007), ' smoke decreased germination by 15% at 40°C but did not influence germination at any other temperature (Figure 3.6f).

69 ab aa aa aa aa aa aa ;e- 60 1 1 ~ ~ ~ l ab p: I DNo Smoke DNo Smoke E 60 I ~ •Smoke •smoke

a: Austrostipa rudis ssp. rudis b: Bothriochloa macra

ab bb aa aa ab aa aa aa ~ 100 l ;e- 40 - 80 1 :E :E iE ~ eo ONo Smoke E 40 •Smoke •Smoke ; 20

c: Echinopogon caespitosus var. caespitosus d: Sporobo/us indicus var. capensis

aa aa aa ab aa ab aa aa

.-.~ 50 J § 40 ~ QI 30 DNo Smoke DNo Smoke C> 20 - c: I •Smoke •Smoke C'G QI 10 ~ ::!E 0 0 40 80 120 0 40 80 120 Temperature (C) Temperature (c)

e: Eragrostis curvula f: Cymbopogon refractus

Figure 3.6: The mean germination (germ) data (%) (back transformed to %) of six species showing significant differences in heat/smoke interactions using Tukey's HSD. Tukey's HSD was performed at each temperature to investigate the impact of smoke. Temperature is in °C. Differences in letters above each temperature indicates a significant difference via Tukey's HSD.

70 3.4 Discussion

There is high variability in how the germination of grasses of the Cumberiland Plain Woodlands respond to different heat treatments and the application of smoke, which simulate changes in the fire regime. These responses appear to be unrelated to taxonomic relationships.

Comparatively, these results reinforce the conclusions of the field study where changes in species richness and cover abundance (Chapter 2) also showed that there was high variability in how the grasses responded to various fire treatments and indicated the necessity for variable fire intervals and seasons to ensure that all species have the opportunity to thrive at some point in time. All of the species tested are perennial grasses, which generally survive fire well, except when very hot fires occur (McGowen, 1997). The regeneration capacity of these grasses after fire will be greater if their seeds can withstand the temperatures and/or smoke, or are buried sufficiently in the soil bed to be protected from any fatal effects of fire. Over time, however the species able to regenerate by way of seed germination may out-compete the species whose seeds are destroyed by fire, such as Aristida ramosa, Aristida vagans and Austrodanthonia racemosa var. racemosa. Further research is necessary to determine how these species are affected by frequent fire over time.

Germination in six species was not affected by either treatment. The response was not associated with a lack of germination in Dichanthium sericeum, which had high germination of over 95 %. This result differs slightly from the response found by Read et al. (1999) who found a 5% increase in germination of this species from 93 to 98.3% after the application of a 10% smoke water solution. However, three species, including Dichelachne micrantha, Digitaria ramularis and Themeda australis germinated between 35- 60 % of seeds suggesting that further after-ripening may be required or there is poor seed viability in these species. Two species experienced poor germination in all treatments. Austrostipa setacea had less than 20% germination and Entolasia stricta had less than 6%. These species are not

71 stimulated by the heat and smoke treatments so dormancy may be long-term and not broken by events such as fire.

3.4.1 Heat Treatments My research shows that grass germination in response to heat is variable. Species that have shown a positive heat response are likely to increase in abundance after wildfire, for example, Austrostipa rudis ssp. rudis, Echinopogon caespitosus var. caespitosus, Eragrostis benthamii, E. curvula and Sporobolus indicus var. capensis. Research indicates that many species from a range of families have adapted to the event of fire and are stimulated to germinate at high temperatures. Most studies have found that temperatures between 80-100 °C show optimal germination (Auld & O'Connell, 1991, Keith, 1997, Smith et al., 1999, Clarke et al., 2000, Gilmour et al., 2000, Kenny, 2000, Morris, 2000, Read et al., 2000). My results suggest that within the Poaceae family, response to heat is highly varied resulting in differential germination depending on the presence of fire.

The depth of the seeds of these species in the soil substrate may influence how they respond to a fire event. Temperature varies with increasing soil depth (Lock & Milburn, 1971, Gillon, 1983, Coutinho, 1990, Auld & O'Connell, 1991). Temperatures at ground level during a fire may be between 300 °C - 540 °C, which will kill any seeds on top of the soil (Hopkins, 1965, Miranda et al., 1993). Temperatures of 59, 37 and 32 °C were recorded in burnt Themeda grassland at Katojo, western Uganda at depths of 0, 1 and 5 cm, respectively, (Lock & Milburn, 1971 ). Coutinho (1990) found that soil temperatures were 74, 47, 33 and 25 °C at 0, 1, 2 and 5 cm soil depth during experimental burn-offs in the Brazilian cerrados. Alternatively, research by Auld and O'Connell (1991) of open heath/open forest in south-eastern Australia found that the temperature of soil at 2 cm might be approximately 80 °C during moderate intensity fire.

Three species showed increased germination at high temperatures. The most remarkable increase was in Eragrostis benthamii at 120 °C, which showed poor germination in all other heat treatments suggesting that this species will

72 germinate strongly following wildfire. Interestingly, Eragrostis curvu/a germination was also higher in the 120 °C treatment, germination increased with increasing temperature when smoke was applied. Similarly, Sporobolus indicus var. capensis germination increased at 120 °C but only in the presence of smoke. These genera produce very small seeds, which are expelled from the parent plant without any enclosing structures. These seeds are most probably found within the very top few millimetres of soil due to a lack of an active awn system. This makes them more susceptible to the effects of fire than other species able to bury themselves and be protected from fire. These species are generally common in the study area and may have adapted to fire by germinating after exposure to high temperatures and smoke. This response was the most important within the heat treatment study as two species are introduced, which may have important implications for future management of the Cumberland Plain Woodland. If fires are less frequent and more intense there may be an increase in the germination of introduced species.

Some studies of the effect of heat on grasses have found that heat treatments did not stimulate or were detrimental to germination (Gonzalez-Rabanal & Casal, 1995, Smith et al., 1999, Read et al., 2000). Similarly, my results show that two species, Aristida ramosa and Austrodanthonia racemosa var. racemosa, had reduced germination after heating suggesting that some species decrease after wildfire. Although total germination of both species was still high ie above 90% it may suggest that germination in A. ramosa is reduced after any burn and germination of A. racemosa is reduced in the event of a hot burn. Eleven species showed no response to heat with six of the.se not responding to smoke either eg. Austrostipa setacea, Dichanthium sericeum, Dichelachne micrantha, Digitaria ramularis, Entolasia stricta and Themeda australis suggesting that regeneration immediately after fire may be limited to resprouting parent plants with seeds only germinating when dormancy is broken and conditions are optimal. Microlaena stipoides var. stipoides showed a varied yet significant response to the heat treatments. Slightly less germination was shown in the 80 °C treatment. This is an unusual result as seeds can obviously withstand higher

73 temperatures as germination did increase at 120 °C. Germination in this species was more than 80 % so slightly reduced germination at 80 °c may have been caused by some other factor such as poor quality seed or moisture loss during incubation.

Paspalidium distans germination was very poor and sporadic with no logical pattern of response to treatments. Responses to the heat and smoke were not clear suggesting that dormancy may be broken after fire in a small proportion of seeds but it is likely that further after-ripening is required as zero germination occurred in the control treatments.

The response of seeds to fires will be affected by seed morphology such as awns, which enable seeds to move below the soil surface. Active awns occur on Austrodanthonia racemosa var. racemosa, Austrostipa rudis ssp. rudis, Austrostipa setacea, Austrodanthonia tenuior, Diche/achne micrantha, Echinopogon caespitosus var. caespitosus and Themeda australis and are likely to result in these species experiencing cooler temperatures. Lock and Milburn (1971) found that the mean depth of Themeda seeds dug up from a burnt area was 11 mm. This species' seeds are able to reach these depths due to active hygroscopic awns allowing the seeds to twist themselves into the soil. This suggests that species possessing similar active awns will enable them to bury deep into the soil and protect them from the heat of fire where they remain dormant until after-ripening and favourable conditions initiate germination (Peart, 1979). The species possessing an active awn system in this report showed variable responses to the treatments. Temperatures of 120 °C only negatively affected A. racemosa var. racemosa while smoke increased the germination of A. rudis ssp. rudis, E. caespitosus var. caespitosus and A. tenuior, the remaining species did not respond to the treatments.

3.4.2 Smoke Treatment Smoke significantly affected five species in this study; one species was negatively affected and four species showed a large increase in germination.

7-l Smoke enhances germination in many plant families (de Lange & Boucher, 1990, Brown, 1993, Baxter et al., 1994, Baxter et al., 1995, Dixon et al., 1995, Keith, 1997, Roche et al., 1997a, Brown et al., 1998, Roche et al. , 1998, Read & Bellairs, 1999, Smith et al., 1999, Tieu et al., 1999, Clarke et al. , 2000, Gilmour et al., 2000, Lloyd et al., 2000, Read et al. 1997, Read et al., 2000,) but can also inhibit germination (Brown, 1993, Dixon et al., 1995 and Read & Bellairs, 1999). The reason for inhibition is unclear but could be due to a toxic effect at high concentration levels.

Panicum effusum and P. simile showed the most significant enhancement of germination (ie over 95% increase in both species), confirming other studies (Read & Bellairs, 1999, Read et al., 1997) where there was an increase in germination of Panicum effusum with the application of smoke. Eragrostis benthamii and E. leptostachya also showed an exceptional response to smoke. Another species within this genus, E. sororia, also showed enhanced germination with smoke (Read et al. , 1997), however, Read and Bellairs (1999) found that germination of Eragrostis elongata was reduced, but not significantly, with the application of smoke.

Both Panicum sp. and Eragrostis sp. produce seeds that do not possess awns to enable burial or enclosing structures to aid wind dispersal. Seeds will drop from the plant and remain close to the soil surface where they will remain until dormancy is broken. The fact that various high temperatures do not negatively affect these species, but are enhanced by the application of smoke suggests that they have adapted to germinate following fire regime in the Cumberland Plain Woodlands. Sporobolus indicus var. capesis is an introduced species in the Cumberland Plain Woodlands and .is closely related to Eragrostis sp. (Table 3.3) with similar seeds but showed a different response to the treatments.

3.4.3 Heat and Smoke Interactions The interaction of heat and smoke was the most significant factor in influencing germination. The interaction responses varied between species. Five species showed an additive effect of heat and smoke where the

75 application of both enhanced germination more than either factor alone. These species are likely to show a positive response to fire, although there was some variability at the temperature that each species responded (Table 3.5, Figure 3.6).

Many studies have found that the additive and independent effect of fire­ related cues increases germination in a range of species (Dixon et al., 1995, Keith, 1997, Gilmour et al., 2000, Morris, 2000). Heat and smoke have also been found as complimentary triggers of soil-stored seed. Read et al. (2000) found that heat and smoke increased the density of different species and enhanced the species richness of different components of the seed bank. The interactive effect of multiple germination cues is obviously a complicated and variable process, which has been shown in this report.

Given these results, in the absence of fire I would predict an increased germination of Aristida vagans in remnants, while frequent fires are likely to favour species such as Austrostipa rudis ssp. rudis, Echinopogon caespitosus var. caespitosus, Eragrostis benthamii, E. curvula and Sporobolus indicus var. capensis. Hot wildfires should see an increase in seedlings of Eragrostis benthamii. E. curvula and Sporobolus indicus var. capensis while cooler control burns will favour Aristida ramosa and Austrodanthonia racemosa var. racemosa. Maintaining the species richness of grasses in the Cumberland Plain Woodlands will require adequate conditions at a range of sites for each species, and my results point to a maintenance of landscape heterogeneity in the fire regime to cater for variable germination requirements.

76 Chapter 4 Concluding Discussion

The aim of this research was to extend our understanding of the effect of fire and grazing on a range of 'lifeform groups in the Cumberland Plain Woodlands, Holsworthy Military Area to help us understand the consequence of changed disturbance regimes in other Cumberland Plain Woodland and grassy woodland communities. Historically, grassy woodlands have evolved under a vadable fire regime and as a consequence contain species able to regenerate at various times and after various fire events. Therefore, management practices that involve burning regularly will eliminate many species from the shrub and herbaceous layers. In this study of the Cumberland Plain Woodlands the fire regime is more important to the maintenance of species richness and diversity in grassy woodlands and grasslands than grazing by native herbivores, which may not play an important role in the dynamics of the grassy woodland ecosystems unless grazing intensity increases. In comparison, South African grasslands that sustain large, migratory ungulates have evolved to survive seasonal grazing at an intermediate level, which plays an impo.rtant role in the dynamics of the grasslands (McNaughton, 1985). It is important to understand the evolution of natural disturbance regimes to determine how to manage endangered ecological communities effectively which points to a necessity to maintain a variable fire regime to preserve species richness and diversity within the Cumberland Plain Woodland and grassy woodland communities.

The results of this research will provide important information indicating the factors necessary for appropriate management strategies for the recovery of the Cumberland Plain Woodlands and other ecologically sensitive grassy woodland communities within Australia and throughout the world. This res.earch included investigating how the vegetation varied over six years in areas subjected to differing fire treatments and grazing. Furthermore, the effect of heat and smoke on the germination response of 22 Poaceae species was investigated to determine how the by-products of bushfires affect the germination of grasses in grassy woodlands.

77 4.1 The effect of fire and grazing on species richness and abundance I found that the summer wildfires were more influential on the shrub species than the autumn control burn with shrub richness increasing after the fires in many species, particularly Pultenaea villosa. However, some shrub species showed a negative response to the summer burns e.g. Astroloma humifusum and Exocarpos strictus. Differences in the capacity of shrub species to regenerate by seed germination may depend on how the seed is dispersed and may suggest that the species that have vertebrate-dispersed seed may be less successful than species that have ant dispersed seeds in fire prone environments (French & Westaby, 1996). Further research is required to determine the dispersal mode and the resprouting potential of the shrub species in the Cumberland Plain Woodlands, Holsworthy to better understand the regeneration capacity of the shrub species after fire. Although dispersal modes may be an important factor of shrub survival it will be of no consequence if shrubs are burnt too frequently and do not have sufficient time between fires to produce flowers and set seed. These results indicate that many shrub species in the Cumberland Plain Woodland areas are sensitive to the current fire regime, particularly the frequency and season of fire. To maintain plant species diversity in grassy woodlands a fire regime that is suitable to woody shrub species is necessary.

Most grasses at Holsworthy favoured burning at some time and were more abundant after increased rainfall. The diversity of grasses and herbs increased during the study period possibly due to frequent fire removing Themeda australis and opening the ground cover layer and allowing the germination and resprouting of other herbaceous species, which was more intense after good rainfall. This suggests that past fire regimes have allowed a range of species to survive in the Cumberland Plain Woodlands at Holsworthy and indicates that variations in the fire regime are likely to . determine the relative abundance of species with some species clearly favouring specific fire frequencies.

The fact that shrub species require less frequent, more intense fire than the herbaceous layer is a difficult situation to overcome. For example, if the fire

78 frequency is reduced and the cover abundance of Themeda australis increases then a reduction in the diversity of grasses and herbs may occur, particularly in periods of reduced rainfall. Alternatively, if fire frequency remains high there will be a loss of many shrub species. This may have important implications for the future management of the Cumberland Plain Woodland areas and other grassy woodlands and points to the necessity of a variable fire regime to incorporate the requirements of individual species and maintain diversity within vegetation communities. Further research is necessary to determine the response of the community with greater lengths of time between fires, in particular the growth of Themeda austra/is and the potential to out-compete other herbaceous species.

The grazing experiment indicated that the impact of grazing may increase after summer burning but currently does not seem to influence community structure. Low scat abundances and limited herbivore sightings indicate that population numbers are low at this site. Further research at other sites is necessary to determine the consequence of increased numbers of herbivores on the vegetation structure.

4.2 The effect of heat and smoke on grass germination The germination experiment also showed a variab'le response of grasses to the heat and smoke experienced during wildfires. The additive effect of heat and smoke was the most influential in enhancing germination. Even though there was some variability at the temperature that each species responded an additive effect suggests that these species would respond positively to fire. If there are frequent fires, then species such as Austrostipa rudis ssp. rudis, Echinopogon caespitosus var. caespitosus, Eragrostis benthamii, E. curvu/a and Sporobolus indicus var. capensis are likely to be more common. The last three species are also likely to be more common after less frequent, more intense fire. If cooler control burns were applied then species such as Aristida ramosa and Austrodanthonia racemosa var. racemosa would be more common while the absence of fire would see an increase in Aristida vagans. This variability suggests that the species in question have evolved to respond to various fire events and is highlighted by the fact that all species possess a

79 variety of morphological structures, which enable seeds to either bury themselves and be protected or remain on the surface and be relatively exposed enabling them to take advantage of suitable conditions. The consequence of frequent and/or intense fire will see an increase in the weed grass species, which will have important implications for the future management of the area as other species also favour such a regime.

Both the field and germination studies clearly show that fire is an important factor in the survival and diversity of grasses in the Cumberland Plain Woodlands, Holsworthy. The species tested generally survive fire well and have shown in the field and in the laboratory that variations in fire frequency, intensity and season is necessary to promote diversity. It is clear from these findings that to maintain the diversity of grasses within the Cumberland Plain

Woodlands and other grassy woodland ~ommunities a variable fire regime is essential.

4.3 Conclusions This study has shown that fire is an important factor in species richness and abundance in the Cumberland Plain Woodlands at Holsworthy and it is necessary to maintain fire regime variability. To fully understand the consequences of the current fire regime in this area further investigations are necessary to determine the impact over a longer term for the following reasons:

• Themeda australis is the most abundant ground cover species in the area, which general~y increases over time possibly to the detriment of other herbaceous species and shrubs. In order to understand the impact of this species on herb and shrub richness, further work is required to determine the competitive dominance of Themeda australis with greater time between fires.

• The impact of fire on the grasses in this area has shown that there is considerable variability in the germination response of species.

80 Therefore, further study of the effect of frequency, intensity and season of fire on the resprouting potential of native and introduced grasses may indicate that many of the species tested will show an ability to resprout and survive after fire. This would provide greater predictability in determining appropriate fire regimes for Cumberland Plain Woodlands.

• This study has shown that there is variability in how shrub species respond to fire, particularly in the presence of frequent fire, which may see a reduction or complete removal of some species if frequent fire continues. These results indicate that it is necessary to investigate the effect of frequency, intensity and season of fire on all shrub species in the field and also germination trails to determine how each species will respond to the current fire regime.

• This study confirn:ied past work that diversity is greatest in the herbaceous layer and shows a variable response in this structural layer to the current fire regime and changes in rainfall patterns. Further investigation is needed to determine how the germination response of herbaceous species to fire and changed rainfall patterns will affect future diversity of the ground cover. This will involve laboratory trails due to the unpredictability of wildfire burning exclosure plots.

• This study shows that the current grazing regime does not have a detrimental effect on the vegetation at Holsworthy but is more intense after summer burning. It must not be assumed that grazing will not affect the woodlands at Holsworthy eventually and requires further monitoring.

The Cumberland Plain Woodlands once covered much of the Cumberland Plain. It is now under constant threat of destruction and has been reduced to small fragments containing many endangered and vulnerable plant species. The Cumberland Plain Woodlands at Holsworthy represents one of the largest

81 remnants of this vegetation community and is an ideal area to study the effect of the disturbances constantly faced by these woodlands. To adequately conserve this area it is necessary to determine the consequences of the current disturbance regime to identify the factors of the recovery process that are needed to develop the management plan of the area. This includes increased experimental and monitoring processes to determine the ongoing health and diversity of the Cumberland Plain Woodlands, which, in turn, will enhance the knowledge of the requirements necessary for the future management of other grassy woodland and grassland communities.

This project has enhanced the knowledge of the effect of fire and grazing on the Cumberland Plain Woodlands at Holsworthy with management recommendations suitable for the recovery of other grassy woodland communities. It is one of the first studies to research the effect of heat and smoke on grass species in grassy woodlands. It provides important and interesting data needed to conserve the present communities of Cumberland Plain Woodlands and grassy woodlands and highlights the need for the continuation of the current experiment at Holsworthy to determine the long­ term effect of the present disturbance regime on the diversity and abundance of plant species.

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93 Appendix 1: List of species identified at the Small Arms Danger Area, Holsworthy between 1997 and 2002. Reg Con =regional conservation status, C =conserved, V = vulnerable, • = introduced species. Species that are considered to be of particular significance in the western Sydney context are shown in bold type (James et al., 1999).

Family Genus Species Reg Con Acanthaceae Brunoniella australis c ipumilio c Anthericaceae Dichopogon strictus v Laxmannia gracilis c Thysanotus tuberosus ssp. tuberosus v Tricoryne elatior c simplex v sp. Apiaceae Centella asiatica c Daucus g/ochidiatus v Hydrocotyle ipeduncu/aris v Platysace linearifolia c Asteraceae !Aster subulatus • Brachycome angustifolia var. angustifolia c Calo tis sp. Cassinia sp. Chrysocephalum apiculatum v Cirsium vu lg are * Conyza albida * bonariensis * Euchiton involucratus v Face/is retusa * Gamochaeta ca/Viceps * spicata * He/ichrysum scorpioides c Hypochaeris radicata * Laganifera gracilis v stipitata c Olearia microphylla c viscidula v Ozothamnus diosmifolius c Senecio hispidulus var. hispidulus v /autus subsp. lanceolatus madagascariensis * Sonchus oleraceus * Vernonia cinerea var. cinerea c Vittadinia cuneata var. cuneata v tenuissima Campanulaceae Whalenbergia gracilis c stricta v

94 Family Genus Species Reg Con Casuarinaeae A/locasuarina littoral is c nana torulosa c Celastraceae Maytenus silvestris v Clusiaceae Hypericum gramineum c Colchicaceae Burchardia umbel/a ta c Convolvulaceae Convolvulus erubescens v Dichondra re pens c Polymeria ca/ycina v Cyperaceae Carex breviclumis v Gahnia aspera c lsolepis cernua v Lepidosperma laterale c Ptilothrix deusta c Schoen us apogon c Dilleniaceae Hibbertia aspera var. aspera c obtusifolia c oedunculata v Droseraceae Drosera pe/tat~ c Epacridaceae Astroloma humifusum c Lissanthe strigosa c Styphelia laeta var. laeta c Euphorbiaceae Breynia oblongifolia c Phy/Ian thus hirtellus c Poranthera microphylla c Fabaceae Acacia decurrens c fa/ca ta c /ongifolia c Bossiaea orostrata c Chorizema oarviflorum v Daviesia u/icifolia c Desmodium varians c Dillwynia parvifolia v retorta c Glycine clandestina c microphyl/a v tabacina c Gompholobium minus c Hardenbergia vio/acea c /ndigofera australis c Jacksonia scorporia c Kennedia rubicunda c Pultenaea pedunculata v retusa v villosa c Senna odorata v

95 Family Genus Species Reg Con Gentianaceae Centaurium erythraea * tenuiflorum * Geraniaceae Geranium solanderi var. solanderi v Goodeniaceae Goodenia hederacea c Haloragaceae Gonocarpus teucrioides c lridaceae Patersonia sericea c Sisyrinchium iridifolium * Juncaceae Juncus ochrocoleus subsecundus v Lobeliaceae Lobelia den ta ta c Pratia purpurascens c Lomandraceae Lomandra confertifolia subsp. rubiginosa c cylindrica c filiformis ssp. coriacea c filiformis ssp. filiformis c gracilis c multif/ora c Myrsinaceae Rapanea variapilis c Myrtaceae IAngophora costata c floribunda c Eucalyptus ere bra c eugenoidies c fibrosa c longifolia c moluccana c 'tpaniculata v 1punctata c tereticornis c Kunzea ambigua c Mela/euca decora c nodosa c Syncarpia glomulifera c Orchidaceae IAcianthus fornicatus c Caladenia catenata c Diuris maculata v 'Gfossodia sp. Micro tis sp. Orchid sp. Pterostylis concinna v sp. Thelymitra sp. Oxalidaceae Oxalis corniculata * radicosa v

96 Family Genus Species Reg Con Phormiaceae Diane/la caerulea c longifolia c revoluta var. revoluta c Pittosporaceae Bil/ardiera scandens var. scandens c Bursaria spinosa var. spinosa c Plantaginceae Plantago gaudichaudii v lanceolatus * Poaceae Agrostis avenacea var. avenaceae c Aristida ramosa var. ramosa c vagans c !Austrodanthonia laevis :pilosa v racemosa var. racemosa v tenuior c Austrostipa rudis subsp. rudis v setacea v Cymbopogon refractus c Dichanthium sericeµm v Dichelachne micrantha c Digitaria ramularis v Echinopogon caespitosus var. caespitosus c Entolasia marginata c stricta c Eragrostis benthamii v leptostachya c sororia lmperata cylindrica c Microlaena stipoides var. stipoides c Notodanthonia longifolia Panicum effusum c simile c Paspalidium distans c Poa Jabillardieri v Seteria gracilis * Sporobolus indicus var. capensis * Themed a australis c Primulaceae !Anagallis arvensis * Proteaceae Banksia spinulosa c Persoonia le vis c linearis c Petrophile sessi/is v Ranunculaceae Clematis aristata c

97 Family Genus Species Reg Con Rubiaceae Ga/ium gaudichaudii v Opercu/aria diphylla c hispida v Pomax umbel/a tum c Richardia stellaris * Rutaceae Boronia oolygalifolia v Santalaceae Exocarpus stricta c Sapindaceae Dodonaea triquetra c Scrophulariaceae Veronica olebeia c Sinopteridaceae Cheilanthes sieberi c Solanaceae Solan um jprinophyl/um c Stackhousiaceae Stackhousia vim in ea c Sterculiaceae Ru/ingia dasphylla v Thymelaeaceae Pime/ea linifolia c Xanthorrhoeaceae IXanthorrhoea concava v Zamiaceae Macrozamia spiralis c

98 Appendix 2: Australian Map Grid references taken from a Global Positioning System of the forty-eight study sites at the Small Arms Danger Area, Holsworthy

i l Site Zone ! Eastin gs i Northings 1 56 307957 6237517 2 56 307884 6237383 3 56 I 308144 6236437 i 4 56 I 308047 6236293 5 56 307961 6236615 il l 6 56 I 307615 6236568 7 56 308084 6236143 I 8 56 ' 307819 6236218 9 56 307781 6236134 10 56 307784 6235473 11 56 307686 6235545 12 56 307576 6234831 13 56 307794 6234946 14 56 307884 6234834 15 56 307511 6234469 16 56 307576 6234627 17 56 307720 6237318 18 56 307788 6237265 19 56 308021 6237277 20 56 308082 6236458 21 56 308040 6236354 22 56 307891 6236214 23 I 56 307775 6235923 24 I 56 307544 6234319 I 307476 6234379 25 i 56 l I 26 i 56 307576 6234627 ' 27 I 56 307860 6234953 l l 307688 6234939 28 I 56 l 29 I 56 307685 6234852 30 56 307710 6236140 31 I 56 307847 6236568 32 I 56 307915 6236442 33 I 56 307565 6234399 6234548 34 I 56 307460 35 56 307832 6234880 36 56 307619 6234890 37 56 307861 6235960 38 56 307713 6236263 39 56 308022 6236182 40 56 308104 6236267 41 56 308057 6236414 42 56 308097 6236528

99 Site Zone Eastin gs Northings 43 56 307707 6236523 44 56 307886 6236514 45 56 307885 6236371 46 56 308066 i 6237352 47 56 307912 I 6237182 I 48 56 307828 I 6237331

100