FIRE AND THE PERSISTENCE OF TUART WOODLANDS

by

Robert D. Archibald

This thesis is presented for the degree of Doctor of Philosophy of Murdoch University 2006

I declare that the work in this thesis is my own account of my research and contains as its main content work that has not been submitted for a degree at any tertiary education institution.

Robert Archibald

ABSTRACT

Tall tuart (Eucalyptus gomphocephala) trees are a defining element of the landscape of Perth and the coastal plain to the north and south. However, with the health of some tuart stands deteriorating, most notably at Yalgorup south of Perth, concerns are heightening that the already fragmented tuart ecosystem will continue to contract, leaving a cultural and ecological scar in the landscape. Like many other eucalypt ecosystems, tuart woodlands have had a long association with fire and are believed to have been frequently burnt by the Aborigines prior to European settlement. Today, fragmentation and European land management practices have led to a lower frequency of fire across most remaining woodlands, but also episodes of intense fire at shorter intervals in some areas. Individual fire events as well as fire regimes have the potential to shape the structure, composition and extent of ecosystems, and eucalypt ecosystems are no different. Thus, the impact of fire and fire regimes on tuart health and regeneration was investigated in this study.

A survey of the woodlands at Yalgorup revealed tuart decline was present across a range of sites with contrasting fire histories. Tuart health was poorest at the longest unburnt site (35 years) and the site burnt by frequent wildfire (three fires in 13 years), suggesting these extreme fire regimes had played a role in the decline. Nevertheless, low ratings of tree health at sites burnt approximately once per decade, point to factors other than fire playing a role in the decline of tuart at Yalgorup. The tuart populations at Yalgorup were dominated by individuals in larger size-classes and there was a significant negative relationship between tree size and the probability of tree mortality. In one stand in Yalgorup National Park, 38 % of the tuart saplings/trees had died. It follows that the regeneration and development of tuart seedlings into adults will need to occur within the next one to two decades if these woodlands are to persist. Seedling counts confirmed that tuart regeneration was virtually absent in unburnt areas, as is common with eucalypt species. Interpretations of size-class distributions suggested controlled burning contributed little to recruitment, whereas in areas subject to wildfire outside Yalgorup, tuart had regenerated in abundance such that size-class distributions were skewed towards the smaller size-classes. A repeat survey of plots established in 1976 supported anecdotal reports that the mid-storey tree peppermint (Agonis flexuosa) was increasing in density and height in Yalgorup. The skew towards the smallest size- class within peppermint populations at Yalgorup and the presence of seedlings ( > 200

i seedlings ha-1) within areas not recently burnt (at least four years since fire) brought into focus the differing mode of regeneration for this species in comparison to tuart. In further contrast to tuart, mature peppermints were in good health with no dead trees reported in the population surveys. A possible role for competition in tuart decline was highlighted by significant negative correlation between peppermint density and tuart health. Together, these results suggest that a general drift from tuart woodland to peppermint forest appears entrenched.

Comparative studies of the bark thickness, fire response and resprouting behaviour of tuart and peppermint illustrated the capacity of individuals of both species to persist with recurrent fire. As adults or juveniles, tuart and peppermint resprouted following complete canopy scorch, and often from crown branches: an ability not uncommon for co-occurring tree species and shrubs in this environment. Survival was between 75 and 100 % for fully-scorched tuarts in size-classes ranging from small saplings to trees across six sites. From the small sample of peppermints available for comparison, seedlings and small saplings appeared to be more vulnerable to mortality by fire; only 45 % of individuals with a diameter at breast height ≤ 1 cm survived following complete canopy scorch from a controlled burn. With thicker bark, a reliance on stem epicormic buds rather than a lignotuber for resprouting and a greater capacity for height growth, the fire resistance and post-fire recovery of tuart would be expected to differ from that for peppermint. Opportunities for managers to exploit these differences in their burning prescriptions so as to address the drift from tuart to peppermint dominance were outlined. In addition, results indicating a positive response in canopy condition for tuarts with < 10 % canopy scorch following a controlled burn imply that tuart vigour may benefit more immediately from fire. But, observations also revealed that spikes in intensity within controlled burns can damage large unhealthy trees and thereby accelerate the decline process. Thus, the application of fire to declining tuart stands needs to be conducted skillfully.

Patterns of growth and survival for tuart and peppermint seedlings were linked to the contrasting ability of the species to establish at burnt and unburnt sites. The purported importance of ashbeds in tuart establishment was demonstrated; at the end of summer (February) at a recently burnt site (Golden Bay), the mean height of tuart seedlings on ashbeds was three times that for those off ashbeds, and the survival rate on and off the ashbeds was 35 % and 15 %, respectively. Therefore, while important, tuart seedlings

ii were not dependent on ashbeds in this instance. The hardy nature of peppermint seedlings when compared to tuart was illustrated when the species were planted in an unburnt area of tuart-peppermint woodland at Yalgorup; nine months after planting, 38 % of the peppermint seedlings survived while only 9 % of the tuart survived. Most of the mortality was linked with the summer-drought period and it was suspected that peppermint seedlings were more drought tolerant. A complementary glasshouse experiment showed that tuart had a greater shoot:root ratio than peppermint, which may be one factor in the greater sensitivity of this species to drought. Further, this experiment revealed that with an increase in nutrient supply, the proportional increment in rooting depth for tuart (1.6 times) was significantly greater than for peppermint (0.3 times). An increased availability of nutrients in ashbeds following fire, particularly for phosphorous, was measured within tuart woodland. Therefore, it was inferred that the availability of nutrients was a critical factor increasing the growth, rooting depth and consequently the greater survival of tuarts on ashbeds. Overall, tuart was concluded to share the Competitor strategy (after Grime) in common with many other eucalypts: rapid growth and the attainment of a large size. On the other hand, peppermint exhibits some of the traits of the Stress Tolerator strategy (after Grime): low minimum resource requirements for growth and survival.

The consequence of the increasing cover of peppermint on tuart establishment was explored in field experiments with planted tuart seedlings. No significant impact on seedling growth or survival was observed. There was a declining trend in the seedling growth rates under the high compared to the low peppermint density treatments as the second summer of the experiment was approached, although a statistical difference could not be shown. Peppermint density also had no significant influence on damage to the seedlings from insect attack or the foliar pathogen Mycosphaerella cryptica. Overall, insects and M. cryptica had only an incidental impact on the seedlings: the mean level of canopy damage per affected seedling was < 10 % for either of these two damage agents. Soil type, specifically whether growth occurred on an ashbed was demonstrated to be the most important factor in tuart growth and survival in the 18 months following planting. Although survival was superior on ashbeds (88 %), mean survival was > 50 % in unburnt soil 18 months after planting indicating the potential for restorative plantings to occur in some woodlands between fires.

iii Tuart decline, and the role of altered fire regimes in the decline is complex and further avenues of research were emphasized. Nevertheless, the findings of this and other studies are sufficient to enable purposeful actions to conserve and restore tuart woodlands; recommendations regarding the application of fire with a view to these goals are presented in the final discussion.

iv ACKNOWLEDGMENTS

I thank my principal supervisors Barbara Bowen and Giles Hardy for their invaluable guidance, suggestions and encouragement. I am also grateful to Lachie McCaw for his supervision and for passing on his knowledge in relation to conducting fire experiments. I also thank the following people for their contributions: John McGrath for co- supervision in the initial months of the project, Paul Barber for support with many facets of the project throughout three years, including his contributions to Chapter 6, Mike Calver for statistical advice, John Fox for the kind provision of unpublished data, photographs and opinions, Dugald Close for helpful discussions and editorial suggestions for Chapters 1 to 3, and Dave Ward for sharing opinions and demonstrating the grasstree fire-dating technique. My thanks also extend to the Department of Environment and Conservation (DEC), particularly Drew Haswell for both his interest and his role in mobilizing Departmental resources for the benefit of the project. And, the efforts of other DEC staff were also appreciated; particular mention goes to Rob Towers, Mike Cantelo, Brian Ingles, John Wheeler, Steve Dutton, Myles Mulvay, Peter Gibson, Rob Doria and Femina Metcalfe. Further, I thank the following people for both assisting with fieldwork and for engaging in useful discussions that benefited the project: Des Archibald, Glenys Archibald, Paul Drake, Brett Gray, Wayne Houden, Chris Howe, Mai Itami, Martin Landolt and Peter Scott. Finally, I express my gratitude for the Australian Postgraduate Award Industry (APAI) Scholarship as part of the Australian Research Council Linkage Grant (LPO346931) for tuart decline research involving the DEC, ALCOA, the City of Mandurah, Bemax Cable Sands, Murdoch University and Edith Cowan University.

v TABLE OF CONTENTS

Definitions……………………………………………………………………………….6 Chapter 1. General Introduction: fire and the persistence of eucalypt populations in southern Australia ...... 9 1.1 Introduction...... 10 1.2 The present condition of forests...... 10 1.3 Description of fire events and regimes...... 12 1.4 Fire history of Australia ...... 15 1.5 The impact of fire on processes that affect the persistence of eucalypt populations ...... 19 1.5.1 Plant damage ...... 19 1.5.2 Soil heating...... 20 1.5.3 Microbiota...... 21 1.5.4 Soil chemistry ...... 22 1.5.5 Post-fire microclimate...... 25 1.5.6 Phytophagy...... 26 1.5.7 Competition...... 28 1.6 Fire and eucalypt population persistence ...... 30 1.6.1 Survival and recovery of individuals exposed to fire...... 30 1.6.2 The growth response of trees surviving fire...... 34 1.6.3 Fire and seedling regeneration ...... 37 1.6.4 Fire and population persistence...... 41 1.7 Conclusion...... 43 Chapter 2. Tuart (Eucalyptus gomphocephala) and tuart woodlands from the 19th to the 21st century...... 45 2.1 Introduction...... 46 2.2 The tuart tree ...... 46 2.3 Distribution ...... 47 2.4 Climate ...... 49 2.5 Landforms and soils ...... 50 2.6 Vegetation ...... 51 2.6.1 Formations ...... 51 2.6.2 Peppermint (Agonis flexuosa) ...... 52

1 2.6.3 Other flora...... 52 2.7 History of anthropogenic disturbance within the distribution of tuart...... 54 2.7.1 Clearing and fragmentation...... 54 2.7.2 Grazing...... 54 2.7.3 Timber cutting...... 54 2.7.4 Mining and quarrying...... 55 2.7.5 Fire ...... 55 2.8 Understorey change...... 57 2.8.1 The pre-settlement character of tuart forest and woodland...... 57 2.8.2 Understorey changes since settlement ...... 59 2.9 Tuart decline...... 60 2.9.1 Tuart tree decline and fire ...... 62 2.9.2 Fire regimes and the lack of tuart regeneration...... 63 2.10 Thesis objectives and structure ...... 64 Chapter 3. Relationships between recent fire history and the structural patterns of tuart and tuart-peppermint stands ...... 67 3.1 Introduction...... 68 3.2 Methods...... 69 3.2.1 Sites and site selection ...... 69 3.2.2 Weather conditions during the study period ...... 74 3.2.3 Tuart and peppermint seedling regeneration at unburnt and recently burnt areas ...... 75 3.2.4 Tuart and peppermint population characteristics in tuart and tuart- peppermint woodlands ...... 78 3.2.5 Changes within vegetation monitoring plots in Yalgorup National Park: 1976 to 2004 ...... 83 3.3 Results...... 84 3.3.1 Tuart and peppermint seedling regeneration at unburnt and recently burnt areas ...... 84 3.3.2 Tuart and peppermint population characteristics in tuart and tuart- peppermint woodlands ...... 84 3.3.3 Changes within vegetation monitoring plots in Yalgorup National Park: 1976 to 2004 ...... 91 3.4 Discussion ...... 95

2 3.4.1 Tuart and peppermint seedling regeneration at unburnt and recently burnt areas ...... 95 3.4.2 Tuart and peppermint population characteristics in tuart and tuart- peppermint woodlands ...... 96 3.4.3 Changes within vegetation monitoring plots in Yalgorup National Park: 1976 to 2004...... 101 3.4.4 Conclusion ...... 102 Chapter 4. Survival and recovery of tuart and co-occurring tree species following exposure to fire...... 105 4.1 Introduction...... 106 4.2 Methods...... 107 4.2.1 General ...... 107 4.2.2 Indirect effect of fire on established tuart and peppermint trees...... 108 4.2.3 Comparison of the anatomical features affecting the capacity of tuart and peppermint to survive fire and recover following fire ...... 111 4.2.4 Observations of survival and resprouting response in tuart, peppermint and co-occurring tree species following complete canopy scorch...... 113 4.2.5 Comparative early canopy recovery of tuart and peppermint seedlings subjected to decapitation...... 119 4.2.6 Canopy area (CA) recovery of tuart trees following complete canopy scorch ...... 120 4.3 Results...... 122 4.3.1 Indirect effect of fire on established tuart and peppermint trees...... 122 4.3.2 Comparison of the anatomical features affecting the capacity of tuart and peppermint to survive fire and recover following fire ...... 126 4.3.3 Observations of survival and resprouting response in tuart, peppermint and co-occurring tree species following complete canopy scorch...... 128 4.3.4 Comparative early canopy recovery of tuart and peppermint seedlings subjected to decapitation...... 133 4.3.5 Canopy area (CA) recovery of tuart trees following complete canopy scorch ...... 134 4.4 Discussion ...... 136 4.4.1 Indirect effect of fire on established tuart and peppermint trees...... 136 4.4.2 Comparison of the anatomical features affecting the capacity of tuart and peppermint to survive fire and recover following fire...... 138

3 4.4.3 Observations of survival and resprouting response in tuart, peppermint and co-occurring tree species following complete canopy scorch...... 140 4.4.4 Comparative early canopy recovery of tuart and peppermint seedlings subject to decapitation...... 142 4.4.5 Canopy area (CA) recovery of tuart trees following complete canopy scorch ...... 142 4.4.6 Conclusion ...... 143 Chapter 5. A comparison of tuart and peppermint seedling establishment...... 145 5.1 Introduction...... 146 5.2 Methods...... 147 5.2.1 General...... 147 5.2.2 Field studies...... 147 5.2.3 Glasshouse and laboratory studies comparing tuart and peppermint....153 5.3 Results...... 159 5.3.1 Field studies...... 159 5.3.2 Glasshouse and laboratory studies comparing tuart and peppermint....167 5.4 Discussion ...... 174 Chapter 6. The influence of peppermint on tuart seedling growth and survival: competition, herbivory and foliar disease...... 183 6.1 Introduction...... 184 6.2 Methods...... 185 6.2.1 General...... 185 6.2.2 Study 1: the influence of peppermint trees on the survival rate of tuart seedlings planted in an unburnt area...... 186 6.2.3 Study 2: the influence of peppermint trees on the growth and survival of tuart seedlings in contrasting soil treatments ...... 187 6.2.4 Study 3: a comparison of the susceptibility of tuart and peppermint leaves to herbivory...... 192 6.3 Results...... 193 6.3.1 Study 1: the influence of peppermint trees on the survival rate of tuart seedlings planted in an unburnt area...... 193 6.3.2 Study 2: the influence of peppermint trees on the growth and survival of tuart seedlings in contrasting soil treatments ...... 194 6.3.3 Study 3: a comparison of the susceptibility of tuart and peppermint leaves to herbivory...... 205

4 6.4 Discussion ...... 206 Chapter 7. General Discussion...... 211 7.1 Management...... 215 7.2 Further research...... 218 7.3 Conclusion...... 219 REFERENCES...... 221 Appendicies……………………………………………………………………………I

5 DEFINITIONS

Adult (plant) Having produced flower buds ie. reached maturity.

Ashbed Where combustion results in soil heating to depth (several centimeters) as well as ash deposition (several millimeters). Usually corresponds with patches where coarse fuels (for example, fallen tree branches) combust ie. where burn-out times are extended.

Burn-out time Period of combustion (McArthur and Cheney 1966).

Canopy a) Total foliage on a plant or b) The upper layer of foliage of the vegetation.

Canopy area (CA) rating A scoring system for tree canopy area used in this thesis and based on Grimes (1978).

Canopy health (CH) rating A scoring system for tree canopy health used in this thesis and based on Grimes (1978).

Crown Spreading branches in upper zone of stem.

Decline a) Long-term loss (years) of leaf area from a crown (tree scale), and/or b) a canopy (stand scale) and/or c) the loss of a species from the stand through a failure to regenerate (ecological drift).

Diameter at breast height Diameter at approximately 1.5 m above ground level. (DBH) over bark (DBHOB) Department of Environment The Department of Conservation and Land Management until 2006. and Conservation (DEC) Department of Conservation Changed to the Department of Environment and Conservation in and Land Management 2006. (DCLM) Dieback Refers to “b” in “Decline”.

Establishment (seedlings) The period from germination to approximately the end of one growth season (Bullock 2000).

6 Fire intensity Standard descriptor for comparison between fires. Represents the energy release at the flaming front of a fire measured as kW m-1 but often qualitatively described as low, medium etc.

Fire severity The degree of heating sustained by the soil and vegetation during a fire. Relates to the quantity of fuel combusted.

Fire regime A description of fire types (intensity, severity, season etc.) within time (frequency) and space (burn extent and patchiness) (see Gill 1975). Simplest descriptions consist of intensity class x average fire return interval.

Forest a) General: vegetation dominated by trees where the potential canopy cover is > 20 % (Department of Agriculture, Forestry and Fisheries 2003). Therefore technically includes some woodlands. b) Technical: Foliage projected cover of trees is > 30 % (Specht 1970).

Germinants The period of seedling development following germination and prior to the formation of true leaves.

Regeneration a) Seedling development from seed and/or b) new stems arising from the plant base or below. In this thesis regeneration refers to “a” only unless otherwise stated.

Recruitment The completed progression from germination to adult. In this thesis the term refers only to the process beginning with seedling regeneration, not vegetative regeneration.

Safe site A location within a site that can support survival and growth sufficient for a seedling to reach adulthood (ie. be recruited).

Sapling For tree species: height above 1.5 m and yet to reach maturity.

Seedlings For tree species: height below 1.5 m and includes germinants.

Unburnt (site) A site that has not been burnt for several years or longer.

7

8 1 Chapter 1. General Introduction: fire and the persistence of eucalypt populations in southern Australia

9 1.1 INTRODUCTION

As a recurrent and sometimes dramatic disturbance in the Australian landscape, fire influences eucalypt stand dynamics. Destructive and nurturitive effects on forests arise from a fire event (for example, tree death and seedling regeneration), with ecosystem processes (for example, nutrient cycling) altered over both the short and longer-term. Therefore, changing fire regimes in parts of Australia since European settlement in 1788 have likely affected eucalypt stand structure, composition and possibly tree health. With this background complexity, plus the emotive issue of fire as a land management tool, the role of fire in shaping the condition of eucalypt forests is the subject of as much speculation as fact. The following examines the association between fire and the short- and long-term condition of eucalypt populations. With a focus on the remnant eucalypt forests (and woodlands) of southern Australia, important fire-related processes relevant to eucalypt survival, growth and regeneration are clarified and knowledge gaps are highlighted. Fire regimes that potentially threaten eucalypt population persistence are then discussed. In Chapter 2 these issues are examined specifically for tuart (Eucalyptus gomphocephala) woodlands. However, to begin, background on forest condition, concepts of fire description as well as the fire history of the study area, is presented.

1.2 THE PRESENT CONDITION OF FORESTS1

The character of Australia’s eucalypt forest now differs from that at European settlement in 1788. Most significantly, a large area (approximately 43 %) has been cleared for urban and agricultural development (State of the Environment Advisory Council 1996). Within the 127 million ha of Eucalypt forest remaining (Department of Agriculture, Fisheries and Forestry 2003) fragmentation, alterations to disturbance regimes (including fire) and the impact of new disturbances (such as rising water tables and species introduction) have affected tree health and forest structure and composition (Norton 1997, Yates and Hobbs 1997). However, information on the extent of areas affected by forest decline agents is lacking due to data inconsistencies and uncertainty of the causes of the range of declines (Department of Agriculture, Fisheries and Forestry 2003). Decline has been severe and widespread in the highly fragmented woodland formations of the agricultural zone (Figure 1.1) with symptoms including poor tree

1 In this chapter, “forest” refers to woodlands and forest. That is, vegetation dominated by trees where the potential crown cover is at least 20% (Department of Agriculture, Forestry and Fisheries 2003) 10 health, low levels of recruitment and understorey modification (Yates and Hobbs 1997). These remnants play an important role in preserving regional biodiversity, but are now under further threat from rising saline water tables (National Land and Water Resources Audit 2001). Within the medium to high rainfall zone across several states, the root pathogen Phytophthora cinnamomi has affected hundreds of thousands of hectares of forest (Podger and Keane 2000) and its expansion poses a further threat to biodiversity (Shearer et al. 2004). In addition, a range of tree declines with unknown, but presumably complex, multi-factored causes associated with drought, insect attack and successional processes are widespread across southern Australia (Old 2000).

Figure 1.1: The eucalypt structural formations of Australia (reproduced from Williams and Brooker 1997).

The contrasting moist-winter, summer-drought climate over much of southern Australia makes for a fire-prone landscape (Luke and McArthur 1978, Walker 1981, McCaw and Hanstrum 2003). Fire burnt nine percent of the country’s forest area in the year to March 1999 and 17 percent in 2000 (Department of Agriculture, Fisheries and Forestry 2003). Regional patterns of fire vary with climate and vegetation type. Annual rainfall

11 positively influences primary productivity and therefore fuel (vegetation and plant debris) structure and quantity, whereas the length and severity (temperature) of the seasonal dry period promotes fuel flammability and suitable weather conditions for combustion (McArthur 1966, Luke and McArthur 1978, Walker 1981) (Figure 1.2).

Figure 1.2: Bushfire zones of Australia (reproduced from State of the Environment Advisory Council 1996).

1.3 DESCRIPTION OF FIRE EVENTS AND REGIMES

A fire event requires description beyond its presence or absence to be ecologically meaningful (McArthur and Cheney 1966, Gill 1975, Burrows 1984, Bond and van Wilgen 1996). Most fire descriptors are interrelated (Figure 1.3). Heat output is perhaps the most ecologically influential component of a fire (Raison 1979, Tolhurst 1995, Moore et al. 1995, Bond and van Wilgen 1996). However, other more easily quantifiable descriptors such as intensity are generally reported in order to compare fires (Cheney 1981, Alexander 1982). Intensity (energy release as kW m-1 at the fire front [Byram 1959]) relates to other fire variables and can be readily estimated from flame dimensions or leaf scorch height (McArthur and Cheney 1966, Cheney 1981, Burrows 1984). Intensity is generally accepted as the minimum standard for fire comparison as

12 general intensity, flame height and vegetation damage relationships have been established from observation (Table 1.1). Flame residence time, that is, the length of time of flaming combustion at the fire front, is another important influence on the degree of heating of soil and plants (McArthur and Cheney 1966, Burrows 1984). However, substantial heating to depth within soils or plant stems generally occurs where coarse, woody fuels combust over prolonged periods behind the fire front (Walker et al. 1986, Burrows et al. 1990). This total period of combustion of all fuels is referred to as “burn-out time” (McArthur and Cheney 1966). The term “severity” is sometimes used to give a qualitative description (high, moderate or low) of the heating or heat damage sustained by the environment (through high intensities and/or long burn-out times), although partial quantification of this variable may be achieved through determining total fuel consumption (Gill 1981a).

Intensity Heat output

Flame residence time Burn-out time

Season Severity

Spatial burn pattern Figure 1.3: A selection of important, ecologically relevant fire descriptors and their interrelationships.

Table 1.1: Fire intensity, flame height and vegetation damage in typical open eucalypt-forests (adapted from Cheney 1981 and Burrows 1984). Intensity Max. flame (kW m-1) height (m) Classification Generalized environmental effects < 500 1.5 Low At upper limit, above ground parts of shrubs and eucalypt saplings killed. 500 - 3000 6 Moderate Canopy scorch complete. Trees stems < 10 cm diameter at breast height (dbh) killed to base and some larger trees may suffer bole and crown damage. 3000 - 7000 15 High Some crown fire occurs. Tree stems < 20 cm dbh killed to base. All trees suffer bole or crown damage. > 7000 > 15 Very high Crown fire occurs. Death of below-ground plant structures can occur.

13 Season of burn has major effects on other fire descriptors by affecting moisture content and rates of drying of large fuels and the litter-bed profile (Boone Kauffman and Martin 1989). For example, spring fires generally burn at lower intensity, have shorter burn-out times and leave a greater number of patches unburnt compared to autumn or summer fires (Boone Kauffman and Martin 1989, Burrows et al. 1990). Additionally, seasonal influences on plant moisture content, plant physiological status and soil moisture affect plant damage and recovery. For example, scorch heights can potentially be halved in southwestern forests if fires of intensity around 500 KWm-1 occur in spring rather than late summer or autumn (Burrows 1984).

Assessing the ecological responses to fire requires a regime as well as an event perspective. The concept of fire regime considers event descriptors (Figure 1.3), plus the element of frequency, specified within spatial and temporal parameters (Gill 1975, Williams et al. 1994). As with event descriptors, there are interrelationships between regime components. For example, intensity and frequency tend to be inversely related due to the accumulation of fuel over time (McArthur 1966). The range of information a fire regime can potentially encompass, plus the spatial and temporal component, makes the task of describing fire regimes a difficult one (Williams et al. 1994, Morgan et al. 2001). Descriptions such as “frequent, low-intensity fire” need qualification to be meaningful. For example, the specification of an intensity range (Table 1.1), the mean fire return interval and any trend in seasonal occurrence, allows the regime to be assessed in the local context of potential fire regime and tolerance of ecosystem components (Gill 1975, Gill 1981a).

Potential fire regimes can be generalized for the forest types of southern Australia. For some dry sclerophyll forests (for example, E. marginata forest), fuels may potentially carry fire as frequent as every three years, particularly if grasses and other annuals dominate understories (Christensen et al. 1981). Grasses can be highly flammable during the summer months and burning often promotes their vigour and spread, thereby maintaining a cycle of frequent fire (Baird 1977, Hobbs 2003). Following long periods of fuel accumulation in these forest types, wildfires can burn at intensities > 60 000 kW m-1 (Christensen et al. 1981). In the lower rainfall (< 500 mm) woodland and mallee areas, fire frequency and intensity is generally limited by the lower rate of fuel accumulation (Walker 1981, Hobbs 2002, Bradstock and Cohn 2002). However, intensities may be as high as 30 000 kW m-1 where heath formations exist (McCaw

14 1997, Bradstock and Cohn 2002). In the high rainfall zone, the tall, wet sclerophyll forests (for example, E. diversicolor and E. regnans) with their moist, shaded understories, naturally burn less frequently but consequently with great intensity due to large fuel loads that accumulate (McArthur 1966, Ashton 1981). Fuel moisture regimes in these forests dictate that fires are generally confined to the mid-summer to autumn period (Ashton 1981).

1.4 FIRE HISTORY OF AUSTRALIA

Fire likely became a regular occurrence in the Australian landscape following the onset of the increasingly arid conditions of the mid-Tertiary period (Kemp 1981, Kershaw et al. 2002). The increased frequency of fire likely shaped the distribution of plant species and vegetation types (Kemp 1981, Kershaw et al. 2002, Hopper 2003). While the spread of eucalypts coincided with the increasing prominence of fire in the landscape, the processes of change in vegetation and fire regimes were undoubtedly complex given the intimate link between the two, and the controlling influence of a drying climate on them both (Kershaw et al. 2002). Claims that the Australian flora has evolved adaptations to fire (for example, Gardner 1957, Mutch 1970) are controversial because of the difficulty in separating the influence of factors such as drought on the selection of traits which also enable persistence within a fire-prone environment (Gill 1975, Hopper 2003).

The arrival of Aboriginal people between 40 000 and 100 000 years ago ensured the greater regularity of fire in the Australian landscape (Nicholson et al. 1981, Hallam 1975, Bowman 1998). Fire was fundamental to traditional Aboriginal society, employed in ways that could have effected landscape scale change, for example: managing and hunting game, increasing the abundance of food plants and facilitating passage across the landscape (Hallam 1975, Nicholson 1981, Bowman 1998, Gott 2005). Early settlers’ accounts (Hallam 1975, Abbott 2003, Gott 2005) and the analyses of charcoal deposits in lake sediments (Singh et al. 1981, Kershaw et al. 2002, Hassell and Dodson 2003) suggest fire occurred more frequently, at least in some parts of Australia, prior to the decline of Aboriginal society in the 19th century. Recently, data from the colour banding on Xanthorrhoea spp. stems (Ward 2000, Ward et al. 2001) and E. marginata stem fire scars (Burrows et al. 1995) have suggested frequent burning (at 3 to 5 year intervals) characterized the fire regime prior to European settlement in E. marginata and E. gomphocephala forests of southwestern Australia. However, the significance of data 15 from the Xanthorrhoea stem method remains controversial (Hopper 2003, Enright et al. 2005, Gill 2006) and the interpretation of fire history from fire scars on eucalypt stems (Burrows et al. 1995) has shortcomings (Burrows et al. 1995, Gill and Moore 1997). Nevertheless, where populations of Aborigines were relatively large, such as in coastal southwestern Australia, the collective evidence suggests that burning was conducted at near maximal frequencies in some parts of the landscape (Hallam 1975, Hassell and Dodson 2003). Consequently, due to limited fuel accumulation periods, many of the fires lit by the Aborigines would have been of low to moderate intensity and relatively small in extent (Burrows et al. 1995, Abbott 2003). A fine-scale mosaic comprising a complex of burn histories, from recently burnt to even long unburnt, was the likely result of this fire regime (Bowman 2003).

There are uncertainties as to the degree and extent of the impact Aboriginal burning had on the vegetation. These stem from inconsistencies in the charcoal/pollen record between sites, the methodological problems associated with deriving these records and the difficulties in interpreting cause and effect given vegetation, and the fact that fire and climatic patterns are interrelated (Bowman 1998, Kershaw et al. 2002). Some suggest the impact was dramatic (for example, Flannery 1994). Authors of the most recent review of available charcoal and pollen data across Australia (Kershaw et al. 2002, p3) propose that the effect of the introduction of the Aboriginal fire regime “..was largely to accelerate existing trends rather than cause a wholescale landscape change..”. Importantly, any fire-related modifications would have occurred in accordance with the intensity of occupation (Hallam 1975, Hassell and Dodson 2003).

Low or grassy understories beneath dry sclerophyll forest reported at the time of European settlement have been attributed to frequent Aboriginal burning (Hallam 1975, Flannery 1994, Ryan et al. 1995 in Benson and Redpath 1997, Gott 2005). However, this view is also not without controversy, for example: Benson and Redpath (1997), Benson and Redpath (2001) and Florence (2001) versus Jurskis (2000), and Keith and Henderson (2002) versus Jurskis (2002). Much of the debate in fact concerns the extent of open and/or grassy understorey formation at the time of European settlement. Limited information and historical accounts has led to subjective judgments to be include in the debate (Benson and Redpath 1997, Lunt 2002). As Benson and Redpath (1997) illustrate, while many accounts of open understorey forest can be found in historical reports, references to dense vegetation and shrubby understories are also

16 prevalent. Further, historic photographs and paintings tend to be biased towards openings in the landscape and are therefore not necessarily conclusive evidence (Reithmaier 2001). Even given the existence of apparently open vegetation types at the time of settlement, it may not always follow that these were maintained by fire. There has been a tendency in the Australian ecological literature for landscape patterns to be inferred on the basis of fire without due consideration to the other possibilities (Lunt 2002). That frequent fire is a major limitation on shrub and tree development in a range of ecosystems is an undeniable fact (Bond and van Wilgen 1996, Pyne 2001, Bond et al. 2005, Bond and Keeley 2005). Nevertheless, open forest formations relatively free of shrubby understorey can also develop under the competitive influence of a healthy tree canopy (Mount 1964, Florence 1996, 2001), as a consequence of climatic and soil factors (Florence 1996, Burrows and Wardell-Johnson 2003) and grazing history (McDougall 2003, Rippey and Hobbs 2003, Lunt and Spooner 2005).

While some large destructive bushfires occurred in the 1800s (Pyne 1991), a detailed fire history for the period from European settlement until the establishment of Forests Departments in the early 1900s has not been produced (Gill 1981b, Gill and Moore 1997). Dramatic shifts in fire regimes undoubtedly occurred as forests were cleared and exploited by settlers (Wallace 1966, Gill 1981b, Shea et al. 1981), burning by Aborigines ceased as their societies broke down and prohibitive burning laws were instituted (Hallam 1975, Ward 2000, Ward et al. 2001). In some areas, the decline in Aboriginal burning practices likely allowed forest fuels to accumulate, and therefore more destructive fires to eventuate (Burrows et al. 1995). In other areas, European graziers continued the practices of the Aborigines by burning regularly to favour the growth of grass and herbs (Barker 1988, Ward 2000).

A significant shift in forest fire regimes occurred during the 1900s as the practices of forest management agencies evolved. Until the middle of the century, a policy of fire protection was generally favoured, with controlled burning carried out on only a limited scale (Luke and McArthur 1978, Pyne 1991). The failure of this strategy to prevent large, severe wildfires saw the introduction of extensive controlled fuel reduction burning (McArthur 1966, Wallace 1966, Gill 1981b, Pyne 1991). This practice has continued to today. For example, in southwestern Western Australia, between 50 000 and 250 000 ha of forest (3 to 15 % of the total forested area in public estate) was burnt annually by prescribed fires in the ten years to 2004-2005 (Department of Conservation 17 and Land Management 2005). To ensure these fires are controllable and do not cause major damage to trees, the majority are lit in spring and autumn to minimize fire intensity (Vines 1968, Luke and McArthur 1978). For example, for southwestern Western Australia in 2004-2005, 52 % and 44 % of the total area burnt by prescribed fire was associated with ignitions in spring and autumn, respectively (Department of Conservation and Land Management 2005). Fires of high severity across limited areas (generally < 10 000 ha year-1 per relevant state) have also been lit by management agencies to encourage regeneration following timber harvesting in wet-sclerophyll forests (Shea et al. 1981).

Even with the advent of fuel reduction burning, wildfires continue to occur. For example, 406 wildfires occurred in southwestern forests in 2004-2005 (Department of Conservation and Land Management 2005). However, in general, the severity and extent of wildfires since prescribed burning was introduced has been reduced (McCaw and Hanstrum 2003). Nevertheless, severe and large-scale wildfires still eventuate periodically. For example, 28 600 ha was burnt in January 2005 by intense wildfires in the forests on the Darling scarp east of Perth (Department of Conservation and Land Management 2005). Both human activity as well as lightning account for today’s wildfires (McCaw and Hanstrum 2003, Department of Conservation and Land Management 2005). Forests surrounded by urban areas are particularly susceptible to frequent wildfire due to the high risk of arson and great potential for accidental ignitions (Walker 1981, Burrows and Abbott 2003). The propensity for introduced annuals, particularly grasses such as Ehrharta calycina, to invade urban remnants further increases the vulnerability of such remnants to frequent wildfire due to the flammability of these plants as they cure with each dry season (Baird 1977, Van Etten 1995).

Today, the fragmentation of forest areas has also reduced the frequency of fires in some remnants as roads and other cleared areas often inhibit the passage of fire across the landscape and humans actively limit the spread of fire into urban and agricultural areas (Yates and Hobbs 1997, Pyne 2001, McCaw and Hanstrum 2003). This pattern is particularly pronounced for the small woodland remnants in the agricultural zones (Yates and Hobbs 1997).

18 1.5 THE IMPACT OF FIRE ON PROCESSES THAT AFFECT THE PERSISTENCE OF EUCALYPT POPULATIONS

Fire has the potential to affect eucalypts directly and indirectly, at both the scale of the individual as well as the population (Figure 1.4). Direct impacts involve the heating or combustion of plant parts. Indirect effects are numerous and also potentially significant, for example: the changes in nutrient availability due to soil heating, the destruction of competitors and the alteration of the microclimate at the soil surface (Burrows and Wardell-Johnson 2003). A selection of important processes is discussed below.

1.5.1 Plant damage

Plant tissues are killed when heated to temperatures between 50° and 60° C (Vines 1968, Warcup 1981, Luke and McArthur 1978). Along with the temperature to which tissues are exposed, additional factors significantly influence the heating process: the duration of exposure, tissue moisture content and initial temperature (Vines 1968, Bond and van Wilgen 1996). While heating of tissues occurs instantly at the flame front of a high intensity fire, continued combustion of coarse fuels behind the front can cause considerable heat damage (Cheney 1981, Burrows et al. 1990, Gill 1997).

HEATING/COMBUSTION Eucalypt

Regeneration other vegetation

Growth

soil Persistence (tree) microbiota microclimate

Persistence competition phytophages (population)

Figure 1.4: The direct (thick arrows) and indirect (thin arrows) effects of fire (heat and/or combustion) on eucalypt regeneration, growth and persistence.

19 The heat load sustained by sensitive plant tissues is dependent on the structural features of the plant. Leaves and buds on tall trees, distant from flames, are exposed to lower temperatures or high temperatures only briefly (Gill 1997). Buds and storage structures (for example, lignotubers) below-ground absorb less heat due to the insulative properties of soil (Jacobs 1955, Gill 1975). Bark is an important structure that insulates epicormic buds and stem vascular tissue from heat with thickness of the layer primarily determining the degree of insulation (Vines 1968, Gill 1997). Similarly, woody capsules provide protection to seeds from heat (Gardner 1957, Gill 1975) depending on the size of the covering structure (Mercer et al. 1994). For all primary plant structures (roots, stems and branches), some resistance to heat damage is afforded by bulk (Bond and van Wilgen 1996). Due to the dissipation of heat energy larger objects must receive greater heat loads for the temperature of the object to rise (Mercer et al. 1994). Thus, large diameter roots, stems and branches would tend to be heated more slowly.

Destructive heating of plant tissues can also arise from the combustion of parts of the plant itself. For example, when the outer layer of eucalypt bark is alight the probability of cambial damage is higher (Gill and Ashton 1968). Likewise, heat from the combustion of leaves would increase the probability of cambial death in supporting small branches. Structural attributes, particularly fineness, quantity and spatial arrangement (Papio and Trabaud 1991), as well as moisture and chemical content are important determinants of the flammability of plants and plant parts. For example, loose, dry stringy barks are more flammable (Gill and Ashton 1968) as are the leaves of eucalypts with their high content of volatile compounds (Mutch 1970, Vines 1981).

1.5.2 Soil heating

Soil chemical and biological changes arise in proportion to the degree of soil heating associated with fire (Raison 1979, Warcup 1981, Humphreys and Craig 1981, Burrows 1999). These changes are detailed in subsequent sections (1.5.3 and 1.5.4). Fire factors, particularly the quantity of fuel combusted, are the primary determinants of soil heating (Humphreys and Craig 1981, Walker et al. 1986, Burrows 1999). However, soil type and moisture content also influence the temperature–depth profile (Walker et al. 1986, Burrows 1999). A generalized temperature–depth profile by fire type is presented in Figure 1.5. Considerable within-site variability in soil heating, mirroring spatial variability in fuel quantity, also adds a further layer of complexity to fire–soil heating patterns at a site (Raison 1979, Gill 1981a, Warcup 1983, Adams et al. 2003). 20 1.5.3 Microbiota

Fungi and other soil microbiota are an integral part of ecosystems (Raison 1979, Warcup 1981). Important, direct associations with eucalypts range from the beneficial (mycorrhizae) to the pathogenic (for example, P. cinnamomi and Armillaria luteobubaliana). Indirectly, the activities of the soil microbiota influence nutrient cycling and soil nutrient availability (Raison 1979, Warcup 1981). Despite their significance, knowledge of even the fungal component of the microbiota in eucalypt ecosystems, let alone the effects of fire upon them, is far from complete (May and Simpson 1997, Robinson and Bougher 2003).

Figure 1.5: Generalized soil temperature–depth profiles for a range of fire types (reproduced from Walker et al. [1986]).

The impact of fire on soil biology is greatest in the surface soil and litter layers where the microbial biomass is at a maximum and the potential for combustion or significant heating is highest (Neary et al. 1999). When soils are heated above 25°C microbial activity is stimulated with peak activity occurring at around 37°C (Walker et al. 1986). From 50°C, soil sterilization begins and is complete at 125°C (Walker et al. 1986). Heating soil to temperatures above 100°C reduces the fungal more than the bacterial

21 component of the microbiota, with heating to 200°C eliminating almost 100 % of viable fungal propagules (Guerrero et al. 2005). The horizontal variation in combustion and therefore heating across a site poses difficulties for predicting the soil microbial community response to fire (Warcup 1983, Robinson and Bougher 2003). However, this mosaic of different degrees of soil heating, sometimes including unburnt patches within burnt areas, is important for the maintenance of microbial diversity at sites subject to periodic fire (Warcup 1981, Robinson and Bougher 2003).

Recovery of microbial communities following fire varies with organism type and degree of soil heating. In comparison to other microbiota, decolonization of heated soil by fungi is slower in both laboratory (Guerrero et al. 2005) and field studies (Renbuss et al. 1973, D’Ascoli et al. 2005). For soils heated above 400°C, recovery of fungal populations may only be transient due to the diminishing supply of available nutrients arising from the loss of organic compounds (Guerrero et al. 2005). Following severe fires, pre-fire fungal communities may not be approximated until more than one year has passed (Renbuss et al. 1973, D’Ascoli et al. 2005). For example, in a study in E. marginata forest, mycorrhizal associations were less abundant one year after intense fire compared to a long-unburnt site (Malajckuk and Hingston 1981). Following a fire, a succession of species within the fungal community generally follows the re- accumulation of leaf litter and soil organic matter, as well as the soil pH change that occurs with time (Robinson and Bougher 2003). Thus, the types of mycorrhizal associations generally differ between recently burnt and long unburnt sites (Malajckuk and Hingston 1981, Ellis and Pennington 1992, Launonen et al. 1999).

1.5.4 Soil chemistry

A temporary increase in plant-available nutrients usually occurs after fire due to the addition of ash and the heating of soil (Raison 1979, Walker et al. 1986, Bond and van Wilgen 1996, Keith 1997, Carter and Foster 2004, Certini 2005). Thus, the proportion of above-ground biomass combusted is often positively related to the increase in available nutrients (Grove et al. 1986, Weston and Attiwill 1990, Adams et al. 2003). However, high temperature combustion of biomass and the prolonged heating of the soil at high temperature may lead to the volatilization of some nutrients, especially nitrogen (N) (Raison 1979). Arguably, the most significant overall site effect of fire on soil chemistry is to increases the spatial heterogeneity in nutrient availability (Adams et al. 2003). 22 The contribution of ash to any increase in nutrient availability differs within and between sites. In a review of studies reporting the nutrient content of ash, Raison (1979) found great variation. For example, N content ranged from 0.03 % to 1.77 %, phosphorus (P) from 0.003 to 3.1 % and potassium (K) from 0.17 to 19.1 %. Many factors contribute to this variability: fire temperature, completeness of combustion, fuel type (for example, leaves, wood or grass) and state (for example, living, dead, decomposed), and species from which the fuel originated (Raison 1979, Groves et al. 1986, Walker et al. 1986, Khanna et al. 1994, Gray and Dighton 2006). Metallic elements are generally concentrated, while N, carbon (C) and sulfur (S) are less concentrated in the lighter coloured ash formed from more complete combustion (Khanna et al. 1994).

As ash is rich in base elements, pH increases in the surface layer of soil are usually associated with the deposition of ash during fires (Walker et al. 1986, Attiwill and Leeper 1987, Pritchett and Fisher 1987, Khanna et al. 1994). Due to the higher concentration of metallic elements, the alkalinity of lighter coloured ash can be up to 20 times as high as darker ash formed from less complete combustion (Khanna et al. 1994). Increases of up to three units at the soil surface are possible following forest fires (Viro 1974, Raison 1979). Such large pH changes have wide-ranging consequences for soil ecology, affecting nutrient availability (Humphreys and Lambert 1965), altering the composition of fungal communities (Robinson and Bougher 2003), affecting microbial activity (Raison and McGarity 1980) and therefore also N mineralization rates (Khanna et al. 1994).

Soil heating can be an important cause of the post-fire pulse in plant-nutrient availability. Soils heated to temperatures as low as 35°C can stimulate soil microbial activity, resulting in greater rates of nitrogen mineralization sufficient to induce plant growth responses (Florence and Crocker, 1964). However, a moderate level of soil heating (< 200°C) benefits plant growth by significantly increasing the quantity of mineral forms of N and P (Pryor 1963, Humphreys and Lambert 1965, Warcup 1981, Walker et al. 1986, Adams et al. 2003). The higher concentration of plant-available N + immediately after fire is mostly present as ammonium (NH4 ) (Warcup 1981, Stock and Lewis 1986, Wan et al. 2001). Biological activity after heating also has the potential to + supply large amounts of NH4 due to changes in both microbial populations and the nature of the organic compounds they break down (Walker et al. 1986). The

23 + concentration of NH4 typically doubles in the months following fire (Wan et al. 2001). - Nitrate (NO3 ) concentrations also increase, but generally peak several months after the fire (Stock and Lewis 1986, Wan et al. 2001). For P, heating increases quantities of available forms by altering the fixation capacity of the soil and through the turnover of labile organic P from soil flora and fauna (Adams et al. 2003). Plant available P may be several times higher in the soil under ashbeds where soil heating and ash additions are at a maximum (Adams et al. 2003).

When soil or plant material is heated to temperatures above 200°C, important plant nutrients may be lost through volatilization (Raison 1979, Walker et al. 1986). However, for most of the soil surface subject to heating from typical wild and controlled forest fire, soil temperatures rarely exceed 200°C, and if they do, usually only in the surface centimeters (Figure 1.5) (Raison 1979, Walker et al. 1986). For N and S, volatilization occurs from 200°C and 300°C, respectively with temperatures of 540°C resulting in the loss of virtually all N (Walker et al. 1986). Consequently, plant growth may be affected where soils are heated excessively. For example, Launonen et al. (1999) recorded poorer growth and lower foliar N concentrations for E. regnans seedlings grown in orange coloured ashbed soils (indicative of high temperature oxidation of the soil) compared to black ashbed and unburnt soils. P can also be volatilized in some quantity when temperatures exceed 300°C, although losses are comparatively smaller than for N (Walker et al. 1986, Certini 2005).

Depending on the intensity/severity of the fire and the nutrient in question, site soil chemistry often approximates the pre-burn state within one or two years after fire (Raison 1979, Attiwill and Adams 1993). For example, inorganic N returned to pre-fire levels 205 days after a low intensity fire and 485 days after a slash fire in eucalypt forest (Weston and Attiwill, 1990). However, where heating was severe, some chemical attributes of patches may not return to their pre-fire state for many years. For example, Humphreys and Lambert (1965) found ashbeds at a site had different chemical attributes including a lower P absorption capacity and higher pH, when compared to adjacent non- ashbed patches nine years after burning.

Nutrients prone to volatilization or leaching, particularly N can theoretically be depleted at sites subject to frequent fires (Raison 1979). A recent study found a significant, although small decline (< 10 %) in total nitrogen for only the surface 0 to 2 cm layer at

24 eucalypt forest sites subject to 3 to 4 controlled fires over 14 years (Hopmans 2003). However, other studies comparing frequently and less frequently burnt eucalypt forests have not demonstrated significant differences (for example, Hatch 1959, Adams et al. 2003). In fact, little evidence supporting fire frequency as a cause of soil fertility decline exists for temperate ecosystems generally (Neary et al. 1999). The potential for losses from the leaching of mineralized N is limited by uptake by microbes and regenerating plants (Weston and Attiwill 1990). Regardless, deposition from the atmosphere and N- fixation can replenish N losses from leaching and volatilization with sufficient time (Walker et al. 1986, O’Connell and Grove 1996). In contrast to N, some propose that losses of P, although small, may be significant given the low rates of replenishment of this element in ecosystems (O’Connell and Grove 1996, Adams et al. 2003). More studies are needed to fully understand the influence of fire regimes on nutrient cycling over the long term (Raison 1979, Adams et al. 1994, O’Connell and Grove 1996, Keith 1997, Wan et al. 2001).

1.5.5 Post-fire microclimate

Vegetation damage and the destruction of leaf litter by fire, alters light, temperature, humidity and soil moisture regimes with the magnitude of the change generally proportional to fire intensity. Alterations in microclimate can have a wide ranging influence, affecting plant growth, animal habitat, soil microbiota, litter decomposition and nutrition cycling. While light and humidity regimes are affected relatively simply in accordance with rates of canopy loss and recovery, the changes in soil temperature and moisture after fire can be more complex.

Temperatures at the soil surface generally increase after fire as the soil surface is blackened and more exposed due to the combustion of leaf litter (Walker et al. 1986). For example, increases of up to 8°C in average summer day-time temperatures were recorded for surface soils in recently burnt compared to unburnt E. pauciflora forest (Raison et al. 1986). Forest soils exposed to sunlight in southern Australia can reach temperatures of 50 to 60°C (Warcup 1983). If such temperatures are sustained for several hours, a mild sterilization effect may occur (Warcup 1983). With extended periods of exposure to high temperatures in summer, the warming effect may influence temperatures as deep as 20 cm in some soils (Viro 1974).

25 The moisture content of surface soil layers declines in the short-term following fire (Chandler et al. 1983). Studies in eucalypt forests concur with this generalization (Hingston and Malajczuk 1981, Woods et al. 1983, York 1999). The extent of removal of the litter layer, which insulates the soil from heating and provides a barrier to evaporation, is one important factor determining the magnitude of the immediate change (Viro 1974, Chandler et al. 1983). The formation or enhancement of water-repellant layers due to fire (Biswell 1974, Humphreys and Craig 1981) is another. Eucalypt forest soils can be water-repellant in the long absence of fire (Doerr et al. 2004). However, when heated, repellency can be induced or increased until temperatures above 250°C are sustained, following which water repellency is eliminated (Doerr et al. 2004).

When long-term assessments are made covering an entire year, surface soil moisture at burnt sites may sometimes be higher than at comparable unburnt sites for several reasons: during the subsequent wet season the reduction in plant and litter cover limits interception losses (Humphreys and Craig 1981, DeMarco et al. 2005), water repellency becomes less significant as more rainfall reaches the soil surface (Biswell 1974, Carter 2002), the input of organic matter derived from the products of combustion may improve soil moisture retention (DeMarco et al. 2005), and a reduced demand for soil moisture occurs for some period until recovery of leaf area at a site is complete (Eamus et al. 2006). As with soil biological and chemical changes following fire, significant within site temporal and spatial complexity can be expected for surface soil moisture according to patterns of fuel consumption and fire intensity.

The relationship between increased groundwater recharge and vegetation removal is well established (Peck 1978). Thus, post-fire increases in groundwater at depth in the soil can occur in the months or initial years after fire (Eamus et al. 2006). In recognition of this view, some management authorities are considering regular burning of forest as a possible method of increasing water resources (Hansard 2005). However, where dense eucalypt regrowth develops following high intensity fire, leaf area index can exceed pre-fire levels, thereby reducing groundwater availability for many decades until a mature stand develops (Eamus et al. 2006).

1.5.6 Phytophagy

As consumers of plants, both vertebrates and invertebrates affect the health and regeneration of plants, sometimes to the extent of shaping landscape vegetation patterns 26 (Hairston et al. 1960 in Bond and van Wilgen 1996). In eucalypt forests, phytophagous invertebrates are a diverse group including leaf chewers, leaf miners, sap suckers, wood borers, gall formers, flower feeders and seed eaters (Landsberg and Cork 1997). Vertebrates include arboreal herbivores (for example, possums [Pseudocheirus spp.]), ground-dwelling native and introduced herbivores (for example, kangaroos [Macropus spp.] and rabbits [Oryctolagus spp.]) and granivores (for example, cockatoos [Calyptorhynchus spp.]) (Landsberg and Cork 1997).

Overall, knowledge of invertebrate responses to fire has been somewhat limited by the diversity of life forms, the complex population patterns in time and space, the variability of fires and the challenges associated with sampling and experimental design (Whelan et al. 2002, van Heurck and Abbott 2003). Ants are one of the invertebrate groups that have been better studied, with many reports of increased or unchanged abundance after fire (Ahlgren 1974a, Whelan et al. 2002). This trend is particularly significant for eucalypt regeneration given the ability of ant populations to remove large quantities of seed (O’Dowd and Gill 1984, Wellington and Noble 1985b, Yates 1994, Ruthrof 2001) (discussed in section 1.6.3).

The use of fire to control insect populations has a long history of application in agriculture (McCullough et al. 1998). Control of some invertebrate pests in eucalypt forest by fire both directly or indirectly by habitat modification is conceivable following the alignment of certain species life-cycle, site, weather, seasonal and fire variable combinations (Campbell 1961, Carne and Taylor 1984, Farr et al. 2004). However, attempts at direct control using controlled fire have generally been unsuccessful due to incomplete mortality from suspected fire intensity variation as well as the rapid recolonization by individuals from surrounding unburnt areas (for example, Campbell 1961, Abbott et al. 1999). Nevertheless, regenerating plants in the middle of larger burn areas may be subject to less grazing than other plants for a period while recolonization takes place (Whelan and Main 1979).

The direct impact of fire on the size of vertebrate populations in Australia is largely unknown (Whelan et al. 2002). Apart from mortality, fast-moving animals may emigrate from an area as a fire burns (Biswell 1974, Whelan et al. 2002). Burning of large areas may lead to significant temporary reduction in vertebrate herbivory (Whelan and Main 1979, Baird 1958 and Storr 1963 in Rippey and Hobbs 2003), with

27 recolonization rates varying between species according to mobility (Whelan et al. 2002).

Indirectly, fire can influence phytophagy in burnt areas by altering habitat or the pattern of food resources (Bendell 1974, Bond and van Wilgen 1986, Whelan et al. 2002). As the lush post-fire regrowth often attracts herbivores, grazing pressure may gradually become higher in burnt than in unburnt areas, particularly where they are small in extent (Cowling and Lamont 1987, Bond and van Wilgen 1996, Cohn and Bradstock 2000). Changes in patterns of herbivory post-fire arising from the attractiveness of resprouting leaves (Landsberg 1990) and the post-fire changes in foliar nutrition are also possible (Hanley and Lamont 2000, Radho-Toly et al. 2001, Rieske 2002). While this is discussed further for seedlings in section 1.6.3., the complex, high order interactions involved with herbivory and the physical and chemical properties of post-fire resprout foliage is considered outside the scope of this introduction.

1.5.7 Competition

Competition has traditionally been viewed as less important in the organization of plant communities where frequent fire (or other disturbance) causes significant mortality (Huston 1979, Bond and van Wilgen 1996). In its simplest form, this view fails to account for the range of fire types and differences in species’ response to fire (Bond and van Wilgen 1996, Huston 2003). For example, a low intensity fire may burn through the ground fuel layer promoting the vigour of certain resprouting shrub species (for example, Lamont et al. 2004) and tree species (for example, Bell and Chatto 2003) in the years following fire. The competition–disturbance frequency tradeoff is most likely found in highly productive environments where rapid growth rates allow species to quickly dominate and exclude others following disturbance (Huston 1979, 2003). Where climatic factors allow, frequent fire may be an important process in enabling the co- existence of certain competitors in such environments (Bond and van Wilgen 1996, Huston 2003). For example, frequent fire is important in the co-existence of trees and grass in savannah formations and variations in disturbance frequency, either higher or lower can lead to fluctuations between grassland and forest (Ellis 1985, Bond and van Wilgen 1996, Higgins et al. 2000, Fensham and Fairfax 2006).

As fire alters resource availability (light, water and nutrients) it has implications for resource competition. Although, with fire the situation is complicated greatly by the 28 temporary loss of plant density and/or size (Bond and van Wilgen 1996) and the fact that the size of the resource pulse is generally proportional to fire severity (sections 1.5.4 and 1.5.5) and the level of plant destruction (section 1.5.1). Those individuals that resprout or regenerate from seed more rapidly after fire are most able to exploit this resource pulse (Pate et al. 1990, Huston 2003), but whether such a process alone represents competition is controversial (Tilman 1987 versus Grime 1977, Craine et al. 2005 versus Tilman 1987, and reviews in Goldberg 1990 and Grace 1990). Nevertheless, the greater occupation of space above and below-ground that accompanies rapid growth or recovery, likely enhances competitive ability (Fowler 1990), at least under a “short-term” (within the life-span of the individual) definition of competition (Grace 1990, Goldberg 1990). For example, occupation of root-space may be important in long-term soil resource capture to the detriment of other plants (Schenk et al. 1999) and taller, larger canopies can capture light and shade-out other plants.

A role for fire in affecting direct competition between plants by chemical exudates (allelopathy) has been suggested by Zackrisson et al. (1996). Their work showed that charcoal formed by fire in the boreal forest soils of Sweden absorbs significant quantities of phenolic allelochemicals for up to 100 years after its formation. Research on species from these forests has revealed allelopathic effects of Empetrum spp. on Pinus silvestris seedling development (Nilsson et al. 1993, Nilsson 1994) including via phenolic compounds (Nilsson et al. 2000). However, in the majority of ecosystems, relatively little progress has been made in demonstrating allelopathy as a significant ecological process under field conditions (Williamson 1990, Inderjit 1995, Florence 1996). A recent review of allelopathy in eucalypt ecosystems reached a similar conclusion (Willis 1999). Rather than direct interference, the indirect effect of allelochemicals on plants via microbial populations and nutrient cycling may be of greater significance (Rice 1984, Nilsson et al. 1993, Nilsson 1994, Wardle et al. 1998, Willis 1999). Considering its influence on soil microbiology (section 1.5.3), soil chemistry (section 1.5.4) and microclimate (section 1.5.5), fire may play an important role in such interactions.

Theories on the interdependence between fire and community types (for example, Mount 1964, Mutch 1970) have led to proposals that flammability may be a form of interspecific competition (Bond and Midgley 1995, Bond and van Wilgen 1996, Bowman 2003). It is argued that “torches” (flammable species) may spread in

29 communities where they mingle with “damps” (less flammable species), if “torches” regenerate more successfully than “damps” in the post-fire environment (Bond and Midgley 1995, Bond and van Wilgen 1996). For example, the retention of dead branches on the stem of the Californian shrub Adenostoma fasciculatum, increases local fire intensity; the species regenerates from soil stored seed after fire and any benefit from dead branch retention is not apparent, this architectural trait promoting flammability could logically have been selected (Schwilk 2003). Conversley, plants may protect themselves from burning by reducing local fuel accumulation by allelopathic means (Williamson 1990). Clearly, processes involving flammability and competition remain mostly theoretical at this stage.

1.6 FIRE AND EUCALYPT POPULATION PERSISTENCE

1.6.1 Survival and recovery of individuals exposed to fire

The ability of many species to resprout is one mechanism ensuring the recovery of Australian sclerophyll vegetation following fire (Gardner 1957, Gill 1975, Christensen et al. 1981, Gill 1981a, c, Specht 1981, Bell 2001). The majority of eucalypts exhibit a strong tendency to resprout following fire (Jacobs 1955, Gill 1997, Nicolle 2006). Notable exceptions include some tall-growing species in the highest rainfall zones of southern Australia (for example, E. delegatensis and E. regnans) and a number of species that grow in the low to medium rainfall zone of southwestern Australia (for example, E. astringens and E. spathulata) (Nicolle 2006). Eucalypt species that resprout after fire have thick bark on the stem and/or a lignotuber (Jacobs 1955, Gardner 1957, Gill 1975). By growing tall, many also avoid damage to sensitive crown branches or total canopy scorch (Luke and McArthur 1978, Gill and Catling 2002).

Stem bark thicknesses and fire intensity principally govern resprouting height in eucalypts surviving fire (McCaw et al. 1994, Grant et al. 1997). Aerial resprouting (above 2 m) is commonly reported for eucalypts (> 10 cm diameter at breast height) exposed to canopy scorching fires (for example, Burrows et al. 1990, McCaw et al. 1994, Grant et al. 1997, Wardell-Johnson 2000). The ability to resprout aerially from tall, surviving stems is advantageous in that rates of refoliation are potentially greater due to the larger number of available meristems (Hodgkinson 1998). In addition, the redevelopment of a canopy at height is less likely to be impeded by shading, and

30 herbivory from ground-dwelling animals (Bond and van Wilgen 1996). A more rapid rate of canopy restoration maximizes the ability to capitalize on the post-fire pulse of soil nutrient and water availability (Pate et al. 1990).

As with many plants, stores of non-structural carbohydrates (NSC) are important in the initiation of resprouts following defoliation in eucalypts (Bamber and Humphreys 1965). NSC are stored in tissues within stems (Bamber and Humphreys 1965), roots (Wildy and Pate 2002) and the lignotubers (Bamber and Mullette 1978, Mullette and Bamber 1978) of eucalypts. The storage tissues within these structures and the vascular connections to buds must be protected from lethal temperatures for resprouting to occur. In eucalypts this is achieved by thick bark and/or below-ground storage of NSC and buds (Jacobs 1955, Gill 1975).

The lignotuber, a swelling that develops at the root crown, is a structure common to a range of plant families in Mediterranean-type environments (James 1984) and is important in the ecology of many eucalypts (Jacobs 1955, Gardner 1957, Abbott and Loneragen 1984, Noble 2001). Except for the wet sclerophyll forest zone, lignotuberous eucalypt species are common and distributed widely (Nicolle 2006). As both a store of NSC and a source of buds, lignotubers enable rapid resprouting following canopy loss (Jacobs 1955, Noble 2001). However, recent findings have suggested a greater proportion of the total pool of NSC is stored in the roots rather than lignotubers (Cruz and Moreno 2001, Wildy and Pate 2002). Thus, the most important function of the lignotuber in eucalypts at the tree stage may be simply as a “bud bank” (Noble 2001, Wildy and Pate 2002). The depth of burial of the lignotuber in the soil, in addition to degree of soil heating from fire, are important interacting factors determining lignotuber survival (Beadle 1940, Jacobs 1955, Noble 1984, Gill 1997). The insulation of the bud bank below the soil (if at sufficient depth) may be a more reliable means of surviving high intensity fire in environments where major factors limiting growth (for example, low rainfall) may restrict the ability to resprout aerially (Noble 2001).

The formation of the lignotuber may be more important in the seedling and young sapling than in subsequent life-stages, as the small size of seedlings makes them particularly vulnerable to mortality from fire, grazing, drought and competition (Abbott and Loneragen 1984, Gill 1997). Thus, the seedlings of species with a lignotuber may be more resilient than those without. For example, E. marginata is a lignotuberous

31 eucalypt with considerable capacity to persist through both disturbance and competition (Abbott and Loneragen 1984). Complete annual defoliations of E. marginata saplings over a period of 15 years did not result in any mortality (Wills et al. 2004). These authors contrasted their findings with those where significant mortality occurred following less frequent defoliation of saplings of non-lignotuberous species E. nitens and E. regnans (Wilkinson and Neilson 1995).

For non-lignotuberous trees, attaining a large size at an early age provides some resistance against fire, grazing and being shaded by understorey vegetation (Ashton 1981, Bond and van Wilgen 1996, Higgins et. al. 2000). The trade-off in investing carbon to ensure the ability to resprout (survivability or persistence) is generally a reduction in above-ground growth (Ashton 1981, Pate et al. 1990, Bowen 1991, Bond and Midgley 2001). Thus, the absence of a lignotuber probably contributes to the ability of seedlings of those species growing in productive environments (for example, E. regnans and E. diversicolor) to gain rapid height growth (Ashton 1981). However, when suppressed by competition, bark thicknesses are limited, increasing the vulnerability of such saplings to mortality from fire (Ashton 1981).

Frequent fire potentially impairs resprouting capacity when interfire periods are insufficient to allow the replenishment of NSC reserves. Repeated defoliation has been demonstrated to deplete plant NSC stores and result in death for shrub species (Bowen and Pate 1993, Canadell and Lopez-Soria 1998), and in eucalypts as seedlings (Wilkinson and Neilson 1995) and possibly trees (Bamber and Humphreys 1965, Cremer 1966). However, the importance of carbohydrate reserve levels as a limitation on resprouting is open to question (Caldwell 1986, Bond and van Wilgen 1996). For example, Wildy and Pate (2002) suspected that falling starch levels were more likely symptomatic of a general decline in health rather than a cause of death in the case of the mallee E. kochii subject to frequent cutting. They inferred that only a small proportion of reserves of below-ground starch in E. kochii was associated with the initiation of resprouting. Levels in recently dead individuals subject to shade and repeated defoliation were only slightly lower than in wild individuals recovering from fire. The physiological consequences of sustaining a reduced shoot:root ratio over an extended time has been proposed as a more important factor in the health of eucalypt mallee species (Noble 2001, Wildy and Pate 2002). For plants subject to incomplete defoliation, such as where low intensity fire incompletely scorches the canopy of tall

32 trees, NSC reserve depletion may be of little relevance as supply of carbohydrate from concurrent photosynthesis may be sufficient to initiate resprouting (Bond and van Wilgen 1996).

High intensity fire is generally the greatest cause of crown damage in eucalypt forests (Neumann et al. 1981, Abbott and Loneragen 1983). Damage can be permanent, particularly to the ash-type eucalypts such as E. regnans (Neumann et al. 1981) and large old trees due to their poorer post-fire resprouting capacity (Bell et al. 1992, Gill 1997). Fire scars and dead branches arising from intense and/or severe fire can provide entry points for fungal pathogens, insect borers and termites (Greaves et al. 1965, McCaw 1983, Carne and Taylor 1984, Pook and Forrester 1984, Ashton 2000a). Intense fires also affect the resistance of surviving stems to subsequent fires due to the loss of bark thickness (Gill 1981a, Bell and Chatto 2003) and the probable trade-off between the restoration of bark thickness and crown recovery (Gill 1980, 1997). Eventual stem death from declining bark thickness associated with frequent fire can be envisaged (Ashton 1981, Gill 1981a). Where fire intensity is sufficient to kill crown branches and lead to stem resprouting, epicormic growth along the tree bole can then connect surface fuels to crown fuels and thus raise the potential for further crown damage in the event of another fire in short succession (Ashton 1981). If the stem cambium is also damaged and resprouts originate basally or below-ground, the recovering canopy is potentially exposed to light competition or further defoliation from ground-dwelling herbivores.

Variability in the literature for survival rates of eucalypts after fire is not surprising considering the number of potential factors involved (size, age, species, health, fire intensity, fire history health). For eucalypt seedlings, few fire survival studies have been undertaken (Fensham and Fairfax 2006). The variability reported in seedling survival studies is attributed to the range of seedling sizes and fire intensities within sites (Noble 1984, Fensham and Fairfax 2006). Similar variability may also occur with saplings, particularly for more fire sensitive (non-sprouter) species such as E. delegatensis (Bowman and Kirkpatrick 1986a). However, as resprouter species reach the sapling- young tree stage they appear to be resistant to most fires. For example, in a study of young trees of resprouter species (E. diversicolor, Corymbia calophylla and E. guilfoylei at seven and 14 years of age), 92 to 100% survival was recorded after high intensity fire (Wardell-Johnson 2000). Survival was also 100 % where high intensity fire burnt 16-year-old E. diversicolor and E. muellerana (McCaw et al. 1994). For

33 mature resprouter-species of eucalypt, it also appears that few trees are killed by fire. For example, in a study of four co-occurring eucalypt resprouter species (E. jacksonii, E. guilfoylei, E. diversicolor and C. calophylla) where moderate and high fire intensity resulted in 100 % canopy scorch, mortality varied between 0 and 12 % depending on the species (Wardell-Johnson 2000); for E. marginata trees at three sites subject to high intensity fire or wildfire (presumably 100 % crown scorch resulted), mortality ranged from 0 to just 6 % (Abbott and Loneragen 1983).

1.6.2 The growth response of trees surviving fire

Growth increments associated with fire have been reported for eucalypts in some studies. Wallace (1966) reported that basal area increments for E. marginata in the years following an intense fire were up to four times as high as in the pre-fire period, but no details of sample size or methodology were noted. In another study of E. marginata, pre-fire and post-fire bole diameter measurements uncovered a significant increase for two of three sites subject to intense fires, but no difference for four sites burnt by frequent (range 3 to 11 years), low-intensity fire (Abbott and Loneragen 1983). In contrast, Mazanec (1999) detected short-term (one year) girth increments following low-intensity spring burning on four out of five occasions across two sites for E. marginata. Bell and Chatto (2003) also found an increase in stem diameter growth for E. obliqua and E. rubida in forest sites subjected to 3 to 4 or 1 to 2 prescribed fires over a 14 year period. Whereas a study of both dry-sclerophyll and wet-sclerophyll eucalypt forests in southeastern Queensland subjected to frequent burning (every 1 to 4 years), produced mixed results (Guinto et al. 1999). While two species (E. tereticornis and C. variegata) at the dry site displayed a positive diameter increment response, remaining species (E. drepanophylla, E. acmenoides, E. pilularis, E. intermedia, E. microcorys and E. resinifera) across both sites were found to be unaffected (Guinto et al. 1999). In a review of several other studies, Shea et al. (1981) concluded that the results for low intensity burning were also mixed. Clearly, a growth stimulus following low-intensity and high-intensity fire can occur, but not always.

Observations suggest older or larger trees do not resprout as vigourously as young trees and saplings (Gill 1975, Baird 1977, Bell et al. 1992, Cochrane 1968 in McCaw et al. 1994). Gill (1997) proposes that some trees may lose buds from basal portions of the bole as they age. Further, the physiological decline accompanying senescence may also be a factor (Gill 1981a). 34 Regardless of the defoliation agent, it is well established that differences in the structure and physiology at both the leaf and whole-plant level following defoliation and resprouting can invoke a vigour response in woody plants (reviewed in Wildy et al. 2004). For example, Cremer (1966) observed that the epicormic crowns of eucalypts recovering after drought had more foliage than before the drought or than on trees that were undamaged. A reduction in shoot:root ratio improves water relations (for example, Crombie 1997) which may enhance leaf photosynthetic rates (for example, Kruger and Reich 1993). The increased carbohydrate sink:source ratio following following defoliation can also contribute to a higher photosynthetic capacity (Medhurst et al. 2006). In addition, resprouting leaves, including those of eucalypts, may also be equipped with a higher photosynthetic capacity enabled by a greater specific leaf area (SLA) and higher nutrient concentrations (Landsberg 1990). Leaf photosynthetic rates may also be unchanged (Oechel and Hastings 1983) or even lower in resprouting leaves if the attainment of a higher SLA results in a lower total chlorophyll content (Wildy et al. 2004). Nevertheless, growth rates may still be greater where SLA is higher as less photosynthate is necessary to allocate per additional unit of leaf area constructed (Wildy et al. 2004). Overall, the mechanisms underlying changes in photosynthetic performance by plants following defoliation vary between species and according to the degree of defoliation (Medhurst et al. 2006).

Where the defoliating agent is fire, increases in soil nutrient availability potentially enhance growth. Plant-available nutrient increases are generally proportional to fire severity (section 1.5.4). The well-known stimulus of nutrient enriched ashbed to seedling growth is outlined below (section 1.6.3). Abbott and Loneragen (1983) implied that the ashbed effect was probably an influential factor in distinguishing the positive effect on the growth of E. marginata trees from high intensity fire versus the absence of an effect from low intensity burning. However, for eucalypts as trees, there are few if any studies that have examined the significance of the pulse in nutrient availability following fire, or whether a long absence of it may limit growth. Turner and Lambert (2005) state that there is little evidence to suggest that nutrient “lock-up” in the biomass is a significant process affecting eucalypt health in native forests, whereas a reduction in nutrient availability with a decline in fire frequency has been proposed as a mechanism reducing the growth of Xanthorrhea preissii in southwestern Australian forests (Swanborough et al. 2003). The significance of regular pulses of increased nutrient availability for eucalypt growth and health deserves further investigation.

35 Much research has linked the beneficial effects of soil sterilization from heating by fire on eucalypt seedling vigour (Florence and Crocker 1962, Renbuss 1973, Ashton and Willis 1982, Warcup 1983) but the work has generally not been extended to trees. Florence (1981) proposed that disease sensitive fine roots may be more able to proliferate within sterilized soil, this being particularly advantageous given the post-fire increase in nutrients. Alternatively, as the work of Ellis and Pennington (1992) indicated, soil conditions favourable to particular mycorrhizal fungi following fire may benefit tree health. They found that E. delegatensis seedlings grew poorly in soil from stands where the trees were in poor health. Subsequent pot inoculation treatments with field soils revealed differences in the mycorrhizal associations. Earlier field trials had illustrated the role of fire in restoring the health of the eucalypt overstorey (Ellis 1981). Thus it was proposed that in the absence of fire or other disturbance, rainforest plants invade the understorey and mycorrhizae beneficial to E. delegatensis are disfavoured under these conditions (Ellis and Pennington 1992).

The contribution of fire in relieving trees from the stress from soil-borne pathogens may be only minor or insignificant. The impact of fire on P. cinnamomi by soil sterilization is minimal as inoculum remains distributed across the site due to the patchiness of soil heating (Podger and Keane 2000). Further, the pathogen survives at considerable depth in the soil, beyond the influence of even the most severe of fires (G. Hardy pers comm.). However, Shea (1979) and Shea et al. (1981) suggested a role for fire in indirectly limiting the impact and spread of the disease, thereby improving tree health. They argued that sufficiently intense fires could alter the composition of the understorey, reducing the abundance of the highly susceptible Banksias spp. and promoting the dense growth of resistant leguminous species. Further, they proposed that this would disfavour the pathogen by reducing spring soil temperatures. However, not only were such burning practices proven to be relatively impractical, they also resulted in serious tree bole damage and were not demonstrated to have led to any reduction in disease impact or spread (Podger and Keane 2000). Nevertheless, the presence of Acacia spp. has been demonstrated to suppress the activity of P. cinnamomi by biological means (D’Souza et al. 2004). Thus, the potential for fire regimes to exert some influence over understorey composition and therefore disease severity at a site remains plausible.

Fire can theoretically result in tree growth increments by reducing understorey competition (Wallace 1966). Competition from understorey development under a

36 regime of less-frequent fire has been suggested as a factor affecting eucalypt tree health in parts of southern Australia (Beard 1967, Withers and Ashton 1977). However, Abbott and Loneragen (1983) disputed understorey competition as significant in their results where a growth response for E. marginata was recorded after high-intensity fire. They considered that frequent low-intensity fire was more likely to reduce understorey development, but no growth response was revealed for trees under such a regime. In addition, the control of healthy overstorey over understorey development in dry- sclerophyll eucalypt forests and woodlands suggests established trees are the dominant competitors in these environments (Lamont 1985, Florence 2001). Therefore, until results of manipulative studies are presented, the significance of understorey competition on eucalypt tree health will remain speculative.

The inconsistent response of tree growth to fire appears to stem primarily from variability in fire intensity and/or severity. There is a probable trade-off between the sufficient release of plant-available nutrients, the rise in pH, the degree of soil sterilization and the increase in soil moisture at depth (sections 1.5.3 to 1.5.5), against the negative impacts of excessive intensity/severity resulting in tree damage (section 1.6.1) including to surface roots (Abbott and Loneragen 1983), as well as the temporary loss of beneficial fungi (section 1.5.3). Further confounding factors appear to involve species differences, tree age/size, initial tree health (section 1.6.1) and the favourability of seasonal conditions following fire. Presumably, the coincidence of favourable rainfall and temperature regimes during the period of recovery would also influence the magnitude of any vigour response (Malanson and Trabaud 1988, McCaw et al. 1994). In support of this, more rapid crown recovery has been observed following spring defoliation of eucalypts by cutting (Wildy and Pate 2002), and of woody shrubs (Baird 1977, Hobbs and Atkins 1990) and eucalypts (Grant et al. 1997) after fire in spring than later in the season. However, determining the effect of seasonal differences of fire occurrence in isolation is difficult because of the interdependence of this variable with intensity and severity.

1.6.3 Fire and seedling regeneration

The link between fire and the successful seedling regeneration of many Eucalyptus spp. across a range of community types has been well documented (for example, Jacobs 1955, Mount 1964, Withers 1978a, Ashton 1981, Wellington and Noble 1985a, Bowman and Kirkpatrick 1986a, Burrows et al. 1990). Following a fire event of 37 sufficient intensity to scorch the overstorey canopy, eucalypt seed is released enmasse from the protective woody capsules as they dry and open in the days or weeks following the fire (Cremer 1965a, O’Dowd and Gill 1984, Burrows et al. 1990). Eucalypt seed released into the immediate post-fire environment has a greater probability of germinating and leading to seedling establishment. The reasons for this are outlined below.

The removal by ants of the majority and sometimes all eucalypt seed in experimental caches can occur within a period of a week (O'Dowd and Gill 1984, Wellington and Noble 1985b, Yates 1994), but predator satiation following the mass release of seed after fire can lead to a decline in rates of removal (granivory) by as much as several orders of magnitude (O’Dowd and Gill 1984). However, the significance of predator satiation on total seed survival and ultimately recruitment is arguable. Eucalypt germinants are often found at undisturbed sites when careful surveys are conducted in winter and spring, but not in summer by which time mortality is virtually complete (Jacobs 1955, Purdie 1977, Wellington and Noble 1985b). Thus, it is claimed that the abundance of “safe sites” (those sites with the potential to support sustained seedling growth) rather than seed predation, is generally a greater limiting factor on recruitment (Wellington and Noble 1985b, Anderson 1989).

The season in which fire occurs, triggering the release of canopy seed stores, has important implications for seedling establishment. If spring fires are followed by suitable germination conditions (governed by temperature and adequate moisture) the newly developing seedlings may be exposed to the summer drought prior to the development of a deep root system (Baird 1977, Cowling and Lamont 1987). Where suitable germination conditions do not occur until the following autumn or winter, the seed population is at greater risk from granivory (Cowling and Lamont 1987). In addition, loss of viability of eucalypt seed can be very high or even complete within the year following release (Cremer 1965b, Ruthrof 2001). Therefore, spring fires appear least favourable for seed to survive until germination, and then for germinants to become established seedlings.

Year to year variability in rainfall within the Mediterranean climate of southern Australia increases the significance of the timing of fire in terms of the success of seedling regeneration (Wellington and Noble 1985a, Bradstock and Bedward 1992,

38 Whelan and Tait 1995, Vesk and Dorrough 2006). Droughting is often reported as the major factor in eucalypt seedling mortality in the pre-establishment period in dry forests and woodlands (for example, Purdie 1977, Withers 1978b, Wellington and Noble 1985a, Stoneman 1992). Therefore, particularly in the lower-rainfall woodland and mallee regions, the coincidence of both fire and favourable seasonal conditions may be irregular and thus significant seedling establishment events rare (Wellington and Noble 1985a).

Ashbeds created by fire especially benefit eucalypt seedling survival and growth (Pryor 1963, Burrows et al. 1990). For example, Burrows et al. (1990) found that E. wandoo seedlings established at mean densities of 9.2 and 5.5 individuals m2 on ashbeds and 0 and 0.6 individuals m2 off ashbeds 14 months after fire at two sites. Further, the mean height of ashbed seedlings was at least 2.5 times those off ashbeds. The ashbed growth stimulus (“the ashbed effect”) is attributed largely to the increase in plant-available soil nutrients (Loneragen and Loneragen 1964, Humphreys and Lambert 1965, Chambers and Attiwill 1994, Launonen et al. 1999). This likely enables greater rooting depth which is important to minimize drought stress in the first year of growth (Wellington and Noble 1985a, Burrows et al. 1990), particularly as soil surface layers are usually drier after fire (Section 1.5.5 above). In high-rainfall zones, the growth stimulus is also important to prevent seedlings being shaded-out by the vigorous understorey plants that also develop post-fire (Ashton 1981, Ashton 2000b). Early attainment of a larger size also reduces the vulnerability of a seedling to destruction from leaf disease or herbivory (Westoby et al. 1997, Kitajima and Fenner 2000).

While seedlings may derive growth benefits from the greater availability of nutrients on ashbeds, there may be negative consequences arising from the nutritious foliage being attractive to herbivores. Increased levels of herbivory by insects on eucalypt leaves with higher foliar nutrient concentrations has been demonstrated (Landsberg 1990). However, few studies of post-fire herbivory have been carried out in Australia (Whelan et al. 2002). Both reductions (Whelan and Main 1979, Hanley and Lamont 2000) and increases (Radho-Toly et al. 2001) have been observed. It is argued that factors such as leaf toughness, moisture content and chemical defense compounds are of greater importance than foliar nutrition on levels of herbivory (Majer et al. 1992, Abbott et al. 1993, Hanley and Lamont 2000, Peeters 2002). Also, the direct effects of fire on herbivore populations may be significant depending on burn area extent and intensity

39 (Section 1.5.6). In summary, there is insufficient information upon which to generalize about how herbivory impacts on seedling development in the post-fire environment.

Sterilization of the soil from heating also contributes to the ashbed effect (Pryor 1963, Renbuss et al. 1973, Warcup 1981). Potential microbiotic competitors for nutrients, and organisms that are antagonistic to root growth, may be eliminated (Florence and Crocker 1962, Renbuss et al. 1973, Ashton and Willis 1982, Guerrero et al. 2005). Those sites where soil moisture is less seasonally variable are believed to be most favourable to the activity of inhibitory soil microbiota (Florence and Crocker 1962). The change in the types of mycorrhizae as well as the increase in hyphal growth rates that occur following the mild heat treatment of soil (60°C for 30 minutes) could also play a role in the ashbed effect (Warcup 1983). In contrast, deep and complete sterilization from prolonged heating of the soil at very high temperatures may have negative consequences for seedlings germinating in ashbeds. Launonen et al. (1999) concluded that seedling growth may be limited in such ashbeds (generally orange in colour) by the combination of soil N depletion via volatilization and the temporary, localized loss of mycorrhizal fungi. For example, the mean dry weight of E. regnans seedlings grown for 15 weeks in soils collected two months after a fire was several orders of magnitude lower in orange ashbed soils compared to black ashbed soils, and less than half of that for unburnt soils (Launonen et al. 1999).

The success of seedling growth on ashbeds may also be assisted by the localized destruction of both surface roots and the soil seed store of potential competitors (Jacobs 1955, Cunnigham 1958 and Campbell 1956 in Burrows et al. 1990). Developing eucalypt seedlings are vulnerable to competition for soil moisture and/or nutrients from a vigorous layer of herbs (Withers 1978a, Vesk and Dorrough 2006) and shrubs (Cunningham and Cremer 1965). Trees may also negatively affect eucalypt seedling growth and survival (Bowman and Kirkpatrick 1986a, c, Stoneman 1992), but the overall effect of the overstorey on eucalypt establishment appears complex. Some researchers have reported little effect of the overstorey on survival in the establishment phase (for example, Dignan et al. 1998, Abbott 1984). In a review of experiments across ecosystems worldwide, Coomes and Grubb (2000) noted a “nurse effect” from the overstorey at the post-emergent stage was commonly reported in dry environments. On the other hand, in moist climates supporting tall forest types shading may result in significant seedling losses by favouring the development of leaf disease (Ashton

40 2000b). Therefore, the net effect of an overstorey on seedling survival from competition versus the reduction in evapotranspiration from shading, likely varies with a range of factors such as a species’ drought tolerance, tree density, climate and soil.

Eucalypt seedlings are considered “intolerant” (of competition) and therefore require the light, soil moisture and nutrients of gaps to develop beyond establishment to the sapling and tree stage (Jacobs 1955, Florence 1996). Thus suppressed saplings and seedlings are common in undisturbed eucalypt forests (for example, White 1971, Ashton 1976, Bowman and Kirkpatrick 1986b, c, Abbott and Loneragen 1984). By damaging or killing overstorey and mid-storey trees, fires have the potential to create gaps where seedlings can develop from seed or redevelop from lignotubers (Abbott and Loneragen 1984); the number and size of the gaps created is dependent upon fire intensity and severity as well as stand composition, age and condition. Those species that do not develop lignotubers and occur in productive environments are particularly reliant on both gaps and ashbeds to develop beyond establishment stage (White 1971, Ashton 1981, Ashton 2000b). Such species (for example, E. regnans) can appear closer to the C-type strategy of along the C-S axis of Grime’s (1977) plant ecological strategy scheme (Ashton 2000b). The C (“competitor”) strategy is based on rapid early growth and maximal capture of resources for successful growth and survival, whereas the S (“stress tolerator”) strategy involves lower minimum resource requirements (Grime 1977).

1.6.4 Fire and population persistence

Given the importance of fire in eucalypt establishment, in the long absence of fire (centuries) eucalypt populations are less likely to persist. As eucalypt seed is only temporarily stored in the soil (around one year maximum) (Cremer 1965b, Ruthrof 2001) seedling establishment within the tree lifespan (estimated as typically around 400 years [Jacobs 1955, Ogden 1978]) is necessary for population persistence. As opposed to stand-replacement recruitment, multiple seedling establishment events within periods less than the lifespan of individuals (gap replacement) can confer resilience to the population, allowing the opportunity for senescent and diseased individuals to be gradually replaced (Jacobs 1955, Florence 1996). While possible for wet-sclerophyll forest types, gap recruitment is more associated with dry-sclerophyll eucalypt stands due to fire regimes (more frequent, less intense), the more open stand structures and the resprouting characteristics of the trees (Florence 1996). Towards the driest regions in 41 which eucalypts occur, rainfall variability limits possible seedling establishment to a proportion of fire events, and therefore modes of recruitment may also become more “wave-like” rather than “step-like” (Wellington and Noble 1985a).

As time elapses after fire, non-eucalypt tree species that are tolerant and therefore recruit in interfire periods, can become increasingly prominent in both dry-sclerophyll formations (for example, Withers and Ashton 1977, Lunt 1998a, b) and in wet- sclerophyll forests (for example, Cunningham and Cremer 1965, Ellis 1985). Hence there are some similarities with the widespread successional-type forest declines of the United States where tolerant tree species are increasing stand densities and achieving dominance over intolerant species in the long absence of fire (for example, Fule and Covington 1998, Gilliam and Platt 1999, Abrams 1992, 2003). These dense stands are believed to be more vulnerable to insect outbreaks and drought (McCullough et al. 1998, Arno and Fiedler 2005). However, eucalypt canopy decline due to stand compositional and structural changes in the absence of fire (for example Withers and Ashton 1977, Jurskis and Turner 2002) is largely conjectural at this stage.

Changed soil conditions from a reduction in fire frequency has been proposed as the common, dominant factor underlying the majority of eucalypt declines across southern Australia (Jurskis and Turner 2002, Jurskis 2005). It is hypothesized that the process of decline involves an excess of soil nutrients and moisture in association with the development of a denser understorey layer (Jurskis and Turner 2002, Turner and Lambert 2005). It is considered that herbivory is often a secondary agent in the decline process (for example bell-miner dieback), due to changes in foliar chemistry arising from the soil conditions (Jurskis and Turner 2002, Turner and Lambert 2005). Sufficient evidence has yet to be presented to support this hypothesis. While the work associated with Ellis (particularly Ellis 1981 and Ellis and Pennington 1992) was frequently cited by Jurskis and Turner (2002) and does provide some basis for the theory (although not necessarily the process), it was carried out in only one ecosystem with characteristics distinct from other areas where eucalypt tree declines occur, such as in Western Australia. In addition, the post-fire vigour response of eucalypts following fire in some instances also suggests a possible role for fire in enhancing tree health (section 1.6.2) but the decline in vigour from an absence of fire does not necessarily follow from these findings. Finally, evidence contrary to the applicability of the Jurskis and Turner (2002) process across extensive areas of southern Australia exists by way of the studies that

42 conclude no major differences in soil chemistry between frequently and long-unburnt sites (Section 1.5.4) and the fact that long-unburnt sites with healthy eucalypt canopies can be found within regions with decline (for example, Edwards 2004). Therefore, considerable research on both pattern and process is needed to evaluate the theory of tree decline and the reduced occurrence of fire as a significant factor in the acceleration of eucalypt tree declines across southern Australia. Even conclusive studies of pattern, attempting to test correlations between fire regime and vegetation change generally, are not abundant due to the lack of long-term studies and the inability to adequately control for site, climate and other non-fire disturbances in the alternative chronosequence (“space for time”) approach (Fensham and Fairfax 2002, Lunt 2002).

Frequent, high-intensity fire appears to be a more obvious threat to eucalypt population persistence, than infrequent burning. Section 1.6.1 outlined several processes whereby the capacity to resist, and recovery after fire may be eroded in a succession of intense/severe fires, especially for non-lignotuberous eucalypts: erosion of bark thickness, secondary damage from insects and fungal pathogens via fire scars and dead branches, progressive hollow-butting, a decline in NSC reserves and the physiological stress of sustaining a high root:shoot ratio. While severe fires can promote waves of seedling establishment through ashbed and gap creation (Section 1.6.3), seedlings and young saplings are vulnerable to mortality from fire (Section 1.6.1) and a period of at least several years is required for seed production capacity to be restored after crown scorch (Gardner 1957, Gill 1975). Thus, a succession of severe fires may both prevent recruitment and lead to the decline of the health of adults, particularly for stands of non- lignotuberous species (Noble and Slayter 1980, Ashton 1981), but possibly also for lignotuberous eucalypts if the fires occur at very short intervals (Noble 2001). Beard (1967) and Florence (2005) both raise the possibility that past post-settlement era of frequent, intense fires along with uncontrolled logging, may have predisposed some forests to the declines in the modern era as a consequence of previous tree damage and stand structural changes.

1.7 CONCLUSION

Fire plays an important role in eucalypt population persistence by influencing the health and regeneration of eucalypts in numerous ways. Considerable complexity arises due to the scope for fire variables to interact with multiple other factors that affect eucalypt 43 health, growth and regeneration (for example, climate, seasonal variability, soil type, pathogens and grazing). A further source of complexity arises from within-site variation in fire behaviour. What is certain is that eucalypt persistence is threatened under the most extreme fire regimes, those being the total absence of fire beyond the life-span of the individual, or very frequent, intense fires. The first is unlikely given climatic factors, the flammability of eucalypt dominated vegetation and the presence of humans. The second is limited to some extent by the fuel accumulation-fire intensity relationship. Nevertheless, the dense regrowth that can develop following intense fires or the presence of introduced annual “fire weeds” in the understorey, ensures flammable fuels are available to support successive intense fire. What is not so clear is how less extreme variations of fire regimes such as infrequent fire, frequent low intensity fire or periodic intense fire impact on the condition of some eucalypt stands. Further research is especially needed to address the issue of whether long periods (decades) without fire are a significant factor in the current extensive declines of eucalypts. Uncertainty and debate about the nature of Aboriginal fire regimes and the effect on the vegetation, is not necessarily relevant from a scientific or a management point of view with regards to fire and tree decline. As noted by Ahlgren (1974b), due to recent and dramatic anthropogenic disturbance of ecosystems, fire now acts on a different “stage”. Duly, there has been much recent interest in the effect of a reduction of fire frequency on the health of eucalypt forests and woodlands in southern Australia. However, as highlighted by Beard (1967) and Florence (2005), the repercussions of the post-settlement exploitation period where severe fires and logging created major disturbance in forests, cannot be ignored either.

44 2 Chapter 2. Tuart (Eucalyptus gomphocephala) and tuart woodlands from the 19th to the 21st century

45 2.1 INTRODUCTION

The health of the tuart (Eucalyptus gomphocephala) tree is deteriorating across an extensive portion of its natural range. While the cause of this decline has not been identified, a wide range of factors including altered fire regimes have been implicated. This chapter reviews the current knowledge of tuart decline. To begin, the tree and its setting are described. Then, the disturbances that have affected tuart woodlands since European settlement in 1829, with particular focus on changing fire regimes, are outlined. Knowledge gaps in relation to the role of fire in tuart health and regeneration are highlighted. These are then linked to the objectives of the research chapters comprising the remainder of the thesis.

2.2 THE TUART TREE

Tuart belongs to the subgenus Symphomyrtus and is found only in the southwest of Western Australia. It can grow as a tall tree (to around 40 m) at the southern end of its extent, but on shallow soils overlying limestone, in exposed coastal positions or at the northern limits of its range, short forms commonly occur (Seddon 1972). Little genetic variation exists across the range of the species, the greatest contrast existing between the small isolated northern stands and the populations at southern localities (Coates et al. 2002). Characteristic morphological features of the tree include the club-shaped operculum, the rough grey box-like bark, light-green foliage and the hard light-coloured wood (Boland et al. 1984) (Figure 2.1).

Flowering occurs from January to April (Boland et al. 1984) with mass flowering events generally occurring once or twice per decade (Keene and Cracknell 1972). As with many other eucalypts, seeds are stored in woody capsules for up to two years, but persist for a shorter period (approximately six months maximum) within the soil after release (Ruthrof 2001). Also in common with other eucalypts, tuart appears to regenerate mostly after fire (Ruthrof 2001), with establishment apparently confined to ashbeds (Keene and Cracknell 1972, Forests Department of Western Australia 1979). When soil and rainfall conditions are suitable, the species is capable of rapid growth, and can form a large tree in around 30 years (Keighery et al. 2002).

46 a) b) Figure 2.1: a) The tuart tree: Eucalyptus gomphocephala, b) tuart leaf, seed capsules and the distinctive flower buds.

The tendency for tuart to regenerate by resprouting is apparent at both the seedling (Forests Department of Western Australia 1979, Ruthrof 2001) and the adult stage (Seddon 1972). Tuart belongs to the small group of eucalypts that do not form a lignotuber (Jacobs 1955): “stem sprouters” (Nicolle 2006). However, this may need revision as Keighery et al. (2002) report the existence of mallee forms on Quindalup dunes south of Bunbury. Like many other eucalypts, resprouting following fire is presumably enabled by the protection afforded to epicormic buds from the bark layer (Jacobs 1955, Gill 1975). Even three-year-old seedlings developed on ashbeds, reportedly resprouted following exposure to wildfire (F. McKinnell, Forests Department of Western Australia, pers. comm.).

2.3 DISTRIBUTION

The natural range of tuart is within a narrow belt coinciding with the Spearwood and the Quindalup dune systems of the coastal plain (Figure 2.2), known as the Swan Coastal Plain (SCP). From Ludlow in the south, the distribution of the tree is virtually continuous until isolated stands occur near Cervantes to the north (Department of Conservation and Land Management et al. 2003). Tuart occurs in a variety of positions in the landscape, from ridges to swales and fringing lakes and swamps, but towards the northern extent of its range, it is often confined to low-lying areas (Keighery et al.

47 2002). Tuart trees have also been successfully grown outside their natural range in Western Australia (for example, Geraldton and Esperance) and in overseas locations (for example, on the Mediterranean coast), the trees being especially valued for their ability to tolerate coastal conditions (Seddon 1972).

Figure 2.2: The current distribution of tuart (data sourced from the Department of Conservation and Land Management et al. 2003) with bounding isohyets labeled. The general location of study sites in this thesis is indicated by the locality names in lighter print.

48 2.4 CLIMATE

Southwestern Australia experiences a Mediterranean-type climate. Within the distribution of tuart, average annual rainfall and mean daily temperature maxima range from 817 mm year -1 and 21.9°C in the south (Busselton) to 560 mm year -1 and 24.7°C in the north (Jurien Bay). Mean monthly rainfall and temperature data are presented for Mandurah (Figure 2.3), being near the centre of the tuart distribution and proximal to many of the study sites that comprise this thesis. The distinct summer drought defines the climate, with less than 10 % of the annual rainfall falling between November and March inclusive. This period is also hot with an average of 18 days above 35°C. Annual average evaporation greatly exceeds rainfall, with the margin increasing to the north. The southern extent of the tuart distribution is located in the 1500 mm annual evaporation zone whereas the northern extremity is in the 2000 mm zone (data sourced from the Bureau of Meteorology 2006).

Research by the Indian Ocean Climate Initiative (Indian Ocean Climate Initiative 2002) indicates that a sharp drop (around 15 to 20 %) in winter rainfall totals occurred during the 1970s. This was offset by increased spring totals, but only partially. In addition, a warming trend has been apparent in the southwest region since the middle of last century (Indian Ocean Climate Initiative 2002).

200 35 180 30 160 140 25 120 20 100 80 15 60 10 Mean rainfall (mm) rainfall Mean

40 Mean max. temp. (°C) 5 20 0 0 Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Month

Figure 2.3: Mean rainfall (■) and daily mean maximum temperature (x) by month for Mandurah (Mandurah Park Station [latitude: -32.5031 S, longitude: 115.7664 E]) from 1889 to 2001. Data sourced from the Bureau of Meteorology 2006.

49 2.5 LANDFORMS AND SOILS

The surface of the western coastal plain is derived from aeolian and marine sediments of Pleistocene and Holocene periods (McArthur and Bettenay 1960). Relief is generally low with heights ranging from near sea-level in the swales and fringes of lakes to in excess of 50 m (rarely) on dune crests (Beard 1990). Units within the dune system can be distinguished on an east-west progression, with older and more highly-leached sands occupying the eastern extents (McArthur and Bettenay 1960). The chemical attributes of representative soil profiles are presented in Table 2.1.

Adjacent to the coast lie the sandy, highly calcareous soils of the Quindalup dunes (McArthur and Bettenay 1960). The youngest, most western Quindalup soils (Q) are lacking in profile, while at the eastern margins of the unit (Q1), organic matter can occur down to 0.5 m in the profile with cemented lime layers common at depth (McArthur and Bartle 1980). The Spearwood dunes are situated to the east of the Quindalup. Soils are podzolic and consist of siliceous sands (Karrakatta sand) over limestone which commonly outcrops in western localities (McArthur and Bartle 1980). Several Karrakatta soil units are distinguished by McArthur and Bartle (1980). They describe these as the grey (Kg) and the yellow (Ky) phases consisting of grey (Kg) or grey- brown (Ky) sand above yellow sand and limestone. The other main feature separating the soil types is the depth of limestone, being within 1 m for Ky and at greater but variable depth for Kg. Further, Ky soils are generally located at higher positions in the landscape. Shallow soils in elevated positions where surface limestone features prominently (Kls) are also distinguished by McArthur and Bartle (1980). An additional soil type known as Spearwood sand (or the Cottosloe association) is found on the western margin of the Spearwood system, this consisting of dark-brown siliceous sand with limestone at shallow depth (McArthur 1999).

From Perth to Busselton, stands of tuart also commonly occur on the Yoongarillup Plain. McArthur and Bartle (1980) describe this feature as an area of low relief, with limestone of marine and estuarine origin, overlain by siliceous sands. They note that tuart grows mostly on the Ys unit of the plain, this being characterized by fine sandy soils with limestone generally less than 1 m deep. Soil colour is described as grey or brown on the surface but vary in colour (grey, yellow, brown, reddish brown) below. On low ridges the limestone can be close to the surface or even exposed (Yls). At the

50 southern end of the tuart distribution (Capel to Ludlow), soils on the Yoongarillup Plain (Wonnerup association) comprise coarse siliceous sands with no podzolic characteristics in the profile (McArthur 1999).

To the north of Perth, the Herdsman association can be found in dune swales and is sandy to clayey in texture and often rich in organic matter (Seddon 1972). South of Perth, McArthur and Bartle (1980) describe similar features which they denote as Ksw.

Table 2.1: Selected chemical characteristics of soils (to between 10 and 25 cm depth) in which tuart is commonly found, adapted from McArthur (1999). Q1 = Quindalup phase 1, Ky and Kg = the yellow and grey phases of the Karrakatta sands, respectively and Y = Yongarillup plain. Quindalup Karrakatta Spearwood Karrakatta Wonnerup

(Q1) (Ky) sand (Kg) assoc. (Y) pH (CaCl2) 7.5 – 7.9 5.5 – 6.5 6.1 – 7.1 4.6 – 5.4 6.0 – 6.1 Org. C (%) 0.75 – 1.1 0.73 – 0.77 1.1 – 2.1 0.23 – 1.0 0.54 – 2.6 Total N (%) 0.08 – 0.11 0.03 0.06 – 0.12 0.01 – 0.04 0.03 – 0.15 Total P (mg kg-1) 280 - 290 44 – 49 140 - 190 12 - 22 130 - 180 Extract. P (mg kg-1) 4 - 9 < 2 4 - 6 < 2 6 - 10 Extract. K (mg kg-1) 12 - 27 12 - 17 21 - 43 10 - 20 13 - 34

2.6 VEGETATION

2.6.1 Formations

Tuart comprises a range of vegetation types ranging from open mallee-like formations to tall forest (Keighery et al. 2002). North of Yanchep, trees are generally stunted and grow within Banksia woodland and scrub heath (Beard 1990). Populations also tend to be smaller in size towards the northern limit of the tree’s range (Keighery et al. 2002). From Yanchep southwards, tuart can form tall open forests (Speck 1952 in Seddon 1972). Seddon (1972) qualifies this by stating that most examples of tall open forest including tuart are only marginally of this classification, and that it represents potentiality rather than reality, except at Ludlow where trees can grow to 40 m tall (Seddon 1972). Thus, there is considerable variety of form south of Yanchep, including significant areas which more accurately fit the description of woodland: trees 10 to 30 m tall with the projected cover of the tree canopy between 10 and 30 % (Specht 1970). Low woodland formations are more prevalent on positions exposed to the coastal winds or on ridge tops where soils are skeletal. On the Spearwood dunes, tuart often co-occurs with jarrah (E. marginata) and marri (Corymbia calophylla) (Seddon 1972). On Kg soil 51 tuart is not usually the dominant species, but becomes co-dominant towards the west in Ky soil (McArthur and Bartle 1980). However, pure stands of tuart do occur on Spearwood dunes, with good examples at Yanchep and around Mandurah (Seddon 1972). On the Yoongarillup plain, tuart is the dominant tree in forests and woodlands on Ys soil units as well as on terraces fringing lakes and inlets (McArthur and Bartle 1980) and in the Ludlow area.

2.6.2 Peppermint (Agonis flexuosa)

North of Mandurah, peppermint (Agonis flexuosa) is only a minor constituent of the midstorey in tuart forests and woodlands, whereas to the south it becomes the dominant element (Seddon 1972, Beard 1990). Within the southern extent of the tuart distribution, peppermint can attain heights of up to 15 m (Seddon 1972). Like tuart, peppermint can grow near the coast, but it extends further south than tuart, growing at Bremer Bay on the south coast (Rippey and Rowland 2004), and further inland, growing within parts of the southern karri (E. diversicolor) and jarrah forests (Boland et al. 1984). The leaves are narrow, dark green and aromatic, and they usually form a dense crown (Seddon 1972) (Figure 2.4). Flowering is from August to December (Rippey and Rowland 2004) and the seed is stored temporarily in small woody capsules that form clusters up to 1 cm in diameter (Boland et al. 1984) (Figure 2.4b). The species regenerates prolifically from both seed and by resprouting from a lignotuber (Bradshaw 2000). Individuals subject to frequent disturbance commonly assume a many-stemmed form (Figure 2.4c).

2.6.3 Other flora

Other species that commonly form a tree layer below tuart include Banksia attenuata, B. grandis, and Allocasuarina fraseriana, particularly on central and northern parts of the coastal plain (Seddon 1972). Xylomelum occidentale is sometimes present as a small tree on Kg soils south of Perth and Melaleuca tree species can be found amongst tuart fringing swamps and lakes (McArthur and Bartle 1980). In long-unburnt tuart woodlands, Acacia spp. can develop into small trees (Powell 1990).

52 b) b)

a) c)

Figure 2.4: a) The peppermint tree: Agonis flexuosa, b) leaves and seed capsules and c) a grove of multi-stemmed adults casting shade upon a dense litter bed.

The shrub and ground layers of undisturbed tuart communities can support a diversity of species (Seddon 1972). A total of 575 vascular plants were recorded in a survey of 12 sites across the north-south distribution of tuart (Keighery 2002). Macrozamia riedlei and Xanthorrhoea preissii are common on the Spearwood dune soils (McArthur and Bartle 1980, Seddon 1972). In the tall shrub layer, species of Acacia, Jacksonia and Hakea are often represented (Seddon 1972). Spyridium globulosum, Templetonia retusa and Diplolaena dampeieri can also be prominent tall shrubs on the western margins of the Spearwood dunes and on the Quindalup dunes, especially where shallow soils overlie limestone (McArthur and Bartle 1980). In the low shrub layer, species of Acacia, Melaleuca, Hibbertia, Petrophile, Synaphea, Grevillea, Dryandra, Calothamnus, Leucopogon, Hemiandra, Conospermum, Oxylobium, Daviesia and Bossiaea and Hovea frequently occur (Seddon 1972). Variations in soil type, soil depth and degree of exposure across quite small distances result in changes in understorey composition and structure (Seddon 1972, Rippey and Rowland 2004). In addition, the composition and density of the understorey in many areas has been significantly influenced by the history of human disturbance (Seddon 1972, Keighery 2002).

53 2.7 HISTORY OF ANTHROPOGENIC DISTURBANCE WITHIN THE DISTRIBUTION OF TUART

2.7.1 Clearing and fragmentation

The greatest disturbance to tuart forests and woodlands since European settlement has been from clearing and the resultant fragmentation. Tuart was estimated to occur across 111 600 ha at the time of European settlement in 1829, whereas today only some 38 800 ha remain (Department of Conservation and Land Management et al. 2003). Within the major urban areas of Perth, Mandurah and Bunbury, tuart woodlands exist in small remnants surrounded by houses and industry.

2.7.2 Grazing

There has been a long history of cattle grazing in the tuart belt (Ward 2000). Grazing in tuart woodland today is mostly confined to pasture established in partially-cleared “parkland” landscapes on the Spearwood and Yoongarillup landforms (McArthur and Bartle 1980). Keighery et al. (2002) notes that very little of the understorey in remaining woodlands has not been disturbed by grazing. For example, grazing continued up until the 1960s and 1970s on land that is now Yalgorup National Park (Bradshaw 2000, Ward 2000). In the Ludlow State Forest and National Park, grazing over most of the area occurred from the 1830s and was continuing in the 1970s (Forests Department 1979). In addition to trampling and eating the understorey vegetation, cattle have also been observed eating tuart seedlings (Forests Department of Western Australia 1979).

2.7.3 Timber cutting

Logging of trees, including tuarts, has been widespread on the coastal plain. In the Ludlow tuart forest, logging began in the 1840’s and continued up until 1975 (Forests Department of Western Australia 1979). During the 1900s at Ludlow, non-commercial trees were removed in silvicultural treatments (L. McCaw pers. comm.). At the locality of Myalup within the Yalgorup area, a sawmill operated between 1953 and 1964 (Department of Conservation and Land Management 1995). While jarrah appears to have been targeted by loggers in the forest blocks around this mill, there is also evidence that tuart trees were cut (pers. obs.). As a consequence of loggers cutting jarrah in preference to other species, the tree species mix has been altered here and elsewhere on the Spearwood dunes (McArthur and Bartle 1980). Additionally, stand 54 structural modifications likely arose from the tallest trees having been cut soon after European settlement (Seddon 1972).

2.7.4 Mining and quarrying

Some disturbance from mining and quarrying has occurred in tuart woodlands. According to McArthur and Bartle (1980), evidence of quarrying was widespread, but relatively limited in area across the extent of Spearwood dunes in the Mandurah to Bunbury region at the time of their study. Bennett (1988) also mentions limestone quarrying as having occurred within Kings Park in the early years of settlement. At Ludlow in 2006, mineral sands were being mined within State Forest that contained tuart.

2.7.5 Fire

Anthropological and historical evidence suggests that much of the Swan Coastal Plain was regularly burnt by the Aborigines until the middle of the 19th century (Hallam 1975, Ward 2000, Abbott 2003). The analysis of colour banding on Xanthorrhoea preissii stems indicate an average fire frequency of two to four years within the Yalgorup tuart woodlands prior to European settlement (Ward 2000). Fires at 2 to 4 year frequencies would have been relatively low in intensity (and severity) given the fuel accumulation– fire intensity relationship (Ward 2000, Abbott 2003). Historical accounts also reveal that fires on the coastal plain and adjacent forests were lit by Aborigines in summer (December to March) and burnt at various scales of extent depending on seasonal conditions and weather (Abbott 2003). Thus a fine-scale complex mosaic of fuel ages likely developed (Ward 2000, Abbott 2003).

While the Xanthorrhoea stem fire-dating technique is in dispute (section 1.4), the debate is currently focused on data from a different species (X. acanthostachya) located in a drier environment (Enright et al. 2005, Gill 2006). Thus, any apparent deficiencies that may have been demonstrated from this latter work are of questionable relevance to the fire history interpretations of X. preissii in tuart and E. marginata forest and woodland (D Ward, pers comm.). Furthermore, historical documents and research on fire behaviour in the region lend strong support to Ward’s (2000) findings (Abbott 2003).

The practice of frequent burning was continued by European graziers across much of the coastal and near-coastal areas of the southwest into the first half of the 20th century 55 (Abbott 2003). Depending on the location, the average fire frequency was thought to vary between two and ten years, with most fires lit in autumn just after the first rains (Abbott 2003). In the Yalgorup area, this practice ceased in some areas as State Forest was declared (around the 1930s) and grazing was subsequently phased out (R Underwood pers. comm.). The region-wide policy at the time of fire-excluded compartments surrounded by regularly-burnt buffer strips (Shea et al. 1981) likely shaped the fire regimes of these State Forest areas. In much of the adjacent area that is now National Park or private property, frequent burning by graziers (average frequency between two and four years) continued until the middle of last century (Ward 2000). At Ludlow, a policy of fire exclusion was implemented in 1921 (Forests Department of Western Australia 1979). Information on the fire regimes for the area of tuart woodland north of Perth for the first half of last century is not readily available, but regimes were probably highly variable given the increasingly fragmented nature of the stands that developed with urban expansion. The management of metropolitan remnants varied according to the authority controlling the area. For example, regular prescribed burning was carried out in Kings Park from the 1930s until the mid-1980s by the management board (Dixon et al. 1995), while at Woodman Point, just south of Perth, the land was owned by the Commonwealth from 1958 until the late 1970s and no prescribed burning was conducted during this period (Powell and Emberson 1981).

Since the middle of the last century, fire regimes within tuart woodlands have varied considerably depending on location, land tenure and period. With the institution of the policy of widespread prescribed burning in the 1950s, areas under Forests Department control were subject to more frequent burning (Shea et al. 1981). Fire exclusion, with the exception of some small scale regeneration burning, continued to be the policy at Ludlow, with grazing relied upon to reduce fuel loads (Forests Department of Western Australia 1979). For Yalgorup, departmental records suggest burning was conducted in the State Forest as frequently as three to six years in the 1960s and 1970s, but less frequently from the 1980s: an eight to ten year minimum interval. Although a number of extensive wildfires occurred in various parts of the adjacent National Park during the first half of the 1970s (Smith 1975), fires have been less frequent in recent decades, with the time since fire in some parts between 20 and 80 years (Ward 2000). Frequent wildfires have occurred in tuart woodlands in and adjoining the Perth metropolitan area such as at Kings Park (Dixon et al. 1995), Star Swamp (Pigott and Loneragen 1995) and Rockingham Lakes (G. Napier Department of Conservation and Land Management,

56 pers. comm.). Since the 1990s, extensive wildfires have also occurred in tuart woodlands north of Perth in Yanchep National Park in 1983 (Department of Conservation and Land Management 1988), 1991 (Department of Conservation and Land Management fire records) and in 2005 (pers obs.). Parts of these northern woodlands are also subject to controlled burning with a minimum interval of 8 to 12 years between fires (L. Sage, Department of Environment and Conservation, pers. comm.).

2.8 UNDERSTOREY CHANGE

2.8.1 The pre-settlement character of tuart forest and woodland

A savannah-like formation consisting of widely-spaced tuarts and a low grassy understorey maintained by frequent Aboriginal firing, is thought to have been common at the time of European settlement at Ludlow (Forests Department of Western Australia 1979, Figure 2.5) and further north at Yalgorup (Ward 2000, Bradshaw 2000). The records of Lieut. H.W. Bunbury’s travels reported in Bunbury and Morrell (1930) support this view (Bradshaw 2000). They represent a reliable source of information about the nature of tuart woodlands landscape between Mandurah and Ludlow at the time of European settlement given they were made soon after settlement (1836 to 1837), after the travel route (passing through Yalgorup, Ludlow and areas between) had been traversed four times and by one who was critical of others who had overstated the suitability of land for grazing to encourage settlement (Bunbury and Morrell 1930). The following recollections of Bunbury in Bunbury and Morrell (1930) describe landscapes that support the view of Ward (2000) and Bradshaw (2000): “..open flat country lightly timbered with Tooarts, with abundance of grass and not many bushes..” (p73-74 referring to just North of the Leschenault Inlet), and “..undulating Tooart country of considerable extent, with plenty of grass and good though light soil..” (p90, referring to a location between Bunbury and Ludlow). Bunbury also notes the connection between the open nature of the landscape and frequent Aboriginal burning: “..by these fires…the country is kept comparatively free from underwood and other obstructions, having the character of an open forest through most parts of which one can ride freely; otherwise in all probability, it would soon become impenetrably thick…” (p105-106).

57

Figure 2.5: A historical scene of a tuart formation at Ludlow (date unknown). Reproduced from Powell (1978).

While tuart no doubt occurred as open formations, it should not be assumed that all landscapes with tuart were similarly so. Benson and Redpath (1997) illustrate how selective quoting of historical accounts has been used to give an over-simplified view of historical vegetation structure in eastern Australia. When other quotes of Bunbury in Bunbury and Morrell (1930) are added to those mentioned above, a more complex portrait emerges. For example, north of the Leschenault Inlet, he also describes passing through stands of tuart with marri and jarrah, but notes the understorey was “..thick with the Furze tree or Stinkwood and other bushes.” (p73). Overall, his reflections upon the route to the inlet do not indicate an “open” vegetation structure: “I reckoned that we had gone over about 25 miles today, much of it through a very scrubby country which is particularly unpleasant to walk through..” (p75). Further, the above quote of “..undulating Tooart country..” continues with: “..but here the thickets of broom and furze and occasional patches of swamp were so thick as to impede my progress very much, rendering it difficult to keep anything like a straight course” (p90). Thus, the mosaic of vegetation structural types thought to arise from the intensive burning activity of Aborigines (Abbott 2003, Bowman 2003), inclusive of the open savannah-like formations, may best describe the pre-European landscape of the coastal plain. The

58 relative proportion of understorey extremes (open versus closed) may have been very dynamic, shifting with variations in rainfall patterns between years and Aboriginal population sizes over decades and centuries. At the time of European settlement, the size of the Aboriginal population on the coastal plain, and thus the intensity of Aboriginal burning activity and the impact this had on the vegetation, was thought to have been at a maximum (Hassell and Dodson 2003).

The existence of dense formations within a landscape subject to frequent burning could be accounted for by several explanations. Firstly, soil factors have an important influence on the structure and composition of understorey vegetation, regardless of fire regime (Callaway and Davis 1993, Florence 1996, Abbott 2003, Burrows and Wardell- Johnson 2003). Secondly, the presence of long-unburnt patches and recently-burnt patches may have co-existed within a landscape subject to frequent fire (Abbott 2003, Bowman 2003). Thirdly, where areas were burnt more severely due to being longer unburnt or burnt under extremely hot and dry conditions, ashbeds likely supported dense woody post-fire regrowth (Bradshaw 2000, Abbott 2003). Shrub species such as Jacksonia furcellata (probably the “stinkwood” referred to by Bunbury as being dense in some areas) and Acacia pulchella regenerate abundantly after fire from soil-stored seed (Powell 1990, Bennett 1988).

2.8.2 Understorey changes since settlement

The understorey is claimed to have become increasingly dense within tuart woodland in the Yalgorup area during the past century (Ward 2000, Bradshaw 2000). The reduction in fire frequency is believed to have been an influential factor in this change (Ward 2000, Bradshaw 2000). Increasingly dense understorey growth has also been observed in tuart woodlands in Kings Park, but attributed to periodic intense fire reducing the competitive influence of trees via repeated canopy damage (Beard 1967).

Ward (2000) and Bradshaw (2000) have suggested that the height and density of the peppermint understorey has increased in the tuart woodlands south of Mandurah. In his diary, Bunbury expresses a particular appreciation for the tree: “..I think it is about the most picturesque tree in the Colony” p73, and makes mention of large attractive examples in two instances (Bunbury and Morrell 1930). He also notes that “...The tree is common near the lagoons and near the coast all the way down the bight of Geographe Bay..” (p73). However, he makes no comment that tree was growing densely under 59 tuart. At Ludlow, the potential for peppermints to form a vigorous understorey and midstorey frustrated forest managers as far back as the beginning of the 20th century (Gardner 1923 and Kessell 1923 in Bradshaw 2000). In contrast to tuart, peppermint is an apparently tolerant species, able to regenerate from seed between fires and develop as a sapling under an established canopy (Bradshaw 2000). Once it has formed a lignotuber, peppermint, like tuart, seems able to persist in the presence of frequent fire (Bradshaw 2000). The apparent ability to rapidly spread and dominate has led some to label the species as a native weed (Bradshaw 2000, Ward 2000). Such weed-like behaviour has been observed for some other Australian natives in their local environment (Burrell 1981, Gleadow and Ashton 1981).

Around 60 % of the current tuart woodland area was recently rated as having a “disturbed” understorey (Department of Conservation and Land Management et al. 2003). The fragmented nature of remaining tuart woodlands, their proximity to urban areas and the disturbances outlined above, have together resulted in significant weed invasion (Keighery 2002). The understorey in tuart stands of the Perth area generally comprises a greater cover of weeds than elsewhere (Keighery 2002). In tuart woodlands at Yalgorup National Park, Fox et al. (1980) recorded 29 introduced species from a total of 210. In the recent survey of sites across the extent of tuart, 28 % of the vascular plants recorded were weeds (Keighery 2002). Any pre-existing native “grass” component of the understorey promoted by regular burning (Bunbury and Morrell 1930, Hallam 1975, Bradshaw 2000, Ward 2000, Abbott 2003) but not evident today, is disputed by Keighery and Keighery (2002). They argue that a number of recent and historical botanical studies do not support such claims.

2.9 TUART DECLINE

The health of tuart trees has deteriorated in the metropolitan area over some decades (Beard 1967, Seddon 1972, Fox and Curry 1981, Pigott and Loneragan 1995) and in Yalgorup since the early 1990s (Bradshaw 2000, Longman and Keighery 2002, Edwards 2004). The decline is characterized by dead branches extending from dead trunks or sparse crowns (Figure 2.6). The most severe decline in terms of progression and extent is in the Yalgorup area (Edwards 2004). Concerns have also been expressed about the lack of tuart seedlings in urban remnants (Fox and Curry 1980), at Ludlow (Keene and Cracknell 1972) and at Yalgorup (Backshall 1983, Bradshaw 2000). The co- 60 occurrence of declining adult trees and poor seedling regeneration, suggests that permanent loss of some tuart woodlands may occur in the absence of human intervention.

a) b)

Figure 2.6: Tuart decline in Yalgorup National Park: a) severe decline near Martins Tank in 2003 and b) moderate decline at the northern end of the Park near White Hill Road.

As with other declines of apparent complex etiology (Old 2000), there are seemingly a range of interacting factors underlying tuart decline (Longman and Keighery 2002). The proximity of tuart to urban areas raises the question as to whether air pollution may be major decline agent. While the study of Chilcott (1992) found an association between poor tuart health and air pollution from the Kwinana Industrial area, the spatial pattern of tree health suggests that air pollution is not linked to the decline of trees outside the Perth metropolitan area (Chilcott 1992, Longman and Keighery 2002, Edwards 2004). Other possible decline agents include fire, drought, fragmentation, fungal pathogens, frost, hydrology, nutrient changes in soil and groundwater, and changes in insect populations and beneficial fungal associations (Longman and Keighery 2002). The most significant factor observed in association with tuart crown decline thus far has been the presence of wood boring insects (particularly Phorocantha impavida and P. semipunctata), but these are regarded as a secondary rather than primary agents of decline (Day 1959a,b, Fox and Curry 1980, Longman and Keighery 2002). The step change in annual rainfall across the region during the 20th century (Indian Ocean Climate Initiative 2002) could be a significant predisposing and contributing factor to tuart decline (Longman and Keighery 2002). Severe droughts in the 1960s were believed to have triggered extensive declines and increased borer damage in eastern Australian eucalypts (Pook and Costin 1966, Pook and Forrester 1984, Old 2000). 61 However, a satisfactory explanation with supporting evidence for tuart decline remains to be presented.

2.9.1 Tuart tree decline and fire

Successive wildfire damage has been suggested as a contributing factor to poor tuart health in the Perth area (Beard 1967, Fox and Curry 1980, Powell and Emberson 1981, Pigott and Loneragan 1995). Excluding the direct damage caused by wildfire and the subsequent damage from borers and pathogens entering the trees via fire scars (Fox and Curry 1981), Beard (1967) argues that a succession of intense fires may have increased the competitive influence of the understorey on overstorey species such as tuart in Kings Park. In was hypothesized that the canopy damage increased light and sub- surface soil moisture availability for the understorey and repeated fire damage diminished the capacity of tree crown recovery. Further, competition from the more vigorous understorey may also have limited both tree recovery and tree seedling regeneration, thereby reinforcing the change. Healthy tuart stands at Woodman Point just south of Perth have been largely spared the periodic wildfire that has occurred in other Perth remnants (Powell and Emberson 1981). Today, the canopy at Woodman Point is still complete and healthy (pers. obs.). Despite the claims linking intense fire with poor tuart health, there have been few published data upon which to evaluate this.

Competition-mediated decline of tuart arising from a dense understorey developed under a regime of less frequent fire (Bradshaw 2000, Ward 2000) is unproven, as is the case for eucalypt ecosystems generally (Chapter 1). In addition to resource competition, Bradshaw (2000) suggests that possible allelopathic effects of peppermint could be a factor in tuart decline. However, it may now be difficult to determine whether the understorey change was a cause of, or a reaction to, the reduction in tuart canopy cover.

Tree health at Yalgorup has possibly been influenced by changes in the chemical and microbiological characteristics of the soil in the long absence of fire (section 1.5.3 and 1.5.4). Jurskis (2005) has proposed that his model associating a long absence of fire with tree decline in eastern Australia (Jurskis and Turner 2002) may be applicable to tuart decline at Yalgorup. However, no studies in tuart woodlands have examined this. In addition, there are no studies that have assessed whether a tree vigour response to fire occurs in tuart woodlands, as has been detected in other eucalypt forests (section 1.6.2).

62 2.9.2 Fire regimes and the lack of tuart regeneration

Where canopy seed stores are not severely limited by the tuart bud weevil (Haplonyx tibialis) (Fox and Curry 1980) or crown decline, the presence or absence of fire is the dominant influence on natural tuart seedling regeneration success (Ruthrof 2001). From observation and forestry trial work at Ludlow, natural tuart seedling regeneration appears dependent on ashbeds (Forests Department of Western Australia 1979, Keene and Cracknell 1972). Competition from introduced grasses was seen as the main factor limiting regeneration off ashbed areas (Forests Department of Western Australia 1979). In contrast to tuart, the apparently increasingly-dominant peppermint presumably also establishes between fires (Bradshaw 2000). Thus the reduction in fire frequency may account for any imbalance between tuart and peppermint regeneration at Yalgorup (Bradshaw 2000, Ward 2000). A study of the comparative mode of establishment for the two species would confirm this.

Little information has been documented on the ability of tuart and peppermint seedlings and saplings to survive fire. Tuart seedlings were thought to be “..extremely fire resistant..” as a result of observations of one wildfire event at Ludlow (Forests Department of Western Australia 1979, P9). It was reported that seedlings less than 2 m resprouted, while those taller seedlings that were only partially defoliated “..continued to survive and flourish.” (Forests Department of Western Australia 1979, p9). The size/age at which peppermints become fire resistant is unknown but presumably it is related largely to the development of the lignotuber (Bradshaw 2000). It has been observed that frequent fire (every two to three years) maintains a peppermint savannah in some southern coastal areas by eliminating most peppermint seedling regeneration (T Muir, pers. comm.). Investigation of the capacity of tuart and peppermint for early fire resistance and post-fire recovery is required to fully understand population dynamics of these species under alternative fire regimes.

There are concerns that an increase in the density of peppermint seedlings could also be impacting on tuart seedling regeneration through competition (Keene and Cracknell 1972, Forests Department of Western Australia 1979, Backshall 1983, Bradshaw 2000). Growth-rate comparisons of peppermint and tuart that developed from seed after a wildfire at Yalgorup National Park revealed that most peppermints were taller than tuart after six years, but that the tallest individuals measured (7 to 8 m) at the site were in fact

63 tuarts (L. McCaw, pers comm.). However, 60 % of the tuarts were suppressed, being less than 2 m tall (L. McCaw, pers comm.). This suggests that tuart may be the superior competitor under the most favourable conditions of light, nutrients and moisture whereas peppermint is a stress tolerator and therefore a more successful competitor under resource-limiting conditions. This proposition is compatible with observations that tuart, like other eucalypts, is an intolerant (gap regenerator and C-type strategist [Grime 1977]) species (Forests Department of Western Australia 1979), and peppermint a tolerant (S-type) species (Bradshaw 2000). Morphological and growth comparison experiments would allow a more definite rating of the relative competitive ability of peppermint and tuart.

An established dense midstorey canopy could also affect tuart seedling establishment by competition. Under a regime of low-intensity fire, peppermint trees may be left unscorched or rapidly refoliate their crowns. Thus, tuart seedling establishment may be negatively affected by insufficient light and possibly a lack of available moisture and nutrients. The influence of the dense peppermint canopy on patterns of phytophagy and foliar disease on seedlings below the canopy may also affect tuart seedling establishment indirectly. There are no published studies addressing these issues in tuart woodland.

2.10 THESIS OBJECTIVES AND STRUCTURE

The aim of this thesis was to further understand the impact of fire and fire regimes on tuart health and regeneration. Most of the knowledge gaps highlighted in sections 2.9.1 and 2.9.2 were addressed in the experimental chapters (Figure 2.7). As discussed in Chapter 1, the potential interactions between fire and other factors are numerous and potentially complex. Therefore, it was not possible to cover all of the important aspects relevant to fire and the persistence of tuart woodlands; for example, soil microbiological studies were not included. Additionally, as a broad ecological study, the investigations were often focused on pattern, with inferences drawn on process after reference to the literature. The first experimental Chapter (Chapter 3) investigated tuart tree health, population structure of tuart and peppermint, and understorey change. The broad hypothesis of this Chapter was that patterns of tuart health and regeneration are associated with fire history and peppermint density. This was tested by way of field surveys across a range of sites with contrasting fire histories, principally in the Yalgorup area where tuart decline is most severe and widespread. Data from historical surveys 64 was also drawn upon to examine changes in time. In Chapter 4, tuart health and survival was followed over time: from before or immediately following fire, until 9 to 12 months after. In addition, post-fire recovery was compared between tuart, peppermint and other co-occurring tree species. These studies were central to testing the hypothesis that individual fires have a long-term impact on tuart health. The final two experimental Chapters in the thesis focused on seedling establishment, and in particular, the effect of fire and overstorey competition on seedling establishment. Chapter 5 tested the hypothesis that differences exist between growth and survival rates for tuart and peppermint seedlings at the establishment phase. Several field studies at burnt and unburnt sites were complemented by a number of glasshouse studies in order to address this hypothesis. The final experimental Chapter (Chapter 6) was based on a field experiment where tuart seedlings were grown in ashbed and unburnt soils under contrasting densities of peppermint canopy cover. This allowed the hypothesis that peppermint trees inhibit tuart seedling growth and survival to be tested. In the concluding Chapter, the main results of the thesis were summarized, management recommendations were outlined and further avenues for research were suggested.

65 Chapter 1: General Introduction: Fire and the persistence of eucalypt populations in southern Australia • Fires and fire regimes can affect eucalypt health and survival in numerous ways. • There is insufficient data to conclude an absence of fire is responsible for tree decline across large areas.

Chapter 2: Tuart and the tuart ecosystem from the 19th to 21st century • The ecosystem has a long history of anthropogenic disturbance. • It has been proposed that the peppermint midstorey has become denser. • Several interacting factors could underlie tuart decline. • The role of fire regimes in tuart decline, including via their influence on the density of the peppermint midstorey, is unclear.

Chapter 3: Relationships between recent fire history and the structural patterns of tuart and tuart-peppermint stands

Hypothesis: Patterns of tuart health and regeneration are associated with fire history and peppermint density.

Chapter 4: Survival and recovery patterns for tuart and co-occurring tree species following exposure to fire

Hypothesis: An individual fire affects tuart health beyond the first growth season after the fire.

Chapter 5: A comparison of tuart and peppermint seedling establishment and early growth

Hypothesis: There are differences in growth and survival rates between tuart and peppermint at the seedling establishment stage.

Chapter 6: The effect of peppermint trees on tuart seedling growth and survival

Hypothesis: The density of peppermint trees inhibits tuart seedling growth and survival.

Chapter 7: Synthesis Summarizes the main findings of the thesis, provides management recommendations and suggests avenues for further research.

Figure 2.7: Thesis structure by Chapter including main findings of Chapters 1 and 2, main hypotheses for each experimental Chapter and the contents of the final Chapter.

66 3 Chapter 3. Relationships between recent fire history and the structural patterns of tuart and tuart-peppermint stands

The following publication was based on part of this Chapter:

Archibald, R. D., B. J. Bowen, G. E. StJ. Hardy, J. E. D. Fox, and D. J. Ward. 2005. Changes to tuart woodland in Yalgorup National Park over four decades. in M. Calver, H. Bigler-Cole, G. Bolton, J. Dargavel, A. Gaynor, P. Horwitz, J. Mills, and G. Wardell-Johnson, editors. Proceedings 6th National Conference of the Australian Forest History Society Inc., September 2004, Augusta, Western Australia. Millpress, Rotterdam.

Note: B.J. Bowen and G.E.StJ. Hardy were supervisors of the PhD project, J.E.D. Fox supplied the historical data and D.J. Ward was a contributer to the original survey work.

67 3.1 INTRODUCTION

The potential for fire to shape the structure and composition of eucalypt forest and woodlands was outlined in Chapter 1. The many direct and indirect impacts of fire events and regimes on seed, regenerating seedlings and established trees were discussed. The importance of event and regime variables in the sign and magnitude of impacts was emphasized. In the final part of the chapter, the theoretical role of extreme fire regimes (frequent intense fire or the long absence of fire) in eucalypt decline was explored.

Chapter 2 highlighted possible roles for extreme fire regimes in the decline of tuart (Eucalyptus gomphocephala). The reduction in fire frequency within Yalgorup National Park has been linked with an increasingly dense understorey of peppermint (Agonis flexuosa) and declining tuart health (Ward 2000, Bradshaw 2000, Longman and Keighery 2002, Jurskis 2005), while in remnants within the Perth area, recurring intense fire has been implicated as a cause of poor tuart tree health (Beard 1967, Fox and Curry 1980, Pigott and Loneragen 1995). Increased competition from understorey development has been proposed as a tree decline factor in both scenarios (Bradshaw 2000, Beard 1967). However, any association between fire regimes, understorey change and declining tree health is largely anecdotal.

In addition to declining tree health, an apparent lack of tuart seedling regeneration has compounded concerns for the future of tuart woodlands (Fox and Curry 1980, Backshall 1983, Bradshaw 2000). For natural tuart seedling establishment, a fire event appears to be a necessary pre-condition (Keene and Cracknell 1972, Forests Department of Western Australia 1979, Bradshaw 2000, Ruthrof 2001), but a comprehensive survey of regeneration patterns across sites with different fire histories has yet to be carried out. In contrast to tuart, seedling establishment for peppermint is suspected to occur at unburnt sites (Forests Department of Western Australia 1979, Bradshaw 2000), but again, data are lacking to demonstrate or quantify the regeneration patterns of this species in relation to fire.

The principal hypothesis of this Chapter is that patterns of tuart health and regeneration can be associated with fire history and peppermint density. For regeneration patterns, surveys of seedling abundance and stand structure were conducted across sites with

68 contrasting times since fire. Tuart tree health was also assessed at these sites and associations with fire history were investigated. In addition, the possible indirect effect of fire regimes on tuart health via an alteration of vegetation structure and composition was addressed by testing for an association between tuart health and peppermint density, and by examining differences in the vegetation between current and historical vegetation data for the Yalgorup area. Hypotheses relating to specific components of the experimental work are presented in the following section.

3.2 METHODS

3.2.1 Sites and site selection

Site descriptions are outlined in Table 3.1 and general site locations follow in Figures 3.1 and 3.2. These sites were used for a range of studies in this Chapter and elsewhere in this thesis. Permanent plots were established at a number of sites (Y1, GB, Ya1 and Ya2) where fires occurred during the study period. Where possible, plots were also set up in areas that served as unburnt comparisons; sites Y2 and Y3 were unburnt comparisons for site Y1; site Ya2 was intended as a control but the area was burnt in 2005; for site GB, there were no comparable unburnt sites nearby. Six sites (Y4 to Y9) in the Yalgorup area that had not been burnt for at least four years were also selected on the basis of initial knowledge of fire history, management history, vegetation and soil type. The aim was to sample a set of sites with a maximum contrast in fire history (from the available range) yet with minimal difference in the other factors, particularly soil type. For the survey outlined in Section 3.2.5, pre-existing plots at sites Y2, Y10 and Y11 were used.

69 Table 3.1: Details of the study sites located at Yalgorup (Y), Golden Bay (GB) and Yanchep (Ya). Tuart trees were present at all sites. Vegetation Fire history6. Site Overstorey Under- Start of ID Soil units1. Structure2. Other trees3. storey4. Recorded fires5. Record Regime description Disturbance history 7. Tenure 8. Y1 Yls Open woodland Af, Alf open 72, 91a, 04 72 Periodic & variable intensity Grazing, quarrying NP Y2 Yls, Ys Open woodland Af, Alf open 72, 76, 85, 93a 72 Periods of frequent wildfire Grazing, quarrying NP Y3 Ys, Yls Open woodland Af, Alf, Em, B open 72 72 Infrequent unknown NP Y4 Ky, Kg Tall open woodland Af, Em, B open 72, 85, 92, 00a 67 Decadal controlled Grazing, logging SF Y5 Ky Tall open woodland Af, Em, B open 72, 85, 92 67 Decadal controlled Grazing, logging SF Y6 Ky, Kls Open woodland Af, Em open 94s 70 Infrequent Grazing NP Y7 Ky, Kls, Ys Open woodland Af, B closed 67 67 Infrequent Grazing NP Y8 Ys, Ky Tall open woodland Af, Em, Cc open 64s, 70 64 Decadal controlled Grazing, logging SF Y9 Ys, Yb Tall open woodland Af, Em open 63s, 66s, 70, 89, 95s 637. Decadal controlled Grazing, logging NR Y10 Ys Open woodland Af, B variable none 65 Very infrequent unknown NP Y11 Ky, Kls Open woodland Af, Em, B variable 72 (part) 60 Infrequent Grazing NP GB Ky, Ksw Open woodland Em, Alf, B variable 95, 03 95 Periodic wildfire unknown RP Ya1 S Tall open woodland B variable 91, 04s 90 Periodic & variable intensity unknown NP Ya2 S Tall open woodland B open 91, 05 90 Periodic & variable intensity unknown NP Ya3 Kg Open woodland unknown open 99w, 05 90 Periodic & variable intensity unknown NP Ya4 Kls Tall open forest none open 05 90 Infrequent unknown NP 1. From McArthur and Bartle (1980). See Chapter 2, Section 2.5 for descriptions. Yb soils are described by these authors as located in depressions in the Yoongarillup Plain and consisting of a shallow layer of dark sandy loam above calcareous marl and shell beds. Predominant type at site mentioned first. 2. Based on Specht 1970. In many cases these ratings are below the potential projected crown cover due to poor crown health. Further, accurate classification is complicated by the midstorey becoming dominant in its contribution to total crown cover at some sites. This fact was not considered in the classifications. Due to within-site variability the classifications represent an estimated average. 3. Agonis flexuosa (Af), Allocasuarina fraseriana (Alf), Banksia spp. (B), Eucalyptus marginata (Em), Corymbia calophylla (Cc).

70 4. Based on the difficulty in walking through the understorey: easy (open) and difficult (closed). 5. Refers to the majority of the site as at July 2005. Wildfires/summer fires denoted in bold type. For other fires, season is denoted where known (a = autumn, s = spring, w = winter). Data sourced from a combination of Department of Conservation and Land Management (DCLM) Fire Management Branch records (Where a two-year figure existed [for example 1976/1977], the latter year is listed), Smith (1975), S. Dutton, J. Wheeler and G. Napier (Department of Conservation and Land Management) pers. comm., G. Behn (CSIRO) pers. comm., F. Masao and W. Nairn pers. comm. 6. From field observation. Excludes fires. 7. The record between 1970 and 1989 was incomplete. 8. National Park (NP), State Forest (SF), Nature Reserve (NR), Regional Park (RP).

71

Figure 3.1: Location of study sites at Yalgorup. Key: site location and identifier (▲Y1), water feature () and roads (). See Figure 2.2 (Chapter 2) for the location of Yalgorup at a larger scale.

72

a)

b) Figure 3.2: Location of study sites at a) Golden Bay and b) Yanchep. Key: site location and identifier (▲GB), water feature () and roads (). See Figure 2.2 (Chapter 2) for the location of Golden Bay and Yanchep at a larger scale.

73 3.2.2 Weather conditions during the study period

Rainfall and temperature data at Mandurah for the study period (January 2003 to May 2006) are shown in Figure 3.3. The broad trends in these data are also applicable to the more distant study sites at Yanchep. There were notable deviations from the long-term averages in rainfall and temperature over the study period. Rainfall totals in early autumn (March to May) 2003, late autumn 2005 and for January 2006, were considerably higher than for the long-term average or the other years of the study period. However, excluding the total for June 2005, winter (June to August) rainfall was well below the long-term average in all years. This decline in winter rainfall totals is consistent with the trend that has occurred across the southwest over past decades (Indian Ocean Climate Initiative 2002). The most prominent feature of the temperature data were the lower than average maximum temperatures throughout spring (September to November) and summer (December to January) of 2005 (Bureau of Meteorology 2006).

300

250

200

150

100 Mean rainfall (mm) 50

0 Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Month a)

30

27

24

21

Mean max. temp.(°C) 18

15 Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Month b)

Figure 3.3: a) Monthly rainfall (mm) and b) mean maximum daily temperature (°C) for Mandurah (Mandurah Station) for 2003 (□), 2004 (x), 2005 (■), January to May 2006 (■) and averages for Mandurah Park Station (latitude: -32.5031 S, longitude: 115.7664 E) from 1889 to 2001 (---). Data sourced from the Bureau of Meteorology (2006). 74 3.2.3 Tuart and peppermint seedling regeneration at unburnt and recently burnt areas

Hypothesis:

• The abundance of tuart and peppermint seedlings differs between recently burnt (< 1 year since fire) and unburnt areas (> 4 years since fire).

Sites All sites in Table 3.1 were surveyed with the exception of sites Y10, Y11, Ya3 and Ya4. Descriptions of the last fire at the recently burnt sites are presented in Table 3.2.

Table 3.2: Fire and burn description for sites burnt within one year prior to seedling regeneration survey. Ave. est. intensity Max. scorch Burnt area Site Date (kW m-1)1. ht (m)2. (%) 3. Comments Y1 12/2004 < 350 10 60 Prescribed fire Ya1 10 - 11/2004 < 350 12 50 Prescribed fire. Lit several times Ya2 01/2005 < 350 6.5 90 Back burn at night for wildfire control4. GB 02/2003 350 - 1000 < 15 95 Back burn for wildfire control5. 1. Estimated according to scorch height or observation in the case of Y1 and Ya2 and with reference to Burrows (1984) and Cheney (1981). 2. Measured within plots at Y1, Ya1 and Ya2. Estimated at GB. 3. Based on visual estimations. At sites Ya1, Ya2 and Y1, multiple estimates were made across plots to derive a mean. 4. J. Wheeler, Department of Conservation and Land Management, pers. comm. 5. G. Napier, Department of Conservation and Land Management, pers. comm.

Sampling design

Permanent plots at sites Y1, Y2, Y3, Ya1, Ya2 and GB Distances between the plots were approximately equivalent to plot size at each site. At sites Y1, Y2, Y3, Ya1 and Ya2, sub-plots nested within the main plot were used to sample seedlings (Figure 3.4). At site GB, there were four 25 x 25 m plots established and seedlings were counted within the entire plot areas. Plot sizes, numbers and sampling dates are summarized in Table 3.3. Variation in plot numbers and plot sizes arose as a result of pooling data from several separate studies that were set-up at various times. A much larger number of plots (43) were set up at site Y7 due to the mistaken belief at the time of the survey that this was two separate sites. The coordinates of the main plots at each site are listed in Appendix 5. 75 50 m 40 m

4 m 4 m 15 m

25 m 40 m m 50

5 m 5 m 25 m

a) b) Figure 3.4: Plot design for a) sites Y1, Y2 and Y3 and b) sites Ya1 and Ya2. Seedling survey plots are shaded. At sites Y1, Y2 and Y3, tuarts and peppermints ≥ 15 cm dbhob (diameter at breast height over bark) were measured within the 50 x 50 m area, tuarts < 15 cm dbhob were measured within the 40 x 40 m area and peppermints < 15 cm dbhob were measured within the 15 x 4 m areas. At Ya1 and Ya2, all trees and seedlings were measured in the 25 x 25 m.

Table 3.3: Survey dates and plot details for vegetation surveys at recently burnt sites (bold type) and sites unburnt for at least 4 years (normal type). n/a = plot size not applicable due to plotless sampling or species not present at site. Survey Dates Tuart: plot size & number1. Peppermint: plot size & number1. Seedlings Saplings Seedlings Saplings & trees Seedlings Saplings & trees Site & trees m2 No. m2 No. m2 No. m2 No. Y1 01/06 04-05/04 16 222. 25002. 6 0.79 222. 603. 24 GB 02/04 02-03/03 625 4 625 4 n/a n/a n/a n/a Ya1 01/06 03-04 25 12 625 4 n/a n/a n/a n/a Ya2 01/06 03-04 25 73. 625 3 n/a n/a n/a n/a Y2 01/06 04-05/04 16 8 25004. 2 0.79 8 605. 8 Y3 01/06 04-05/04 16 4 25004. 1 0.79 4 605. 4 Y4 02/06 02/06 12.6 27 n/a 20 12.6 27 n/a 20 Y5 02/06 02/06 12.6 29 n/a 20 12.6 29 n/a 20 Y6 01–02/06 01–02/06 12.6 28 n/a 20 12.6 28 n/a 20 Y7 01–02/06 01–02/06 12.6 59 n/a 43 12.6 59 n/a 43 Y8 11/05 11/05 12.6 56 n/a 40 12.6 56 n/a 40 Y9 01–02/06 01–02/06 12.6 16 n/a 12 12.6 16 n/a 12 1. For plotless sampling (plots Y4 to Y9), number refers to sampling points. 2. Completely unburnt plots were not surveyed. 3. Two seedling plots sprayed with herbicide were excluded. 4. Individuals < 15 cm dbhob were assessed within 1600 m2 plots nested within the main plot. 5. Individuals ≥ 15 cm dbhob were assessed within the entire 2500 m2 main plot.

76 Plotless sampling at sites Y4 to Y9 Temporary plots (4 m diameter circular areas) were located at random points along transect lines. The minimum distance separating plots along the transect lines was 10 m and the maximum distance was 110 m. Transect lines were separated by at least 75 m. The lengths of transect lines were variable, being governed by site boundaries, vegetation type (areas without tuart or peppermint were avoided), soil-type boundaries and time. Where either tuart or peppermint saplings or trees (dead or live) were not present within 30 m of the sampling point, it was deemed that the position was not in tuart-peppermint woodland and therefore no sampling occurred at the point. Sampling dates are listed in Table 3.3. The bounding coordinates for the areas sampled are listed in Appendix 5.

Measurements For sites Y4 to Y9, seedlings were only counted if between 0.1 and 1.5 m tall. The size range was set due to the time required to search for small individuals amongst other plants and debris. The woody nature of the stems of some of the individuals in this size range appeared to indicate that they were more than one year old. At all other sites, seedlings were those that had established within the previous year as judged by size (< 1.5 m) and form (stems not woody).

Calculations For data collected in plots an attempt was made to determine means and errors based on data distributions (Poisson or negative binomial) using QUADRAT SAMPLING in ECOLOGICAL METHODOLOGY Version 6.1 (Exeter Software 2002). However, either a small sample size, small number of plots or a lack of fit of the data to a known distribution (Poisson or negative binomial) prevented this. Thus, estimates from data collected from plots were based on simple conversions without regard to spatial arrangement of sampled individuals. That is, the number of individuals (adults or seedlings) per hectare was calculated by:

Number of individuals x (total area sampled / 1 ha)

77 3.2.4 Tuart and peppermint population characteristics in tuart and tuart- peppermint woodlands

Hypotheses:

• The following are related to site fire history: o size-class structure for tuart and peppermint, o tuart and peppermint density, o tuart health. • Peppermint density is related to tuart tree health. • Tuart mortality is related to tree size in declining areas.

Sampling design

Permanent plots at sites Y1, Y2, Y3, GB, Ya1 and Ya2 Trees and saplings were sampled as outlined as in Figure 3.4 and Table 3.3. Sampling occurred prior to the most recent fire except in the case of site GB. Plot locations are specified in Appendix 5.

Plotless sampling at sites Y4 to Y9 Sampling occurred from points along transects (as outlined in Section 3.2.3 above). An advantage of this method was that samples could be distributed over large areas more efficiently and therefore the impact of any patch to patch variability in past fires was likely to have been lessened. At least 20 points were sampled per site, more if time permitted, and less if the site was of insufficient area (Y9). Trees were not sampled at points which were < 40 m from another sampling point along a transect so as to ensure that trees were not sampled twice. The three nearest tuart trees or saplings from each sampling point were measured, except at site Y8 where only the third nearest tree was measured. At all sites, the third nearest peppermint sapling or tree from each sampling point was also measured. Thus, as outlined in Table 3.3 there were more seedling assessment plots than tree/sapling assessment points. Sampling dates are outlined in Table 3.3 and bounding coordinates of the sampling areas are listed in Appendix 5.

Measurements

General The dbhob of all stems of saplings and trees (> 1.5 m tall) sampled was measured. Separation of individual peppermint trees was sometimes difficult owing to the presence 78 of thickets with multiple stems arising from below-ground. On some occasions, judgement was used to distinguish trees from the positioning and patterning of stems. Where uncertainty arose, the stems were deemed to belong to a single tree. Therefore, for peppermint, there may have been a slight overestimation of size and a slight underestimation of density. Tuart saplings/trees > 5 cm dbhob were assessed for canopy health (detailed in the following section) except at site GB where sampling occurred immediately following a fire. At sites Y4 to Y9, the distance to the third nearest sapling or tree of both tuart and peppermint was measured in order to calculate sapling and tree density for each species (see “Calculations” section below). For these sites, dead trees were also identified (where some part of the structure was standing above 1.5 m in height) and sampled. The measurements taken at the various sites are summarized in Table 3.4.

Table 3.4: Measurements made on tuart and peppermint saplings and trees. dbhob = diameter at breast height over bark. Sites Sampling method Measurements Y1 – Y3 Plots • dbhob live tuart and peppermint (> 1.5 m tall) • Canopy health1. live tuart trees (> 5 cm dbhob) Ya1 – Ya2 Plots • dbhob of live tuarts (> 1.5 m tall)2. • Canopy health1. of live tuart trees (> 5 cm dbhob) GB Plots • dbhob of live and dead tuarts (> 1.5 m tall) 2. Y4 – Y9 Plotless/transects • dbhob: three nearest tuarts3. to sampling point and third closest peppermint • Canopy health1. of three nearest tuarts (> 5 cm dbhob) • Distance from sampling point to third nearest tuart3. and peppermint 1. Based on a modified version of the method of Grimes (1978). 2. No height data were recorded for some seedlings/saplings. Further, where the dbhob was recorded as ≤ 1 cm, stems were sometimes < 1.5 m tall. Consequently, individuals with a dbhob ≤ 1 cm were deemed to be < 1.5 m tall and not included in the analysis. 3. At site Y8, only the third nearest tuart was measured.

Tree health Tuart tree ratings for canopy area (CA) and canopy health (CH) were adapted from Grimes (1987). The CA measure represented a slight modification of the canopy density component of Grimes’ (1987) scheme (Figure 3.5). The density of foliage within the crown structure was the main element assessed for this measure. However, some adjustment was made to the score in cases where the crown structure was reduced by

79 major branch loss (Figure 3.6a). In this case the original crown frame was estimated and visualized based on the remaining structure. Another issue with this measure related to accounting for epicormic foliage arising from the tree base or stem, particularly when trees were resprouting after fire. The approach taken here was to consider this foliage by visually projecting it onto the crown structure of the tree (Figure 3.6b). The overall rating for CH was taken as the sum of the three elements in Figure 3.5. Due to uncertainty about the ability to accurately rate small individuals, or compare the ratings of very small and large individuals, only trees with a dbhob > 5 cm were assessed. Some photographs of tuart trees with accompanying ratings are presented in Appendix 1.

Calculations The dbhob of each stem of each tree/sapling was converted to cross-sectional area and the sum of the cross-sectional areas was summed to give basal area (ba) per tree/sapling. Given the multi-stemmed habit of both species, particularly peppermint, this provided a better representation of size/age than dbhob. The relationship between ba and dbhob with a classification of growth stage is presented in (Table 3.5).

Table 3.5: Diameter at breast height over bark (dbhob) class for equivalent basal area (ba) class, assuming one stem. Generalized growth stage/dominance status listed for each size class. Where a plant comprised multiple stems, the ba per tree was calculated as the sum of the cross- sectional areas of the stems. ba per tree (cm2) dbhob (cm) Growth stage/dominance status ≤ 10 ≤ 3.5 small sapling 10.1 - 100 3.6 – 11.0 sapling - pole 100.1 - 1000 11.1 – 35.0 pole – tree 1000.1 - 10000 35.1 – 113.0 co-dominant - dominant tree ≥ 10000.1 ≥ 113.1 large dominant tree

The density of saplings and trees for sites Y4 to Y9 was determined according to the Ordered Distance Method (Morista 1957 and Pollard 1971 in Krebs 1999). The calculation was performed using DISTANCE METHODS contained in ECOLOGICAL METHODOLOGY Version 6.1 (Exeter Software 2002). Tree density at other sites was not determined. It was deemed that the different sampling method (plots) would account for sufficient error as to invalidate comparison between sites sampled with plots and sites sampled with the Ordered Distance Method.

80 a) b)

c) 9 points 7 points 5 points 3 points Figure 3.5: The scoring scheme for three components of Grimes (1987) system for rating canopy health (CH) used in this study. a) Dead branches: from 2 (main growing braches dead) to 5 (no dead branches); scores < 2 were recorded for more severe branch death, including 0 where all branches were dead. b) Epicormic growth: ranged from 1.0 (severe on crown and stem) or 1.5 (severe on crown or stem) to 3 (no epicormic growth). c) Crown density (named canopy area [CA] in this study and modified according to Figure 3.6): from 3 (sparse) to 9 (very dense); scores < 3 were recorded for sparser canopies, including 0 for where no canopy was present. Images were reproduced from Grimes (1978).

81

a)

b)

Figure 3.6: Method of rating canopy density where deviations from standard canopies or complete crowns were present. a) Where significant basal or stem resprouting occurred, the volume of the lower foliage was added to the crown-foliage volume. Note that the profile view in this demonstration exaggerates the compensation effect of the addition of the resprouting foliage, unless it is equivalent to the volume of crown lost. b) Where major branch loss had occurred, current canopy density was estimated on the basis of an idealized or assumed original crown size based on the branch structure and bole dimensions.

Analyses MINITAB Release 13 (Minitab Inc. 2000) was used for all statistical tests. Where multiple comparisons were made, α was modified in accordance with the Sequential Bonferroni method (Holm 1979 in Quinn and Keough 2002).

Size-class distributions for each population were interpreted and compared visually. Classes were arranged according to basal area on a log scale. For sites where the dbhob 82 of dead trees was also measured (Y4 to Y9), the relationship between size and mortality was determined using logistic regression with the model checked for Pearson, Deviance and Hosmer-Lemeshow Goodness of fit test results as provided by the MINITAB output. Comparison of sites Y4 to Y9 was also carried out using Principle Components (PC) Analysis. Where possible, PCs were interpreted and the spread of PC scores were examined in relation to fire history. For sites Y4 to Y9, associations between tuart tree health (for trees with a dbh > 5 cm), peppermint density and tuart density, were tested using correlation analysis. For clusters of sites other than Y4 to Y9 where tree health ratings had been recorded in the absence of recent fire (Y1, Y2 and Y3, and Ya1 and Ya2), medians were compared between adjacent sites using Mood’s Median Test.

3.2.5 Changes within vegetation monitoring plots in Yalgorup National Park: 1976 to 2004

Hypothesis:

• The structure and composition of the vegetation within Yalgorup National Park has changed since the 1970s.

Methods A search was made for vegetation assessment plots established in Yalgorup National Park by John Fox and Curtin University Ecology students in 1976/1977. Plots were re- assessed where the following conditions were satisfied: the plots were located in tuart woodland, the original data were available and reliable, and the original plot centre could be located. Thirteen plots arranged in four transects met these conditions. These plots were located at sites Y2, Y10 and Y11 (Table 3.1). Plot coordinates are listed in Appendix 5. The original assessment methods used in 1976/77 were repeated in 2004. The height and diameter of trees (defined as a stem greater than 20 cm dbhob) were recorded within a 500 m2 circular area and the species and projected cover of shrubs were noted for the central 100 m2 circular area. Braun-Blanquet cover classes (BBCC) were used to classify cover estimates (Kent and Coker 1992). The original and 2004 data were summarized by listing: the species of the dominant tree/s (accounting for number and size), the species and height of the tallest tree, the species and BBCC for tree saplings/seedlings, and the species and BBCC for the dominant shrubs. For some plots, the original data did not allow a BBCC to be determined for the dominant shrub. A time series of aerial photographs covering the plots at site Y11 from 1957 to 2003 was also examined for major vegetation structural change. 83 3.3 RESULTS

3.3.1 Tuart and peppermint seedling regeneration at unburnt and recently burnt areas

At sites unburnt for > 4 years, tuart seedlings were mostly absent (Table 3.6). At the two unburnt sites where tuarts were found, there were fewer than 30 seedlings ha-1 estimated. In contrast, tuart seedlings were found at all recently burnt sites and in quantities at least six times as great as the unburnt sites. Peppermint seedlings were present at both burnt and unburnt sites. As only one burnt site (Y1) with peppermint was examined, it was difficult to compare densities between burnt and unburnt sites. However, compared to tuart, the density of peppermint seedlings was greater by a factor of > 10 at the two unburnt sites and > 25 at the recently burnt site.

Table 3.6: Results of seedlings surveys in tuart woodlands between 2004 and 2006. Sites in bold type were burnt within approximately one year of the survey. Other sites had been unburnt for > 4 years. Sampling area (m2) No. seedlings1. Est. no. seedlings ha-1 Site Tuart Peppermint Tuart Peppermint Tuart Peppermint Y1 352 17.3 14 19 398 10 982 GB 2500 n/a 48 n/a 192 n/a Ya1 300 n/a 16 n/a 533 n/a Ya2 175 n/a 15 n/a 857 n/a Y2 128 6.3 0 10 0 15 873 Y3 64 3.1 0 3 0 9677 Y4 339 339 0 14 0 413 Y5 364 364 0 14 0 384 Y6 352 352 0 5 0 227 Y7 741 741 2 37 27 499 Y8 703 703 2 20 28 284 Y9 201 201 0 5 0 249 1. At sites Y4 to Y9, seedlings were defined as individuals between 0.1 and 1.5 m tall. Elsewhere, seedlings were those that had established within the previous year as judged by size (< 1.5 m tall) and form (stems not woody).

3.3.2 Tuart and peppermint population characteristics in tuart and tuart- peppermint woodlands

The size-class distributions of tuart and peppermint stands in the Yalgorup area had different shapes (Figures 3.7 and 3.8). With the exception of site Y5, the smallest size

84 class (< 10 cm2 ba) contained the greatest proportion of individuals in the peppermint populations (Figure 3.8a). For tuart, the proportion of individuals in the larger size classes was greater (Y6, Y7 and Y8) or more of a normal-type distribution prevailed (Y1, Y2, Y3, Y4, Y5 and Y9). In contrast to Yalgorup, a large proportion of small saplings were a feature of the tuart populations at Golden Bay and in Yanchep National Park (Figure 3.7d, e and f).

Another feature distinguishing tuart and peppermint was the presence of dead individuals for tuart and the virtual absence of dead trees for the peppermint populations (Figure 3.8). Of the sites where dead trees were counted, the two in National Park (Y6 and Y7) had the highest proportion of dead tuarts: 34 % and 28 %, respectively. A logistic regression for dbhob as a predictor of death revealed a positive and highly significant result (Table 3.7). Initially, basal area per tree was used as a predictor variable, but regression model assumptions were violated (P < 0.05 for Deviance and Hosmer-Lemeshow χ2 tests). Therefore, the dbhob of the largest stem for each tree was used in preference.

Table 3.7: Logistic regression results for tree death (dependent variable) versus diameter at breast height (dbhob) of largest stem (main trunk) per tree (n = 249) for tuarts at six sites in the Yalgorup area (sites Y4 to Y9). Regression was significant: G(1) = 10.109, P = 0.001. Coefficient Odds 95 % CI Variable SE Z P ratio Lower Upper Constant -1.58 0.234 -6.74 <0.001 dbhob (cm) 0.011 0.003 3.15 0.002 1.01 1.00 1.02

A summary of a range of additional data pertaining to the tuart and peppermint populations at the six sites surveyed in Yalgorup in 2005 and 2006 (sites Y4 to Y9) are presented in Tables 3.8 and 3.9. A notable feature of the data is the high ratio of peppermint:tuart density at all sites, ranging between 3.6 and 11.9 (Table 3.9).

85 ) ) ) 80 80 80 70 70 70 60 60 60 50 50 50 40 40 40 30 30 30 20 20 20 10 10 10

(% population of Proportion 0 (% population of Proportion 0 (% population of Proportion 0 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ 2 2 2 a) Basal area class max. (cm ) Basal area class max. (cm ) d) Basal area class max. (cm ) ) ) ) 80 80 80 70 70 70 60 60 60 50 50 50 40 40 40 30 30 30 20 20 20 10 10 10

Proportion of population (% population of Proportion 0 (% population of Proportion 0 (% population of Proportion 0 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ b) 2 2 2 Basal area class max. (cm ) Basal area class max. (cm ) e) Basal area class max. (cm ) ) ) ) 80 80 80 70 70 70 60 60 60 50 50 50 40 40 40 30 30 30 20 20 20 10 10 10 Proportion of population (% population of Proportion Proportion of population (% population of Proportion 0 (% population of Proportion 0 0 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ 2 c) Basal area class max. (cm2) Basal area class max. (cm2) f) Basal area class max. (cm )

Figure 3.7: Proportion of sample population in basal area classes for live tuart (■) and live peppermint (■) ≥ 1.5 m tall. Sites sampled and n (tuart, peppermint) were a) Y1 (98, 63), b) Y2 (21, 28), c) Y3 (7, 27), d) GB (371, n/a), e) Ya1 (140, n/a), f) Ya2 (186, n/a). Sampling occurred in April to May 2004 (Y1, Y2 and Y3), February 2003 (GB) and April 2003 (Ya1 and Ya2). Basal area class maximum is denoted on x axis, for example: “100” = 10.1 – 100 cm2.

86 ) ) ) ) 80 80 80 80 70 70 70 70 60 60 60 60 50 50 50 50 40 40 40 40 30 30 30 30 20 20 20 20 10 10 10 10

0 (% population of Proportion 0

(% population of Proportion 0 (% population of Proportion (% population of Proportion 0 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ 10 100 1000 10000 10000+

2 2 2 a) Basal area class max. (cm ) Basal area class max. (cm ) d) Basal area class max. (cm ) Basal area class max. (cm2) ) ) ) ) 80 80 80 80 70 70 70 70 60 60 60 60 50 50 50 50 40 40 40 40 30 30 30 30 20 20 20 20 10 10 10 10

Proportion of population (% population of Proportion 0 (% population of Proportion 0 Proportion of population (% population of Proportion (% population of Proportion 0 0 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ 2 2 b) Basal area class max. (cm ) Basal area class max. (cm2) e) Basal area class max. (cm ) Basal area class max. (cm2) ) ) ) ) 80 80 80 80 70 70 70 70 60 60 60 60 50 50 50 50 40 40 40 40 30 30 30 30 20 20 20 20 10 10 10 10

0 (% population of Proportion 0 Proportion of population (% population of Proportion 0 0 Proportion of population (% population of Proportion (% population of Proportion 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ 10 100 1000 10000 10000+ c) 2 2 f) 2 2 Basal area class max. (cm ) Basal area class max. (cm ) Basal area class max. (cm ) Basal area class max. (cm ) Figure 3.8: Proportion of sample population in basal area classes for tuart (■) and peppermint (■) ≥ 1.5 m tall. Proportion of dead individuals is denoted by □. Sites sampled and n (tuart, peppermint) were a) Y4 (60, 20), b) Y5 (60, 20), c) Y6 (60, 20), d) Y7 (129, 43), e) Y8 (40, 40), f) Y9 (36, 12). Sampling occurred from November 2005 to March 2006. Basal area class maximum is denoted on x axis, for example: “100” = 10.1 – 100 cm2.

87 Table 3.8: Characteristics of tuart population (individuals > 1.5 m height) sampled at sites in the Yalgorup area from November 2005 to March 2006. ba = basal area, dbhob = diameter at breast height over bark of the largest stem per tree. CI = confidence interval. Tree refers to trees and saplings Site Mean ba Mean Density Dead Median tree health1. tree-1 (cm2) ± SE dbhob (cm) ± SE (trees ha-1) 95 % CI (%) 95 % CI n All IQ range n Live IQ range n Y4 4384 ± 836 54 ± 6.5 30 23 - 39 18 9.5 – 30.4 60 6.5 3.5 – 9.0 53 7.5 5.5 – 9.1 42 Y5 3706 ± 928 42 ± 7.0 25 19 - 32 5 1.0 – 13.9 60 8.5 5.9 – 10.8 42 9 6.0 – 11.5 39 Y6 2071 ± 375 36 ± 4.4 72 56 - 93 38 26.1 – 51.8 60 4 0.0 – 9.0 51 8.5 6.0 – 11.4 28 Y7 3306 ± 401 47 ± 3.5 42 35 - 50 24 16.9 – 32.3 129 5 0.0 – 7.5 107 6 4.0 – 8.5 76 Y8 1642 ± 541 31 ± 5.1 33 28 - 38 5 0.6 – 16.9 40 7.5 6.0 – 9.5 33 7.5 6.1 – 9.5 32 Y9 1404 ± 333 32 ± 4.2 15 11 - 21 6 0.7 – 18.7 36 8.5 7.0 - 10 34 8.75 7.0 – 10.0 32 1. For saplings or trees with a dbhob > 5 cm. Rating is based on a modified version of the method of Grimes (1978).

Table 3.9: Characteristics of peppermint population (individuals > 1.5 m height) sampled at sites in the Yalgorup area from November 2005 to March 2006. ba = basal area, dbhob = diameter at breast height over bark of the largest stem per tree. CI = confidence interval. Tree refers to trees and saplings. Site Mean ba Mean Density Dead trees Peppermint:tuart tree-1 (cm2) ± SE dbhob (cm) ± SE (trees ha-1) 95 % CI (%) 95 % CI n density Y4 89 ± 38 6 ± 1.4 326 251 - 419 0 0.0 – 13.9 20 10.9 Y5 1460 ± 661 11 ± 2.6 211 163 - 271 0 0.0 – 13.9 20 8.4 Y6 268 ± 101 9 ± 1.8 258 199 – 331 0 0.0 – 13.9 20 3.6 Y7 249 ± 82 7 ± 1.2 450 378 – 535 0 0.0 – 6.7 43 10.7 Y8 152 ± 62 6 ± 1.0 281 235 – 336 0 0.0 – 7.2 40 8.5 Y9 69 ± 23 5 ±1.2 115 82 - 159 0 0.0 – 22.0 12 7.7

88 Results of Principle Components (PC) Analysis (Table 3.10) revealed that site Y5 was most distinct from the others when scores for PC1 and PC2 were plotted (Figure 3.9). This site is the longest unburnt site within the group of sites outside the National Park area. Tuart health ratings (inclusive of dead individuals) and tree density (both tuart and peppermint) appeared to be an important element of PC1 (Table 3.10). Towards the right end of the PC1 axis, the sites generally had higher tuart health ratings in combination with a lower density of both tuart and peppermint trees (Figure 3.9). Sites Y6 and Y7, ranked lowest along this axis, were both located in National Park and had been subject to few fires in recent decades. The contrast between the proportion of tuarts and peppermints in the < 10 cm2 basal area class was influential in determining PC2 scores (Table 3.10). Consequently, with the combination of a greater proportion of tuarts and a lesser proportion of peppermints in this size class compared to the other sites, sites Y5 and Y6 were ranked highest for PC2 (Figure 3.9). The coefficients for PC3 were not readily interpretable (Table 3.10), so score plots were not examined.

Table 3.10: Results of Principle Components (PC) analysis for a selection of variables measured for the tuart and peppermint populations at sites Y4 to Y9. Results for the first three PC only are presented. Tree refers to trees and saplings. Coefficients Variable PC1 PC2 PC3 Mean basal area per tuart (cm2 tree-1) -0.040 0.320 0.513 Mean basal area per peppermint (cm2 tree-1) 0.332 0.420 0.141 Tuarts ≤ 10 cm2 ba (%) 0.157 0.529 0.205 Tuart density (trees ha-1) -0.334 0.308 -0.382 All tuarts (> 5 cm dbh) median health rating1. 0.453 -0.224 0.201 Live tuarts (> 5 cm dbh) median health rating1. 0.395 0.086 -0.413 Dead tuarts (%) -0.392 0.262 -0.300 Peppermints ≤ 10 cm2 ba (%) -0.278 -0.444 0.260 Peppermint density (trees ha-1) -0.399 0.151 0.400 Eigenvalue 3.808 2.840 1.856 Cumulative proportion of variance (%) 42.3 73.9 94.5 1. For saplings or trees with a dbhob > 5 cm. Rating is based on a modified version of the method of Grimes (1978)

89 3 Y5 Y6 2

1 Y7 Y4 PC 1P 0 -3-2-10123 -1 Y8 -2 Y9

-3

P 2 PC

Figure 3.9: First and second principle components (PC) scores relating to data on tuart and peppermint populations at sites surveyed in the Yalgorup area in November 2005 to March 2006. Variables and PC coefficients are presented in Table 3.10.

Given the importance of tree health ratings in distinguishing sites in the above PC analysis, the health ratings for live tuart trees at other sites were compared within each locality (Table 3.11). Trees at site Y2 had significantly poorer canopy health than those at site Y1. This value was also lower than any of the health ratings recorded for live trees at the other six sites surveyed in Yalgorup in 2005 to 2006 (Table 3.8). Sample sizes for sites Y3, Ya1 and Ya2 were probably too small for any differences to be revealed in the comparisons.

Table 3.11: Median health rating for live tuart trees (> 5 cm diameter at breast height over bark) at adjacent sites at Yalgorup and Yanchep National Parks in 2004. Medians with different letters are significantly different for sites within each locality using Mood’s Median Test with Bonferroni corrections. Locality Site Median health rating1. IQ range n Yalgorup Y1 8.0 a 6.5 – 9.1 70 Y2 5.0 b 4.1 – 7.0 20 Y3 9.0 ab 5.8 - 10.0 6 Yanchep Ya1 9.3 a 7.5 – 10.5 16 Ya2 8.8 a 8.4 – 10.6 10 1. Rating is based on a modified version of the method of Grimes (1978)

The health of live tuarts (> 5 cm dbhob) was negatively related to peppermint density for sites Y4 to Y9 (Figure 3.10). The association was significant after a Bonferroni correction was applied for four comparisons (correlation coefficient = -0.912, P = 90 0.011, n = 6). The health of live tuart trees was not significantly correlated with tuart density. Correlations between the health ratings of both live and dead tuart trees combined, with peppermint density or tuart density, were also not significant. However, with only six sample points these tests were severely lacking in statistical power.

10

8

6

4 tuart trees tuart

2

Median health rating for live for rating health Median 0 0 100 200 300 400 500

Peppermint density (trees ha-1)

Figure 3.10: Median health rating for live tuart trees (> 5 cm diameter at breast height) versus peppermint density (saplings and trees > 1.5 m tall) across six sites (Y4 to Y9) in the Yalgorup area in November 2005 to March 2006. Rating was based on a modified version of the method of Grimes (1978). Tree refers to saplings and trees.

3.3.3 Changes within vegetation monitoring plots in Yalgorup National Park: 1976 to 2004

In 6 of 13 plots, the dominant tree species recorded in 2004 differed to that in 1976/1977 (Table 3.12). In each case, peppermint replaced the original dominant (tuart, Banksia attenuata or Allocasuarina fraseriana), or formed a tree layer where none existed in 1976/77 (using the definition of tree applied in the original survey: stems greater than 20 cm). Where a tree layer of peppermint developed during the monitoring period, seedlings or saplings of this species were recorded in 1976/1977 for two of the three plots (plots 141and 144). There were also three plots (139, 140 and 143) where peppermint was present in 2004 (as seedlings or saplings) but not in 1976/1977.

91 Table 3.12: Changes within vegetation monitoring plots between 1976/1977 (normal type) and 2004 (bold type). Dominance (Dom.) defined as the greatest projected (canopy) cover. An absence of tree species is denoted by “-”. Braun-Blanquet cover classes1. are in brackets (last two columns) where available. Plot Site Trees Shrubs ID ID Dom. spp. Max. height (m) Saplings/seedlings Dom. Spp. 362. Y11 Ba Ba (11) Af (1), Ba (1) Hh (1) Ba Ba (15) Af (+) Sg (2) 372. Eg Eg (9) - Ma (2) Eg Eg (9) - Ma (2) 382. Eg Eg (14) - Ma (2), Gt (2) Eg Eg (9) - Sg (2) 39 Eg, Ba, Af Eg (15) Af (+), Ba (+) Hh (1), Mr (1) Eg, Ba, Af Eg (13) - Hh (1), Mr (1) 40 Ba Eg (19) Af (1) Hh (2), Mr (2) Af Af (15) Af (1) Tr (2) 51 Y10 Eg Eg (21) Eg (+) Ap (1) Eg Eg (22) Eg (1) Xp (2) 52 Eg Eg (20) Af (2) Gt (1) Af Eg (21) Af (2), Eg (1) Xp (2) 138 Y2 Alf Alf (12) Eg , Afl , Af Hh Af Eg (13) Alf , Af Hh 139 Eg Eg (17) - Hh Eg Eg (20) Eg (+), Alf (1), Af (1) Hh 140 Alf Alf (8) - Hh Alf Alf (14) Af (+) Hh 141 - Af (2), Ar (1) Hh (1) Af Af (13) Af (+), Ar (2) Hh (1) 143 - - Ar (1) Ap (+), Hh (+) Af, Ar Af (10) Af (1), Ar (3) Hh (1) 144 - - Af (1) A (1), Gt (1) Af Af (10) Af (3) Tr (2) Species abbreviations: Af (peppermint: Agonis flexuosa), Alf (Allocasuarina fraseriana), A (unknown Acacia spp.), Ap (Acacia pulchella), Ar (Acacia rostellifera), Ba (Banksia attenuata), Eg (tuart: Eucalyptus gomphocephala), Gt (Grevillea thelemanniana), Hh (Hibbertia hypericoides), Ma (Melaleuca acerosa), Mr (Macrozamia riedlei), Sg (Spyridium globulosum), Tr (Templetonia retusa), Xp (Xanthorrhoea preissii) 1. Braun-Blanquet classes for projected cover: + (< 1%), 1 (1-5 %), 2 (6-25 %), 3 (26–50 %), 4 (51– 75 %) and 5 (76-100%). 2. From reference to student observations in 1976 (records held by J. Fox, Curtin University), assumed unburnt in 1972. These plots occur near the boundary of the 1972 fire reported by Smith (1975).

92 The structure and composition of the vegetation at site Y11 changed substantially from 1976/1977 to 2004, with a decline in the height of tuart trees and the death of B. attenuata trees (Table 3.12). For example, in Plot 40, the dominant species in 1977 was B. attenuata and the tallest tree was tuart (19 m), but in 2004 peppermint was both the dominant and tallest tree (15 m), B. attenuata was absent and tuart had a severely reduced canopy. Numbers of living B. attenuata trees fell from 19 to just five in these five plots over the monitoring period. The poor condition of some tuart canopies in these plots was also observed in the 1976/1977 survey and associated with the presence of fire and borer damage (D. Ward pers. comm.).

The data in Table 3.12 also reveal differences in the shrub layer between 2004 and 1977. Tall growing species: Spyridium globulosum (plots 36 and 38), Templetonia retusa (plots 40 and 144) and Acacia rostellifera (plots 141 and 143) have become more prominent.

For site Y11, visual evidence in the form of an aerial photographic time-series confirms the trend towards a denser understorey (Figure 3.11). This is most noticeable in the once more open expanses on the western half of the area. The increasingly textured appearance represents understorey growth. On-ground observations also revealed that shrubs such as S. globulosum were attaining heights above 2 m and were forming dense thickets (see Appendix 2 for a related study). Understorey and/or midstorey development over time is also discernable between tree crowns in the eastern half of the area within the photos (Figure 3.11c and d). The increased height and density of the lowerstorey appears to have occurred largely after the mid-1970s.

93

Figure 3.11: Aerial photographs of land between Lake Preston (left) and Lake Yalgorup (right) in Yalgorup National Park (site Y11) taken in a) 1957, b) 1976, c) 1991 and d) 2003. Increasing vegetation density over time can be seen within the open areas on the western (left) half of the photographs and within the woodland areas on the eastern (right) half of the photographs.

94 3.4 DISCUSSION

The results in this Chapter revealed a number of important trends within the tuart woodlands of Yalgorup. Peppermint seedling establishment has been abundant in both burnt and unburnt areas, whereas tuart seedling establishment has been largely confined to recently burnt areas. Tuart seedling establishment has been most abundant at sites outside Yalgorup that were burnt by wildfire. The contrasting size-class structures of tuart and peppermint in Yalgorup reflect this difference in regeneration mode; individuals in the large size-classes dominate for tuart, whereas the peppermint populations are dominated by individuals in the smaller size-classes. The tuart and peppermint populations also differ markedly in that many of the tuarts in the larger size- classes are either dead or in poor health. The relationship between fire regime and tuart health is complex. At the longest-unburnt site (unburnt since 1967) the trees are in poor health, but health is also poor at sites that have been burnt on an approximately decadal rotation of controlled fires. Additionally, trees were in much poorer health at a site that had been subject to frequent wildfire, compared to an adjacent site that had not. These data support the view that extreme fire regimes may be detrimental to tree health but they also suggest factors other than fire are contributing to tuart decline. As the health of tuart trees was negatively associated with peppermint density, competition cannot be ruled out as a factor. Data from vegetation monitoring plots showed that peppermint along with other understorey species have been increasing in density and height within Yalgorup National Park since the 1970s. These trends and their implications are elaborated upon below.

3.4.1 Tuart and peppermint seedling regeneration at unburnt and recently burnt areas

The results of the seedling regeneration surveys clearly indicate the importance of fire in the success of natural tuart seedling regeneration, whereas peppermint regeneration is abundant both following fire, and in the absence of fire. These findings support those of Ruthrof (2001) for tuart and the views of the Forests Department of Western Australia (1979) and Bradshaw (2000) for tuart and peppermint. While only one recently burnt site with peppermint was available for examination (Y1), the large number of seedlings present suggests that low-intensity fire results in no major limitation, if not a benefit, to peppermint seedling regeneration. In fact, a high density of peppermint seedling regeneration (> 650 seedlings ha-1) has also been found at a site following intense 95 wildfire (L. McCaw pers. comm.). This temporal and spatial overlap in seedling regeneration between the species raises the possibility of interspecific competition at the seedling stage. Without standardizing the results for the density of adult trees and seed supply, it is difficult to compare between sites or comment further on the role of fire intensity or season in the regeneration success of either species based on these data alone.

The discovery of two tuart seedlings at each of two unburnt sites highlights the possible role for infrequent events in time and uncommon events in space to contribute to the recruitment of long-lived species (Wellington and Noble 1985a, Higgins et al. 2000). At site Y7, the seedlings were found in March 2006. Thus, having survived the summer period, these seedlings had a high chance of establishing. The 2005-2006 summer appeared more favourable than normal for seedling survival with lower than average December temperatures (22°C compared to the average of 27°C) and an unusually high rainfall total for January (75 mm versus the average of 10 mm) (Figure 3.2). When calculating the density of these seedlings on an area basis, sufficient numbers were present to enable the approximate replacement of the dominants, thereby ensuring the persistence of tuart woodland. However, with virtually all sampling plots recording seedling absence, the error of this density estimate was clearly very wide. Further, recruitment is also dependent on these seedlings surviving additional hazards (for example, competition) preceding and following the sapling stage. Nevertheless, the possibility that important levels of recruitment may occur within the life span of tuart in some areas in the total absence of fire or other disturbance cannot be refuted. Just 20 to 30 seedlings ha-1 growing to maturity over a period of 200 to 400 years would arguably be sufficient recruitment for the persistence of a stand. Additionally, disturbances other than fire, such as tree-fall associated with windstorms, can promote eucalypt seedling establishment between fires (Yates 1994).

3.4.2 Tuart and peppermint population characteristics in tuart and tuart- peppermint woodlands

The high proportion of tuart in the larger size-classes (stem cross-sectional areas larger than 100 cm2) for sites at Yalgorup suggests that little recruitment or seedling establishment has occurred for tuart in the last one to two decades at these sites. As tuart seedling establishment was shown to occur almost exclusively following fire and as most sites had been burnt in the 10 to 15 year period prior to sampling, this result was 96 unexpected. While it is possible that some of the individuals that developed following fires in the 1990s could have exceeded 100 cm2 size by the time of sampling, data from other tuart woodlands (sites GB, Ya1 and Ya2) where abundant post-fire regeneration was apparent 10 to 15 years after, suggests the majority of the post-fire cohorts would probably have developed more slowly. An accumulation of large numbers of suppressed eucalypt seedlings/saplings in the decades following significant regeneration events has also been observed in other eucalypt forests (for example, Ashton 1976) and in a tuart woodland six years after an intense fire (L. McCaw pers. comm.). Therefore, episodes of tuart seedling regeneration in large numbers should be reflected in the size-class structure for periods of at least 10 to 15 years, especially if dead individuals are included in the surveys as was the case here.

These findings cast doubt on the usefulness of a single, standard controlled burn for tuart recruitment. Like most other eucalypts (Florence 1996), tuart is generally considered a “gap” regenerator (Forests Department of Western Australia 1979, Bradshaw 2000) and has an apparent dependence on ashbeds for establishment (Keene and Cracknell 1972, Forests Department of Western Australia 1979). However, as the preceding fires were in most cases controlled fires, a proportion of which were lit in spring, it is possible that establishment was limited by the lack of ashbeds. Coarser woody debris is generally less likely to burn in spring than later in the season due to the inverse relationship between fuel size and drying rate (Luke and McArthur 1978). Burrows et al. 1990 concluded that early autumn fires were required to maximize the regeneration of the ashbed-dependent E. wandoo. Controlled burning in spring as opposed to autumn is also less likely to result in large crops of seedling regeneration for other reasons. These include the possibility of germination just prior to the summer drought period (Baird 1977, Hobbs and Atkins 1990) or the greater probability of seed predation (Cowling and Lamont 1987) or the loss of seed viability in the months before the following wet season (Cremer 1965b, Ruthrof 2001). However, sites in Yalgorup burnt in autumn in the 10 to 15 year period prior to sampling, also had as proportionally few saplings in the population as the sites where burn season was unknown or in spring. One site (Y4) was burnt in autumn only six years prior to sampling, yet had as proportionally few saplings as at other sites, including a long unburnt site (Y7). Thus, failure to establish and/or failure to progress from the seedling to sapling stage could limit recruitment at sites burnt by controlled fires. However, the results of the seedling survey indicate that seedling establishment can occur following controlled fires,

97 although it is uncertain as to the extent the unusually mild summer enhanced survival; continued monitoring would have been necessary to confirm establishment actually took place. Failure to progress from the seedling to sapling stage may be the “bottleneck” at sites burnt by controlled fire. The relative importance of the ashbed versus the gap in this is not clear.

The potential abundance of post-fire regeneration is undoubtedly also determined by adult tree-canopy health. The site burnt in autumn 2000 (Y4) had markedly poorer tuart canopy health and proportionally fewer tuart saplings than a nearby site with healthier trees burnt in 1992 (Y5). Similarly, the tuart population at site Y2 had relatively few saplings, and the trees were in significantly poorer health than the adjacent site Y1. Both of these sites were burnt in autumn and within two years of one another (1993 and 1991, respectively). Tuart canopy decline as a factor in reduced seedling regeneration has probably become more important in recent years as the decline has progressed. However, at other sites where decline is present today, seed supply may not have limited regeneration to an ecologically significant extent in the early years of the decline (the first half of the 1990s). A wildfire that burnt through a section of tuart woodland on the corner of Old Coast Road and Preston Beach Road in 1996, resulted in tuart seedlings and saplings at densities > 650 ha-1 six years later (L. McCaw pers. comm.). This site is within the general area of most severe tuart decline at present but canopy condition prior to the fire is unknown.

In contrast to the Yalgorup sites, tuarts in the smaller size-classes greatly outnumbered those in the larger size-classes at sites north of the Yalgorup area (GB, Ya1 and Ya2). These sites were also burnt once in the 10 to 15 year period prior to sampling, but by wildfires, not controlled fires. Consequently, the area of ashbed formed was likely to have been more extensive, thereby enabling more abundant tuart seedling establishment. In addition, the fires possibly resulted in greater destruction of the pre-fire vegetation, thus creating more gaps within which seedlings and saplings could develop over the longer term. The high proportion of individuals in the smallest size-class at these sites highlights the fact that most saplings become suppressed by competition following establishment, even after severe fire.

For a tuart woodland or forest to persist, there is little requirement for large numbers of saplings to develop where healthy dominants exist (Bradshaw 2000). Thus in healthy

98 stands, the low level of recruitment accompanying controlled burning has little ecological significance. Nevertheless, over the long term, a “trickle” of recruitment may be important in the maintenance of tuart dominance. Where rapid, irreversible tuart tree decline is occurring at a site, abundant seedlings or saplings following a previous severe fire would reduce the probability of a permanent loss of tuart dominance.

None of the fire regimes examined appears to have provided protection to tuart stands in Yalgorup from the decline process. A long unburnt site (Y7), a site where controlled burns have been conducted on a decadal rotation (Y4) and a site where three wildfires occurred between 1972 and 1985 (Y2), all had stands with poor canopy health. The two National Park sites (less-frequently burnt) assessed as part of a six-site survey (Y4 to Y9) had poorer tree health compared to other sites when dead trees were included in the results. However, the effect that other differences in management history (grazing and logging) between National Park and State forest sites may have had was unknown. Considering live tree health for the sites in National Park, the trees at Y6 (burnt in 1998) were considerably healthier than at Y7 (unburnt since 1967). While at site Y2 (subject to frequent wildfire), health was significantly poorer than at the adjacent Y1 site which has been burnt by wildfire and controlled fire, and burnt less frequently. On the basis of fire history alone, these trends support the theories linking canopy decline processes with long periods without fire (Jurskis and Turner 2002) or frequent intense fire (Beard 1967), but properly replicated studies would be needed to confirm this. Regardless, the finding that trees are also in poor health at sites that have been subject to decadal low intensity fire indicates that other factors unrelated to fire are important in the decline process.

The greater incidence of tuart death in larger tree size classes in the Yalgorup area prompts speculation on the ways in which age or size-related factors may be interacting with the decline process. Older trees in the process of senescencing naturally, may simply be more vulnerable to the impact of any number of secondary decline agents. However, when size is considered in isolation, it is conceivable that a change in site and/or climatic conditions resulting in a reduced availability of resources may impose supply-demand pressures on trees that have grown to their maximum potential. There is a strong correlation between tree size and available moisture in southern Australia (Specht and Specht 1989), but in the southwest of Western Australia a step reduction in annual rainfall totals has occurred since the middle of last century (Indian Ocean

99 Climate Initiative 2002). At the same time, the density of the lowerstorey vegetation in the Yalgorup area appears to have increased from during the same period as was noted by this study and by Bradshaw (2000). Thus, two separate processes acting to reduce available soil moisture (and possibly nutrients) have been operating. Trees that attained sizes under climatic and management regimes earlier last century, may have died or severely declined as moisture regimes changed. Therefore, could the large dead tuarts that stand in parts of Yalgorup today be monuments to the potential of a previous age?

In comparison to tuart, the peppermint populations at the six sites surveyed in Yalgorup were healthy, with virtually no dead individuals present, and the size-class structure was suggestive of a stable or increasing population. This pattern was also reported for peppermint in areas of Yalgorup National Park in the late 1970s (Backshall 1983). There was no indication that fire regime had affected the recruitment or the survival of peppermint saplings or trees. The formation of a lignotuber greatly increases the probability for post-fire recovery of peppermints beyond the young seedling stage (Bradshaw 2000). In addition, as demonstrated by the results of the seedling regeneration surveys, seedling regeneration post-fire and in interfire years would compensate to some degree for any losses due to periodic fire.

Sites Y5 and Y9 were distinct in that there were fewer peppermint in the 100 cm2 basal area class relative to other sites. As sites Y5 and Y9 also supported the healthiest tuart canopies, it is conceivable that the tuart overstorey was maintaining some suppressive effect on peppermint development there. This implies that declining tree health at other sites has allowed peppermint to develop. The finding of a significant negative association between tuart health and peppermint sapling/tree density also supports this proposition. Alternatively, this trend may be indicative of peppermint competing for soil moisture and nutrients to the detriment of tuart. Therefore, it is difficult to draw conclusions as to whether there is a competitive imbalance between the tuart overstorey and the peppermint lowerstorey, and what the nature of it may be. The superior competitor (peppermint or tuart) may differ between sites and over time. The possibility of peppermint trees competing with tuart seedlings and limiting tuart recruitment has been raised (Backshall 1983). The high density of peppermint saplings/trees (at least three times as great as tuart) combined with their assumed capacity to recover after fire, could be a factor contributing to the lack of recruitment in the tuart-peppermint stands

100 burnt by controlled fire. Manipulative field experimentation accompanied by long-term monitoring would be the only way to clarify the issues relating to competition.

3.4.3 Changes within vegetation monitoring plots in Yalgorup National Park: 1976 to 2004

The changes in the dominant tree species and the height of the tallest tree within vegetation plots in Yalgorup National Park between 1976/1977 and 2004 were consistent with general observations of the decline of tuart and the increasing prominence of peppermint (Bradshaw 2000, Ward 2000, Longman and Keighery 2002). The data indicate that where peppermint has become the dominant tree, seedlings/saplings were generally established at the time of the first assessment. However, three plots (at site Y2) where the species was not recorded in 1976/1977 now have trees or seedlings/saplings growing within them, suggesting some of the peppermint expansion has occurred in recent decades. Elsewhere, peppermint was absent from plots 37, 38 and 52, despite being located within 100 m of other plots where the species was recorded at both sampling times. Seed supply was unlikely to have limited the advance. Adult trees produce seed annually (pers. obs.) and the small winged seed (Boland et al. 1984) is undoubtedly carried some distance with the wind. When the findings of the seedling regeneration surveys are considered, it is not clear why seedlings have failed to establish in these plots. The absence of fire, beginning from well before the plots were established, could have limited the opportunity for seeds to land on suitable substrate due to the intact vegetation-litter layer. In contrast, at site Y2, the three fires since 1976 possibly created greater opportunity for seedling establishment. Fire disturbance has been linked with the invasion of native perennials elsewhere in Australian ecosystems (Burrell 1981).

The marked decline in the abundance of living Banksia attenuata trees at site Y12 may have arisen through natural senescence coupled with a lack of fire events (regeneration opportunities). At the time of sampling in 2004, the plots had not been burnt for 30 years. Like tuart, regeneration of B. attenuata seedlings is closely tied with fire events (Cowling and Lamont 1987). Alternatively, the loss may be due to disease as Banksia spp. are especially susceptible to Phytophthora cinnamomi (Wills 1993, Shearer et al. 2004) and the pathogen may have been introduced in association with plot establishment and subsequent surveys.

101 The substantial cover of the tall growing shrub species Templetonia retusa (plots 36 and 38) and Spyridium globulosum (plots 40 and 144) where they were not recorded in the original surveys, and the increased cover of Acacia rostellifera (plots 141 and 143) which can develop as a tree (Powell and Emberson 1981), together represent a general trend towards understorey development. This trend is not surprising given the period of time since fire or the presence of episodic intense fire at the sites examined. The expansion of some shrub species within the National Park may also have been assisted by the cessation of grazing around the middle of the last century (Bradshaw 2000). As with the increased peppermint density, the growth of the shrub layer could have exerted competitive pressures on tuart. More certain is that the increased height and density of the lowerstorey has raised the probability of tuart crown damage in wildfire or in controlled fire carried out at low fuel-moisture conditions.

3.4.4 Conclusion

In conclusion, associations were revealed between patterns of tuart regeneration and fire history and between tuart health and peppermint density, but it was not possible to demonstrate a clear association between tuart health and fire history alone. Obligate and facultative modes of seedling regeneration in relation to fire were confirmed for tuart and peppermint, respectively. However, the contribution to tuart recruitment of infrequent seedling establishment between fires is unknown when century-scale time frames are considered. In contrast to intense/severe fire events, controlled low-intensity fires in recent decades appear to have resulted in little tuart recruitment. It was speculated that an insufficient area of ashbed and/or a lack of gaps has imposed limitations on tuart seedling establishment and/or sapling development. While the presence of seedling regeneration in the summer at two sites following low-intensity controlled burning points to the lack of gaps being the greater limitation, the unusually mild summer in 2005-2006 prevents any firm conclusions from being drawn. Regular low-intensity fire within the life-span of tuart could result in incremental recruitment (“trickle” recruitment) maintaining tuart dominance. Presumably, this is how tuart persisted under the fire regimes maintained by Aborigines and European graziers. However, in the presence of severe tuart canopy decline and the possibly inhibitive effect of an increasingly dense peppermint midstorey, such rates of replacement are likely to be insufficient to prevent local extinction. Contrasts in the abundance, the size- class structures and the health of tuart and peppermint populations at the majority of sites examined in Yalgorup are indicative of succession towards a peppermint 102 dominated community there. Trends observed in time support this finding. If periodic (decadal) low-intensity fire has insulated some sites from tuart canopy decline and the successional trend more generally, it has not done so for all. Thus non-fire processes are undoubtedly involved in the tuart decline at Yalgorup. Data are supportive of a competitive interaction between tuart trees and the peppermint mid-storey, but further work is required to identify the superior competitor and the consequences of the interaction. Possible competitive interactions between tuart and peppermint at the seedling stage and between established peppermints and tuart seedlings also require investigation to understand tuart-peppermint stand dynamics.

103

104 4 Chapter 4. Survival and recovery of tuart and co-occurring tree species following exposure to fire

Paul Drake was a co-contributer to the design, collection and measurements for Study 2 outlined in section 4.2.2.

105 4.1 INTRODUCTION

Fire has the potential to dramatically accelerate or reverse the process of tree decline. As discussed in Chapter 1, individual fires can affect tree health, influence competitive interactions between trees and alter tree population structure by killing particular species or size-classes within a species. Therefore, the management of fire, even a single fire, in tuart woodlands with canopy decline, requires careful consideration. Damage to trees from intense fire, particularly fires at short intervals, may be a factor in tuart canopy decline (Beard 1967, Pigott and Loneragen 1995, Fox and Curry 1981) and Chapter 3 provided some evidence to support this. Conversely, fires can also increase the growth of established eucalypt trees (Wallace 1966, Mazanec 1999, Bell and Chatto 2003) although not consistently (see Shea et al. 1981, Abbott and Loneragen 1983, Guinto et al. 1999). Where a growth stimulus has occurred changes in soil nutrient availability (Abbott and Loneragen 1983) and/or possibly soil microbiology (Florence 1981, Ellis and Pennington 1992) are suspected to be important. Given the possibility of both positive and negative effects of fire, there is some uncertainty about the net effect of individual fires on tuart tree health. However, as noted in Chapter 1, factors such as fire intensity, tree size and initial tree health are likely to be influential.

The increasing prominence of peppermint could be a factor in the declining health of tuart trees and a limitation to tuart recruitment (Chapters 2 and 3). Greater knowledge of the effects of fire on established seedlings, saplings and trees of tuart in relation to other tree species in tuart woodland could enable the manipulation of woodland structure by fire to favour tuart health and recruitment. For eucalypts that resprout from lignotubers, mortality is rare even following high-intensity fire (Chapter 1). However, with the exception of a report that three-year-old tuart seedlings resprouted following a wildfire (Forests Department of Western Australia 1979), few if any observations on the fire response of tuart or peppermint have been published.

The hypothesis that an individual fire affects tuart health beyond the first growth season after the fire was addressed in this Chapter. The response of tuart to fire was compared with that for co-occurring midstorey tree species, particularly peppermint. One study tested whether the health of tuart trees changed in the year following exposure to low intensity fire. Another assessed whether exposure to low intensity fire influenced the physical and chemical properties of leaves of tuart and peppermint trees. Seedlings,

106 saplings and trees of both tuart and peppermint were also compared for bark thickness, resprouting ability and post-fire recovery. The studies of post-fire recovery extended to other co-occurring midstorey species to provide a further reference for the responses revealed for tuart and peppermint. Finally, the rate of canopy recovery of tuart trees subjected to full canopy scorch was monitored to evaluate the direct role of fire in tuart canopy decline. Further background and the specific hypotheses for the experiments are outlined in the following section.

4.2 METHODS

4.2.1 General

Sites Unless otherwise stated, sampling occurred in the plots or along the transects as specified in sections 3.2.3 and 3.2.4 (Chapter 3). Details for sites Y1 to Y9, Ya1 and Ya2 and GB were outlined in section 3.2 (Chapter 3).

Measurements Methodology for ratings of canopy health (CH) and canopy area (CA) were detailed in section 3.2.4 (Chapter 3). Only those trees with a diameter at breast height over bark (dbhob) > 5 cm were rated. This was due to uncertainty about the ability to accurately rate small individuals, or compare the ratings of very small and large individuals. Canopy scorch of a tree following fire was estimated visually as a percentage of the total canopy area of the tree.

Potting mix For studies where seedlings were grown under glasshouse conditions, the potting mix comprised coarse sand, composted pine bark and “coco peat” in a 2:2:1 mix (Richgro Garden Products, Perth) plus 4 g of dolomite, 4 g of calcium carbonate and 2 g of a slow release fertiliser (Osmocote Exact®, Scotts International, Netherlands: nitrogen 16 %, phosphorus 5 % and potassium 9.2 % plus trace elements) in each 9 L bucket of potting mix prepared.

Statistical analyses MINITAB Release 13 (Minitab Inc. 2000) was used for all statistical calculations and tests unless otherwise stated. Data for t-tests and models for linear regression, ANOVA

107 and MANOVA, were checked for assumptions of independence, constant variance and normality; transformations were performed where necessary. Logistic regression was used where survival or death was the dependent variable and Pearson, Deviance and Hosmer-Lemeshow Goodness of fit test results as provided by the MINITAB output were checked. Significance (α) was set to 0.05 unless otherwise stated. However, for multiple tests within data sets, α was modified in accordance with the Sequential Bonferroni method (Holm 1979 in Quinn and Keough 2002) where appropriate.

Changes in CH and CA ratings over time were compared using paired-sign tests in STATVIEW Version 5.0 (SAS Institute Inc. 1998). These were used in preference to paired t-tests because some data sets could not be normalized. Medians and interquartile ranges, rather than means and standard errors, were reported for these data. This reflected both the non-normality of the data, as well as the fact that the data were somewhat crude, being derived from visual estimation of a rating scale.

4.2.2 Indirect effect of fire on established tuart and peppermint trees.

Background Two studies were conducted to determine the indirect effect of tree canopies to fire. By selecting trees within burnt areas with no or only minor canopy scorch, the direct impact of fire on trees was thought to have been eliminated or reduced. Study 1 assessed the changes in the canopy size and health of tuart trees in a burnt and unburnt area over time. Study 2 involved the measurement of leaf-scale differences between tuart and peppermint trees in burnt and unburnt areas.

Study 1: canopy size and health of tuart trees

Hypothesis:

• Canopy area and canopy health ratings are higher for minimally-scorched tuart trees in the year following fire.

Sites The study occurred at sites Y1 (burnt site), Y2 and Y3 (unburnt sites) in Yalgorup National Park.

108 Sampling design One recently burnt area (site Y1) with 22 sub-samples (trees) and one unburnt site (sites Y2 and Y3 pooled) with 26 subsamples were available. Trees with canopy scorch, but with ≤ 10 % of the canopy scorched, comprised the sub-samples in the burnt area. Trees with some scorch were selected to indicate that burning had occurred within their vicinity; the 10 % level was arbitrarily chosen as an upper-limit for minimal direct damage.

Measurements Pre-fire ratings of CH and CA, and measurements of dbhob, were made from March to May 2004. The fire occurred at site Y1 in December 2004 and estimations of canopy scorch were recorded for each tree later in that month. Post-fire ratings of CA were made in January 2006.

Statistical analysis Paired-sign tests for pre-fire and post-fire measurements of CH and CA were calculated. Tests were performed separately for the burnt and unburnt samples. In testing the comparability of burnt and unburnt samples, medians for dbhob were compared using the Mann-Whitney test as the data sets could not be normalized.

Study 2: leaf attributes of tuart and peppermint trees

Hypotheses:

• The physical and chemical properties of tuart and peppermint leaves are different between burnt and unburnt patches. • There is an interaction between species and patch (burnt or unburnt) for the physical and chemical properties of leaves.

Sites The study occurred at sites Y1 (burnt site) and site Y3 (unburnt site) at Yalgorup, but was not confined to within the permanent plots.

Sampling design There were two factors: species and patch type. The species were tuart and peppermint and the patch-type factor comprised three alternatives: unburnt patches within site Y1, burnt patches within site Y1 and patches within site Y3 (unburnt site). There were eight replicates trees per species per patch. From each tree, 10 youngest-fully-expanded 109 leaves (sub-samples) were collected. Leaves from from outer, mid-height canopy positions were selected in order to standardize for leaf type between replicates.

Sampling methodology Replicates were selected by searching for suitable tree patch-type combinations. Suitable combinations had burnt or unburnt ground present within at least the projected canopy area of a tree, and trees between 3 and 15 m tall, in at least moderate health, displaying no evidence of post-fire resprouting (assumed to be indicative of no or minimal direct fire damage). Relatively small, healthy trees were chosen as they were thought to be more able to respond rapidly to environmental changes than larger or less- healthy trees. That the fire was low intensity fortuitously provided burnt and unburnt patches at the site, as well as small trees without scorched canopies within burnt patches.

Sample processing and measurements Sampling occurred in December 2005, 12 months after the fire. Leaf thickness was measured immediately upon collection using digital calipers aligned parallel to the mid- vein and positioned mid-way between the mid-vein and leaf edge. Leaves were then sealed in plastic bags and stored on ice. On return from the field, they were refrigerated. Within 48 hours they were scanned and leaf area measured using ASSESS software (American Phytopathological Society 2002). After drying at 70°C for 48 hours, they were weighed and specific leaf area (SLA) was calculated. The dried material was then finely ground in a ball mill in preparation for the determination of nitrogen (N) concentration, 15N:14N and 13C:12C. Analysis was carried out at Edith Cowan University (Joondalup) using a continuous flow mass spectrometer (model 20:20, PDZ Europa Crewe, UK). The isotopic ratios, represented in relation to known standards, were calculated as:

15 13 δ N or δ C (‰) = Rsample/Rstandard – 1) x 1000

15 14 13 12 where, Rsample and Rstandard are the N: N or C: C ratios of the leaf sample and the standards, respectively.

Statistical analysis Data from the unburnt site (Y3) were not compared statistically to the data from the burnt site (Y1) owing to the distinct separation of the sites in space. In contrast, the 110 samples for the burnt and unburnt patches within site Y1 were compared as they were relatively well dispersed. MANOVA was performed initially. P values from Pillai’s trace were reported given that this measure is less affected by departures from normality and constant variance compared to the others (Townend 2002). Nevertheless, as serious departures from model assumptions could not be rectified, adjustment of α to 0.01 was made to reduce the probability of Type 1 errors (Tabachnik and Fidell 1996). Following a significant result, ANOVA was then performed for each response variable separately.

4.2.3 Comparison of the anatomical features affecting the capacity of tuart and peppermint to survive fire and recover following fire

Background Three studies were carried out to compare the bark thickness and the location of buds (above-ground or below-ground) for resprouting in tuart and peppermint. Differences in the responses of the species to fire (survival and resprouting origin) are likely to be related to species differences in bark thickness and bud location (Gill 1975). Glasshouse seedlings were measured for bark thickness in Study 1, while saplings and trees growing at the study sites were measured in Study 2. In Study 3, the stems of seedlings in the glasshouse were decapitated to observe resprouting patterns and thereby infer the location of buds.

Study 1: bark thickness of seedlings grown under glasshouse conditions

Hypothesis:

• For seedlings, bark thickness differs between tuart and peppermint.

Methods Tuart (n = 5) and peppermint (n = 5) seedlings were gown in potting mix under glasshouse conditions (evaporatively cooled). The seedlings were measured for bark thickness at approximately two years of age. Measurements of stem diameter were made with vernier calipers at 5 cm and 20 cm above the root–stem junction. The stem was then remeasured after the bark was peeled away. Half of the difference between the two measurements was recorded as the bark thickness. Stem diameter and bark thickness data were compared between the species using the t-test.

111 Study 2: bark thickness of saplings and trees

Hypothesis

• For saplings and trees, bark thickness differs between tuart and peppermint.

Sites Trees and saplings were sampled at sites Y1, Y6, Y7 and Y9 in Yalgorup.

Sampling The trees and saplings comprised a selection of those from the transect surveys at sites Y6, Y7 and Y9 (sections 3.2.1 and 3.2.4 in Chapter 3) and from Study 2 at site Y1 as outlined in section 4.2.2. Individuals selected were > 1.5 m tall, had one stem from ground-level to breast height and were at least moderately healthy in appearance. At site Y1, only those trees with bark not burnt in the 2004 fire were sampled. Across the sites, there were 22 tuarts and 17 peppermints sampled between December 2005 and March 2006.

Measurements A small, flat-blade screwdiver was used to penetrate the bark and thickness was determined by the depth of penetration under firm hand pressure. Two measurements were made on opposite sides of each stem at both breast height and at ground-level. The average depth of each pair of measurements was recorded for both heights.

Statistical analyses Comparison between the species of both the slopes and the intercepts of linear relationships for bark thickness against stem diameter was achieved by incorporating the two data sets into a multiple regression (Townend 2002). This was carried out for both the basal and the breast height data separately.

Study 3: the location of buds on seedlings raised in the glasshouse

Hypothesis:

• Two-year-old tuart and peppermint seedlings resprout when the stems are cut within 1 cm of the root-stem junction.

112 Methods The stems of tuart (n = 8) and peppermint (n = 8) seedlings germinated and raised in potting mix under glasshouse conditions (evaporatively cooled) for approximately two years were cut off at 1 cm above the root-stem junction. The presence or absence of resprouts was observed over two months after cutting. During this period temperatures ranged from 11°C to 40°C and water was supplied twice daily by sprinkler irrigation.

4.2.4 Observations of survival and resprouting response in tuart, peppermint and co-occurring tree species following complete canopy scorch.

Background Three studies were conducted to observe patterns of survival and resprouting position (crown or below) in tuart, peppermint and co-occurring tree species following fires. Study 1 focused on the survival rates for seedlings and small saplings (that is, individuals ≤ 1 cm dbhob), Study 2 examined the survival rates of saplings and trees (that is, individuals > 1 cm dbhob) and Study 3 compared the ability of the species to resprout from crown branches.

General

Sites Six burnt sites covering four separate fires were examined. These sites were Y1, Ya1, Ya2, Ya3, Ya4 and GB (Table 4.1). Plots established prior to planned burns (Y1 and Ya1) were on the whole subjected to only low-intensity fire which resulted in minor canopy scorch. Some measurements were therefore taken opportunistically following unplanned fires (GB, Ya2, Ya3 and Ya4). At GB and Ya2, sampling occurred in plots (see section 3.2.3 in Chapter 3) whereas at sites Ya3 and Ya4 trees were sampled along transect belts that were approximately 20 m wide (locations specified in Appendix 5). In addition, further trees and saplings were selected in two small areas (< 0.5 ha) in zones of higher fire intensity (as indicated by greater scorch heights) outside the main plots within site Y1 (locations specified in Appendix 5).

Control sites (unburnt sites) were monitored during the study period at site Y2 and Y3 (adjacent to Y1) only. A control site (Ya2) was established for the cluster of sites at Yanchep, but was burnt in a subsequent unplanned fire (back-burn for wildfire control).

113 For site GB, there were no nearby unburnt areas that could have acted as controls. Some control-like comparisons were achieved by comparing trees with fully-burnt canopies against trees with partially scorched or unburnt canopies within burnt areas. Arguably, the interspersion of these “controls” within the “treatment” area was preferable to having controls segregated at different sites.

Table 4.1: Fire and burn description for sites burnt during the study period. Ave. est. intensity Max. scorch Burnt area Site Date (kW m-1)1. ht (m)2. (%) 3. Comments Y1 12/2004 < 350 10 60 Prescribed fire Ya1 10 - 11/2004 < 350 12 50 Prescribed fire. Lit several times Ya2 01/2005 < 350 6.5 90 Back burn at night for wildfire control4. Ya3 01/2005 2000 - 4000 > 20 100 Wildfire Ya4 01/2005 3000 - 7000 > 20 100 Wildfire GB 02/2003 350 - 1000 < 15 95 Back burn for wildfire control5. 1. Estimated according to scorch height or observation in the case of Y1 and Ya2 and with reference to Burrows (1984) and Cheney (1981). 2. Measured within plots at Y1, Ya1 and Ya2, estimated at GB, Ya3 and Ya4. 3. Based on visual estimations. At sites Ya1, Ya2 and Y1, multiple estimates were made across plots to derive a mean. 4. J. Wheeler, Department of Conservation and Land Management, pers. comm. 5. G. Napier, Department of Conservation and Land Management, pers. comm.

Sampling Across all the sites, the species monitored were tuart, peppermint, Allocasuarina fraseriana, Banksia grandis, Banksia attenuata and Melaleuca raphiophylla. Site and species details are summarized in Table 4.2. All individuals (except germinants) of tree species present within plots at sites Y1, Y2, Y3 and GB were selected for assessment and monitoring. At site Y1, Ya1, Ya2 (burnt sites), Y2 and Y3 (unburnt sites) individuals were tagged prior to the most recent fire. At Ya1 and Ya2 (burnt sites), not all individuals were tagged within plots when densities were high. In these instances, an individual from each species closest to the northwestern corner of a 5 x 5 m sector of the main plot, was selected, tagged and remonitored after the fire. Assessments at sites GB, Ya3 and Ya4 began from immediately after the most recent fire. The additional sampling of tuarts at site Y1 also commenced following the 2004 fire. The primary aim of this additional sampling at sites Y1, Ya3 and Ya4, was to examine the response of large tuart trees to 100 % canopy scorch. There was little difficulty in identifying the

114 species for scorched or defoliated stem structures. It was assumed that all stems standing after the fires were living prior to the fires. However, at site GB where individuals across all sizes were sampled, seedlings that were incinerated would not have been included.

Measurements The proportion of foliage scorched in the canopy was estimated for individual trees, seedlings and saplings at all burnt sites. An emphasis on instances of 100 % scorch was given to the presentation of results on survival and resprouting patterns. This was because 100 % scorch is recommended as the standard minimum condition for assessing plant species response to fire (Gill 1981c). In some cases, particularly for seedlings, foliage was not only scorched, but also incinerated, sometimes completely. Therefore, technically the 100 % scorch category refered to the minimum level of plant damage. Survival was recorded as the presence of living resprouts at the final assessment times (7 to 12 months after the fire event). A minority of tagged plants (mostly seedlings) could not be located after the fires. While in most cases the role of fire in the loss of individuals could be deduced, on some occasions this could not be determined with certainty. Where there was uncertainty the data were excluded. The patchy nature of prescribed burns together with the necessity to be opportunistic in the event of unplanned fires meant that there were some inconsistencies in methodology and sampling times between sites. Details of the measurements taken during the surveys are summarised in Table 4.2.

Analyses Statistical analyses comparing species responses were not conducted as there were small sample sizes for some species at some sites, differences in the size distributions of individuals between species, variations in fire severity within sites and as most species (except tuart) were found at some sites but not others.

115 Table 4.2: Summary of measurements (trees, saplings and seedlings) by site and species (Spp) for survey of the impact of fire on survival and recovery. dbhob = diameter at breast height over bark. CA = canopy area rating, CH = canopy health rating, n/a = not applicable or not available. Date of canopy scorch assessment Pre-fire measurements and other post-fire measurements Final measurements Site Spp1. n2. Date Date Other measurements Date Comments Y1 Eg 105 03-05/04 dbhob, CA, CH, height3. 12/04 Burn severity at base4. 01/06 Survival, CA, resprout origins Plants tagged pre-fire Af 105 dbhob, height3. Burn severity at base4. Survival, resprout origins Alf 46 Estimated height/dbhob n/a Survival Y1 Eg 22 n/a n/a 01/05 dbhob, CA, CH 01/06 Survival, CA, resprout origins 100% scorch plants selected post-fire Y2 Eg 28 03-05/04 dbhob, CA, CH, height3. n/a n/a 01/06 Survival, CA Data from these two unburnt sites were & Af 138 dbhob, height3. n/a Survival pooled Y3 Alf 21 Estimated height/dbhob n/a Survival GB Eg 430 n/a n/a 02/03 dbhob 09/03 Survival, resprout origins Plants tagged post-fire Alf 10 n/a Estimated height/dbhob Survival Ba 160 n/a dbhob Survival, resprout origins Bg 55 n/a dbhob Survival, resprout origins Mr 81 n/a dbhob Survival, resprout origins Ya1 Eg 63 04/03 & dbhob, CA, CH, height3. 11/04 Burn severity at base4. 01/06 Survival, CA, resprout origins Plants tagged pre-fire Bg 34 09/04 dbhob Burn severity at base4. Survival, resprout origins Ya2 Eg 33 04/03 & dbhob CA, CH, height3. 01/05 Burn severity at base4. 01/06 Survival, CA, resprout origins Plants tagged pre-fire Bg 19 09/04 dbhob Burn severity at base4. Survival, resprout origins Ya3 Eg 18 n/a n/a 02/05 dhob, CA, CH 01/06 Survival, CA, resprout origins 100% scorch plants selected post-fire Ya4 Eg 18 n/a n/a 02/05 dhob, CA, CH 01/06 Survival, CA, resprout origins 100% scorch plants selected post-fire 1. Agonis flexuosa (Af), Allocasuarina fraseriana (Alf), Banksia attenuata (Ba), Bg = Banksia grandis (Bg), Eucalyptus gomphocephala (Eg) Melaleuca raphiophylla. (Mr).

116 2. This number represents all plants that were assessed. For burnt sites, not all plants were necessarily scorched completely or at all. 3. Height was only measured for individuals < 5 m tall. 4. Within 0.5 m radius of the base of tuarts and peppermints < 5 m tall. Severity was rated as unburnt, partially burnt, completely burnt (little or no ash) or burnt with ash formation.

117 Study 1: survival of seedlings and small saplings

Hypothesis:

• When subjected to 100 % canopy scorch, the majority of seedlings and small saplings of tuart and co-occurring tree species survive.

Methods Survival of all individuals ≤ 1 cm dbhob with 100 % canopy scorch across the various sites was assessed. Sample sizes were 77 (tuart), 33 (peppermint), 2 (A. fraseriana), 9 (B. grandis), 65 (B. attenuata) and 66 (Melaleuca raphiophylla). For tuart ≤ 1 cm dbhob at sites Y1, Ya1 and Ya2 and peppermint at site Y1, survival rates were compared on the basis of local fire severity and seedling height. Heights were measured either prior to the fire or immediately after the fire. Fire severity around the base of the stem (approximately 0.5 m radius) was rated immediately after the fire as unburnt, partially burnt, completely burnt with little or no ash, or completely burnt with ash formation. There were 17 tuart and 33 peppermint seedlings examined in this way. The growth rings on cut stems of a small sample of tuart (n = 4) and peppermint seedlings (n = 5) at site Y1 and Y3, and tuart seedlings and saplings (n = 14) at site GB, were also counted at the final measurement time in order to estimate stem age (see Table 4.2 for sampling dates and species-site combinations).

Study 2: survival of saplings and trees

Hypothesis

• When subjected to 100 % canopy scorch, the majority of saplings and trees of tuart and co-occurring tree species survive.

Methods Survival of all individuals > 1 cm dbhob that were subjected to at least 100 % foliage scorch was assessed for all sites listed in Table 4.2. Survival rates were calculated for each species by dbhob class. However, A. fraseriana saplings and trees were not categorized into dbhob classes because of the great irregularities in stem shape arising from previous fire damage. Sample sizes by species for individuals > 1cm dbhob were 377 (tuart), 18 (peppermint), 78 (B. attenuata), 51 (B. grandis), 15 (M. raphiophylla) and 29 (A. fraseriana). A further number of incompletely-scorched tuarts (n = 125 across sites Y1, Ya1, Ya2 and GB) and peppermints (n = 39 at site Y1) along with 118 unscorched tuarts (n = 107 across sites Y1, Y2, Y3, Ya1 and Ya2) and peppermints (n = 162 across sites Y1, Y2 and Y3) were monitored for comparative purposes.

Study 3: ability to resprout from the crown following 100 % canopy scorch

Hypothesis:

• When subjected to 100 % canopy scorch, surviving saplings and trees of tuart and co-occurring tree species resprout from the crown.

Methods The presence or absence of crown-branch resprouts was noted for all completely scorched, surviving saplings and trees (dbhob > 1 cm) at the final post-fire assessment time (Table 4.2). Sample sizes were 349 (tuart), 18 (peppermint), 70 (B. attenuata), 29 (B. grandis) and 7 (M. raphiophylla). A. fraseriana were not assessed. Data were presented for each species on the basis of dbhob class.

4.2.5 Comparative early canopy recovery of tuart and peppermint seedlings subjected to decapitation

Background A glasshouse experiment was conducted to compare the rate of recovery of the foliage of tuart and peppermint seedlings following decapitation. This experiment also provided data that complimented the previous study comparing the location of buds between the species (Study 3, section 4.2.3).

Hypotheses

• Two-year-old tuart and peppermint seedlings resprout when stems are decapitated within 5 cm of the root-stem junction. • Following decapitation, two-year-old tuart and peppermint seedlings refoliate at different rates

Experimental design The experiment was arranged in a randomized block design. There were four blocks and one factor (species) consisting of either tuart or peppermint. The blocks covered different positions on the glasshouse bench. There was one seedling of each species (replicate) per block. Thus, n = 4 for each species.

119 Growth conditions The seedlings had been raised in potting mix under glasshouse conditions (evaporatively cooled) for approximately 21 months before being repotted into large containers (approximately 25 000 cm3 in volume and 22 cm deep) filled with yellow sand. Two seedlings of the same species were placed in each container with one seedling per container randomly selected for the decapitation treatment; the other seedling in the container was not used in the experiment. Stems were cut off at approximately 5 cm above the root-stem junction four weeks after the seedlings were potted into the containers. A 500 ml quantity of 2 g L-1 fertilizer solution (Thrive®, Arthur Yates and Co. Ltd. Australia: 27 % nitrogen, 5.5 % phosphorus, 9.0 % potassium plus trace elements) was added to each container at the time the experiment was set up and again at two and four weeks later. During the course of the experiement, glasshouse temperatures ranged from 11°C to 40°C and water was supplied twice daily from drippers.

Measurements and calculations Leaves were weighed at the time of the initial cut and the resprouted leaves were weighed 57 days following cutting. Weights were determined after the leaves were dried at 70°C for 48 hours. For each seedling, the weight of the resprout leaves was converted to a proportion of the weight of the leaves at the initial cut.

Statistical analyses The mean weights of the resprouting leaves as a proportion of the weight of the leaves at the time of the stem-decapitation treatment were compared between the species by ANOVA.

4.2.6 Canopy area (CA) recovery of tuart trees following complete canopy scorch

Background The rate of canopy recovery following fire was assessed for tuart trees as the restoration of pre-fire canopy area is one method of assessing the impact of fire on plants (Gill 1997). Trees with lower canopy area ratings, indicative of poorer tree health, may be slower to recover after fire (Chapter 1) and tuart health is poor across large areas (Chapters 2 and 3). Hence, the importance of pre-fire canopy area on the post-fire canopy recovery of tuarts was also explored. Further, as tree size and the proportion of

120 canopy scorch also influence the rate of tree recovery after fire (Chapter 1), these factors were also considered in the study.

Hypotheses:

• Tuart trees recover their pre-fire canopy area within the year following 100 % canopy scorch. • Recovery of canopy area is related to pre-fire canopy area and tree size.

Sites and sampling Trees with 100 % canopy scorch (n = 60), 11 to 90 % (n = 15) and no canopy scorch (n = 67) were sampled across sites Y1, Y2, Y3, Ya1, Ya3 and Ya4. The 11 to 90 % category was devised with a 10 % gap between the other categories (100 % and no scorch) to ensure a clear distinction between the three categories. All trees were > 5 cm dbhob.

Measurements All trees were rated for pre-fire CH and CA as well as CA at approximately one-year post-fire. Ratings were also made over the same period at the unburnt sites (Y2 and Y3). Only seven of the trees in the 100 % scorch category had been tagged and assessed before the fire. As the majority of trees were assessed immediately following the fire on scorched canopies, pre-fire CH and CA was not assessed directly. Thus, there was some uncertainty as to the pre-fire canopy condition in these cases. In particular, leaf incineration may have resulted in underestimates of CA for smaller trees, or trees with basal resprouts prior to the fires. Photographic examples of CH and CA ratings of scorched and unscorched canopies are presented in Appendix 1. One year after the fires, all trees were rated for CA, but not CH. This was because the quantity of epicormic foliage is negatively related to the CH score in the CH rating system, but when comparing trees with complete canopy scorch, healthier trees would necessarily be those with a greater quantity of epicormic foliage. Therefore, CA was thought to represent the best indicator of tree health after fire. Measurement dates by site are presented in Table 4.2.

Statistical analyses Changes in CA ratings over time were compared within each scorch category using paired-sign tests. For trees with 100% scorch, relationships between pre-fire CH and CA

121 ratings, dbhob and post-fire CA rating were tested using multiple linear regression. The “best subsets” method was employed for initial variable selection (Townend 2002).

4.3 RESULTS

4.3.1 Indirect effect of fire on established tuart and peppermint trees

Study 1: canopy size and health of tuart trees The canopy health of tuarts exposed to fire that scorched ≤ 10 % of the canopy foliage, was higher one year post-fire at site Y1 (Figure 4.1). Comparison of pre and post-fire ratings by the paired-sign test were highly significant for CA (P = 0.008) and significant for CH (P = 0.049). While not representative of true controls due to their separation in space, the median CA and CH ratings of trees in adjacent unburnt areas (sites Y2 and Y3) also increased over the same time period, but not significantly (P = 0.189 for both data sets). As the sample size for the unburnt trees was slightly larger, the comparative lack of statistical significance was not attributable to a difference in power. However, the median canopy health rating was initially lower for the unburnt trees (Figure 4.1b). The median dbhob of unburnt trees was also less than that for the burnt trees (18.9 and 26.5 cm, respectively) but significant differences were not revealed from a Mann- Whitney test: W = 565.5, P = 0.142.

122 12 7 b b g 6 a 10 a a a 5 a a 8 4 6 3 4 2 2

Median canopyrating area 1 Median canopy health ratin health canopy Median 0 0 burnt unburnt burnt unburnt Site Site a) b)

Figure 4.1: Median pre-fire (■) and one-year post-fire (□) ratings of the canopy condition of tuart trees (> 5 cm diameter at breast height over bark) after slight (≤10 %) canopy scorch (site Y1) or located in unburnt areas (sites Y2 and Y3). Health ratings are for a) canopy area and b) canopy health, both based on Grimes (1978) with modifications. For the burnt trees, n = 22 and the unburnt trees, n = 26. Within each data set (burnt or unburnt) samples with different letters are significantly different following paired-sign tests. With Bonferroni corrections applied across charts but within site type, differences remain significant. Interquartile range bars are presented.

123

0.4 110 2.4 ) 0.35 100

) 90 2 0.3 -1 80 0.25 70 1.6 60 0.2 1.2 50 0.15 40 0.8 0.1 30 g (cm SLA Mean 20 0.4 Mean thickness leaf (mm 0.05 10

0 0 concentration MeanN leaf (%) 0 Burnt patch Unburnt patch Unburnt site Burnt patch Unburnt patch Unburnt site Burnt patch Unburnt patch Unburnt site

Location Location Location a) b) c) Location Location Burnt patch Unburnt patch Unburnt site Burnt patch Unburnt patch Unburnt site

0 -24 -0.4 -25 -0.8 N (‰)N (‰) C -26 15 -1.2 13 δ δ -1.6 -27 -2

Mean leaf Mean leaf Mean leaf -28 -2.4 -2.8 -29

d) e) Figure 4.2: Selected attributes of youngest-fully-expanded leaves from tuart (■) and peppermint (□) trees in burnt and unburnt patches at site Y1 and an adjacent unburnt site (Y3). Attributes were a) thickness, b) specific area (SLA), c) N concentration, d) δ15N and e) δ13C. For each species in each location, n = 8. Trees were between 3 and 15 m tall. MANOVA/ANOVA indicated significant difference between the species in all cases, and between locations for d. No significant species – location interactions were detected (Tables 4.2 and 4.3). The unburnt site was not included in the analyses. Standard error bars are indicated.

124 Study 2: leaf attributes of tuart and peppermint trees Within the species, selected attributes of the youngest-fully-expanded leaves were little different between small to medium-sized unscorched trees, regardless of whether they were located in burnt or unburnt patches at site Y1 (Figure 4.2). One exception was for δ15N. In the leaves of trees in burnt patches, N was less depleted in the 15N isotope and the difference was significant (Table 4.4). Means for δ15N at the unburnt site (Y3) were also more similar to the means for the unburnt patches than the burnt patches at site Y1.

When the species were compared, overall differences for the attributes were apparent (Figure 4.2) and significant (Tables 4.3 and 4.4); tuart leaves were thicker, had a higher nitrogen content and a higher δ13C, but a lower specific leaf area and a lower δ15N than peppermint leaves. However, no significant species patch-type interaction was evident when all the results were assessed together (Table 4.3).

Table 4.3: MANOVA results for thickness, specific area, N concentration, δ15N and δ13C of youngest-fully-expanded leaves from small to medium tuart and peppermint trees (3 to 15 m tall) in burnt and unburnt patches at site Y1. Due to violations of model assumptions, α = 0.01. Significant values are denoted in bold. Effect P

Species (S) F(5, 24) = 41.148 < 0.001

Patch (P) F(5, 24) = 4.007 0.009

S x P F(5, 24) = 1.255 0.315

Table 4.4: ANOVA results for thickness (Box-Cox transformed), specific area (SLA [Box-Cox transformed]), N concentration (log transformed), δ15N and δ13C of youngest-fully-expanded leaves from small to medium tuart and peppermint trees (3 to 15 m tall) in burnt and unburnt patches at site Y1. Significant values denoted in bold where significant effects were also present in MANOVA results (Table 4.3). F Leaf thickness SLA % N δ15N Effect df1, df2 F P F P F P F P Species (S) 1, 28 35.77 <0.001 9.90 0.004 45.58 <0.001 56.04 <0.001 Patch (P) 1, 28 1.63 0.212 0.01 0.915 1.33 0.258 11.15 0.002 S x P 1, 28 2.61 0.125 0.21 0.652 0.34 0.563 5.26 0.030 F δ13C Effect df1, df2 F P Species (S) 1, 28 10.18 0.003 Patch (P) 1, 28 0.14 0.709 S x P 1, 28 0.57 0.458

125 4.3.2 Comparison of the anatomical features affecting the capacity of tuart and peppermint to survive fire and recover following fire

Study 1: bark thickness of seedlings grown under glasshouse conditions Within a small sample of two-year-old seedlings raised under glasshouse conditions, the stem bark thickness of tuart was greater than for peppermint, although not significantly different (Table 4.5). The seedlings were similar in size with stem diameters 5 cm above the roots of 1.35 cm ± 0.13 SE for tuart and 1.31 ± 0.29 SE for peppermint (t(8) = 0.21, P = 0.836).

Table 4.5: Mean bark thickness (mm) of two-year-old tuart (n = 5) and peppermint seedlings (n = 5) raised under glasshouse conditions. Height above Bark thickness (mm) root-stem Tuart Peppermint junction mean ± SE mean ± SE t(8) P 5 cm 2.1 ± 0.38 1.6 ± 0.29 0.96 0.366 20 cm 1.2 ± 0.17 0.9 ± 0.16 1.04 0.327

Study 2: bark thickness of saplings and trees Tuarts generally had thicker bark than peppermint when bark thicknesses on a range of stems sizes (from 1 to 30 cm dbhob) were compared among naturally occurring individuals (Figure 4.3). Comparison of the regression lines between dbhob and bark thickness at the stem base revealed a significant difference between the species across the 1 to 30 cm dbhob range for the intercepts, but not for the slopes (Table 4.6). In contrast, comparison of the regression lines for bark thickness at breast height revealed a significant difference between the species across the 1 to 30 cm dbhob for the slopes, but not the intercepts (Table 4.6). This difference represented the greater increment in bark thickness at breast height per unit dbhob for tuart when compared to peppermint (Table 4.7). Separate regressions for bark thickness at breast height or at the stem base against dbhob (< 30 cm) were significant for tuart and peppermint (Table 4.7). From the few samples of bark thickness measured for tuarts with dbhob > 30 cm, a general leveling off of the dbhob bark-thickness relationship, along with substantial variability, was observed (Figure 4.3a).

126 Table 4.6: Comparison of the bark-thickness diameter at breast height over bark (dbhob) linear relationships between tuart and peppermint. Bark thickness (mm) was square-root transformed. dbhob > 1 < 30 cm were compared (Figure 4.3). Significant values are denoted in bold following adjustment of α using the Sequential Bonferroni Method. Regressions were 2 significant: for breast height, F(3, 29) = 68.02, P = < 0.001, r = 0.876, and for the base F(3, 29) = 69.94, P = < 0.001, r2 = 0.879. Bark thickness at breast ht. Bark thickness at base Effect Coeff. t P Coeff. t P Constant 1.642 6.30 < 0.001 3.588 13.46 < 0.001 dbhob 0.150 7.93 < 0.001 0.075 3.87 0.001 Species (S) -0.269 -0.88 0.384 -1.484 -4.76 < 0.001 dbhob x S -0.061 -2.53 0.017 -0.011 -0.43 0.668

40 ) 35 30 25 20 15 10 5 Bark thickness (mm Bark ht. at breast 0 0 102030405060708090100 dbhob (cm) a)

35

) 30

25

20

15

10

Bark thickness at base (mm base at thickness Bark 5

0 0 102030405060708090100 dbhob (cm) b)

Figure 4.3: Relationship between diameter at breast height over bark (dbhob) and bark thickness at a) breast height and b) the stem base, for tuart (■, n = 22) and peppermint (□, n = 17) at sites Y1, Y6, Y7 and Y9. Vertical dashed line represents maximum of range of values for which regression lines for species were determined and compared (Table 4.6). Regression equations are presented in Table 4.7. 127 Table 4.7: Regression equations for the relationship between bark thickness and diameter at breast height over bark (dbhob) for tuart and peppermint with a dbhob > 1 < 30 cm at sites Y1, Y6, Y7 and Y9. Species Bark thickness at Equation Df P r2 tuart breast height = (0.164 + 0.015*dbhob)2 15 < 0.001 0.791 peppermint breast height = (0.137 + 0.009*dbhob)2 16 <0.001 0.746 tuart base = 1.28 + 0.064*dbhob 15 0.005 0.435 peppermint base = 0.43 + 0.004*dbhob 16 <0.001 0.683

Study 3: the location of buds on seedlings raised in the glasshouse When the stems of eight two-year-old seedlings each of tuart and peppermint were decapitated at 1 cm above the root-stem junction, all the tuarts failed to resprout whereas seven of the peppermints resprouted. Further studies of tuart and peppermint seedling resprouting behaviour following stem decapitation higher on the stem are reported in Section 4.3.4 and Appendix 3.

4.3.3 Observations of survival and resprouting response in tuart, peppermint and co-occurring tree species following complete canopy scorch

Study 1: survival of seedlings and small saplings The majority of tuart (65 %, n = 17), and a considerable proportion of peppermint (46 %, n = 33) seedlings and small saplings survived following exposure to fire that scorched or incinerated all foliage. Seedlings and small saplings of both species (< 1.0 m tall) did not survive fire where combustion of coarse fuels (as indicated by ash) occurred near stems.

Examination of the stem growth rings of four tuarts ranging from < 1 cm to 1.7 cm dbhob at site Y1 suggested a proportion of the seedlings/small saplings were between four and six years old. Given the time of the last fire at the site was much earlier (14 years prior) these seedlings appear to have established in an interfire period. The growth rings on five peppermint stems from < 1 to 1.05 cm dbhob were indicative of stem ages between five and eight years. However, the plant age of these peppermints could not be estimated because of the lignotuberous growth form.

When observations from site GB were included for tuart, the relative survival rate for seedlings subject to complete foliage damage from fire was clearly greater at 75 % than for peppermint with only 45 % survival (Table 4.8). However, whether the two figures 128 can be compared is questionable due to the different sites sampled. Additionally, as a large proportion of the tuarts sampled at GB were not tagged pre-fire, the possibility existed that some fire-killed individuals were not assessed because of incineration or severe damage making the recognition of species impossible. Examination of growth rings within 14 tuart stems ranging from < 1cm to 7.6 cm dbhob at this site suggested ages of between six and eight years.

3.5

3 . 2.5

2

1.5

1 Seedling height (m) height Seedling 0.5

0 01234Unburnt Part-burnt Burnt Ash Fire severity around base of seedling b)

3.5

3 . 2.5

2

1.5

1 Seedling height (m) height Seedling 0.5

0 01234Unburnt Part-burnt Burnt Ash Fire severity around base of seedling a)

Figure 4.4: Height and survival of seedlings and small saplings of a) tuart (n = 17) and b) peppermint (n = 33) following exposure to fire of sufficient intensity to scorch 100 % of the foliage. □ = killed, ■ = living at one year after fire. Seedling/sapling diameter at breast height over bark was ≤ 1cm. Categories on x axis represent a burn severity index based on burn completeness and the presence of ash. All peppermint seedlings were from site Y1 whereas tuart seedlings were from sites Y1, Ya1 and Ya2.

129 Study 2: survival of saplings and trees Survival rates for tuart following complete foliage damage were high across all size- classes (Table 4.8). The few deaths in the larger size-classes (20.1 to 50.0 cm and 50.1 to 100 cm) occurred for trees in comparatively poor health. The CH ratings for these trees were 2.5, 3.5 and 5.0, whereas the median CH rating for the surviving trees in the two classes was 8.5 with 75 % of all individuals rating higher than 8.0.

There were no clear differences in the ability of tuart to survive relative to a range of co- occurring tree species (Table 4.8). The data for peppermint were relatively deficient in respect to sample numbers and the fact the observations could only be made at one site (Y1) where fire intensity was mostly low. No peppermint trees with a dbhob > 5.0 cm within the sampling plots at this site were subject to 100 % crown scorch.

Table 4.8: Percentage survival (% surv) and 95 % confidence intervals (95 % CI) for tree species in diameter at breast height over bark (dbhob) classes following fire in tuart woodland. Each individual suffered 100 % foliage scorch and was assessed for survival within 7 to 12 months after fire. Sites are listed in notes. Tuart1. Peppermint2. Banksia attenuata3. dbhob (cm) % surv (95 % CI) n % surv (95 % CI) n % surv 95 % CI N ≤ 1 75 (64 – 84) 77 45 (28 – 64) 33 100 (95 – 100) 65 1.1 – 5.0 94 (90 – 97) 224 100 (85 – 100) 18 92 (80 – 98) 49 5.1 – 10.0 98 (92 – 100) 89 n/a n/a n/a 82 (60 – 95) 22 10.1 – 20.0 100 (87 – 100) 22 n/a n/a n/a 75 (19 – 99) 4 20.1 – 50.0 96 (79 – 100) 24 n/a n/a n/a 100 (37 – 100) 3 50.1 – 100.0 88 (64 – 99) 17 n/a n/a n/a n/a n/a n/a 100 > 100 n/a 1 n/a n/a n/a n/a n/a n/a Banksia grandis4. Melaleuca raphiophylla3. Allocasuarina fraseriana5. dbhob (cm) % surv (95 % CI) n % surv (95 % CI) n % surv 95 % CI N ≤ 1 56 (22 – 86) 9 100 (96 – 100) 66 50 (13 – 99) 2 1.1 – 5.0 60 (26 - 88) 10 100 (74 – 100) 10 5.1 – 10.0 58 (37 - 77) 26 100 (37 – 100) 3 10.1 – 20.0 90 (55 - 100) 10 100 n/a 1 93 (77 – 99) 29 20.1 – 50.0 92 (55 – 100) 5 100 n/a 1 50.1 – 100.0 n/a n/a n/a n/a n/a n/a 100 > n/a n/a n/a n/a n/a n/a 1. Sites GB (predominantly), Y1, Ya3, Ya4, Ya1 and Ya2 (in order of contribution). 2. Site Y1. All individuals were tagged pre-fire and post-fire in contrast to the other species 3. Site GB 4. Sites GB, Ya1 and Ya2. 5. Sites Y1 and GB. dbhob was not presented as many boles were highly irregular in shape due to past fire damage. Heights ranged from approximately 2 to 10 m tall.

130 For tuart and peppermint, very few deaths occurred during the monitoring period for individuals that either were not scorched, or had < 100 % foliage scorch (Table 4.9). The one unburnt tuart that died was a small, suppressed sapling (dbhob = 3.5 cm, canopy health [CH] = 2.5) growing within a peppermint tree thicket. The stem growth rings within this sapling hinted that it was around ten years of age. The other tuart that died was a sapling (dbhob = 8.6 cm) that suffered only 20 % canopy scorch but was unhealthy (CH = 4 and 60 % of the foliage necrotic) and well shaded by other trees. The two unburnt peppermints that died were small seedlings, being 0.1 m and 1.0 m in height. The one peppermint that died following partial (20 %) canopy scorch had a dbhob of 6.7 cm; no distinguishing circumstances were noted in this case.

Table 4.9: Percentage survival (% surv) and 95 % confidence intervals (95 % CI) for unburnt and incompletely-scorched (< 100 % scorch) tuart and peppermint seedlings, saplings and trees monitored for a period > 7 months between 2003 and 2006. Sites are listed in notes. Tuart1. Peppermint2. % surv (95 % CI) n % surv (95 % CI) n unburnt 99 (95 - 100) 107 99 (96 – 100) 162 < 100 % scorch 99 (96 – 100) 125 97 (87 – 100) 39 1. Sites GB, Y1, Y2, Y3, Ya1 and Ya2. 2. Sites Y1, Y2 and Y3.

Study 3: ability to resprout from the crown following 100 % canopy scorch Tuart, peppermint and a selection of co-occurring tree species, all displayed an ability to resprout from the crown after complete canopy scorch (Figure 4.5). While insufficient sample numbers precluded a proper assessment of the crown-resprouting ability of M. raphiophylla and A. fraseriana, crown resprouting was observed in trees of these species following complete canopy scorch. In all the other species examined, crown resprouting occurred for stems as small as 1.05 to 3 cm dbhob.

There was a high probability for resprouts to form from the crown of tuart when the dbhob was > 6 cm (Figure 4.5a and b). However, some tuarts with a dbhob > 6 cm (11 %, n = 116) only resprouted along the bole below the crown. There was no relationship evident between pre-fire canopy health and the post-fire formation of crown resprouts according to logistic regression analysis on a subsample of completely-scorched tuarts (dbhob > 5 cm, n = 55) for which pre-fire canopy health was rated pre-fire or estimated post-fire: G(1) = 1.154, P = 0.283.

131

120 10 9 100 8 7 80 6 60 5 4

dbhob (cm) dbhob (cm) dbhob (cm) 40 3 2 20 1 0 0 0.5 1.5 0.5 Absent 1.5 Present Absent Present Crown resprouts Crown resprouts a) b)

25 10 9 8 20 7 6 15

5 4 10 dbhob (cm) (cm) dbhob dbhob (cm) (cm) dbhob 3

2 5 1 0 0 0.5 1.5 0.5 Absent 1.5 Present Present Crown resprouts Crown resprouts c) d)

25 25

20 20

15 15

10 10 dbhob (cm) dbhob (cm) dbhob (cm) dbhob (cm)

5 5

0 0 0.5 1.5 0.5 Absent 1.5 Present Absent Present Crown resprouts Crown resprouts e) f)

Figure 4.5: Presence of resprouting in the crown following 100 % canopy scorch of surviving saplings and trees with a diameter at breast height over bark (dbhob) >1 cm. Species observed were: a) tuart (n = 349), b) tuart from “a” with a dbhob from 1.1 to 10.0 cm (n = 291), c) peppermint (n = 18), d) Banksia attenuata (n = 70), e) Banksia grandis (n = 29) and f) Melaleuca raphiophylla (n = 7). See Table 4.8 for site details.

132 4.3.4 Comparative early canopy recovery of tuart and peppermint seedlings subjected to decapitation.

From a small sample of two-year-old seedlings raised under glasshouse conditions, the rate of recovery of foliage (by weight) two months after cutting the stems approximately 5 cm above the root-stem junction, was greater for tuart than for peppermint (Figure 4.6). This was despite the fact that the mean pre-cut foliage weight was markedly greater for the peppermint seedlings. A significant difference between the species’ rates of recovery was established (Table 4.10).

40 ) 35 30 25 20 15 10 5 Mean dry weight of foliage (g foliage of weight dry Mean 0 Tuart Peppermint Species

Figure 4.6: Pre-cut (■) and 57th-day post-cut (□) means for dry weight of two-year-old tuart (n = 4) and peppermint seedlings (n = 4) raised under glasshouse conditions. Stems were cut approximately 5 cm above the root-stem junction. A significant difference was detected between the species for proportional rate of recovery (Table 4.10). Standard error bars are indicated.

Table 4.10: ANOVA results for proportional recovery of foliar dry weight 57 days after cutting the stems of two-year-old tuart and peppermint seedlings raised under glasshouse conditions. Stems were cut approximately 5 cm above the root-stem junction. Significant values are denoted in bold. Effect P

Block F(3, 3) = 1.34 0.408

Species F(1, 3) = 33.79 0.010

133 4.3.5 Canopy area (CA) recovery of tuart trees following complete canopy scorch

Tuart CA ratings were generally lower one year after fire for trees subject to 100 % canopy scorch, but not for trees with 11 to 90 % canopy scorch or no canopy scorch (Figure 4.7). The results of paired-sign tests for before and after ratings were significant for the 100 % scorch sample (P = 0.004) but not for the 0 % (P = 0.489) or 11 to 90 % (P = 0.388) scorch-category samples. Comparison between canopy scorch categories was not carried out due to the differences in sites and sampling methods. However, the results in Table 4.11 suggest the data sets had similar mean dbhob, and that the 0 % and 100 % scorch-category trees were similar for pre-fire health. The sample set for the 11 to 90 % scorch category was solely from site Y1 and thus comparable to the results in Figure 4.1.

12

10

8 a a a 6 a a b

4

2 Median canopyrating area

0 0 11 - 90 100 Canopy scorch (%)

Figure 4.7: Median pre-fire (■) and one-year post-fire (□) canopy area rating (based on Grimes [1978] with modifications) of tuart trees (> 5 cm diameter at breast height) according to canopy scorch categories: 0 % (n = 67), 11 to 90 % (n = 15) and 100 % (n = 60). Further details of the sample sets are presented in Table 4.11. Within each canopy scorch category, sample sets with different letters are significantly different following paired-sign tests. Interquartile range bars are indicated.

Table 4.11: Summary of data for sample sets used to assess the relationship between percentage canopy scorch and recovery of canopy area (Figure 4.7). dbhob = diameter at breast height over bark. Canopy Mean dbhob Median pre- Interquartile scorch (%) (cm) ± SE fire health1. range n Sites 0 33 ± 3.5 7.5 5.5 - 9.0 67 Y1, Y2, Y3, Ya1, Ya2 11 – 90 30 ± 7.3 8.5 7.0 – 12.0 15 Y1 100 38 ± 3.3 9.0 7.6 – 10.5 60 Y1, Ya1, Ya3, Ya4 1. Rating of canopy health based on Grimes (1978) with modifications. 134 One year after fire that caused complete canopy scorch, an inverse relationship between dbhob and change in CA rating was apparent for tuart trees > 5 cm dbhob (Figure 4.8). A dbhob of 20 cm marked a point above which few trees rated higher for CA one-year post-fire. Both dbhob and pre-fire CA rating were significant factors in a linear regression model with post-fire CA rating as the dependent variable (Table 4.12). Both variables were negatively related to post-fire CA rating. The slope of the regression line was significant (F(2, 57) = 54.50, P < 0.001) and the two factors accounted for 33.9 % of the variation (R2 adjusted) in post-fire CA rating. Pre-fire CA rating accounted for the greatest proportion (59 %) of this variation.

r 6

4

2

0 after fire -2

-4

Change in canopy ratingone area yea -6 0 10 20 30 40 50 60 70 80 90 100 110 120 130 dbhob (cm)

Figure 4.8: Change in canopy area (pre-fire and one-year post-fire ratings based on Grimes [1978] with modifications) versus diameter at breast height over bark (dbhob) for trees (dbhob > 5 cm) with 100 % crown scorch. The points are differentiated for individuals selected and assessed prior to controlled burning (■, n = 8) and individuals selected and assessed immediately following controlled burning and wildfire (□, n = 52). Sites surveyed were Y1, Ya1, Ya3 and Ya4. dbhob was a significant explanatory variable in canopy area change (Table 4.12).

Table 4.12: Multiple linear regression coefficients for diameter at breast height over bark (dbhob) and estimated pre-fire canopy area rating (based on Grime [1978] with modifications) versus change in canopy area rating one year after fire. Subjects were tuart saplings and trees (> 5 cm dbhob) with 100 % canopy scorch from sites Y1, Ya1, Ya3 and Ya4. Regression was 2 significant: F(2, 57) = 54.50, P < 0.001 and r (adjusted) = 0.339. Variable Coefficient Proportion SE T P of r2 Constant 3.543 0.816 4.34 <0.001 n/a Pre-fire canopy area rating -0.634 0.145 -4.33 <0.001 0.59 dbhob -0.036 0.009 -4.03 <0.001 0.41

135 4.4 DISCUSSION

Tuart, peppermint and co-occurring tree species of tuart woodlands are clearly strong resprouters. High proportions (≥ 45 %) of individuals in small and large size-classes of all species were able to survive and recover following complete canopy scorch, often by resprouting from crown branches. Stem decapitation experiments with tuart and peppermint seedlings indicated that resprouting capacity can develop as early as two years of age. However, field observations indicated that peppermint seedlings and small saplings were more vulnerable to fire-caused mortality than tuart. Therefore, fire at high frequency (two to four years) may favour tuart over peppermint recruitment. For large tuart trees with complete canopy scorch, the observed deaths or slow rates of recovery add further weight to concerns that intense fires can exacerbate tuart tree decline (reviewed in Chapter 2 and supporting evidence in Chapter 3). Alternatively, the reported stimulatory effect of fire on eucalypt trees (reviewed in Chapter 1), was apparent from the monitoring of tuart canopy areas at a burnt site. In total, these results highlighted both the positive and negative, and direct and indirect effect of fire on established trees depending on fire intensity, species and plant-size factors. Further, they point to a role for prescribed fire to maintain and restore tuart health, including by preventing successions of intense fire.

4.4.1 Indirect effect of fire on established tuart and peppermint trees.

There was some evidence that established tuart trees in a burnt area (site Y1) displayed a positive response to fire. This inference was based on the significant increase in canopy area ratings for tuart trees with minimal canopy damage in contrast to the non- significant increase for the unburnt trees in adjacent areas. These results support the findings of others (Mazanec 1999, Guinto et al. 1999, Bell and Chatto 2003) although some believe that low-intensity fire is less likely to stimulate tree growth (Shea et al. 1981, Abbott and Loneragen 1983). The study in this thesis was distinguished from most others by accounting for the confounding direct effect of fire (defoliation) and by assessing canopies rather than stems diameters. Canopy changes are probably more sensitive indicators of short-term variations in resources than stem diameters. While the biological significance of any increase in canopy area over the short-term (one season) may not be large, growth benefits may accrue over the long-term if the response was to occur regularly (by frequent burning). The regular pulse of plant-available nutrients

136 from frequent fire has been proposed to increase the growth of Xanthorrhoea preissii over the long-term (Swanborough et al. 2003).

Some element of the increase in the canopy density rating for trees in both and unburnt and burnt areas could have been due to seasonal effects. The initial assessment period occurred in autumn whereas the final assessment took place in summer. Eucalypts in southern Australia are renowned for their summer growth flush (Specht and Specht 1999). However, the most serious shortcoming of the survey was the lack of replication (one fire only) and the segregated control. This is a characteristic problem of all but those fire studies where considerable resources can be invested to ensure multiple, planned fires occur at desired intensities.

Sampling of interspersed controls (trees in unburnt patches) and treatment trees (trees in burnt patches) at the same site, did not detect any effect of burning on leaf P or N content. Increased concentrations of N are often attributed to greater post-fire growth (Bond and van Wilgen 1996). By sampling patches where fire behaviour was mild (as indicated by the ≤ 10 % tree canopy scorch), the soil in these patches may not have been as enriched in available forms of N or P, relative to those where more fuel was burnt. Nevertheless, there was a significant difference in δ15N between the burnt and unburnt treatments suggesting that contrasting nutrient sources may have been accessed by the trees on the burnt compared to the unburnt patches. Other studies have attributed post- fire increases in foliar δ15N to changes in the forms of nitrogen availability in the soil (for example, Grogan et al. 2000, Cook 2001), but as a complex of factors influence plant δ15N, interpretation of plant access to specific forms of N is difficult (Hogberg 1997). The near-significant result for a species x burn-treatment interaction for δ15N points to an avenue for further work. If access to nutrients is altered differently between the species following fire, the relative performance of the species may differ under contrasting fire-frequency regimes. Any changes in the types or functioning of mycorrhizal associations following fire may also play an important role in this.

An increased nutrient availability in the soil due to fire (Chapter 1, section 1.5.4) may not necessarily be reflected by increased concentrations of nutrients in leaves or in any other leaf characteristic; greater N supply can lead to a larger plant photosynthetic area through changes at the scale of the leaf or the canopy as a whole (Kozlowski et al. 1991). For example, fertilizer addition in a young E. grandis plantation, increased foliar

137 biomass and N concentrations in stems and roots, but elevated concentrations of N were not detected in the leaves (Campion et al. 2006). For the study in this thesis, there was no indication that leaf morphology was affected and as noted above, larger canopy areas were suspected to have formed as a result of burning. So, the possibility remains that any change in N availability may not have been detected by the analysis of leaf concentrations due to the greater total leaf area (whole-canopy response). Clearly, the best approach would be to assess both leaf properties and canopy area before and after fire on the same trees. However, in this study suitable trees for sampling could only be determined post-fire.

4.4.2 Comparison of the anatomical features affecting the capacity of tuart and peppermint to survive fire and recover following fire.

Stem decapitation experiments showed that two-year-old seedlings of both tuart and peppermint had dormant buds from which shoot growth could be activated after defoliation. As the observed resprouting behaviour was for seedlings raised under the favourable environment of the glasshouse, and was for a cutting, not a burning treatment, it is best described as a potentiality. Burning treatments often lead to higher mortality than cutting treatments as a consequence of bud death (Bond and van Wilgen 1996, Hodgkinson 1998). Regardless, the finding is of particular significance to tuart, given that the recruitment of this species is strongly tied to fire (Chapter 3), and that the minimum possible fire interval for tuart woodland is probably around two to four years (Christensen et al. 1981, Ward 2000). Under a regime of frequent fire eucalypt seedling mortality may be high, but not complete due to spatial variations in both fire intensity and the capacity of seedlings to resprout (Noble 1984, Fensham and Fairfax 2006). In addition to spatial variations, temporal variations in intensity are also important in determining recruitment patterns in frequently-burnt systems (Higgins et al. 2000). Such variability offers an explanation for the co-existence between fire frequencies as high as two to four years in tuart woodland (Ward 2000) and the perpetuation of tuart trees.

While peppermint exhibited an ability to resprout from below-ground within two years (assuming the lignotuber was buried), tuart relies solely on the insulation afforded by bark to protect its reserve buds which are situated on the stem above the soil. Bark thicknesses of two-year-old seedlings were comparable between the species. Therefore, peppermint seedlings may be able to survive more intense fires than tuart at the seedling stage if the peppermint lignotuber is sufficiently buried. However, under field 138 conditions, seedling vigour and spatial variability in fire intensity overwhelm species- specific influences on seedling survival (Noble 1984). A strong, positive relationship between bark thickness and stem size for both tuart and peppermint was revealed in the present study. Thus, growth rather than species would likely be the primary determinant of resistance to fire for tuart and peppermint seedlings.

The fact that successful tuart regeneration apparently occurs mostly on ashbeds (Chapter 2 section 2.9.2), may account for the ability of the species to survive and grow under a very frequent fire regime. This is because seedlings with larger stems and therefore thicker bark would develop more rapidly due to the growth stimulus of the ashbed (Chapter 1 section 1.6.3). In addition, ashbeds would represent sites of low fire severity in a subsequent fire event soon after the first as coarse fuels would have already been consumed. Further, the height gains enabled by the growth stimulus of the ashbed may also be important for the seedling to develop as saplings beyond the flame zone. This check on tree seedling height from frequent fire (and competition) due to the presence of grasses, has been termed the “Gulliver effect” in savannah systems (Bond and van Wilgen 1996). Occasional “escapes” from the scorch zone by rapid height growth can ensure the recruitment of tree species in these systems (Bond and van Wilgen 1996, Higgins et al. 2000) or eucalypt invasion of grassland (Fensham and Fairfax 2006). Thus, ashbeds may represent “escape routes” for tuart seedlings in frequently burnt woodlands.

The considerable bark thickness on the trunks of tuart trees implies a capacity for this species to persist with fire despite the lack of a lignotuber. While density and chemical properties influence the insulative effectiveness of bark, thickness is generally recognised as the most important factor (Gill and Ashton 1968, Vines 1968). All tuarts assessed with a dbhob > 5 cm, had at least a 15 mm layer of bark at the base. From laboratory experiments with three types of eucalypt bark, Gill and Ashton (1968) showed that for 15 mm of bark, at least five minutes of exposure to heat comparable to that generated from a fire, was required for cambial temperatures to rise by 40°C. Where, bark does not ignite, the duration of heating necessary to attain this temperature rise for the same bark thickness may be as long as ten minutes (Gill and Ashton 1968). No peppermint stems (all < 30 cm dbhob) had bark > 15 mm in thickness at the base of the stem. Thus, the greater bark thickness for tuart, allows one to predict with some

139 confidence that tuart stems would have a greater capacity to resist fire damage than peppermint when compared on a stem-size basis.

Maintaining stem function after fire provides a competitive advantage by enabling recovery of foliage at height above the resprouting foliage of other plants with thinner bark (Hodgkinson 1998) such as peppermint. As the bark-thickness stem-size relationship at breast height was more pronounced for tuart, they would be more able to retain the ability to resprout aerially than comparably sized peppermints subject to a similar fire intensity/severity. But, at intensities sufficient to cause cambial death above the base, this competitive advantage would be lost as resprouts would originate at the same height as lignotuberous species, or not at all if cambial death was also complete around the base.

4.4.3 Observations of survival and resprouting response in tuart, peppermint and co-occurring tree species following complete canopy scorch

For seedlings and small saplings (dbhob ≤ 1 cm) with completely scorched or incinerated foliage, more tuarts survived than peppermints. Distinguishing seedlings on the basis of size and fire severity supports the strength of the related proposal that tuart seedlings/small saplings were more resistant to fire than peppermint. However, without a more specific indication of local fire intensity (difficult to achieve with observation in natural areas) and stem age, definite conclusions on species differences across all species were not possible. The high rates of survival (75 % and above) for tuart, B. attenuata and M. raphiophylla seedlings and small saplings no doubt over-represent survival rates that would occur for moderate and high intensity fires.

For the few tuart seedlings examined at site Y1, stem age did not correspond with the time of the last fire. This was unexpected considering that the seedling regeneration mode of this species is strongly tied to fire (Chapter 3). If aging of the stems was inaccurate (eucalypts are notoriously difficult to age using growth rings [Ogden 1978]), then these seedlings may have represented the highly suppressed individuals of the immediate post-fire cohort. Either way, the survival behaviour of the tuart seedlings at site Y1 (and possibly the other sites) may have been atypical representations of most tuart seedlings which have the potential to grow to maturity, that is, seedlings on an ashbed and in a gap. Growth limitations would affect resprouting capacity by not only

140 influencing seedling size and bark thickness, but also by limiting carbohydrate reserves, particularly if the seedlings were shaded (Bowen and Pate 1993, Olano et al. 2006).

For stems > 1 cm dbhob, the majority (56 to 100 %) of tuart and all the co-occurring tree species examined displayed a strong tendency to survive and recover after complete foliage scorch. Thus, these species are fire-resistant. Survival of trees after fire across a range of intensities in southwestern Australia (with some tree species common to this study) also report high survival rates (Baird 1977, Grant et al. 1997, Bell et al. 1992, Wardell-Johnson 2000). Conclusions concerning the resistance of peppermint trees to fire can only be limited because few completely-scorched individuals > 1 cm dbhob (n = 18) and none > 5.1 cm dbhob were examined. However, this may not be necessary as the presence of the lignotuber would virtually guarantee post-fire recovery, assuming part of this structure is buried to sufficient depth.

The three tuart deaths observed in the > 20.1 dbhob cm size-classes following full crown scorch, all were rated at lower than average health, suggesting factors in addition to fire contributed to their demise. Apart from limits imposed by physiological status or possibly low carbohydrate reserves, trees with incomplete canopies may have thinner bark and therefore be more vulnerable to cambial damage (Gill 1980, Gill 1997), or even a less than complete covering of thick bark, which could also raise the risk of fire damage (Bell et al. 1989). However, the failure of these large tuart trees (> 50 cm dbhob) to resprout relative to smaller tuart trees may have been an artifact of a positive correlation between tree size and fire intensity within the data set. This was an unavoidable consequence of searching for large trees with complete scorch in order to supplement the data set in the study. Nevertheless, these deaths underlie the potential for high-intensity fire to cause tree death and decline, or from the regeneration perspective: to create gaps. Only where there is no regeneration of the species within the gap, should any decline in the overstorey be viewed as permanent.

Tuart and the other tree species all exhibited a capacity to resprout from crowns after complete crown scorch, even on stems < 5 cm dbhob. The data for bark thickness suggested peppermint trees were less likely than tuart to resprout aerially (section 4.4.2). However, the small sample size for peppermint prevented a definite assessment of this strength of the tendency of this species to resprout from the crown. For tuart, a dbhob of around 6 cm marked the size above which individuals reliably resprouted from

141 the crown. Thus, for sites with tuart regeneration at this size, low-intensity controlled burning could safely be carried without diminishing potential recruitment. However, where a densely regenerating stand encompasses individuals from < 6 cm dbhob to above, it may be desirable to apply fire to assist the release of the larger individuals from competition; with careful control of intensity, the larger individuals would resprout aerially, thereby recovering their pre-fire canopies more rapidly and at positions above the smaller individuals resprouting from stem bases. This could also apply to situations where a lower or midstorey layer is suspected to be competing with a tuart overstorey.

4.4.4 Comparative early canopy recovery of tuart and peppermint seedlings subject to decapitation.

The results of the decapitation experiment in the glasshouse implied that tuart seedlings may be more vigorous resprouters than peppermint. This aligns with tuart as a C-type and peppermint more of an S-type strategist according to the system of Grime (1977) as discussed in Chapter 2. The ability of tuart to restore its canopy more rapidly than peppermint may enable it to exploit a greater share of the post-fire pulse of resources than peppermint. Further studies carried out over the longer-term would be needed to test if any competitive advantage ensues. In addition, studies would ideally involve larger size-classes to evaluate the full ecological significance of any such effect.

4.4.5 Canopy area (CA) recovery of tuart trees following complete canopy scorch

The recovery of tuart tree canopies following 100 % scorch was not complete one year after fire. Gill (1978), reported complete recovery of E. dives within eight months of full canopy scorch, although at 5 to 7 m tall these trees were relatively small and therefore potentially more vigorous compared to the average of those in the present study. Overall, ratings of CA were generally only slightly below pre-fire levels and thus trees were possibly still recovering at the time of assessments one year after fire.

The negative relationship between increment in CA rating and dbhob for fully-scorched tuarts could have arisen from a combination of several factors. Firstly, as outlined in section 4.2.3, there was probably a positive correlation between tree size and fire intensity arising from the sampling method. Secondly, there may have been some element of older trees declining in their capacity to resprout (Gill 1975, Baird 1977, Bell et al. 1992, Cochrane 1968 in McCaw et al. 1994, Gill 1997). Thirdly, the point raised above in relation to younger trees being more vigorous may also apply. However, the 142 inverse relationship between initial CA rating and the change in CA rating was unexpected and cannot be explained readily; intuitively, it follows that higher pre-fire CA ratings would have represented healthier and therefore, more resilient trees.

The other results suggested that the recovery of canopy area for trees scorched between 11 and 90 % was on average complete within one year. Ideally, examination of the trends within sub-categories of scorch-classes may be more revealing, but the small sample size did not permit this. By the end of the monitoring period, the lack of a difference in canopy area change between the trees with 11 to 90 % scorch and the unscorched trees was surprising considering that a sub-set of the trees comprising the unscorched trees data-set (trees at site Y1, Y2 and Y3) exhibited an increase in median canopy rating. This highlights the need for replication across sites, and season if possible.

4.4.6 Conclusion

An individual fire may affect the health of tuart trees 9 to 12 months after fire, both positively and negatively depending on fire intensity. A controlled fire at low intensity is likely to have a minimal impact on the population structures of a range of tree species in tuart woodland. Tuart and the other tree species examined displayed considerable resilience to complete canopy scorch, even as seedlings. However, not surprisingly, seedling mortality was markedly higher for tuart and peppermint as seedlings, than as adults. Therefore, repeated controlled burning has the potential to alter the population structure of these species over the long-term by reducing seedling/small sapling survival. The studies comparing the species suggest that the effect may be greater for peppermint than for tuart. Some evidence for a stimulatory effect on the health of tuart trees from low-intensity burning as well as the potential for intense fire to kill large unhealthy tuarts was apparent. Thus, regular controlled burning may not only benefit tuart tree health, but also reduce the likelihood of intense, damaging fires which at short intervals appear detrimental to tuart health (Beard 1967, Fox and Curry 1981, Pigott and Loneragen 1995, Chapter 3). However, the results presented in this chapter were based on observations from single fire events. Long-term studies monitoring the cumulative outcome of multiple fires would be valuable.

143

144 5 Chapter 5. A comparison of tuart and peppermint seedling establishment

145 5.1 INTRODUCTION

With the health of tuart trees in decline in the Perth metropolitan area (Beard 1967, Fox and Curry 1980, Pigott and Loneragan 1995) and at Yalgorup (Bradshaw 2000, Longman and Keighery 2002, Edwards 2004, Chapter 3), there is a need for tuart seedling regeneration to occur in these areas so as to prevent the further fragmentation of tuart woodlands. However, tuart seedlings are generally scarce within tuart woodlands (Keene and Cracknell 1972, Fox and Curry 1980, Backshall 1983, Bradshaw 2000, Chapter 3) while peppermint seedlings are abundant where the species grow together (Backshall 1983, Bradshaw 2000, Chapter 3). Thus, a process of species drift, as has been reported in other eucalypt woodlands (for example, Withers and Ashton 1977), is apparent. With a link between recruitment and periodic severe fire for many eucalypts (Chapter 1), including tuart (Chapters 2 and 3), a reduction in fire frequency may have facilitated this process in some areas (Withers and Ashton 1977, Bradshaw 2000, Chapter 3)

Tuart seedling establishment is much more common in recently burnt areas than in unburnt areas (Ruthrof 2001, and Chapter 3), whereas peppermint establishment appears to have less specific requirements (Bradshaw 2000, Chapter 3). Tuart seedlings also display a strong reliance on ashbeds for establishment (Keene and Cracknell 1972, Forests Department W.A. 1979), a feature in common with other non-lignotuberous eucalypts in productive environments (Ashton 1981). However, few, if any published data are available to compare the early growth and survival rates of tuart and peppermint. Such data would assist in explaining the differences in the patterns of establishment between the species. This knowledge could also assist restorative works aimed at promoting tuart establishment in areas of tuart tree decline.

This Chapter tests the hypothesis that there are differences in growth and survival rates between tuart and peppermint at the seedling establishment stage. Early growth and survival of tuart and peppermint seedlings was compared under controlled conditions, and under field conditions in burnt and unburnt environments. Differences were discussed in terms of ecological strategies and the altered availability of resources (particularly nutrients and water) in the year after fire. Specific hypotheses for each study that was conducted are outlined in the following section.

146 5.2 METHODS

5.2.1 General

Seed source In experiments involving seeds, seedlings in glasshouse experiments or seedlings planted into the field, seed was sourced from the Lake Clifton area (tuart) and the Perth to Bunbury area (peppermint).

Statistical analyses MINITAB Release 13 (Minitab Inc. 2000) was used for all statistical tests and significance (α) was set to 0.05 unless otherwise stated. Where multiple responses (time series or different attributes) were available for testing, MANOVA was performed initially. When significant values were calculated, ANOVA was then performed for each response variable separately in conjunction with Tukey’s pairwise comparisons. With all MANOVA and ANOVA analyses, model residual plots were diagnosed for serial correlation, constant variance and normality. Data transformations were made where necessary. P values from Pillai’s trace were reported given that this measure is less affected by departures from normality and constant variance compared to the others (Townend 2002). Regardless, in instances where departures from model assumptions could not be rectified for MANOVA, adjustment of α to 0.01 was made to reduce the probability of Type 1 errors (Tabachnik and Fidell 1996). t-tests were used for the comparison of two means after checking the data for normality and equal variance. Data were transformed where necessary. For proportions, comparisons were made using the z-test. For multiple comparisons of two proportions or means, α was modified in accordance with the Sequential Bonferroni Method (Holm 1979 in Quinn and Keough 2002).

5.2.2 Field studies

Background There were five field studies undertaken. Studies 1 to 3 focused on the growth and survival of tuart and peppermint seedlings; seedlings were examined at burnt sites, including on and off ashbeds, and at an unburnt site. The aim of Studies 4 and 5 was to assess the importance of changes in soil moisture and soil chemistry following fire in the differences between seedling growth and survival at burnt and unburnt sites.

147 Sites Studies were carried out at sites Y1, Y2, Y3 and GB. Details for these sites, including fire history and plot layouts, were presented in sections 3.2.1 and 3.2.3 (Chapter 3).

Study1: survival and growth of tuart seedlings that developed from natural germinants on and off ashbeds

Hypothesis

• Survival and growth rates for tuart seedlings are different on ashbeds to off ashbeds at a site that has been recently burnt.

Design All seedlings on ashbeds (initial n = 68) were compared with seedlings off ashbeds (initial n = 158) within four plots at a recently-burnt site. Sampling occurred on five (height) or six (survival) occasions.

Sampling and measurments Following the February 2003 fire at site GB, tuart seedling growth and survival was monitored within the four 25 x 25 m plots. In June 2003, all germinants within the plots were tagged and their location as either on or off ashbed was noted. Height and survival was recorded periodically until March 2005.

Statistical analyses Ashbed and non-ashbed seedlings were compared using the z-test (survival proportions) and the t-test (mean height). For the August 2003 height data, Mood’s Median Test was used as the t-test assumptions of normality and equal variance could not be satisfied.

Study 2: survival and growth of tuart and peppermint seedlings at a burnt site

Hypothesis:

• Survival and growth rates for tuart and peppermint seedlings are different at a site that has been recently burnt.

Design Within one plot at a recently-burnt site, tuart seedlings on ashbeds (initial n = 16) and off ashbeds (initial n = 51), and peppermint seedlings on ashbeds (initial n = 13) and off ashbeds (initial n = 51) were compared for survival and height on two occasions.

148 Sampling and measurements The sampling occurred within the central 20 x 20 m area of plot 6 at site Y1 between August and November 2005. This plot was chosen due to the observed abundance of post-fire regeneration for both tuart and peppermint. In August 2005, every tuart germinant within the area was identified, tagged, measured for height and the substrate was noted (that is, ashbed or non ashbed). As peppermint germinants were more abundant than tuart, the nearest peppermint to every tuart was tagged and also measured for height and the substrate was recorded. Germinants on unburnt patches of ground were excluded from the analysis. Height and survival was recorded again 81 days later in November 2005.

Statistical analyses Survival proportions for the species were compared using the z-test. As sample sizes were too small to compare the species on ashbeds (a warning message was generated in MINITAB), only the seedlings off ashbeds were compared. For height, ANOVA was performed for height at each measurement time. This was deemed to be preferable to a MANOVA incorporating the two measurement times because the data for August had serious departures from normality and constant variance. Species (tuart and peppermint), substrate (on ashbed and off ashbed) and species x substrate were the factors in the ANOVA models.

Study 3: survival of tuart and peppermint seedlings planted at an unburnt site

Hypothesis:

• Survival rates for planted tuart and peppermint seedlings differ at a site that has not been recently burnt.

Design The survival rates of tuart (n = 79) and peppermint seedlings (n = 79) were compared at one site. There were eight monitoring times.

Site, experimental layout and measurements Given the difficulty in locating sufficient numbers of tuart and peppermint seedlings to monitor in unburnt areas, seedlings were planted at site Y2 (unburnt for 11 years) in and around the general area of plot 1. The seedlings were arranged at 20 m intervals along transects aligned from north to south. At each 20 m point, one tuart and one peppermint

149 seedling were planted approximately 0.25 m either side of the transect line. There were six transects spaced 20 m apart, each ending before an ecotone to the south that was defined by the presence of Banksia spp. and a decline in peppermint density. Consequently, the transect lines had different lengths. Monitoring of seedling survival occurred monthly from July 2004 until January 2005. A final assessment took place in April 2005. At each time, the suspected causes of seedling deaths were noted.

Seedlings and experiment establishment Initially, 1 to 2 week-old germinants of tuart and peppermint seed were placed in 64 cell trays filled with potting mix comprising composted pine-bark, coarse sand and “coco peat” in a 2:2:1 mix (Richgro Garden Products, Perth). Approximately 4 g of dolomite and 4 g of calcium carbonate was mixed thoroughly into each 9 L bucket full of potting mix. Seedlings were tended under glasshouse conditions (evaporatively cooled) for five months. A solution of soluble fertilizer (Thrive®, Arthur Yates and Co. Ltd. Australia: 27 % nitrogen, 5.5 % phosphorus and 9.0 % potassium plus trace elements) was applied to foliage and soil periodically over this time. In June 2004, the seedlings were planted at the experimental site at five months of age using a “pottiputki” planting device (Lannen Plant Systems).

Statistical analysis The survival proportions were compared between the species at each monitoring time using the z-test.

Study 4: The effect of fire on soil chemistry and soil moisture

Hypothesis:

• There are differences in soil chemistry and soil moisture between ashbed, burnt- non-ashbed and unburnt patches within tuart woodland that has been recently burnt.

Design There were five blocks (discrete sampling locations) within one recently-burnt site. Each block comprised an ashbed, burnt-non-ashbed and unburnt patch “treatment”. For each patch-type, soil samples from 0 to 5 cm and 5 to 10 cm depth were collected for chemical analysis, and soil samples from 20 to 25 cm depth were collected for moisture content determination.

150 Site and sampling Soil was collected at site Y1 approximately 10 months after the December 2004 fire. A 5.3 cm diameter steel tube was used to extract the soil sample. Ashbeds were defined as a layer of grey or white ash covering at least 1 m2. Five ashbeds were located across the site and sampled accordingly. The closest (but > 2 m away) patch of burnt and unburnt ground was sampled from each ashbed. For the analysis of chemical attributes, each replicate of each soil-type x soil-depth combination comprised a bulked sample of twelve sub-samples; samples were air-dried upon return to the laboratory. For the determination of soil moisture content, only one sample was collected per replicate; immediately upon unearthing, samples were placed in aluminum containers and sealed with plastic tape.

Measurements Chemical analysis was carried out by Cumming Smith British Petroleum (CSBP) Ltd using the methods in Table 5.1. Gravimetric water contents were determined in the laboratory after weighing the soils both before and after drying at 105°C for 48 hours.

Table 5.1: Methods of soil analysis used by Cumming Smith British Petroleum (CSBP) Ltd, Bibra Lake, Western Australia. n/a = not applicable. Component Procedure Attribute Abbreviation reported name Reference Phosphorus P Available Colwell Rayment and Higginson 1992 Potassium K Available Colwell Rayment and Higginson 1992 - Nitrate NO3 n/a n/a Searle 1984 + Ammonium NH4 n/a n/a Searle 1984 Sulphur S Extractable n/a Blair et al. 1991 Organic Carbon Organic C n/a Walkley-Black Walkley and Black 1934 Electrical Conductivity EC 1:5 soil solution 1:5 Rayment and Higginson 1992 ph ph CaCl2 CaCl2 Rayment and Higginson 1992

Statistical analyses For soil chemistry, a MANOVA with the eight soil attributes (Table 5.1) as responses and subsequent ANOVA with each attribute (response) individually, were performed with the following factors: block (sampling location), soil (ashbed, burnt-non-ashbed or unburnt patch), sample depth (0 to 5 cm or 5 to 10 cm) and sample depth x soil. For soil moisture content, an ANOVA was carried out with block (sampling location) and soil (ashbed, burnt-non-ashbed or unburnt patch) as the factors. 151 Study 5: The effect of fire on the moisture content of sub-surface layers of the soil

Hypothesis:

• The moisture content in the sub-surface layers of the soil differs between recently burnt and not-recently-burnt sites.

Design Comparisons were made for two depths (50 cm and 100 cm) between one burnt and two unburnt sites; as there was only one burnt site, data were pooled for the two unburnt sites and henceforth an unburnt “site”, rather than “sites” is referred to. There were six samples for each site-type x depth combination, except for the unburnt site at 50 cm where there were eight samples collected. Measurements were made on two occasions prior to burning and once following burning.

Site, basic sampling procedure and measurements The samples were collected at site Y1 (burnt site) and sites Y2 and Y3 (unburnt sites) at the corners of the permanent plots (see sections 3.2.1 and 3.2.3 in Chapter 3 for full site, plot and fire details). Collections were made in November 2003 and October 2004 (pre- fire) and in November 2005 approximately one year after the fire. Immediately upon unearthing, samples were placed in aluminum containers and sealed with plastic tape. Gravimetric water contents were determined in the laboratory after weighing the soils before and after dying at 105°C for 48 hours.

Additional details of sampling Sampling was designed such that positions were well dispersed. A position within 5 m of the northwest corner of each permanent main plot was sampled. However, due to fewer plots being located at the unburnt sites, the southeast corner of the plot was also sampled at the unburnt site. But at both the burnt and the unburnt site, where a shallow rock layer prevented sampling to 50 cm, one of the adjacent plot corners was randomly selected and sampled instead. At the burnt site, sampling occurred at two additional plots (7 and 8) that were established along with the six main plots (1 to 6) set up for various experimental objectives (the plot coordinates are listed in Appendix 5). These two plots were not used for any other purpose after being set up. Within the site where the fire was lit, there were two positions (associated with plots 2 and 7) located in large unburnt patches and where soil was sampled to only 0.5 m depth (limestone prevented

152 sampling to greater depth). The data for these samples were added to the unburnt data for the 0.5 m depth.

Survey 5: Statistical analysis A MANOVA was conducted with the three sampling times as the response variables and the following as the factors: depth (50 cm or 100 cm), site-type (burnt or unburnt) and depth x site-type.

5.2.3 Glasshouse and laboratory studies comparing tuart and peppermint

Background Five studies were conducted to compare selected attributes of tuart and peppermint from the seed to the young-seedling stage. The aim was to further understand why patterns of seedling establishment differ between the species. Study 1 examined seed weights and Study 2, germination rates. In Study 3, the seedling growth rates of the two species were compared in burnt and unburnt soils collected from the field. Selected morphological attributes of seedlings were compared between the species in Study 4. And in Study 5, the shoot and root growth of of seedlings was examined under contrasting fertilizer and watering regimes.

Study 1: seed weight

Hypothesis:

• Tuart and peppermint seeds have different weights.

Design The weights of three replicate batches of 50 seeds for each species were compared.

Methods Seeds of healthy appearance (that is, not shriveled or diseased) were selected under a dissecting microscope. After sorting, the seed was dried at 70°C for 48 hours and then weighed. The weight per seed for each replicate of species was determined by dividing the batch weight by the number of seeds: 50. The weights for each species were compared using the t-test.

153 Study 2: germination rate

Hypothesis:

• Germination rates of tuart and peppermint are different.

Design Germination rates within three replicate batches of 50 seeds each for tuart and peppermint were compared. Each replicate was set up at a different time (separated by seven or ten days) and assessed ten times during the experiment.

Experimental set-up and measurements Seed of healthy appearance was selected under a dissecting microscope. Each replicate batch of seeds per species was placed on filter paper which was underlain by sponge and sealed in a Petri dish. The filter paper was maintained at full saturation with a weak fungicide solution (0.015 % Previcur® [Bayer CropScience, Australia] in deionized water) by maintaining a constant pool of the solution in the bottom of the Petri dish. Petri dishes were placed in full darkness inside a box within a ventilated room, although the seeds were exposed to light for brief periods (minutes) when each of the ten assessments were made. The experiment began in May 2006, a month in which germination commonly occurs in the region, with a replicate of peppermint and tuart. An additional replicate of each species was set up seven and ten days later. Each replicate was inspected every third day over the 30 days after being set up. As each seed germinated (that is, when the radicle emerged from the seed) it was recorded and removed. Seeds that were infected by mould were also removed, but excluded from the analysis. In addition, seeds that had not germinated after 30 days were squashed then checked for the presence of white endosperm, which if absent also resulted in the seed being excluded from the analysis. Maximum and minimum temperatures were recorded daily and ranged from 12 to 22°C, respectively.

Statistical analysis Statistical analysis was unnecessary as at each time at least one species had either 100 % or 0 % germination for all replicates.

Study 3: seedling growth rates

Hypotheses:

• Tuart and peppermint seedlings have different early growth rates. 154 • There is a significant interaction between species (tuart or peppermint) and soil type (ashbed, burnt-non-ashbed or unburnt) for early growth rate.

Design The experiment comprised the following factors: block (sampling location), species (tuart and peppermint) and soil (ashbed, burnt-non-ashbed or unburnt patch). There were five blocks, each with one replicate of each soil type. For each replicate of soil type, there were three potted seedlings (sub-samples) of each species. In total, there were 45 tuart seedlings and 44 peppermint seedlings in the experiment; one sub-sample for peppermint in burnt-non-ashbed soil could not be included due to insufficient soil. The potted seedlings were set out in a completely randomized design over the course of the experiment

Soil sampling and processing Excess soil collected from survey 3 (section 5.2.2 above) was used in this experiment. After the samples were air-dried, they were passed through a 2 mm sieve. After sieving, plastic pots 9 cm tall were filled with approximately 240 cm3 of soil to a depth of 8 cm; the surface to 5 cm layer of soil originated from the 0 to 5 cm depth in the field and the lower 5 to 8 cm in the pot was from the 5 to 10 cm depth in the field. The pots were free draining, but a layer of hessian covered the drainage holes to ensure the soil did not fall out. One week prior to the start of the growth experiment, 35 ml of a 0.5 % solution of a soil wetting agent (Hydrasoil®, Organic Crop Protectants Pty Ltd, Australia) was added to each pot. This was to minimize any contrast in water repellency between the soil types as a potential source of variation in the the growth of the seedlings. Following the addition of this solution, the soil was then watered twice daily by a sprinkler system.

Experiment establishment, measurements and conditions Sufficient batches of seed were moistened at different times to enable all germinants of both tuart and peppermint to be placed in soil within one to two days of germination. The experiment began as the germinants were transplanted. Two germinants were placed in each pot and after approximately one week seedlings were thinned to one per pot. Every two weeks seedling height (height to the apical bud) was measured and the pots were re-asigned to a random position. Water was supplied twice daily by a sprinkler irrigation system and the experiment ran over 70 days in an evaporatively- cooled glasshouse where temperatures ranged from 8°C to 31°C.

155 Statistical analysis A MANOVA was carried out with the heights for the five measurement times as the response variables. The model included the following factors: block (field sampling position), species (tuart and peppermint), soil (ashbed, burnt-non-ashbed and unburnt patch) and the species x soil interaction.

Study 4: seedling morphology

Hypotheses:

• Tuart and peppermint seedlings differ in respect to a range of morphological attributes. • There is a significant interaction between species (tuart or peppermint) and soil type (ashbed, burnt-non-ashbed or unburnt) in respect to morphological attributes of seedlings.

Design This study was a continuation of study 3. Thus the experimental design was as outlined for 3 above. However, only 11 seedlings of each species were studied. These seedlings were identified by selecting four (of the five) blocks at random, then selecting one seedling of each species at random from each soil type within the block (sampling position). There were only 11 seedlings of each species as the data for one peppermint and one tuart seedling were lost; these lost data came from the burnt-non-ashbed treatment. As a result of the missing data and the non-significant effect of block in the results for Study 3, block was not considered as a factor in this study.

Experiment establishment, conditions and measurements The experimental conditions were as outlined for Study 3 above. However, this latter study was allowed to run for a further eight days. At day 78, the seedlings were harvested and separated into root and shoot components. Shoots were weighed immediately for “fresh weight” (FW). Leaf thickness was then measured with digital calipers placed parallel to the mid-vein and half-way between the mid-vein and leaf edge. Next, all leaves were scanned and leaf areas were measured using ASSESS software (American Phytopathological Society 2002). Finally, all shoot and root material was dried at 70°C for 48 hours after which dry weights (DW) were determined. In total, there were seven attributes that were either measured or derived: shoot DW,

156 root DW, shoot:root ratio (DW), leaf FW:DW ratio, leaf area, leaf thickness and specific leaf area (SLA).

Statistical analyses A MANOVA was conducted with the seven morphological attributes as the response variables. The factors were species (tuart and peppermint), soil (ashbed, burnt-non- ashbed and unburnt patch) and the species x soil interaction. Following a significant result, ANOVA were performed for each attribute using the same factors as for the MANOVA.

Study 5: shoot and root growth-responses of seedlings to contrasting fertilizer and watering regimes

Hypotheses:

• For root and shoot growth, there are differences between tuart and peppermint seedlings. • There is an interaction between species (tuart and peppermint) and the rate of fertilizer addition for root-growth and shoot-growth responses. • There is an interaction between species and rate of water addition for root- growth and shoot-growth responses. • There is three-way interaction between species, rate of fertilizer addition and rate of water addition for root-growth and shoot-growth responses.

Design A randomized block design comprising three blocks (positions within the glasshouse), and 32 replicates each for a watering factor (high or low), a fertilizer factor (high or low) and a species factor (peppermint or tuart) was established. All factors were crossed. However, the final number of replicates at the end of the experiment was altered due to the death of two peppermint seedlings (one high-fertilizer low-water treatment and one low-fertilizer low-water treatment) and the loss of data for the shoot weights of two peppermint seedlings (high-fertilizer high-water treatments).

Experiment establishment and general conditions The experimental set-up represented a modified version of that for examining the response of root growth to drought as outlined by Reader et al. (1993). Seedlings were grown in washed, white silica-sand contained in polyvinylchloride (PVC) pipes of 1 m

157 length and 8.8 cm diameter. Initially, seed was germinated on either filter paper or white sand. Sufficient germinants of both tuart and peppermint were available to be placed in soil within one to two days of germination. The experiment commenced on the day the germinants were transplanted. Two germinants were placed in each pipe and thinned to one seedling per pipe after approximately one week. Clear plastic film was placed over the mouth of the pipe to maintain a favourable moisture environment around the germinants for the first three days. Additional measures were also put in place to ensure drought stress did not kill the developing seedlings; black plastic discs (approximately 8.5 cm diameter) with a central hole (approximately 1 cm diameter) were placed on the soil surface to reduce evaporation from the soil until day 20, shadecloth was placed over the pipes during the midday hours during two days of high temperature in the first week and the seedlings were also periodically misted with DI water over the first week of the experiment. The experiment ran over 41 days within an evaporatively cooled glasshouse in which temperatures ranged from 12°C to 30°C.

Watering and fertilizer treatments Treatments for nutrient availability comprised a “low” and a “high” dose of fertilizer (Thrive®) in DI water applied to the soil. Initially the low dose was 0.02 g L-1 and the high dose was 0.2 g L-1. The recommended application rate of this product for seedlings is approximately 1 g L-1, with 4.5 L of solution applied to 1 m2 of seedlings every two weeks. Immediately before the start of the experiment, 1 L of the solutions was added to each pipe. Thereafter, 25 ml day-1 was added to each pipe for the first four days. Given the high temperatures being experienced, the volume of solution applied to each seedling was increased to 35 ml day-1 and the solution strengths were lowered to 0.15 g L-1 (high) and 0.015 g L-1 (low). At day 8, concerns about the health of the low-nutrient treatment seedlings led to the high dose being applied to all seedlings for that day only. From day 10, the concentration in the low dose was doubled to 0.03 g L-1.

At day 20, the watering treatment was applied: half of the seedlings of each species per fertilizer treatment per block were selected at random and were not watered (drought treatment) until the end of the experiment (day 40). However, on day 21 fertilizer solutions were added in error to the watered treatments. From then on, watered treatments consisted only of DI water.

158 Harvesting and measurements Seedlings were harvested on day 41. Pipes were cut open and the extent of rooting depth was determined by carefully removing soil around the roots at progressive depths. The roots were then separated from the soil by washing and sieving. Shoot length was measured, then both the root and shoot material was dried at 70°C for a minimum of 48 hours and weighed

Statistical analysis A MANOVA was carried out with six response variables: shoot length, shoot DW, root depth, root DW, shoot length:root depth ratio and shoot DW:root DW ratio. The factors in the model were block (position in the glasshouse), fertilizer treatment (high and low), watering treatment (high and low), species (peppermint and tuart) and four interactions comprising all combinations of the last three mentioned factors. Following a significant result, an ANOVA was performed for each response variable.

5.3 RESULTS

5.3.1 Field studies

Study 1: survival and growth of tuart seedlings that developed from natural germinants on and off ashbeds Tuart seedlings off ashbeds had less than half the rate of survival (15 %) as for seedlings on ashbeds (35 %) at the end of the first summer following post-fire germination (Figure 5.1). Rates of mortality for both ashbed and non-ashbed seedlings then stabilized from this time, although the contrast in survival rates continued to widen; at approximately 21 months (March 2005) after germination, survival rates were 22 % (ashbeds) and 4 % (off ashbeds). The differences between survival proportions on and off ashbeds were significant at all times following the initial survey in July 2003 (Figure 5.1 and Table 5.2).

The seedlings on ashbeds were taller than the seedlings off ashbeds, particularly over the first year of growth following the fire (Figure 5.2). At the end of the first summer, the mean height for seedlings on ashbeds was three times as great as those that were off ashbeds. Significant differences existed for the August 2003 and February 2004 measurements. After 20 months, the mean height of ashbed seedlings was still greater,

159 but a significant difference was not evident. However by this time, the sample size was much reduced due to mortality: 15 (ashbeds) and 7 (off ashbeds). Thus, statistical power was lower by the end of the experiment.

a a 100 a 90 a 80 b 70 60 a 50 a a 40 a 30 Mean survival (%) 20 b 10 b b bb 0 Jul-04 Jul-03 Jan-05 Jan-04 Oct-04 Oct-03 Jun-04 Sep-04 Feb-05 Sep-03 Feb-04 Dec-04 Dec-03 Apr-04 Mar-05 Mar-04 Nov-04 Nov-03 Aug-04 Aug-03 May-04 Date

Figure 5.1: Mean percentage (%) survival of tuart seedlings that germinated on ashbeds (■) and off ashbeds (□) following the fire at site GB in February 2003. In July 2003, n = 68 for ashbeds and n = 164 for off ashbeds. At each sampling time, proportions with different letters are significantly different using z-tests with adjustment of α using the Sequential Bonferroni Method. 95 % confidence intervals are indicated as bars.

40 a 35 a a 30 25 a 20 15 a a Mean height(cm) 10 a a 5 b b 0 Jul-04 Jan-05 Jan-04 Oct-04 Oct-03 Jun-04 Sep-04 Feb-05 Sep-03 Feb-04 Dec-04 Dec-03 Apr-04 Mar-05 Mar-04 Nov-04 Nov-03 Aug-04 Aug-03 May-04 Date

Figure 5.2: Mean height (cm) of tuart seedlings that germinated on ashbeds (■) and off ashbeds (□) at site GB following a fire in February 2003. Means with different letters are significantly different following t-tests with data transformations where appropriate and non-parametric tests where data could not be suitably transformed (see Table 5.2); adjustment of α using the Sequential Bonferroni Method was then conducted. Total seedling number at each measurement time in chronological order was 68, 24, 17, 17 and 15 for ashbeds, and 158, 24, 8, 8 and 7 for off ashbed sites. Standard error bars are indicated.

160 Table 5.2: Comparison of survival proportions and mean height (or median height for August 2003) for tuart seedlings that germinated on ashbeds and off ashbeds at site GB following a fire in February 2003. The February 2004 data for height were log transformed. August 2003 data were analyzed using Mood’s Median Test. Significant values are denoted in bold after adjustment of α using the Sequential Bonferroni Method for the height and survival data separately. n/a = not available. Survival Height On ashbed Off ashbed Date z P Test and result P (n) (n)

2 August 2003 -2.70 0.007 χ (1) = 28.87 <0.001 68 158 November 2003 -5.14 <0.001 n/a = n/a n/a 46 54

February 2004 -3.11 0.002 t(46) = 6.65 <0.001 24 24

November 2004 -3.65 <0.001 t(23) = 1.63 0.117 17 8

January 2004 -3.65 <0.001 t(23) = 1.48 0.151 17 8

March 2005 -3.38 0.001 t(20) = 1.26 0.223 15 7

Study 2: survival and growth of tuart and peppermint seedlings at a burnt site For the August to November 2005 period, tuart survival (77 %) was greater than peppermint survival (43 %) for seedlings that germinated on burnt non-ashbed patches within plot 6 at site Y1 following the fire in December 2004 (Figure 5.3). This difference was significant (z = 3.65, P < 0.001). Survival proportions between the species on ashbeds were not contrasted statistically due to the small sample sizes of both tuart (n = 16) and peppermint (n = 13). However, mean survival proportions were higher on ashbeds for both species: 81 % (tuart) and 61 % (peppermint).

100 a 90 80 70 b 60 50 40 Survival (%) 30 20 10 0 Tuart Peppermint Species

Figure 5.3: Percentage (%) survival of tuart and peppermint seedlings that germinated off ashbeds, but on burnt ground, between August and November 2005 at site Y1 (plot 6) following the fire in December 2004. For both species, n = 51. Proportions with different letters are significantly different using the z-test. 95 % confidence intervals are indicated as bars. 161 Tuart seedlings were also several times taller than the peppermint seedlings over the August to November period and the difference was highly significant on both occasions (Figure 5.4 and Table 5.3). An initial MANOVA including both measurement times was not performed due to the inability to overcome serious violations of model assumptions in the August data. Non-constant variance could not even be rectified when the August data were treated in isolation. Nevertheless, given the high level of significance (P < 0.001), the difference between species was accepted for both dates.

For tuart, growth on ashbeds was greater than off ashbeds. In November, the mean height of tuarts on ashbeds was 9.5 cm compared to 6.1 cm for those off ashbeds, but the difference was not significant. The mean heights of peppermint on ashbeds (2.0 cm) and off ashbeds (2.1 cm) were more similar and also not significantly different.

12 a 11 10 9 8 7 a 6 5 4

Mean height (cm) height Mean a 3 a b 2 b 1 b 0 b 5-Augtuart peppermint 25-Nov Date

Figure 5.4: Mean height (cm) of tuart and peppermint seedlings between August and November 2005 at site Y1 (plot 6) following the fire in December 2004. The seedlings germinated on ashbeds (■ = tuart, ■ = peppermint) and off ashbeds (□ = tuart, □ = peppermint). On ashbeds, n = 13 (tuart) and n = 8 (peppermint), and off ashbeds, n = 39 (tuart) and n = 22 (peppermint). Seedlings that died between August and November were excluded from the analysis. Means with different letters at each time are significantly different using Tukey pairwise comparisons on Box-Cox transformed data (August) and log-transformed data (November) in ANOVA (Table 5.3). Standard error bars are indicated.

162 Table 5.3: ANOVA for tuart and peppermint seedling height following germination on ashbeds and off ashbeds (soil) following a fire in December 2004 at site Y1 (plot 6). Box-Cox (August) and log (November) transformations were applied. Due to the two measurement periods, α was adjusted using the Sequential Bonferroni Method. Significant values are denoted in bold. For the August data α = 0.01 due to violation of model assumptions. Species Soil Species x Soil

Date F(1, 78) P F(1, 78) P F(1, 78) P August 188.76 <0.001 0.07 0.793 4.19 0.044 November 55.96 <0.001 1.26 0.265 1.24 0.070

Study 3: survival of tuart and peppermint seedlings planted at an unburnt site Mortality of tuart seedlings was higher than for peppermint seedlings planted at five months of age at an unburnt site (Y2) (Figure 5.5). Ten months after planting, more than four times as many peppermints than as tuarts had survived. With the onset of the summer-drought period (December), the rate of mortality accelerated for both species. However, the rate was greater for tuart between November and January. From December (7 months after planting) until the conclusion of monitoring, the difference in survival proportions between species was significant (Figure 5.5 and Table 5.4). It was inferred from observation that drought caused the majority of seedling deaths. While some of the seedlings were grazed, this appeared to be only a minor cause of death as most dead individuals remained with leaves intact.

90 a a a a a a a 80 a 70 a a a 60 a a a 50 a 40 30 b 20 No. surviving seedlings surviving No. b 10 b 0 June July Aug Sept Oct Nov Dec Jan Feb March April Month (2003-2004)

Figure 5.5: Number of surviving tuart (■) and peppermint (□) seedlings after planting at five months of age at an unburnt site (Y2) in June 2004. At each sampling time, values with different letters are significantly different using the z-test with adjustment of α using the Sequential Bonferroni Method. 95 % confidence intervals are indicated as bars.

163 Table 5.4: z-tests of survival proportions for tuart and peppermint seedlings planted at five months of age at an unburnt site (Y2) in June 2004. Significant values are denoted in bold after adjustment of α using the Sequential Bonferroni Method. n/a = not applicable. Tuart Peppermint Date z P (n) (n) June 2004 n/a n/a 79 79 July 2004 -1.01 0.314 78 79 August 2004 -1.77 0.077 76 79 September 2004 -2.05 0.040 75 79 October 2004 -2.31 0.021 74 79 November 2004 -1.68 0.093 74 78 December 2004 -5.18 <0.001 49 74 January 2005 -6.99 <0.001 26 64 April 2005 -4.60 <0.001 7 30

Study 4: the effect of fire on soil chemistry and soil moisture At ten months after fire at site Y1, pH, EC and concentrations of available P and K and extractable S were all significantly higher in the surface 0 to 10 cm of ashbeds than in burnt-non-ash and unburnt patches (Figure 5.6). Mean available P concentrations were many times greater in ashbed soils, while available K and EC were approximately + - double when compared to non-ashbed soils. In contrast, levels of NH4 , NO3 and organic C were similar and not significantly different between the patch-types.

Overall, MANOVA results indicated that soil chemistry varied significantly according to fire severity (ashbed, burnt-non-ashbed or unburnt) and soil depth (Table 5.5). However, univariate analyses revealed that significant differences were only between the ashbed and the other two soil-types, with no significant differences between the burnt-non-ash and unburnt soil for any of the chemical attributes tested (Table 5.6). For changes with depth, the only significant trends were for the lower concentrations of P and K in the 5 to 10 cm layer than in the 0 to 5 cm layer of ashbeds. The univariate analyses highlighted those attributes that had significant background spatial variability: P and pH (Table 5.6). Such a finding did not follow from the MANOVA, but given that model assumptions were met for the ANOVA but not MANOVA, and that the differences were still significant after Sequential Bonferroni corrections, the ANOVA results were accepted in preference.

164

5 70 l l a a a 4.5 60 4 a 3.5 a 50 ) c ) a -1 -1 3 a a 40 a 2.5 a a 2 a 30 (mg kg (mg kg (mg a c 1.5 b b b b a 20 1 a a 10 b b b b b b b b Mean concentration in soi concentration Mean

Mean concentration insoi concentration Mean 0.5 0 0 0-5 cm 5-10 cm 0-5 cm 5-10 cm 0-5 cm 5-10 cm 0-5 cm 5-10 cm 0-5 cm 5-10 cm - + NO3 NH4 P (available) K (available) S (extractable)

Nutrients and soil depth Nutrients and soil depth a) b)

a 0.16

3 ) ) 9 a a -1 a 8 0.14 a 2.5 ) a 2 b b b a 7 b 0.12 2 a a 6 0.1 a 5 1.5 0.08 4 b 0.06 b 1 3 b b 0.04 2 0.5 Maen soil pH (CaCl 1 0.02 Mean concentration in soil (% in soil concentration Mean m conductivity soil (dS Mean 0 0 0 0-5 cm 5-10 cm 0-5 cm 5-10 cm 0-5 cm 5-10 cm Soil depth Soil depth Soil depth c) d) e)

- + -1 Figure 5.6: a) NO3 , NH4 , b) available P, available K and extractable S concentrations (mg kg ), and c) organic C concentration (%), d) pH levels and e) EC (ds m1) in soil collected at 0-5 cm and 5-10 cm depths from ashbed (■), burnt-non-ashbed (■) and unburnt (□) patches ten months after the fire in December 2004 at site Y1. Means with different letters across depth classes for each attribute are significantly different using Tukey’s pairwise comparisons of log-transformed data except for pH (untransformed) and organic C (cube-root transformed). Standard error bars are indicated. 165 - + Table 5.5: MANOVA results for concentrations of NO3 , NH4 , available P, available K, extractable S, organic C, and levels of electrical conductivity and pH (all square-root transformed) in soil collected at 0-5 cm and 5-10 cm depths from ashbed, burnt-non-ashbed and unburnt patches ten months after a fire in December 2004 at site Y1. Significant values are denoted in bold. Due to violation of model assumptions: α = 0.01. Block Soil Depth Soil x Depth

F(32, 64) P F(16, 28) P F(8, 13) P F(16, 28) P 1.888 0.015 5.607 <0.001 11.576 <0.001 3.000 0.005

Table 5.6: ANOVA results for chemical attributes of soil collected at 0-5 cm and 5-10 cm depths from ashbed, burnt-non-ashbed and unburnt patches ten months after a fire in December 2004 at site Y1. Data were log transformed with the exception of pH (untransformed) and organic carbon (cube-root transformed). Significant values are denoted in bold. Block Soil Depth Soil x Depth

Soil attribute F(4, 20) P F(2, 20) P F(1, 20) P F(2, 20) P - NO3 2.55 0.071 2.20 0.137 0.150 0.707 1.60 0.226 + NH4 0.40 0.803 0.26 0.772 1.07 0.314 4.57 0.023 P (available) 5.08 0.005 133.39 <0.001 26.36 <0.001 6.26 0.008 K (available) 1.26 0.319 167.26 <0.001 3.70 0.069 4.18 0.030 S (extract.) 1.00 0.432 56.43 <0.001 3.30 0.084 0.49 0.621 Organic C 0.45 0.770 0.37 0.694 8.10 0.010 0.20 0.822 EC 2.52 0.074 75.48 <0.001 11.39 0.003 1.74 0.200 pH (CaCl2) 6.58 0.002 26.51 <0.001 1.12 0.303 1.13 0.342

Mean gravimetric water contents varied little (from 4.9 to 5.3 %) at 20 to 25 cm under the ashbed, burnt-non-ashbed and unburnt patches in September 2005, ten months after the fire at site Y1. An ANOVA did not produce a significant result for either patch-type

(F(2, 8) = 0.36, P = 0.710) or sampling location (F(4, 8) = 1.82, P = 0.218).

Study 5: The effect of fire on the moisture content of sub-surface layers of the soil Sub-soil water contents in the burnt (Y1) and unburnt areas (Y2 and Y3) were contrasted in the October/November period of consecutive years; the greatest change at both 50 cm and 100 cm soil-depth occurred between the years rather than in the year following burning (Figure 5.7). Prior to burning, mean soil moisture increased 1 to 2 % at both depths within the designated burnt and unburnt areas between 2003 and 2004. For the samples collected after the fire (November 2005), mean moisture contents were within 0.2 % of the 2004 measurements except for the burnt area at 50 cm where a 0.5 % increment was recorded. However, variation with depth and site could not be

166 distinguished statistically (Table 5.7). The data could not be transformed to overcome minor departures from normality; regardless, none of the reported P values were in proximity to 0.05.

5.5

5.0

4.5

4.0

3.5

3.0 e Fir 2.5 Mean soil moisture (% w/w) (% moisture soil Mean 2.0 Nov2003 2003 Oct 2004 2004 Nov 2005 2005 Date

Figure 5.7: Gravimetric moisture content of soil (%) in October (2004) and November (2003 and 2005) at sites Y1 (burnt in December 2004) and sites Y2 and Y3 (unburnt). Samples were collected from 50 cm depth at burnt (▲) and unburnt (▲) sites, and 100 cm depth for burnt (■) and unburnt (■) sites. n = 6, except for unburnt sites at 50 cm where n = 8. There were no significant differences between the means. Standard error bars are indicated.

Table 5.7: MANOVA for gravimetric moisture content of soil at 50 cm and 100 cm in October (2004) and November (2003 and 2005) at site Y1 (burnt in December 2004) and sites Y2 and Y3 (unburnt). Due to violation of model assumptions: α = 0.01.

Factor F(3, 20) P Depth 0.679 0.575 Burnt/unburnt (B) 1.619 0.217 Depth x B 0.148 0.930

5.3.2 Glasshouse and laboratory studies comparing tuart and peppermint

Study 1: seed weight From three samples of 50 seeds for each species, the mean weight per seed was 1.7 ± 0.10 mg for tuart, but only 0.1 ± 0.01 mg for peppermint. This difference was significant: t(2) = 16.29, P = 0.004.

167 Study 2: germination rate Tuart seed germinated sooner than peppermint seed (Figure 5.8). All tuart seed germinated within six days of moistening, whereas less than 10 % of peppermint seed had germinated by this time. After six days, the rate for peppermint increased rapidly such that by 12 days more than half of the seeds had germinated. The rate then declined, but by 30 days almost all peppermint seeds had germinated (95 ± 0.8 % SE).

100 90 80 70 60 50 40 30 20 10

Mean proportion of germinants (%) germinants of proportion Mean 0 0 3 6 9 12 15 18 21 24 27 30 Days since water added

Figure 5.8: Mean germination rate (%) for tuart (■) and peppermint (□) seed on moistened filter paper in darkness and under ambient temperatures (May and June 2006). There were three replicate batches of 50 seeds for each species. Standard error bars are indicated.

Study 3: seedling growth rate In the 70 days after germination, tuart seedlings were at least twice as tall as peppermint seedlings when grown under glasshouse conditions in soils collected ten months after fire from site Y1 (Figure 5.9). The MANOVA result for all measurement times (Table 5.8) and the ANOVA result for the final measurement (Table 5.9) indicated highly significant differences between the species. However, no significant difference for the effect of block (the collection location within site) or soil (ashbed, burnt-non-ash or unburnt) was revealed.

168 a 5.5 5.0 ) 4.5 4.0 3.5 3.0 b 2.5 2.0 1.5

Mean seedling height (cm 1.0 0.5 0.0 0 12 29 42 56 70 Days since germination

Figure 5.9: Mean height (cm) of peppermint (□, n = 44) and tuart (■, n = 45) seedlings grown for 70 days under glasshouse conditions in soil collected from tuart woodland (site Y1). At day 70, means with different letters are significantly following ANOVA (Table 5.9). Standard error bars are indicated.

Table 5.8: MANOVA for peppermint and tuart seedling height (log transformed) grown for 70 days under glasshouse conditions in soil collected from tuart woodland (site Y1). Soil was collected from ashbed, burnt-non-ashbed and unburnt patches ten months after the December 2004 fire. Measurements were made at 12, 29, 42, 56 and 70 days after germination. Significant values are denoted in bold. Block Soil Species Species x Soil

F(20, 76) F(10, 34) F(5, 16) F(10, 34) 0.417 0.985 1.137 0.365 214.175 <0.001 0.713 0.706

Table 5.9: ANOVA for peppermint and tuart seedling height grown for 70 days under glasshouse conditions in soil collected from tuart woodland (site Y1). Soil was collected from ashbed, burnt-non-ashbed and unburnt patches ten months after the December 2004 fire. Significant values are denoted in bold. Block Soil Species Species x Soil

F(4, 20) F(2, 20) F(1, 20) F(2, 20) 0.59 0.675 0.08 0.922 32.50 <0.001 0.29 0.748

Study 4: seedling morphology After a further eight days of growth, 11 tuart and peppermint seedlings were randomly selected and harvested to compare a range of morphological attributes other than height. Leaf area, root dry weight and shoot dry weight were significantly greater for tuart, while leaf fresh weight:dry weight and specific leaf area were significantly greater for

169 peppermint (Table 5.10). Mean shoot:root ratio was 25 % greater for tuart but this difference was not significant. Initially, a MANOVA was conducted incorporating block (that is, sampling location). However, this factor was not significant, so it was excluded to enable sufficient degrees of freedom for the incorporation of the species x soil interactive term into both the MANOVA and ANOVA models. Significant differences for species, but not for soil treatment, were found (Tables 5.11 and 5.12).

Table 5.10: Mean and standard errors for various morphological traits of tuart and peppermint seedlings (n = 11) grown for 78 days following germination under glasshouse conditions in soil collected from tuart woodland (site Y1). Soil was collected from ashbed, burnt-non-ashbed and unburnt patches ten months after the December 2004 fire. DW = dry wight, FW = fresh weight, SLA = specific leaf area, * = significant differences as reported in Table 5.12. Tuart Peppermint Seedling attributes mean ± SE mean ± SE Shoot DW (g) 0.16 ± 0.032 0.05 ± 0.019 * Root DW (g) 0.07 ± 0.015 0.03 ± 0.008 * Shoot:Root (DW) 2.4 ± 0.29 1.8 ± 0.13 Leaf FW:DW 2.7 ± 0.13 4.4 ± 0.15 * Leaf area (cm2) 14.9 ± 2.79 7.4 ± 2.4 * Leaf thick. (mm2) 0.23 ± 0.007 0.25 ± 0.013 SLA (cm g-1) 114 ± 10.5 167 ± 8.2 *

Table 5.11: MANOVA results for square-root transformation of leaf area, leaf thickness, shoot dry weight, root dry weight, leaf fresh weight:dry weight, leaf area:dry weight (SLA) and shoot:root dry weight of peppermint and tuart seedlings grown for 78 days following germination under glasshouse conditions in soil collected from tuart woodland (site Y1). Soil was collected from ashbed, burnt-non-ashbed and unburnt patches ten months after the December 2004 fire. Significant values are denoted in bold. Due to violation of model assumptions: α = 0.01. Species Soil Species x Soil

F(7, 10) P F(14, 22) P F(14, 22) P 13.418 <0.001 0.975 0.506 0.286 0.991

170 Table 5.12: ANOVA results and data transformations for various morphological traits of tuart and peppermint seedlings grown for 78 days following germination under glasshouse conditions in soil collected from tuart woodland (site Y1). Soil was collected from ashbed, burnt-non- ashbed and unburnt patches ten months after the December 2004 fire. Significant values are denoted in bold. Seedling attributes Data Species Soil Species x Soil

transformation F(1, 16) P F(2, 16) P F(2, 16) P Shoot DW Cube root 9.50 0.007 0.04 0.965 0.31 0.737 Root DW Cube root 7.51 0.015 0.10 0.908 0.47 0.634 Shoot:Root (DW) Log 2.98 0.104 1.29 0.303 0.11 0.898 Leaf FW:DW None 70.87 <0.001 0.21 0.754 0.89 0.611 Leaf area Log 4.96 0.041 0.00 1.00 0.56 0.581 Leaf thickness Box-Cox 1.37 0.259 0.40 0.676 0.18 0.840 SLA 1 Box-Cox 18.94 <0.001 0.08 0.923 0.45 0.645 1. Minor violation of normality assumption. However, with untransformed data, P was also < 2 0.001 (χ (1)= 14.73) for Mood’s Median Test of species: the only factor in model.

Study 5: shoot and root growth-responses of seedlings to contrasting fertilizer and watering regimes Tuart shoots and roots were significantly larger in weight and dimension than for peppermint in a seedling experiment conducted for 40 days in the glasshouse (Figure 5.10). Shoot and root weights and dimensions of both species responded to the contrasting fertilizer treatments. However, the proportional gain in weight and length/depth for roots and shoot was greater for tuart than for peppermint with the higher fertilizer rate. Tuart seedlings produced approximately 2.9 times as much shoot on a dry weight basis under the high fertilizer rate whereas for peppermint the gain was only around 1.8 times. For shoot length the increments were approximately 1.6 times for tuart, but only 0.4 times for peppermint. The contrast in root weight increment with the high fertilizer rate was even more dramatic between the species, with an increase of 4.9 times for tuart and 1.3 times for peppermint. Depth of rooting was also 1.6 times as great for tuart, but only 0.3 times as great for peppermint under the high fertilizer rate. MANOVA results for species and fertilizer factors as well as their interaction were highly significant (Table 5.13). The interaction between species and fertilizer treatment was also highly significant in univariate analyses of shoot and root weight (Table 5.14) and shoot length and rooting depth (Table 5.15).

171

25 4.0 a 3.5 a 20 3.0

15 2.5 2.0 10 1.5 c c 1.0 b 5 b d length shoot (cm) Mean d 0.5 Mean shoot dry weight (mg) (mg) weight dry shoot Mean 0 0.0 High Low High Low Fertilizer rate Fertilizer rate a) b)

25 35 a a 30 20 25 15 20 b 10 15 bc 10 c bc b

5 (cm) length root Mean

Mean root dry weight (mg) weight dry root Mean c 5

0 0 High Low High Low Fertilizer rate Fertilizer rate c) d)

Figure 5.10: Mean a) shoot dry weight (mg), b) shoot length (cm), c) root dry weight (mg) and d) root depth (cm) of tuart (■) and peppermint (□) seedlings grown under glasshouse conditions for 40 days with high and low rates of fertilizer addition. For tuart, n = 16 for each treatment. For peppermint n = 15 for each treatment except for the high-fertilizer rate shoot-weight combination where n = 13. Means with different letters are significantly different following ANOVA and Tukey’s pairwise comparisons of transformed data (Tables 5.14 and 5.15). Standard error bars are indicated.

Shoot:root ratio was significantly larger for tuart than for peppermint with regards to weight and where shoot-length and root-depth were measured (Figure 5.11 and Tables 5.14 and 5.15). Shoot:root dry weight ratio decreased for both species when the drought treatment was imposed (Figure 5.11). Conversely, shoot length:root depth increased for both species with the drought treatment. Tuart displayed the greatest response of the two species, especially when shoot:root dry weight ratio was measured, although there were no significant species x watering treatment interactions (Tables 5.13). A significant

172 effect for the watering treatment was only detected when shoot:root dry weight ratio was measured, not when shoot length:root depth was measured (Table 5.14 and 5.15)

2.0 0.14 1.8 0.12 1.6 1.4 0.10

1.2 0.08 1.0 0.8 0.06

0.6 0.04 0.4

Mean shoot:root DW ratio ratio DW shoot:root Mean 0.02 0.2

0.0 depth ratio length:root shoot Mean 0.00

Normal Drought Normal Drought Watering treatment Watering treatment a) b)

Figure 5.11: Mean a) shoot:root dry weight (DW) ratio and b) shoot length:root depth ratio of tuart (■) and peppermint (□) seedlings grown under glasshouse conditions for 40 days with contrasting watering treatments: daily watering (normal) and daily watering until day 20 only (drought). For tuart, n = 16 for each treatment. For peppermint n = 14 for each treatment except for the normal watering treatment with shoot length:root depth where n = 16. There were no significant differences for the species x watering treatment interactions, but there was a significant difference for species, and for the watering treatment in the case of shoot:root DW ratio (Tables 5.14 and 5.15). Standard error bars are indicated.

Table 5.13: MANOVA for square-root transformation of shoot length, shoot dry weight, root depth, root dry weight, shoot length:root depth ratio and shoot:root dry weight ratio of peppermint and tuart seedlings grown under glasshouse conditions for 40 days with contrasting fertilizer rates and watering regimes. Significant values are denoted in bold. Due to violation of model assumptions: α = 0.01. Factor P

Block F(12, 92) = 1.059 0.403

Species (S) F(6, 45) = 73.133 <0.001

Water (W) F(6, 45) =3.658 0.005

Fertilizer (F) F(6, 45) = 36.757 <0.001

S x W F(6, 45) = 1.515 0.195

S x F F(6, 45) = 16.272 <0.001

W x F F(6, 45) = 2.33 0.048

S x W x F F(6, 45) = 2.069 0.076

173 Table 5.14: ANOVA for shoot dry weight (cube-root transformed), root dry weight (Box-Cox transformed) and shoot:root dry weight ratio (S:R) (log transformed) of peppermint and tuart seedlings grown under glasshouse conditions for 40 days with contrasting fertilizer rates and watering regimes. Significant values are denoted in bold. Shoot dry weight Root dry weight S:R Factor F P F P F P

Block F(2, 52) 0.26 0.774 2.18 0.123 1.36 0.266

Species (S) F(1, 52) 339.52 <0.001 197.23 <0.001 10.89 0.002

Water (W) F(1, 52) 3.28 0.076 2.06 0.157 9.48 0.003

Fertilizer (F) F(1, 52) 151.98 <0.001 158.17 <0.001 2.52 0.119

S x W F(1, 52) 0.01 0.916 2.56 0.115 1.75 0.192

S x F F(1, 52) 30.04 <0.001 15.65 <0.001 3.64 0.062

W x F F(1, 52) 0.04 0.839 2.42 0.126 1.04 0.312

S x W x F F(1, 52) 0.29 0.595 1.7 0.198 1.06 0.309

Table 5.15: ANOVA for shoot length (log transformed), root depth (square-root transformed) and shoot length:root depth (S:R) (Box-Cox transformed) of peppermint and tuart seedlings grown under glasshouse conditions for 40 days with contrasting fertilizer rates and watering regimes. Significant values are denoted in bold. Shoot length Root depth S:R Factor F P F P F P

Block F(2, 52) 0.67 0.514 0.78 0.464 0.46 0.634

Species (S) F(1, 52) 359.40 <0.001 204.04 <0.001 7.17 0.010

Water (W) F(1, 52) 0.33 0.571 1.71 0.197 3.10 0.084

Fertilizer (F) F(1, 52) 137.02 <0.001 115.50 <0.001 0.21 0.646

S x W F(1, 52) 0.68 0.413 0.03 0.857 0.07 0.786

S x F F(1, 52) 40.53 <0.001 53.64 <0.001 0.00 0.983

W x F F(1, 52) 0.09 0.762 0.580 0.580 1.56 0.218

S x W x F F(1, 52) 1.43 0.237 0.216 0.216 0.12 0.726

5.4 DISCUSSION

The studies in this chapter revealed important differences between tuart and peppermint from the germination to the establishment stage and these are summarized in Table 5.16. The ecological implications arising from these differences help to explain the contrasting ability of tuart and peppermint to establish between fires (Forests Department 1979, Bradshaw 2000 and Chapter 3). The results of the field studies also reafffirmed the importance of ashbeds within burnt areas for seedling establishment.

174 Table 5.16: Comparison of selected attributes of tuart (T) and peppermint (P) seedlings from the germination to establishment phase, and the ecological implications of these differences. The statements are based on the findings of this Chapter. A particular advantage for regeneration at a recently-burnt site is denoted by ■ and no particular advantage at a recently- burnt site by □. Attribute Species comparison Ecological implications: tuart in comparison to peppermint Germination • Rate: T > P ■ ↑ exploitation of resource pulses1. ■ ↓ exposed/vulnerable to competition post-fire ■ ↓ risk of seed predation • Variability in rate: □ ↓ potential loss of germinants from unseasonal dry spells2. P > T □ ↑ chance of some seed germinating in unseasonal wet spells Seed size • T > P □ ↓ dispersal distance □ ↑ risk of predation (more visible and less likely to enter the soil) ■ ↑seedling size3.4. (see below) Seedling size • T > P ■ ↑ resource capture and less vulnerable to competition ■ ↑ ability to survive herbivory5. □ ↑ resistance to hazards generally4 Response to • T > P ■ ↑ potential for growth (and survival) on nutrient enriched sites nutrient supply (eg: ashbeds) □ ↓ tolerance of nutrient-poor/undisturbed sites 1. Pate et al. 1990 2. Grubb 1977 3. Stock et al. 1990 4. Westoby et al. 1997 5. Kitajima and Fenner 2000

The germination rate for tuart was more rapid than peppermint; advantages associated with rapid germination in the post-fire environment can be envisaged. Firstly, seedlings that begin growth earlier can capitalize on the greater availability of resources temporarily available following fire (Pate et al. 1990). Elevated concentrations of some + nutrients, such as NH4 , may be available for only a few months after fire (Stock and Lewis 1986, Wan et al. 2001). Further, as the vegetation recovers with time since fire, competition for light, water and nutrients intensifies. Secondly, more rapid germination would reduce the period to which seeds are exposed to granivorous animals. As with other eucalypts, losses of tuart seed to ants can be high, with the complete removal of some proportion of experimental batches within four days (Ruthrof 2001). The loss of peppermint seed by granivory is unknown. However, there is no reason to suspect that it would not occur. In contrast to tuart, the more staggered germination rate of peppermint has an advantage in that risk is lower of all potential germinants dying due to an extreme weather variation, such as an extended dry spell in winter (Grubb 1977).

175 The seed weight of tuart was many times greater than for peppermint and there are several implications arising from this. One is that dispersal vectors and therefore distribution patterns would differ between the species (Westoby et al. 1997). Peppermint seed, with its small mass and wing-like structure (Boland et al. 1984) would be dispersed more widely in the wind than tuart. Wind dispersal would appear to be the main dispersal mode for tuart because little dispersal of eucalypt seed by vertebrates or invertebrates occurs due to the seeds being consumed by these agents (for example: Gill and O’Dowd 1984, Wellington and Noble 1985b, Stoneman 1992, Yates 1994). Wider dispersal could improve the probability that at least some germinants would occupy niches where conditions were more favourable for establishment. This may be particularly important for establishment in unburnt areas where establishment niches are rare. However, following germination, the larger seed mass of tuart may provide it with an advantage over peppermint due to the positive relationships between seed size, seedling size and early survival (Stock et al. 1990, Westoby et al. 1997, Kitajima and Fenner 2000). In the field and glasshouse experiments, tuart seedlings were more than double the size of peppermint seedlings in the early growth stage, and under field conditions (site Y1) the survival rate of tuart was approximately double that for peppermint in the initial months after germination. In addition to being more resilient to disease, grazing and trampling, larger seedlings may develop roots that penetrate to greater depth in the soil where moisture is more abundant and less subject to sudden drying (Bowen and Pate 1991, Westoby et al. 1997).

The reports that tuart seedlings have greater survival and growth rates on ashbeds (Keene and Cracknell 1972, Forests Department of Western Australia 1979) were confirmed in this study, particularly at site GB. However, some seedlings (5 %) established off ashbeds at site GB. Thus, establishment may not be exclusive to ashbeds. The observations of the dependence of tuart regeneration on ashbeds (Keene and Cracknell 1972, Forests Department of Western Australia 1979) were made at Ludlow where a grassy layer was present due to a long history of grazing and disturbance. One important role of the ashbed was believed to be the elimination of grass competition in the first year of seedling growth (Keene and Cracknell 1972, Forests Department of Western Australia 1979). Therefore, in woodlands where the cover of annuals is not the dominant feature of the understorey, establishment off ashbeds may occur more frequently.

176 For those seedlings at site GB that established off ashbeds, the mean heights were still less than the ashbed seedlings, but not statistically different from the ashbed seedlings in the second summer after germination. This suggests either or both of the following: the waning influence of the ashbed effect after one year or the importance of growth for ensuring the survival of seedlings off ashbeds; growth to a larger size by whatever mechanism may enable a seedling to develop a deeper root system and thereby survive the summer drought (Wellington and Noble 1985a, Burrows et al. 1990) as well as buffer the seedling from the competition, grazing and disease (Westoby et al. 1997, Kitajima and Fenner 2000).

While the survival rate of newly emerged seedlings was greater for tuart than for peppermint in a recently-burnt area (site Y1), for older seedlings planted at an unburnt site (Y2), the opposite was observed. The coincidence of a rising mortality rate with the onset of the summer period implied that droughting was a significant cause of seedling death at the unburnt site. The observations made on seedlings soon after they had died supported the proposition that drought was the cause. Eucalypt seedlings are particularly vulnerable to droughting in their first year of growth, especially at undisturbed sites where competition for water can be intense (Jacobs 1955, Purdie 1977, Wellington and Noble 1985a, Stoneman 1992). While not simulating normal circumstances, particularly for tuart (Chapter 3), this experiment created the scenario where seedlings develop early in a favourable microsite at an unburnt site under ideal seasonal conditions. The patterns of survival suggested that peppermint seedlings avoided or tolerated the summer drought more effectively than tuart; this may account for the contrasting ability of the species to establish at unburnt sites.

Studies at one burnt site (Y1) suggested that the fire had little impact on soil moisture within the seedling root-zone (to 1 m) in the spring after the fire. As noted in Chapter 1, the effect of fire on soil moisture content would be greatest where fire intensity or severity is high. Overall, the fire at the study site was mild in intensity/severity, but even under small areas of ashbed, representing those patches where fire severity was high, the soil moisture levels were little different. The limited sampling conducted at one time post-fire, only allowed for an analysis which was insensitive to variations in time and space. However, as the largest change in mean soil moisture occurred between years, the effect of burning, particularly when fire intensity is low, may be comparatively minor in comparison to variations in annual rainfall.

177 The chemistry of the surface soil over the majority of the burnt area (site Y1) appeared little different from the unburnt patches in the spring following the low-intensity fire. However, the soil chemistry of the ashbeds was distinct, especially with regards to the greater concentration of available P. This result was in agreement with other studies and reviews (for example: Raison 1979, Walker et al. 1986, Grove et al. 1986, Adams + - et al. 2003). Concentrations of available forms of N (NH4 and NO3 ) were similar in the ashbed, burnt-non-ashbed and unburnt soils at the time of sampling. Other studies suggest that elevated levels of these forms of N are present initially after a fire, but then diminish in the proceeding months (Stock and Lewis 1984, Weston and Attiwill 1990, Wan et al. 2001). Therefore, as sampling occurred ten months after fire in this study, there is a distinct possibility that higher concentrations of available N existed in the months following the fire and then declined to pre-fire levels by the time of sampling.

Small, but biologically significant changes in soil chemistry following burning may be more widespread than on ashbeds only. These may only be detected by sampling the surface 1 to 2 cm layer of soil (Grove et al. 1986). Further, the distinction between ashbed and burnt-non-ashbed is an artificial one in reality. A considerable area within the burnt site was likely to have had nutrient concentrations at intermediate values between the burnt-non-ashbed patches and ashbeds that were targeted for sampling.

Nutrient enrichment associated with ashbed formation increases seedling growth (for example: Loneragen and Loneragen 1964, Chambers and Attiwill 1994, Launonen et al. 1999), but no such phenomenon occurred when seedlings were planted in ashbed soils brought into the glasshouse. Two possibilities could account for this. One was that the disturbance from handling, drying and rewetting resulted in nutrient loss from ashbed soils and/or the release of nutrients from organic matter in the non-ashbed soils. Yates (1994) speculated that disturbance may have contributed to an inability to detect an ashbed effect in a similar type of experiment. A second possibility is that the low N availability in the ashbed soils prevented any benefit from the elevated P levels. Experiments with eucalypt seedlings indicate that only a small growth response follows from additions of P without additions of N (Loneragen and Loneragen 1962, Florence and Crocker 1964). Seedlings in the field are likely to germinate earlier in the wet season when soil nitrogen levels may be higher and then contend with declining levels of availability as the season progresses. Launonen et al. (1999) argues that the development of mycorrhizal associations is important to sustain seedling growth on

178 sites where high-temperature combustion depletes soil nitrogen (orange ashbeds). Similarly, mycorrhizae may play an important role in sustaining growth on ashbeds where available P remains elevated, but where available N has declined relative to the level in unburnt soils, as was the nutrient status of the ashbed soils examined in this study within tuart woodland ten months after fire. However, it was not assessed as to whether the seedlings planted in these soils were able to develop mycorrhizal associations during the course of the glasshouse experiment.

In a glasshouse experiment involving contrasting fertilizer regimes, the growth increment for tuart was more than for peppermint with the higher fertilizer rate. This difference between the species under the fertilizer treatments may simply have reflected the size difference between the seeds and therefore the size difference between the germinants (Stock et al. 1990, Westoby et al. 1997); tuart being larger may have had a greater ability to acquire additional resources. Nevertheless, this disparity has significant implications for competition if the “pre-emptive’ exploitation of resources has a negative effect on peppermint growth and survival.

It is proposed that the results for the higher survival of tuarts on ashbeds in the field surveys and the large increment in rooting depth (around 300 %) when under the high nutrient treatment in the glasshouse experiment, can be related. Elevated levels of nutrients in ashbeds were also thought to enable E. wandoo seedlings to survive drought over the first summer by driving growth sufficient to attain an adequate rooting depth (Burrows et al. 1990). It follows that limited nutrient availability may be a factor limiting the rooting depth and consequently seedling establishment of tuart in unburnt areas. Fertilizer additions have been shown to increase eucalypt seedling survival in the field, with an increase in rooting depth hypothesized to be the underlying factor (Wellington and Noble 1985b).

The shoot:root weight ratios for both tuart and peppermint (< 2.5) were within the range of those of southwestern Australian “resprouter” species (Pate et al. 1990). ‘Resprouters’ are distinguished from ‘seeders’ on not only the regeneration mode, but also for shoot:root ratio (higher for seeders), carbohydrate storage (higher in resprouters) and overall productivity (higher in seeders) (Pate et al. 1990). The higher shoot:root ratio of the tuart seedlings may have been a factor enabling the larger growth response to the fertilizer additions than peppermint in the glasshouse experiment.

179 Additionally, the larger initial size of the tuart germinants may have contributed to a different capacity to acquire and use the additional nutrients that were available. The shoot:root ratio (on a dry-weight basis) of both the tuart and peppermint seedlings in the glasshouse experiment declined as the water supply was reduced. This is a common growth response that aids plant survival by reducing transpirational demand and/or increasing access to water (Reader et al. 1993). With a higher shoot:root ratio, tuart seedlings may be more prone to droughting than peppermint seedlings. This again points to the importance of the ashbed effect for tuart to achieve sufficient growth rates to attain an adequate root system to cope with the summer drought. However, for firm conclusions on the relative importance of the interspecific difference in shoot:root ratio, physiological traits (for example: stomatal regulation) of both species would need to be examined (Reich 2002). The contrasting specific leaf area and leaf fresh weight:dry weight between the species may be indicative of important differences in this regard.

The higher shoot:root ratio, the larger seedling size and the greater responsiveness of the seedlings to an increased nutrient supply, places tuart closer to the ‘C’ end of Grime’s (1977) ‘C-S’ plant strategy axis than peppermint. Thus, tuart is more reliant on maximizing resource uptake, attaining a large size and occupying maximal space. Therefore, it is on ashbeds where tuart establishment is most common.

In conclusion, there are clear differences between the growth and survival rates of tuart and peppermint at the seedling establishment stage. As suspected, natural tuart establishment occurs predominantly on ashbeds whereas peppermint establishment occurs more widely. The concentration of nutrients in ashbeds is proposed to drive the necessary root growth for tuart seedlings to survive the summer-drought period. The growth to a larger size may also assist survival by buffering the seedling against competition, grazing and disease. However, given that available N is little different in ashbeds by the spring following a summer fire, the formation of mycorrhizal associations may be necessary for seedlings to fully benefit from the higher available P and other nutrients in ashbeds; this requires further investigation. Regardless, artificial planting and seeding works for tuart restoration should incorporate fertilizer treatments to maximize establishment success. Some natural tuart establishment was observed off ashbeds and some artificial tuart establishment occurred in an unburnt area. Given the size and the lifespan of tuart trees, rare instances of tuart establishment between fires could still be ecologically significant. For peppermint, the greater ability to establish at

180 unburnt sites, the lower shoot:root ratio and the smaller growth increment with increased nutrient supply, confirms the more conservative ‘S’-type strategy (Grimes 1977). Thus, it is unsurprising that peppermint seedlings are abundant in long-unburnt areas where adults are also abundant. No consideration for differences in seed supply was given in this Chapter. However, this factor no doubt also influences differences in the establishment patterns between the species and is worthy of further study.

181

182 6 Chapter 6. The influence of peppermint on tuart seedling growth and survival: competition, herbivory and foliar disease

The establishment of the experiment, the seedling measurements, and the assessment and identification of forms of leaf damage in Study 2, were carried out with Paul Barber. Paul Drake contributed to the design, collection and measurements for Study 3.

183 6.1 INTRODUCTION

The increasing cover of peppermint trees (Chapter 3) could be a factor limiting tuart recruitment in the Yalgorup area as a result of competition. This concern was first raised by Backshall (1983). With the current widespread decline in tuart tree health at Yalgorup, serious implications follow for tuart persistence if tuart seedling establishment and development are being thwarted by the increasing density of peppermint.

In dry-sclerophyll eucalypt forests, the competitive effect of the overstorey on soil moisture is generally regarded as the most significant factor in the suppression of the regenerating eucalypt seedlings (Jacobs 1955, Bowman and Kirkpatrick 1986c, Stoneman 1994) and the understorey generally (Lamont 1985). The significance of nutrient competition in eucalypt forests in this regard is largely unknown (Florence 1996), although some studies have concluded it to be of only secondary importance to soil moisture (Lamont 1985, Bowman and Kirkpatrick 1986c). Eucalypts are generally light demanding (Florence 1996), particularly as seedlings progress to the sapling stage (Ashton 2000b). However, in the more open forest and woodland types of southern Australia competition for light may also be of lesser significance than that for soil moisture (Abbott 1984). While competition for one resource may be more important than another in structuring communities, studies from a range of vegetation types reveal that interactions between competition for light, nutrients and soil moisture are often influential (Coomes and Grubb 2000).

In addition to resource competition, other forms of competitive interactions between plants include direct competition (allelopathy) and apparent competition (Holt 1977, 1984, Connell 1990). Some research suggests that eucalypts and other native Australian plants have allelopathic potential (for example: del Moral et al. 1978, Withers 1978a, May and Ash 1990), but studies are yet to convincingly demonstrate direct allelopathy as a significant process in a natural community in Australia (Willis 1999). With apparent competition in the case of plants and herbivores, the scenario is presented where a change in the abundance or location of one plant species results in a change in the rate of herbivory of another due to both plant species having a common enemy herbivore (Holt 1977, 1984, Connell 1990). Presumably, this process could also apply to pathogens. More complex cases of apparent competition are also discussed in

184 Connell (1990), but are not considered here. An overall change in vegetation structure may also affect the fate of seedling regeneration indirectly as a result of micro-climatic changes. For example, insect herbivory on plants can be affected by within-site environmental variation, particularly whether plants are within shade or sunlight (Louda et al. 1990). Shade, as well as its influence on humidity, can also be an important determinant of the success of important foliar pathogens of eucalypt seedlings (Ashton 2000b). Patterns of herbivory at least, have been demonstrated as a major confounding factor in field experiments where resource competition is assessed (Ladd and Facelli 2005).

The broad hypothesis tested in this Chapter is that the density of peppermint trees inhibits tuart seedling growth and survival. The main experiment involved tuart seedlings planted in various soil/litter cover treatments under different combinations of peppermint density. Patterns of insect and pathogen damage on the seedlings were necessarily examined in order to consider these potentially masking influences separately from direct and resource competition. Complementary studies investigated the survival of tuart seedlings in relation to peppermint leaf litter cover at an unburnt site and differences in the levels of herbivory on tuart and peppermint leaves. Further background and the hypotheses for each experiment are outlined in the following section.

6.2 METHODS

6.2.1 General

Background Three field studies were conducted. Two involved tuart seedlings planted out in the field. The third study compared levels of insect herbivory on tuart and peppermint leaves from naturally growing trees.

Tuart seedlings planted into the field experiments The seedlings used in these experiments were germinated from seed that originated from the Lake Clifton area. Germination occurred on white sand or filter paper. Germinants of one to two weeks of age were placed in 64 cell trays filled with potting mix. This potting mix comprised coarse sand, composted pine bark and “coco peat” in a

185 2:2:1 mix (Richgro Garden Products, Perth) plus 4 g of dolomite, 4 g of calcium carbonate mixed thoroughly into each 9 L bucket of potting mix. Seedlings were tended under glasshouse conditions (evaporatively cooled) for five months. A solution of soluble fertilizer (Thrive®, Arthur Yates and Co. Ltd. Australia: 27 % nitrogen, 5.5 % phosphorus and 9.0 % potassium plus trace elements) was applied to foliage and soil periodically over this time. The seedlings were planted at the experimental sites when at five months of age using a “pottiputki” planting device (Lannen Plant Systems).

Statistical analyses MINITAB Release 13 (Minitab Inc. 2000) was used for all statistical tests and significance (α) was set to 0.05 unless otherwise stated. For logistic regression analyses, the results of Pearson, Deviance and Hosmer-Lemeshow Goodness of fit tests from the MINITAB output were checked.

6.2.2 Study 1: the influence of peppermint trees on the survival rate of tuart seedlings planted in an unburnt area

Background A field experiment at site Y2 (Chapter 5, section 5.2.2, Study 3) was re-analyzed using additional data on visual estimates of the cover of peppermint leaf litter. Peppermint leaf litter cover was measured as a surrogate for the presence and quantity of peppermint canopy. The relationship between mortality of the planted tuart seedlings and peppermint leaf litter cover was assessed in order to explore the effect of the presence of peppermint trees on tuart seedling survival.

Hypothesis:

• Peppermint leaf litter cover is related to tuart seedling survival.

Experimental design There were 79 tuart seedlings planted along transects at an unburnt site. Mortality at the end of the experiment was analyzed.

Experiment set up and assessments The full details of the set up and assessments for this experiment were outlined in Chapter 5, section 5.2.2, Study 3. The seedlings were planted in June 2004. Mortality was assessed at eight times over the period: monthly until January 2005, with a final

186 assessment at the end of March 2005. Peppermint leaf-litter cover as a percentage was estimated visually for the 0.5 m radius area around each seedling in March 2005.

Statistical analysis The significance of the relationship between seedling mortality and peppermint leaf- litter cover was tested using logistic regression.

6.2.3 Study 2: the influence of peppermint trees on the growth and survival of tuart seedlings in contrasting soil treatments

Background A field experiment was established to determine the effect of peppermint trees on tuart seedlings. In addition to resource competition, the possibility of apparent competition and/or the indirect influence of microclimate were considered by including an assessment of insect herbivory and foliar pathogen (Mycosphaerella cryptica) attack. Soil treatments, including an ashbed treatment, were also imposed given the importance of the ashbeds for tuart seedling survival and early growth (Keene and Cracknell 1972, Forests Department of Western Australia 1979 and Chapter 5). Further, an unburnt- disturbed soil treatment was included to test how such a treatment compares to ashbeds and untreated soils for tuart seedling survival and growth.

Hypotheses:

• With increasing peppermint density: o soil moisture is lower, o canopy cover is greater, o tuart seedling survival rates are lower, o tuart seedling growth is reduced, o the incidence and severity of foliar disease on tuart seedlings is greater, o the incidence and severity of insect attack on tuart seedlings is greater. • There are differences between tuart growth and survival on ashbeds, unburnt- disturbed soils and control (unburnt-undisturbed) soils. • There are significant interactions between peppermint density and soil treatment for survival rates, growth, and the incidence and severity of foliar disease and insect attack on tuart seedlings.

187 Experimental design The experiment was a split-plot design arranged in four blocks. Within each block, there was a replicate of each level of peppermint density: high, intermediate and low. Within each peppermint density replicate, there was one replicate of each soil treatment: ashbed, unburnt-disturbed and unburnt-undisturbed. There were eight seedlings (sub- samples) planted within each soil replicate at the start of the experiment.

Site and experimental set-up Within site Y1, an area of tuart-peppermint woodland in Yalgorup National Park (see Chapter 3 section 3.2 for site details), four areas (blocks) were identified where a pronounced gradient in the density of peppermint trees occurred within 20 to 30 m (locations specified in Appendix 5). Observations suggested that these small-scale gradients were apparently unrelated to soil conditions, slope or any other obvious factor. The reliance on the natural density of peppermint trees for the location of the peppermint density treatments prevented these treatments from being truly randomized. Each peppermint density plot was contiguous and occupied an area of approximately 20 x 10 m (Figure 6.1a). At the non-adjoining perimeters of each such plot, any large roots from trees other than peppermints were severed if within 20 to 30 cm of the soil surface in April 2004. For each plot with an intermediate peppermint competition treatment, all trees taller than 3 m in height were removed by cutting at the base and applying herbicide to the stump. All peppermint trees and seedlings were removed from the low peppermint competition plots and all roots within 20-30 cm of the soil surface of the entire perimeter of the plot were severed. No large trees of any other species were present within the treatment areas.

The soil treatment sub-plots were approximately 1 x 2.5 m and each was separated by approximately 5 m (Figure 6.1a). Coarse woody debris and existing shrubs in each sub- plot were removed. Ashbeds were created by piling available woody debris from the site, including cut peppermint stems and crowns, onto the designated subplot areas. Piles were lit in May 2004 and the majority of the debris was consumed. Combustion was complete within 24 hours leaving a layer of grey ash and charcoal. Orange colouration of surface soil, indicating high temperature combustion was present in some ashbeds but generally did not account for the majority of any subplot. Any large unburnt debris was removed from the plot surface. Unburnt-disturbed soils were created in April 2004 by removing the litter layer and raking the soil surface to a depth of approximately

188 5 cm. In June 2004, all subplots were sprayed with a 1 % solution of glyphosate (Roundup®, Monsanto) to reduce weed growth, but this measure proved largely ineffective. Hand removal of weeds immediately around each planting position occurred in block 1 in spring 2004, but not in the remaining blocks due to the time consuming nature of this activity. Seedlings were planted in each sub-plot in June 2004 with 0.5 m separating each planting position. Chicken-wire cages were placed around each seedling to prevent vertebrate damage (Figure 6.1b). A chronology of activities related to the experimental set up and measurements are summarised in Table 6.1.

10 m

2.5 m

1 m

20 m

High Intermediate Low a)

b) Figure 6.1: a) Layout of peppermint density treatments (high, intermediate and low) within a stylized example block (one of four) of the experiment (Study 2) at site Y1 showing idealized peppermint canopies (■) and where severing to 20 to 30 cm soil depth occurred for all tree roots (▬) and for all except peppermint tree roots (▪▪▪). The soil treatments (ashbed, unburnt- disturbed and unburnt-undisturbed) were imposed within the plots represented by the smallest rectangles (lightly shaded). b) View of an ashbed plot with tuart seedlings enclosed in chicken- wire cages.

189 Table 6.1: Chronology of key activities undertaken in the experiment (Study 2) at site Y1. At each measurement time, seedling survival and height, and leaf damage due to insects, Mycosphaerella cryptica and drought were assessed unless otherwise stated. Date Activity Jan 2004 Seedlings germinated and potted. April 2004 Severed roots entering experimental area in surface soil. Cut and removed shrubs from sub-plots. Peppermints (> 3 m tall) removed from within intermediate peppermint density treatments and all peppermints removed from low peppermint density treatments. Created unburnt-disturbed soil treatment. May 2004 Created ashbeds. June 2004 Sprayed sub-plots with glyphosate. Planted seedlings and installed cages to protect from large vertebrate grazers. Sept 2004 Measurement 1 (Insect and drought damage not assessed). Nov 2004 Measurement 2. Feb 2005 Measurement 3. April 2005 Measurement 4. Sept 2005 Measurement 5. Nov 2005 Samples collected for water content in soil under peppermint competition treatments. Dec 2005 Measurement 6 (seedling stem diameter and seedling canopy diameter also recorded). Images taken for the assessment of mid and overstorey canopy cover.

Measurements of seedling attributes At approximately three-monthly intervals, seedling height (to the growing tip) was measured. In addition, foliage was assessed for damage from insects, Mycosphaerella cryptica and drought. A visual rating system was used to estimate the damage from each agent as a proportion of the total leaf area of each seedling (Stone et al. 2003). At the final assessment date (December 2005), 18 months after planting the seedlings, stem diameter (at 5 cm above ground level) was measured with vernier calipers and canopy diameter (east-west dimension) was measured using a tape measure.

Measurements of peppermint density treatment effects

Gravimetric soil moisture content To test for soil moisture differences between peppermint density treatments, a minimum of 100g of soil was collected in November 2005 from 50 cm depth at a point equidistant between the sub-plots within each peppermint density treatment plot. Thus, two sub- samples were collected for each replicate of each peppermint competition treatment.

190 Each sample was immediately sealed in an aluminum tin. Gravimetric water content was determined by weighing the samples before and after drying at 105°C for 48 hours.

Canopy cover Canopy cover was assessed from digital images taken in December 2005. A digital camera (Canon Powershot A-70) fitted with a “fisheye” lens (Seimax) was leveled on a tripod at a height of approximately 1 m, and directed skyward. Cloud cover was complete at the time of photography. One photograph was taken from the centre of every sub-plot in the experiment, to give three images for each replicate of the peppermint density treatments. A black mask with a circular opening was created in PHOTOSHOP 7.0 (Adobe Systems Inc. 2002) and overlaid on each image such that central 75 % of the field of view was assessed. The proportion of canopy cover was then determined using ASSESS software (American Phytopathological Society 2002). The average cover for the set of three images taken in each peppermint density replicate was used to compare cover differences between the peppermint density treatments. Canopy- cover data included trees within and outside the main plot due to the wide field of view. Therefore, the canopies of both peppermint and other tree species were included in the images, but in general peppermint canopies were predominant.

Statistical analyses Where multiple responses (time series or different attributes) were available for testing, MANOVA was performed initially. However, for time-series data where the number of replicates changed with time due to seedling deaths, a series of ANOVA were calculated with α modified in accordance with the Sequential Bonferroni method (Holm 1979 in Quinn and Keough 2002). When significant values were calculated from MANOVA, an ANOVA was then performed for each response variable separately in conjunction with Tukey’s pairwise comparisons. With all MANOVA and ANOVA analyses, model residual plots were diagnosed for serial correlation, constant variance and normality. Data transformations were made where necessary. P values from Pillai’s trace were reported for MANOVA given that this measure is less affected by departures from normality and constant variance compared to the others (Townend 2002). Regardless, in instances where departures from model assumptions could not be rectified for MANOVA, adjustment of α to 0.01 was made to reduce the probability of Type 1 errors (Tabachnik and Fidell 1996).

191 In the MANOVA and ANOVA models involving tuart seedling growth, survival or damage from the various agents as responses, a different error term was specified for the peppermint density factor because of the split-plot design. In MINITAB, this was achieved by specifying the error term: block x peppermint density for the block and peppermint density factors. Alternatively, for ANOVA calculations this was achieved by specifying the block treatment as a random variable, with a block x peppermint density term included in the model.

Factors related to seedling mortality and leaf drought damage were examined using logistic regression. Only those periods in which either mortality or drought damage were at a maximum were examined.

6.2.4 Study 3: a comparison of the susceptibility of tuart and peppermint leaves to herbivory

Background Further measurements were made on the leaf samples collected at sites Y1 and Y3 in Study 2 of section 4.2.2 (Chapter 4) in order to assess the differences in the susceptibility of the species to herbivory. Any differences in this regard may also point to wider ecosystem implications for nutrient cycling where the relative abundance of the species is changing (Wardle et al. 1998).

Hypotheses:

• The amount of damage from leaf chewing insects is different between tuart and peppermint. • The species (tuart or peppermint) with the greatest amount of damage from leaf chewing insects has a higher specific leaf area and a lower carbon:nitrogen content.

Design There were 24 replicate leaf samples each for tuart and peppermint. Each replicate (tree) comprised ten subsamples. The total proportion of damage from leaf chewing insects, the mean specific leaf area and the mean carbon:nitrogen ratio was measured for each replicate sample.

192 Sampling and measurements The sampling details for this study were outlined Chapter 4 (section 4.2.2, Study 2). A sample of the youngest-fully-expanded (YFE) leaves from tuart and peppermint trees were collected. Differences in leaf age can have a significant bias on assessments of herbivore damage (Abbott et al. 1993); while the ages of the leaves compared between the species was not known, it was felt that the YFE leaves would be most comparable considering both age and structure. The proportion of damage from leaf chewing insects on each leaf was estimated visually and the mean proportion for the leaves in each replicate sample was calculated. The data for specific leaf area and nitrogen content (as determined in Study 2, section 4.2.2 in Chapter 4) were pooled by species, rather than separated into species x site/patch-type (burnt or unburnt). The carbon and nitrogen contents of the leaves were measured at Edith Cowan University (Joondalup) using a continuous flow mass spectrometer (model 20:20, PDZ Europa Crewe, UK).

Statistical analyses t-tests were used for the comparison of the means for peppermint and tuart after checking the data for normality and equal variance. Data were transformed where necessary, but Mood’s Median Test was used to compare data where a suitable transformation was not possible. Due to the multiple comparisons of data, α was modified in accordance with the Sequential Bonferroni Method (Holm 1979 in Quinn and Keough 2002).

6.3 RESULTS

6.3.1 Study 1: the influence of peppermint trees on the survival rate of tuart seedlings planted in an unburnt area

The pattern of survival for tuart seedlings planted in an unburnt area (site Y2) was significantly and positively related to peppermint leaf-litter cover (Table 6.2).

193 Table 6.2: Logistic regression results for tuart seedling survival in relation to estimated peppermint (Pep.) leaf-litter cover. Seedlings (n = 79) were planted at five months of age at an unburnt site (Y2) in June 2004 and survival was recorded at April 2005. G(1) = 5.377, P = 0.020. Significant values are denoted in bold. Coefficient Odds 95 % CI Variable SE Z P ratio Lower Upper Constant -1.36 0.393 -3.46 0.001 Pep. leaf-litter cover (%) 0.015 0.007 2.28 0.023 1.02 1.00 1.03

6.3.2 Study 2: the influence of peppermint trees on the growth and survival of tuart seedlings in contrasting soil treatments

Peppermint density treatments Mean canopy cover was greater with each increase in peppermint density class and the differences were significant (Figure 6.2 and Table 6.3). Mean soil moisture content at 50 cm depth in November 2005 was 3.4 % in the high density treatment, whereas levels in both the intermediate and low treatments were approximately 1 % greater, but not significantly different (Table 6.3).

50 a 45 40 35 30 25 b 20 c 15 10 Mean canopy cover (%) 5 0 High Intermediate Low Peppermint density treatment

Figure 6.2: Mean canopy cover (%) as determined by the analysis of photographs taken in December 2005 for the low, intermediate and high peppermint density treatments (n = 4) in the field experiment at site Y1. Means with different letters are significantly different using Tukey pairwise comparisons of log-transformed data. Standard error bars are indicated

194 Table 6.3: ANOVA results for canopy cover (log transformed) in December 2005 and gravimetric moisture content of the soil (50 cm depth) in November 2005 within the peppermint (Pep.) density treatments in the field experiment at site Y1. Significant values are denoted in bold. Peppermint canopy cover Soil moisture content Effect P P

Block F(3, 6) = 2.92 0.122 F(3, 6) = 4.06 0.068

Pep. F(2, 6) = 30.64 < 0.001 F(2, 6) = 2.69 0.146

Seedling survival Mean survival of the seedlings varied little between the peppermint density treatments, ranging from 66 % (high density) to 71 % (low density) by the end of the experiment (Figure 6.3). Greatest mortality occurred during the seedling’s first summer. MANOVA results for the six measurement times from September 2004 did not indicate a significant effect of the peppermint density treatment (Table 6.4). However, there was a significant difference between soil treatments, including for the final measurement time

(December 2005) following an ANOVA: MSresidual = 0.0304, F(2,18) = 12.75, and P < 0.001 for the soil factor. At this time survival was significantly higher on ashbeds (88 ± SE 4.5 %) compared to unburnt-disturbed (64 ± SE 7.3 %) and unburnt-undisturbed soil treatments (53 ± SE 8.2 %) according to Tukey’s pairwise comparisons. There was no significant interaction between the peppermint density and the soil treatments.

100 90 80 70 60 50 40 30 Mean survival (%) 20 10 0 Jul-05 Jul-04 Jan-05 Oct-05 Oct-04 Jun-05 Jun-04 Sep-05 Sep-04 Feb-05 Dec-05 Dec-04 Apr-05 Mar-05 Nov-05 Nov-04 Aug-05 Aug-04 May-05 Date

Figure 6.3: Mean survival for tuart seedlings under high (■), intermediate (x) and low (□) peppermint density treatments (n = 4) in the field experiment at site Y1 over 19 months. Seedlings were planted at five months of age in June 2004 in ashbed, unburnt-disturbed and unburnt-undisturbed soils. There were no significant differences between treatments (Table 6.4). Standard error bars are indicated.

195 Table 6.4: MANOVA results for survival (arcsine transformed) of tuart seedlings under contrasting peppermint density treatments (Pep.) and soil treatments in field experiment at site Y1 over 19 months. There were six measurement times. Seedlings were planted at five months of age in June 2004 in ashbed, unburnt-disturbed and unburnt-undisturbed soils. Significant values are denoted in bold. Due to violations of model assumptions: α = 0.01. Effect P

Block F(18, 9) = 0.880 0.611

Pep. F(12, 4) = 0.560 0.803

Block x Pep. F(36, 108) = 1.362 0.114

Soil F(12, 28) = 3.162 0.006

Pep. x Soil F(24, 64) = 1.218 0.261

Seedling growth The mean height of the seedlings under the high peppermint density treatment was less than that for seedlings under the low (Figure 6.4). At December 2005, such a pattern was also evident for crown diameter and stem diameter (Table 6.5). However, in both cases these differences were not significant (Table 6.6). Mean values for seedlings under the intermediate and low treatments were comparable. From around April 2005, a trend of increasing difference emerged between the heights of seedlings under the low and intermediate peppermint density levels compared to the high (Figure 6.4). In contrast, the disparity in height between ashbed and unburnt soils occurred during the first spring but changed little thereafter. The soil treatment had a significant influence on seedling growth (Table 6.6), including for height at the end of the experiment in

December 2005: MSresidual = 0.396, F(2,17) = 24.15 and P < 0.001 for the soil factor. The mean height of ashbed seedlings in December 2005 was 88 ± 7.4 cm, which was significantly different from the means for seedlings in unburnt-disturbed soil (52 ± 5.2 cm) and unburnt-undisturbed soil (51 ± 3.9 cm) soil following Tukey’s pairwise comparisons.

Table 6.5: Canopy diameter and stem diameter (5 cm above ground level) of two-year-old tuart seedlings under high, intermediate and low peppermint (Pep.) density treatments (n = 4) in the field experiment at site Y1. Seedlings were planted at five-months of age in June 2004 in ashbed, unburnt-disturbed and unburnt-undisturbed soils. There were no significant differences between peppermint density treatments (Table 6.6). Pep. density. Canopy diameter Stem diameter treatment mean ± SE mean ± SE low 45 ± 6.1 0.9 0.16 int. 44 ± 5.3 0.9 0.15 high 32 ± 4.1 0.5 0.10

196 100 90 80 a a 70 60 a 50 40 30 Mean height (cm) height Mean 20 10 0 Jul-04 Jul-05 Jan-05 Jun-04 Oct-04 Jun-05 Oct-05 Sep-04 Feb-05 Sep-05 Apr-05 Dec-04 Dec-05 Mar-05 Nov-04 Nov-05 Aug-04 Aug-05 May-05 Date a)

100 90 a 80 70 60 b 50 b 40 30 Mean height (cm) height Mean 20 10 0 Jul-04 Jul-05 Jan-05 Jun-04 Oct-04 Jun-05 Oct-05 Sep-04 Feb-05 Sep-05 Apr-05 Dec-04 Dec-05 Mar-05 Nov-04 Nov-05 Aug-04 Aug-05 May-05 b) Date

Figure 6.4: Mean height of tuart seedlings for the: a) high (■), intermediate (x) and low (□) peppermint density treatments (n = 4); and the b) ashbed (■, n = 12), unburnt-disturbed (x, n = 12) and unburnt-undisturbed (□, n = 11) soil treatments in the field experiment at site Y1. The seedlings were planted at five months of age in June 2004. Means for the final measurements are significantly different following ANOVA and Tukey’s pairwise comparisons. Standard error bars are indicated.

Table 6.6: MANOVA results for height, canopy diameter (di.) and stem diameter (at 5 cm above ground) in December 2005, and height over six measurement times (September 2004 to December 2005) for tuart seedlings in the field experiment at site Y1. Seedlings were planted at five months of age in June 2004 in ashbed, unburnt-disturbed and unburnt-undisturbed soils under high, intermediate and low peppermint (Pep.) density treatments. Data were cube-root transformed. Significant values are denoted in bold Height, canopy di. and Effect stem di.: Dec 05 P Height: Sept 04-Dec 05 P

Block F(9, 18) = 1.631 0.180 F(18, 9) = 2.251 0.107

Pep. F(6, 10) = 1.764 0.204 F(12, 4) = 1.702 0.322

Block x Pep. F(18, 51) = 2.384 0.008 F(36, 102) = 1.476 0.067

Soil F(6, 32) = 6.846 <0.001 F(12, 26) = 2.843 0.012

Pep. x Soil F(12, 51) = 1.035 0.432 F(24, 60) = 0.902 0.598

197 Leaf damage Damage expressed as necrotic tissue towards the margins of leaves, particularly older leaves, was observed and inferred to be drought related after noticing similar patterns in tuart seedlings deprived of water under glasshouse conditions (Figure 6.5a). While sometimes severe, this form of damage was not widespread with only 9 % of the seedlings exhibiting these symptoms by the end of the experiment in December 2005. Drought-related necrosis was highest in April 2005 when 19 % of seedlings were affected. A logistic regression was constructed to explain the patterns of observed drought damage at this time (Table 6.7). Due to evidence that the model did not fit the data initially, the factor relating to peppermint density was removed because it had the highest, non-significant P value in the model. The pattern whereby less drought damage occurred on seedlings in the unburnt-disturbed soil compared to other soil treatments was significant. Only 7 % of seedlings on this soil-type displayed drought damage, compared to 20 % and 30 % on ashbeds and on unburnt-undisturbed soil, respectively.

↑

← ↑ ←

a) b)

Figure 6.5: a) The leaf of a tuart seedling showing drought damage. b) A tuart seedling with damage from a leaf chewing insect (↑) and necrotic spots resulting from Mycosphaerella cryptica (↑).

Attack from insects and the foliar pathogen Mycosphaerella cryptica (Figure 6.5b) comprised the majority of instances where direct damage to the leaves of the seedlings was observed. Leaf chewing was the most common form of insect attack (Figure 6.5b). Blisters, tunneling, mining, skeletonizing, galls and various forms of damage from wasps were also recorded. By the conclusion of the experiment, the majority of seedlings had some level of damage from both insects and M. cryptica (Figure 6.6). Damage from insects or M. cryptica was not a significant explanatory variable of

198 seedling mortality for the period in which the greatest mortality occurred (Table 6.8); soil and block factors, exerted the largest effects on mortality.

Table 6.7: Logistic regression for predictors of drought damage (event) to leaves of tuart seedlings (n = 192) in the field experiment at site Y1 in April 2005. The seedlings were planted at five months of age in June 2004 in ashbed, unburnt-disturbed and unburnt-undisturbed soil treatments under various peppermint density treatments (not included in the final model due to the insignificant effect of this factor in the original model which displayed a poor fit to the data). For the regression: G(5) = 26.71, P < 0.001. Ref = Reference (the variable to which the other variables are compared). Significant values are denoted in bold. Coefficient Odds 95 % CI Variable SE Z P ratio Lower Upper Constant -2.60 0.637 -4.08 <0.001 Block (Ref: Block 1) 2 0.75 0.811 0.93 0.355 2.12 0.43 10.39 3 2.15 0.668 3.23 <0.001 8.62 2.33 31.90 4 1.04 0.720 1.45 0.147 2.84 0.69 11.62 Soil (Ref: Ashbed) Unburnt-disturbed -1.37 0.609 -2.25 0.024 0.25 0.08 0.84 Unburnt-undisturbed 0.42 0.436 0.96 0.338 1.52 0.65 3.57

100 90 80 70 60 50 40 30

Seedlings affected (%) Seedlings affected 20 10 0 Jul-05 Jan-05 Oct-04 Oct-05 Jun-05 Sep-04 Feb-05 Sep-05 Dec-04 Apr-05 Dec-05 Mar-05 Nov-04 Nov-05 Aug-05 May-05 Date

Figure 6.6: Proportion of surviving tuart seedlings with Mycosphaerella cryptica (□) and insect damage (■) on foliage in the field experiment at site Y1 from September 2004 to December 2005. Seedlings were planted at five months of age in June 2004 in ashbed, unburnt-disturbed and unburnt-undisturbed soils under high, intermediate and low peppermint density treatments. Data for insect damage in September 2004 was unavailable. n = 265, 251, 199, 191, 189 and 187 at each measurement time in chronological order.

199 Table 6.8: Logistic regression for predictors of tuart seedling mortality (event) in the field experiment at site Y1 between November 2004 (n = 198) and February 2005. Seedlings were planted at five months of age in June 2004 in ashbed, unburnt-disturbed and unburnt- undisturbed soils under high, intermediate and low peppermint (Pep.) density treatments. G(9) = 63.61, P < 0.001. Ref = Reference (the variable to which the other variables are compared). Significant values are denoted in bold. Coefficient Odds 95 % CI Variable SE Z P ratio Lower Upper Constant -2.58 0.619 -4.17 <0.001 Block (Ref: Block 1) 2 0.80 0.435 1.84 0.066 2.22 0.95 5.21 3 -1.66 0.609 -2.73 0.006 0.19 0.06 0.63 4 -0.73 0.483 -1.51 0.130 0.48 0.19 1.24 Pep. density (Ref: Low) Intermediate 0.25 0.422 0.59 0.553 1.28 0.56 2.94 High -0.13 0.448 -0.30 0.764 0.87 0.36 2.10 Soil (Ref: Ashbed) Unburnt-disturbed 2.19 0.551 3.99 <0.001 8.98 3.05 26.42 Unburnt-undisturbed 2.12 0.557 3.81 <0.001 8.33 2.80 24.83 Mycosphaerella cryptica -0.65 0.380 -1.71 0.088 0.52 0.25 1.10 damage (Nov 04) Insect damage (Nov 04) 0.39 0.586 0.67 0.502 1.48 0.47 4.67

The incidence of M. cryptica damage varied little between the peppermint density treatments, at least during the second half of the experiment (Figure 6.7a). During the first spring period, a lower proportion of surviving seedlings under the high peppermint density treatment were damaged by this pathogen, but this difference was not significant (Table 6.9). Soil type had a significant effect in February 2005 (Table 6.9). The mean proportion of seedlings affected in the unburnt-disturbed soils (53 ± 10 %) was significantly higher than for seedlings in unburnt-undisturbed (37 ± 9 %) and ashbed (28 ± 8 %) according to Tukey’s pairwise comparison of arcsine-transformed data. No significant interaction between soil and peppermint density treatment was apparent.

The severity of M. cryptica attack was greatest in December 2005 (Figure 6.7b). The mean level of damage was at least twice as high for seedlings under the low peppermint density treatment at this time. However, the difference was not significant (Table 6.10). Due to an inability to overcome serious violations of model assumptions or the presence of a significant imbalance in data, ANOVA was unable to be performed for the data between November 2004 and September 2005 inclusive. 200 100 90 80 70 60 50 (%) 40 30 20 10 0 Mean propoportion of seedlings affected affected seedlings of propoportion Mean Jul-05 Jan-05 Oct-05 Jun-05 Oct-04 Feb-05 Sep-05 Sep-04 Dec-05 Apr-05 Dec-04 Mar-05 Nov-05 Nov-04 Aug-05 May-05 a) Date

10 9 8 7 6 5 4 seedling(%) 3 2 1

Mean leaf area damaged affected per Mean area leaf 0 Jul-05 Jan-05 Oct-05 Oct-04 Jun-05 Sep-05 Sep-04 Feb-05 Dec-05 Dec-04 Apr-05 Mar-05 Nov-05 Nov-04 Aug-05 May-05 Date b)

Figure 6.7: Mycosphaerella cryptica damage on tuart seedling foliage in the field experiment at site Y1 from September 2004 to December 2005 expressed as a) the mean proportion (%) of surviving seedlings affected and b) mean proportion (%) of leaf area damaged per affected seedling. Means are distinguished for high (■), intermediate (x) and low (□) peppermint density treatments (n = 4). Seedlings were planted at five months of age in June 2004 in ashbed, unburnt-disturbed and unburnt-undisturbed soils. There were no significant differences between the means. Standard error bars are indicated.

201 Table 6.9: ANOVA for the mean proportion of surviving tuart seedlings with Mycosphaerella cryptica damage on foliage in the field experiment at site Y1 from September 2004 to December 2005. Seedlings were planted at five months of age in June 2004 in ashbed, unburnt- disturbed and unburnt-undisturbed soils under high, intermediate and low peppermint (Pep.) density treatments. Data for February, April and December 2005 were arcsine transformed and data for September 2005 were square-root transformed. Significant values are denoted in bold after adjustment of α using the Sequential Bonferroni method for the six measurement times. F Sept 04 Nov 04 F Feb 05 Effect df1, df2 F P F P df1, df2 F P Block 3, 6 1.09 0.422 3.11 0.110 3, 6 0.76 0.557 Pep. 2, 6 0.49 0.635 0.56 0.600 2, 6 0.02 0.981 Block x Pep. 6, 18 7.70 <0.001 3.35 0.021 6, 17 6.69 0.001 Soil 2, 18 1.98 0.167 3.69 0.045 2, 17 7.32 0.005 Pep. x Soil 4, 18 1.94 0.148 0.57 0.690 4, 17 1.45 0.260 F April 05 Sept 05 Dec 05 Effect df1, df2 F P F P F P Block 3, 6 0.96 0.468 2.23 0.181 1.87 0.235 Pep. 2, 6 0.03 0.973 1.05 0.403 2.67 0.147 Block x Pep. 6, 17 3.03 0.033 0.97 0.399 2.03 0.053 Soil 2, 17 1.02 0.382 0.25 0.954 3.50 0.118 Pep. x Soil 4, 17 0.24 0.910 0.06 0.992 1.73 0.189

Table 6.10: ANOVA for mean proportion of leaf area damaged by Mycosphaerella cryptica per affected seedling in the field experiment at site Y1 in September 2004 and December 2005. Seedlings were planted at five months of age in June 2004 in ashbed, unburnt-disturbed and unburnt-undisturbed soils under high, intermediate and low peppermint (Pep.) density treatments. Data for November 2004, and February and April 2005, were unable to be satisfactorily transformed and therefore were not analyzed. Significant values are denoted in bold after adjustment of α using the Sequential Bonferroni method for the two measurement times. Effect Sept 04 P Dec 05 P

Block F(3, 6) = 2.68 0.128 F(3, 6) = 1.22 0.381

Pep. F(2, 6) = 1.10 0.381 F(2, 6) = 2.88 0.126

Block x Pep. F(6, 14) = 1.10 0.361 F(6, 17) = 4.19 0.009

Soil F(2, 14) = 3.24 0.070 F(2, 17) = 0.06 0.942

Pep. x Soil F(4, 14) = 3.46 0.036 F(4, 17) = 1.78 0.179

The proportion of seedlings with insect damage was consistently less in the low peppermint density treatment over the monitoring period (Figure 6.8a). From November 2004 to April 2005, this difference was approximately 20 to 30 %. However, a significant difference was only detected for the February 2005 measurement (Table

202 6.11). No evidence existed for a significant interaction between the soil and peppermint density treatments.

100 a 90 a 80 70 60 b 50 (%) 40 30 20 10 0 Mean proportion of seedlings affected affected seedlings of proportion Mean Jul-05 Jan-05 Oct-05 Jun-05 Sep-05 Feb-05 Dec-05 Dec-04 Apr-05 Mar-05 Nov-05 Nov-04 Aug-05 May-05 Date a)

13 12 11 10 9 8 7 6 5 seedling (%) seedling 4 3 2 1

Mean leaf area damaged affected per Meanarea leaf 0 Jul-05 Jan-05 Oct-05 Jun-05 Sep-05 Feb-05 Dec-05 Apr-05 Dec-04 Mar-05 Nov-05 Nov-04 Aug-05 May-05 Date b)

Figure 6.8: Insect damage on tuart seedling foliage in the field experiment at site Y1 from September 2004 to December 2005 expressed as a) the mean proportion (%) of surviving seedlings affected and b) the mean proportion (%) of leaf area damaged per affected seedling. Means are distinguished for high (■), intermediate (x) and low (□) peppermint density treatments (n = 4). Seedlings were planted at five months of age in June 2004 in ashbed, unburnt-disturbed and unburnt-undisturbed soils. Data for September 2004 is unavailable. Means with different letters are significantly different after adjustment of α using the Sequential Bonferroni method; all unlabelled means are not significant. Standard error bars are indicated.

The severity of insect attack was highest in February and April 2005 and in the high peppermint density treatment (Figure 6.8b). However, there was considerable error in estimations of mean values and again no significant differences for attack severity were detected for the main treatments over the monitoring period (Table 6.12).

203

Table 6.11: ANOVA for mean proportion of surviving tuart seedlings with insect damage on the foliage in the field experiment at site Y1 from November 2004 to December 2005. Seedlings were planted at five months of age in June 2004 in ashbed, unburnt-disturbed and unburnt- undisturbed soils under high, intermediate and low peppermint (Pep.) density treatments. Data for April 2005 and December 2005 were square-root and arcsine-transformed, respectively. Significant values are denoted in bold after adjustment of α using the Sequential Bonferroni method. F Nov 04 F Feb 05 April 05 Effect df1, df2 F P df1, df2 F P F P Block 3, 6 0.61 0.631 3, 6 0.69 0.587 0.25 0.856 Pep. 2, 6 4.15 0.074 2, 6 31.07 <0.001 4.12 0.075 Block x Pep. 6, 18 0.89 0.522 6, 17 0.13 0.990 3.80 0.014 Soil 2, 18 2.43 0.116 2, 17 2.49 0.112 5.87 0.012 Pep. x Soil 4, 18 0.35 0.840 4, 17 0.66 0.630 3.28 0.036 F Sept 05 Dec 05 Effect df1, df2 F P F P Block 3, 6 3.04 0.112 4.28 0.061 Pep. 2, 6 4.72 0.056 0.43 0.669 Block x Pep. 6, 17 0.92 0.418 1.59 0.211 Soil 2, 17 0.41 0.862 2.05 0.160 Pep. x Soil 4, 17 0.44 0.775 1.28 0.316

204 Table 6.12: ANOVA for mean proportion of leaf area damaged by insects per affected seedling in field experiment at site Y1 from November 2004 to December 2005. Seedlings were planted at five-months of age in June 2004 in ashbed, unburnt-disturbed and unburnt-undisturbed soils under high, intermediate and low peppermint (Pep.) density treatments. Data were cube-root transformed except for February 2005 which were square-root-transformed. Excluding the November 2004 data, violation of the constant variance assumption occurred. After adjustment of α using the Sequential Bonferroni method, there were no significant values. Nov 04 Feb 05 Effect P P

Block F(3, 6) = 6.07 0.026 F(3, 6) = 1.23 0.379

Pep. F(2, 6) = 1.32 0.382 F(2, 6) = 3.70 0.085

Block x Pep. F(6, 16) = 1.10 0.407 F(6, 17) = 4.07 0.010

Soil F(2, 16) = 3.22 0.145 F(2, 17) = 0.29 0.761

Pep. x Soil F(4, 16) = 1.35 0.295 F(4, 17) = 1.66 0.205 April 05 Sept 05 Effect P P

Block F(3, 6) = 0.55 0.665 F(3, 6) = 0.50 0.693

Pep. F(2, 6) = 1.70 0.281 F(2, 6) = 5.71 0.683

Block x Pep. F(6, 16) = 1.72 0.180 F(6, 17) = 0.69 0.657

Soil F(2, 16) = 0.36 0.718 F(2, 17) = 0.59 0.596

Pep. x Soil F(4, 16) = 1.30 0.311 F(4, 17) = 0.49 0.743 Dec 05 Effect P

Block F(3, 6) = 5.27 0.037

Pep. F(2, 6) = 1.20 0.396

Block x Pep. F(6, 16) = 1.26 0.330

Soil F(2, 16) = 2.32 0.213

Pep. x Soil F(4, 16) = 1.45 0.263

6.3.3 Study 3: a comparison of the susceptibility of tuart and peppermint leaves to herbivory

The level of damage from leaf chewing insects was many-fold greater on young tuart than on young peppermint leaves collected from trees (Figure 6.9a). This difference 2 (Mood’s Median Test: χ (1) < 0.001) was highly significant. In comparison to tuart, the peppermint leaves had a significantly higher carbon:nitrogen ratio: t(46) = 6.99, P < 0.001 for log-transformed data (Figure 6.9 b), and a significantly greater specific leaf area: t(46) = -4.01, P < 0.001 (Figure 6.9c).

205 2.4 40 b a 2.1 35 1.8 30 o a 1.5 25 1.2 20 0.9 15 0.6 Mean rati C:N 10 0.3 b 5

Mean (%) chew proportion leaf of 0.0 0 Tuart Peppermint Tuart Peppermint Species Species a) b)

b 100 a

) 80 -1 g 2 60

40

Mean SLA (cm Mean SLA 20

0 tuart peppermint Species c)

Figure 6.9: Comparison of youngest-fully-expanded leaves from tuart (n = 24) and peppermint (n = 24) trees (3 to 15 m tall) for a) the proportion of the leaf lost from chewing by insects, b) carbon:nitrogen (C:N) ratio and c) specific leaf area (SLA). The leaves were collected at sites Y1 and Y3 in December 2005. Means with different letters are significantly different using Mood’s Median Test (a) and t-tests (b and c) with adjustment of α using the Sequential Bonferroni Method. Means for C:N ratio were log transformed prior to testing for a significant difference.

6.4 DISCUSSION

The presence of peppermint trees did not have a negative effect on tuart seedling establishment. In fact, there was some evidence to suggest that the litter and/or shade of peppermint trees had a positive influence on tuart establishment. Overall, soil factors, particularly whether or not the seedlings grew on an ashbed, were demonstrated to be the most important influence on both growth and survival; damage by insects and leaf disease had a much smaller impact in comparison. Furthermore, patterns of insect attack

206 and leaf disease were largely unrelated to either the peppermint density or soil treatments.

Although not directly assessed, peppermint appears unlikely to be strongly allelopathic in relation to tuart (see Appendix 4 for a supporting study). Willis (1999) not only argues that there is a lack of evidence for allelopathy as an important factor in native Australian ecosystems, he also proposes on theoretical grounds that direct allelopathy may only be significant between species without a history of co-evolution. Given these points and the findings of this chapter, any further work should focus on the more indirect consequences that may be connected with the type and concentration of secondary chemicals in peppermint. The much lower level of herbivory damage displayed on peppermint leaves when compared to tuart leaves could be a pointer to a higher concentration or more potent types of secondary chemicals in peppermint leaves, especially as the peppermint leaves had a higher specific leaf area. However, the higher carbon:nitrogen ratio could also be a factor contributing to their relative unattractiveness to herbivores (Landsberg 1990, Majer et. al 1992, Peeters 2002).

The apparent absence of an impact of trees on seedling establishment via resource competition has been reported in some studies in eucalypt forests (Abbott 1984, Dignan et al. 1998), but not others (Bowman and Kirkpatrick 1986a, c, Stoneman 1994). A complex of seasonal, site and species factors must account for these differences. For the seedlings planted in the unburnt area in Study 1, survival was higher where the cover of peppermint litter was greater. This may have been due to the shading from the peppermint canopies and/or mulching effect of the leaf litter. Coomes and Grubb (2000) noted that such a “nurse effect” of the overstorey at the post-emergent stage was commonly reported in dry environments and Facelli and Ladd (1999) attributed a beneficial effect of litter on eucalypt seedling emergence to the more favourable temperature and moisture regime prevailing under litter. A further experiment would be necessary to tease apart the relative importance of shade from canopy versus leaf litter in the case of tuart establishment. However, in Study 2 drought damage was higher for tuarts planted in soils where the leaf litter was not removed, survival was no different depending on whether the seedlings were planted outside or under peppermint canopy, and survival was higher overall in this experiment even when the results from the ashbed plantings were excluded. Thus, the results were somewhat inconsistent between the two experiments. Apart from unknown factors associated with site, there are two

207 other possible explanations that could account for some of the inconsistencies. Firstly, shrubs were cut and removed in Study 2 but not in Study 1. Therefore, soil moisture and nutrient availability could have been higher in Study 2. For the greater drought damage of the seedlings on the undisturbed (litter retained) soil treatment in Study 2, soil wettability factors could have been important. The undisturbed soil treatment was contrasted with ashbed and disturbed (raked and litter removed) soil beds; the wettability of the surface soil is known to be improved following heating at high temperature (Doerr et al. 2004), the mechanical disturbance (raking) could have broken up any repellent layer that was present in the surface centimetres, and water repellence is most pronounced at the end of the seasonal drought period (Carter 2002); the assessments of drought damage occurred in April following only some light, early season rain. Further work addressing the role of soil wettability and leaf-litter cover could enable the development of techniques to improve the artificial establishment of tuart. Disturbance without the removal of litter may be beneficial to seedling establishment.

With the combined level of damage from insects and M. cryptica at less than 10 % of seedling foliage for most assessment times, these agents were minor or insignificant factors in the growth or survival of tuart seedlings. Further, as there were no relationships between seedling damage from these agents and the peppermint density factor, any indirect impact of peppermint density on tuart seedlings from apparent competition or microclimate, was not evident. However, with only four replications of each peppermint density treatment, statistical power was such that only large differences between treatments would have been possible to detect.

No consistent relationships between the level of insect damage and whether seedlings were in burnt or unburnt soils, were revealed either. While the absence of a fire- herbivory relationship has also been found in other studies (Adams and Rieske 2003), others report reduced (Whelan and Main 1979, Hanley and Lamont 2000) or increased insect herbivory on trees in burnt areas in association with higher foliar nutrient concentrations (Radho-Toly et al. 2001). The experiment reported in this chapter was distinguished from those listed above by the fact that burning was highly localized and thus had little if any direct impact on the invertebrate population. One related question of interest was not resolved in the current study: does the rapid growth of ashbed seedlings from the enhanced nutrient status of the soil dilute the concentration of

208 nutrients in the foliage, thereby averting greater insect attack? Further work is needed to address this.

The effect of the soil treatments in Study 2, specifically whether the seedlings grew in ashbed or non-ashbed soils, had the greatest influence on growth and survival, at least over the first year of the experiment. This is consistent with other studies that have demonstrated the benefits to seedlings from the ashbed effect (Pryor 1963, Humphreys and Lambert 1965, Loneragen and Loneragen 1964, Renbuss 1973, Burrows et al. 1990 and Chapter 5). Surprisingly, reasonably high survival rates (> 50 %) were also attained on the unburnt soils. Reduced competition by the cutting and removal of shrub stems in the immediate plot area may have been important in this regard. The fact that unburnt soils disturbed by the removal of leaf litter and raking conferred little if any benefit to seedling survival was also unexpected; observations to the contrary have been widely reported (Jacobs 1955, Abbott 1984, Bowman and Kirkpatrick 1986a, Stoneman 1992, Florence 1996). Any such benefit may not apply to larger seedlings unless the disturbance extends beyond the surface few centimeters of soil.

The relative influence of the soil versus the competition treatments in Study 2 appeared to switch towards the end of the monitoring period, possibly due to the increasing resource demands of the seedlings and the diminishing ashbed effect. There was a trend emerging in the second summer for the growth rate of tuart seedlings to decline under the high peppermint density treatments relative to the intermediate and low treatments, although this was not supported statistically, possibly because of the low power of the experiment in relation to this factor. Regardless, others report that suppression of eucalypts by the overlying canopy becomes more important following the establishment stage (Florence 1996, Dignan et al. 1998). In similar open forest types in the same region as tuart, soil moisture limitations are the primary explanation for the suppressive effect of the overstorey (Lamont 1985, Stoneman 1994). While a large soil moisture differential at 0.5 m depth was not detected between the peppermint competition treatments at the conclusion of Study 2, smaller yet important differences could have been present, including at other sampling times and at other soil depths. As the seedlings progress to the sapling stage, competition for moisture as well as light and nutrients will no doubt intensify.

209 In conclusion, the increasing density of peppermint is unlikely to be limiting tuart seedling establishment at Yalgorup: a possibility raised in Chapter 3. Any suppressive effect of peppermint trees, if it occurs, is probably at the sapling stage, but long-term studies are needed to confirm this. Once again, the ashbed effect was shown to be the most influential factor in tuart seedling survival and growth during the first year. However, when tuart seedlings are planted, survival rates at the end of the first summer in even unburnt areas may be sufficient to warrant restoration plantings of tuart in declining woodlands in the absence of fire. Planting in gaps, away from shrubs and where a leaf litter layer is present, may increase the survival of such seedlings, although measures would be necessary to address weed competition at some sites.

210 7 Chapter 7. General Discussion

211 This thesis has demonstrated direct and indirect influences of fire on tuart health and confirmed the important role for fire in tuart regeneration. Sites subjected to extreme fire regimes, either a long-absence of fire or frequent wildfire, had the poorest tree health of the sites examined within Yalgorup. However, the data indicated that decadal low-intensity fire has not prevented tuart tree decline, highlighting the probable role of non-fire factors in the decline process; the increasing density of the peppermint midstorey within Yalgorup appeared to be one factor associated with poorer tuart tree health. Most tuart and peppermint saplings and trees were shown to recover after low- intensity to moderate-intensity fire. The greatest number of deaths observed due to fire occurred for small saplings and seedlings in areas where fire severity was higher. However, some large unhealthy tuart trees were also killed where fire intensity was high. For tuart in patches where fire intensity was low, an increase in tree canopy area was observed in the year after the fire, implying tree health can benefit from the indirect effect of fire. The studies also confirmed the importance of fire in tuart establishment, in contrast to peppermint establishment which can occur between fires. Experiment and observation suggested the nutritional stimulus of the ashbed is critical for tuart seedling establishment. In the first one to two years of tuart seedling growth in a field experiment, the presence or absence of the peppermint midstorey was shown to be a minor factor in comparison to whether or not the seedlings were on ashbeds, although growth trends suggested that the peppermint midstorey was beginning to exert a suppressive effect as the seedlings approached two years of age.

As with the tuart woodlands at Yalgorup, declining tree health and denser understorey growth has occurred in forests and woodlands in eastern Australia (Withers and Ashton 1977, Ellis 1985, Jurskis and Turner 2002) and extensively in North America (for example, Fule and Covington 1998, Gilliam and Platt 1999, Abrams 1992 and 2003, Arno and Fiedler 2005) in association with altered fire regimes, particularly a decline in fire frequency. The Australian and American circumstances share obvious similarities in that indigenous peoples applied fire intensively within the landscape until European settlement (Pyne 2001). The critical role of regular fire in the maintenance of the structure and composition of a range of ecosystem types across the globe, including savannah systems, has been confirmed by recent field research and modeling (Bond et al. 2005). The term fire dependent ecosystem (FDE) has been used to describe such ecosystems (Chandler et al. 1983, Bond et al. 2005). Certainly, tuart in the savannah- like formations that apparently existed at the time of European settlement and

212 maintained at Yalgorup into the last century was a FDE. But, are tuart woodlands generally FDE? Due to the strong link between fire and seedling regeneration, like many other eucalypts, tuart persistence is undoubtedly more assured in fire prone environment. In addition the possibility of premature tree decline, through competition or changes in the soil ecosystem in association with a denser understorey (processes yet to be understood), is reduced by frequent, low-intensity fire restricting understorey development. In the presence of peppermint, or any other species with both a contrasting regeneration ecology and the capacity to attain a large size in a competitive environment, tuart woodlands are probably at their most fire-dependent.

When the traits of tuart and peppermint are considered within the framework of plant ecology strategy schemes (Westoby 1998), it is clear why peppermint is favoured by long periods without fire or moderately-frequent intense fire, relative to tuart (Figure 7.1). As seedlings, peppermint are closer to the S (stress-tolerator) strategy of Grime (1977) due to their lower shoot:root ratio (Chapter 5) and the greater ability to establish from seed and continue to grow at unburnt sites (Chapters 3 and 5). The higher specific leaf area of peppermint in Westoby’s (1998) Leaf Height Seed (LHS) scheme (Figure 7.1) could be suggestive of greater shade tolerance than tuart, thereby contributing to the contrasting establishment patterns, although physiological studies would also be needed to support this (Atwell et al. 1999). The apparent C (competitor) strategy (Grime 1977) of tuart is in common with other non-lignotuberous eucalypts that have potential for rapid early growth (Ashton 2000b). Rapid early growth is critical for tuart to attain sufficient rooting depth and thereby access an adequate supply of water to ensure seedling establishment; nutrient enriched ashbeds are important in this regard (Chapter 5). There is also a marked difference in the size of tuart and peppermint seedlings, reflecting the contrasting C and C-S strategies. The greater seed weight and more rapid germination rate of tuart (Chapter 5) sets up this difference from an early stage, bestowing tuart seedlings with the size to maximize the acquisition of the abundant resources of the post-fire environment. At the sapling stage, continued growth to access light and further drive the necessary growth to progress to the tree form, is critical for most eucalypts (Florence 1996). The attainment of a large size, particularly height, is important for tuart from the perspective of buffering it from interspecific competition, but also allowing it resistance to and/or resilience with low-intensity and moderate- intensity fire.

213

1. Grime’s CSR

C 2. Westoby’s LHS X Tuart

X Peppermint

SR Tuart ●

Height Peppermint ●

SLA 3. Noble and Slayter’s Vital Attributes Tuart Peppermint P P Seed weight

EEJJ EEJJ

M M

Figure 7.1: The plant ecology strategy schemes of 1) Grime (1977), 2) Westoby (1998) and 3) Noble and Slayter (1980) incorporating tuart and peppermint based on the information from this thesis and elsewhere. For 1) C = competitor, S = stress tolerator and R = ruderal. In 2) SLA = specific leaf area; the intention of Westoby (1998) was to quantify the differences between species, but only a qualitative representation of differences is shown here. For 3) P = propagules, J = juvenile, M = mature, E = extinct; the arrows represent outcomes in the absence () or presence of high-intensity (----) or low-intensity ( ) fire, although Noble and Slayter (1980) referred to disturbance generally, as opposed to fire only. In addition, although they discussed the importance of the intensity of disturbance, they did not distinguish it in their model representations.

Tuart and peppermint can be fitted to, and contrasted under, modified versions of models within the Vital Attributes scheme of Noble and Slayter (1980). These models illustrate the response of populations to disturbance based on the key features of species relating to survival and regeneration (by seed or vegetatively) by life-stage (Figure 7.1). In acknowledgement of the contrasting mode of establishment, it is indicated that tuart can be lost from an area in the absence of fire (or other disturbance) beyond the lifespan of the species, whereas peppermint will persist. The other major difference relates to the fire response of the adults. By nature of the thicker bark (Chapter 4) and taller height at maturity (Boland et al. 1984), tuart has a greater capacity than peppermint to maintain 214 reproduction and growth following low intensity fire. Chapter 4 also suggested that there is a difference in the fire response of the juveniles, tuart being more fire resistant.

7.1 MANAGEMENT

Tuart health, tuart health research, the conservation of biodiversity and possibly wildfire control would all benefit from reinstating a diversity of fire regimes within the remaining area of tuart woodlands (Table 7.1). In particular, areas subject to frequent (three to five years) low intensity burning should be restored to a portion of the landscape. These areas should encompass both healthy and declining stands in order to address the research questions outlined above as well as the proposition of Jurskis and Turner (2002), Jurskis (2005) and Turner and Lambert (2005) that an excess of soil moisture and nutrients following a reduction in burning frequency underlies eucalypt decline. The reintroduction of frequent, low-intensity burning would be a means by which the tuart savannah-like formation maintained by fire could be revived. Mature tuart, with taller stems and with thicker bark than mature peppermint trees, would largely retain intact canopies and therefore also seed production capacity under low- intensity fire. On the other hand, regular scorching to around 10 to 15 m could reduce and maintain peppermint canopies to within the scorch-zone; once within the scorch- zone, the regular return of fire would limit peppermint leaf area and seed production capacity. Widely spaced trees with a patchy, low understorey may be the most compatible formation for tall trees if the current downward trend for rainfall continues. The reappearance of tuart in savannah-like formations would also increase structural diversity across the landscape, supporting species conservation at the regional scale. Additionally, these areas would represent low-fuel buffers which would assist wildfire control operations and reduce the likelihood of destructive wildfires further damaging the health of tuart trees across extensive areas. Ideally, these frequently burnt areas would be strip-like to maximize both the ecotonal effect between the contrasting vegetation structures and the distribution of the low-fuel buffer.

215 Table 7.1: A representation of the aerial extent of fire regimes prior to European settlement (18th century), and at present (21st century): a greater extent is represented by darker shading. Fire regime1. Pre-European Current Proposed 1. Frequent, low intensity2. 2. Moderately frequent, low intensity 3. Moderately frequent, intense 4. Infrequent, intense 5. Variable: shifting between regimes3. 1. For frequency, the return interval for fire = every 1 to 5 years (frequent), every 6 to 15 years (moderate frequency) and > 16 years (infrequent). 2. Sites burnt at high frequency should occur in strips to maximize the ecotonal effect and the distribution of low-fuel buffers 3. Arguably this has occurred across extensive areas following Europen settlement by way of a shift from regime 1, to 2 or 4. For the “proposed” column, there would be no shift to regime 3 given the likely negative impact of this regime to tuart health and possibly other plant and animal species, especially given the current highly-fragmented nature of the ecosystem.

The increased density of the understorey at Yalgorup, or in tuart formations elsewhere, may now not support regular, low-intensity fire. Due to the increased shading of the litter layer under the denser understorey, fire may only carry under dry-summer conditions and then burn intensely because of the elevated-fuel layer. Where tuart regeneration is urgently needed to replace declining adults, establishment of tuart may not be impeded (Chapter 6), but as tuart is a gap regenerator, the sapling-tree transition is unlikely to occur under a dense midstorey (Forests Department of Western Australia 1979, Bradshaw 2000). Therefore, more active intervention by way of vegetation removal will be needed initially in some areas, including those areas where a more rapid change in woodland structure is desired. Bradshaw (2000) outlined a range of prescriptions for restoration treatments involving understorey clearing, and these should be trialed. Where the trees are in poor health, the reintroduction of frequent burning would need to be delayed until seedlings (planted or from sown seed) have developed sufficient size to resist or recover rapidly following fire. The results in Chapter 4 suggested that once developed as a small tree with a stem diameter > 6 cm at breast height, tuarts should be very resilient with low-intensity to moderate-intensity fire. The examination of growth rings at Golden Bay, north of Mandurah, indicated that tuarts have the potential to grow to this size within eight years (Chapter 4).

Should intense wildfires occur within areas of severe decline at Yalgorup, then these events should be viewed as restoration opportunities. In the United States, severe fire is 216 considered to be one possible tool for restoring Pinus ponderosa forests by reducing vegetation density and increasing species diversity (Fule et al. 2004). Ashbeds should be planted or seeded to ensure tuart can once again dominate in areas where seed production is poor or no longer occurs due to declining tree health. The recovering shoots of resprouting midstorey species such as peppermint could be treated chemically to arrest the redevelopment of the midstorey until the tuarts are sufficiently developed for fire to be reintroduced. Following an intense wildfire, recovering foliage would be arising from the lignotuber and thus could be treated relatively efficiently in this way. If vigorous resprouting continues, the treatment could be limited to areas immediately around the regenerating tuarts.

From the perspective of general fuel-reduction burning and tuart health issues, the current practice of periodic controlled burning need only be modified to the extent that the seasonal bias should be towards autumn for those areas with declining tuart trees that still produce seed. As discussed in Chapters 3 and 5, autumn fires would increase the chance of seedling establishment. However, if tuart health responds positively in areas subjected to frequent, low-intensity fires, then the current frequency of burning for broad-scale fuel reduction should be increased in a proportion of tuart woodlands where this practice would not threaten the persistence of other species. Where controlled burning is carried out in declining areas, planting or seeding of tuarts in ashbeds in gaps or under declining trees, would be desirable.

The maintenance of long unburnt areas will also continue to be an important component of the modern tuart landscape, maintaining diversity at the regional scale and providing reference sites for assessing the impact of management practices. From a tuart conservation viewpoint, it would be desirable for only healthy stands to exist within long-unburnt areas as without fire or human intervention, declining tuarts over a vigorous midstorey would not be replaced by natural seedling regeneration. On the other hand, from a scientific and historical perspective, should we not also have a diversity of successional examples within the landscape? Like any other community, the values of an ex-tuart, peppermint climax-community of recent origin deserve consideration.

Where fire is not desirable (for example, because threatened fire-sensitive species, research projects or a weed problem exists in an area), tuart restoration goals could still

217 be achieved, although certain measures would need to be implemented to ensure the adequate regeneration of tuart seedlings. In Chapter 6, it was demonstrated that > 50 % of planted tuart seedlings established at one site in unburnt soil. Annual weeds were not a major problem at this site and some control of shrub species around the planting area was carried out. These factors will no doubt inhibit establishment at other sites. A passive restoration approach may succeed where the strategic planting of seedlings in understorey and overstorey gaps and by targeting favourable microsites such as depressions or the leeward side of logs. Fertilizer could be added to the planting spots to replicate the nutrient stimulus of the ashbed, ensuring that growth is sufficient to ensure adequate rooting depth (Chapter 5). In areas where a dense layer of annuals is present, herbicides would need to be applied at the planting spots. Where natural gaps are absent at a site, tree and shrub removal around planting areas would be required for saplings to develop as trees, as trials at Ludlow attest (Forests Department of Western Australia 1979) and as the results of Chapter 6 hint. Given the size to which a tuart tree can grow, as few as 15 to 20 well dispersed, vigorous saplings ha-1, may be sufficient to eventually restore a tuart woodland. Thus, a strategic, intensive yet small-scale effort in the absence of fire may be effective in some areas.

7.2 FURTHER RESEARCH

Several important research questions need to be addressed for tuart, and other eucalypt forests and woodlands that have declined with changing fire regimes. These all concern the significance of processes that may accelerate the premature mortality of the eucalypt dominants. Evidence for a significant negative relationship between tree health and peppermint density was presented in Chapter 3. Thus, the role of competition by understorey vegetation for water and nutrients in determining the condition of tuart trees needs detailed study; long-term manipulative field experiments will be the most informative. In Chapter 4, tree canopies with little direct damage from fire were observed to have grown larger in the year following burning than comparable trees in an unburnt area. The soil analysis results in Chapter 5 and the general literature (Chapter 1), point to increased concentrations of available nutrients in the year after fire, especially for phosphorous. Therefore, further research should examine if the recurrence of the post-fire resource pulse, particularly in relation to nutrients, is necessary to sustain the long-term growth and functioning of tuart trees at some sites. The importance of such a process has been proposed for the long-lived, co-occurring 218 Xanthorrea preissii (Swanborough et al. 2003). Both research questions posed above may be intertwined; greater access to water, possible under an undeveloped understorey, may be necessary before pulses of greater plant-available nutrients can be fully exploited. A final and also not unconnected avenue of research that should be pursued, concerns the impact on nutrient cycling from changes in the quality and quantity of leaf litter accompanying the increase in peppermint cover at Yalgorup. Contrasts in leaf chemistry and the level of damage from insect herbivores between tuart and peppermint leaves (Chapter 6) could be reflected in different rates of leaf decomposition and alternative associations of microbiota (Wardle et al. 1998). The type of mycorrhizal associations that develop in association with peppermint dominated litter, as opposed to litter originating from several species, would ideally be investigated given the links made between mycorrhizal associations and eucalypt decline (Ellis and Pennington 1992). In addition, the work of Ellis (1981) and Ellis and Pennington (1992) implies that the influence of fire on the types of mycorrhizal associations that develop with tuart should also be studied and related to tree health.

7.3 CONCLUSION

Tuart woodlands have coexisted with, if not been maintained by, thousands of years of direct human intervention in the disturbance regimes, namely the fire regime. These woodlands may or may not be FDE, but frequent low-intensity fire is highly compatible with their persistence. Today’s fragmented tuart woodlands, in declining health in some areas, and subject to threats such as invasive weeds, introduced disease and possibly climate change, are now undoubtedly management dependent ecosystems. Fire, the ancient tool, will have an important role in the future management of tuart woodlands.

219

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