Determining the effects of drought and fire on regeneration in the Northern

This thesis is presented for the degree of Bachelor of Science (Honours)

TRAVIS RASMUSSEN

BSc ENVIRONMENTAL MANAGEMENT AND SUSTAINABILITY

College of Environmental and Conservation Sciences

Supervisors: Dr Joseph Fontaine, Dr Lewis Walden

and Dr Katinka Ruthrof

This document is referenced according to Forest Ecology Management journal format

31ST OCTOBER 2020

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DECLARATION.

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

I state, all previous data used was with the permission and approval of the researchers, Lewis Walden and Katinka Ruthrof.

Travis Michael Rasmussen

Word count: 19908 (Excluding reference list)

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ACKNOWLEDGEMENTS Firstly, I would like to thank my supervisors Joe Fontaine, Katinka Ruthrof and

Lewis Walden. Each of you has individually guided my learning experience throughout my education at Murdoch University since 2016. For this, I am forever grateful for the opportunities you have presented me, the knowledge you have passed down and the experiences we have shared.

Secondly, I would like to thank Billi “Nighthawk” Veber, the experience you have passed down to me with patience and understanding over the years has increased my knowledge on identification, field surveying techniques and data collection has allowed me to appreciate the environment and what is required so much more

A very big thankyou to volunteers and friends, Nate Anderson, Ebony Cowan,

Jason Paterson, Tom Mansfield and Luisa Ducki for keeping me on track during my thesis and throughout my undergraduate major. I would not have been able to get through the years without your help, you are all amazing

I want to thank the Department of Biodiversity, Conservation and Attractions and

Trillion for the scholarship and financial support, which allowed me to undertake this project.

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ABSTRACT Forest ecosystems in fire-prone regions globally have a high level of resilience.

However, as the climate changes, the level of frequency and severity of disturbances is expected to increase. This may influence the likelihood of multiple disturbance events occurring and impeding the ability of a forest to regenerate. Tree regeneration has been previously examined, although evidence-based data assessing tree regeneration following multiple disturbances are rare, limiting the capability to predict forests recovery and persistence once a disturbance occurs.

This study expanded upon two earlier studies to quantify the regeneration of overstorey and midstorey species ten years after drought-induced canopy die-off and three years following a mixed-severity wildfire that burnt under extreme fire weather. The effects of prescribed burning were also examined by including prescribed burning activity of the same time since fire as the wildfire event. This study investigated three main questions. (1) How is tree regeneration affected by drought and fire? (2) How do growth stages of tree regeneration vary by drought and fire severity? Lastly, (3) How does tree regeneration respond to fire over time?

The study found that firstly, under a drought and fire interaction, regeneration density significantly increased, secondly, lignotuberous seedlings and seedling coppice were the most dominant growth stage identified with significantly higher densities following moderate severity fire. Thirdly, moderate severity fire displayed higher densities over the three years since wildfire, however, prescribed burning and high severity fire showed less regeneration abundance by comparison. This study highlights the need for future research in prescribed burning techniques and an increase in forest management practices due to the increased risk of higher fire severities. Furthermore, it illustrates the necessity of understanding the long-term trajectory of tree regeneration following multiple disturbance events.

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CONTENTS

Declaration...... i Acknowledgements ...... ii Abstract ...... iii 1 Introduction ...... 1 2 Literature Review: Influence of drought and fire on forest ecosystems – with reference to prescribed burning ...... 5 2.1 Drought effects on forest dynamics ...... 5 2.2 Fire regimes ...... 9 2.3 Prescribed burning in ...... 17 2.4 Eucalypt species and adaptive traits to disturbance ...... 20 2.5 Conclusion ...... 26 3 Methods ...... 28 3.1 Study area: The Northern Jarrah Forest...... 28 3.1.1 Location and vegetation distribution ...... 28 3.1.2 Geology and topography ...... 29 3.1.3 Climate ...... 30 3.2 Study design ...... 33 3.2.1 Site selection and plot establishment ...... 33 3.2.2 Plot measurements ...... 38 3.2.3 Statistical design ...... 41 Results ...... 45 4.1 How is tree regeneration affected by drought and fire? ...... 45 Tree regeneration of all study species following drought and fire... 45 Overstorey tree regeneration following drought and fire ...... 47 Midstorey tree regeneration following drought and fire ...... 48 4.2 How do growth stages of tree regeneration vary by drought and fire severity? ...... 50 How drought and fire influence frequency distribution ...... 50 How drought and fire influence resprouting stages three and four .. 52 4.3 How does tree regeneration respond to fire over time? ...... 56 Prescribed burning regeneration change at time since fire ...... 56 Wildfire severity regeneration change and time since fire...... 59 Discussion ...... 64 5.1 How is tree regeneration is affected by drought and fire?...... 65 Drought and fire influence on forest level tree regeneration ...... 65 iv | P a g e

Drought and fire influence on overstorey species regeneration ...... 66 5.2 How do growth stages of tree regeneration vary by drought and fire severity? ...... 68 Overstorey tree regeneration of germinants and lignotuberous seedlings following drought and fire - stages one and two ...... 68 Overstorey tree regeneration of seedling coppice and ground coppice following drought and fire - stages three and four...... 69 5.3 How does tree regeneration respond to fire over time? ...... 70 Tree regeneration response to time since prescribed burn ...... 70 Tree regeneration response to time since wildfire ...... 71 Overstorey species regeneration response to time since fire for germinants and lignotuberous seedlings – stage one and two ...... 73 5.4 Midstorey species response to drought and fire ...... 74 5.5 Limitations and management implications ...... 75 5.6 Conclusion ...... 79 Appendix A –Studied stages of growth ...... 80 Appendix B – Studied species ...... 81 References ...... 82

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1 INTRODUCTION Climate change is expected to alter the structure, composition and functionality of forest ecosystem suitable for specific plant species (Thomas et al., 2004). The alteration of forest ecosystems is primarily driven by the level of disturbance (intensity and duration) (Buma et al., 2013). As the climate changes, natural disturbance regimes (e.g. patterns of severity, frequency, and timing (Seidl et al., 2011)) are increasing, however, the uncertainty is, whether and how ecosystems recover (that is, their resilience) their structure and function after disruption (Turner, 2010). Disturbance regimes are described as the space-time characteristics of the disturbance agent (e.g. drought, fire, pathogen outbreaks) (Loehman et al., 2018). Nonetheless, disturbances can fundamentally change ecosystem states and increase the likelihood of degradation and more disturbances

(Scheffer et al., 2001). Understanding the effects of changing disturbance regime and how they alter ecosystem dynamics is critical for conservation and management (Turner, 2010;

Trumbore et al., 2015).

The severity and frequency of disturbances are often climate-sensitive and appear to be increasing under climate change (Dale et al., 2001a). Disturbance patterns are influenced by the climate inducing an effect over entire biogeographic regions (e.g. altered wildfire regimes in the Rocky Mountains (USA), due to higher than average temperatures

(Loehman et al., 2018), variation in community abundance in swamp forests (West

Africa), due to rainfall reduction and temperature rises (Igu and Marchant, 2018) and an increase in post-fire mortality of resprouting vegetation in Mediterranean-type ecosystems (southwestern Australia), following fire disturbance (Nicholson et al., 2017).

Forest structure, composition and function contribute considerable amounts of unpredictability to regional disturbance regimes (Batllori et al., 2017; Lucash et al., 2018).

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Therefore, multiple disturbance interactions can impede the ability of forests to recover and hence affect long-term community trajectories (Lamothe et al., 2019)..

Drought induced canopy die-off has the potential to severely affect forest structure, composition, and functioning. Drought disturbances have been recorded in numerous forest ecosystems (Allen et al., 2010; Allen et al., 2015; McDowell et al., 2020). Canopy die-off events caused by drought have been shown to impact the forest dynamics, leading to structural change or accelerating successional dynamics (Clark et al., 2016; Martínez-

Vilalta and Lloret, 2016). Forests stands can experience slower growth rates from alterations to trees, recovery abilities, including delayed effects and changes to their level of resilience following droughts (Trouvé et al., 2017; Vitali et al., 2017). Further, the type,

(e.g. drought or pests), the duration and the frequency of disturbances can also impede the tree recovery process following additional disturbances (Dale et al., 2001a). For example, in the Mediterranean-type climate of Catalonia, in northeastern Spain, Serra-

Maluquer et al. (2018) studied forests of mixed pine species following three consecutive droughts events. The affected trees revealed that recovery, resilience and the resistance varied between each drought event, indicating the forest population recovered slower, was less resilient and less resistant after each drought occurrence. The consequences of the weakening of the trees resistance can increase tree mortality and magnify the susceptibility of other disturbances such as pathogen outbreaks (Caldeira, 2019) and fire

(Pausas and Fernández-Muñoz, 2012). These changes in forest resilience levels and recovery ability need to be addressed and closely monitored to understand the long-term trajectory and response of forest health with a drying climate.

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Fire, as a disturbance mechanism is an important contribution to many forest ecosystems

(Bowman et al., 2009), highly affecting plant the type of vegetation cover and the level of biodiversity (Whelan, 1995; Taylor et al., 2014). Climate models have predicted the frequency, duration and intensity to change with a shifting climate (Jolly et al., 2015).

Frequent fires can reduce plant regeneration and influence change in species assemblages.

The shifts in disturbance patterns can assist species to develop a forest composition that may be suited to the current disturbance regime, such as a fire-prone environments

(Donato et al., 2016). Drought occurrences expected to increase, this can exacerbate these changes, influencing further challenges of forest recovery and regeneration ability (Littell et al., 2016). For instance, the post-fire and post-drought regeneration study conducted by (ref) in the fire-prone subapline forests of the Rocky mountains (USA), Harvey et al.

(2016) saw reduced tree seedling establishment following a severe post-fire drought. The cause of the seedling reduction may be linked to the distance between seed sources such as mature trees for the studied species (Harvey et al., 2016). The post-fire seedling regeneration and establishment of tree species may continue to shift, inducing a compositional change in the forest environment. Therefore, further studies of the interaction of disturbance events may be beneficial in understanding forest recovery and regeneration persistence following such disturbances.

In southwestern Australia, the Northern Jarrah ( Sm) Forest (NJF), a Mediterranean climate-type forest, experienced long periods of drought, causing chronic water stress in tree species over a 30-year period (BoM, 2011). During this period an acute drought event in 2010 occurred with a series of heatwave events in early 2011

(BoM, 2011). The event caused mass canopy collapse, over 16, 515 ha in early 2011

(Matusick et al., 2013). The region also experienced a drought in 2015: with annual rainfall for the Dwellingup region approximately 35% below the long-term average 780 3 | P a g e mm, (McCaw et al., 2016). On January 5th 2016, a lightning strike ignited the Yarloop wildfire, burning approximately 70 000 ha of native forest, farmland and the small town of Yarloop (McCaw et al., 2016). From 2016 to 2018, annual prescribed burning practices were undertaken by the Department of Parks and Wildlife, now Department of

Biodiversity, Conservation and Attractions (DBCA) within areas that had been subject to the drought-induced die-off event, in the NJF. The combination of different fire severities and prescribed burns within drought-affected areas presented a valuable opportunity to determine the effects of drought and fire on tree regeneration in the NJF. This study aims to quantify differences in the regeneration of the two main overstorey and three main midstorey canopy species in the NJF. By focusing on three fire severities (prescribed burn, and moderate and high severity wildfire) and the factors that can also influence the forests ability to regenerate (time, fire return intervals and drought) after multiple disturbances.

Primarily my thesis will focus on three key research questions:

1. How is tree regeneration affected by drought and fire?

2. How do growth stages of tree regeneration vary by drought and fire severity?

3. How does tree regeneration respond to fire over time?

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2 LITERATURE REVIEW: INFLUENCE OF DROUGHT AND FIRE ON FOREST ECOSYSTEMS – WITH REFERENCE TO PRESCRIBED BURNING

2.1 DROUGHT EFFECTS ON FOREST DYNAMICS

The Anthropocene epoch, is defined as the beginning of near-permanent change from human impact on the world’s biosphere including human induced climate change (Lewis and Maslin, 2015). This new period is also dominated by climate change that is recognised to impact precipitation, evaporation, temperatures and carbon dioxide (CO2) fluxes (IPCC,2013). The changes are predicted to influence environmental drivers, triggering ‘impact cascades’, a series of independent ecological disturbances (e.g. drought, fire, pathogen outbreaks) compounded across ecosystems and creating an interacting effects (Cramer, 2014). Due to these compounding disturbances, a feedback loop is created influencing the frequency, intensity and duration of disturbance(Allen et al., 2010). As a relevant example drought events have been recognised to create a feedback loop to ecosystem services (Allen et al., 2010). They are often a consequence of rainfall reduction over prolonged periods, that are commonly accompanied with extreme heatwave events, that can then lead to water stress in (Allen et al., 2010).

The ramifications of drought weaken tree vigour, allowing for potential insect infestations, or disease (pathogen outbreaks) and increased risk of fire (Allen et al., 2010).

A global review suggests there is growing evidence that forest mortality, caused by drought events is increasing with the changing climate Allen et al. (2010). The increase of drought frequency has a wide range of negative consequences, such as lowering species, ability to withstand pathogen outbreaks following the drought event (Allen et al.,

2010). For instance, the pine moth (Thaumetopoea pityocampa) outbreaks in the

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Mediterranean climate of Andalusia (southern Spain) were studied by Hódar et al. (2012).

The study showed when experiencing low rainfall and the pine moth outbreak, three of the five pine species examined (P. sylvestris (6%), P. halepensis (15%), P. thunbergia (20%)) had a significant defoliation decrease in defoliation.. However, the

Pinus pinea and P. pinaster showed limited defoliation loss under the same disturbance conditions. Inclusively, pathogens can produce systemic effects on trees roots and carbohydrate storages (Klein, 2015), a consequence of warming and drying, independently affecting biotic stressors (Klein, 2015). Some pathogens may decline with drying environments, whereas some bark beetle species (Dendroctonus rufipennis), may benefit from warming temperatures and increase the size of the population (Bentz et al.,

2010). Increased pathogen and drought events may cause negative changes to forest structure, leading to increased solar radiation reaching the forest floor, thus, increasing drought severity (Allen et al., 2010; Allen et al., 2015). The increase in drought events and pathogen outbreaks may produce long-term ramifications on the forest ecosystem.

Therefore, more research in the field of disturbance interaction will deepen our understanding on the flow-on effects of tree mortality for forest dynamics.

Long-term (decadal) forest ecosystem response to drought and heatwave events has not been widely investigated. Understanding the patterns and causes of changes in forest structure and the responses of forests to long-term drought events is important in determining forest adaption strategies to global climate change (Anderegg et al., 2013).

Large trees comprise a large proportion of the forest structure and functionality (Slik et al., 2013); however, they are also expected to be impacted by drought-induced tree mortality at a greater scale (Lutz et al., 2009). For example, McIntyre et al. (2015) used historical data (over 70 years) from regions that has experienced drought stress in the

Mediterranean climate region of California, USA. The study compared all tree species in

6 | P a g e a mixed oak (Quercus) and pine (Pinus) forests, revealing larger trees were between 30% to 50% lower in density than recorded in historical records, indicating the larger trees were more at risk of drought induced mortality than smaller or immature species

(McIntyre et al., 2015). The density differences may also be associated with fire suppression, and a decrease in moisture between individuals during drought periods, thus contributing to the decrease in density of larger tree density loss. In contrast, Hanson et al. (2001) measured the stem diameters of Acer rubrum, Cornus florida, Liriodendron tulipifera, Nyssa sylvatica, Q. alba and Q. prinus in Walker Branch Watershed,

Tennessee (USA), a humid subtropical type climate. The study showed juvenile trees had a significantly lower stem diameter than to the control plots, higher rates of tree mortality

(7.2%) following drought events, than larger trees which remained unaffected. A possible cause for the stem decline and mortality rates could potentially be poorer access to below groundwater supply, a resource accessed by larger root systems (Hanson et al., 2001).

Considerable research needs to be undertaken to determine how forest ecosystems respond to historical drought events and whether potential interactions, such as species composition and structure may change over time.

Chronic and acute drought events can increase tree mortality, through the reduction of available water, this changing the structural dynamics of forests ecosystems (Anderegg et al., 2014; Pangle et al., 2015). Changes in forest structure, such as the reduction in overstorey canopy cover, have been known to promote an abundance of juvenile individuals (Matusick et al., 2016). For example, following a drought-induce canopy collapse over 16, 515/ha in the NJF, due to drought and heat events in 2010/2011,

Matusick et al. (2013) reported on forest changes such as structural variance with a decrease in stand diameter, a greater number of juvenile classes and an increase of multi- stemmed individuals. Die-off events have been recorded in Eucalyptus globulus 7 | P a g e plantations, 100 km south of Perth, Western Australia, in 1994 where Harper et al. (2009) reported tree mortality was influenced by soil depth (<2m; 22% vs >2m; 70%), following regular drought events. The long-term consequences of drought events in forest ecosystems are poorly understood. However, mortality can affect site-level microhabitat, ground-level temperatures, solar radiation and wind speed (Ruthrof et al., 2016). Large fluctuations in rainfall and temperature are predicted due to climate change, that may (or likely to) result in structurally mismatched within the forest sites due to lower water availability, thus increasing dead ground floor biomass (Jump et al., 2017). Therefore, long-term monitoring of areas such as the drought-affected areas of the NJF will be beneficial to be able to predict changes in ecosystem processes following drought events.

Structural change within a forest is dependent on the traits of dominant tree species

(Frelich, 2016). Where structural changes do occur, these can influence changes in compositional dynamics (McDowell et al., 2008). Early changes in composition have been observed in the NJF, between two co-dominant canopy species (E. marginata and

Corymbia callophylla) and partly within midstorey species ( fraseriana,

Banksia grandis and longifolia) (Matusick et al., 2013; Ruthrof et al., 2016)

Sixteen months after the drought and heatwave events, the studies revealed that E. marginata was more susceptible to drought effects, while C. calophylla had a greater abundance, suggesting a higher chance of survival. Midstorey species showed drought had negatively affected densities. Changes in climate and the occurrence of drought and heatwave events are an important study area. How forest ecosystems respond to structural and compositional change in the long term is not fully understood

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2.2 FIRE REGIMES

Wildfires are natural disturbances and have had a strong influence on ecological and evolutionary adaptations in many terrestrial ecosystems (Pausas and Keeley, 2009).

Vegetation-fire regime relationship has evolved, creating a "habitat life-span and spatial scale" feedback loop, were plant flammability can be a major driver of flora evolution and distribution which is primarily driven by the ecosystems fire regime (He et al., 2011;

Pausas and Schwilk, 2012). The fire-regime has six essential interlinked parameters which include; fire size, type, frequency, seasonality, intensity and severity (Flannigan et al., 2000; Keeley, 2008). Definitions of these are summarised below:

• Fire size is commonly used to determine the distribution and landscape spread

(Flannigan et al., 2000; Ryan, 2002).

• Fire-type, is the vertical height to which flames reach (surface to overstorey canopy)

and is often driven by fire intensity and surrounding fuel characteristics, structure and

moisture availability (Flannigan et al., 2000).

• Fire frequency (fire return interval): the time between each fire event (Ryan, 2002).

Fire frequency can affect an ecosystem's response through influencing species

interaction, individual life cycles and vegetation dynamics in an ecosystem

(Flannigan et al., 2000; Bradstock, 2010).

• Fire Seasonality: fire seasons vary across ecosystems, and can influence the

successional trajectories of the ecosystem following the disturbance (Flannigan and

Wotton, 2001). Therefore, the time of year, seasonal weather patterns (e.g. wind,

temperature, humidity and rainfall), geographical location and vegetation type can

determine the fire intensity and severity levels (Figure 1) (Lin et al., 2013). For

example, fire seasons in southwestern Australia (occurs in late summer/autumn –

December to April (BoM, 2020a)) are regulated by dry fuel loads and climatic

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conditions, while in Northern Australia they are typically characterised by grass fires

as a result of biomass accumulation following the wet season (October and April

(BoM,2012)) (Figure 1) (BoM, 2020a).

• Fire intensity is the energy released during a fire within a single burn. Fire intensity

measurement requires on-ground surveys of topography, vegetation, fuel type and

loading, and the previous disturbance history (e.g. drought) (Flannigan et al., 2000;

Keeley, 2009).

• Fire severity is the quantified measurement of above-ground (crown volume scorch

or remaining twig diameter on branches) and below ground (surface burn depth and

soil organic layers) consumptions (Keeley, 2009).

Figure 1: (A) Map of Australia displaying major fuel types, Source: (Lin et al., 2013). (B) Fire seasons in Australia, Source: (BoM, 2020d)

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The study of climate and fire-regime interaction will assist in the long-term trajectory of plant species in fire prone ecosystems such as Mediterranean type ecosystem. Further, the interaction between both drivers may trigger irrecoverable biodiversity loss (Bradstock et al., 2002; Enright and Fontaine, 2014b).

Fire severity is impacted by weather, fuel and physical related variables (Broncano and

Retana, 2004; Ndalila et al., 2018). Through the influence of weather conditions, each mentioned variable can deliver a positive influence on increase of fire severity (Keeley,

2009). Heatwaves, low humidity, strong surfaces winds and unstable air patterns assist in creating extreme fire weather, while high humidity and cool conditions will deliver an opposite effect (Ndalila et al., 2018). Lower severity fires are related to lower average daily temperature and higher average daily relative humidity, and tend to occur in wetter forest regions (Ndalila et al., 2018). Patterns of fire severity can vary depending on several physical parameters such as fuel load, dominant vegetation type, and landscape characteristics (Pausas et al., 2003). For instance, the forest structure has several fuel layers (Figure 2), and over time, fuel loads accumulate, dry and thus, can increase fire severity and spread, depending on physical dryness of the fuel (Burrows and McCaw,

2013). Topography can determine the fire severity pattern through variations in windspeed and spread of the fire (Broncano and Retana, 2004), with upper- and mid- slopes tending to burn at a higher severity than valleys (Ndalila et al., 2018). Lecina-Diaz et al. (2014) investigated the contribution of weather, fuel, water availability, and topography to crown-fires across Mediterranean Pine forests in Catalonia, North East

Spain, finding that crown-fires were primarily located on steeper slopes (p<0.042) and occurred during dry weather (p<0.006). However, fuel biomass (p<0.593) and water availability (p<0.412) had no significant influence on fire spread. Continued investigation

11 | P a g e of crown-fires and the association between fuel loading and fire weather will assist in future fire outcomes in regions with steep topography regions.

Figure 2: Layers of fuel within a forest that can be identified visually. The grey scale on the left indicates the relative bulk density of each layer. The Vesta dry eucalypt forest fire model fuel layers are listed on the right Source: (Gould et al., 2008)

The interaction of fire severity and contributing variables such as topography influence, and wind pattern direction and microclimate changes during a fire can influence the post- fire succession and species regeneration (Pausas et al., 2003; Lentile et al., 2007). Knox and Clarke (2012) examined Eucalyptus species in three different forest types, rainforest

(RF), wet sclerophyll forest (WSF), and dry sclerophyll forest (DSF) in the New England

Tableland Bioregion of eastern Australia. Woody species displayed a faster recovery consistent with wetter forest types with 86% of RF type recovering to pre-disturbance level, WSF (80%) and DSF (72%). Similarly, Heath et al. (2016) examined the post-fire recovery of mixed eucalypts in the Greater Blue Mountains region south-eastern Sydney basin. Areas affected by lower severity fire recovered to pre-wildfire conditions in two years. In both studies, areas affected by high severity fires required significantly longer time to regenerate and occurred mainly in DSF environments. The implications of fire

12 | P a g e severity on post-fire succession and vegetation dynamics may impact fitness in seeders or resprouters differently. Therefore, a greater understanding of fire severity on a wide range of traits would be worthwhile.

Fire frequency and severity can lead to long-term changes to forest dynamics and alter fuels, vegetation distribution and vegetation persistence (Dale et al., 2001b; Steel et al.,

2015; Burton et al., 2019). Fire severity, is dependent on fuel accumulation rate

(decomposition and additions) and vegetation type (Steel et al., 2015). For example, in the temperate climate of Tasmania, Australia Ndalila et al. (2018) found altered fire severity patterns in DSF (high fire) and Pinus plantations (very high), while wet forest and eucalypt plantations remained largely unburnt (Table 1). Similarly, in the Australia

Capital Territory (ACT), Vivian et al. (2008) found fire severity to variably influence mortality amongst different eucalypts, with Eucalyptus delegatensis displaying a 99% mortality rate under high severity conditions, with low severity only resulting in 17%. In contrast, E. fastigata, E. dalrympleana and E. dives were able to persist in both fire severity conditions. Fuel accumulation also varied following fire with fire of low severity displaying significantly higher percentage than high severity (50.23% vs 21.80%). As disturbance regimes increase with the shifting climate, further research is needed on the interacting effects of fire frequency and severity, to understand the long-term and future trajectory of fire-prone forest ecosystems.

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Table 1: Percentage area of each severity classes in each vegetation type in an aerial photo classification, including the unburnt class. Severity class with the highest percentage area in each vegetation type is indicated in red. The total area of each vegetation type is indicated in hectares and as a percentage. Source: (Ndalila et al., 2018)

Fire Severity Vegetation type Dry Wet Eucalyptus Pine Non- Forest Forest Plantation Plantation forest Unburnt 15 57 40 2 2 High 32 9 23 27 27 Very High 26 7 10 54 54 Area (ha) 15,484 2,345 1,044 910 6,158 Vegetation (%) 60% 9% 4% 3% 24%

Species regeneration is driven by fire severity and the other parameters of the fire regimes

(Stephens et al., 2013). Increased fire severity can increase or limit regeneration some species in fire-prone environments are fire-dependent and require triggers such as heat to germinate or resprout (Stephens et al., 2013). In southwestern Australia, Etchells et al.

(2020) observed recruitment patterns of E. diversicolor, following different fire severities, revealing high severity fire having the most significant influence on density regeneration (germination, height and growth) (Table 2). In the Mediterranean climate of northwest Spain, Vega et al. (2008) found low severity conditions were more beneficial to seedling success, at 24months post-fire for Pinus pinaster (1.4/m vs 6.78/m respectively between control and post-fire sites). The low germination rates following low severity fire may be linked to older or non-viable serotinous seeds and short-fire return interval. The published evidence surrounding fire regimes are often explored individually. Therefore, it may be beneficial to examine how the parameters of the fire regime) (size, type, frequency, season, intensity and severity) operate in conjunction with one another, thus gaining an understanding of how a changing climate influences fire- regimes.

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Table 2: Regeneration dimensions (density, height, and growth) of in southwestern Australia, following high and low severity fires, including unburnt areas at 13 months post-fire. Red depicts highest regeneration variables. Source: (Etchells et al., 2020)

Fire Severity Seedling regeneration Density per m2 Height per m2 Growth per m2 (cm) Unburnt 0.00 0.00 0.00 Low 0.20 21.03 4.21 High 0.55 32.64 17.95

Near-annual large wildfires are becoming a regular occurrence in several global regions, such as Australia (McRae et al., 2015; Field et al., 2016), North America (Halofsky et al.,

2020) and Mediterranean Europe (Turco et al., 2019). Longer fire seasons have been linked to climatic changes such as higher temperatures, longer heatwaves and reduced precipitation rates (Bradstock, 2010; Flannigan et al., 2013; Jolly et al., 2015). For example, Harvey (2016) reported climate change to be a potential driver in increased fire activity, with increased temperatures, lower rainfall and humidity rates. Flannigan et al.

(2005) reported, climate change and increased CO2 has the potential to influence fire occurrence through the of vegetation growth, fuel abundance, rainfall, and ignition sources, likely leading to structure and species composition transformation. These changes in structural and composition are expected to modify fire regimes, and post-fire regeneration.

Fuel moisture content has been reported as a major factor influencing fire behaviour

(Matthews, 2014). Fuel moisture is dependent on several environmental conditions

(weather, topography, length of day) and vegetation characteristics (Miller and Urban,

2000). Finer fuels are more responsive to atmospheric moisture, whereas, coarser and denser fuels take longer to respond, as the fuel dries the loadings can generate extreme heat, assisting in the increase of fire severity (Miller and Urban, 2000). Temperature

15 | P a g e increases, reduced precipitation rates and extended drought and heatwave events, are likely to reduce moisture availability within the biomass (Moritz et al., 2012; Westerling,

2016). Dry fuel loads and vegetation increase fire severity, intensity and size (Flannigan et al., 2016; Nolan et al., 2016). Fire activity and spread are considerably weather dependent (Keeley, 2009). Therefore, climatic alterations are likely to shift fire regimes, with warming conditions predicted to promote fire activity in many regions (Moritz et al.,

2012). Rising global temperatures, extended drought and heatwaves in the fire-prone areas promoting fire weather through increasing the annual number of extreme fire weather days being recorded (Dale et al., 2001b). These seasonal changes are expected to increase in temperature, evapotranspiration rates, and to lower soil and fuel moisture substantially during summer in fire-prone ecosystems, creating longer fire seasons (IPCC,

2013)

Seasonal changes in weather conditions can influence the likelihood of increased fire activities through natural ignition sources. Lightning strikes are the leading cause of natural ignition all over the world and are part of the natural fire regime (Komarek, 1964).

Other sources of ignition include volcanic eruptions and gas emissions; however, these are restricted to a few specific locations (Komarek, 1964). Lightning strikes are a result of thunderstorm activity, through the creation of a negative charge in in the lower cloud region and a positive charge in the upper cloud region (Van Wagtendonk and Cayan,

2008). The separate polarity creates a discharge within the clouds, the air and the ground when the electrical discharge reaches the ground this ignites the fire (Van Wagtendonk and Cayan, 2008). Several studies have provided evidence that lightning strikes are likely to increase with climate change. For example, Veraverbeke et al. (2017) found lightning activity has increased between 1975 and 2014, recording greater burn area and emitting higher carbon emissions in the Boreal forest, North America. A more recent study has 16 | P a g e suggested a decrease in lightning activity, found a 15% decrease in total lightning flash rate under a strong global warming scenario by 2100 Finney et al. (2018)

2.3 PRESCRIBED BURNING IN WESTERN AUSTRALIA

Prescribed burning consists of conducting low intensity burns of dry and live understorey surface fuel under controlled conditions (Burrows and McCaw, 2013), in an attempt to reduce wildfire severity, intensity and spread (Burrows et al., 2010; Enright and Fontaine,

2014a). Prescribed burning can also be described as a social objective is to ensure communities within fire-prone regions remained protected, and fire crews are able to manage fires more effectively. This type of burning has also been endorsed by land management agencies as a vital practice for the sustainable management of fire-prone ecosystems (forests, woodlands and shrublands) (Burrows and McCaw, 2013). However,

Stephens et al. (2012) recommends the most effective way to manage and reduce wildfire severity is to use a combination of techniques, such as mechanical thinning and prescribed burning. Mechanical thinning has been seen to reduced negative impacts on plant communities in various forest environments (e.g. mixed conifer, Ponderosa pine, southeastern pine and Boreal forest) (Stephens et al., 2012). Once trees are effectively removed via mechanical thinning, prescribed burning techniques are used to target ladder fuels (i.e. biomass that allows the fire to climb to the canopy) within the intermediate zones of the forest structure. The combined use of mechanical thinning, prescribed burning and indigenous cultural burning (i.e. the use of traditional indigenous knowledge of fire, vegetation, time of year and conditions for small scale low intensity burns

(Russell-Smith et al., 2013) may have advantages in reducing fire severity needs more investigation

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Determining when to conduct prescribed burning requires understanding of the regional and local fire regimes, thorough knowledge of fire season and current climatic conditions

(Fernandes, 2015). Burning at the wrong time of year may increase the possibility of the fire burning out of control impacting the safety of the surrounding community sand changing the vegetation dynamics for affected ecosystems (Whittaker and Mercer, 2004).

Due to the natural fire season and the conditions which they create, prescribed burn practices in southwestern Australia are often conducted during the autumn (March – May) and spring months (September – November), while conditions are moist assisting in delivering low-intensity burns (Penman et al., 2011). The local knowledge of timing, frequency and vegetation and understanding the natural fire regime is an important factor for consideration. However, with a changing climate, the timing of prescribed burning and vegetation conditions may need further research to reduce the risk of prescribed burning becoming out of control wildfires.

Vegetation type and weather influence the severity and frequency of the local and regional fire regime. To evaluate the effect of prescribed burn on reducing size, severity and frequency of wildfires, land managers use a metric called “leverage”, which is the ratio between wildfire burn area and area burned by prescribed burning. That is how much land would have to be prescribed burnt in order to reduce wildfire hazard to an acceptable level

(Boer et al., 2009). This number is based on the vegetation type and climate conditions for each region (Boer et al., 2009). In addition, the frequency at which the land area needs to be burned is equally as important to consider ensuring effectiveness of the practice as vegetation types and climate will influence the rate of fuel accumulation (Boer et al.,

2009). Once both considerations are calculated the areas are burnt to assist in reducing the severity, spread and frequency of wildfire (Gill et al., 1987; Boer et al., 2009). For example, to reduce the wildfire hazard in Western Australia, an annual burn target 18 | P a g e

(leverage) of 200 000 ha has been calculated, which is made up of approximately 5-10% of the NJF area (Fernandes, 2015), or approximately a 180 000 ha of the annual burn target of the public forest estate (~1,127,600/ha), on a six year rotation (Havel, 1975a;

Bradshaw et al., 2018). In the Warren Region of southwestern Australia, Boer et al.

(2009) analysed prescribed burning and wildfire activity over 52-years (1953/54–

2004/05), and found prescribed burning significantly reduced wildfire size, frequency and distribution by 24-71% annually in a six-year burn time interval. However, the prescribed burn area has significantly reduced over time. For example, between 1962-1990 prescribed burning averaged at 12.5% of the regional area, resulting in a 0.3% annual wildfire burn area. However, during a more recent period (1991-2012), the prescribed burn area reduced to 6.6%, increasing the wildfire burn area average to 1.1%. Burrows and McCaw (2013). Thus, this displays how prescribed burning contributes to reducing wildfire burn area. Given the changing climate, fuel load accumulation and population growth into the outer region areas, research is needed into how “leverage” metrics and alternative prescribed burning and fuel reduction methods may assist in reducing wildfire activity and spread.

The primary objective of conservation through prescribed burning is to maintain a diverse representation of the ecosystem and habitat conditions (e.g. riparian zones, rainforests and aquatic ecosystems) (Whelan et al., 2002; Burrows and McCaw, 2013). Fire-sensitive ecosystems are distinguished by areas containing fire-killed species, with extended maturation periods (Burrows and McCaw, 2013). For example, many plant species require long fire intervals to develop and reproduce (Ooi et al., 2006). Regarding fire tolerant species. Burrows et al. (2010) reported the mean tree growth of E. marginata and C. calophylla of the NJF was faster and had a healthier canopies following prescribed burning. This contrasted with long unburned plots (25yrs) displaying slower 19 | P a g e growth rates, poorer canopy condition and higher mortality rates. In a study examining wildfires effect on eucalyptus spp. regeneration after prescribed burning, in the Great

Dividing Range, Victoria, Australia Bennett et al. (2016) found areas affected by low wildfire severity had significantly fewer live seedlings compared to high severity plots

(21/ha vs 12874/ha, respectively). In addition, stem mortality rates differed between wildfire severity conditions (Unburnt: 70%, Low: 80% High: 45%). However, Bennett et al. (2016) concludes suggesting prescribed burning had no apparent effect on forest resistance to wildfire, but decreased site-level resilience (via recruitment) by increasing mortalities of small stems. Several knowledge gaps remain surrounding tree mortality and regeneration and the influence of fire severity. Thus, research into the effects of fire severity on life-history traits of the forest species is an important area of research in conserving native flora.

2.4 EUCALYPT SPECIES AND ADAPTIVE TRAITS TO DISTURBANCE

Disturbances such as fire have had a strong influence on environmental and evolutionary adaptations in terrestrial ecosystems (Pausas and Keeley, 2009), but there are also a wide range of other disturbances such as defoliation, pathogens and drought (Burrows, 2013).

Eucalyptus species are extensive across the Australian terrestrial landscape, with >800 recorded species, with less than 10% being recognised as obligate-seeder species and the remainder having resprouting capabilities (Knox and Clarke, 2005; Nicolle, 2006;

Burrows, 2013). Obligate seeders are species often killed by crown-fire disturbance, and are dependent upon seed for regeneration. Due to absence of resprouting features, the species are able to rapidly grow and gain earlier maturity (Nicolle, 2006; Knox and

Clarke, 2012). Resprouters, by contrast, are slower growing, due to the disturbance adaptive traits the species possess, such as the thick outer bark layer and lignotuber, which

20 | P a g e protects the buds and meristems, allowing for epicormic and basal resprouting (Clarke et al., 2012; Burrows, 2013).

Serotiny is another adaptive trait and has been important for eucalyptus species, success following disturbance, where the closed fruits are retained in the crown canopy (Figure

3; (Lamont et al., 1991; Smith et al., 2014). Fire free intervals facilitate development of stored seed - short intervals may prevent the development of stored seed, longer return intervals may lead to less seed being viable in the canopy, influencing the persistence of fire-killed species (Pausas et al., 2004a; Franklin et al., 2005). A shifting climate may compromise seed development and viability, due to a decrease in the interval between fires (Enright et al., 2015). Further research into fire killed species and seed properties would be worthwhile to understand the factors influencing seed production in a larger range of species.

Figure 3: Factors affecting seedling establishment in obligate species and the positive and negative influences Source: (Smith et al., 2014)

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Stages of germination and seedling establishment, are often accompanied by high mortality rates in the plant life cycle, therefore for germination and seedling establishment success fire can increase the abundance of ‘safe sites’, which provides an area for the seed to germinate and seedling establishment (Bradstock, 1991). There are four ways ‘safe sites’ are created. (1) Through reduced shading and removal of vegetative competition, increasing the bare-ground cover can significantly favour establishment through reducing nutrient and light competition (Fernandes et al., 2018). (2) The removal of biotic influences (e.g. pathogens, herbivores and microorganisms), that may prevent establishment and inhibit seedling growth (Ashton and Willis, 1982). (3) The removal of some of the forest floor physical characteristics (e.g. woody debris), this altered soil characteristics improving water filtration and availability required for establishment

(Loneragan and Loneragan, 1964; Bond and Van Wilgen, 2012). (4) Fire can assist in the episodic release of nutrients into the topsoil layer, increasing regeneration success (Grove et al., 1986). Some of these factors are recognised as essential parameters of the ‘ash-bed’ hypothesis, which enhance the growth of plants on soil following the fire (Pryor, 1963;

Ashton and Willis, 1982). Following a laboratory experiment using soil from Minniberri

Reserve, Western Australia, Yates et al. (1996) reviewed the safe site conditions of E. salmonophloia and found that seedbed conditions, depth of seed burial, water availability and temperature/light influenced a successful establishment (Table 3). The study revealed germination survival of E. salmonophloia germinants was enhanced by all the ‘safe site’ factors, the most significant being nutrients, the ash-bed effect and water availability. The environmental parameters surrounding safe site variables exert a strong influence over post-fire recruitment due to the ash-bed effect.

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Table 3: Influences of site conditions such as soil conditions, water availability, seed depth and temperature on the Eucalyptus salmonophloia, Minniberri Reserve, southwestern Australia Source: (Yates et al., 1996)

Influence Minimum germination Maximum germination Soil condition – watered 10% under litter 78% in ash-bed Soil condition – unwatered 9% under litter 14% in ash-bed Seed depth 1% at 20 mm 84% on the surface Temperature/light 49% at 15°C 98% at 25°C

Fire return intervals can play a crucial role in the survival of many plant species. Obligate seeders are vulnerable to altered fire cycles (McCarthy et al., 1999; Pausas et al., 2004b).

This is because if the fire return intervals are shorter than their development to maturity, fire can limit the plant species ability, to produce viable seed (McCarthy et al., 1999;

Pausas et al., 2004b). On the other hand, if the interval is too long, viable seed can be lost to old age (Pausas et al., 2004b). In either scenario, the alteration of a fire regime to one different from the one to which the plant species are adapted to can cause localised extinction (Enright et al., 2015). In a post-fire study between fire return intervals (FRI), fire severities and obligate seeder recruitment, (Smith et al., 2014) found E. regnans had higher seedling abundance in unburnt and moderately burnt sites, following a long FRI, in Central Highlands region of Victoria, Australia. However, high severity sites displayed significantly higher abundance (4.49 times higher) than an intermediate FRI compared to a short FRI. Understanding the relative importance and the comparisons between ‘safe sites’ and FRI is critical to predicting how future climate will influence the distribution and abundance of Eucalyptus and other serotinous species and deliver a clearer insight into post-fire succession among species.

Species depend on abiotic influences such as temperature, and moisture to germinate

(Fernandes et al., 2018), and biotic factors such as seed-predators can determine how

23 | P a g e many seeds are available for germinate and the distribution of eucalyptus seeds (O'Dowd and Gill, 1984). In the canopy eucalyptus seeds are predated by birds as a primary food source (Valentine et al., 2014). Upon release to the ground, ants and rodents heavily harvest the seeds (O'Dowd and Gill, 1984; Nereu et al., 2019). Several Eucalyptus species store an extensive canopy seed bank and, following a crown-fire, the seed is released, temporarily saturating the ground floor, increasing the probability of recruitment

(O'Dowd and Gill, 1984). The ‘predator-satiation hypothesis’, Janzen

(1974) Hypothesised “synchronous reproduction in long-lived species produces more seed than can be consumed by seed predators in mast years, and starves the seed predators in non-seeding years”. O'Dowd and Gill (1984) reviewed this hypothesis and compared E. delegatensis regeneration between burnt and unburnt sites, in the

Brindabella Range, ACT. Burnt sites had more seed germination, and lower mortality rates at 72 weeks (post-fire). In addition, ant presence increased by 74%. Thus this hypothesis should be investigated, such as distinguishing the different effects each seed- predator would have on the Eucalyptus species and in-turn influence on germination.

Within the fire-prone ecosystem of the NJF, plant persistence strategies include resprouting bud capabilities, lignotubers (bulbous tissue mass located at the stem base of several resprouting species (Kerr, 1925)) and great bark thickness. Thicker bark and lignotubers offer protection and nutrient storage. These traits allow for epicormic resprouting to occur following disturbance (Bamber and Mullette, 1978; Pausas, 2015).

Epicormic resprouting can be driven by disturbances such as herbivory, drought and fire

(Burrows, 2008). For example, in the fire-prone ecosystem of southwestern Australia

Wardell‐Johnson (2000) examined the resprouting capabilities of different Eucalyptus species under different fire severities and revealed that all mature plants resprouted when experiencing 100% canopy scorch from the fire, regardless of fire

24 | P a g e severity. However, some immature eucalypts experienced stem-death and were able to resprout via the lignotuber.

Lignotuberous plants is frequently found in sclerophyll communities around the world.

They are characterised by being hard leaved, and having woody stems, with short internodes (Canadell and Zedler, 1995; Claßen-Bockhoff, 2016). They are most commonly in dry open forest and woodlands in Australia (Canadell and Zedler, 1995).

Lignotubers can be observed at the early stages of plant development (Kerr, 1925; Beadle,

1968; Abbott and Loneragan, 1984). The Eucalyptus lignotuber develops at the cotyledon level in the first or second growing season (Noble, 1984). The lignotuber forms from a pair of outgrowths gradually encompassing the stem Kerr (1925). Upon the final stage of development, bud clusters develop into groups which are incorporated into woody tissue (Kerr, 1925). For example, in E. marginata from southwestern Australia, the lignotuber develops between the first and second stage of growth, allowing for persistence following disturbance due to the carbohydrate reserves (Table 4) (Abbott and Loneragan,

1984). Lignotuber qualities in terms of adaptive traits following various frequencies of disturbance between species may differ substantially, and research into the area may assist in closing knowledge gaps surrounding individual species, that are adaptive to disturbance regimes.

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Table 4: Natural regeneration in virgin (uncut) Eucalyptus marginata (Jarrah) forest consists of six stages, Modified after Stoate and Helms (1938). Source: (Abbott and Loneragan, 1984)

Growth stage Description (1) Seedling Often less than one year old, cotyledons present, and without any obvious signs of a lignotuber.

(2) Lignotuberous The cotyledons are absent, and the lignotuber is obvious. The seedling growth stage between stage one and stage three can be slow as the lignotuber and root system develop.

(3) Seedling Repeated fires and other disturbances killing off the main stem coppice can promote the above-ground growth forcing the species to develop a multistems individual.

(4) Ground coppice The individual is can be viewed either as “incipient advanced growth” without a definite leader or “dynamic advanced growth” with a definite leader at >1.5m.

(5) Sapling >1.5m, with a diameter breast height over bark (D.O.B) <1.5cm

2.5 CONCLUSION

This review indicated the research on forest ecosystems in terms of drought and wildfire disturbance are broad. Increased drought occurrences and associated mortality events are changing forest dynamics, such as structure and composition. Similar observations have been acknowledged for wildfire. The increase in fire severity that is apparent within

Mediterranean –type ecosystems such as the southwest Australia, Mediterranean Europe, and California (USA) is concerning. How prolonged droughts and heatwave events influences the changes in fire regimes and the long-term effects of multiple disturbances on forest ecosystems clearly needs further work.

Wildfire and prescribed burning techniques such as fuel reduction can impact forest dynamics depending on the fire return interval. However, if the calculated burn area

(leverage) and vegetation types are not considered with the changing climate and evolving 26 | P a g e fire-regimes to properly manage fuel reduction, wildfire activity may increase. Recent examples of wildfires include bushfires such as Yarloop Fire (2016) in Western Australia, and in California (USA). These occurrences highlight the importance for further research in these environments. Southwestern Australia is particularly susceptible, experiencing a reduction in rainfall and an increase in mean temperature. A better understanding of forest recovery from disturbances such as prescribed burning is essential to understanding how ecosystems will shift with a changing climate.

Eucalyptus species persist through multiple disturbances, through adaptive traits, such as serotinous seeds, epicormic resprouting and lignotubers. Much of the available research on Eucalyptus species and disturbances investigate the effects of drought and wildfire within several ecosystems. Despite this, studies suggest knowledge gaps and a multitude of opportunities for new research in terms of the influences of current prescribed burning.

For example, the influences of prescribed burning on forests have been reported on public safety and conservation management, albeit mainly concentrating on leverage, seasonality, fuel loadings and fire return intervals required for the protection of public safety. Nevertheless, the influences of prescribed burning on drought-affected areas within the Mediterranean climate of southwestern Australia, are insufficiently studied.

Research on tree species regeneration is vital for a more in-depth understanding of forest structure development and composition.

Primarily my thesis will focus on three key research questions:

1. How is tree regeneration affected by drought and fire?

2. How do growth stages of tree regeneration vary by drought and fire severity?

3. How does tree regeneration respond to fire over time?

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3 METHODS

3.1 STUDY AREA: THE NORTHERN JARRAH FOREST 3.1.1 Location and vegetation distribution The southwestern region of Australia is one of five Mediterranean-type climate biodiversity hotspots worldwide (Myers et al., 2000). Within the floristic region one of the largest vegetation types is the jarrah (E. marginata) forest (Specht et al., 2012). The

NJF extends over three million ha ranging from 30.8-33.5° S and 115.8-117.8° E; (Figure

4) (Havel, 1975b). The vegetation structure ranges from tall open and dry sclerophyll forest in the north, to a tall closed forest in the south where annual rainfall is higher (Havel et al., 1989). Two trees species, Eucalyptus marginata and dominate the dry sclerophyll forests on the uplands of the NJF and range between 20 and

30 m in height (Yates et al., 2003). In the more damp western locations, E. marginata regularly exists as pure stands in the higher topographic regions (Hingston et al., 1980).

Historically, C. calophylla is the less dominant of the two overstorey species, with the downslope distribution of C. calophylla increasing by approximately one-third compared by the upper slopes (Hingston et al., 1980). The lower topographic areas contain other eucalypt species such as yarri (E. patens Benth.) and bullich (E. megacarpa F. Muell.), with wandoo (E. wandoo Blakely) and powder bark wandoo (E. accedens W.V. Fitzg.) alongside jarrah and marri. The midstory contains several smaller trees (between 4 and 7 meters tall), namely she-oak (Allocasuarina fraseriana Miq.), bull ( Willd.), and snotty gobble (Persoonia longifolia R. Br., P. elliptica R. Br.) (Havel et al., 1989).

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Figure 4: Map of the Northern Jarrah Forest, southwestern Australia (Shaded).

3.1.2 Geology and topography Developing on the Archaean crystalline rocks of the Yilgarn Block and forming the western edge of the Darling Range, the Darling Scarp is an expansive geological structure known as the Darling fault (Havel et al., 1989; Copp et al., 2011). The Darling scarp rises steeply to the east approximately 300m above sea level, with a north-south alignment.

The Darling scarp and plateau is dominated with shallow rolling valleys and gentle rises

(Havel et al., 1989), separating the plateau from the west of the scarp.

The sequence of geographical bands, soil and the topography of the NJF varies (Havel et al., 1989). The overlaying components of the geomorphic structure consist of different soil profiles, formed from the parent materials of the region. For instance, the overlaying and underground granite and metamorphic rocks such as gneiss, are regularly found to

29 | P a g e encroach up to 30 meters in thickness (Havel et al., 1989; Copp et al., 2011). Over time, the granite and metamorphic profile has weathered, forming the laterite profiles, which covers the Darling Scarp (Dell et al., 2012). Various processes such as erosion, deposition and leaching of the lateritic profile have assisted in the formation of five different nutrient-poor soil profiles (Table 5) (Churchward and Dimmock, 1989; Wardell‐Johnson,

2000; Dell et al., 2012).

Table 5: Five broad soil groups are identified within the Northern Jarrah Forest region, southwest Australia source: (Dell et al., 2012)

Soil Type Description Sandy gravels Comprised of pisolitic and nodular ferruginous gravels, mixed with an array of yellowish-brown sand

Siliceous sand consisting of two sub-groups with uniformly yellow to yellowish- brown sands, containing ironstone gravels. While the second sub- group is primarily grey soil with either yellowish-brown sand with dark brown loose nodules or dark brown to black sandy pan-like material, commonly known as “coffee rock”.

Earths The soil is derived from a scale of successive change in soil material identified by two separate colours (red or yellow).

Non-calcareous similar to the Earths previously described (above), however, the group differs in having a uniformly textured profile, while holding loams an orange coloured appearance.

Duplex soils The soil is characterized by a sharp break in texture between the sand to sandy loam.

3.1.3 Climate The NJF has a Mediterranean-type climate, experiencing warm dry summers and cool wet winters with the majority of the annual rainfall (80%) occurring between May and

August and a seasonal drought lasting between four to seven months (Bates et al., 2008).

The study region has an estimated average annual rainfall of 1222.6mm. The forest boundaries consist of a decreasing rainfall rate, with approximately 1300 mm on the scarp to 700 mm in the east and north. The maximum average temperature are 21.9 ̊ C and a

30 | P a g e minimum average temperature of 9.6 ̊ C (BoM, 2020c). Recently, a significant climatic shift in temperature and rainfall has been observed, with an increase in successive dry days, and a maximum of a 5-day precipitation reduction (IPCC, 2014) Locally, winter rainfall has reduced by 21% compared to the average precipitation rates since the 1960s

(Smith et al., 2000). Furthermore, simulation results by (Wandres et al., 2018) have shown an overall significant decline in rainfall east of the Darling Scarp. This is particularly in the winter months, the dominant rainfall season (Wandres et al., 2018)(ref). Beyond the long-term averages, regular occurrence of extreme weather events and changes to climatic variance is anticipated (i.e. heat waves, droughts and storms) (IPCC, 2013).

In the winter of 2010 the NJF experienced 40-50% below average rainfall, becoming the driest winter on record (BoM, 2011). Further, in the summer of 2010-11, multiple heatwave events occurred, and the region recorded the highest number of heatwave days since the 1960s (BoM, 2011; Ruthrof et al., 2018), The impacts of these events on populations dynamics of dominant plant species in the NJF is a key aim of this study.

The environmental response to the weather events of 2010-11, generated abrupt biotic disturbances of mortality, demographic shifts and species distribution traversing the terrestrial and marine biological systems (Ruthrof et al., 2018). Mortality increased almost 10-fold for shrubs and trees species (Ruthrof et al., 2018), with coastal marine systems also experienced a decline in health seen through coral bleaching. However, perhaps one of the most extreme and large scale consequence of the drought event, an estimated area of 16,000 ha, suffered canopy die-off in the NJF (Figure 5; Brouwers et al. (2013a); Matusick et al. (2013)). The co-dominant overstorey species (E. marginata and C. calophylla) experience either partial or complete crown dieback,

31 | P a g e resulting in resprouting amongst some individuals. The sites most affected were identified to occur within regions featuring xeric sites, granite outcrops, and low water holding capacity (Brouwers et al., 2013b; Andrew et al., 2016). Additionally the affected areas had significantly higher fine fuel loads than the non-drought affected areas, potentially increasing fire severity and spread. (Ruthrof et al., 2016)

Figure 5: Aerial photo of an area of the Northern Jarrah (Eucalyptus marginata) Forest, southwestern Australia (Site 74), at (a) three and (b) 16 months following drought. Scale bar represents approximately 200m, Source: (Ruthrof et al., 2015)

In Mediterranean climates such as southwestern Australia, fire is considered an endogenic disturbance, and is credited for maintaining species diversity and evolving adaptive traits within plants to promote persistence (Burrows, 2013). Dominant traits employed by species include, resprouting from underground and above ground storage organs (i.e. fire surviving) and obligate seeding (i.e. fire killed), with several dependent on fire for floristic and structural diversity (Wardell‐Johnson, 2000; Burrows and Wardell-Johnson,

2003). Fire regimes within the region are heterogeneous, with the average inter-fire intervals differing from frequent surface fires (2-3 times per decade) to 50-100 years dependent on vegetation type and bioregion (Burrows and Wardell-Johnson, 2003;

Enright et al., 2005).

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Forest surrounding the town of Yarloop, approximately 125km south of Perth is co- dominated by E. marginata and C. calophylla. On January the 5th 2016, a lightning strike ignited the Yarloop wildfire, due to unstable weather conditions of hot and dry offshore winds and a maximum temperature of 37 ̊ C being reached (McCaw et al., 2016).

Conditions prior to the wildfire were very dry; annual rainfall for the Dwellingup region in the preceding year was 780 mm, approximately 35% below the long-term average and the third lowest rainfall on record (McCaw et al., 2016). By the7th of January, the wildfire increased in magnitude, developing smoke plumes, with cumulus and pyrocumulus clouds developing lightning strikes and a new fire head, ahead of the main fire front

(13km east). The total area burned by the Yarloop fire was approximately 69,165 ha

(691.65 km2) of agricultural and native vegetation and destroying much of the town of

Yarloop. The study area for the thesis lays within the NJF, and the largely burned area in the first 2-3 days of the Yarloop wildfire (McCaw et al., 2016).

My study focussed on drought-affected areas in the NJF affected by the 2011 drought and

2016 fires. A particular focus is the co-dominant overstorey species E. marginata and

Corymbia calophylla, and a combined of midstorey species composed of Banksia grandis, Allocasuarina fraseriana, and two Persoonia spp. (P. longifolia and P. elliptica) species.

3.2 STUDY DESIGN 3.2.1 Site selection and plot establishment Drought affected forest site selection This study takes advantage of pre-existing plot networks affected by both drought/heat waves and fire established by (Matusick et al., 2013; Walden, 2020). Here, I describe the

33 | P a g e prior study designs and location of existing sites (Figure 6, 7), and then detail my own sampling approach

Figure 6: Timeline of previous research experiments (green; drought and wildfire) and climate and prescribed burn events (orange) leading to this study of tree regeneration in 2019-2020

Figure 7: Map of existing research sites. Wildfire denoted in red and Drought and Prescribed Burn sites denoted in yellow. Insert, map of Australia outlining the Northern Jarrah Forest location, southwestern Australia Matusick et al. (2013) aimed to understand the impacts of drought/heatwaves (2010/2011) on forest structure (regrowth and regeneration). Following the 2010-2011 drought/heatwave event, 235 drought-induced forest die-off patches were selected through observational aerial and field investigations (Figure 7), ranging in area from 0.3-

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16.8 ha (Matusick et al., 2013). Of these, 20 study sites were selected field study by

Matusick et al. (2013). They found that affected stems (>1cm Diameter Breast Height

(DBH)) in drought-induced plots displayed >70 % canopy collapse (mean 74 ±3 %) as opposed to less-affected areas plots <11% (±2 %) canopy collapse, the unaffected areas were considered as reference ‘control’ conditions.

Each plot had four consecutive visits at 3, 6, 12, and 26-month intervals after the drought

/heatwave event, measuring and documenting the initial impacts and changes the forest induced after drought (Matusick et al., 2013). Parameters measured included species, density, and mature tree resprouting in six healthy (control) and three drought-affected plots (Figure 8). Approximately six years after the event (2017), the sites were revisited for remeasurement of existing regeneration.

Between spring and autumn of 2016-2018 some of the drought-affected and control sites were treated by prescribed burning as part of a broader management program by the relevant environmental department (now DBCA) (Figure 6,7). Prescribed burning mainly occurred during the spring months of 2017, with autumn burns conducted at two sites.

From 2017-2018. In total 14 sites were used within the study, consisting of 96 plots. The information was later used for the purpose of the study into prescribed burn effects. The information of location and size of each prescribed burn was collected from DBCA

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Figure 8: (A) Plot schematic of previous (2016, unpublished) regeneration count (2m) and 2019/20 regeneration count (3m). (B) Plot locations in accordance to the drought affected sites in the Northern Jarrah Forest, southwestern Australia

Wildfire affected forest site selection Following the 2016 Yarloop Wildfire, field evaluations were undertaken to examine the impacts of repeated disturbances (drought and wildfire) on forest structure, mortality and recruitment in a resprouting forest type (Walden, 2020). Therefore, sites containing low and high drought probabilities before the Yarloop wildfire were identified. Geographical and climatic variables including elevation, slope, distance to rocky-outcrop, rainfall and temperature data were used to infer drought and heat sensitivity across a landscape

Brouwers et al. (2015). This model was used to determine whether the areas suffered drought stress before wildfire occurrence. Low drought probability sites were indicated through midstorey species occurrence using methods by Matusick et al. (2013). Indicator plant species such as B. grandis, are known to be susceptible to drought and heatwave events causing die-off (Matusick et al., 2013; Walden, 2020). High drought probability sites were predicted through structural assessment of the dominant overstorey species (E. marginata and C. calophylla), including dead branches and epicormic sprouts that would have developed following the drought event (Matusick et al., 2013; Walden, 2020).

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Measuring the size and development of epicormic sprouts after the wildfire, allowed for the identification of whether resprouting occurred pre- or post- 2010-2011 drought disturbance (Walden, 2020). The different fire severities recorded following the Yarloop wildfire were determined through stand assessments of overstorey and midstorey tree species (Walden, 2020). Once obtained, fire regions were grouped into three main categories: (i) severe surface fire with canopy scorch (moderate severity). (ii) crown fire with complete canopy consumption (high severity), and (iii) unburnt forests (control)

(Walden, 2020).

The experimental design by Walden (2020) followed a factorial model and containing 36 sites, with 144 plots (Figure 9). The field evaluations to determine drought and wildfire severity were undertaken at six-months, three and a half years and four-years post-fire intervals. For plot design establishment, a modified Forest Inventory and Analysis assemblage (Bechtold and Scott, 2005) was used (Figure 9). Each site consisted of four plots, with a centre plot and the remaining three plots, 35 m from plot centre at angles 0°

(North), 120° and 240°. The configuration was altered if forest condition altered rapidly or subplot was situated on track or road (Walden, 2020).

Figure 9: (A) Regeneration plot schematic for 2017 and 2019/20 study, (B) Plot locations in accordance to the drought and wildfire affected sites in the Northern Jarrah Forest, southwestern Australia

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3.2.2 Plot measurements Tree survey (drought affected sites) In summer 2017 each of the 99 drought-affected plots established by Matusick et al.

(2013) were monitored, recording the overstorey live vegetation within a six-meter radius and sampling individuals >1cm DBH of E. marginata and C. calophylla. Tree stem measurements were taken for post-drought tree structure. Five observations were recorded: (i) live and dead height (m), (ii) Diameter at breast height, (iii) basal sprout height (m), (iv) highest and lowest epicormic sprout point origin (m), and (v) canopy health and cover score for canopy loss during the drought. The crown condition, outlined by Worrall et al. (2008) and used by Matusick et al. (2013) was measured in all trees within each plot, in that each tree fell into 1 of 4 classes (Table 6). (Matusick et al., 2013;

Ruthrof et al., 2016). The health score rating was on a scale between 0-100, with 0 representing no canopy loss and 100 representing total crown loss or displaying no evidence of regrowth for each individual tree.

Table 6: Crown condition observations (after Ruthrof et al. (2016)) for the Northern Jarrah Forest, southwestern Australia.

Class Description 1 Healthy trees with foliage, primarily green

2 Dying the foliage is primarily dry and discolouring

3 Recently killed foliage is red and dead

4 Long dead absence of leaves, twigs and the outer bark layer was lifting

Tree survey (wildfire affected sites) Within the stand-level measurements by Walden (2020), each plot consisted of four subplots, with a variable plot radius, dependent on the tree stocking rate. Three radii were established and adjusted accordingly to ensure a minimum of 60 trees with a DBH of >10

38 | P a g e cm in each plot. Each radius was for one of the three different DBH classes; >30cm, 10-

30cm, and <10cm. To avoid overlapping with each subplot, a maximum of a 15m radius was established, remaining constant across each subplot area within the plot.

Tree stem measurements were taken for post-fire tree structure. Five observations were recorded in the process: (i) live and dead height (m), (ii) Diameter at breast height, (iii) basal sprout height (m), (iv) highest and lowest epicormic sprout point origin (m), and (v) canopy health and cover score for canopy loss during the fire. The health score rating was on a scale between 0-100, with 0 representing no canopy loss and 100 representing total crown loss or displaying no evidence of regrowth for each individual tree. A resprouting vigour rating was recorded if the tree had complete canopy resprouting. a score was determined between 0-100, with 100 representing complete canopy resprouting, one-year post fire (Walden, 2020).

Drought-affected and prescribed burn regeneration counts Using the existing radial metrics and microplot locations from Matusick et al. (2016), seven 1m2 subplots were set up at two-meter radial circumference from plot centre at random intervals. The radial circumference for this study was increased to three meters, allowing for prescribed burn and wildfire tree regeneration comparison. The regeneration was recorded in each plot and the seven 1m2 microplots, documenting tree species, and growth stage (Table 7, Figure 10) classified by Abbott and Loneragan (1984). From collecting tree regeneration densities, I was able to determine the effects of drought and prescribed burning on regeneration.

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Table 7: The natural regeneration in virgin (uncut) Jarrah (Eucalyptus marginata) Forest, consisting of six growth stages (Abbott and Loneragan, 1984)

Growth stage Description Germinant Often less than one year old and with cotyledons present, and without seedling any obvious signs of a lignotuber. Lignotuberous The cotyledons are absent, and the lignotuber is obvious. The growth seedling stage between stage one and stage three can be slow as the lignotuber and root system develop. Seedling Repeated fires and other disturbances killing off the main stem can coppice promote the above-ground growth forcing the species to develop a multistems individual. Ground The individual is can be viewed either as “incipient advanced growth” coppice without a definite leader or “dynamic advanced growth” with a definite leader at >1.5m. Sapling The individual is >1.5m, with a diameter breast height over bark (D.O.B) <1.5cm

Figure 10: The growth through growth stages of Eucalyptus marginata: stage one (germinant) grows into a stage two (lignotuberous seedling). Stage three (seedling coppice), is a pre-established resprout, grows into stage four (ground coppice). Stage five is formed when either stage two or stage four reaches >1.5m in height. Stage six is a developed small tree which has a diameter breast height of >2cm.

Wildfire regeneration counts At each subplot in the Yarloop plots, a three- or four-meter radius plot originating from plot centre was used for tree regeneration counts recording the tree species, and documenting the stage of growth (Table 7, Figure 10) classified by Abbott and Loneragan

(1984). From collecting tree regeneration densities, I was able to determine the effects of drought and wildfire on regeneration. 40 | P a g e

3.2.3 Statistical design This study aimed to determine how drought and fire severities influenced the tree regeneration by asking the following research questions (1) How is tree regeneration affected by drought and fire? (2) How do growth stages of tree regeneration vary in relation to drought and fire severity? (3) How does tree regeneration respond to fire over time? Post-fire tree regeneration was determined through analysing collected data from three separate datasets. Two from past data collection (Matusick et al. (2016), Walden

(2020)) and my current study data from 2019-2020. The data sets were collated and organised in Microsoft Access (2016) and Microsoft Excel (2016). Once obtained and organised, the mean estimate of measured variables was calculated delivering the output and a 95% confidence interval within each plot. Tree regeneration densities were calculated (equation 1). All statistical analysis models and visual graphs were designed and created using RStudio (Version 1.2.5033) by RStudio Inc. 2009-2019.

Equation 1: Equation used to calculate the regeneration density within each plot 푹풆품풆풏 풄풐풖풏풕풔 푹풆품풆풏풆풓풂풕풊풐풏 풅풆풏풔풊풕풚 = 풙 ퟏퟎ, ퟎퟎퟎ 퐓퐨퐭퐚퐥 퐩퐥퐨퐭 퐚퐫퐞퐚 (퐦)

Poisson modelling was used for the statistical analysis of total tree regeneration and overstorey regeneration. However, due to the limited densities counts recorded during the survey period, the midstorey species were left out and only graphical analysis was used.

Tree regeneration analysis between drought and fire severity To explore the influence of drought and fire type on tree regeneration density and species,

I used the observed densities from my survey period. The information was examined in one of three variations, total tree species abundance, overall species (E. marginata and C. calophylla), and midstorey species abundance (A. fraseriana, B. grandis, and Persoonia 41 | P a g e spp.). A Poisson regression model was developed to study the regeneration of tree species to the significance and influencing factors to tree regeneration. The first model (Table 8; equation 2) was to investigate the effects of drought intensity and fire severity on the total tree regeneration that was observed in my survey period. The second model (Table 8; equation 3) was used to understand the effects drought and fire severity has on the regeneration of overstorey species forest regeneration.

Growth stage influence analysis between drought and fire severity To examine the influence of drought and fire severity on stages one (germinant) and two

(lignotuber seedling) density abundances within my survey period, percentages were then quantified for each overstorey (E. marginata and C. calophylla) and midstorey (A. fraseriana, B. grandis, and P. longifolia) species. These percentages were then graphed for visual examination.

The analysis of the influence of drought and fire on stages three (seedling coppice) and four (ground coppice), used the densities recorded from my survey period. Firstly, two graphs were created, separating the overstorey and midstorey species, and categorised the drought condition and fire severity of the regeneration densities between the two stages.

This was followed by statistical analysis through a Poisson model (Table 8; equation 4) of the two investigated growth stages of the overstorey species regeneration recorded during my survey period.

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Table 8: Equation used for statistical analysis. cnts: counts, spp: species, dc: drought condition, fs: fire severity, gs: grow stage, TSF: time since fire

Equation Response Explanatory Model equation No. 2 Total tree counts Drought condition 풄풏풕풔 ~ 풅풄 푿 풇풔 of regeneration and fire severity 3 Species counts of Drought condition 풔풑풑 풄풏풕풔 ~ 풅풄 푿 풇풔 tree regeneration and fire severity

4 Resprout counts of Drought condition 풔풑풑 풄풏풕풔 ~ 풅풄 푿 풇풔 + 품풔 species and growth and fire severity stages 5 Total tree counts Drought condition, 풄풏풕풔 ~ 풅풄 푿 푻푺푭 푿 풇풔 of regeneration fire severity and time since fire 6 Seedling counts of Drought condition, 풔풑풑 풄풏풕풔 ~ 풅풄 푿 푻푺푭 푿 풇풔 species and growth fire severity and time stages since fire

Tree regeneration analysis for time since fire Changes in tree regeneration over time for prescribed burn plots were analysed by using data previously collected in 2017, and my survey period in 2019, and analysed from time since the prescribed burn (TSF). The total tree regeneration was first used, followed by investigation of the germinant and lignotuber growth stages (one and two), to visually identify how drought condition may have influenced growth and species regeneration within prescribed burned plots. The data were grouped by year since prescribed burn

(1.5years, 2.5years and 3.5years). A Poisson regression model (equation 5) was used to identify how the total tree regeneration changed over time, and another model (Table 8; equation 6) was used to determine the significance of the interacting variables on overstorey species growth stages one and two.

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How tree regeneration changes over time within moderate and high severity wildfire plots was analysed using data collected from Walden (2020) and current collected data for my current study between autumn of 2019 and summer of 2020. The data were grouped according to the years they were collected including pre-fire (control), 2016/2017 (6- months to 1-year post-fire) and 2019/2020 (3.5-years to 4-years post-fire). Total tree regeneration abundance was analysed using graph and Poisson modelling (Table 8; equation 5). Then to analyse the overstorey species growth stages were separated, with graphs and Poisson modelling (Table 8; equation 6) was used to determine the changes in density between growth stages one (germinant) and two (lignotuber seedling).

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RESULTS

4.1 HOW IS TREE REGENERATION AFFECTED BY DROUGHT AND FIRE? Tree regeneration of all study species following drought and fire Approximately 3.5 years, following fire, a significant difference (as indicated by a lack of overlap in confidence interval) was observed in tree regeneration between the control plots (healthy 1925/ha vs drought affected 3903/ha) and drought plots (Figure 11). These species were Eucalyptus marginata, Corymbia calophylla, Allocasuarina fraseriana,

Banksia grandis and Persoonia longifolia. Overall, there were significantly higher regeneration counts in drought-affected plots within wildfire plots when compared to healthy (non-drought affected) conditions within the wildfire plots. Tree regeneration in healthy plots with moderate fire severity showed significantly higher (p<0.004, with a multiplicative rise of +1.25; Table 9) regeneration counts compared to control (1925/ha) plots with a density of 7840/ha (Figure 11). Regeneration in drought-affected plots across all fire severities were significantly higher compared to healthy fire affected plots, with moderate fire severity recording the highest observed regeneration counts (healthy: 7840 vs drought-affected: 11780/ha; p<0.004, +0.87). Plots burnt at high fire severity were also significantly different from healthy fire affected plots (healthy: 2101/ha vs drought- affected: 5793/ha; p<0.001, +1.44). Finally, drought-affected plots with a prescribed burn were significantly different compared to healthy fire affected plots of the same severity

(healthy: 3198/ha vs drought-affected: 3562/ha; p<0.031, +0.77).

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Figure 11: The mean total tree regeneration (count/ha) of tree species (Eucalyptus marginata, Corymbia calophylla, Allocasuarina. fraseriana, Banksia grandis and Persoonia longifolia) across different drought conditions (heathy and drought affected), among varied fire severities (prescribed burn, moderate and high severity) and control plots in the Northern Jarrah Forest, southwestern Australia. Values include 95% confidence intervals.

Table 9: Results from a Poisson regression model used to quantify the total observed tree species (Eucalyptus marginata, Corymbia calophylla, Allocasuarina. fraseriana, Banksia grandis and Persoonia longifolia) regeneration density amongst varied fire severities (prescribed burn, moderate and high wildfire severities) in the Northern Jarrah Forest, southwestern Australia. Models tested probability differences in tree regeneration across differing drought conditions and fire severities, using covariate from the site to regeneration level. Values include 95% confidence intervals.

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Overstorey tree regeneration following drought and fire An interactive effect was observed between drought condition and fire severity. Densities differed between fire severities with moderate severity plots being significantly higher regardless of drought condition, and of the recorded highest regeneration densities

(healthy: 5699/ha and drought-affected: 7619/ha; Figure 12). However, the influence of high fire severity (p<0.001, +1.54), produced higher abundance of regeneration in drought-affected plots than plots with a healthy conditions (Figure 12; Table 10).

Figure 12: The mean total tree regeneration comparison for overstorey tree species (Eucalyptus marginata and Corymbia calophylla) density among varied drought conditions (healthy and drought-affected)and fire severities (prescribed burn, moderate and high severity) and control plots in the Northern Jarrah Forest, southwest Australia. Values include 95% confidence intervals.

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Table 10: Results from a Poisson regression model used to quantify the overstorey species (Eucalyptus marginata and Corymbia calophylla) tree regeneration density among fire severities (prescribed burn, moderate and high severity) in the Northern Jarrah Forest, southwest Australia. Models test probability differences of overstorey tree regeneration among varied drought conditions and fire severities, using covariates from the site to regeneration level. Values include 95% confidence intervals

Corymbia calophylla did not display any significant change in healthy conditions within any fire severity influence. However, in drought-affected plots, fire displayed an interactive effect producing significantly higher densities in plots where wildfire severities, were moderate (4061/ha; p<0.001, +1.75; Figure 12; Table 10) and high

(2217/ha; p<0.001, +1.85).

Midstorey tree regeneration following drought and fire Healthy conditions had slightly higher regeneration counts compared to drought-affected plots. Prescribed burn plots contained the highest midstorey species densities (for A. fraseriana) for each drought condition (healthy: 353/ha, drought-affected: 337/ha; Figure

13). However, for B. grandis, recorded regeneration only in healthy plots

(251/ha). Banksia grandis displayed the highest density in moderate fire severity plots, with a considerable difference plots drought plots (healthy: 369/ha; drought-affected:

87/ha), and in high severity plots (133/ha) in healthy conditions. However, the records were highly variable as indicated by the wide confidence Intervals (CI)

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Figure 13: The mean total tree regeneration comparison of midstorey tree species (Allocasuarina fraseriana, Banksia grandis and Persoonia longifolia) density among varied drought conditions (healthy and drought-affected), and fire severities (prescribed burn, moderate and high severity) and control plots, in the Northern Jarrah Forest, southwestern Australia, including the 95% confidence interval.

Persoonia longifolia was the only species recorded under each drought condition and fire severity (Figure 13). The highest densities were within healthy conditions with a moderate fire severity (183/ha). Control plots displayed similar densities regardless of drought condition (healthy: 72/ha; drought-affected: 62/ha), likewise for high severity (healthy:

103/ha; drought-affected: 74/ha). No P. elliptica was located, and thus removed from the remaining analysis.

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4.2 HOW DO GROWTH STAGES OF TREE REGENERATION VARY BY DROUGHT AND FIRE SEVERITY? How drought and fire influence frequency distribution 4.2.1.1 Overstorey tree regeneration of germinants and lignotuberous seedlings following drought and fire – stages one and two Eucalyptus marginata, germinants occurred at similar frequencies between the control

(long-time since burnt) and prescribed burn plots, with little differences among drought conditions ranging from 17% to 27% of the total recorded density distribution (Figure 14) and were not found in the moderate and high severity plots. The lignotuber seedlings, had similar frequencies throughout the plots regardless of drought condition and fire severity, ranging from 17% to 26% of the full density.

Figure 14: The mean percentage of growth stages of the overstorey tree species (Eucalyptus marginata and Corymbia calophylla) between the drought condition (healthy and drought-affected) and fire severity (prescribed burn, moderate wildfire and high wildfire) and control plots in the Northern Jarrah Forest, southwestern Australia.

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Corymbia calophylla germinants were also only visible in the control and prescribed burn severities in both drought conditions, displaying a wide frequency range of 11% to 30% of the total observed density (Figure 14), depending on drought condition and fire severity. For example, prescribed burn plots showed high frequencies than control plots, and drought-affected conditions illustrating slightly higher frequencies than healthy conditions. Unlike the germinants, lignotuberous seedlings were recorded throughout each plot, with similar frequencies under both drought and fire severity condition, ranging from 15% to 26% of the full density.

4.2.1.2 Midstorey tree regeneration of germinants and lignotuberous seedlings following drought and fire – stages one and two Midstorey species reflected a variation between all species. Germinants of A. fraseriana were only present in prescribed burn plots in both drought conditions (Healthy: 31%, and drought-affected 46%; Figure 15). Lignotuberous seedlings were present in prescribed burned plots (both drought conditions), and healthy conditions with a moderate fire severity, ranging between 25%-40% of the total regeneration distribution. Banksia grandis, only recorded germinants in control (31%) and prescribed burn severity (25%) plots with healthy condition (Figure 15). The lignotuberous seedlings for B. grandis were absent in prescribed burn and high fire severities within drought-affected plots. However, where present, frequencies ranged from 25% to 50%, with high severity (50%) and moderate severity (42%) displaying the highest frequencies. Persoonia longifolia germinants where only present in healthy condition plots with prescribed burning (30%;

Figure 15). The P. longifolia, lignotuberous seedlings, were present in all fire severity plots with a healthy condition, with high fire severity showing the highest frequency

(42%).

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Figure 15: The mean percentage/ha of growth stages of the midstorey tree species (Allocasuarina fraseriana, Banksia grandis and Persoonia longifolia) between drought conditions (healthy and drought-affected) and fire severity (prescribed burn, moderate wildfire and high wildfire) and control plots in the Northern Jarrah Forest, southwestern Australia.

How drought and fire influence resprouting stages three and four 4.2.2.1 Overstorey tree regeneration of seedling coppice and ground coppice following drought and fire – stages three and four. Eucalyptus marginata displayed significantly higher densities (stages combined) in healthy conditions with a moderate severity wildfire (p<0.007; +0.95; Figure 16; Table;

11), sharing a similar significance as drought-affected areas by comparison. Drought- affected conditions also revealed significantly higher densities for prescribed burn plots

(1656/ha; p<0.044; +1.24) and high severity (1463/ha; p<0.010; +1.66), when compared to healthy conditions (seedling coppice and ground coppice combined). The ground coppice growth, had significantly lower regeneration counts, compared to seedling coppice growth (p<0.001, -1.11) or approximately 6010/ha vs 2322/ha, a difference of

61%. Notably, high fire severity showed similar densities by comparison between growth stages, regardless of drought condition (16%) 590/ha vs 490/ha (drought-conditions combined).

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Figure 16: The mean total tree regeneration comparison for overstorey tree species (Eucalyptus marginata and Corymbia calophylla) density among varied growth stages seedling and coppice ground coppice), drought conditions (healthy and drought-affected), and fire severities (prescribed burn, moderate and high severity wildfire) and control plots in the Northern Jarrah Forest, southwestern Australia, including the 95% confidence interval.

Table 11: Results from a Poisson regression model, used to quantify total observed overstorey tree species (Eucalyptus marginata and Corymbia calophylla) regeneration density across growth stages (seedling coppice and ground coppice) and fire severities (prescribed burn, moderate and high severity wildfire) in the Northern Jarrah Forest, southwestern Australia. Models tests probability differences of tree regeneration across varied fire severities and drought conditions, using covariate from the site to regeneration level, including the 95% confidence intervals.

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Corymbia calophylla, showed significantly higher abundance in healthy conditions for moderate fire severity (811/ha; p<0.037, +0.95 Figure 16, Table 11). However, under drought conditions it showed similar density abundance and no significant differences in regeneration were apparent when compared against healthy conditions. Similar to E. marginata, ground coppice also displayed significantly fewer regeneration counts when compared to seedling coppice densities (p<0.001, -0.57) with an estimated average of

2679/ha vs 1675/ha, a 37% difference. Similarly, E. marginata, densities in high fire severity plots were similar between growth stages in healthy conditions (16%; (194/ha vs

223/ha), and widely different in drought-affected conditions (46%; 425/ha vs 223/ha).

In comparisons between C. calophylla and E. marginata, C. calophylla recorded the higher densities only in drought-affected control plots with a combined stage density of

157/ha vs 72/ha (Figure 16), a difference of 54%. In all other plots, E. marginata appeared to be the more dominant under all fire severity conditions. The greatest difference was within healthy plots under moderate severity conditions, a density comparison of 1907/ha vs 812/ha, a difference of 57%. The smallest difference was also within healthy plots in high fire severity conditions 697/ha vs 427/ha a difference of 38%.

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4.2.2.2 Midstorey tree regeneration of seedling coppice and ground coppice following drought and fire – stages three and four Allocasuarina fraseriana was recorded in prescribed burn plots for both stages in each drought condition, with seedling coppice displaying the highest densities (healthy 68/ha; drought-affected; 60/ha; Figure 17). A. fraseriana also was recorded in healthy plots following moderate and high fire severities, but drought-affected plots failed to contain any regeneration. Banksia grandis was recorded in each healthy plot and growth stage, with prescribed burn severity showing the highest density count (seedling coppice 131/ha; ground coppice 85/ha). However, B. grandis had no counts in drought affected conditions, except for seedling coppice in moderate severity plots (60/ha). Persoonia longifolia had the highest density counts for high severity plots for each growth stage in both drought conditions, with each growth stage showing similar averages regardless of drought condition (seedling coppice 37/ha; ground coppice 40/ha). However, the species recorded the lowest counts over all other fire severities.

Figure 17: The mean total tree regeneration comparison of for midstorey tree species (Allocasuarina fraseriana, Banksia grandis, and Persoonia longifolia) density among grow stages (seedling coppice and ground coppice), different drought conditions (healthy and drought-affected) and fire severities (prescribed burn, moderate and high severity wildfire) and control plots in the Northern Jarrah Forest, southwestern Australia, including the 95% confidence interval.

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4.3 HOW DOES TREE REGENERATION RESPOND TO FIRE OVER TIME? Prescribed burning regeneration change at time since fire 4.3.1.1 Changes in regeneration under differing time since fire when undergoing prescribed burns Higher tree regeneration densities were recorded in drought-affected conditions following prescribed burning at each time since fire period (p<0.015, +0.39; Table 12). Overall, healthy plots showed significantly less regeneration between 2.5-years (p<0.001, -1.39) and 3.5-years (p<0.001, -1.88) post-fire interval measurements when compared to the control (unburnt) plots. Drought-affected conditions resulted in similar measurements by comparison, however 1.5-years post-fire, there was significantly less regeneration compared to healthy conditions (p<0.001, -0.65; Figure 18; Table 12).

Figure 18: The mean total tree regeneration comparison of all tree species (Eucalyptus marginata, Corymbia calophylla, Allocasuarina fraseriana, Banksia. grandis, and Persoonia longifolia) density among drought conditions (healthy and drought-affected), prescribed burn plots and unburnt, over three-and-a-half-year period from time since fire in the Northern Jarrah Forest, southwestern Australia, including the 95% confidence interval.

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Table 12: Results from a Poisson regression model, used to quantify the total tree species (Eucalyptus marginata, Corymbia calophylla, Allocasuarina fraseriana, Banksia. grandis, and Persoonia longifolia) regeneration density among drought conditions (healthy and drought affected), prescribed burn and unburnt and time since fire in the Northern Jarrah Forest, southwestern Australia. Model tests probability differences of tree regeneration differing in time since fire and drought conditions using covariates from the site to regeneration level, including the 95% confidence intervals.

4.3.1.2 Overstorey regeneration change of germinant and lignotuber seedling at time since prescribed burn Following 1.5-years since fire, E. marginata did not produce any germinants in either drought condition, but germinants were present in non-drought affected plots at 2.5-years

(Figure 19). No E. marginata germinant regeneration was observed 3.5-years post-fire

(p<0.001, -2.89; Table 13). Eucalyptus marginata lignotuberous seedlings were significantly higher than germinants overall (p<0.001, +1.64), with lignotubers recorded in almost each year post-fire in both drought conditions, with no significant loss.

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Figure 19: The mean total tree regeneration comparison of overstorey species (Eucalyptus marginata and Corymbia calophylla) density among drought conditions (healthy and drought affected), prescribed burn and control plots for growth stages (germinant and lignotuberous seedling), over a three and a half year period from time since fire in the Northern Jarrah Forest, southwestern Australia, including the 95% confidence interval.

Table 13: Results from a Poisson regression model, used to quantify the overstorey tree species (Eucalyptus marginata and Corymbia calophylla) regeneration density across drought conditions (healthy and drought affected), prescribed burn and control plots, for growth stages (germinant and lignotuberous seedling), over a three and a half year period from time since fire in the Northern Jarrah Forest. Model tests probability differences of tree regeneration differing in time since fire and drought conditions using covariates from the site to regeneration level, including the 95% confidence intervals.

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Corymbia calophylla, revealed a similar trend, displaying no germinants 1.5yrs post-fire, however at 2.5yrs germinants were recorded in healthy plots (Figure 19). However, 3.5yrs post-fire, a significantly lower was witnessed (p<0.045, -2.19; Table 13). Lignotuber seedlings were significantly higher than germinants overall (p<0.001, +1.80), but much lower than control plots. Further lignotuber seedlings were observed in almost each year post-fire.

Wildfire severity regeneration change and time since fire 4.3.2.1 Total species regeneration and time since wildfire The regeneration densities between healthy and drought-affected plots were significantly different with the droughted plots displaying a higher density (p<0.009, +1.12), a 46% difference in regeneration density (Figure 20, Table 14). Under healthy conditions, no significant increase in regeneration was detected during time since fire (severities combined), demonstrated by the overlap of the CI from the high severity. However moderate severity showed a significantly (p<0.010, +1.34) higher regeneration by the lack of overlap between years compared to the control plot (control 1481/ha vs 3-years

7629/ha) showing a difference of 81%. While high severity illustrated no significant increase under healthy conditions. Similarly, drought-affected plots also showed no signs of significance (severities combined) over time since fire, signified by the overlap from the high severity (Figure 20). However, moderate severity revealed a considerable difference in regeneration densities compared to healthy conditions of the same severity

(healthy 7,628/ha vs drought-affected 11,779/ha) at 3 years since fire (63% difference).

Likewise, high severity plots showed a significant difference also (P<0.004, +0.37; healthy 2,101/ha vs drought-affected 5,795/ha) at 3 years since fire (64% difference). In other words, drought-affected plots revealing higher densities counts by comparison.

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Figure 20: The mean total tree regeneration/ha comparison of all tree species (Eucalyptus marginata, Corymbia calophylla, Allocasuarina fraseriana, Banksia. grandis, and Persoonia longifolia) density across drought conditions (healthy and drought affected), wildfire severity (moderate and high) and control plots over the 2016-2020 period from time since fire in the Northern Jarrah Forest, southwestern Australia, including the 95% confidence interval.

Table 14: Results from a Poisson regression model, used to quantify all tree species (Eucalyptus marginata, Corymbia calophylla, Allocasuarina fraseriana, Banksia. grandis, and Persoonia longifolia) regeneration density across drought conditions (healthy and drought affected), wildfire severity (moderate and high) and control plots, over the 2016-2020 period from time since fire occurred in the Northern Jarrah Forest, southwestern Australia. Model tests probability differences of tree regeneration differing in time since fire and drought conditions using covariates from the site to regeneration level, including the 95% confidence intervals.

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4.3.2.2 Overstorey regeneration change between growth of germinant and lignotuberous seedling at time since wildfire Eucalyptus marginata regeneration in wildfire severity plots showed no significant difference between drought conditions (p<0.427). However, 1-year post-fire a significantly higher germinant density count was displayed (p<0.004, +2.40; Figure 21-

22; Table 15). Notably, 3-years post-fire no germinants was recorded in either drought condition or fire severity, also showing a significantly less (p<0.003, -2.59). The lignotuber seedling density was significantly higher 3-years post-fire (p<0.001, +3.08), with lignotuber seedlings overall being significantly higher, compared to germinantes

(p<0.001, +1.47; Figure 21-22; Table 15).

Figure 21: The mean total tree regeneration/ha comparison of overstorey species (Eucalyptus marginata and Corymbia calophylla) density across drought conditions (healthy and drought affected), moderate severity and control plots for growth stages (germinant and lignotuber seedling), over a three year period from time since fire in the Northern Jarrah Forest, southwestern Australia, including the 95% confidence interval.

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Figure 22: The mean total tree regeneration/ha comparison of overstorey species (Eucalyptus marginata and Corymbia calophylla) density across drought conditions (healthy and drought affected), high severity and control plots for growth stages (germinant and lignotuber seedling), over a three year period from time since fire in the Northern Jarrah Forest, southwestern Australia, including the 95% confidence interval.

Table 15: Results from a Poisson regression model, used to quantify the overstorey tree species (Eucalyptus marginata and Corymbia calophylla) regeneration density across drought conditions (healthy and drought affected), wildfire severity (moderate and high) and control plots for growth stages (germinant and lignotuberous seedling), over a three year period from time since fire in the Northern Jarrah Forest, southwestern Australia. Model tests probability differences of tree regeneration differing in time since fire and drought conditions using covariates from the site to regeneration level, including the 95% confidence intervals.

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Corymbia calophylla also revealed no significant difference between drought conditions.

However significantly higher germinant counts were recorded 1-year post-fire (p<0.001,

+3.47), with no significant difference at 3-years (Figure 21-22, Table 15). Lignotuberous seedling density was significantly higher 3-years post-fire (p<0.001, +3.60), and lignotuberous seedling density was significantly higher than that of germinants (p<0.001,

+1.78). Between fire severities, healthy conditions with high fire severity, showed significantly less regeneration (p<0.001, -2.95), while drought-affected conditions with high fire severity revealed significantly higher densities (p<0.035, +1.49; Figure 21-22;

Table 15) compared to control plots.

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DISCUSSION This study has shown that drought and fire severity can likely drive tree regeneration in the Northern Jarrah Forest, southwestern Australia. The results highlight that drought- affected condition was a major driver in overstorey tree regeneration abundance. Fire disturbance was also an effective driver in tree regeneration, with a moderate fire severity being the most influential to overstorey tree regeneration. These two drivers produced an interaction effect, where the highest tree regeneration densities were observed in plots consisting of the drought-fire relationship. This interaction effect was seen in each examined growth stage, and time since fire for sites burnt under a moderate fire severity, thus highlighting how drought and fire can contribute to greater regeneration densities.

Drought and fire disturbances have also been found to influence regeneration in other Eucalypt forests and growth stages throughout Australia (Wardell‐Johnson, 2000;

Lunt et al., 2011; Bennett et al., 2016; Matusick et al., 2016). However, other studies in the Northern Jarrah Forest, have been inconsistent with growth stage regeneration (Stoate and Helms, 1938; Abbott and Loneragan, 1984), Higher abundance of regeneration following one or both disturbances in all Eucalyptus species were seen in all consistent studies. However, highest regeneration densities were most common in plots in wildfire severities affected areas, rather than in areas prescribed burns. The inconsistency in growth stage regeneration can be attributed to other factors such as environmental cues

(e.g. light and water stress), mature tree density, and seed availability (Bell, 1994; Clinton et al., 1994; Torres et al., 2006). This study highlights the response of tree regeneration following a drought-fire disturbance. Regeneration growth stages have not been examined previously following multiple disturbance events in the NJF or other

Mediterranean forest ecosystems, rather studies have examined them as standalone disturbances (Vivian et al., 2008) or based on silviculture and logging stocking densities

(Stoate and Helms, 1938; Abbott and Loneragan, 1984),

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5.1 HOW IS TREE REGENERATION AFFECTED BY DROUGHT AND FIRE? Drought and fire influence on forest level tree regeneration Following the surveying of drought-affected plots in the NJF, healthy (non-drought affected) plots showed almost half as much regeneration as in drought-affected plots.

Drought has the potential to significantly alter the tree canopy (Fernandes et al., 2018), which can influence biotic (e.g. micro-flora and micro-fauna, herbivore) and abiotic (e.g. light availability, soil and air temperature, soil nutrients and water content) factors. Thus, thereby influencing regeneration and growth rate. Such changes may also affect the understorey composition and structure (Dietze and Clark, 2008; Muscolo et al., 2014), as drought can increase fine woody (FWD) and coarse woody debris (CWD) loadings, affecting germination and establishment (Ruthrof et al., 2016; Fernandes et al., 2018).

Previous studies have found CWD to facilitate regeneration in some species through increased resource availability such as water and nutrients (Marañón-Jiménez et al.,

2013). Similarly, drought has known to increase the number of small stems via resprouting (Nano and Clarke, 2011; Matusick et al., 2016). My studies have revealed a similar outcome and showing higher regeneration density compared to healthy conditions.

Fire is a common disturbance in Mediterranean type climates, with varying fire severities often causing different ecological impacts. I investigated the effect of prescribed burn

(typically low fire severity), moderate, and high severity wildfire on tree regeneration in the NJF. The tree regeneration varied amongst fire severity conditions, as well as drought, exposing an interaction effect. Fire is known to promote the germination and growth of several Eucalyptus tree species (Abbott, 1984; Wardell‐Johnson, 2000; Smith et al.,

2014). The removal of understorey competition, FWD and CWD during a fire, can expose the canopy floor and provide nutrients, and enhance water availability thereby assisting

65 | P a g e in the germination and establishment of species (Fernandes et al., 2018). In this study, forest densities were seen to be higher in drought-affected plots compared to healthy plots. An interaction effect further influenced this with different fire severities, and drought-affected conditions showing higher regeneration densities compared to healthy conditions of the same fire severity. Interactions between drought-induced canopy gaps and fire severity in resprouting forests can increase the availability of nutrients and resources, including, light and water, allowing for the establishment of new germinants and pre-existing resprouters (Muscolo et al., 2014). For example, in a mixed Eucalypt forest, in Victoria, Australia, regeneration counts differed between moderate and high severity fire, with high severity sites displaying substantially higher densities by comparison, including lower post-fire mortality rates (Bennett et al., 2016).

In this study, plant densities were seen to be higher in drought-affected plots compared to healthy plots. This was further influenced by different fire severities, with drought- affected conditions showing higher regeneration densities compared to healthy conditions of the same fire severity. This study demonstrates that drought-induced canopy die-off can positively influence tree regeneration regardless of the level of fire severity.

Drought and fire influence on overstorey species regeneration In this study, moderate fire severity was shown to have the highest regeneration densities of Eucalyptus marginata, although the influence was greatest within the drought-affected plots. However, prescribed burn and high fire severity plots were shown to have greater regeneration counts when burned in drought-affected conditions, compared to the healthy plots. Corymbia calophylla showed a similar interaction effect in moderate and high fire severities in drought-affected conditions. Differences in regeneration densities are possibly influenced by environmental cues such as fire regime parameters (e.g season and timing), temperature, light, soil nutrients and water availability (Bell, 1994).

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Unfavourable levels of soil nutrients, light and temperature, can restrict, postpone or delay the successful germination, establishment or survival of new or established regeneration

(Bell et al., 1999; Cochrane, 2017). However, soil moisture content is recognised as the greatest cause of regeneration mortality, owing to the overhead canopy and species competition (Stoneman, 1994).

The results of this study are consistent with past research of drought-induced canopy die- off and fire influence, where the drought-affected conditions have been shown to increase seedling and coppice densities. For example, Matusick et al. (2016) observed a 53% higher density of E. marginata and C. calophylla regeneration under conditions where drought-induced canopy die-off had occurred, compared to a healthy canopy structure. In addition, Vivian et al. (2008) reported differences in regeneration amongst fire severities with high severity significantly increasing densities compared with low fire severity in eastern Australian mixed Eucalypt (E. delegatensis and E. fastigata) forests. This study has shown positive interactions between drought and fire, revealing moderate fire severity in a drought-affected environment had 68% higher establishment than control (drought- affected) when overstorey species are combined. This highlights a positive interaction between drought and fire on the dominant tree species. However, other factors may also be important such as, past disturbances (time since disturbance; (Guinto et al., 1999), individual species requirements (Vivian et al., 2008), and environmental conditions such as the availability of ‘safe sites’ (Bradstock, 1991; Pausas et al., 2003; Fernandes et al.,

2018).

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5.2 HOW DO GROWTH STAGES OF TREE REGENERATION VARY BY DROUGHT AND FIRE SEVERITY? Overstorey tree regeneration of germinants and lignotuberous seedlings following drought and fire - stages one and two In the NJF, E. marginata and C. calophylla germinants (stage one) and lignotuberous seedlings (stage two) were both found within prescribed burn and control (long since burnt) plots, regardless of drought and fire influence. Further, they displayed similar regenerative growth stages. However, moderate and high severity fire plots only displayed lignotuberous seedlings, with both displaying similar frequencies. Eucalyptus marginata and C. calophylla, share similar disturbance adaptations, such as resprouting, fire persistence, lignotuber development (Burrows, 2002, 2013; White-Toney, 2020), thereby assisting in explaining their germination abilities.

Microclimate conditions associated with different fire severities can influence regeneration. For example, E. marginata germination is restricted to cooler and wetter conditions, with a higher level of survival and growth in exposed soil environments than in drier conditions (Abbott and Loneragan, 1986; Stoneman, 1992; Bell, 1994; Bond and

Van Wilgen, 2012). The microsite condition influenced by litter cover and location

(warmer and drier) of prescribed burn plots may differ in comparison to the wildfire plots

(less litter cover, and located in a cooler and wetter region), so late germination in the prescribed burn plots is likely (McCaw et al., 2011). The Yarloop bushfire occurred approximately three years before this study survey. Given the lignotuber develops over approximately two years (Abbott and Loneragan, 1984), the absence of germinants in wildfire plots is possible, as surviving individuals would have grown into a lignotuberous seedlings, potentially explaining the different frequencies. In previous studies in the NJF, the stocking rates of lignotuberous seedlings were considerably lower, compared to the results in this study, for example, 8% (1935) (Abbott and Loneragan, 1984), 7% (1938)

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(Stoate and Helms, 1938) and 0% (1982) (Abbott and Loneragan, 1984). By comparison, to the frequencies reported in this study were approximately 20% - 25%. Contributory factors such as changes in the forest structure (timber harvesting, drought and fire), can increase the potential of resource availability, kill off or promote regeneration growth following the disturbance (Abbott and Loneragan, 1986). The NJF has significantly changed in the past 40 years due to the extensive logging, which may have exposed the ground floor to higher levels of solar radiation, promoting growth.

Overstorey tree regeneration of seedling coppice and ground coppice following drought and fire - stages three and four In this study, when seedling coppice (stage three) and ground coppice (stage four) densities are combined, both overstorey species, E. marginata and C. calophylla showed significantly higher resprouting plants in different drought and fire conditions.

Further, Eucalyptus marginata presented a higher abundance of regeneration after moderate severity fires compared to other fire severities in both drought conditions.

However, an interaction effect between drought-affect and fire can be seen in the prescribed burn and high fire severity plots, when burned. Corymbia calophylla also showed a moderate fire severity influenced regeneration in healthy and drought-affected plots. Ground coppice was significantly less abundant than seedling coppice regeneration for both species. Lignotuber seedling and sapling densities can influence the seedling coppice rates following fire disturbance. The pre-fire seedlings will coppice once burnt, forming a seedling coppice and ground coppice growth, due to the presence of the lignotuber, thus, increasing coppice density (Abbott, 1984). This is likely to increase resprouting rates, as previous lignotuberous seedling distribution drives the densities.

Due to species resource allocation, repeated or lower severity fires can restrict the growth rate more than less frequent fires with higher severity. (Burrows, 2013).

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This is consistent with observations by Werner (2012), where stem heights were reduced and increased basal resprouting were found to follow repeated prescribed burns on various Eucalyptus species in Kakadu National Park (Northern Territory, Australia).

Furthermore, Fairman et al. (2018) reported similar findings with short fire intervals increasing mortality and lowering stem growth in a mixed dry sclerophyll Eucalyptus forest in southwestern Victoria. Limited literature exists on these resprouting stage regeneration densities following drought or fire disturbance; nevertheless, the current literature may clarify the differences in seedling and ground coppice differences. The continuous defoliation of resprouting Eucalyptus species may negatively impact the resprouting ability by exhausting carbohydrate reserves, hence restricting growth and the replacement of shed foliage (Collett and Neumann, 2002; Wills et al., 2004). This study reveals an opening for further research in the effects of fire severity on juvenile resprouting species as fire regimes change in conjunction with prescribed burning techniques.

5.3 HOW DOES TREE REGENERATION RESPOND TO FIRE OVER TIME? Tree regeneration response to time since prescribed burn Prescribed burning techniques were undertaken from 2016 to 2018. Regeneration densities in my study were influenced by an interaction between drought condition and time since prescribed burn. Drought-affected plots showed greater regeneration densities than healthy (non-drought affected) plots. However, both drought conditions displayed less regeneration between 1.5years and 3.5years post-fire. The influence of lower precipitation rates can promote water stress, coupled with higher temperature and species competition can restrict regeneration or increase regeneration mortality (Stoneman and

Dell, 1993; Marañón-Jiménez et al., 2013). In the Jarrahdale region of the NJF (central to my study; Figure 7), annual precipitation rates have declined since 2016, with the highest 70 | P a g e reduction in rainfall between 2018 and 2019, 1213mm vs 878mm (BoM, 2020b). A decrease in water availability can limit the success of germination or cause mortality before developing the lignotuber (Yates et al., 1996) Additionally, the mean maximum summer monthly temperatures for the region have increased between 2016 and 2019 by an average 1.5 °C (BoM, 2020b). These changes in short-term weather conditions could explain the significantly lower density of regeneration between 1.5years and 2.5years periods.

FORESTCHECK monitoring (Anon, 2012), is an integrated forest monitoring project to highlight changes and trends in key elements of forest biodiversity in relation to silviculture, FORESTCHECK recommends stocking rates of the NJF (eastern region) of approximately 5,605/ha E. marginata and C. calophylla regeneration (McCaw, 2011;

McCaw et al., 2011; Anon, 2012). My study recorded considerably fewer regeneration counts in each year since prescribed burn. Moreover, the unburnt (control) conditions, indicated the highest regeneration densities were also substantially less (healthy: 649/ha and drought-affected: 1,035/ha) in comparison to FORESTCHECK monitoring. The potential reasons behind this variance are the silviculture studies may recommend higher stocking rates and may not reflect the forests, natural carrying capacity for the species.

Tree regeneration response to time since wildfire In the drought-affected plots of NJF, fire severity has influenced tree regeneration density in this study. Regeneration in drought-affected conditions was significantly higher than healthy (non-drought affected) plots. High fire severity displayed significantly less regeneration than moderate severity and control plots. However, forest regeneration in the drought-affected conditions under moderate fire severity was significantly higher in

71 | P a g e each survey period (2016 to 2020). The higher regeneration counts in the wildfire plots could be a consequence of southern region weather conditions (McCaw, 2011). The mean annual precipitation rates for the Dwellingup region (closest climate location to the study region) have reduced over time, the largest reduction occurring between 2017 and 2019 with 1272mm vs 841mm, (BoM, 2020b). Additionally, the summer average maximum monthly temperature increased by 1.2 °C between 2016 and 2019 (BoM, 2020b). On the other hand, the eastern region displayed a slightly higher rainfall average, yet a higher temperature variance in the same years recorded in Jarrahdale (closest climate location to the study region; rainfall: 1197mm vs 876mm, temperature: 1.4 °C) (BoM, 2020b) The variations between rainfall and temperature between the two regions may influence survival ability and mortality rate amongst overstorey species. This may assist in explaining why the eastern regions having a lower survival rates after germination than the southern region.

Wildfire severity plots may have produced differences in the microsite conditions to the prescribed burn plots such as a higher level of soil exposure or greater canopy gap (Abbott and Loneragan, 1986; Stoneman, 1992; Bell, 1994; Bond and Van Wilgen, 2012).

FORESTCHECK stocking for the southern region is different from the eastern region, with approximate densities of 12,000/ha (McCaw, 2011; McCaw et al., 2011; Anon,

2012). The results reported in this study show drought-affected conditions under moderate severity 3-years post-fire displayed similar stocking levels (11,780/ha). These were the only plots and time period to reach capacity, according to the relevant

FORESTCHECK densities. The decrease in rainfall and change in temperatures potentially change the forests ability to regenerate (Bell, 1994; Dale et al., 2001a; Torres et al., 2006). This potentially explaining why lower than normal regeneration rates were recorded. The lower overstorey regeneration densities can create a successional shift 72 | P a g e within the plant species community. Thus, creating windows of opportunity for alternative states (Batllori et al., 2017). Therefore, additional research involving fire

(prescribed burning and wildfire disturbance) needs to be investigated, to review

FORESTCHECK monitoring densities. Ongoing research and examination of the long- term trajectory in all regions of the NJF is required to document forest health regeneration and potential changes.

Overstorey species regeneration response to time since fire for germinants and lignotuberous seedlings – stage one and two This study showed there germinants were present in approximately 2.5 years following prescribed burn and one years following wildfire for both overstorey species. However, as the germinant densities decreased there was an increased presence in lignotuberous seedling densities increased overtime, predominately amongst wildfire in both species’ populations. Differences in germinant and lignotuberous seedling densities could be attributed to biotic and abiotic factors (e.g. rainfall, canopy cover, fire severity and herbivory predation). This likely influenced germinant mortality and lignotuber development (Yates et al., 1996; Allcock and Hik, 2004). In the E. luehmanniana dominated mallee region of Garigal National Park, Sydney, Tozer and Bradstock (1997) found germination mortality to occur in the first 200 days after emergence, in open canopy conditions. Other studies showed similar results in other mixed Eucalypt forests, of

Sydney and Australian Capital Territory (Faunt et al., 2006; Vivian et al., 2008). This suggests that reduced germinant density could be due to higher mortality rates.

The FORESTCHECK monitoring report outlines overstorey species distribution to be

55% (E. marginata): 45% (C. calophylla) ratio, with stocking rates differing in the eastern and southern regions (eastern 4,750/ha vs southern 9,500/ha) (McCaw, 2011). However

73 | P a g e reports by Stoate and Helms (1938) suggest lignotuberous seedlings have a density of

7,400/ha – 14,800/ha (E. marginata) and 6,660/ha – 13,320/ha (C. calophylla). Both reports showed an inconsistency with my studies in prescribed burn and high severity fires, where densities did not reach stocking limits. However, moderate fire severity in drought-affected conditions reached stocking rates at 3-years post-fire (E. marginata:

6,537/ha, and C. calophylla: 3187/ha).

There is no evidence to suggest that logging history has consequences to seedling germination or establishment in southwestern Australia McCaw (2011). Therefore, other negative interactions may lower regeneration density; for example, the timing of fires, fire severity and seed viability. One report suggests E. marginata viable seed can be as low as 31% following disturbance, though this was on the Swan Coastal Plain (Ruthrof et al., 2002). There is limited information in this field that reviews different growth stages in tree regeneration. This highlights various opportunities for future studies in new regeneration survival for the NJF, and the drivers contributing to those density changes compared to current silviculture records.

5.4 MIDSTOREY SPECIES RESPONSE TO DROUGHT AND FIRE I also investigated the effects of drought and fire severity on midstorey species

(Allocasuarina fraseriana, Banksia grandis, Persoonia longifolia and ) in the NJF. This investigation showed an absence of P. elliptica and this species was removed from the analysis. This study revealed the midstorey species to be the least resistant to drought, fire or a combination of both compared to the overstorey species.

There were significantly lower regeneration density counts, thus impeding the ability to perform statistical analysis to determine a clear understanding of the effects of drought

74 | P a g e and fire severity. However, through graph analysis and observation, drought was the biggest influencing factor in decreasing regeneration density. The studied midstorey species are highly susceptible to drought and water stress once established (Matusick et al., 2013; Chia, 2016), with the midstorey species studied here being used as an indicating variable for high drought conditions (Matusick et al., 2012; Walden, 2020). In addition, the midstorey species such as B. grandis and Persoonia spp. require higher soil moisture for germination than the overstorey species surveyed (Abbott, 1984). Other contributing factors exist that may determine the low-density counts. For example, studies on

P.longifolia found the germination conditions to be very narrow, and they suffer high mortality rates caused by a combination of water stress and herbivory, as well as low seed viability (Koch and Samsa, 2007; Chia, 2016). Banksia grandis has been reported to be negatively influenced by fire and thus may be less likely to resprout after severe fires (Burrows, 1985) and some Allocasuarina species have been reported to suffer high mortality rates within and around rocky outcrops (Benwell, 2007). The factors outlined here may have led to low levels of regeneration. Therefore, further investigation into the midstorey species of the NJF is advised to investigate other factors limiting establishment.

5.5 LIMITATIONS AND MANAGEMENT IMPLICATIONS In this study, I used two main factors, fire severity (long-since burn, prescribed burn, moderate and high severity wildfire) and drought condition (non-drought affected and drought-affected) to understand germination and lignotuber seedling responses to a drought and fire severities in the NJF, southwestern Australia. I found germination, lignotuberous seedling counts and resprouting ability were affected by fire severity and drought condition. Midstorey species were suspectable to drought with a lower regeneration count recorded in drought-affected plots when compared to healthy conditions.

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Whilst this study provided some supporting evidence of regeneration differences in response to drought and fire, it is not without its limitations and results may have the potential to change, depending on the importance of other abiotic variations not recorded.

Abiotic factors such as fire intensity, canopy cover, soil moisture and soil nutrient availability are likely to change, thereby potentially producing different results as they influence plant responses to drought and fire. Other factors such as fire scorch height, canopy cover percentage, soil moisture content and phosphate and nitrate levels may provide greater insight into the conditions influencing regeneration density and should be examined in the future.

One major limitation of this study was the lack of existing data on the detailed regeneration stages. This provided challenges in distinguishing density volumes when compared to my data. Furthermore, the use of FORESTCHECK monitoring and several studies of the stocking rates of the NJF were for the benefit of silviculture, thus perhaps producing above the natural carrying capacity of the forest. Drought and fire disturbances have been extensively researched, along with their effect on forest regeneration.

However, limited studies exist on their interaction effect to forest tree regeneration densities.

The limited available data highlights a new area of possible research on forest regeneration. A thorough investigation of the different stages of growth may provide further insight into how species develop and respond over time. This could help to determine a more current natural carrying capacity as the NJF structure changes.

Determining densities that are more common in a natural forest environment may also 76 | P a g e promote the understanding of how the forest is likely to change in the future based on the results of my study and predicted climatic changes for the region. This may be helpful in the future if the species that die out to help their population be more representative of a natural forest, and therefore support the forest dynamics.

It is difficult to determine the specific level of canopy reduction from drought and fire that may have influenced tree regeneration, as canopy cover percentage was not recorded between plots, and I relied upon a baseline of "healthy" and "drought-affected" conditions. Nonetheless, this study was a starting point to understand the interaction effect, particularly for overstorey species (E. marginata and C. calophylla). As the climate changes and disturbances such as drought and fire are becoming more common, determining the interaction is a vital component for future forest persistence. Increased research on tree regeneration needs to occur, potentially investigating the long-term trajectory of survival on each growth stage. This information will be vital in informing forest health and composition as the climate changes and drought become more frequent.

The density abundance was challenging to interpret, due to the counts that were recorded throughout the survey period. Nevertheless, this study highlighted the low-density distribution and the absence of P. elliptica regeneration within the eastern and southern region of the Northern Jarrah Forest. As highlighted previously, midstorey species are highly susceptible to drought, and fire disturbance. With a changing climate, drought and fire extremes are expected to become more frequent so it will be beneficial to map the midstorey populations of these species. Through doing so, we can use the information in several ways to promote the population of drought susceptible midstorey species. Firstly, to determine the trajectory of these species as regeneration and mature standing trees.

Secondly, to build a greater understanding of the environmental drivers that enables

77 | P a g e germination and establishment in regeneration. Third, to assist in prescribe burning practices to avoid densely populated areas and use other fuel reduction techniques that may be more suited for these species. Persoonia spp. in the region may benefit from these understandings

Mitigation techniques to manage wildfires and to assure stocking rates are kept at a recommended capacity in the NJF will rely on decreasing the extent of high severity wildfire. The use of prescribed burning is one method; however, the evidence in this thesis has shown low intensity prescribed burns have little effect on increasing abundance of E. marginata and C. calophylla seedlings. Therefore, higher prescribed burn severities at longer burn intervals assist in elevating germination densities, and the increase of saplings or juvenile trees in regions where mature tree densities are low, thus allowing for a higher seed bank. The techniques of fuel reduction and fire mitigation opens gateways into several areas of future research. Within this study, the prescribed burns were undertaken in the eastern region of the NJF, where the region is predominately drier and warmer than the southern region. Investigating the regeneration response of prescribed burning against

(Aboriginal) cultural burning and other fire severities (low, moderate and high) in the different regions (east, south, west), seasons (spring and autumn) of the NJF and Southern

Jarrah Forest may deliver new insights and review of our current methodologies.

There are still several areas of the fire regime we still do not understand, and with a changing climate the knowledge gap is widening however, this is a starting point and the research suggestions outlined in this thesis should be considered for future management options.

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5.6 CONCLUSION Drought and fire can influence tree regeneration, through limiting or promoting growth following disturbance. This study found evidence indicating the potential for higher tree regeneration in the fire-tolerant eucalypt forests following drought and fire in the NJF or

Southern Jarrah Forest. However, several questions remain unanswered such as the long- term trajectory and mortality of regeneration following establishment and sapling stage.

Drought-affected areas can influence tree regeneration significantly following fire; however, the level of fire severity determines the species densities response. Moderate fire severity having the greatest influence on early regeneration for seedling germination in overstorey species, resprouting capabilities may be more variable and dependent on previous seedling establishment. This study will help create a baseline of responses to drought and fire interaction and develop an understanding of regeneration responses to a drying climate. In addition, it will assist in examining potential structural change occurring in the NJF as rainfall declines and temperature increases in the future.

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APPENDIX A –STUDIED STAGES OF GROWTH The stages of growth of Eucalyptus marginata, in accordance with (Abbott and

Loneragan, 1984). Including a closeup image of a Lignotuber

Stage of growth Germinant Lignotuberous seedling Seedling coppice

Ground coppice Sapling Lignotuber

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APPENDIX B – STUDIED SPECIES All species studied for the thesis. From left to right (Eucalyptus marginata, Corymbia calophylla, Banksia grandis, Allocasuarina fraseriana, Persoonia longifolia, and

Persoonia elliptica - photo no available).

Studied species Eucalyptus marginata Corymbia calophylla

Lignotuberous seedling Lignotuberous seedling

Banksia grandis Allocasuarina fraseriana

Seedling coppice Seedling coppice

Persoonia longifolia (Chia, 2015) Persoonia elliptica

SPECIES NOT FOUND

Seedling

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