Spatio-Temporal Variation in the Germinable Soil Seed Bank Communities Characterizing East- Mediterranean Woodlands

Spatio-temporal variation in the germinable soil seed bank communities characterizing East- Mediterranean woodlands

Thesis submitted in partial fulfillment

of the requirements for the degree of

“DOCTOR OF PHILOSOPHY”

Neta Manela

Submitted to the Senate of

Ben-Gurion University of the Negev

February 2019

Beer Sheva Spatio-temporal variation in the germinable soil seed bank communities characterizing East- Mediterranean woodlands

Thesis submitted in partial fulfillment

of the requirements for the degree of

“DOCTOR OF PHILOSOPHY”

Neta Manela

Submitted to the Senate of

Ben-Gurion University of the Negev

February 2019

Beer Sheva

Approved by the advisor …………………………………………

Approved by the Dean of the Kreitman School of Advanced Graduate Studies

…………………………………………………………………………. This work was carried out under the supervision of:

Professor Ofer Ovadia, Department of Life-Sciences, Faculty of Natural Sciences, Ben-Gurion University of the Negev

Consultant:

Doctor Hagai Shemesh, Department of Environmental Sciences, Tel-Hai College Research-Student's Affidavit when Submitting the Doctoral Thesis for Judgment

I, Neta Manela, whose signature appears below, hereby declare that:

X I have written this Thesis by myself, except for the help and guidance offered by my Thesis Advisors.

X The scientific materials included in this Thesis are products of my own research, culled from the period during which I was a research student.

This Thesis incorporates research materials produced in cooperation with others, excluding the technical help commonly received during experimental work. Therefore, I am attaching another affidavit stating the contributions made by myself and the other participants in this research, which has been approved by them and submitted with their approval.

Date: 25/02/2019 Student's name: Neta Manela Signature: Table of contents

Abstract 1

1. General introduction 5

1.1. Chapter 1: The effects of fire season and microhabitat on the 11 composition of the germinable soil seed bank community in East- Mediterranean woodlands

1.2. Chapter 2: The effects of smoke and fire history on the soil seed bank 14 germination in Mediterranean woodlands

1.3. Chapter 3: Long-term effect of a pulse perturbation experiment 16 involving canopy opening on an endangered population of Paeonia mascula in an East-Mediterranean forest

2. Methods and materials 19

2.1. General methods 19

2.1.1. Soil sampling and seedling emergence monitoring 19

2.2. Experiment 1: Fire season and microhabitat specific methods 20

2.2.1. Study site 20

2.2.2. Experimental design 20

2.2.3. Soil sampling and seedbank assessment 21

2.2.4. Statistical analyses 22

2.3. Experiment 2: Fire frequency and smoke specific methods 23

2.3.1. Study site 23

2.3.2. Fire history 23

2.3.3. Soil sampling and experimental design 24

2.3.4. Aerosol smoke treatment 25

2.3.5. Statistical analyses 25

2.4. Experiment 3: Tree clearing (Paeonia) specific methods 26

2.4.1. Study site 26 2.4.2. Experimental design 26

2.4.3. Paeonia mascula population measurements 27

2.4.4. Soil sampling and seedling emergence of neighboring 27 cooccurring species

2.4.5. Canopy closure 27

2.4.6. Statistical analyses 27

3. Results 29

3.1. Experiment 1: Fire season and microhabitat results 28

3.1.1. Fire intensity and severity were consistent between the two 28 burning seasons

3.1.2. The effect of fire treatment and microhabitat type on the 28 germinable soil seed bank density, richness and diversity

3.1.3. Life forms 32

3.1.4. Community composition 33

3.1.5. Unique species 35

3.2. Experiment 2: Fire frequency and smoke results 36

3.2.1. Community composition 39

3.2.2. Species level 40

3.3. Experiment 3: Tree clearing (Paeonia) results 42

3.3.1. Paeonia mascula population 42

3.3.2. Germinable soil seed bank 43

3.3.3. Canopy closure 45

4. Discussion 47

4.1. Discussion Experiment 1: Fire season and microhabitat 47

4.2. Discussion Experiment 2: Fire frequency and smoke 50

4.3. Discussion Experiment 3: Tree clearing (Paeonia) 53

4.4. General discussion 56

5. Cited literature 62

6. Supporting information 78

List of Figures

Figure 1: Paeonia mascula (a) flower and (b) seedling at the study site in Ein 18

Hazaken, Mount Meron, Israel

Figure 2: Experiment 1: The study site in Mount Ya'aran, Israel 21

Figure 3: Experiment 2: The study area in Mount Carmel, Israel 24

Figure 4: Experiment 1: Density (a), richness (b) and diversity (c) of the 31 germinable soil seed bank community (mean± SE) characterizing each of the three fire treatment groups and microhabitat type

32 Figure 5: Experiment 1: Annual herbaceous density (a) and richness (b), and dwarf-shrub density (c) and richness (d) (mean ± SE) in each of the three fire treatment groups.

Figure 6: Experiment 1: Canonical analysis of principal coordinates (CAP) based 35 on Bray-Curtis dissimilarity matrix, best discriminating samples among the unburned control (circles), autumn (squares) and spring burned (triangles) germinable soil seed bank communities, including correlations of individual plant species with the canonical axes.

Figure 7: Experiment 2: Density and richness of the germinable soil seed bank 37 community as a function of fire frequency and smoke treatment. The smoke treatment included: control (black), exposure to smoke (light grey), and watering followed by exposure to smoke (dark grey). Graphs represent the total germinable soil seed bank (a) density, and (b) richness, and the annual (c) density and (d) richness (mean± SE)

Figure 8: Experiment 2: Canonical analysis of principal coordinates (CAP) based 40 on Bray-Curtis dissimilarity matrix, best discriminating samples among the different fire frequency by smoke treatment combination groups. Different colors represent sites that have been subjected to one fire (white), three fires (grey), or no fire (black) during the last four decades. Different shapes represent the smoke treatment (triangle), watering and smoke treatment (circle) and control no-smoke

8 treatment (square).

Figure 9: Experiment 3: Density of P. mascula seedlings and mature plants in plots 42 that were cleared from trees in 1997 and closed plots in the first and second years of survey

44 Figure 10: Density (mean± SE) of the P. mascula soil seed bank in plots that were cleared from trees in 1997 and closed plots in the first and second years of survey.

44 Figure 11 The percentage (mean± SE) of flowering P. mascula in plots that were cleared from trees in 1997 and closed plots in the first and second years of survey.

45 Figure 12 Germination (1) density and (2) richness (mean± SE) from the soil seed bank in plots that were cleared from trees in 1997 and closed plots in the first and second years of survey.

46 Figure 13 A box plot showing the mean (dashed lines) and median (solid lines) LAI (Leaf Area Index) in plots that were cleared from trees in 1997 and closed plots in the first and second years of survey.

List of Tables

Table 1: Experiment 1: Summary of the generalized linear mixed models testing 30 for the effect of fire treatment and microhabitat type on the germinable soil seed bank density, richness and diversity (Fisher’s alpha)

Table 2: Experiment 1: Results of permutational multivariate analysis of variance 34 (PERMANOVA) testing for the effect of fire treatment and microhabitat type on the composition of the germinable soil seed bank community

Table 3: Experiment 2: Summary of Generalized linear mixed models testing the 38 effect of fire frequency and smoke treatment on germinable soil seed bank density, richness and diversity (calculated by Fisher's alpha).

Table 4: Experiment 2: Results of permutational multivariate analysis of variance 42 (PERMANOVA), testing for the effect of fire frequency and smoke treatments on the composition of the germinable soil seed bank community.

Table 5: Summary of the generalized linear mixed models testing tree 43 clearing treatment, sampling year, maturation stage of P. mascula and their interaction on P. mascula density, flower percentage and on P. mascula seed density.

Table 6: Summary of the generalized linear mixed models testing for the effect 45 of tree clearing treatment on germinable soil seed bank density and species richness.

Table 7: Summary of the generalized linear mixed models testing the effect of 46 tree clearing treatment on the leaf area index.

Abstract

Plant community dynamics is highly influenced by germination from the soil seed bank. This relationship may be more important in ecosystems that are characterized by a seasonal climate, such as Mediterranean ecosystems. In such a dynamic environment, germination is limited to the wet season, and thus the composition of the soil seed bank community can strongly influence plant community dynamics. Furthermore, Mediterranean ecosystems have a long history of disturbances, and owing to variation in disturbance type, frequency and intensity, the soil seed bank in these ecosystems has the potential to play an important role in shaping plant community composition and structure. Moreover, changes in a disturbance regime can cause an ecosystem shift between alternative stable states. Such a shift usually manifests in drastic changes in vegetation formation, consequently threatening species that are less adapted to the abiotic and biotic conditions characterizing the newly established stable state. Preventing a shift in an ecosystem state for protecting species that are more sensitive to the changes in the structure and function of the altered ecosystem requires active management.

My doctoral research aimed at investigating the effect of disturbance timing, history and regime on germinable soil seed bank community composition. The soil seed bank plays a major role in maintaining plant species’ diversity by serving as a belowground reservoir for the regeneration of seeder species that may disappear from the aboveground community due to disturbances. Consequently, disturbance regime may be an important factor affecting the soil seed bank composition and its germination patterns, shaping the post-disturbance plant community. My research comprises three main chapters exploring if and how changes in germination patterns from the soil seed banks are affected by: 1) seasonal fires, 2) fire history and smoke cue, and 3) a management action involving a pulse canopy opening performed ~20 years ago. In the third experiment, I also explored the long-term consequences of the pulse canopy opening on the population of the endangered species Paeonia mascula.

The first chapter of my thesis aimed to explore if and how fire season and microhabitat (i.e., under Pistacia, under Cistus, or open gap) influence the composition of the germinable soil seed bank community in a typical East-Mediterranean woodland. To the best of my knowledge, this is the first empirical attempt to investigate the

1 effect of fire season on different attributes of the plant community, characterizing the East-Mediterranean basin, using a large- scale field experiment involving prescribed burns. Specifically, I documented seed germination patterns using soil samples collected from different microhabitats within burned and adjacent unburned control areas. Fire caused an overall reduction in the germinable soil seed bank density, richness and diversity. The reductions in germinable soil seed bank richness and diversity were significantly stronger under Pistacia and Cistus shrubs/dwarf shrubs subjected to autumn burnings and this reduction was mainly observed among annuals. Dwarf shrub density was higher in samples collected from burned sites, and this pattern was significantly stronger in samples collected under Pistacia and Cistus shrubs/dwarf shrubs. Together with the appearance of unique species, fire season led to significant changes in the composition of the germinable soil seed bank community. My results illustrated, for the first time, that fire season interacted with spatial heterogeneity to influence the composition of the germinable soil seed bank community mostly via differential effects on annual and dwarf shrub densities. These findings suggest that the overall increase in fire events during the autumn and spring seasons, which has been evident in this region during the last few decades, may translate into a shift in eco- evolutionary selection pressures, favoring plant species that regenerate after fires, thus having an advantage in the post fire environment.

In the second chapter I examined, for the first time, the interactive effect of aerosol smoke and fire history on the germinable soil seed bank community in East-Mediterranean woodlands. Specifically, I collected soil samples from sites that have been subjected to different fire frequencies during the last four decades and exposed them to aerosol smoke, with or without watering. By documenting the seed germination patterns characterizing these samples, I could examine the changes in the density and species richness of the germinable seeds in the soil. Total germinable soil seed bank density was higher in sites that were burned more frequently during the last four decades. A common notion is that smoke does not affect seed germination in the Mediterranean basin since most disturbances in this region are anthropogenic – a relatively recent development in evolutionary time scales. However, my results showed, for the first time, that exposure to aerosol smoke increased the germinable soil seed bank density, and this pattern was more pronounced in samples originating from sites burned more frequently as well as among annual species. My study highlights the importance of exploring germination responses using intact soil

2 samples rather than synthetic seed communities. Moreover, my findings emphasize the important role smoke plays in shaping post-fire succession processes in the Mediterranean basin, mainly by stimulating the germination of annual species.

The third chapter of my doctoral research aimed to examine the long-term effects of a pulse canopy opening on the germinable soil seed bank density and richness, and on population density and reproduction of the locally endangered herbaceous perennial plant Paeonia mascula. Remarkably, studies involving long- term monitoring of management actions are relatively rare, mostly focusing on short-term responses. Specifically, I quantified the germination density and richness of plant species from the soil seed bank in plots that were cleared from trees ~20 years ago, and in plots whose canopy remained closed. In addition, I monitored P. mascula plant density, flowering fraction and soil seed bank density within these plots. No differences in germinable soil seed bank density and richness were detected between the cleared and closed plots. Notably, P. mascula density was significantly lower in plots that were cleared from trees two decades ago than in plots that remained closed. In contrast to the positive short-term effects of canopy opening on P. mascula flowering, lower number of flowering plants and smaller number of seeds were documented in the cleared than in the closed plots, 20 years after the clearing took place. These results illustrate that the short-term positive effect of a pulse canopy opening can be outweighed and even over-compensated for the respective long-term negative effect. My findings also indicate that the population of P. mascula in mount Meron is stable in sites that remained closed; therefore, no intervention is necessary for its persistence.

Generally, the results obtained during my doctoral studies can have important implications for our understanding of how disturbance timing, history and regime interact to influence the composition of the germinable soil seed bank community.

Germination from the soil seed bank is a mandatory process in the establishment of a plant community, especially in the post-disturbance environment. My findings can largely improve our ability to accurately predict changes in plant community dynamics under the ongoing climatic changes occurring in the Mediterranean basin.

Key-words: Active management, Aerosol smoke, Annuals, Canopy opening, Disturbance regime, Endangered species, Fire season, Germination, Mediterranean woodlands, Microhabitat, Pulse perturbation experiment, Seeder species, Soil seed bank, Spatial historic fire regime

1. General introduction Plant community dynamics is determined by the interplay between disturbance nature, intensity and regime, and species traits, life history trade-offs and the representation of different species in the seed bank (Cowling et al. 1996, Kneitel and Chase 2004, Fenner 2012). Among all, seed bank composition is fundamental in determining the short- and long-term vegetation dynamics, especially in fluctuating environments (Fenner 2000, Fenner 2012). Moreover, a large body of empirical studies have illustrated that the soil seed bank plays a major role in maintaining plant species’ diversity by serving as a belowground reservoir for the regeneration of seeder species that may disappear due to disturbances (Pakeman and Small 2005, Anderson et al. 2012). Consequently, the disturbance regime may be an important factor generating such variations in the soil seed bank composition as well as in germination patterns characterizing the post-disturbance environment.

Soil seed bank is commonly defined as all viable seeds in the soil profile (Saatkamp et al. 2014). Soil seed bank includes both seeds that germinate within a year of initial dispersal (i.e., the transient component) and seeds that remain dormant in the soil for longer time periods (i.e., the persistent component) (Nathan and Muller-Landau 2000). Seed dormancy provides a mechanism blocking germination of viable seeds under conditions when the probability of seedlings survival and growth is low. A diverse range of dormancy mechanisms have been documented in a variety of environments and habitats (Baskin and Baskin, 2004). Dormancy classification is based on the developmental state of the embryo during dispersal (i.e., morphological dormancy), the physical traits of the seed (i.e., physical dormancy), or the physiological responses of seeds to environmental stimuli (i.e., physical dormancy) (Baskin and Baskin, 2004). For example, physical dormancy is caused by a water-impermeable seed coat and it is common in the Fabaceae family (Willis et al. 2014). Controlling the timing of germination by seed dormancy can strongly affect processes of colonization, adaptation, speciation, and extinction. Under natural selection pressures, dormancy can regulate the timing of germination, determining population colonization and distribution (Donohue et al., 2010).

Generally, the dynamics of the soil seed bank is determined by complex interactions and processes dictating its inputs, losses and continuity. The inputs are mostly attributed to dispersal into the soil seed bank. Seed dispersal patterns can be 5 determined by environmental factors, species characteristics, and genetics. Seed losses can occur mostly due to seed germination and predation (in large seeds), but also by seed decaying, naturally by senescence or by exterior predation, pathogen attack, or environmental conditions facilitating decomposition of organic matter (e.g., high soil moisture and warm temperatures) (Leck 2012). Continuity refers to dormant abilities of seeds in the seed bank, allowing their persistence in space and time. Although these processes are inherent in any plant species, the soil seed bank composition often varies greatly between habitats and ecosystems, along with the relative contribution of each process to the soil seed bank dynamics (Leck 2012).

Germination dynamics from the soil seed bank have been studied for many years due to their influence on plant community structure and function (Izhaki et al. 2000, Thompson 2000). Investigating seed germination dynamics is fundamental for predicting the potential changes in vegetation composition in response to environmental changes and recurrent disturbances (Koller and Kozlowski 1972, Walck et al. 2011, Santana et al. 2014). Specifically, soil seed banks may be more important in areas characterized by a seasonal and variable climate (Jiménez and Armesto 1992, Bonis et al. 1995, Anderson et al. 2012) than in areas where seedling recruitment is not limited to one season (Lavorel et al. 1993). In these dynamic ecosystems, the seed bank can increase the time to extinction by functioning as a buffer of plant population against environmental variability (Kalisz and McPeek 1993). Such environmental variability can occur in time and space with species differentially influenced by each scale. Seed banks can also have an important role when the environment fluctuates between years due to variation in climate, herbivory, pathogens and disturbances (Saatkamp et al. 2014).

History of disturbances, including fire, grazing and clear-cutting, is an inherent characteristic of the Mediterranean basin (Le Houérou 1973, Thompson 2005). Such disturbances most likely affected the plant evolution and still affect current dynamics of both standing vegetation and its recruitment from the soil seed bank (Bond and Keeley 2005, Verdú et al. 2007). Also, due to its location among three continents, huge topographical diversity and a unique climate, vegetation diversity within the Mediterranean region is relatively high, with around 25,000 vascular plant species comprising ~8% of the world’s flora (Perevolotsky 2005, Valavanidis and Vlachogianni 2011). Different from other Mediterranean climate ecosystems, the East-Mediterranean 6 basin has been subjected to a longer history of intensive land uses (Thompson 2005). As a result, most disturbances in this region are mainly anthropogenic, a relatively recent development in evolutionary time scales (Ne'eman et al. 2012). Consequently, plant species in this area are assumed to be less adapted to fires (i.e., owing to the shorter evolutionary time scale it is most likely that less fire-related traits, such as post-fire flowering, and smoke-induced seed germination, have evolved in our region) compared to in other Mediterranean ecosystems, and so the germination patterns are expected to be different in the post-disturbance environment (Ne'eman et al. 2012).

Many species in this area have a large accumulated soil seed bank, which is important for the regeneration of the vegetation in this dynamic ecosystem. Moreover, many studies have investigated seed traits, which enable plant regeneration in the post- disturbance environment (Izhaki et al. 2000, Goubitz et al. 2002, Crosti et al. 2006). An important example is regeneration after fire, which was well studied in the Mediterranean basin (Keeley et al. 2011a). Resprouting is a general adaptation of plants to aboveground damage which can be caused by disturbances including fire. Resprouting occurs by the activation of dormant vegetative buds to generate new shoots (Pausas, 2001, Pausas et al, 2014). Common resprouters in Mediterranean woodlands and shrub-lands are mainly broad-leaved shrubs and trees belonging to the genus Pistacia, Quercus and Rhamnus, all of which resprout in response to grazing, clearcutting and fire, and are dominant species in the East-Mediterranean basin (Pausas et al. 2009, Keeley et al. 2011a). A major strategy adopted by plants in diverse environments is seedling recruitment (Izhaki et al. 2000). Species that recruit solely from seeds germinate in response to fire, or to fire-related cues, and their seeds are often capable of staying dormant for long time periods (Keeley and Fotheringham 2000). Prevalent examples in the Mediterranean basin are the common species families Cistaceae and Fabaceae whose seeds are found in physiological dormancy, characterized by a hard coat that can be broken by multiple triggers, including acid scarification, cold, and heat (Bradshaw et al. 2011). The germination of seeds of different species in these two families is heat-dependent to the extent that their respective populations are subjected to local extinction in the absent of fire events (González-Rabanal and Casal 1995, Herranz et al. 1998, Izhaki et al. 2000, Moreira et al. 2010). Notably, most studies investigating the effect of fire on germination in the East-Mediterranean basin have focused on germination response to fire or to heat (Ne'eman and Izhaki 1999, Izhaki et al. 2000),

7 while the effect of germination in response to other fire cues such as smoke is considered to be less common (Ne'eman et al. 2012). However, during the last decade studies in the Mediterranean basin showed that smoke can have a positive effect on the germination of perennial and annual species (Çatav et al. 2012, Tormo et al. 2014, Chamorro et al. 2017, Moreira and Pausas 2018). The effect of smoke can trigger seed germination directly by penetrating the seeds, or indirectly by absorption onto soil particles and later releasing chemicals in vapors or aqueous leachate (Keeley and Fotheringham 2000).

Although fire is a major force that shapes plant communities in the Mediterranean basin, there are many plant species that are able to persist under other disturbance regimes such as grazing and clear-cutting, which open canopies and in so doing maintain the populations of shade-intolerant species (Perevolotsky and Seligman 1998, Perevolotsky 2005). Such disturbances have become an integral part of the Mediterranean ecosystem (Grubb and Hopkins 1986), meaning that changes in their regime could have important implications on community dynamics. Changes in disturbance frequency, timing, severity and other attributes can exert strong selective pressure on plant traits that can directly and indirectly affect their regeneration (Keeley et al. 2011b). Consequently, regeneration strategies evolve in response to disturbance regimes rather than individual disturbance events (Keeley et al. 2011b).

The ability of an ecosystem to recover after disturbances is generally determined by its resilience. Strong resilience of the plant community regulates the ability to return to its original state when the disturbance causes a shift in community composition. However, when a disturbance regime undergoes a radical change, a shift between alternative stable states can occur. This shift can take place when a severe disturbance breaks the resilience of the ecosystem (Liu et al. 2018). Common examples for alternative stable states are woodlands and grasslands. Landscapes dominated by grasses can be maintained by fires, grazing and clearcutting. When these events become less common, a woodland can be established and once established, these woodlands cannot be eliminated by grazing (Scheffer et al. 2001). However, a drastic management regime can reverse this change. With the aim of increasing the resilience of the ecosystem while weakening the changes driven by a disturbance regime shift, ecosystem management requires a comprehensive understanding of the impacts of disturbances on different aspects of ecosystem biodiversity (Thom and Seidl 2016). 8

An example of processes commonly involving active management are canopy gaps. In forested areas, canopy gaps are a common outcome of natural or anthropogenic disturbances, including treefall, grazing and thinning (Whitmore, 1989, Bullock, 2000). Due to their ubiquitous nature, the temporal dynamics of gap closure and seedling establishment in canopy gaps have been studied in a variety of ecosystems, including forests, grasslands and woodlands (Peart, 1989, Koukoulas and Blackburn, 2005, Kneeshaw and Bergeron, 1998). Gaps are usually defined as time-limited competition free spaces (Bullock, 2000). The key point in this definition is that canopy gaps are temporary, and that the re-establishment of competitively superior species is just a matter of time. The colonization potential of a gap is a consequence of gap characteristics, including size, shape and spatial position, as well as species-specific light requirements, ensuring seedling establishment and survival inside the gap, and in its close surroundings (Whitmore, 1989). Species colonizing gaps often differ from the surrounding vegetation, mainly in terms of relative abundances. These differences stem from the differential ability of species to disperse into and to establish in the gaps (Whitmore, 1989). Such gap colonizers are often characterized by small seed size and high dispersal ability (Pakeman and Small, 2005). However, the correlation between seed size and establishment responses in canopy gaps, do not always exist, especially when more than two species are taken into consideration (Leishman et al., 2000).

Gap traits are likely to interact with plant traits in determining the community composition of plants in the colonized gap. For example, seeds of shade-intolerant species are more likely to emerge in larger gaps, where light intensity is higher, while shade tolerant seeders are more likely to dominate small gaps (<5 m2) (Yamamoto, 2000, Poulson and Platt, 1989). Time to establishment can also correlate with plant traits, with pioneer species often being shade intolerant, while late successional species being shade-tolerant, often found in the understory and in old gaps (Bullock, 2000).

Owing to the complex dynamics of the soil seed bank and the difficulty of distinguishing between the input and output factors dictating seed bank dynamics, my doctoral study aimed to characterize seed germination patterns in Mediterranean woodlands. Specifically, the principle aim of my study was to examine how the composition of germinating soil seed bank varies at different spatiotemporal scales in response to several disturbances occurring in East-Mediterranean woodlands. Three common processes linked to seed germination dynamics in Mediterranean ecosystems 9 were investigated:

a) Chapter 1: The interaction between the effects of fire season and microhabitat on the composition of germinable the soil seed bank in East-Mediterranean woodlands.

b) Chapter 2: The interaction between the effects of fire history and smoke on the composition of the soil seed bank germination in East-Mediterranean woodlands.

c) Chapter 3: The long-term effect of a pulse canopy opening on the locally endangered population of Paeonia mascula, and on the respective germination of neighboring plant species in an East-Mediterranean woodland.

My study aimed to investigate different aspects related to seed germination dynamics in the East-part of the Mediterranean basin. This highly heterogeneous region in its climate topography and is characterized by a rich flora (Rundel et al. 2013), has long been subjected to a various of disturbances including fire, drought and other anthropogenic and natural catastrophes (Thompson 2005). These characteristics suggest that most Mediterranean ecosystems are likely to be in a non-equilibrium state, whereas germination from the soil seed bank has the potential to play an important role in shaping plant community composition. Such heterogeneity can allow the establishment of suppressed individuals in space and time, also at the time of disturbance or immediately after it, therefore changing the plant community dynamics (Keeley et al., 2005). Previous studies in this region have illustrated that germination from the soil seed bank can be affected by changes in environmental conditions, while also highlighting the importance of its dynamics to ecosystem conservation and management (Ne'eman and Izhaki 1999, Santana et al. 2014). The present study aimed to fill gaps in our knowledge regarding seed germination dynamics in East-Mediterranean ecosystems, and to improve our understanding of the short- and long-term consequences of natural and anthropogenic disturbances for plant community composition.

The next section includes the specific introduction of each of the three main research topics I investigated as part of my doctoral studies.

1.1. Chapter 1: The effects of fire season and microhabitat on the composition of the germinable soil seed bank community in East-Mediterranean woodlands

Disturbances in general and fire in particular, can strongly influence the composition, structure and function of plant communities (Lloret 1998, Sottile et al. 2015). Post-fire succession is generally determined by fire intensity and frequency combined with community resilience (Fletcher et al. 2014). These factors regulate the ability of the plant community to return to its original pre-fire state when the fire causes a shift in community composition. However, when fire frequency and/or intensity cross a certain threshold, plant community composition can shift from its original stable state to an alternative one (Beisner et al. 2003). Fire-prone ecosystems may be more resilient to fire, retaining their community composition even after high intensity and/or frequent fires. These ecosystems are highly dependent on large seed banks, acting as a buffer of plant community composition against environmental variability and disturbances (Kalisz and McPeek 1993).

Mediterranean-type regions serve as an example for soil seed bank-dependent ecosystems (Verdú 2000). Many seeder species that are shade intolerant can survive fires solely through germination from soil seed bank (Fenner 2000). For instance, Cistus salviifolius is a common pioneer seeder species in Mediterranean ecosystems. A high percentage of C. salviifolius seeds are found in physical dormancy, which can be broken by heat (Moreira and Pausas 2012). Due to this high dependency on heat, exposure to fire is often important for its population persistence in Mediterranean regions (Eshel et al. 2000, Izhaki et al. 2000).

Seasonal changes in plant community composition in fire-prone ecosystems suggest that fire season may be critical for predicting the post-fire seedling recruitment (Knox and Clarke 2006). The effect of fire on vegetation recovery can be a direct result of fire intensity and/or timing relative to the plants' phenological stage (Knapp 2010). For example, summer fires occurring in mixed conifer forests inflicted stronger negative effects on plant species richness and densities compared to spring fires (Knapp et al. 2006, Marriner 2015). This differential fire season effect was attributed to lower fuel loads during the spring season resulting in lower soil heating (Knapp et al. 2006,

Marriner 2015). Furthermore, differential fire season effects can be caused by a temporal shift in the soil seed bank size and composition (Lavorel et al. 1993, Cespedes et al. 2012). In Mediterranean woodlands, both seed density and species richness are higher during the spring, right after seed dispersal (Onaindia and Amezaga 2000). Such variation is mainly the outcome of seed dispersal and predation with the possible addition of some seed germination in late autumn. All of these can vary among seeder species characterized by different life forms (Walck et al. 2005, Marriner 2015).

Considering only fire intensity, owing to lower fuel moisture content autumn fires are expected to be more intense than spring fires, allowing higher seedling recruitment of seeder species and lower recruitment of fire-killed species. However, most Mediterranean woodland species flower during the spring. Therefore, spring fires can strongly influence the seed bank by burning plant canopies while they are still carrying pre-dispersed and possibly immature seeds. This phenological effect can lead to greater reductions in germination density and species diversity of soil seed bank after spring than after autumn fires. Moreover, hard-coated seeds of shrubs and dwarf shrubs, whose physical dormancy is broken by spring fires, should survive the long, dry summer unprotected to successfully germinate in the rainy season (autumn and winter).

The spatial distribution of seeds may have important ecological consequences for vegetation dynamics in fire-prone regions (Ne'eman and Izhaki 1999, Gomez- Aparicio et al. 2005). Generally, Mediterranean woodlands are heterogeneous landscapes, consisting of a mosaic of trees, shrubs, herbaceous vegetation, and canopy gaps. Such heterogeneity also translates into spatial variation in fuel load and composition, creating microhabitats characterized by varying levels of flammability (Loepfe et al. 2010). One relevant example is the occurrence of canopy gaps between shrubs or trees with lower fuel load, consequently lowering fire intensity (Carrington 2010). These open gaps are safer for seeds during fire events, contributing to the maintenance of plant species’ abundance and richness. Other examples are the patches of herbaceous vegetation that may experience higher fire intensities in late spring or early summer, having reached their biomass peak during this season (Sternberg et al. 2000), while shrub patches are likely to be more flammable during autumn burns due to lower fuel moisture content. Clearly, these differences in moisture content are highly dependent on the annual rainfall regime.

Previous studies in the Mediterranean basin exploring the effects of fires on plant community regeneration were mostly based on natural wildfires (Ne'eman et al. 1992, Lloret 1998, Izhaki et al. 2000, Van Leeuwen et al. 2010, Santana et al. 2014, Tessler et al. 2015). Recently, a study involving prescribed fires investigated the effect of fire season on seed germination and vegetation recruitment in the West-Mediterranean basin (Cespedes et al. 2012, Cespedes et al. 2014). This study illustrated that fire season can differentially affect plant regeneration, mainly by changing the relative germination density of dominant seeder species. Remarkably, mistral winds and drought events are more common in the East-Mediterranean region (Saaroni et al. 1998). Furthermore, the East-Mediterranean basin is highly rich with different phytogeographical plant types and includes many plant species that solely regenerate from the soil seed bank (Hegazy and Lovett-Doust 2016). Therefore, it is highly important to understand how fire events affect the germination from the soil seed bank, and if these effects are influenced by the interaction with other environmental factors. The present study is the first large-scale experimental attempt to examine how fire season and microhabitat interact to influence the composition of the germinable soil seed bank community in the East-Mediterranean basin.

I hypothesized that fires should inflict a stronger positive effect on the germinable soil seed bank density, richness and diversity of species whose germination is stimulated by fire (e.g. mainly dwarf shrub species), and a higher negative effect on species with fire-sensitive seeds (e.g. mainly annual species), thus resulting in changes in the composition of the germinable soil seed bank community. Specifically, as a result of higher moisture content, spring fires should result in lower burn temperatures than autumn fires, allowing more species to survive and germinate in the post-fire environment. In other words, considering only fire intensity, autumn fires should inflict a stronger negative effect on germinable seed abundance, richness and diversity of fire- sensitive plant species. Notably, most species occupying Mediterranean woodlands flower during the spring. Therefore, spring fire can reduce the germinable soil seed bank density also by damaging the developing seeds on the plants. Hence, during the spring an additional reduction in the germinable seed abundance should be evident. Clearly, this phenological effect can be strong to the extent that spring fires would cause in greater reductions in germinable seed abundance, richness and diversity, compared to autumn fires. I also posited that this differential fire season effect would be more

13 pronounced under shrubs and weakest in the open microhabitats. Clearly, this pattern should be more pronounced after the presumably more intense autumn fires.

1.2. Chapter 2: The effects of smoke and fire history on the soil seed bank germination in Mediterranean woodlands

(This part of my doctoral research has been published in Journal of Plant Ecology, Manela, N., Dagon, E., Semesh, H., & Ovadia, O (2018).

Fire plays a major role in determining plant community dynamics in many terrestrial biomes, consequently affecting ecosystem functioning (Kruger 1984, Keeley and Fotheringham 2000, Pausas and Keeley 2009). The effect of fire on plant regeneration is more common in fire-prone areas, where selection for species that are stimulated by fires occurs (Keeley 1995, Zedler 1995). These species are often characterized by large soil seed banks (Enright et al. 1997, Ne'eman and Izhaki 1999), which largely influence the post-fire succession processes.

The complex interaction between fire disturbance and plant community dynamics is dictated by the interplay between environmental conditions, plant community structure and fire characteristics (Kruger 1983, Ne'eman and Izhaki 1999, Kruger 2014). Attributes of fire characteristics can have important consequences on plant community composition. For instance, previous studies illustrated that frequent fires can increase species evenness at the community level (Trabaud and Galtié 1996, Tessler et al. 2016a). Other studies showed that higher fire frequency could affect community composition by increasing the abundance of herbaceous plants characterized by a short life-span, while reducing the abundance of trees (Lloret and Vilà 2003, Delitti et al. 2005). A main concern was referred to fires occurring at high frequency, preventing the complete regeneration of long lived-species, and threatening their persistence (Delitti et al. 2005, Malak et al. 2015).

The long history of disturbances, including fire and grazing, is an inherent characteristic of the East-Mediterranean ecosystem (Cowling et al. 1996). Such disturbances most likely affect the dynamics of both the standing vegetation and the germinable soil seed bank (Bond and Keeley 2005, Verdú et al. 2007). Among all, fire is the most frequent large-scale disturbance in Mediterranean ecosystems (Ne'eman et al. 2012, Kruger 2014). Moreover, fire frequency is expected to increase in the near future (Moriondo et al. 2006). Due to the pivotal importance of this disturbance type in 14

Mediterranean systems, it is assumed that life history traits of plant species occupying this ecosystem have evolved in response to selective forces associated with fires. These traits allow plants to survive burn injury, or to successfully regenerate in the post-fire environment (Pausas et al. 2006). Plant species in which germination is stimulated by fire can be divided into two major classes: 1) species characterized by hard coated seeds in which physical dormancy is broken by fire heat (Moreira and Pausas 2012), and 2) species having dormant seeds that germinate in response to fire cues, such as smoke (Brown and Van Staden 1997). Smoke stimulated germination has been documented in several Mediterranean ecosystems including California, South Africa and Australia (Brown 1993, Dixon et al. 1995, Pierce et al. 1995, Keeley and Fotheringham 1998).

Although it has been largely assumed that the effect of smoke on seed germination is less common in the East-Mediterranean basin (Ne'eman et al. 2012), recent studies have illustrated that smoke may be an important germination cue in the Mediterranean basin (Çatav et al. 2012, Tormo et al. 2014, Chamorro et al. 2017, Moreira and Pausas 2018). For instance, a laboratory study indicated that liquid smoke (i.e., smoke-water solution) could stimulate the germination of two common dwarf-shrub species in the Mediterranean basin (i.e., Sarcopoterium spinosum and Satureja thymbra) (Çatav et al. 2012). A more recent field study, conducted in the western Mediterranean basin, have shown that exposure to liquid smoke leads to an increase in the abundance of annual species (Tormo et al. 2014). In addition, the effect of smoke on seed germination can vary within an ecosystem (Moreira and Pausas 2012), suggesting that the response of seeds to smoke should be dependent on environmental characteristics, geographic region and plant life history traits (Baldwin and Morse 1994, Brown et al. 2003). Remarkably, the East-Mediterranean basin represents the dry edge of the Mediterranean ecosystem, with unique environmental conditions and plant community composition (Hegazy and Lovett-Doust 2016). This region is quite rich in terms of different phytogeographical plant types, including a high proportion of annual species, which solely regenerate from the soil seed bank (Hegazy and Lovett-Doust 2016). Moreover, a main difference between the East-Mediterranean basin and other Mediterranean type regions is that this ecosystem has been subjected to a long history of land use, and that fires in this area are mostly anthropogenic (Keeley et al. 2011a). Hence, a key step in predicting plant community dynamics in this unique ecosystem is quantifying how germination potential is affected by fire cues and by the local fire history.

The aim of the present study was to examine the interactive effect of aerosol smoke and fire history on the composition of the germinable soil seed bank community in East- Mediterranean woodlands. In contrast to the traditional approach, I used intact soil samples containing the natural seed bank community, rather than exposing seeds of selected species to liquid smoke in vitro. Specifically, soil samples collected from sites subjected to different fire frequencies (during the last four decades) were treated with aerosol smoke with or without watering. They were then placed in a greenhouse for the eight-month period, during which the number and identity of germinating seeds were documented.

I hypothesized that exposure to aerosol smoke (without fire heat) should particularly stimulate the germination of plant species having no physical dormancy, and that this pattern would be more pronounced in areas subjected to frequent fires. Since it is assumed that the chemical trigger of smoke is water soluble transferred to the seeds in smoke aerosol, aqueous leachate, or smoke vapors (Brown and Van Staden 1997, Keeley and Fotheringham 1998), watering the soil samples, prior to smoking, should enhance the positive effect of smoke on seed germination. From an ecological point of view, this scenario is less realistic because fires commonly occur under dry conditions.

1.3. Chapter 3: Long-term effect of a pulse perturbation experiment involving canopy opening on an endangered population of Paeonia mascula in an East- Mediterranean forest

Disturbances are important drivers of plant biodiversity (Thom and Seidl 2016, Conell, 1978), selecting for plant species that can quickly regenerate in the post- disturbance environment (Sousa 1984). For instance, Mediterranean ecosystems have long been subjected to fires; consequently, they comprise a large number of fire-adapted plant species which regenerate after stimulation by fire or on fire cues (Keeley et al. 2011a). Such selective forces driven by disturbances increase the ecosystem resilience (Johnstone et al. 2016).

A shift in the disturbance regime that manifests as changes in the frequency, intensity or timing of the disturbance can move the ecosystem into an alternative stable- state (Beisner et al. 2003, Johnstone et al. 2016). Hence, species that cannot tolerate the new environment will most likely go extinct. For example, owing to cultivation

16 practices and tree cutting, combined with high grazing intensity, Mediterranean areas in Israel (e.g. Mount Meron nature reserve) have long comprised a mosaic of a few closed woodland fragments and many open areas (Ne'eman 2003). However, in the past 70 years, grazing and tree harvesting have been minimized and vegetation formation in many areas has turned into closed and dense woody vegetation (Carmel and Flather 2004). This ecosystem shift facilitated an increased abundance of woody and shade- tolerant plants while threatening the existence and reproduction of those that could not tolerate the shading under the tree canopy (Perevolotsky 2005, Agra and Ne’eman 2009).

Preventing a shift in an ecosystem state or protecting the species that are more likely to disappear owing to such reorganization in ecosystem structure, requires management actions (Folke et al. 2004). Indeed, management actions are frequently performed to improve environmental conditions for a certain species group and to avoid their local extinction (Kahmen et al. 2002, Bakker 2005, Klimek et al. 2007).

Many studies have shown that the manipulation of different environmental factors, such as increasing light availability, reducing competition and manipulating plant- animal interaction, could have a strong positive impact on population densities of rare or endangered plant species (Kearns et al. 1998, Andrieu et al. 2013, Matsushita et al. 2016). Nevertheless, management actions that seem to positively affect population persistence of the targeted species groups over short time scales, the long-term effects are often unknown.

The aim of the present study was to examine the long-term effect of a pulse (past) perturbation management action, involving one time canopy opening on the population density and reproduction of the locally endangered southernmost population (Shmida and Pollak 2007) of Paeonia mascula (Paeoniaceae) (Figure 1) inhabiting a small area in Mount Meron, Israel. Paeonia mascula is a perennial rhizomatous herb, with thick, curved tubers, extending deeply into the ground down to 100 cm. The flower color ranges from bright pink to purple. Each flower has 5-8 large, dense petals. Beetles pollinate the flowers, but some of the flowers in nature can be sterile. The fruit has 3-5 hairy white follicles, which are red in their inner side when they open. Fertile seeds are dark blue to black and sterile seeds are red. This color combination apparently attracts birds that eat and disperse the seeds. In Israel, P. mascula reproduces mostly asexually,

17 particularly in closed woodlands where most of the plants do not bloom. In well-lit areas, a higher percentage of plants flower and produce fruit. Goats sometimes eat the flowers and leaves of P. mascula (Shmida and Pollak 2007).

A previous study carried out ~20 years ago illustrated that flowering P. mascula plants can be found in small canopy gaps, where they are exposed to higher levels of active radiation (Ne'eman 2003). Specifically, Ne'eman (2003) examined the effect of canopy opening and light increment on P. mascula flowering. In his study, Ne'eman marked plots with non-flowering P. mascula plants found under closed canopies. Five of the plots were cleared from trees in 1997, and five plots remained closed. The results of that study indicated that increasing the radiation to ~70% by tree cutting positively affected the flowering percentage. No long-term monitoring was done to estimate how this pulse perturbation management action affected the P. mascula population over a longer time scale.

I hypothesized that the dense germination of fast-growing and light- dependent plant species (i.e., annual herbs), which was evident soon after the clearing (Ne'eman, 2003), should have led to a recharge of the soil seed bank with seeds of annual herbs. Although these species can compete with P. mascula over resources, they are expected to be more sensitive to shading within the cleared plots. As for P. mascula, since the canopy clearing resulted in a short-term increase in the percentage of flowering P. mascula plants (Ne'eman, 2003), I posited that there should be an increase in the number of P. mascula seeds in the soil seed bank, and consequently a larger number of P. mascula plants in plots cleared 20 years ago.

(a) (b)

Figure 1 Paeonia mascula (a) flower and (b) seedling at the study site in Ein Hazaken,

Mount Meron, Israel.

2. Methods and materials

2.1. General methods

Owing to the complex dynamics of the soil seed bank and the difficulty of distinguishing between the factors responsible for the input and output of seeds, dictating seed bank dynamics, I focused on characterizing seed germination dynamics. To assess the regeneration potential from the soil seed bank, most of the experiments conducted throughout my research included monitoring and documenting seedling emergence from soil samples collected from the different study sites. This methodological approach allowed me to better examine the germination potential from the soil seed banks included in my studies.

2.1.1. Soil sampling and seedling emergence monitoring

To assess the germination potential of the soil seed bank, soil samples were collected during the autumn season, before the start of the rainy season, when germination in the fields naturally occurs. All soil samples were collected from the top five centimeters of the soil, where most seeds are commonly present (Traba et al. 2004). Soil samples were sieved (2 mm mesh) and homogenized and spread over germination trays on a 3 cm layer of vermiculite to improve moisture retention and aeration in soil.

Next, germination trays were randomly placed within the greenhouse and were irrigated on a daily basis by an automatic, overhead, sprinkler system. To monitor airborne seed contamination, trays containing vermiculite mixed with potting soil were randomly placed within the greenhouse.

To monitor seed germination, every newly emerged seedling was documented and removed from its tray to reduce seedling density and allow new seeds to germinate. To achieve a high identification level of seedling species, every morphological type of emerged seedling was tagged with a unique identity. Then, a subset of seedlings from each morphological type were grown in separate pots until they could be identified to the species level or to the closest taxa level. Monitoring seedling emergence from germination trays lasted between six to eight months, until no new seedling emerged for more than two weeks.

19 The next few sections include specific methods and materials for each of my main research topics:

2.2. Experiment 1: Fire season and microhabitat specific methods

2.2.1. Study site

The study site was located in Mount Ya'aran (31° 42'N, 35° 1'E) on the Judean Mountains, Israel (600 m ASL, Figure 2). The climate in this region is typical east Mediterranean, with hot and dry summers lasting at least from June to September and mild- wet winters. The average annual rainfall is 540 mm and the average daily temperature is 19°C. The annual minimum temperature (5.4 °C on average) occurs in January, and the maximum temperature (34 °C on average) occurs in August. Local soils are red Terra-Rossa developed on a hard Cenomanian dolomite bedrock of 'Weradim' formation (Singer, 2007). Vegetation formation in the study site is of a typical eastern-Mediterranean woodland (maquis) with scattered low trees (e.g. Quercus calliprinos), shrubs (e.g. Pistacia lentiscus, Rhamnus lycioides and Calicotome villosa), dwarf shrubs (e.g., C. salviifolius, Cistus creticus and Teucrium divaricatum), patches of herbaceous vegetation, and exposed rock. The vegetation is 0.5 m high on average, with a mean volume of 0.02 m3 per 1 m2. Vegetation phenology is generally similar among dominant plant species germination following first autumn rain and flowering during the spring season. The study area has been subjected to heavy grazing pressure for years; however, during the last decade this pressure has been minimized to allow natural recovery of the vegetation.

Analyzing the fire history of the Judean Mountains revealed that more than 3,800 fires occurred between the years 1987 and 2009, i.e., an average rate of 170 fires per year (Tessler et al. 2010, Tessler 2012). More recent KKL-JNF (Jewish National Fund) fire records showed that one year prior to our prescribed fires (2013), there were at least

48 fire events in this geographical region, burning an area of 8,550 m2 in total. Over half (54.2%) of the fires occurred during summer, 8.3% during autumn, 4.2% during winter, and 33.3% during spring (Tsafrir A. and Ovadia O. unpublished data).

2.2.2. Experimental design

The experiment included 12 50×30 m plots that were randomly assigned to one of the following three fire treatments (four plots per treatment): spring burning (due to

20 exceptionally late rains, spring burning was conducted on the 1st June 2014), autumn burning (September 2014), and unburned control plots (Figure 2). Eight sampling sub- plots of 5×5 m were randomly situated within each of the 12 plots (a total of 96 sampling sub-plots).

2.2.3 Soil sampling and seedbank assessment

Soil samples were collected in September 2014, after the fires and before the rainy season. In each of the eight sub-plots, soil samples of 10×10 cm and 5 cm depth were collected in three dominant microhabitats: under P. lentiscus shrubs, assumed to be flammable due to relatively high fuel load with high phenolic content (Benhammou et al. 2008); under C. salviifolius, also considered flammable since it has a lower fuel load and foliage heat content (Dimitrakopoulos 2001); and open canopy gaps (no woody vegetation cover), representing the lowest fuel load microhabitat. Seedling emergence monitoring was conducted with the 288 soil samples in total (3 fire treatments × 4 experimental plots × 8 sub-plots × 3 microhabitats).

Figure 2. The study site in Mount Ya'aran, Israel. Each of the 12 plots (50×30 m) was randomly assigned to one of the following fire treatment groups: spring burning (black), autumn burning (grey) and unburned control (plaid pattern). Each plot consists of eight 5×5 m subplots (white).

21 2.2.4. Statistical analyses

To analyze the effect of fire treatment, microhabitat type and their interaction on the germinable soil seed bank density, richness and diversity per 1 m2 of soil area, as well as on the annual and dwarf shrub germinable soil seed bank density and richness per 1 m2 of soil area, generalized linear mixed model (GLMM) followed by Bonferroni corrected pairwise comparisons was performed using Stata version 14.2 (StataCorp LP). Due to overdispersion of the data, negative binomial distribution and log link function were used to analyze the germinable soil seed bank density. Species richness (count data) was analyzed using a Poisson distribution and a log link function. Diversity of germinable seeds was calculated using Fisher's alpha index and data was analyzed using the identity link function. Generalized linear mixed model included density, richness or diversity as response variables, and fire season and microhabitats as fixed explanatory variables. To account for spatial dependency, I included both the sampling plots and subplots as random factors, with the latter being nested within the former.

Multivariate analyses were performed using PRIMER v.6 (Clarke and Warwick 1994, Clarke and Gorley 2006). The relative species densities were fourth- root transformed (Clarke and Warwick, 1994). To illustrate the main axes discriminating between treatments, a permutational MANOVA (PERMANOVA) (Anderson and Walsh, 2013) followed by non-metric multidimensional scaling (nMDS) was performed, based on Bray–Curtis dissimilarity matrix (Clarke 1993, Clarke and Warwick 1994, Anderson et al. 2008).

Similarity percentages (SIMPER) analysis was used to quantify the contribution of each plant species to the overall dissimilarity in community composition between different treatment combinations (Clarke, 1993). Finally, a canonical analysis of principal coordinates (CAP) (Anderson et al., 2008) was used to find the axes that best discriminate between a priori defined groups (i.e., unburned control, spring fire and autumn fire) and to correlate (>0.25 correlation using Preason correlation coefficient) between species (vectors) and groups (i.e., control, spring fire and autumn fire).

Venn diagram was performed in R 3.2.1 using the VennDiagram package (Chen and Boutros 2011). This analysis was used to show the number of shared and unique species between different treatment combinations.

22 2.3. Experiment 2: Fire frequency and smoke specific methods

2.3.1. Study site

The study site was located in Mount Carmel (32°45’E, 32°00’N), a coastal mountain range of approximately 23000 ha in size, at the northern region of Israel (500 m ASL, Figure 3). The climate in this region is typical East-Mediterranean, with hot and dry summers lasting from June to September, and mild wet winters. The average annual precipitation ranges from 550 mm to 750 mm, with yearly minimum temperature (13 °C on average) occurring in January, and maximum temperature (28 °C on average) occurring in August (Wittenberg et al. 2007). Local soils are brown or gray Rendzina formed on compact Upper Cenomanian chalk of the Shamir formation (Lavee et al. 1995). The vegetation in the study site is typical eastern-Mediterranean woodland, composed of evergreen trees (e.g. Quercus calliprinos) and shrubs (e.g. Pistacia palaestina, P. lentiscus, C. salviifolius, Cistus creticus and Calicotome villosa), and high diversity of annual species. Previous studies have illustrated that vegetation cover at Mount Carmel is higher in unburned areas compared to areas subjected to recurrent fires, especially those burned in 2010 (Tessler et al. 2016a). More than 30% of the unburned areas in Mount Carmel are covered with trees, while in burned areas trees comprise less than 20% of the vegetation. Moreover, the relative coverage of exposed rocks, herbaceous plants, shrubs, and dwarf shrubs are all higher in recently burned areas (Tessler et al. 2016a).

2.3.2. Fire history

Fire records in Mount Carmel documented more than 620 forest fires between 1974 and 2010, with annual average of 30 fires (Tessler 2012). Large wildland fires (larger than 120 hectare) are predicted to occur in Mount Carmel every six years, while small fires (smaller than 10 hectare) occur every year (Tessler et al. 2014). The largest documented fire event occurred in December 2010, consumed 2500 ha (Tessler 2012).

23 Figure 3. The study area in Mount Carmel, Israel, comprising nine sites that have been subjected to one fire (grey), three fires (black), or no fire (white) during the last four decades.

2.3.3. Soil sampling and experimental design

The experiment included nine sites representing three different fire frequencies records: 1. sites that have not been burned during the last four decades (i.e., unburned), 2. Sites burned on 2010 (i.e., one fire), and 3. sites burned three times during the last four decades (i.e., three fires), on 1983, 1989/1978 and 2010. The area of each sampling site was approximately 100 m2, and they were distributed over an area of ~3000 m2.

In each site, 20 soil samples of 20×10 cm and 5 cm depth were collected. Each soil sample was divided into three sub-samples that were randomly assigned to one of following smoke treatments: 1. no treatment (control), 2. smoke treatment (smoke), and 3. watering followed by smoke treatment (water & smoke). Before treatments were applied, sub-samples were spread over germination trays covered by a 3 cm layer of vermiculite. Prior to the smoke treatment, sub-samples that were assigned to watering followed by smoke treatment were watered manually.

24 Immediately after the smoke treatment, germination trays were randomly placed in a greenhouse, under natural light and temperature conditions, and with daily irrigation by an automatic overhead sprinkler system to maintain soil moisture. To monitor airborne seed contamination, trays containing vermiculite mixed with potting soil were randomly placed within the greenhouse. Seedling emergence was monitored every three days to decrease seedling density. Seedlings were identified to the species level, or to closest taxa level, and were also classified by their life form (Du Rietz 1931). Unidentifiable seedlings were transplanted to pots until flowering. The seedling emergence stage lasted eight months, until no new seedling had emerged for two weeks. Although most of the seedlings (>60%) emerged in the first two months, the experiment lasted longer time (six months) to allow the emergence of late germinating species.

2.3.4. Aerosol smoke treatment

The smoke treatment took place in a smoke generator system (Tieu et al. 1999). The smoke was generated in a metal drum by combustion of a mixture of dried plant material, including branches and foliage of native species including Quercus, Pistacia, and Pinus species, simulated the natural smokes produced by wildland fire. Air inlet fan pushed the smoke through a metal pipe into a sealed tent where germination trays (soil samples) were placed. The metal pipe was cooled with water during the smoke treatment. The temperature inside the sealed tent increased by maximum of 2 °C. Soil samples were exposed to the cooled aerosol for two hours (Landis 2000). Then, the trays were moved to a greenhouse where seedling emergence took place.

2.3.5. Statistical analyses

To test for the effects of aerosol smoke and fire history on germinable soil seed bank density (per 1 m2 of soil area), species richness and annual density (per 1 m2 of soil area) and richness, I used GLMMs, followed by Bonferroni corrected pairwise comparisons. The GLMMs included density, richness or diversity as response variables, and fire frequency and smoke treatment as fixed explanatory variables. To account for spatial dependency, I included both the sampling plots and the samples as random factors, with the latter being nested within the former. Due to overdispersion of the data, a negative binomial distribution and a log link function were used to analyze the germinable soil seed bank density A Poisson distribution with a log link function were used when analyzing species richness (count data). These analyses were performed

25 using Stata version 14.2 (StataCorp LLC, Texas).

Permutational multivariate analysis of variance (PERMANOVA) (Anderson and Walsh 2013), based on Bray–Curtis similarity matrix was used to test for the combined effect of fire frequency and smoke treatment on GSSB community composition. Then, a canonical analysis of principal coordinates (CAP) (Anderson et al. 2008), was used to find the axes best discriminating between different levels of the fire frequency factor (i.e. unburned, one fire and three fires).

2.4. Experiment 3: Tree clearing (Paeonia) specific methods

2.4.1. Study site

The study area was located at Mount Meron nature reserve (35°25’E, 32°58’N), northern Israel. Since it was declared as a nature reserve in 1965, most of the woodland canopy in this area has gradually closed to form a dense woody canopy, mainly due to a reduction in grazing and deforestation (Carmel and Kadmon 1999, Ne'eman 2003, Carmel and Flather 2004). The area is characterized by an east Mediterranean climate with hot and dry summers and cool winters with average annual rainfall of ~900 mm (Agra et al, 2015). Mount Meron is the East-Mediterranean southern edge of several plant species such as Juniperus oxycedrus, Eriolobus trilobatus and Paeonia Mascula. It is also rich with evergreen species dominated mainly by oak maquis (Quercus calliprinos–Pistacia palestina association) (Ne’eman, 2003). The vegetation cover is composed of trees (Arbutus andrachne, Q. calliprinos, Quercus boissieri, and P. palaestina), shrubs (Spartium junceum, Cistus salviifolius, Cistus creticus, and Sarcopoterium spinosum), climbers (Asparagus aphyllus, Smilax aspera) and herbaceous plants mainly occupying open patches.

2.4.2. Experimental design

The study included ten plots that were marked by Ne’eman (2003) in 1997 as part of a study whose aim was to examine the effect of canopy opening on P. mascula flowering. These 5×5 m plots comprised five untreated plots located under closed canopy (hereafter, ‘closed’), and five plots that were cleared from trees by woodcutting in November 1997 (hereafter, ‘cleared’). In addition, we have marked five new plots located in open gaps adjacent the closed woodland. These new plots enabled the characterization of the germinable soil seed bank community surrounding the closed and cleared plots (hereafter open).

26 2.4.3. Paeonia mascula population measurements

In each of the plots, the number of plants and flowers per plant were counted during April 2017 and 2018 at the peak of the flowering season. P. mascula plants were categorized by their maturation stage (seedling/mature plant). P. mascula soil seed bank density was estimated in September 2017 by sieving ten random 20×20×5 cm soil samples in each of the closed and cleared plots and counting peony seeds.

2.4.4. Soil sampling and seedling emergence of neighboring cooccurring species

In each plot, nine random soil samples 10×10 cm were collected from the top five centimeters of the soil. All samples were sieved to remove stones and were spread within germination trays (25×10×10 cm) covered by a thin layer of vermiculite. Then, trays were placed in a greenhouse for seedling emergence monitoring.

2.4.5. Canopy closure

LAI (leaf area index) was used to measure plant canopy cover in each plot using SunScan Canopy Analysis System (SunSCAN, Delta‐T Devices Ltd, Cambridge, UK). Measurements were taken during the autumn season (September 2018). In each plot, LAI measurements were taken at 16 points positioned one meter apart.

2.1.6 Statistical analyses

To analyze the long-term effect of pulse canopy opening on P. mascula population (i.e. plant number, flowering percentage and soil seed bank), the density (per 1 m2 of soil area) and species richness of the germinable soil seed bank of neighboring plant species, LAI, and proportion of flowers per plot, a generalized linear mixed model (GLMM) was performed using Stata version 14.2 (StataCorp LP).

Negative binomial distribution with a log link function was used to analyze the density (per 1 m2 of soil area) of the germinable soil seed bank of neighboring plant species. A Poisson distribution with a log link function was used to analyze the species richness of the germinable soil seed bank community. Flowering percentage was distributed binomially and was thus analyzed using logit link distribution, and LAI was analyzed using an identity link function.

GLMMs included density and richness of the germinable soil seed bank of neighboring plant species, LAI, and flowering percentage as fixed explanatory

27 variables, while both the sampling plots and the samples were included as random factors, with the latter being nested within the former. Sampling year was added as a crossed random factor.

28 3. Results

3.1. Experiment 1: Fire season and microhabitat results

3.1.1. Fire intensity and severity were consistent between the two burning seasons

Fire properties were quantified during the prescribed burnings (Tsafrir et al. 2018). No significant differences in flame height (i.e., proxy of fire intensity; spring: 1.81±0.16 m; autumn: 2.22±0.18 m; mean±1 SE; F1,58=3.73, P=0.058) were detected between spring and autumn burnings. Soil temperature (i.e., proxy of fire intensity) was also measured during the burnings. Eight thermocouples were buried in the soil within two sampling sub-plots in two of the twelve plots (four thermocouples per plot). Soil temperature was measured 10 cm below the top ground layer. There were no significant differences in average soil temperature between the two burning seasons (spring; 90.94

±13.3 °C; autumn; 102 ± 21.03 °C; mean±1SE; F1,42 = 0.298, P = 0.587). Finally, the proportion of burned area (i.e., proxy of fire severity) did not differ significantly between the two burning seasons (spring: 45.98±4.16%; autumn: 37.1±4%; mean±1 SE;

F1,58=1.87, P =0.182).

3.1.2. The effect of fire treatment and microhabitat type on the germinable soil seed bank density, richness and diversity

A total of 5601 seedlings from 145 species of vascular plants emerged from the soil samples. The germinable soil seed bank density varied considerably between treatments, with 1501±232, 1392±210 and 2455±367 (mean±1SE) seedlings per 1 m2 for spring fire, autumn fire, and unburned control plots, respectively (Figure 4a). Generally, compared to the unburned control, the germinable soil seed bank density was significantly affected by the spring fire, while its diversity was significantly affected by the autumn fire (Table 1). Specifically, the germinable soil seed bank density, richness and diversity were significantly higher in unburned samples than in samples subjected to autumn and spring burning, respectively (Figure 4, see Supporting Information Table S1, S2 & S3). No differences in germination density, richness and diversity were observed between microhabitats in the unburned samples and in samples subjected to spring fire (Figure 4, see Supporting Information Table S1, S2 & S3). Notably, samples subjected to autumn fire showed a significant lower germination density and richness under shrubs (i.e., Pistacia and Cistus) compared to open microhabitat (Figure 4a, b and c; see Supporting Information Table S1 & S2). 29 Table 1: Summary of the generalized linear mixed models testing for the effect of fire treatment and microhabitat type on the germinable soil seed bank density, richness and diversity (Fisher’s alpha).

Density Species richness Diversity

P P z P Estimates SE z value Estimates SE z value Estimates SE value value value value Treatment Autumn fire -0.299 0.293 -1.020 0.308 -0.254 0.207 -1.230 0.219 -2.313 1.065 -2.170 0.030 Spring fire -0.686 0.299 -2.300 0.022 -0.290 0.209 -1.390 0.166 -0.974 1.088 -0.900 0.371 Microhabitat Pistacia -0.409 0.238 -1.720 0.085 -0.040 0.089 -0.450 0.655 0.579 0.945 0.610 0.540 Cistus -0.156 0.236 -0.660 0.508 -0.032 0.089 -0.360 0.721 -0.998 0.945 -1.060 0.291 Treatment × Microhabitat Autumn fire × Pistacia -0.595 0.343 -1.740 0.083 -0.924 0.167 -5.540 <0.001 -2.719 1.349 -2.020 0.044 Autumn fire × Cistus -0.125 0.337 -0.370 0.709 -0.304 0.143 -2.130 0.034 0.483 1.337 0.360 0.718 Spring fire × Pistacia 0.234 0.345 0.680 0.499 -0.186 0.144 -1.290 0.197 -2.187 1.365 -1.600 0.109 Spring fire × Cistus 0.338 0.342 0.990 0.323 -0.069 0.141 -0.490 0.623 -0.518 1.365 -0.380 0.705 Constant term 3.405 0.208 16.360 <0.001 1.989 0.144 13.800 <0.00 5.929 0.753 7.870 <0.001 1 Log of the dispersion parameter/ln(alpha) -0.233 0.107 -2.180 0.029 Plot Variance 0.053 0.042 0.053 0.031 0.415 0.468 Subplot(Plot) Variance 0.060 0.068 0.112 0.027 0.529 0.964

30 Figure 4. Density (a), richness (b) and diversity (measured by Fisher’s alpha index) (c) of the germinable soil seed bank community (mean± SE) characterizing each of the three fire treatment groups and microhabitat type. Different letters above the bars indicate significant differences as indicated by Bonferroni post-hoc test (n=32, N=288).

31 3.1.3. Life forms

The germinable soil seed bank communities were composed mainly of annuals and dwarf shrubs, which together comprised more than 80% of the species and more than 90% of the germinable soil seed bank density in all treatment groups. The major difference between the fire treatments was the representation of species life forms in the germinable soil seed bank community (see Supporting Information Table S4, S5, S6 S7 & S8). Specifically, density of annual herbaceous was lower and dwarf shrub density was higher in samples subjected to fires compared to unburned plots. (Figure 5a and b, see Supporting Information Table S4, S6).

Figure 5. Annual herbaceous density (a) and richness (b), and dwarf-shrub density (c) and richness (d) (mean ± SE) in each of the three fire treatment groups. Different letters above the bars indicate significant differences as indicated by Bonferroni post-hoc test (n=32, N=288).

32 Similarly, annual richness of the germinable soil seed bank was lower in the burned than unburned plots (Figure 5c, see Supporting Information Table S5). This trend in annual richness was significantly stronger in samples collected under Pistacia and Cistus shrubs from plots subjected to autumn burnings (Figure 5c). Differently from annuals, the germinable soil seed bank density of dwarf shrub was significantly higher in samples subjected to autumn fire compared to unburned samples (Figure 5b), with no significant difference in species richness (Figure 5d). Of all Cistaceae species, F. arabica and C. salviifolius, were the main players contributing to the higher germinable soil seed bank density of dwarf shrubs in samples subjected to fires (See Supporting Information Table S9). As expected, no germination of Pistacia sp. was evident (this species is not present in the soil seed bank).

3.1.4. Community composition Non-metric multidimensional scaling indicated a clear separation in the germinable soil seed bank community composition between burned and unburned plots (see Supporting Information Figure S1). PERMANOVA based on the Bray- Curtis similarity matrix indicated that the community composition of the germinable soil seed bank was significantly affected by both fire treatment and microhabitat type, and their interaction (Table 2). A significant dissimilarity in the germinable soil seed bank community composition occurred between the unburned control and autumn fire treatment plots, as indicated by pairwise comparisons (see Supporting Information Table S10), with average dissimilarity of 70.05%. A SIMPER analysis also indicated that the species A. sterilis, Sedum rubens and Erophila minima generated 12.44% of this dissimilarity. A dissimilarity of 66.85% was observed between the unburned control and spring fire plots, with A. sterilis, E. minima and S. rubens having a total contribution of 11.8% to this dissimilarity.

Canonical analysis of principal coordinates (CAP) was used to examine patterns of community composition that can be explained by fire treatment (Figure 6). Permutation analysis showed that squared canonical correlations δ2 = 0.96 and δ2 = 0.79 were 1 2 significant (p<0.001 for both tests using 9999 permutations), and that the first canonical axis better separated between the three fire treatment groups, compared to the second axis. The unburned control was the most distinct group, with a 92% classification success rate in a cross validation, whereas the autumn and spring fire treatment groups were less distinct, with classification success rates of 67% and 58%, respectively.

33 Correlations of species with CAP axes are shown in Figure 6. Several correlations were noted between species and each of the three fire treatment groups. There were positive correlations (>0.25) of annual species, including E. minima, Minuartia hybrid (Caryophyllaceae family), S. rubens and A. sterilis, with the unburned control group. The abundant perennials F. arabica (dwarf shrub) and C. salviifolius (dwarf shrub, Cistaceae family) were positively correlated with the autumn fire treatment group, and Anagalis arvensis (annual herb) and S. spinosum (dwarf shrub) were positively correlated with the spring fire treatment group.

Table 2. Results of permutational multivariate analysis of variance (PERMANOVA) testing for the effect of fire treatment and microhabitat type on the composition of the germinable soil seed bank community.

Degree Source of SS MS Pseudo-F P(perm) perms freedom Fire treatment 2 12828 6413.8 2.3751 0.020 4744

Microhabitat 2 5747.7 2873.9 2.3706 <0.001 9878

Plot (Fire treatment) 9 24304 2700.5 2.2276 <0.001 9797 Fire treatment × 4 6464.4 1616.1 1.3331 0.030 9824 Microhabitat Res 18 21821 1212.3

Total 35 71165

Figure 6. Canonical analysis of principal coordinates (CAP) based on Bray-Curtis dissimilarity matrix, best discriminating samples among the unburned control (circles), autumn (squares) and spring burned (triangles) germinable soil seed bank communities, including correlations of individual plant species with the canonical axes.

3.1.5 Unique species

Out of 145 species that germinated, 28 species were unique to the unburned control plots, 14 species were unique to plots subjected to spring burnings, and 6 species were unique to plots subjected to autumn burnings (see Supporting Information Figure S2). In all treatment groups, most of the unique species were annuals germinating either from a single plot or a single sample. Of these unique species, the most abundant were the annual species: Torilis nodosa, Trifolium spp., and Alyssum simplex in the unburned control plots (n>10); and Heliotropium myosotoides, and Coronilla scorpioides in plots subjected to autumn and spring burnings, respectively.

35 3.2 Experiment 2: Fire frequency and smoke results

A total of 13,672 seedlings from 159 species emerged from the soil samples. The germinable soil seed bank density was significantly higher in sites that have been subjected to fires compared to unburned sites (Table 3, Figure 7a, see Supporting Information Table S11). Specifically, 2279 ± 166 seeds per m2 (mean±1 SE) germinated in the unburned sites, compared to 4128 ± 451 and 5405 ± 424 seeds per m2 (mean±1 SE) in sites that were burned once and three times during the last four decades, respectively (Figure 7a).

Smoke caused a significant increase in germinable soil seed bank density (Table 3; Figure 7a). Notably, this increase was higher in samples collected from sites that have been subjected to fires compared to unburned sites (Table 3; Figure 7a). In unburned sites, there was no significant difference in germinable soil seed bank density between samples that were watered prior to the smoke treatment and those that were directly exposed to smoke (Figure 7a; See Supporting Information Table S11). Nevertheless, in burned sites, germinable soil seed bank density was lower in samples that were watered prior to the smoke treatment, compared to those that were directly exposed to smoke (See Supporting Information Table S11). Germinable soil seed bank richness tended to be higher in samples collected from sites that have been subjected to three fires during the last four decades (Figure 7b, See Supporting Information Table S11). Furthermore, exposure to smoke tended to increase the germinable soil seed bank richness in these samples collected from frequently burned sites (P =0.073, Figure 7b; See Supporting Information Table S11). Notably, no differences in germinable soil seed bank diversity were detected between sites, or among smoke treatments, nor was there a significant interaction between these two factors (Table 3).

Annual species were the most abundant life form, comprising more than 90% of germinable soil seed bank density. germinable soil seed bank density of annuals was significantly higher in samples collected from sites that have been subjected to fires, compared to unburned sites (\Figure 7c, See Supporting Information Table S12, S13). Moreover, smoke caused a significant increase in germinable soil seed bank density of annuals (See Supporting Information Table S12, S13). This pattern was evident in all samples irrespective of fire. Similarly to the overall effect on the entire germinable soil seed bank community, watering prior to exposure to smoke, caused a significant decrease in the germinable soil seed bank density of annuals (See Supporting

36 Information Table S12, S13). germinable soil seed bank richness of annuals tended to increase by smoke (Figure 7d, See Supporting Information Table S12, S13). Notably, the density of the seven shrubs and dwarf-shrubs, which germinated from the soil seed bank, did not vary significantly among sites or between smoke treatments (See Supporting Information Table S12, S13).

Figure 7. Density and richness of the germinable soil seed bank community as a function of fire frequency and smoke treatment. The smoke treatment included: control (black), exposure to smoke (light grey), and watering followed by exposure to smoke (dark grey). Graphs represent the total germinable soil seed bank (a) density, and (b) richness, and the annual (c) density and (d) richness (mean± SE). Different letters indicate Bonferroni corrected significant differences (n=60, N=540).

Table 3. Summary of Generalized linear mixed models testing the effect of fire frequency and smoke treatment on germinable soil seed bank density, richness and diversity (calculated by Fisher's alpha).

Density Richness Diversity z z P Estimates SE value P value Estimates SE value value Estimates SE z value P value Fire frequency One Fire 0.757 0.296 2.56 0.011 0.261 0.192 1.36 0.175 0.063 0.806 0.080 0.938 Three Fires 1.004 0.295 3.40 0.001 0.348 0.192 1.81 0.070 0.275 0.806 0.340 0.733 Smoke Treatment Smoke 0.245 0.111 2.20 0.028 0.137 0.079 1.75 0.080 1.009 0.722 1.400 0.162 Water & Smoke 0.137 0.112 1.23 0.219 0.079 0.080 0.99 0.323 -0.066 0.722 -0.090 0.927 Fire frequency× Treatment One Fire × Smoke -0.013 0.152 -0.09 0.932 -0.060 0.106 -0.57 0.569 -1.516 1.020 -1.490 0.137 One Fire × Water & Smoke -0.499 0.155 -3.22 0.001 -0.257 0.110 -2.34 0.019 -0.407 1.020 -0.400 0.690 Three fires × Smoke 0.034 0.150 0.23 0.820 -0.014 0.103 -0.14 0.890 -1.070 1.020 -1.050 0.294 Three fires × Water & Smoke -0.355 0.151 -2.35 0.019 -0.181 0.107 -1.70 0.090 -0.880 1.020 -0.860 0.389 Constant term 2.165 0.211 10.27 <0.001 1.503 0.137 10.93 <0.00 4.075 0.570 7.150 <0.001 1 Log of the dispersion parameter/ln(alpha) -1.416 0.097 -14.56 <0.001 Plot Variance 0.076 0.056 0.037 0.023 0.074 0.218 <0.001 23.883 Sample(Plot) Variance 0.743 0.096 0.184 0.028 2.386 0.905 1.135 5.018

38 3.2.1. Community composition

PERMANOVA based on Bray-Curtis similarity matrix, indicated that germinable soil seed bank community composition was significantly affected by smoke (P<0.001, Table 4). Notably, the interaction between the effect of fire frequency and smoke was also significant (P<0.01; Table 4). Canonical analysis of principal coordinates (CAP), at the plot-level, was used to examine patterns of community composition that can be explained by fire frequency (Figure 8). Permutation analysis showed that squared 2 2 canonical correlations 훿1 = 0.9 and 훿2 = 0.845 were significant (p<0.001 for both tests using 99999 permutations). Fire frequency groups were distinct from each other with 100% classification success rate in a cross validation for unburned sites and for those burned three times, and with 88% classification success rate for sites burned only once during the last four decades. Notably, a clear separation between the smoke treatment groups was observed in sites burned once and in sites burned three times during the last four decades (Figure 8). Moreover, the cluster of samples collected from burned sites was positively correlated (>0.25) mainly with annual species. Specifically, Trifolium campestre (annual), Poa infirma (annual) and Cistus spp. (perennial dwarf shrubs), were positively correlated with samples originating from sites that were burned three times, while Anagallis arvensis (annual), Lotus spp., (annuals) and Bromus spp. (annuals) were positively correlated with samples collected from burned sites (i.e., one and three fires) that were exposed to smoke. (annuals) and Bromus spp. (annuals) were positively correlated with samples collected from burned sites (i.e., one and three fires) that were exposed to smoke.

Figure 8 Canonical analysis of principal coordinates (CAP) based on Bray-Curtis dissimilarity matrix, best discriminating samples among the different fire frequency by smoke treatment combination groups. Different colors represent sites that have been subjected to one fire (white), three fires (grey), or no fire (black) during the last four decades. Different shapes represent the smoke treatment (triangle), watering and smoke treatment (circle) and control no-smoke treatment (square).

3.2.2. Species level

Germinable soil seed bank density of several species was differentially affected by both fire history and smoke treatments. Out of all documented species, A. arvensis was the most abundant one in all sites, constituting more than 20% of the total germinable soil seed bank density (See Supporting Information Table S14). Other species including P. infirma, Chrysanthemum coronarium and Mercurialis annua were more abundant in samples originating from sites burned more frequently, and in samples treated with smoke. However, owing to relatively low abundances, the respective statistical power was low.

Table 4 Results of permutational multivariate analysis of variance (PERMANOVA), testing for the effect of fire frequency and smoke treatments on the composition of the germinable soil seed bank community.

Source df SS MS Pseudo-F P(perm) perms P(MC) Fire Frequency 2 50943.00 25471.000 1.169 0.294 280 0.259 Smoke treatment 2 9977.80 4988.900 2.716 <0.001 9871 <0.001 Plot(Fire frequency) 6 130730.00 21788.000 8.156 <0.001 9771 0.008

Fire Frequency × Smoke <0.001 treatment 4 10656.00 2664.100 1.450 0.007 9813 Sample (Plot) 171 7.53 4399.500 2.395 <0.001 9115 <0.001 Res 525 1402500.00 2671.500 Total 539 1604800.00

41 Experiment 3: Tree clearing (Paeonia) results

3.2.1 Paeonia mascula population

In total, 300 and 366 peony plants were recorded in the study site in 2017 and 2018, respectively. The density of P. mascula plants and soil seed bank were found to be affected by the clearing treatment (Table 5). The number of peony plants was significantly higher in closed plots than in the cleared plots (Figure 9). Specifically, the total density of peony population in closed plots was ~290% and ~420% higher than that recorded in the cleared plots in 2017 and 2018, respectively (Figure 9). Notably, a higher density of mature peony plants than seedlings was observed in 2017 in cleared and closed plots (Figure 9). However, the opposite pattern was evident in the following year in closed plots, whilst no difference was observed between seedling and mature plant density (Figure 9). Significantly higher peony seed density was recorded in the closed plots compared to cleared plots (P= 0.021; Figure 10). Similarly, the percentage of flowering plants was higher in closed than in cleared plots, but these differences were significant only in 2018 (Figure 11). Notably, no flowers were observed in cleared plots

Figure 9 Density (mean± SE) of P. mascula seedlings and mature plants in plots that were cleared from trees in 1997 and closed plots in the first and second years of survey. Different letters above the bars indicate significant differences as indicated by Tukey's post-hoc test. Dashed line and grey area represent the average density and standard error, respectively, of P. mascula plants documented by Ne'eman (2003) (n=5, N=10). documented by Ne'eman (2003).

Table 5: Summary of the generalized linear mixed models testing for the effect of tree clearing treatment, sampling year, maturation stage of P. mascula and their interaction on P. mascula density, flower percentage and on P. mascula seed density.

Density Flower percentage Seed bank density Estimates SE z value P value Estimates SE z value P value Estimates SE z value P value Year 2.388 0.714 3.340 0.001 0.079 0.043 1.860 0.062 Treatment 3.243 1.357 2.390 0.017 -0.023 0.098 -0.230 0.814 1.468 0.637 2.310 0.021

Maturation stage 2.814 0.691 4.070 <0.001 Year × Treatment -0.194 0.767 -0.250 0.800 0.087 0.060 1.450 0.147

Year × Maturation stage -1.407 0.408 -3.450 0.001

Treatment × Maturation stage -1.047 0.751 -1.390 0.163 Year × Treatment × Maturation stage 0.038 0.448 0.090 0.932 Constant term -2.957 1.262 -2.340 0.019 -0.079 0.069 -1.140 0.254 -2.672 0.638 4.190 <0.001 Plot Variance 0.111 0.057 0.001 0.002 2.843 1.249

43 *

Figure 10 Density (mean± SE) of the P. mascula soil seed bank in plots that were cleared from trees in 1997 and closed plots in the first and second years of survey (n=5, N=10).

Figure 11 The percentage (mean± SE) of flowering P. mascula in plots that were cleared from trees in 1997 and closed plots in the first and second years of survey. Different letters above the bars indicate significant differences as indicated by Tukey's post-hoc test (n=5, N=10).

44 3.2.2 Germinable soil seed bank

Only 164 seedlings emerged from soil samples collected from the closed and cleared plots together. Specifically, 90 and 74 seedlings from 44 families emerged from the cleared and the closed plots, respectively. Both germination density and richness were not affected by the tree clearing treatment (Table 6, Figure 12).

Table 6: Summary of the generalized linear mixed models testing for the effect of tree clearing treatment on germinable soil seed bank density and species richness. Density (m2) Richness Estimates SE z value P value Estimates SE z value P value Treatment 0.708 0.901 0.790 0.432 0.443 0.810 0.550 0.585 Constant term -1.991 0.808 -2.460 0.014 -1.438 0.677 -2.120 0.034 Plot Variance 1.474 1.263 1.122 0.885 Sample (Plot) Variance 5.123 1.881 1.909 0.840

Figure 12 Germination (1) density and (2) richness (mean± SE) from the soil seed bank in plots that were cleared from trees in 1997 and closed plots in the first and second years of survey. No significant differences in germination density and richness were detected between samples collected from cleared and closed plots (n=45, N=90).

45 3.2.3 Canopy closure

LAI measurements were not significantly different between cleared and closed plots (Table 7, Figure 13).

Table 7: Summary of the generalized linear mixed models testing for the effect of tree clearing treatment on the leaf area index

Estimates SE z value P value

Treatment 0.072 0.131 0.55 0.582 Constant term 1.640 0.094 17.47 <0.001

Plot Variance 0.0311 0.02

Sample(Plot) Variance 4.09E-35 1.11E-18

Figure 13 A box plot showing the mean (dashed lines) and median (full lines) LAI (Leaf Area Index) in plots that were cleared from trees in 1997 and closed plots in the first and second years of survey (n=80, N=160).

46 4. Discussion

4.1 Discussion Experiment 1: Fire season and microhabitat

Fire disturbance plays a key role in dictating plant community dynamics (Bond and Keeley 2005). Recent studies have shown that fire risk in the Mediterranean basin can also be high during spring and autumn (Giannakopoulos et al. 2009). I carried out a large-scale field experiment whose aim was to explore if and how the effects of fire season and microhabitat interact to influence the composition of the germinable soil seed bank community in typical East-Mediterranean woodlands. I demonstrate that the germinable soil seed bank community composition is influenced by fire in general, and by fire season through changes in the relative abundance and richness of dwarf shrub and annual species.

Despite the overall reduction in the germinable soil seed bank density and the richness of annual species, their species richness in the germinable soil seed bank community higher after spring than after autumn burnings. An opposite pattern was detected among dwarf shrubs, demonstrating a higher germinable soil seed bank density after fire, with higher density after autumn burnings. Since the intensity and severity of prescribed fires was largely consistent between the two burning seasons, it is most likely that the observed differences in the germination responses of annuals and dwarf shrubs to fire season were related to the post-fire environmental conditions combined with phenological effects.

On the one hand, the time lag between the spring burns and germination season (i.e., rainy season) was longer than that related to autumn burns, extending the exposure of seeds to predators, especially in the open canopy gaps created by fires (Torre and Díaz 2004). In addition, the hard-coated seeds of shrubs and dwarf shrubs, which their physical dormancy was broken by spring burns, had to survive the long dry summer unprotected to successfully germinate in the rainy season (i.e., autumn and winter), resulting in lower germination density of these species. This may explain the lower germination density of dwarf shrubs in areas subjected spring burns compared to areas subjected to autumn burns. Notably, a parallel study examining seasonal fire effects on the aboveground perennial plant community in the same experimental plots, illustrated that seeders experienced a stronger reduction in their abundance after spring than after autumn burns (Tsafrir et al. 2018). On the other hand, during the time lag between the

47 spring burns and the germination season, seeds in the surrounding matrix could have dispersed into the burned areas, compensating for the local loss of seeds caused by fires. This main explain why the species richness of annuals was higher in sites subjected to spring than to autumn burns. Notably, the richness and diversity of dwarf shrubs in the germinable soil seed bank community were not affected by fires, suggesting that the diversity of this functional group is less sensitive to fire compared to that of annuals.

My results showed that the germinable soil seed bank density of dwarf shrubs was higher under pistacia shrubs located in sites subjected to autumn than to spring burns, and that autumn fires enhanced the dominance of C. salviifolius and F. arabica. These results are inconsistent with findings of a study done in the western Mediterranean basin (Cespedes et al. 2012), illustrating that the germinable soil seed bank density of dwarf shrubs was higher after spring than after autumn fire, and that the dominance of C. salviifolius was higher after autumn than after spring fire. These differences illustrate that although plant communities in the eastern and western Mediterranean basin are quite similar, these two ecosystems differ in their responses to fires in general, and in particular to fire season. Such differences may be a consequence of lower precipitation and more severe summer droughts in the East-Mediterranean basin (Saaroni et al. 1998, Keeley et al. 2011a, Kruger 2014).

I illustrated that the effects of fire season and microhabitat interact to influence the total germinable soil seed bank density. Specifically, autumn fire acted to produce a variation in the germinable soil seed bank community composition between the open and closed microhabitats by reducing the germinable soil seed bank density, richness and diversity under Pistacia and Cistus shrubs/dwarf shrubs. Notably, this variation in germinable soil seed bank density, richness and diversity was not evident in samples subjected to spring fire. This can be attributed to seed dispersal from the surrounding matrix into burned areas after spring burnings. Specifically, the closed canopy microhabitat may act as a seed trap during seed dispersal (Jiménez and Armesto 1992, Ne'eman and Izhaki 1999), thus increasing the germinable soil seed bank density mainly in the closed canopy microhabitats located in areas subjected to spring burnings. Also, lower germination under Pistacia shrubs, especially after the autumn burns could be due to lower water content and higher terpene concentration in the Pistacia shrubs during the dry season, resulting in higher flammability (Alessio et al., 2008). High flammability can increase the survival probability of Pistacia by reducing competition

48 with other species in the post fire environment.

Fires are known as extinction trigger for many plant species (Bradstock et al. 1997). The absence of 25 species from plots subjected to burnings, being solely present in the unburned control plots, suggests that fires can reduce plant community diversity in the Mediterranean basin, as in some other Mediterranean type regions (Montenegro et al. 2004, Pausas et al. 2009). This reduction was more pronounced in species richness of annual species, mainly after autumn burnings, indicating that fires during this season can increase the risk of extinction of annuals, especially rare ones such as A. simplex and Crucianella latifolia, which were solely present in the unburned control plots. Remarkably, the existence of 25 species appearing exclusively in burned plots emphasizes the importance of fires for species conservation. The germination of these unique species may have a great impact on the variation observed in species richness between burned and unburned control plots. Since the densities of these unique species were relatively low, these findings should be treated with caution.

Several important points should be considered when interpreting the results of my study. First, prescribed burns are usually carried out when the environmental conditions allow the fire team to better control the fires, and to avoid their spread into unwanted areas. Consequently, both the intensity and severity of prescribed burns are lower than that of wild fires. Second, ideally, prescribed burns should be done in different days within each season, and at different years. However, owing to many logistical constraints associated with executing such a large-scale experiment, all prescribed burns were done in a single day within each season.

Despite these limitations, my findings have important value. First, my study is the first large-scale field experiment in the East-Mediterranean basin involving prescribed burns. Second, my results clearly indicate that fire season differentially affects the composition of the germinable soil seed bank community. A parallel study showed a differential effect of seasonal fires on the aboveground vegetation community (Tsafrir et al. 2018). Third, fire frequency is predicted to increase in the Mediterranean basin, along with drastic changes in seasonality. Specifically, the winter season is expected to be shorter by 56%, while both summer and autumn lengths are expected to increase by 49% and 24%, respectively (Hochman et al. 2018). Given the high spatio-temporal variation characterizing fires in the East-Mediterranean basin, the differential fire season effects observed in my study are most likely to act as a major force shaping the

49 composition of plant community in this region over large spatial scales.

Conclusions

I suggest that post-fire succession in East-Mediterranean woodlands can be highly dependent on the spatio-temporal scales shaping the post-fire germinable soil seed bank communities. Fire season acted as a source for variation in the composition of the germinable soil seed bank community, which was attributed to the increased germinable soil seed bank density of dwarf-shrubs together with the decreased germinable soil seed bank density and richness of annuals in the post-fire environment. The observed variation in the density of annuals, which comprised most of the germinable soil seed bank density, between areas subjected to spring and autumn burnings should be attributed to the combined effect of the post-fire environmental conditions and vegetation phenological stage, rather than to fire intensity or severity, which were largely consistent between the two burning seasons. The spatial heterogeneity in microhabitat type interacted with the fire season effect to differentially influence the post-fire germination dynamics in Mediterranean woodlands. This interaction mainly affected germination under shrubs, suggesting that the expected increase in fire frequency during spring and autumn (Giannakopoulos et al. 2009) may translate into a shift in the eco-evolutionary selection pressures (Gilman et al. 2010) operating on plants inhabiting this unique Mediterranean ecosystem.

4.2 Discussion Experiment 2: Fire frequency and smoke

Although smoke is a reliable cue of fire, relatively little is known about its effect on germination from the soil seed bank, especially in the East-Mediterranean basin (Keeley and Baer-Keeley 1999, Ne'eman et al. 2012). I demonstrated here, that the composition of the germinable soil seed bank community in the East- Mediterranean basin was differentially influenced by aerosol smoke, and that this effect could be also dependent on fire history.

My results illustrated that similarly to other fire-prone ecosystems such as California chaparral (Keeley and Fotheringham 1998) Fynbos in Africa (Brown et al. 2003) and mixed forests in Mexico (Zuloaga-Aguilar et al. 2011), species in the East- Mediterranean basin showed responsiveness to smoke. However, germination stimulation by smoke is not restricted to fire-prone areas and is evident in fire-free ecosystems (Figueroa and Cavieres 2012). Since smoke was recognized as germination

50 stimulated cue in 1990 (de Lange and Boucher 1990), there were several active compounds that have been identified (Flematti et al. 2004, van Staden et al. 2004) including karrikinolide and glyceronitrile (Flematti et al. 2004, Dixon et al. 2009, Flematti et al. 2011), however, how these compounds promoting or inhibiting seed germination is not completely clear (Chen 2014), suggesting that there are more components in smoke that can affect germination.

I showed that germinable soil seed bank density was higher in sites that were burned more frequently during the last four decades. This germination patterns were similar to those observed in other Mediterranean ecosystem, showing increased germinable soil seed bank density among obligate seeders, with increased fire frequency (Santana et al. 2014). A similar pattern was described for the aboveground shrub and dwarf-shrub community in Mount Carmel, despite the overall decrease in total vegetation cover (Tessler et al. 2016a). These drastic differences between the responses of the natural plant community to fires and its potential to regenerate via germination, illustrate the strong impact of environmental filtering and species sorting (Herrera 1992). For instance, in areas characterized by low fire frequency and intensity, the presence and post-fire recovery of shrubs and dwarf-shrubs can simply hamper the establishment of annuals, preventing them from properly utilizing their germination potential (Bohlman et al. 2016). However, such negative effects on the germination of annuals are expected to be weaker in areas subject to high fire frequency and intensity, where the survival of most shrubs and dwarf-shrubs is low.

Notably, the effect of smoke on germinable soil seed bank density of annuals (but not of dwarf shrubs) was greater in samples originating from areas burned more frequently during the last four decades. A higher density of dwarf-shrubs might have been observed should we allowed fire heat to break the physical dormancy of the hard- coated seeds of some dwarf shrub species (e.g. of the Cistaceae family) prior to the exposure to smoke (Izhaki et al. 2000). In contrast to woody species, only little is known about smoke stimulated germination in annuals inhabiting the East- Mediterranean basin (Eshel et al. 2000, Keeley et al. 2011a). Annual species in this region are mostly known as post-fire invaders, colonizing open gaps created by fires (Eshel et al. 2000). Together with previous studies performed in the Mediterranean basin (Santana et al. 2014, Tormo et al. 2014); my results suggest that the germinable soil seed bank of annual species plays an important role in the post-fire environment

51 characterizing the East- Mediterranean basin, where their density and richness are significantly enhanced by smoke.

Examining germinable soil seed bank density of annuals at the species-level, I illustrated that A. arvensis germination density was significantly increased in response to smoke in samples collected from sites subjected to fires during the last four decades. However, it is important to note that more species in the community, and not only A. arvensis, were affected by the aerosol smoke. Notably, the germinable soil seed bank density of other annuals increased in response to aerosol smoke, but mainly in sites that were burned more frequently during the last four decades. These observed higher germination densities in response to smoke suggest that these annual species accumulate a soil seed bank that is triggered to germinate by smoke. Such local increases in the soil seed bank of annuals can be attributed to an increase in their establishment in vegetation gaps created by frequent fires, which in turn resulted in increased annual density (Keeley and Fotheringham 2000).

In addition to the effect of aerosol smoke on germination, my experimental setup also enabled me to examine the importance of the temporal order of two key environmental events, increase in soil moisture and exposure to smoke. When the soil was watered prior to the exposure to aerosol smoke, the positive effect of the latter on seed germination was eliminated. This was true even though the watering of the two treatments differed by only a few hours. These findings suggest that the water in the soil can buffer the effect of aerosol smoke either chemically, by enabling the soil to interact with the active components of the smoke and/or physically by occupying the pores within the soil, thus reducing the penetration of smoke into the soil. My study also highlights the importance of using aerosol smoke on dry soils harboring seeds, rather than exposing seeds within a synthetic matrix to smoke. In addition, the effect of smoke on the germination of intact natural soil samples has been rarely investigated, with most studies examining the direct effect of aqueous smoke on germination patterns of selected seeds in vitro (Pérez-Fernández and Rodríguez- Echeverría 2003, Reyes and Trabaud 2009, Moreira et al. 2010, Çatav et al. 2012). Studies examining the effect of aerosol smoke on intact soil seed communities, preferably in the field, should result in the highest ecological validity, regarding the effect of smoke on the germination potential from the soil seed bank.

In conclusion, my findings indicate that fire selects for species that are stimulated

52 by fire and fire cues, resulting in a large germinable soil seed bank community dominated by smoke stimulated species. Moreover, the synergistic effect of fire history and smoke on germinable soil seed bank community composition indicates that smoke can play an important role in shaping the post-fire succession in this ecosystem, mainly by stimulating the germination of annual species. Clearly, this pattern would be more pronounced in areas subjected to frequent fires, where competition with perennial shrubs and dwarf-shrubs should be weaker.

Focusing on the applicative value of this study, I showed that differences in fire history can differentially influence the germinable soil seed bank community composition, potentially affecting the vegetation composition. I thus suggest that promoting variability of fire mosaic (i.e., fire frequency, fire severity, fire patchiness) can play a key role in maintaining biodiversity.

4.3. Discussion Experiment 3: Tree clearing (Paeonia)

I quantified the long-term effect of a pulse canopy opening on the population of P. mascula and on the germinable soil seed bank of its neighboring plant species from the soil seed bank. I found that the short-term positive effect of a pulse canopy opening on the flowering of P. mascula turned into a negative effect over a long-time scale.

The long-term effect of a pulse canopy opening was to reduce the population density of P. mascula, the proportion of flowering plants, as well as the density of its respective long-term negative effects of the canopy opening. Furthermore, the number of P. mascula plants prior to the canopy opening (i.e., 20 years ago within the same experimental plots) was higher than that detected in our study. This reduction indicates that the long-term effect of a pulse canopy opening was not restricted only to the reproduction of P. mascula, but also to its regeneration and survival rate at the population level.

Several studies have illustrated that canopy opening can positively affect the reproduction of different Paeonia species (Ne'eman 2003, Andrieu et al. 2013, Andrieu et al. 2017), while the shading of open areas occupied by Paeonia regressed the individual plants into a vegetative stage (Andrieu et al. 2013). Specifically, the response of Paeonia plants to canopy opening and shading was relatively fast (i.e., one- or two- years following canopy opening/shading treatment). The results of these studies may explain the long-term negative effect of a pulse canopy opening on the reproduction of

53 P. mascula. Specifically, I interpreted this to mean that the short- term changes in resource allocation towards higher reproduction and photosynthetic rate (Zhang et al. 2003) occurring when the canopy is open may be costly, thus resulting in decreased reproduction performance. This is evident over a longer time scale, when the canopy returns to its closed state (Pugnaire and Valladares 2007).

Although I could not pinpoint the ecological factors that caused a reduction in the reproduction of P. mascula in the long-term, this result can be attributed to several processes related to negative effects of the canopy opening. Notably, Mount Meron is the southernmost range of distribution of Paeonia species, as well as the warmest region that is populated with P. mascula. Remarkably, in Mount Meron, P. mascula plants inhabit mainly northern slopes of closed canopy woodlands. However, contrary to its southern distribution, in northern distribution regions, Paeonia species inhabit mainly open habitats (Passalacqua and Bernardo 2004, Andrieu et al. 2013). This can suggest that the population in Mount Meron may avoid heat and dry conditions that could be associated with an open canopy. Thus, although the immediate response of P. mascula plants to the canopy opening was positive, unsuitable conditions (e.g. heat and dry conditions) could evolve at the same time. Consequently, these conditions could negatively affect the reproduction and survival rate of P. mascula population in the long run. Moreover, between-year variation in climate conditions such as dryness, heat and precipitation may possibly affect the cleared more than the closed plots.

Pulse canopy opening acted to increase the density of annual and perennial herbaceous plants in the short term (i.e., four years after the canopy opening) (Ne'eman 2003). I therefore hypothesized that it should also increase the input of herbaceous plant seeds into the soil seed bank in the cleared plots. However, I could not detect such a long-term effect on the germinable soil seed bank of neighboring plant species, as no differences were evident in the germination density and richness of the soil seed bank between the opened and closed plots. This may be due to a rapid decay of these shade- intolerant herbaceous plant species during the process of canopy closure (Anderson et al. 1969, Halpern and Lutz 2013). Moreover, the total density of the seed germination was relatively small (i.e. 164 emerged seedlings), indicating that the germination potential from the soil seed bank in the closed canopy areas in Mount Meron is low, or that a high percentage of the seeds are dormant. These results also suggest that long- range seed dispersal is less likely to occur into these cleared plots that are surrounded by

54 dense maquis.

As many natural landscapes undergoing a major regime shift, potentially threatening the existence of different plant species, understanding the long-term implications of management actions aimed to prevent such shifts is of high importance for biodiversity conservation. My results indicate that P. mascula population can be maintained under shade conditions for long time periods. Moreover, a recent report of a long-term survey (Oron 2017) showed that the population size of P. mascula in unmanaged sites at Mount Meron nature reserve was stable over the last 45 years, suggesting that similarly to my results, P. mascula plants can persist under shade conditions in the long run. Notably, this area is subjected to successional processes involving the establishment of tall trees with natural open gaps that can enhance the flowering of the P. mascula. Therefore, passive management should be recommended for the conservation of the P. mascula in Mount Meron nature reserve. However, if active management involving canopy opening is considered, changes in microclimate of the canopy gaps should be taken into account. Exploring the effect of processes associated with habitat preference of Paeonia species in its southernmost distribution range (e.g., Mount Meron, Israel) can improve the understanding of the effect of climate change and global warming on its distribution and management in northern areas.

55 4.3 General discussion

The Mediterranean basin comprises a great diversity of plant species, making this area the third richest hotspot in the world in terms of its plant biodiversity (Mittermeier 2004). The main factors contributing to this high diversity are the complex geological history, the seasonal climate and the long history of human activity (Thompson 2005, Keeley et al. 2011a). These characteristics, in combination with a long history of disturbances (including fires, grazing and soil disturbances), have exposed the plant community to different selective forces which ultimately determined the structure and function of this ecosystem. Hence, plant regeneration via germination has been subjected to strong selection pressures in the Mediterranean basin, resulting in a plant community that is disturbance-resilient. However, major shifts in the disturbance regime that have been documented during the last century (Grubb and Hopkins 1986, Pausas 2004) suggest that different characteristics of the plant community are expected to change, including the community compositions of the soil seed bank and the respective germination patterns from the soil seed bank (Santana et al. 2014). The main objectives of my PhD study were to investigate the effect of disturbance timing, history and regime on germination patterns from the soil seed bank.

The first chapter of my doctoral research focused on exploring if and how seasonal fires affect germination patterns from the soil seed banks. To the best of my knowledge, it is the first attempt to examine this question in the East- Mediterranean basin using a large-scale field experiment involving prescribed burns. My findings clearly illustrate that fire season differentially influences the composition of the germinable soil seed bank community. This differential fire season effect should be attributed to phenological and environmental effects, rather than to fire intensity/severity effects, and it can have important implications for plant community dynamics in this unique ecosystem. This may include selection for early seed dispersal in areas subjected to spring fires, or for hard coated seeds in areas exposed to autumn fires.

The second chapter of my study focused on exploring the synergistic effect of fire history and smoke cue on the germinable soil seed bank community. My findings demonstrate for the first time that, contrary to the accepted notion, smoke cue can stimulate the germination of plant species in the East- Mediterranean basin, especially of annuals. Moreover, I illustrate that this pattern is more pronounced in

56 areas that have been subjected to more fires during the last four decades, i.e., smoke cue and fire history act synergistically to affect the composition of the germinable soil seed bank community. Finally, in contrast to most studies, I examined the response of intact natural soil samples to aerosol smoke, rather than examining the responses of selected seeds in vitro to aqueous smoke; doing so increases the ecological validity of my findings.

In the third chapter, I examined the long-term effect of a pulse canopy opening on the germinable soil seed bank density and richness and on the density and reproduction of the endangered species P. mascula. Specifically, I monitored plots that were cleared from trees ~20 years ago by Ne'eman (2003), and from plots that remained closed. In his work, Ne'eman found that higher radiation increased the number of P. mascula flowering plants and the density of neighboring herb species. My study aimed to understand if these positive effects of canopy opening can persist over long time scales. These kinds of studies involving the monitoring of seed germination and plant populations in areas subjected to management actions over long time scales are relatively rare. Although I could not detect a significant effect of the pulse canopy opening on the germinable soil seed bank density and richness after 20 years, there was clear evidence for a negative effect on the density and reproduction of the endangered species P. mascula. Still, it is not clear if the negative effect of the pulse canopy opening resulted from processes related to canopy closure, or to changes in microenvironmental conditions after the clearing treatment. Nevertheless, my findings indicate that the population of P. mascula is stable in sites that remained closed, therefore no intervention is necessary for its persistence. However, if management is required for maintaining the population of P. mascula, a press (i.e., continuous active interference) rather than pulse management (i.e., one-time perturbation) should be applied. This active management plan should be performed in short intervals to ensure the penetration of adequate light required for the reproduction of P. mascula through the woodland canopy.

Mediterranean ecosystems are considered resilient to fire, which is often essential for maintaining optimal diversity conditions (Keeley et al. 2011a). An alteration in disturbance regime (e.g. changes in timing, severity or frequency) can potentially trigger a shift between alternative stable states in this apparently resilient ecosystem. The disturbance regime is a major source of variation in the composition of the germinable

57 soil seed bank community, manifested by the relative representation of two main life forms: annuals and dwarf shrubs. My results show that variation in the fire regime (i.e., fire season and fire history/frequency) can differentially affect the composition of these communities by weakening the dominance of one group while increasing the representation of the other. Specifically, dwarf shrub species with hard-coated seeds rapidly colonizing after fires, mostly because their physical dormancy can be broken by fire heat (Valbuena et al. 1992, Moreira et al. 2010, Keeley et al. 2011a), were mainly abundant in the germinable soil seed bank community characterizing areas that were subjected to fires in general, and those subjected to autumn fires in particular. The differences in dwarf shrub germination density between areas subjected to autumn and spring burns represent the proportion of seeds whose physical dormancy was broken during spring burns but that did not survive the long and dry summer to germinate in the rainy season. Similar to the prescribed fire experiment, a higher density of dwarf shrub species was observed in the germinable soil seed bank community characterizing sites that have been subjected to fires during the last four decades, in comparison to that of unburned sites. These results support previous studies reporting higher germination of shrubs, dwarf shrubs and other seeder species in Mediterranean woodlands subjected to recurrent fires (Izhaki et al. 2000, Moreira and Pausas 2012). Specifically, the main dwarf-shrub species mostly contributed to increased abundance after fires were Cistus spp., which are very common in the aboveground vegetation characterizing the woodlands in the East-Mediterranean basin (i.e. Mount Yaaran and Mount Carmel) (Tessler et al. 2016b).

In contrast to dwarf-shrub species, the germination of annuals was lower after the prescribed burns, indicating that the immediate effect of fire is to reduce the germinable soil seed bank of annuals. However, in contrast to this immediate response to fire, increased fire frequency acted to induce the germination of annual species over longer time periods (i.e. five years after the last fire). These results suggest that fire history has the capacity to modify germination patterns from the soil seed banks in the eastern Mediterranean ecosystem through different mechanisms. Specifically, my results suggest that although annual germination was reduced by fire over a short time scale of few years, fire history and smoke can act to stimulate the germination of annuals in fire prone areas (Manela et al. 2018), thus allowing them to successfully refill their seed bank. Differences in the density and richness of annuals were clearly observed between unburned areas and areas burned three times in the last four decades, while areas that

58 suffered only one fire exhibited intermediate compositional and density values between these two extremes. Therefore, these results strengthen the previous study claiming that fire frequency can act to enhance the representation of small-sized, quick-reproducing life forms in the soil seed bank (i.e. annual species) (Santana et al. 2014). Such a local increase in the soil seed bank of annuals can be attributed to an increase in their establishment in vegetation gaps created by frequent fires, which in turn result in increased annual density. My findings clearly illustrate that fire regime with short inter- fire intervals can modify the balance of the germinable soil seed bank community to an annual-dominated community.

Studies in Mediterranean regions showed that the germination of some annual herbaceous species can be directly stimulated by fire. For instance, grasslands in California are relatively resilient to frequent fires (Reiner 2007). In these habitats, low fire intensity resulting from low fuels lead to high seed survivorship. Notably, there are some annual species (e.g., Trifolium spp., Lotus spp and Silene spp.) distributed in both grasslands and chaparral, but their post-fire germination response is different (Keeley & Davis 2007). Other example are Mediterranean ecosystems of the Northern hemisphere including shrublands and woodlands in South Africa and Southwest Australia that were developed due to the absence of disturbances, including fires (Keeley et al. 2011a). The understory in these areas with low fertility soils comprises a low diversity of annual and perennial herbaceous but in the soil, they accumulate a rich diversity of dormant seeds that should increase their aboveground community diversity after fire. These responses suggest of ecotypic variation between habitats and ecosystems, suggesting that the differences in germination responses of annuals between areas subjected to fires and unburned areas may be related to variation in dormancy characteristics.

Even though post-fire germination communities are commonly characterized by the representation of mainly dwarf shrub species, a higher abundance of annual species in the germination community was also evident in areas exposed to smoke cue. These results indicate that although annuals are likely to be more sensitive to fire, many annual species can maintain their populations by germination from the soil seed bank (Manela et al. 2018). These results also suggest that the germinable soil seed bank community of annual species plays an important role in shaping the post- fire environment characterizing the East- Mediterranean basin. In contrast to previous studies, I tested the effect of smoke on the germination of intact natural soil samples. The main advantage

59 of this method is the ability to evaluate the germination potential of the entire soil seed bank community, and to identify candidate species that are affected by smoke cue. Indeed, Anagallis arvensis, an annual herb that was previously reported not to be affected by smoke (Ne'eman et al. 2012), showed a positive germination response to smoke. Germination stimulation by smoke was also observed in other species such as Poa infirma and Mercurialis annua, not as before reported to be affected by smoke. Clearly, further research is needed to understand the species-specific response to smoke and the mechanism behind these responses.

During the last century, disturbance regime characterizing many Mediterranean ecosystems has intensified and the anticipated future climatic changes are expected to further amplify this worrying pattern (Seidl et al. 2011). These changes are increasingly challenging the main objectives of forest and woodland ecosystem management, which are to maintain the biological diversity of the ecosystem. Nature conservation policies can be considered to be composed of two contrasting views including passive and active management (Williams 2011). The active approach sets up specific goals to be achieved; for example, the conservation of biodiversity and populations of endangered species, or even the reintroduction of recently extinct species. In most cases, this approach will affect the ecosystem in the desired direction. However, it is highly important to understand the impact of the management action on the entire ecosystem in time and space. Studying the long-term effect of a pulse canopy opening clearly showed that although the clearcutting treatment did not affect the germination from the soil seed bank, it negatively influenced the population of the endangered species P. mascula. Furthermore, these results demonstrate the importance of studying the long-term consequences of management actions. Specifically, although Mount Meron has undergone a major shift from mosaic of close woodlands and many open patches into a closed and dense woody vegetation, my results indicated that the population of P. mascula is resistant to these changes.

Research significance and future directions

By studying germination patterns of disturbed Mediterranean woodlands, I demonstrated that a shift in disturbance regime can strongly affect the composition of the germinable soil seed bank community. I also showed that smoke cue is an important player in the germination of Mediterranean plant species, especially in areas subjected to fires. Specifically, my study indicated that the changes in fire season and fire

60 frequency have the potential to cause a shift in plant community composition, driven by plant species that regenerate from seeds stored in the soil seed bank. These findings illustrate how fire selects for species that are stimulated by fire and fire cues, resulting in a large germinable soil seed bank community dominated by fire and smoke stimulated species. Still, the mechanism of germination stimulation by smoke is not completely understood and further studies involving characterization of species and germination responses to smoke are required. My study highlights the importance of examining the response of intact soil samples, rather than that seeds within a synthetic matrix to smoke cue. Future studies should examine the effect of smoke and fire history at the species level in order to better understand the mechanism of smoke and fire stimulating germination.

Finally, I was able to demonstrate that long-term studies are crucial for evaluating the consequences of management actions aimed at counteracting a shift in the disturbance regime. The decline of P.mascula population 20 years after the pulse canopy opening raises the need to investigate and carefully monitor our management actions over long time scales, and specifically to study the mechanisms causing these negative effects on the density and reproduction of P.mascula in areas cleared 20 years ago. Furthermore, in order to understand why canopy opening resulted in a shift in the population size of P. mascula, a follow-up study involving the examination of multiple- scenarios including growing plants in both open and close canopy conditions, together with different possible environmental conditions such as low/high irrigation, and low/high temperatures should be conducted. Such a study may allow understanding why the distribution of P. mascula in Israel is limited.

61 5. Cited literature

Agra, H. e., and G. Ne’eman. 2009. Woody species as landscape modulators: their effect on the herbaceous plants in a Mediterranean maquis. Plant Ecology.

Agra He, Ne’eman G, Shachak M, Segoli M, Gabay O, Perevolotsky A, Arnon A, Boeken B, Groner E, Walczak M. 2015. Canopy structure of woody landscape modulators determines herbaceous species richness along a rainfall gradient. Plant Ecology 216:1511-1522.

Alessio G.A., Peñuelas J, Llusià J, Ogaya R, Estiarte M, De Lillis M. 2008. Influence of water and terpenes on flammability in some dominant mediterranean species. International Journal of Wildland Fire 17:274- 286.205:165-177.

Anderson, M., R. N. Gorley, and R. K. Clarke. 2008. Permanova+ for Primer: Guide to Software and Statisticl Methods. Primer-E Limited.

Anderson, R., O. Loucks, and A. M. Swain. 1969. Herbaceous response to canopy cover, light intensity, and throughfall precipitation in coniferous forests. Ecology 50:255-263.

Anderson, T. M., M. Schütz, and A. C. Risch. 2012. Seed germination cues and the importance of the soil seed bank across an environmental gradient in the Serengeti. Oikos 121:306-312.

Andrieu, E., A. Besnard, H. Fréville, V. Vaudey, P. Gauthier, J. D. Thompson, and M. Debussche. 2017. Population dynamics of Paeonia officinalis in relation to forest closure: From model predictions to practical conservation management. Biological Conservation 215:51-60.

Andrieu, E., H. Fréville, A. Besnard, V. Vaudey, P. Gauthier, J. D. Thompson, and M. Debussche. 2013. Forest-cutting rapidly improves the demographic status of Paeonia officinalis, a species threatened by forest closure. Population ecology 55:147-158.

Bakker, J. P. 2005. Vegetation conservation, management and restoration. Vegetation ecology:309-331.

62 Baldwin, I. T., and L. Morse. 1994. Up in smoke: II. Germination ofNicotiana attenuata in response to smoke-derived cues and nutrients in burned and unburned soils. Journal of chemical ecology 20:2373-2391.

Baskin, J. M., & Baskin, C. C. (2004). A classification system for seed dormancy. Seed science research, 14(1), 1-16.

Beisner, B. E., D. T. Haydon, and K. Cuddington. 2003. Alternative stable states in ecology. Frontiers in Ecology and the Environment 1:376-382.

Benhammou, N., F. A. Bekkara, and T. K. Panovska. 2008. Antioxidant and antimicrobial activities of the Pistacia lentiscus and Pistacia atlantica extracts. African Journal of Pharmacy and Pharmacology 2:022-028.

Bohlman, G. N., M. North, and H. D. Safford. 2016. Shrub removal in reforested post-fire areas increases native plant species richness. Forest ecology and management 374:195-210.

Bond, W. J., and J. E. Keeley. 2005. Fire as a global ‘herbivore’: the ecology and evolution of flammable ecosystems. Trends in ecology & evolution 20:387- 394.

Bonis, A., J. Lepart, and P. Grillas. 1995. Seed bank dynamics and coexistence of annual macrophytes in a temporary and variable habitat. Oikos:81-92.

Bradshaw, S. D., K. W. Dixon, S. D. Hopper, H. Lambers, and S. R. Turner. 2011. Little evidence for fire-adapted plant traits in Mediterranean climate regions. Trends in plant science 16:69-76.

Bradstock, R., M. Tozer, and D. Keith. 1997. Effects of high frequency fire on floristic composition and abundance in a fire-prone heathland near Sydney. Australian Journal of Botany 45:641-655.

Brown, N. 1993. Promotion of germination of fynbos seeds by plant‐derived smoke. New Phytologist 123:575-583.

Brown, N., and J. Van Staden. 1997. Smoke as a germination cue: a review. Plant Growth Regulation 22:115-124.

Brown, N., J. Van Staden, M. Daws, T. Johnson, and A. Van Wyk. 2003. Patterns in the seed germination response to smoke in plants from the Cape

63 Floristic Region, South Africa. South African Journal of Botany 69:514- 525.

Bullock, J. M. 2000. Gaps and seedling colonization. Seeds: the ecology of regeneration in plant communities, 2, 375-395.

Carmel, Y., and C. H. Flather. 2004. Comparing landscape scale vegetation dynamics following recent disturbance in climatically similar sites in California and the Mediterranean basin. Landscape Ecology 19:573-590.

Carmel, Y., and R. Kadmon. 1999. Effects of grazing and topography on long- term vegetation changes in a Mediterranean ecosystem in Israel. Plant Ecology 145:243-254.

Carrington, M. E. 2010. Effects of soil temperature during fire on seed survival in Florida Sand Pine Scrub. International Journal of Forestry Research 2010.

Çatav, Ş. S., I. Bekar, B. S. Ateş, G. Ergan, F. Oymak, E. D. Ülker, and Ç. Tavşanoğlu. 2012. Germination response of five eastern Mediterranean woody species to smoke solutions derived from various plants. Turkish Journal of Botany 36:480-487.

Cespedes, B., I. Torres, B. Luna, B. Perez, and J. M. Moreno. 2012. Soil seed bank, fire season, and temporal patterns of germination in a seeder- dominated Mediterranean shrubland. Plant Ecology 213:383-393.

Cespedes, B., I. Torres, B. Perez, B. Luna, and J. M. Moreno. 2014. Burning season does not affect post‐fire regeneration but fire alters the balance of the dominant species in a seeder‐dominated Mediterranean shrubland. Applied Vegetation Science 17:711-725.

Chamorro, D., B. Luna, J. M. Ourcival, A. Kavgacı, C. Sirca, F. Mouillot, M. Arianoutsou, and J. Moreno. 2017. Germination sensitivity to water stress in four shrubby species across the Mediterranean Basin. Plant Biology 19:23-31.

Chen, H., and P. C. Boutros. 2011. VennDiagram: a package for the generation of highly-customizable Venn and Euler diagrams in R. BMC bioinformatics 12:35.

64 Chen, Y. 2014. The effect of smoke on seed germination: Global patterns and regional prospects for the Southern High Plains (Doctoral dissertation).

Clarke, K., and R. Gorley. 2006. PRIMER version 6: user manual/tutorial. PRIMER- E, Plymouth, UK 192.

Clarke, K., and R. Warwick. 1994. An approach to statistical analysis and interpretation. Change in Marine Communities 2.

Clarke, K. R. 1993. Non‐parametric multivariate analyses of changes in community structure. Australian Journal of Ecology 18:117-143.

Connell, J. H. 1978. Diversity in tropical rain forests and coral reefs. Science, 199(4335), 1302-1310.

Cowling, R. M., P. W. Rundel, B. B. Lamont, M. K. Arroyo, and M. Arianoutsou. 1996. Plant diversity in Mediterranean-climate regions. Trends in ecology & evolution 11:362-366.

Crosti, R., P. Ladd, K. Dixon, and B. Piotto. 2006. Post-fire germination: the effect of smoke on seeds of selected species from the central Mediterranean basin. Forest Ecology and Management 221:306-312.

De Lange, J. H., & Boucher, C. 1990. Autecological studies on audouiniacapitatabruniaceae I. plant-derived smoke as a seed germination cue. South African Journal of Botany, 56(6), 700-703.

Delitti, W., A. Ferran, L. Trabaud, and V. R. Vallejo. 2005. Effects of fire recurrence in Quercus coccifera L. shrublands of the Valencia Region (Spain): I. plant composition and productivity. Plant Ecology 177:57-70.

Dimitrakopoulos, A. 2001. A statistical classification of Mediterranean species based on their flammability components. International Journal of Wildland Fire 10:113-118.

Dixon, K. W., Merritt, D. J., Flematti, G. R., & Ghisalberti, E. L. 2009. Karrikinolide - a phytoreactive compound derived from smoke with applications in horticulture, ecological restoration and agriculture. In M. Johnston, M. J. O. Dragovic, & R. A. Criley (Eds.), Vi International Symposium on New Floricultural Crops (Vol. 813, pp. 155-170, Acta

65 Horticulturae).

Dixon, K. W., S. Roche, and J. S. Pate. 1995. The promotive effect of smoke derived from burnt native vegetation on seed germination of Western Australian plants. Oecologia 101:185-192.

Donohue, K., Rubio de Casas, R., Burghardt, L., Kovach, K., & Willis, C. G. 2010. Germination, postgermination adaptation, and species ecological ranges. Annual review of ecology, evolution, and systematics, 41, 293- 319.

Du Rietz, G. E. 1931. Life-forms of terrestrial flowering plants, 1. Sv. växtgeografiska sällsk.

Enright, N., D. Goldblum, P. Ata, and D. Ashton. 1997. The independent effects of heat, smoke and ash on emergence of seedlings from the soil seed bank of a heathy Eucalyptus woodland in Grampians (Gariwerd) National Park, western Victoria. Australian Journal of Ecology 22:81-88.

Eshel, A., N. Henig-Sever, and G. Ne'eman. 2000. Spatial variation of seedling distribution in an east Mediterranean pine woodland at the beginning of post- fire succession. Plant Ecology 148:175-182.

Fenner, M. 2000. Seeds: the ecology of regeneration in plant communities. Cabi. Fenner, M. 2012. Seed ecology. Springer Science & Business Media.

Figueroa, J. A., & Cavieres, L. A. (2012). The effect of heat and smoke on the emergence of exotic and native seedlings in a Mediterranean fire-free matorral of central Chile. Revista Chilena De Historia Natural, 85(1), 101- 111.

Flematti, G. R., Ghisalberti, E. L., Dixon, K. W., & Trengove, R. D. 2004. A compound from smoke that promotes seed germination. Science, 305(5686), 977-977.

Flematti, G. R., Merritt, D. J., Piggott, M. J., Trengove, R. D., Smith, S. M., Dixon, K. W., et al. (2011). Burning vegetation produces cyanohydrins that liberate cyanide and stimulate seed germination. Nature Communications, 2, 360.

66 Fletcher, M.-S., S. W. Wood, and S. G. Haberle. 2014. A fire‐driven shift from forest to non‐forest: evidence for alternative stable states? Ecology 95:2504-2513.

Folke, C., S. Carpenter, B. Walker, M. Scheffer, T. Elmqvist, L. Gunderson, and C. S. Holling. 2004. Regime shifts, resilience, and biodiversity in ecosystem management. Annu. Rev. Ecol. Evol. Syst. 35:557-581.

Giannakopoulos, C., P. Le Sager, M. Bindi, M. Moriondo, E. Kostopoulou, and C. M. Goodess. 2009. Climatic changes and associated impacts in the Mediterranean resulting from a 2 degrees C global warming. Global and Planetary Change 68:209-224.

Gilman, S. E., M. C. Urban, J. Tewksbury, G. W. Gilchrist, and R. D. Holt. 2010. A framework for community interactions under climate change. Trends in ecology & evolution 25:325-331.

Gomez-Aparicio, L., J. M. Gomez, and R. Zamora. 2005. Microhabitats shift rank in suitability for seedling establishment depending on habitat type and climate. Journal of Ecology 93:1194-1202.

González-Rabanal, F., and M. Casal. 1995. Effect of high temperatures and ash on germination of ten species from gorse shrubland. Vegetatio 116:123-131.

Goubitz, S., M. Werger, and G. Ne'eman. 2002. Germination response to fire- related factors of seeds from non-serotinous and serotinous cones. Plant Ecology 169:195-204.

Grubb, P., and A. Hopkins. 1986. Resilience at the level of the plant community. Resilience in Mediterranean-type ecosystems. Springer. 21-38

Halpern, C. B., and J. A. Lutz. 2013. Canopy closure exerts weak controls on understory dynamics: a 30‐year study of overstory–understory interactions. Ecological Monographs 83:221-237.

Hegazy, A., and J. Lovett-Doust. 2016. Plant ecology in the Middle East. Oxford University Press.

Herranz, J. M., P. Ferrandis, and J. J. Martínez-Sánchez. 1998. Influence of heat on seed germination of seven Mediterranean Leguminosae species. Plant

67 Ecology 136:95-103.

Herrera, C. M. 1992. Historical effects and sorting processes as explanations for contemporary ecological patterns: character syndromes in Mediterranean woody plants. The American Naturalist 140:421-446.

Hochman, A., T. Harpaz, H. Saaroni, and P. Alpert. 2018. The seasons’ length in 21st century CMIP5 projections over the eastern Mediterranean. International Journal of Climatology.

Izhaki, I., N. Henig-Sever, and G. Ne'eman. 2000. Soil seed banks in Mediterranean Aleppo pine forests: the effect of heat, cover and ash on seedling emergence. Journal of Ecology 88:667-675.

Jiménez, H. E., and J. J. Armesto. 1992. Importance of the soil seed bank of disturbed sites in Chilean matorral in early secondary succession. Journal of Vegetation Science 3:579-586.

Johnstone, J. F., C. D. Allen, J. F. Franklin, L. E. Frelich, B. J. Harvey, P. E. Higuera, M. C. Mack, R. K. Meentemeyer, M. R. Metz, and G. L. Perry. 2016. Changing disturbance regimes, ecological memory, and forest resilience. Frontiers in Ecology and the Environment 14:369-378.

Kahmen, S., P. Poschlod, and K.-F. Schreiber. 2002. Conservation management of calcareous grasslands. Changes in plant species composition and response of functional traits during 25 years. Biological Conservation 104:319-328.

Kalisz, S., and M. A. McPeek. 1993. Extinction dynamics, population growth and seed banks. Oecologia 95:314-320.

Kearns, C. A., D. W. Inouye, and N. M. Waser. 1998. Endangered mutualisms: the conservation of plant-pollinator interactions. Annual review of ecology and systematics 29:83-112.

Keeley, J., and M. Baer-Keeley. 1999. Role of charred wood, heat-shock, and light in germination of postfire phrygana species from the eastern Mediterranean basin. Israel Journal of Plant Sciences 47:11-16.

Keeley, J. E. 1995. Seed-germination patterns in fire-prone Mediterranean-climate

68 regions. Pages 239-273 Ecology and biogeography of Mediterranean ecosystems in Chile, California, and Australia. Springer.

Keeley, J. E., W. J. Bond, R. A. Bradstock, J. G. Pausas, and P. W. Rundel. 2011a. Fire in Mediterranean ecosystems: ecology, evolution and management. Cambridge University Press.

Keeley, J.E. & Davis, F.W. (2007) Chaparral. In Terrestrial vegetation of California, third edition (eds M.G. Barbour, T. Keeler-Wolf & A.A. Schoenherr), pp. 339–366. University of California Press, Los Angeles, California.

Keeley, J. E., and Fotheringham, C. J. 1998. Smoke‐induced seed germination in California chaparral. Ecology 79:2320-2336.

Keeley, J. E., and Fotheringham, C. J. 2000. Role of fire in regeneration from seed. Seeds: the ecology of regeneration in plant communities 2:311-330.

Keeley, J. E., Fotheringham, C. J., & Baer‐Keeley, M. 2005. Factors affecting plant diversity during post‐fire recovery and succession of Mediterranean‐ climate shrublands in California, USA. Diversity and distributions, 11(6), 525-537.

Keeley, J. E., J. G. Pausas, P. W. Rundel, W. J. Bond, and R. A. Bradstock. 2011b. Fire as an evolutionary pressure shaping plant traits. Trends in plant science 16:406-411.

Klimek, S., M. Hofmann, and J. Isselstein. 2007. Plant species richness and composition in managed grasslands: the relative importance of field management and environmental factors. Biological Conservation 134:559- 570.

Knapp, E. E. 2010. Ecological effects of prescribed fire season: a literature review and synthesis for managers. DIANE Publishing.

Knapp, E. E., D. W. Schwilk, J. M. Kane, and J. E. Keeley. 2006. Role of burning season on initial understory vegetation response to prescribed fire in a mixed conifer forest. Canadian Journal of Forest Research 37:11-22.

Kneeshaw, D. D. & Bergeron, Y. 1998 Canopy gap characteristics and tree

69 replacement in the southeastern boreal forest. Ecology, 79, 783-794.

Kneitel, J. M., and J. M. Chase. 2004. Trade‐offs in community ecology: linking spatial scales and species coexistence. Ecology Letters 7:69-80.

Knox, K. J., and P. J. Clarke. 2006. Fire season and intensity affect shrub recruitment in temperate sclerophyllous woodlands. Oecologia 149:730- 739.

Koller, D., and T. Kozlowski. 1972. Environmental control of seed germination. Seed biology 2.

Koukoulas, S. & Blackburn, G. A. 2005 Spatial relationships between tree species and gap characteristics in broad-leaved deciduous woodland. Journal of Vegetation Science, 16, 587-596.

Kruger, F. 1984. Effects of fire on vegetation structure and dynamics. Pages 219- 243 Ecological effects of fire in South African ecosystems. Springer.

Kruger, F. 2014. Fire in Mediterranean Ecosystems: Ecology, Evolution and Management. International Forestry Review 16:113-116.

Kruger, F. J. 1983. Plant community diversity and dynamics in relation to fire. Pages 446-472 Mediterranean-type ecosystems. Springer.

Landis, T. D. 2000. Where there's smoke… There’s Germination? Native Plants Journal 1:25-29.

Lavee, H., P. Kutiel, M. Segev, and Y. Benyamini. 1995. Effect of surface roughness on runoff and erosion in a Mediterranean ecosystem: the role of fire. Geomorphology 11:227-234.

Lavorel, S., M. Debussche, J.-D. Lebreton, and J. Lepart. 1993. Seasonal patterns in the seed bank of Mediterranean old-fields. Oikos:114-128.

Le Houérou, H. N. 1973. Fire and vegetation in the Mediterranean Basin. FAO. Leck, M. A. 2012. Ecology of soil seed banks. Elsevier.

Leishman, M. R., Wright, I. J., Moles, A. T. & Westoby, M. 2000. The evolutionary ecology of seed size. Seeds: the ecology of regeneration in plant communities, 2, 31-57.

70 Liu, H., C. Gao, and G. Wang. 2018. Understand the resilience and regime shift of the wetland ecosystem after human disturbances. Science of the Total Environment 643:1031-1040.

Lloret, F. 1998. Fire, canopy cover and seedling dynamics in Mediterranean shrubland of northeastern Spain. Journal of Vegetation Science 9:417-430.

Lloret, F., and M. Vilà. 2003. Diversity patterns of plant functional types in relation to fire regime and previous land use in Mediterranean woodlands. Journal of Vegetation Science 14:387-398.

Loepfe, L., J. Martinez-Vilalta, J. Oliveres, J. Piñol, and F. Lloret. 2010. Feedbacks between fuel reduction and landscape homogenisation determine fire regimes in three Mediterranean areas. Forest Ecology and Management 259:2366- 2374.

Malak, D. A., J. G. Pausas, J. E. Pardo-Pascual, and L. A. Ruiz. 2015. Fire recurrence and the dynamics of the enhanced vegetation index in a Mediterranean ecosystem. International Journal of Applied Geospatial Research (IJAGR) 6:18-35.

Manela, N., E. Dagon, H. Semesh, and O. Ovadia. 2018. Smoke interacts with fire history to stimulate soil seed bank germination in Mediterranean woodlands. Journal of Plant Ecology.

Marriner, N. 2015. Fire in Mediterranean Ecosystems: ecology, evolution and management. Méditerranée 121:111-111.

Matsushita, M., S. Setsuko, I. Tamaki, M. Nakagawa, N. Nishimura, and N. Tomaru. 2016. Thinning operations increase the demographic performance of the rare subtree species Magnolia stellata in a suburban forest landscape. Landscape and ecological engineering 12:179-186.

Mittermeier, R. A. 2004. Hotspots revisited. Cemex.

Montenegro, G., R. Ginocchio, A. Segura, J. E. Keely, and M. Gomez. 2004. Fire regimes and vegetation responses in two Mediterranean-climate regions. Revista chilena de historia natural 77:455-464.

Moreira, B., and J. Pausas. 2018. Shedding light through the smoke on the germination of Mediterranean Basin flora. South African Journal of Botany

71 115:244-250.

Moreira, B., and J. G. Pausas. 2012. Tanned or burned: the role of fire in shaping physical seed dormancy. PLoS One 7:e51523.

Moreira, B., J. Tormo, E. Estrelles, and J. Pausas. 2010. Disentangling the role of heat and smoke as germination cues in Mediterranean Basin flora. Annals of botany 105:627-635.

Moriondo, M., P. Good, R. Durao, M. Bindi, C. Giannakopoulos, and J. Corte-Real.

2006. Potential impact of climate change on fire risk in the Mediterranean area. Climate Research 31:85-95.

Nathan, R., and H. C. Muller-Landau. 2000. Spatial patterns of seed dispersal, their determinants and consequences for recruitment. Trends in Ecology & Evolution 15:278-285.

Naveh, Z. 2007. Transdisciplinary challenges in landscape ecology and restoration ecology-an anthology. Springer Science & Business Media.

Ne'eman, G. 2003. To be or not to be—the effect of nature conservation management on flowering of Paeonia mascula (L.) Miller in Israel. Biological Conservation 109:103-109.

Ne'eman, G., and I. Izhaki. 1999. The effect of stand age and microhabitat on soil seed banks in Mediterranean Aleppo pine forests after fire. Plant Ecology 144:115-125.

Ne'eman, G., H. Lahav, and I. Izhaki. 1992. Spatial pattern of seedlings 1 year after fire in a Mediterranean pine forest. Oecologia 91:365-370.

Ne'eman, G., S. Lev-Yadun, and M. Arianoutsou. 2012. Fire-related traits in Mediterranean basin plants. Israel Journal of Ecology & Evolution 58:177- 194.

Onaindia, M., and I. Amezaga. 2000. Seasonal variation in the seed banks of native woodland and coniferous plantations in Northern Spain. Forest Ecology and Management 126:163-172.

Oron, T. 2017. Trends in the population size of the Wild Peony in the upper Kziv stream in the period 1973-2017. The findings of new surveys. Kalanit 4.

Pakeman, R., and J. Small. 2005. The role of the seed bank, seed rain and the timing

72 of disturbance in gap regeneration. Journal of Vegetation Science 16:121- 130.

Passalacqua, N., and L. Bernardo. 2004. The genus Paeonia L. in Italy: taxonomic survey and revision. Webbia 59:215-268.

Pausas, J., J. Keeley, and M. Verdú. 2006. Inferring differential evolutionary processes of plant persistence traits in Northern Hemisphere Mediterranean fire‐prone ecosystems. Journal of Ecology 94:31-39.

Pausas, J. G. 2004. Changes in fire and climate in the eastern Iberian Peninsula (Mediterranean basin). Climatic change 63:337-350.

Pausas, J. G. 2001. Resprouting vs seeding–a Mediterranean perspective. Oikos, 94(1), 193-194.

Pausas, J. G., and J. E. Keeley. 2009. A burning story: the role of fire in the history of life. BioScience 59:593-601.

Pausas, J. G., & Keeley, J. E. (2014). Evolutionary ecology of resprouting and seeding in fire‐prone ecosystems. New Phytologist, 204(1), 55-65.

Pausas, J. G., J. Llovet, A. Rodrigo, and R. Vallejo. 2009. Are wildfires a disaster in the Mediterranean basin?–A review. International Journal of Wildland Fire 17:713-723.

Peart, D. R. 1989. Species Interactions in a Successional Grassland .3. Effects of Canopy Gaps, Gopher Mounds and Grazing on Colonization. Journal of Ecology, 77, 267-289.

Perevolotsky, A. 2005. Integrating landscape ecology in the conservation of Mediterranean ecosystems: the Israeli experience. Israel Journal of Plant Sciences 53:203-213.

Perevolotsky, A., and N. a. G. Seligman. 1998. Role of grazing in Mediterranean rangeland ecosystems. Bioscience 48:1007-1017.

Pérez-Fernández, M., and S. Rodríguez-Echeverría. 2003. Effect of smoke, charred wood, and nitrogenous compounds on seed germination of ten species from woodland in central-western Spain. Journal of Chemical Ecology 29:237- 251.

Pierce, S., K. Esler, and R. Cowling. 1995. Smoke-induced germination of

73 succulents (Mesembryanthemaceae) from fire-prone and fire-free habitats in South Africa. Oecologia 102:520-522.

Poulson, T. L. & Platt, W. J. 1989. Gap Light Regimes Influence Canopy Tree Diversity. Ecology, 70, 553-555.

Pugnaire, F., and F. Valladares. 2007. Functional plant ecology. CRC Press.

Reiner, R.J. (2007) Fire in California grasslands. In California grasslands: ecology and management (eds M.R. Stromberg, J.D. Corbin & C.M. D’Antonio), pp. 207–217. University of California Press, Los Angeles, California.

Reyes, O., and L. Trabaud. 2009. Germination behaviour of 14 Mediterranean species in relation to fire factors: smoke and heat. Plant Ecology 202:113.

Rundel, P. W., G. Montenegro, and F. M. Jaksic. 2013. Landscape disturbance and biodiversity in Mediterranean-type ecosystems. Springer Science & Business Media.

Saaroni, H., B. Ziv, A. Bitan, and P. Alpert. 1998. Easterly wind storms over Israel.

Theoretical and Applied Climatology 59:61-77.

Saatkamp, A., P. Poschlod, and D. L. Venable. 2014. 11 The Functional Role of Soil Seed Banks in Natural Communities.

Santana, V. M., J. G. Alday, and M. J. Baeza. 2014. Effects of fire regime shift in Mediterranean Basin ecosystems: changes in soil seed bank composition among functional types. Plant Ecology 215:555-566.

Scheffer, M., S. Carpenter, J. A. Foley, C. Folke, and B. Walker. 2001. Catastrophic shifts in ecosystems. Nature 413:591.

Seidl, R., M. J. SCHELHAAS, and M. J. Lexer. 2011. Unraveling the drivers of intensifying forest disturbance regimes in Europe. Global Change Biology 17:2842-2852.

Shmida, A., and G. Pollak. 2007. Red data book: endangered plants of Israel, Vol. 1.

Israel Nature and Parks Authority, Jerusalem 495.

Singer, A. (2007). The soils of Israel. Springer Science & Business Media.

Sottile, G. D., P. E. Meretta, M. S. Tonello, M. M. Bianchi, and M. V. Mancini. 2015. Disturbance induced changes in species and functional diversity in southern Patagonian forest-steppe ecotone. Forest Ecology and Management

74 353:77- 86.

Sousa, W. P. 1984. The role of disturbance in natural communities. Annual review of ecology and systematics 15:353-391.

Sternberg, M., M. Gutman, A. Perevolotsky, E. D. Ungar, and J. Kigel. 2000. Vegetation response to grazing management in a Mediterranean herbaceous community: a functional group approach. Journal of Applied Ecology 37:224- 237.

Tessler, N. 2012. Documentation and analysis of wildfire regimes on Mount Carmel and the Jerusalem hills. Horizons in Geography:184-193.

Tessler, N., D. Malkinson, L. Wittenberg, and N. Greenbaum. 2010. Wildfires of the Mt. Carmel and Jerusalem Mountains: Documentation and Temporal Analysis. Horizons in Geography:157-165.

Tessler, N., Y. Sapir, L. Wittenberg, and N. Greenbaum. 2015. Recovery of Mediterranean vegetation after recurrent forest fires: insight from the 2010 forest fire on Mount Carmel, Israel. Land Degradation & Development.

Tessler, N., Y. Sapir, L. Wittenberg, and N. Greenbaum. 2016a. Recovery of Mediterranean vegetation after recurrent forest fires: insight from the 2010 forest fire on Mount Carmel, Israel. Land Degradation & Development 27:1424-1431.

Tessler, N., L. Wittenberg, and N. Greenbaum. 2016b. Vegetation cover and species richness after recurrent forest fires in the Eastern Mediterranean ecosystem of Mount Carmel, Israel. Science of the Total Environment 572:1395-1402.

Tessler, N., L. Wittenberg, E. Provizor, and N. Greenbaum. 2014. The influence of short-interval recurrent forest fires on the abundance of Aleppo pine (Pinus halepensis Mill.) on Mount Carmel, Israel. Forest ecology and management 324:109-116.

Thom, D., and R. Seidl. 2016. Natural disturbance impacts on ecosystem services and biodiversity in temperate and boreal forests. Biological Reviews 91:760- 781.

Thompson, J. D. 2005. Plant evolution in the Mediterranean. Oxford University Press on Demand.

Thompson, K. 2000. The functional ecology of soil seed banks. Seeds: the ecology

75 of regeneration in plant communities 2:215-235.

Tieu, A., J. A. Plummer, K. A. Dixon, K. Sivasithamparam, and I. M. Sieler. 1999. Germination of four species of native Western Australian plants using plant- derived smoke. Australian Journal of Botany 47:207-219.

Tormo, J., B. Moreira, and J. Pausas. 2014. Field evidence of smoke‐stimulated seedling emergence and establishment in M editerranean B asin flora. Journal of Vegetation Science 25:771-777.

Torre, I., and M. Díaz. 2004. Small mammal abundance in Mediterranean post-fire habitats: a role for predators? Acta Oecologica 25:137-142.

Traba, J., F. M. Azcárate, and B. Peco. 2004. From what depth do seeds emerge? A soil seed bank experiment with Mediterranean grassland species. Seed Science Research 14:297-303.

Trabaud, L., and J.-F. Galtié. 1996. Effects of fire frequency on plant communities and landscape pattern in the Massif des Aspres (southern France). Landscape Ecology 11:215-224.

Tsafrir, A., Y. Osem, H. Shemesh, Y. Carmel, C. Soref, and O. Ovadia. 2018. Fire season modifies the perennial plant community composition through a differential effect on obligate seeders in eastern Mediterranean woodlands. Applied Vegetation Science.

Valavanidis, A., and T. Vlachogianni. 2011. Ecosystems and biodiversity hotspots in the Mediterranean basin threats and conservation efforts. Sci Adv Environ Toxicol Ecotoxicol:1-24.

Valbuena, L., R. Tarrega, and E. Luis. 1992. Influence of heat on seed germination of Cistus laurifolius and Cistus ladanifer. International Journal of Wildland Fire 2:15-20.

Van Leeuwen, W. J., G. M. Casady, D. G. Neary, S. Bautista, J. A. Alloza, Y. Carmel, L. Wittenberg, D. Malkinson, and B. J. Orr. 2010. Monitoring post- wildfire vegetation response with remotely sensed time-series data in Spain, USA and Israel. International Journal of Wildland Fire 19:75-93. van Staden, J., Jager, A. K., Light, M. E., & Burger, B. V. 2004. Isolation of the major germination cue from plant-derived smoke. South African Journal of Botany, 70(4), 654-659

76 Verdú, M. 2000. Ecological and evolutionary differences between Mediterranean seeders and resprouters. Journal of Vegetation Science 11:265-268.

Verdú, M., J. G. Pausas, J. G. Segarra‐Moragues, and F. Ojeda. 2007. Burning phylogenies: fire, molecular evolutionary rates, and diversification. Evolution: International Journal of Organic Evolution 61:2195-2204.

Walck, J. L., J. M. Baskin, C. C. Baskin, and S. N. Hidayati. 2005. Defining transient and persistent seed banks in species with pronounced seasonal dormancy and germination patterns. Seed Science Research 15:189-196.

Walck, J. L., S. N. Hidayati, K. W. Dixon, K. Thompson, and P. Poschlod. 2011. Climate change and plant regeneration from seed. Global Change Biology 17:2145-2161.

Williams, B. K. 2011. Passive and active adaptive management: approaches and an example. Journal of environmental management 92:1371-1378.

Willis, C. G., Baskin, C. C., Baskin, J. M., Auld, J. R., Venable, D. L., Cavender‐ Bares, J., & NESCent Germination Working Group. 2014. The evolution of seed dormancy: environmental cues, evolutionary hubs, and diversification of the seed plants. New Phytologist, 203(1), 300-309.

Whitmore, T. 1989 Canopy gaps and the two major groups of forest trees. Ecology, 70, 536-538.

Wittenberg, L., H. Kutiel, N. Greenbaum, and M. Inbar. 2007. Short-term changes in the magnitude, frequency and temporal distribution of floods in the Eastern Mediterranean region during the last 45 years—Nahal Oren, Mt. Carmel, Israel. Geomorphology 84:181-191.

Yamamoto, S.-I. 2000. Forest gap dynamics and tree regeneration. Journal of Forest Research, 5, 223-229.

Zedler, P. H. 1995. Fire frequency in southern California shrublands: biological effects and management options. Brushfires in California wildlands: ecology and resource management. International Association of Wildland Fire, Fairfield, Washington, USA:101-112.

Zhang, S., K. Ma, and L. Chen. 2003. Response of photosynthetic plasticity of Paeonia suffruticosa to changed light environments. Environmental and Experimental Botany 49:121-133.

77 Zuloaga-Aguilar, S., Briones, O., & Orozco-Segovia, A. (2011). Seed germination of

montane forest species in response to ash, smoke and heat shock in Mexico. Acta Oecologica, 37(3), 256-262.

78 6. Supporting information

Figure S 1: Experiment 1:

Non-metric multidimensional scaling (nMDS) ordination, based on Bray-Curtis dissimilarity matrix, examining for differences in the composition of the germinable soil seed bank (GSSB) community among the unburned control (circles), autumn burns (squares) and spring burns

(triangles).

Figure S 2 : Experiment 1: Venn diagram of shared species among the three fire treatment groups.

80 Table S 1: Experiment 1: Summary of the pairwise comparison of the germinable soil seed bank density (per 1 m2 of soil area) between the different fire treatment groups and microhabitats. Treatment Contrast Std. Err. z P>z Autumn fire × Unburned -61.858 22.696 -2.73 0.006 Spring fire × Unburned -54.915 22.866 -2.4 0.016 Spring fire × Autumn fire 6.942 22.945 0.3 0.762

Treatment × Microhabitat Unburned × Cistus Unburned × Open -16.679 24.771 -0.67 0.501 Unburned × Pistacia Unburned × Open -40.350 24.935 -1.62 0.106 Autumn fire × Open Unburned × Open -33.743 30.450 -1.11 0.268 Autumn fire × Cistus Unburned × Open -60.278 30.403 -1.98 0.047 Autumn fire × Pistacia Unburned × Open -148.582 30.943 -4.8 <0.001 Spring fire × Open Unburned × Open -75.549 30.849 -2.45 0.014 Spring fire × Cistus Unburned × Open -55.186 30.785 -1.79 0.073 Spring fire × Pistacia Unburned × Open -91.040 30.816 -2.95 0.003 Unburned × Pistacia Unburned × Cistus -23.672 24.596 -0.96 0.336 Autumn fire × Open Unburned × Cistus -17.064 30.390 -0.56 0.574 Autumn fire × Cistus Unburned × Cistus -43.599 30.359 -1.44 0.151 Autumn fire × Pistacia Unburned × Cistus -131.903 30.876 -4.27 <0.001 Spring fire × Open Unburned × Cistus -58.870 30.783 -1.91 0.056 Spring fire × Cistus Unburned × Cistus -38.507 30.713 -1.25 0.21 Spring fire × Pistacia Unburned × Cistus -74.362 30.770 -2.42 0.016 Autumn fire × Open Unburned × Pistacia 6.608 30.534 0.22 0.829 Autumn fire × Cistus Unburned × Pistacia -19.927 30.543 -0.65 0.514 Autumn fire × Pistacia Unburned × Pistacia -108.231 31.073 -3.48 <0.001 Spring fire × Open Unburned × Pistacia -35.199 30.940 -1.14 0.255

Spring fire × Cistus Unburned × Pistacia -14.835 30.898 -0.48 0.631 Spring fire × Pistacia Unburned × Pistacia -50.690 30.909 -1.64 0.101 Autumn fire × Cistus Autumn fire × Open -26.535 25.226 -1.05 0.293 Autumn fire × Pistacia Autumn fire × Open -114.839 25.774 -4.46 <0.001 Spring fire × Open Autumn fire × Open -41.806 30.984 -1.35 0.177 Spring fire × Cistus Autumn fire × Open -21.443 30.92275 -0.69 0.488 Spring fire × Pistacia Autumn fire × Open -57.298 30.958 -1.85 0.064 Autumn fire × Pistacia Autumn fire × Cistus -88.304 25.716 -3.43 0.001 Spring fire × Open Autumn fire × Cistus -15.271 31.004 -0.49 0.622 Spring fire × Cistus Autumn fire × Cistus 5.092 30.979 0.16 0.869 Spring fire × Pistacia Autumn fire × Cistus -30.763 30.826 -1 0.318

Spring fire × Open Autumn fire × Pistacia 73.033 31.434 2.32 0.02 Spring fire × Cistus Autumn fire × Pistacia 93.396 31.329 2.98 0.003 Spring fire × Pistacia Autumn fire × Pistacia 57.541 31.511 1.83 0.068 Spring fire × Cistus Spring fire × Open 20.363 25.717 0.79 0.428 Spring fire × Pistacia Spring fire × Open -15.491 26.046 -0.59 0.552 Spring fire × Pistacia Spring fire × Cistus -35.854 25.850 -1.39 0.165

Table S 2: Experiment 1: Summary of the pairwise comparison of the germinable soil seed bank species richness between the different fire treatment groups and microhabitats.

Treatment Contrast Std. Err. z P>z Autumn fire × Unburned -3.758 1.313 -2.860 0.004 Spring fire × Unburned -2.904 1.380 -2.100 0.035 Spring fire × Autumn fire 0.854 0.947 0.900 0.367

Treatment × Microhabitat Autumn fire × Cistus Unburned × Open -5.897 1.312 -4.490 <0.001 Autumn fire × Cistus Unburned × Cistus -5.739 1.289 -4.450 <0.001 Autumn fire × Cistus Unburned × Pistacia -5.549 1.261 -4.400 <0.001 Spring fire × Cistus Unburned × Open -3.688 1.444 -2.550 0.011 Autumn fire × Cistus Autumn fire × Open -3.623 0.730 -4.960 <0.001 Autumn fire × Pistacia Unburned × Open -3.609 1.446 -2.500 0.013 Spring fire × Cistus Unburned × Cistus -3.530 1.423 -2.480 0.013 Autumn fire × Pistacia Unburned × Cistus -3.451 1.425 -2.420 0.015 Spring fire × Cistus Unburned × Pistacia -3.341 1.398 -2.390 0.017 Autumn fire × Pistacia Unburned × Pistacia -3.261 1.400 -2.330 0.02 Spring fire × Pistacia Unburned × Open -2.930 1.504 -1.950 0.051 Spring fire × Pistacia Unburned × Pistacia -2.772 1.484 -1.870 0.062 Spring fire × Open Unburned × Open -2.600 1.532 -1.700 0.09 Spring fire × Pistacia Unburned × Pistacia -2.583 1.460 -1.770 0.077 Spring fire × Open Unburned × Cistus -2.442 1.512 -1.620 0.106 Autumn fire × Cistus Autumn fire × Pistacia -2.288 0.553 -4.140 <0.001 Autumn fire × Open Unburned × Open -2.274 1.555 -1.460 0.144 Spring fire × Open Unburned × Pistacia -2.253 1.489 -1.510 0.13 Autumn fire × Open Unburned × Cistus -2.116 1.536 -1.380 0.168 Autumn fire × Open Unburned × Pistacia -1.927 1.513 -1.270 0.203 Spring fire × Cistus Autumn fire × Open -1.414 1.154 -1.230 0.22 Autumn fire × Pistacia Autumn fire × Open -1.335 0.584 -2.290 0.022 Spring fire × Cistus Spring fire × Open -1.088 0.570 -1.910 0.056 Spring fire × Cistus Spring fire × Pistacia -0.758 0.547 -1.390 0.166 Spring fire × Pistacia Autumn fire × Open -0.656 1.22976 -0.530 0.594 Unburned × Pistacia Unburned × Open -0.347 0.689 -0.500 0.614 Spring fire × Pistacia Spring fire × Open -0.330 0.570 -0.580 0.563 Spring fire × Open Autumn fire × Open -0.326 1.264 -0.260 0.796 Unburned × Cistus Unburned × Open -0.158 0.692 -0.230 0.819 Spring fire × Cistus Autumn fire × Pistacia -0.080 1.000 -0.080 0.937 Unburned × Cistus Unburned × Pistacia 0.190 0.684 0.280 0.782 Spring fire × Pistacia Autumn fire × Pistacia 0.679 1.086 0.630 0.532 Spring fire × Open Autumn fire × Pistacia 1.009 1.125 0.900 0.37 Spring fire × Cistus Autumn fire × Cistus 2.209 0.790 2.800 0.005 Spring fire × Pistacia Autumn fire × Cistus 2.967 0.897 3.310 0.001 Spring fire × Open Autumn fire × Cistus 3.297 0.944 3.490 <0.001

Table S 3: Experiment 1: Summary of the pairwise comparison of the germinable soil seed bank diversity between the different fire treatment groups and microhabitats. Treatment Contrast Std. Err. z P>z Autumn fire × Unburned -3.162 0.671 -4.710 <0.001 Spring fire × Unburned -2.130 0.678 -3.140 0.002 Spring fire × Autumn fire 1.032 0.678 1.520 0.128

Treatment × Microhabitat Unburned × Pistacia Unburned × Open -0.734 0.717 -1.020 0.306 Unburned × Cistus Unburned × Open -0.016 0.717 -0.020 0.982 Autumn fire × Open Unburned × Open -2.376 0.890 -2.670 0.008 Autumn fire × Pistacia Unburned × Open -3.070 0.890 -3.450 0.001 Autumn fire × Cistus Unburned × Open -4.790 0.890 -5.380 <0.001 Spring fire × Open Unburned × Open -1.587 0.902 -1.760 0.078 Spring fire × Pistacia Unburned × Open -2.630 0.902 -2.920 0.004 Spring fire × Cistus Unburned × Open -2.924 0.902 -3.240 0.001 Unburned × Cistus Unburned × Pistacia 0.718 0.717 1.000 0.317 Autumn fire × Open Unburned × Pistacia -1.643 0.890 -1.850 0.065 Autumn fire × Pistacia Unburned × Pistacia -2.336 0.890 -2.620 0.009 Autumn fire × Cistus Unburned × Pistacia -4.057 0.890 -4.560 <0.001 Spring fire × Open Unburned × Pistacia -0.853 0.902 -0.950 0.344 Spring fire × Pistacia Unburned × Pistacia -1.896 0.902 -2.100 0.035 Spring fire × Cistus Unburned × Pistacia -2.190 0.902 -2.430 0.015 Autumn fire × Open Unburned × Cistus -2.360 0.890 -2.650 0.008 Autumn fire × Pistacia Unburned × Cistus -3.053 0.890 -3.430 0.001 Autumn fire × Cistus Unburned × Cistus -4.774 0.890 -5.360 <0.001 Spring fire × Open Unburned × Cistus -1.571 0.902 -1.740 0.081 Spring fire × Pistacia Unburned × Cistus -2.614 0.902 -2.900 0.004 Spring fire × Cistus Unburned × Cistus -2.907 0.902 -3.220 0.001 Autumn fire × Pistacia Autumn fire × Open -0.693 0.717 -0.970 0.333 Autumn fire × Cistus Autumn fire × Open -2.414 0.717 -3.370 0.001 Spring fire × Open Autumn fire × Open 0.789 0.902 0.880 0.381 Spring fire × Pistacia Autumn fire × Open -0.254 0.902 -0.280 0.778 Spring fire × Cistus Autumn fire × Open -0.547 0.90277 -0.610 0.544 Autumn fire × Pistacia Autumn fire × Cistus -1.721 0.717 -2.400 0.016 Spring fire × Open Autumn fire × Pistacia 1.483 0.902 1.640 0.1 Unburned × Pistacia Autumn fire × Pistacia 0.440 0.902 0.490 0.626 Spring fire × Cistus Autumn fire × Pistacia 0.146 0.902 0.160 0.871 Spring fire × Open Autumn fire × Cistus 3.203 0.902 3.550 <0.001 Spring fire × Pistacia Autumn fire × Cistus 2.160 0.902 2.400 0.017 Spring fire × Cistus Autumn fire × Cistus 1.867 0.902 2.070 0.038 Spring fire × Pistacia Spring fire × Open -1.043 0.740 -1.410 0.159 Spring fire × Cistus Spring fire × Open -1.337 0.740 -1.810 0.071 Spring fire × Cistus Spring fire × Pistacia -0.294 0.740 -0.400 0.691

Table S 4: Experiment 1: Summary of the pairwise comparison of germinable soil seed bank density of annuals between the different fire treatment groups and microhabitats. Treatment Contrast Std. Err. z P>z Autumn fire × Unburned -183.152 86.507 -2.120 0.034 Spring fire × Unburned -180.157 86.671 -2.080 0.038 Spring fire × Autumn fire 2.995 38.241 0.080 0.938

Treatment × Microhabitat Spring fire × Pistacia Unburned × Open -124.878 93.179 -1.340 0.180 Unburned × Cistus Unburned × Open -58.787 94.857 -0.620 0.535 Autumn fire × Open Unburned × Open -137.392 133.558 -1.030 0.304 Spring fire × Pistacia Unburned × Open -312.788 119.069 -2.630 0.009 Autumn fire × Cistus Unburned × Open -282.939 119.309 -2.370 0.018 Autumn fire × Open Unburned × Open -214.525 124.150 -1.730 0.084 Spring fire × Pistacia Unburned × Open -262.992 120.150 -2.190 0.029 Autumn fire × Cistus Unburned × Open -246.619 121.267 -2.030 0.042 Unburned × Cistus Unburned × Pistacia 66.092 75.115 0.880 0.379 Autumn fire × Open Unburned × Pistacia -12.514 97.666 -0.130 0.898 Autumn fire × Pistacia Unburned × Pistacia -187.910 73.348 -2.560 0.010 Autumn fire × Cistus Unburned × Pistacia -158.061 74.316 -2.130 0.033 Spring fire × Open Unburned × Pistacia -89.646 82.706 -1.080 0.278 Spring fire × Pistacia Unburned × Pistacia -138.113 75.991 -1.820 0.069 Spring fire × Cistus Unburned × Pistacia -121.741 77.829 -1.560 0.118 Autumn fire × Open Unburned × Cistus -78.606 115.197 -0.680 0.495 Autumn fire × Pistacia Unburned × Cistus -254.002 97.295 -2.610 0.009 Autumn fire × Cistus Unburned × Cistus -224.152 97.727 -2.290 0.022 Spring fire × Open Unburned × Cistus -155.738 103.834 -1.500 0.134 Spring fire × Pistacia Unburned × Cistus -204.205 98.842 -2.070 0.039 Spring fire × Cistus Unburned × Cistus -187.833 100.197 -1.870 0.061 Autumn fire × Pistacia Autumn fire × Open -175.396 66.315 -2.640 0.008 Autumn fire × Cistus Autumn fire × Open -145.547 60.809 -2.390 0.017 Spring fire × Open Autumn fire × Open -77.132 79.338 -0.970 0.331 Spring fire × Pistacia Autumn fire × Open -125.600 72.318 -1.740 0.082 78 Unburned × Cistus Autumn fire × Open -109.227 74.215 -1.470 0.141 Autumn fire × Cistus Autumn fire × Pistacia 29.850 14.455 2.070 0.039 Spring fire × Open Autumn fire × Pistacia 98.264 42.294 2.320 0.020 Spring fire × Pistacia Autumn fire × Pistacia 49.797 25.027 1.990 0.047 Spring fire × Cistus Autumn fire × Pistacia 66.169 30.546 2.170 0.030 Spring fire × Open Autumn fire × Cistus 68.414 44.564 1.540 0.125 Spring fire × Pistacia Autumn fire × Cistus 19.947 29.016 0.690 0.492 Spring fire × Cistus Autumn fire × Cistus 36.320 33.841 1.070 0.283 Spring fire × Pistacia Spring fire × Open -48.467 33.714 -1.440 0.151 Spring fire × Cistus Spring fire × Open -32.094 33.744 -0.950 0.342 Spring fire × Cistus Spring fire × Pistacia 16.373 24.758 0.660 0.508

Table S 5: Experiment 1: Summary of the pairwise comparison of germinable soil seed bank richness of annuals between the different fire treatment groups and microhabitats. Treatment Contrast Std. Err. z P>z Autumn fire × Unburned -3.641 1.418 -2.570 0.010 Spring fire × Unburned -2.997 1.477 -2.030 0.042 Spring fire × Autumn fire 0.643 0.824 0.780 0.435

Treatment × Microhabitat Unburned × Pistacia Unburned × Open -0.752 0.643 -1.170 0.242 Unburned × Cistus Unburned × Open -0.687 0.641 -1.070 0.284 Autumn fire × Open Unburned × Open -2.603 1.722 -1.510 0.131 Autumn fire × Pistacia Unburned × Open -5.464 1.510 -3.620 <0.001 Autumn fire × Cistus Unburned × Open -4.293 1.567 -2.740 0.006 Spring fire × Open Unburned × Open -3.199 1.663 -1.920 0.054 Spring fire × Pistacia Unburned × Open -3.872 1.602 -2.420 0.016 Spring fire × Cistus Unburned × Open -3.359 1.647 -2.040 0.041 Unburned × Cistus Unburned × Pistacia 0.065 0.601 0.110 0.913 Autumn fire × Open Unburned × Pistacia -1.851 1.578 -1.170 0.241 Autumn fire × Pistacia Unburned × Pistacia -4.712 1.340 -3.520 <0.001 Autumn fire × Cistus Unburned × Pistacia -3.542 1.406 -2.520 0.012 Spring fire × Open Unburned × Pistacia -2.447 1.513 -1.620 0.106 Spring fire × Pistacia Unburned × Pistacia -3.120 1.444 -2.160 0.031 Spring fire × Cistus Unburned × Pistacia -2.608 1.495 -1.740 0.081 Autumn fire × Open Unburned × Cistus -1.916 1.590 -1.200 0.228 Autumn fire × Pistacia Unburned × Cistus -4.777 1.355 -3.530 <0.001 Autumn fire × Cistus Unburned × Cistus -3.607 1.420 -2.540 0.011 Spring fire × Open Unburned × Cistus -2.513 1.525 -1.650 0.100 Spring fire × Pistacia Unburned × Cistus -3.186 1.458 -2.190 0.029 Spring fire × Cistus Unburned × Cistus -2.673 1.508 -1.770 0.076 Autumn fire × Pistacia Autumn fire × Open -2.861 0.757 -3.780 <0.001 Autumn fire × Cistus Autumn fire × Open -1.691 0.576 -2.930 0.003 Spring fire × Open Autumn fire × Open -0.596 1.171 -0.510 0.611 Spring fire × Pistacia Autumn fire × Open -1.269 1.078 -1.180 0.239 Spring fire × Cistus Autumn fire × Open -0.757 791.148 -0.660 0.510 Autumn fire × Cistus Autumn fire × Pistacia 1.170 0.401 2.920 0.004 Spring fire × Open Autumn fire × Pistacia 2.265 0.801 2.830 0.005 Spring fire × Pistacia Autumn fire × Pistacia 1.592 0.654 2.440 0.015 Spring fire × Cistus Autumn fire × Pistacia 2.104 0.765 2.750 0.006 Spring fire × Open Autumn fire × Cistus 1.094 0.914 1.200 0.231 Spring fire × Pistacia Autumn fire × Cistus 0.421 0.790 0.530 0.594 Spring fire × Cistus Autumn fire × Cistus 0.934 0.884 1.060 0.291 Spring fire × Pistacia Spring fire × Open -0.673 0.445 -1.510 0.130 Spring fire × Cistus Spring fire × Open -0.160 0.438 -0.370 0.714 Spring fire × Cistus Spring fire × Pistacia 0.513 0.427 1.200 0.230

Table S 6: Experiment 1: Summary of the pairwise comparison of germinable soil seed bank density of dwarf-shrubs between the different fire treatment groups and microhabitats. Treatment Contrast Std. Err. z P>z Autumn fire × Unburned 4.102 1.139 3.600 <0.001 Spring fire × Unburned 2.219 0.934 2.380 0.018 Spring fire × Autumn fire -1.883 1.298 -1.450 0.147

Treatment × Microhabitat Unburned × Pistacia Unburned × Open -1.007 0.440 -2.290 0.022 Unburned × Cistus Unburned × Open 0.976 0.505 1.930 0.053 Autumn fire × Open Unburned × Open 2.006 0.988 2.030 0.042 Autumn fire × Pistacia Unburned × Open 1.838 0.969 1.900 0.058 Autumn fire × Cistus Unburned × Open 8.429 1.796 4.690 <0.001 Spring fire × Open Unburned × Open -0.522 0.729 -0.720 0.474 Spring fire × Pistacia Unburned × Open 2.323 1.042 2.230 0.026 Spring fire × Cistus Unburned × Open 4.824 1.358 3.550 <0.001 Unburned × Cistus Unburned × Pistacia 1.983 0.530 3.740 <0.001 Autumn fire × Open Unburned × Pistacia 3.014 0.922 3.270 0.001 Autumn fire × Pistacia Unburned × Pistacia 2.845 0.901 3.160 0.002 Autumn fire × Cistus Unburned × Pistacia 9.437 1.765 5.350 <0.001 Spring fire × Open Unburned × Pistacia 0.485 0.631 0.770 0.442 Spring fire × Pistacia Unburned × Pistacia 3.330 0.979 3.400 0.001 Spring fire × Cistus Unburned × Pistacia 5.831 1.313 4.440 <0.001 Autumn fire × Open Unburned × Cistus 1.031 1.066 0.970 0.334 Autumn fire × Pistacia Unburned × Cistus 0.862 1.048 0.820 0.411 Autumn fire × Cistus Unburned × Cistus 7.454 1.835 4.060 <0.001 Spring fire × Open Unburned × Cistus -1.498 0.836 -1.790 0.073 Spring fire × Pistacia Unburned × Cistus 1.347 1.116 1.210 0.228 Spring fire × Cistus Unburned × Cistus 3.848 1.414 2.720 0.007 Autumn fire × Pistacia Autumn fire × Open -0.168 0.592 -0.280 0.776 Autumn fire × Cistus Autumn fire × Open 6.423 1.167 5.510 <0.001 Spring fire × Open Autumn fire × Open -2.528 0.956 -2.650 0.008 Spring fire × Pistacia Autumn fire × Open 0.317 1.203 0.260 0.792 Spring fire × Cistus Autumn fire × Open 2.818 80 1.480 1.900 0.057 Autumn fire × Cistus Autumn fire × Pistacia 6.591 1.182 5.580 <0.001 Spring fire × Open Autumn fire × Pistacia -2.360 0.936 -2.520 0.012 Spring fire × Pistacia Autumn fire × Pistacia 0.485 1.188 0.410 0.683 Spring fire × Cistus Autumn fire × Pistacia 2.986 1.468 2.030 0.042 Spring fire × Open Autumn fire × Cistus -8.952 1.781 -5.030 <0.001 Spring fire × Pistacia Autumn fire × Cistus -6.107 1.913 -3.190 0.001 Spring fire × Cistus Autumn fire × Cistus -3.605 2.088 -1.730 0.084 Spring fire × Pistacia Spring fire × Open 2.845 0.653 4.360 <0.001 Spring fire × Cistus Spring fire × Open 5.346 0.961 5.570 <0.001 Spring fire × Cistus Spring fire × Pistacia 2.501 0.746 3.350 0.001

Table S 7: Experiment 1: Summary of the pairwise comparison of germinable soil seed bank richness of dwarf-shrubs between the different fire treatment groups and microhabitats.

Treatment Contrast Std. Err. z P>z Autumn fire × Unburned 0.229 0.163 1.410 0.160 Spring fire × Unburned 0.271 0.165 1.640 0.100 Spring fire × Autumn fire 0.042 0.165 0.250 0.800

Treatment × Microhabitat Unburned × Pistacia Unburned × Open 0.438 0.221 1.980 0.048 Unburned × Cistus Unburned × Open 0.719 0.221 3.260 0.001 Autumn fire × Open Unburned × Open 0.625 0.243 2.570 0.010 Autumn fire × Pistacia Unburned × Open 0.094 0.243 0.390 0.700 Autumn fire × Cistus Unburned × Open 1.125 0.243 4.630 <0.001 Spring fire × Open Unburned × Open 0.656 0.247 2.660 0.008 Spring fire × Pistacia Unburned × Open 0.556 0.247 2.260 0.024 Spring fire × Cistus Unburned × Open 0.756 0.247 3.070 0.002 Unburned × Cistus Unburned × Pistacia 0.281 0.221 1.270 0.203 Autumn fire × Open Unburned × Pistacia 0.188 0.243 0.770 0.440 Autumn fire × Pistacia Unburned × Pistacia -0.344 0.243 -1.410 0.157 Autumn fire × Cistus Unburned × Pistacia 0.688 0.243 2.830 0.005 Spring fire × Open Unburned × Pistacia 0.219 0.247 0.890 0.375 Spring fire × Pistacia Unburned × Pistacia 0.119 0.247 0.480 0.630 Spring fire × Cistus Unburned × Pistacia 0.319 0.247 1.290 0.196 Autumn fire × Open Unburned × Cistus -0.094 0.243 -0.390 0.700 Autumn fire × Pistacia Unburned × Cistus -0.625 0.243 -2.570 0.010 Autumn fire × Cistus Unburned × Cistus 0.406 0.243 1.670 0.095 Spring fire × Open Unburned × Cistus -0.062 0.247 -0.250 0.800 Spring fire × Pistacia Unburned × Cistus -0.162 0.247 -0.660 0.510 Spring fire × Cistus Unburned × Cistus 0.038 0.247 0.150 0.879 Autumn fire × Pistacia Autumn fire × Open -0.531 0.221 -2.410 0.016 Autumn fire × Cistus Autumn fire × Open 0.500 0.221 2.260 0.024 Spring fire × Open Autumn fire × Open 0.031 0.247 0.130 0.899 Spring fire × Pistacia Autumn fire × Open -0.069 0.24781 -0.280 0.781 Spring fire × Cistus Autumn fire × Open 0.131 0.247 0.530 0.594 Autumn fire × Cistus Autumn fire × Pistacia 1.031 0.221 4.670 <0.001 Spring fire × Open Autumn fire × Pistacia 0.563 0.247 2.280 0.023 Spring fire × Pistacia Autumn fire × Pistacia 0.463 0.247 1.880 0.061 Spring fire × Cistus Autumn fire × Pistacia 0.663 0.247 2.690 0.007 Spring fire × Open Autumn fire × Cistus -0.469 0.247 -1.900 0.057 Spring fire × Pistacia Autumn fire × Cistus -0.569 0.247 -2.310 0.021 Spring fire × Cistus Autumn fire × Cistus -0.369 0.247 -1.500 0.135 Spring fire × Pistacia Spring fire × Open -0.100 0.228 -0.440 0.661 Spring fire × Cistus Spring fire × Open 0.100 0.228 0.440 0.661 Spring fire × Cistus Spring fire × Pistacia 0.200 0.228 0.880 0.380

Table S 8 : Experiment 1: Summary of the generalized linear mixed models testing the effect of fire treatment and microhabitat type on annual and dwarf- shrub germinable soil seed bank density and species richness. Annual density Annual species richness Dwarf shrubs density Dwarf shrub species richness Est SE z P Estimat SE z P Estimat SE z P Estimates SE z P i valu value es valu val es valu value valu va mat e e u e e e l es ue Treatment Autumn fire - 0.475 - 0.19 -0.445 0.260 - 0.08 0.695 0.225 3.100 0.00 0.425 0.312 1.360 0.1 5 2 0.61 1.300 1.720 6 73 6 Spring fire - 0.477 - 0.03 -0.523 0.261 - 0.04 -0.014 0.236 -0.060 0.95 0.163 0.330 0.490 0.6 6 3 1.00 2.100 2.000 5 22 1 Microhabitat Pistacia - 0.254 -1.82 0.06 -0.016 0.128 - 0.89 -0.305 0.154 -1.970 0.04 0.211 0.326 0.650 0.5 9 8 0.46 0.130 8 17 2 Cistus - 0.251 -0.83 0.40 -0.085 0.130 - 0.51 0.395 0.130 3.040 0.00 0.211 0.326 0.650 0.5 8 2 0.20 0.650 5 17 8 Treatment × Microhabitat Autumn fire × Pistacia - 0.376 - <0.0 -0.887 0.246 - <0.0 0.099 0.187 0.530 0.59 -0.984 0.478 - 0.0 01 01 7 1.38 3.680 3.610 <0.0 2.060 39 01 3 Autumn fire × Cistus - 0.365 - 0.00 -0.500 0.229 - 0.02 0.241 0.157 1.540 0.12 -0.251 0.430 - 0.5 4 4 1.04 2.850 2.190 9 0.580 60 2 Spring fire × Pistacia - 0.370 - 0.87 -0.010 0.207 - 0.96 0.947 0.198 4.790 <0.0 -0.072 0.447 - 0.8 3 01 0.05 0.160 0.050 2 0.160 73 9 Spring fire × Cistus - 0.365 - 0.68 -0.069 0.212 - 0.74 0.627 0.175 3.590 <0.0 -0.163 0.452 - 0.7 5 01 82

0.14 0.410 0.330 4 0.360 19 8

Constant term - 0.333 - <0.0 3.400 0.177 19.20 0.00 3.184 0.167 19.11 <0.00 1.447 0.243 5.970 0.0 01 1 2.14 6.440 0 0 0 00 5 Log of the dispersion 0.02 0.121 0.230 0.820 parameter/ln(alpha) 8 Plot

Variance 0.25 0.152 0.126 0.074 5.18E- 2.09E- 3.73E-28 3.26E 3 32 17 -14 Subplot(Plot)

Variance 0.41 0.152 0.266 0.061 0.37281 0.06817 2.19E-34 1.54E 5 9 5 -18

Table S 9 : Experiment 1: Germinable soil seed bank (germinable soil seed bank) density of common annual and dwarf-shrub species in the three fire treatment groups. The dwarf-shrubs and annuals included in this table comprised 30.3% and 44.9% of the total germination density, respectively.

Unburned control Autumn fire Spring fire P value

Dwarf-shrubs

Calicotome villosa 0 0.958 ± 0.96 1.917± 1.92 0.596

Fumana arabica 102.542± 36.43 310.5± 70.96 174.417± 48.37 0.012

Cistus salviifolius 75.708± 28.67 274.083± 76.93 238.625± 74.44 0.006

Micromeria sp 101.583± 31.42 41.208± 16.77 24.917± 10.95 0.202

Sarcopoterium spinosum 31.625± 10.97 27.792± 8.89 51.75± 17.13 0.741

Satureja thymbra 3.833± 2.16 5.75± 3.87 26.833± 11.44 0.293

Teucrium capitatum 2.875± 1.50 5.75± 3.32 2.875± 2.06 0.830

Teucrium divaricatum 52.708± 14.46 62.292± 27.86 25.875± 6.95 0.456

Annuals

Anagallis arvensis 47.917± 20.36 36.417± 16.19 76.67± 22.44 0.568

Anthemis pseudocotula 29.70± 12.297 15.33± 14.32 24.97± 15.69 0.470

Avena sterilis 287.5± 119.15 115± 105.68 20.125± 11.25 0.041

Catapodium rigidum 81.46± 41.21 42.167± 20.96 32.58± 12.73 0.852

Erophila minima 227.125± 85.76 53.67± 35.029 4.79± 2.99 0.035

Filago pyramidata 115± 27.12 121.71± 55.01 90.08± 27.12 0.419

Minuartia hybrida 116.92± 38.76 5.75± 3.88 31.625± 16.63 0.076

Phagnalon rupestre 53.67± 19.37 19.17± 5.17 32.58± 8.35 0.177

Plantago afra 65.17± 16.22 23.96± 19.12 46± 24.23 0.020

Plantago cretica 45.04± 18.37 0 10.54± 8.66 0.009

Sedum rubens 284.625± 94.3 45.04± 21.58 50.79± 22.45 0.179

Theligonum cynocrambe 42.17± 24.81 7.67± 4.09 20.125± 8.15 0.412

Torilis tenella 44.08± 22.59 8.62± 6.8 97.75± 86.4 0.346

Valantia hispida 29.71± 11 13.42± 13.417 4.79± 2.99 0.080

Table S 10 : Experiment 1: Pairwise comparisons of the germinable soil seed bank community composition between the different fire treatment groups that followed the permutational MANOVA (PERMANOVA) done on the Bray–Curtis dissimilarity matrix.

Groups t P(perm) perms Unburned control, Autumn fire 2.001 0.0264 35 Unburned control, Spring fire 1.5458 0.0549 35 Autumn fire, Spring fire 0.92144 0.5147 35

Table S 11: Experiment 2: Pairwise comparison for the combined effect of fire frequency and smoke treatment on germinable soil seed bank density (per 1 m2 of soil area) and species richness

Density (m2) Species richness Fire Frequency Contrast Std. Err. z P>z Contrast Std. Err. z P>z 1 fire × Unburned 1890.53 944.48 2 0.045 0.93 1.07 0.87 1 3 fires × Unburned 3353.01 1213.61 2.76 0.006 1.78 1.15 1.55 0.36 3 fires × 1 fire 1462.48 1387.13 1.05 0.292 0.85 1.22 0.7 1 Fire Frequency × Treatment Unburned × Smoke Unburned × Control 545.90 273.74 1.99 0.046 0.74 0.43 1.71 1 Unburned × Water&Smoke Unburned × Control 289.66 243.34 1.19 0.234 0.41 0.42 0.98 1 1 fire × Control Unburned × Control 2228.63 974.21 2.29 0.022 1.50 1.11 1.34 1 1 fire × Smoke Unburned × Control 3323.18 1186.01 2.8 0.005 2.02 1.17 1.73 1 1 fire × Water&Smoke Unburned × Control 955.35 741.02 1.29 0.197 0.43 1.01 0.43 1 3 fires × Control Unburned × Control 3406.14 1201.86 2.83 0.005 2.09 1.17 1.78 1 3 fires × Smoke Unburned × Control 5135.29 1549.51 3.31 0.001 3.02 1.27 2.38 0.63 3 fires × Water&Smoke Unburned × Control 2353.16 995.92 2.36 0.018 1.40 1.10 1.26 1 Unburned × Water&Smoke Unburned × Smoke -256.24 269.36 -0.95 0.341 -0.33 0.43 -0.76 1 1 fire × Control Unburned × Smoke 1682.73 1024.26 1.64 0.1 0.76 1.17 0.64 1 1 fire × Smoke Unburned × Smoke 2777.28 1226.16 2.27 0.024 1.28 1.22 1.05 1 1 fire × Water&Smoke Unburned × Smoke 409.45 807.94 0.51 0.612 -0.31 1.08 -0.29 1 3 fires × Control Unburned × Smoke 2860.25 1241.29 2.3 0.021 1.35 1.23 1.1 1 3 fires × Smoke Unburned × Smoke 4589.39 1578.63 2.91 0.004 2.28 1.32 1.72 1 3 fires × Water&Smoke Unburned × Smoke 1807.26 1044.87 1.73 0.084 0.66 1.16 0.56 1 1 fire × Control Unburned × Water&Smoke 1938.97 999.36 1.94 0.052 1.08 1.15 0.95 1 1 fire × Smoke Unburned × Water&Smoke 3033.51 1206.08 2.52 0.012 1.61 1.20 1.34 1 80

1 fire × Water&Smoke Unburned × Water&Smoke 665.69 775.00 0.86 0.39 0.02 1.05 0.02 1 3 fires × Control Unburned × Water&Smoke 3116.49 1221.56 2.55 0.011 1.68 1.20 1.39 1 3 fires × Smoke Unburned × Water&Smoke 4845.63 1563.97 3.1 0.002 2.61 1.30 2.01 1 3 fires × Water&Smoke Unburned × Water&Smoke 2063.50 1020.50 2.02 0.043 0.99 1.14 0.87 1 1 fire × Smoke 1 fire × Control 1094.54 540.60 2.02 0.043 0.52 0.48 1.08 1 1 fire × Water&Smoke 1 fire × Control -1273.28 463.26 -2.75 0.006 -1.06 0.47 -2.27 0.84 3 fires × Control 1 fire × Control 1177.51 1416.27 0.83 0.406 0.59 1.29 0.46 1 3 fires × Smoke 1 fire × Control 2906.66 1714.92 1.69 0.09 1.52 1.38 1.1 1 3 fires × Water&Smoke 1 fire × Control 124.53 1252.09 0.1 0.921 -0.10 1.23 -0.08 1 1 fire × Water&Smoke 1 fire × Smoke -2367.83 649.70 -3.64 <0.001 -1.59 0.50 -3.17 0.06 3 fires × Control 1 fire × Smoke 82.97 1563.13 0.05 0.958 0.07 1.34 0.05 1 3 fires × Smoke 1 fire × Smoke 1812.12 1835.38 0.99 0.323 1.00 1.42 0.7 1 3 fires × Water&Smoke 1 fire × Smoke -970.01 1418.42 -0.68 0.494 -0.62 1.28 -0.49 1 3 fires × Control 1 fire × Water&Smoke 2450.79 1276.22 1.92 0.055 1.65 1.21 1.37 1 3 fires × Smoke 1 fire × Water&Smoke 4179.94 1604.74 2.6 0.009 2.58 1.30 1.99 1 3 fires × Water&Smoke 1 fire × Water&Smoke 1397.81 1087.65 1.29 0.199 0.96 1.14 0.85 1 3 fires × Smoke 3 fires × Control 1729.15 722.04 2.39 0.017 0.93 0.52 1.79 1 3 fires × Water&Smoke 3 fires × Control -1052.98 540.23 -1.95 0.051 -0.69 0.49 -1.42 1 3 fires × Water&Smoke 3 fires × Smoke -2782.13 816.06 -3.41 0.001 -1.62 0.53 -3.04 0.09

Table S 12: Experiment 2: Summary of the generalized linear mixed models testing the effect of fire frequency and smoke treatment on annual and dwarf shrub germinable soil seed bank density, richness and diversity (calculated by Fisher's alpha).

Annuals Dwarf shrubs Density Richness Density Richness z P z P z P z Estimates SE value value Estimates SE value value Estimates SE value value Estimates SE value P value Fire frequency One Fire 1.058 0.422 2.51 0.012 0.294 0.210 1.40 0.163 -0.022 0.618 -0.04 0.971 -0.213 0.486 -0.44 0.66 Three Fires 1.139 0.421 2.71 0.007 0.336 0.210 1.60 0.110 0.896 0.596 1.50 0.133 0.610 0.448 1.360 0.173 Smoke Treatment Smoke 0.243 0.052 4.69 <0.00 0.150 0.084 1.79 0.074 -0.028 0.360 -0.08 0.938 0.182 0.303 0.60 0.547 1 Water&Smoke 0.224 0.052 4.30 <0.00 0.121 0.085 1.43 0.154 -0.308 0.376 -0.82 0.414 -0.051 0.320 -0.16 0.873 1 Fire frequency× Treatment One Fire × Smoke 0.116 0.063 1.84 0.066 -0.053 0.112 -0.47 0.636 -0.293 0.533 -0.55 0.582 -0.247 0.470 -0.53 0.599 One Fire × Water&Smoke -0.489 0.067 -7.31 <0.00 -0.306 0.117 -2.63 0.009 -0.715 0.575 -1.24 0.213 -0.323 0.506 -0.64 0.523 1 Three fires × Smoke 0.060 0.060 0.98 0.325 -0.047 0.111 -0.42 0.671 0.158 0.470 0.34 0.736 -0.020 0.382 -0.05 0.959 Three fires × Water&Smoke -0.334 0.062 -5.35 <0.00 -0.252 0.114 -2.20 0.028 0.140 0.488 0.29 0.774 0.021 0.403 0.05 0.958 1 Constant term 1.214 0.297 4.09 <0.00 1.343 0.151 8.910 <0.00 -0.878 0.442 -1.99 0.047 -1.215 0.335 -3.63 <0.001 1 1 Plot Variance 1.214 0.297 0.045 0.028 0.278 0.199 0.176 0.113 Plot(Goto and Watanabe) Variance 0.218 0.129 0.210 0.033 1.445 0.310 <0.001 <0.00 1

Table S 13: Experiment 2: Pairwise comparison for the combined effect of fire frequency and smoke treatment on germinable soil seed bank density (per 1 m2 of soil area) and species richness of annual species

Annual Density (m2) Annual Species richness Fire Frequency Contrast Std. Err. z P>z Contrast Std. Err. z P>z 1 fire × Unburned 1819.68 904.48 2.01 0.044 0.94 1.05 0.89 0.372 3 fires × Unburned 3010.11 1134.56 2.65 0.008 1.29 1.09 1.19 0.236 3 fires × 1 fire 1190.43 1310.38 0.91 0.364 0.35 1.17 0.3 0.763 Fire Frequency × Treatment Unburned × Smoke Unburned × Control 577.25 256.25 2.25 0.024 0.71 0.41 1.74 0.082 Unburned × Water&Smoke Unburned × Control 519.90 248.78 2.09 0.037 0.56 0.40 1.4 0.161 1 fire × Control Unburned × Control 2180.42 912.02 2.39 0.017 1.49 1.08 1.38 0.169 1 fire × Smoke Unburned × Control 3346.09 1151.66 2.91 0.004 2.08 1.15 1.82 0.069 1 fire × Water&Smoke Unburned × Control 1029.68 686.75 1.5 0.134 0.50 0.98 0.51 0.609 3 fires × Control Unburned × Control 3114.02 1103.42 2.82 0.005 1.74 1.11 1.57 0.117 3 fires × Smoke Unburned × Control 4839.74 1469.14 3.29 0.001 2.40 1.18 2.03 0.042 3 fires × Water&Smoke Unburned × Control 2173.71 909.30 2.39 0.017 0.99 1.03 0.96 0.335 Unburned × Water&Smoke Unburned × Smoke -57.35 252.82 -0.23 0.821 -0.15 0.41 -0.36 0.715 1 fire × Control Unburned × Smoke 1603.17 963.87 1.66 0.096 0.78 1.14 0.68 0.494 1 fire × Smoke Unburned × Smoke 2768.84 1191.45 2.32 0.02 1.38 1.21 1.14 0.253 1 fire × Water&Smoke Unburned × Smoke 452.43 757.27 0.6 0.55 -0.21 1.04 -0.2 0.843 3 fires × Control Unburned × Smoke 2536.77 1145.22 2.22 0.027 1.03 1.17 0.88 0.377 3 fires × Smoke Unburned × Smoke 4262.50 1498.55 2.84 0.004 1.69 1.24 1.37 0.171 3 fires × Water&Smoke Unburned × Smoke 1596.46 961.65 1.66 0.097 0.28 1.09 0.26 0.795 1 fire × Control Unburned × Water&Smoke 1660.52 958.24 1.73 0.083 0.93 1.13 0.82 0.41 1 fire × Smoke Unburned × Water&Smoke 2826.19 1187.09 2.38 0.017 1.53 1.19 1.28 0.201 1 fire × Water&Smoke Unburned × Water&Smoke 509.78 749.74 0.68 0.497 -0.06 1.03 -0.06 0.955 3 fires × Control Unburned × Water&Smoke 2594.12 1140.65 2.27 0.023 1.18 1.16 1.02 0.307 3 fires × Smoke Unburned × Water&Smoke 4319.84 1495.32 2.89 0.004 1.84 1.23 1.5 0.133 3 fires × Water&Smoke Unburned × Water&Smoke 1653.80 955.95 1.73 0.084 0.43 1.08 0.4 0.689 1 fire × Smoke 1 fire × Control 1165.67 529.15 2.2 0.028 0.60 0.46 1.28 0.2 1 fire × Water&Smoke 1 fire × Control -1150.74 434.41 -2.65 0.008 -0.99 0.45 -2.2 0.027 83

3 fires × Control 1 fire × Control 933.60 1319.54 0.71 0.479 0.25 1.24 0.2 0.84 3 fires × Smoke 1 fire × Control 2659.33 1629.57 1.63 0.103 0.91 1.31 0.7 0.485 3 fires × Water&Smoke 1 fire × Control -6.71 1168.74 -0.01 0.995 -0.50 1.17 -0.43 0.67 1 fire × Water&Smoke 1 fire × Smoke -2316.41 646.87 -3.58 0 -1.58 0.49 -3.23 0.001 3 fires × Control 1 fire × Smoke -232.07 1487.95 -0.16 0.876 -0.34 1.30 -0.27 0.79 3 fires × Smoke 1 fire × Smoke 1493.65 1765.27 0.85 0.397 0.32 1.36 0.23 0.816 3 fires × Water&Smoke 1 fire × Smoke -1172.38 1358.56 -0.86 0.388 -1.09 1.23 -0.89 0.373 3 fires × Control 1 fire × Water&Smoke 2084.34 1185.59 1.76 0.079 1.24 1.15 1.08 0.281 3 fires × Smoke 1 fire × Water&Smoke 3810.06 1527.59 2.49 0.013 1.90 1.22 1.56 0.12 3 fires × Water&Smoke 1 fire × Water&Smoke 1144.03 1011.19 1.13 0.258 0.49 1.07 0.46 0.648 3 fires × Smoke 3 fires × Control 1725.72 695.09 2.48 0.013 0.66 0.47 1.4 0.162 3 fires × Water&Smoke 3 fires × Control -940.31 499.00 -1.88 0.06 -0.75 0.45 -1.66 0.096 3 fires × Water&Smoke 3 fires × Smoke -2666.04 791.85 -3.37 0.001 -1.41 0.49 -2.87 0.004

Table S 14 :Experiment 2: Mean germinable soil seed bank density in the fire frequency sites and smoke treatments for the abundant species.

Unburned 3 fires 1 fire P-value

Fire Fire Smoke hist history Contr Smoke Smoke + Smoke treatme Species Family Smoke Control Smoke Control Smoke ory ×Smoke ol + Water Water + Water nt effe treatment effect ct effect 4.33 ± 1.67 ± 5.33 ± 4.67 ± 1.25 ± 4.33 ± 0.5 ± Crepis sp. Asteraceae 4 ± 1.77 5 ± 0.88 0.95 0.30 2.22 0.51 2.79 0.84 0.55 0.19 0.29 2.67 ± 3.33 ± 2.67 ± 5.25 ± 8.33 ± Torilis nodosa Apiaceae 4 ± 1.77 0 7 ± 3.76 0 0.72 0.57 0.78 0.84 1.39 1.54 1.78 4.53 7.5 ± 7.67 ± 7.67 ± Conyza sp. Asteraceae 6 ± 1.15 7 ± 0.58 6 ± 0.33 7 ± 0.66 5 ± 0.66 11 ± 0.14 0.61 0.60 0.38 0.51 1.02 3.33 ± 0.67 ± 0.75 ± 2.33 ± 1.33 ± Crepis sancta Asteraceae 1 ± 0.33 5 ± 1.53 0.19 0.39 1 ± 0.33 0.28 0.39 0.51 1 ± 0.10 0.43 0.23 0.33 ± 2.67 ± 21.67 ± 16.25 ± 5.5 ± Filago pyramidata Asteraceae 0 7 ± 1.33 0 4 ± 2.03 0.27 0.44 0.83 0.19 0.96 9.76 4.89 2.6 0.33 ± 1.33 ± 9.33 ± 6.25 ± 3.33 ± 9.33 ± 1.5 ± Mercurialis annua Asteraceae 2 ± 3.61 7 ± 1.33 0.21 0.02 0.13 0.33 2.03 3.84 5.21 2.00 1.50 2.25

5.33 ± 8.67 ± 5.67 ± 4.75 ± 4.67 ± 8.67 ± 1.5 ± Sonchus oleraceus Asteraceae 4 ± 0.33 7 ± 1.45 0.96 0.27 0.77 1.71 3.91 1.54 0.68 0.84 1.07 0.29

Brassicacea 0.33 ± 18.75 ± 27.33 ± 29.33 ± Cardamine hirsuta 3 ± 1.73 3 ± 1.15 3 ± 1.45 6 ± 3.18 0 0.45 0.72 0.49 e 0.19 7.76 15.21 16.93

Rapistrum rugosu Brassicacea 6.33 ± 4.33 ± 1.75 ± 7.67 ± 3 ± 0.66 5 ± 2.08 2 ± 0.66 11 ± 2.19 3 ± 1.73 0.54 0.25 0.47 m e 2.17 2.22 1.01 1.39

Brassicacea 0.33 ± 1.67 ± 1.33 ± 0.33 ± 0.25 ± 3.5 ± Sinapis alba 0 4 ± 1.2 2 ± 0.88 0.02 0.72 0.51 e 0.19 0.39 0.51 0.19 0.14 0.87

Caryophylla 6.33 ± 3.67 ± 4.33 ± 4.75 ± 5.67 ± 4.5 ± Silene sp. 5 ± 0.33 9 ± 2.19 7 ± 1.33 0.71 0.30 0.87 ceae 0.69 0.19 0.69 0.89 0.84 1.44 Stellaria cupanian Caryophylla 3.67 ± 7.33 ± 1.33 ± 0.67 ± 2.33 ± 0.67 ± 0 5 ± 2.89 0 0.71 0.98 0.28 a ceae 2.12 2.91 0.77 0.39 1.35 0.39 11.33 ± 16.33 ± 10.67 ± 27 ± 29.67 ± 14.25 ± 11 ± Cistus sp. Cistaceae 10 ± 2.96 6 ± 0.53 0.14 0.85 4.82 7 2.52 7.86 7.12 3.65 4.41 Convolvulus coele Convolvula 1.33 ± 1.33 ± 0.33 ± 6.67 ± 2 ± 0.88 0 0 5 ± 2.89 9 ± 5.2 0.46 0.93 0.58 syriacus ceae 0.51 0.51 0.19 3.85 Convolvulus penta Convolvula 8.67 ± 1.67 ± 3.33 ± 0.25 ± 2.5 ± 4 ± 1.77 5 ± 1.85 7 ± 2.73 6 ± 3.18 0.65 0.08 0.47 petaloides ceae 2.5 0.96 1.92 0.14 1.44 Crassulacea 5.67 ± 3.33 ± 0.25 ± 1.33 ± Sedum hispanicum 4 ± 1.2 0 0 1 ± 0.33 1 ± 0.58 0.19 0.73 0.71 e 2.46 1.35 0.14 0.77 Cephalaria joppen Dipsacacea 8.67 ± 2.33 ± 2.5 ± 1.67 ± 9 ± 2.33 6 ± 1.85 0 7 ± 3.48 9 ± 4.62 0.34 0.80 0.22 sis e 2.17 0.51 0.55 0.69 Chrysanthemum c Dipsacacea 1.33 ± 1.33 ± 1.33 ± 39.67 ± 12.67 ± 19.33 ± 7.5 ± 3 ± 1.45 4 ± 1.13 0.40 0.31 0.54 oronarium e 0.19 0.19 0.39 21.75 4.19 3.03 3.18 1.33 ± 7.67 ± 10.33 ± 1.75 ± 4.67 ± Linum pubescens Linaceae 2 ± 0.66 1 ± 0.33 9 ± 2.73 5 ± 1.15 0.32 0.17 0.43 0.77 2.37 3.47 0.64 1.26 Lotus Papilionace 0.33 ± 4.67 ± 4.33 ± 3.75 ± 11.67 ± 1 ± 0.58 0 4 ± 1.53 7 ± 2.31 0.00 0.36 0.54 longesiliquosus ae 0.19 1.17 0.51 0.95 2.5 Scorpiurus murica Papilionace 4.33 ± 2.67 ± 1.75 ± 0.33 ± 0.67 ± 0.5 ± 1 ± 0.58 0 0 0.33 0.87 0.14 tus ae 1.95 1.02 0.83 0.19 0.19 0.29

Trifolium campestr Papilionace 1.67 ± 0.33 ± 30 ± 6.67 ± 13.25 ± 3.67 ± 5.67 ± 0 1 ± 0.58 0.29 0.46 0.33 e ae 0.39 0.19 13.48 2.78 2.88 1.58 1.58

Trifolium clypeatu Papilionace 1.33 ± 0.67 ± 0.67 ± 2.67 ± 2.5 ± 3.67 ± 7.33 ± 1 ± 0.58 4 ± 2.31 0.26 0.51 0.25 m ae 0.77 0.39 0.19 1.26 0.89 1.58 1.54 Aegilops geniculat 3.33 6.67 1.67 Poaceae 4 ± 2.31 5 ±1.85 2 ±1.15 2 ±1.15 4.67 ±2.7 0 0.95 0.85 0.67 a ±1.02 ±3.85 ±0.51

2.67 ± 0.33 ± 0.33 ± 0.25 ± 15 ± 31 ± Avena sterilis Poaceae 0 0 1 ± 0.58 0.27 0.48 0.34 1.54 0.19 0.19 0.14 8.66 17.03 Brachypodium dist 0.33 ± 2.67 ± 4.33 ± 5.5 ± 3.67 ± 1.33 ± 0.5 ± Poaceae 8 ± 2.08 4 ± 1.33 0.57 0.63 0.19 achyon 0.19 1.54 1.65 1.71 1.58 0.77 0.29 0.67 ± 73.67 ± 18 ± 79 ± 38.67 ± 26.33 ± 0.5 ± Bromus sp. Poaceae 0 8 ± 4.34 0.53 0.46 0.50 0.39 42.24 6.17 28.04 22.33 12.7 0.29 Bromus lanceolatu 5.67 ± 13.67 ± 18.67 ± 7.67 ± 12 ± 13.25 ± 16 ± 19.33 ± 11 ± Poaceae 0.81 0.44 0.95 s 1.71 3.03 8.8 2.14 5.29 4.6 3.71 9.72 1.73 Bromus madritensi 4.67 ± 1.33 ± 0.67 ± 41 ± 5.33 ± Poaceae 4 ± 1.33 2 ± 0.66 8 ± 4.62 1 ± 0 0.32 0.82 0.94 s 0.51 0.19 0.39 22.15 2.22 Catapodium rigidu 24.67 ± 24.33 ± 41.33 ± 21 ± 47.33 ± 37 ± 25.67 ± 28.67 ± 27.5 ± Poaceae 0.93 0.68 0.88 m 4.67 6.17 8.26 4.66 24.45 7.18 5.93 4.86 1.44 21.33 ± 17.33 ± 188 ± 243 ± 79.25 ± 2.33 ± 57.33 ± 1.5 ± Poa infirma Poaceae 7 ± 1.53 0.58 0.44 0.54 9.43 9.15 105.08 135.1 41.92 1.35 32.24 0.87 41.67 ± 83.33 ± 70.33 ± 79.33 ± 198.33 ± 76.5 ± 95.33 ± 212.33 ± 72.5 ± Anagallis arvensis Primulaceae 0.50 0.00 0.05 7 9.05 0.51 4.28 5.67 9.01 30.51 53.95 25.69 Crucianella macro 1.33 ± 1.33 ± 2.67 ± 4.33 ± 13 ± 2.75 ± 2.33 ± 0.5 ± Rubiaceae 6 ± 1.33 0.95 0.30 0.17 stachya 0.51 0.51 0.69 1.39 5.13 1.4 1.07 0.29 Scrophularia rubri Scrophulari 0.33 ± 0.33 ± 3.33 ± 2.25 ± 0 0 4 ± 2.31 9 ± 4.91 4 ± 2.31 0.56 0.19 0.34 caulis aceae 0.19 0.19 1.39 0.64 Theligonum cynoc Theligonace 6.67 ± 3.33 ± 6.5 ± 1.67 ± 1.67 ± 2 ± 1.15 2 ± 0.88 5 ± 1.53 1 ± 0.58 0.65 0.20 0.23 rambe ae 0.51 1.17 2.36 0.69 0.39

תקציר

הדינמיקה של חברת הצומח מושפעת באופן ניכר מתהליכי נביטה מבנק הזרעים בקרקע. תהליכים אלו חשובים במיוחד באזורים ים-תיכוניים המאופיינים באקלים עונתי, בהם תהליכי נביטה מבנק הזרעים מוגבלים בעיקר לעונה הרטובה, ולכן להרכב בנק הזרעים צפוי להיות תפקיד חשוב בקביעת הדינמיקה של חברת הצומח. מעבר לכך, מערכות ים-תיכוניות הן בעלות היסטוריה ארוכה של הפרעות מסוגים שונים, וכתוצאה ממגוון ההפרעות המתחוללות באזור זה והשונות הגדולה בתדירותן ועוצמתן, הנביטה מבנק הזרעים בקרקע מהווה תהליך מרכזי בעיצוב חברת הצומח. בנוסף, שינוי במשטר ההפרעה עלול לגרום למעבר בין מצבים חלופיים יציבים (alternative stable states). שינויים כאלו בדרך כלל באים לידי ביטוי גם בשינויים בתצורת הצומח המאפיינת את הסביבה המופרעת. מניעה של שינויים קיצוניים במערכת האקולוגית והגנה על מינים הרגישים לשינויים אלו דורשים שימוש בממשק אקטיבי.

מטרת עבודת הדוקטורט שלי הייתה לחקור את ההשפעה של משטרי הפרעה שונים על מבנה חברת הנביטה מבנק הזרעים בקרקע. לבנק הזרעים בקרקע יש חשיבות גדולה בשימור מגוון מיני הצומח משום שהוא משמש כמאגר להתחדשות של מינים רבים המתחדשים מזרעים, אשר יכולים להיעלם כתוצאה מהפרעות שונות. לכן, משטר הפרעה הוא גורם חשוב היכול לייצר שונות בהרכב בנק הזרעים ובדגמי הנביטה המאפיינים סביבות מופרעות. עבודת הדוקטורט שלי מורכבת משלושה פרקים הבוחנים כיצד שינויים בדגמי נביטה מבנק הזרעים מושפעים מ: 1) עונת השריפה, 2) האינטראקציה בין היסטוריית השריפה ואות של עשן, 3) משטר ניהול של החורש הכולל פתיחה חד פעמית של הנוף שבוצעה לפני כ20- שנה. בניסוי השלישי חקרתי גם את ההשלכות ארוכות הטווח של פתיחת הנוף החד-פעמית על האוכלוסייה של המין אדמונית החורש הנמצא בסכנת הכחדה.

מטרת הפרק הראשון של העבודה הייתה לחקור האם וכיצד עונת השריפה, מיקרו-בית הגידול (תחת אלות, לוטמים ובשטחים פתוחים) והאינטראקציה ביניהם, משפיעים על הרכב חברת הצמחים הנובטים מבנק הזרעים בקרקע בחורש טיפוסי בחלק המזרחי של אגן הים-התיכון. למיטב ידיעתי, זהו הניסיון הראשון לבחון את ההשפעה של שריפות אביב וסתיו על מרכיבים שונים של חברת הצומח, בחלק המזרחי של אגן הים- התיכון על ידי ניסוי שדה רחב ממדים הכולל שריפות מבוקרות. בניסוי זה תיעדתי דגמי נביטה מדגימות קרקע שנלקחו ממיקרו- בתי גידול שונים בחלקות שרופות ומחלקות ביקורת סמוכות. מצאתי כי שריפות גרמו לירידה בצפיפות, עושר במינים ובמגוון של מיני הצמחים שנבטו מבנק הזרעים בקרקע. העושר ומגוון המינים מבנק הזרעים היו נמוכים באופן משמעותי תחת שיחי אלה ולוטם אשר נשרפו בסתיו, והדגם הזה בלט בעיקר בנביטה של מינים חד-שנתיים. צפיפות הנביטה של מיני בני-שיח הייתה גבוהה בדגימות שנאספו משטחים שנשרפו, ודגם זה בלט באופן מובהק בדגימות שנאספו תחת שיחי לוטם ואלה. יחד עם ההופעה של מינים ייחודיים בעונות השונות, עונת השריפה גרמה לשינויים משמעותיים בחברת הנביטה מבנק הזרעים, בעיקר על ידי השפעה דיפרנציאלית על צפיפות המינים החד שנתיים ובני השיח. תוצאות אלו הראו, בפעם הראשונה, כי

האינטראקציה בין עונת השריפה וההטרוגניות המרחבית משפיעים על הרכב חברת הצמחים שנובטים מבנק הזרעים בקרקע, בעיקר על ידי השפעה דיפרנציאלית על הנביטה של מינים חד שנתיים ובני שיח. ממצאים אלו מציעים כי שינויים בעונת השריפה שתועדו בעשורים האחרונים, עלולים להוביל לשינוי בכוחות הסלקציה הפועלים על הצומח במערכת הים-תיכונית.

בפרק השני בחנתי בפעם הראשונה את ההשפעה של יחסי הגומלין בין אות של עשן והיסטוריית השריפות על חברת הצמחים שנובטים מבנק הזרעים בקרקע של החורש הים תיכוני. דגמתי קרקעות מאתרים שנשרפו בתדירות שונה במהלך ארבעת העשורים האחרונים וחשפתי אותם לעשן, עם וללא הרטבה. על ידי תיעוד דגמי הנביטה, בחנתי את השינויים בצפיפות ובעושר המינים של הצמחים שנבטו מבנק הזרעים בקרקע. סך צפיפות הנביטה מבנק הזרעים הייתה גבוהה באתרים שנשרפו בתדירות גבוהה במהלך ארבעת העשורים האחרונים. מקובל לחשוב כי ההשפעה של עשן על נביטת זרעים אינה נפוצה באגן הים התיכון בשל העובדה כי מרבית ההפרעות באזור זה הן אנתרופוגניות, משמע הן הופיעו מאוחר יחסית בקנה מידה אבולוציוני. למרות זאת, תוצאות העבודה הראו כי החשיפה לעשן הגבירה את צפיפות הנביטה מבנק הזרעים, וכי דגם זה בלט יותר בדגימות שנאספו מאתרים שנשרפו בתדירות גבוהה יותר, ובמיוחד בדגמי הנביטה של מינים חד-שנתיים. תוצאות אלו מדגישות את חשיבותם של מחקרים הבוחנים את תגובת הנביטה של חברת הזרעים מהקרקע לעומת מחקרים הבוחנים את תגובת הנביטה הישירה של זרעים שנאספו מהשדה. מעבר לכך, ממצאים אלו מבליטים את התפקיד החשוב של העשן בעיצוב תהליכי סוקצסיה לאחר שריפה באגן בים התיכון, בעיקר על ידי עידוד הנביטה של מינים חד-שנתיים.

מטרת הפרק השלישי הייתה לבחון את ההשפעה ארוכת הטווח של הפרעה חד-פעמית כחלק מממשק, הכולל פתיחה של הנוף, על צפיפות ועושר המינים של הנביטה מבנק הזרעים בקרקע, ועל צפיפות האוכלוסייה וההתרבות של המין המקומי הנמצא בסכנת הכחדה אדמונית החורש. מחקרים הכוללים ניטור ארוך טווח הבוחנים את ההשפעה של משטרי ממשק הם נדירים משום שבמרבית המקרים ממשק אקטיבי נמדד בטווחי זמן קצרים יותר. כימתי את צפיפות הזרעים בבנק הזרעים בקרקע בחלקות בהם כרתו/גזמו את הצומח מעל הקרקע לפני כעשרים שנה, ובחלקות שלא טופלו והחורש בהם נשאר סגור. בנוסף, עקבתי אחר צפיפות צמחי האדמונית, אחוזי הפריחה וצפיפות זרעי האדמונית בקרקע באותן החלקות. יש לציין כי לא נמצאו הבדלים בצפיפות ובעושר המינים של כלל הצמחים שנבטו מבנק הזרעים בקרקע. אולם, בניגוד להשפעה החיובית של פתיחת הנוף החד-פעמית על אחוזי הפריחה של האדמונית שנראתה בטווח הקצר, מצאתי, כי אחוז הפריחה, בנוסף לצפיפות זרעי האדמונית בקרקע היו נמוכים בחלקות בהן הנוף נפתח לעומת החלקות הסגורות. תוצאות אלו מעידות כי ההשפעה החיובית של פתיחת החורש בטווח הקצר, יכולה להיעלם בטווח הארוך ואף להפוך לשליליות עם הזמן בעקבות תהליכים שיכולים להתרחש בזמן הפתיחה או הסגירה של החורש. בנוסף, ממצאים אלו מעידים כי אוכלוסיית האדמוניות בשמורת הר מירון יציבה באזורים בהם החורש סגור ולא הופרע בעבר, ולכן אין צורך בהתערבות אקטיבית לשימור האוכלוסייה.

לתוצאות עבודת הדוקטורט שלי יש השלכות חשובות להבנה של יחסי הגומלין בין משטרי הפרעה שונים לבין הרכב חברת הנביטה מבנק הזרעים. תהליך הנביטה מבנק הזרעים הכרחי להתבססות חברת הצומח, במיוחד לאחר הפרעות. לכן, ממצאי עבודת הדוקטורט שלי יוכלו לשפר את היכולת לחזות שינויים בדינמיקה של חברת הצומח לנוכח השינויים האקלימיים שמתרחשים באגן הים התיכון החשוף להשפעות אנתרופוגניות רבות.

מילות מפתח: בנק הזרעים בקרקע, היסטוריה מרחבית של משטר השריפה, חורש ים תיכוני, מינים בסכנת הכחדה, מינים זריעים, מינים חד-שנתיים, מיקרו בית גידול, משטר הפרעה, נביטה, ניהול אקטיבי, עונת שריפה, עשן, פתיחת נוף.

תודות,

הייתה לי זכות גדולה לבצע את עבודת המחקר תחת הנחייתו של פרופ' עופר עובדיה, שהפך את המעבר הלא פשוט מעולם הביולוגיה המולקולארית לעולם האקולוגיה למלא בעניין ועומק. תודה על הסבלנות הרבה והשעות הלא מבוטלות שהשקעת בגיבוש רעיונות, תכנון ניסויים וכתיבה משותפת. בארבעת השנים האחרונות זכיתי ללמוד מהידע העצום שלך בתחומים רבים, היית ותהיה מקור השראה בשבילי בהמשך הדרך. ברצוני להודות גם לדר' חגי שמש על הליווי במחקר התיאורטי ולא מעט פעמים גם בשטח, על הרעיונות היצירתיים והטובים שהתגבשו למחקרים, ועל ההכוונה והמיקוד בכל שלבי המחקר והכתיבה. חלק לא מבוטל מעבודת המחקר התבצע בשיתוף עם סטודנטיות ממכללת תל-חי ומאוניברסיטת בן גוריון. היה לי כיף גדול לעבוד עם אלה דגון שבחריצות והשקעה רבה עבדה על המחקר בשמורת הכרמל, העבודה בשטח ובבית רשת היו שילוב טוב של לימוד והנאה. תודה נוספת לענבל אילון על שהמשיכה את העבודה בכרמל, ושבמאמץ גדול ודקדקנות רבה הצליחה להוציא מהפרויקט תוצאות מפתיעות ומעניינות. חשוב לי להודות גם לפרופ' לאה ויטנברג ופרופ' דן מלקינסון מאוניברסיטת חיפה על העזרה באיתור שטחי המחקר בשמורת הכרמל. תודה גדולה לחברי המעבדה ודים, יוני, עדי, ענת, סתו ואילון על החברות הטובה, העלאת המורל והעצות המועילות. תודה לוועדה המלווה פרופ' ירון זיו ודר' מרב סייפן על הביקורת הבונה והייעוץ לאורך הדרך. בשנים בהם ביצעתי את עבודת המחקר הכרתי ולמדתי נושאים שונים הקשורים לתחום האקולוגיה ולמחקר בכלל, והרבה בזכות חוקרים ומורים רבים שפגשתי לאורך הדרך ביניהם פרופ' יוחאי כרמל, דר' יגיל אוסם, דר' איתמר גלעדי, דר' חיים סיוון ופרופ' קטיה טילברגר. תודה אחרונה וחשובה למשפחה הקרובה, שהייתה העוגן שלי לאורך המסע הזה, על הגב והתמיכה בדרכים בהם אני בוחרת ללכת.

הצהרת תלמיד המחקר עם הגשת עבודת הדוקטורט לשיפוט

אני החתום מטה מצהיר/ה בזאת:

X חיברתי את חיבורי בעצמי, להוציא עזרת ההדרכה שקיבלתי מאת מנחה/ים.

X החומר המדעי הנכלל בעבודה זו הינו פרי מחקרי מתקופת היותי תלמיד/ת מחקר.

בעבודה נכלל חומר מחקרי שהוא פרי שיתוף עם אחרים ,למעט עזרה טכנית הנהוגה

בעבודה ניסיונית .לפי כך מצורפת בזאת הצהרה על תרומתי ותרומת שותפי למחקר ,שאושרה על

ידם ומוגשת בהסכמתם.

תאריך: 25/02/2019 שם התלמיד/ה : נטע מנלה חתימה

העבודה נעשתה בהדרכת:

פרופ' עופר עובדיה

במחלקה למדעי החיים,

בפקולטה למדעי הטבע.

אוניברסיטת בן גוריון בנגב.

יעוץ חיצוני:

דר' חגי שמש

במחלקה ללימודי הסביבה,

מכללת תל חי.

שונות בזמן ובמרחב בהרכב חברת הנביטה המאפיינת את החורש באגן הים תיכוני המזרחי

מחקר לשם מילוי חלקי של הדרישות לקבלת תואר" דוקטור לפילוסופיה"

מאת

נטע מנלה

הוגש לסינאט אוניברסיטת בן גוריון בנגב

אישור המנחה

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אדר תשע"ט פברואר 2019

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