Restoring the Biodiversity of Canopy Species Within Degraded Spekboom Thicket

Restoring the Biodiversity of Canopy Species within Degraded Spekboom Thicket

Marius Lodewyk van der Vyver

Submitted in fulfilment of the requirements for the degree Magister Scientae in the faculty of Science at the Nelson Mandela Metropolitan University

January 2011

Supervisors: Prof. R.M. Cowling & Prof. E.E. Campbell

Acknowledgements

“In the name of Allah, the most compassionate, the merciful”.

Firstly I would like to thank the One, the All-Knowing, the All-Mighty, to whom all praise is due. I am very grateful to my parents for their unconditional love and support, I know I don‘t appreciate them enough. Prof. Richard Cowling, my supervisor, who had been instrumental in making this thesis happen, has opened up for me wonderful new worlds and horizons on many levels. I would like to thank him and his wife Shirley for their love, support and guidance, without which this thesis would not have been possible. Prof. Eileen Campbell is thanked for her encouragement, valuable advice and direction. I thank the Restoration Research Group that provides scientific oversight of the Subtropical Thicket Restoration Project (STRP). Dr. Anthony Mills, whose work has been a model of excellence to aspire to, deserves thanks for his valuable comments and help with the soil sampling procedures, carbon sequestration analyses and his sympathetic ear. Dr. Ayanda Sigwela is thanked for his valuable questions, help and comments. Mike Powell is thanked for his help and suggestions regarding the allometric sampling procedures.

Funding for this research was provided by the Department of Water and Environment‘s (DWE) Working for Woodlands which initiated the subtropical thicket restoration project (STRP) of which Gamtoos Irrigation Board (GIB) is the implementing agent. I sincerely would like to thank these organizations for the opportunity and their assistance with the project. Dr. Christo Marais from DWE, and Andrew Knipe, Yolande Vermaak, Pippa Holm, and Victoria Wilman of GIB are thanked in this regard and also the Working for Woodlands contractor teams that fenced off my experimental plots and performed the practical soil sampling procedures. Victoria Wilman is especially thanked as she was instrumental in the propagation and preparation of the propagules, and has provided valuable advice and encouragement. Yolande Vermaak has helped considerably with organizing the fencing of my experimental plots and Pippa Holm has been a great help with handling and sorting the soil samples.

I want to thank Dr. Mark Difford for his guidance with statistical analyses, he has proven himself to be a true guru, not just through his thorough analyses and knowledge of statistical intricacies, but also displaying superior patience and

i providing clear guidance to an inspired but mostly confused R novice. The late Graham Slater and his wife Rhoda are thanked for their cheerful assistance, friendly conversation and many cups of tea during my fieldwork at Krompoort. I would also like to thank Mr. Desmond Slater for his advice and friendly conversation during long hours of solitary fieldwork. I thank Mr. Warren Rudman and his wife Belinda for their encouragement and their warm, hospitable reception and treatment of me on their farm Rhinosterhoek. Both Mr. Warren Rudman and Mr. Desmond Slater have provided valuable rainfall data, practical advice and personal observations of the ecosystem within which they have grown up and make a living in. Without Mr. Rudman (snr.) and Mr. Graham Slater‘s vision to plant spekboom 50 and 33 years ago respectively, this research and our better understanding of biodiversity regeneration within spekboom thicket would not have been possible. I hereby would like to thank them for this, albeit posthumously.

Clayton Weatherall-Thomas and Merika Louw are thanked for their camaraderie and help with fieldwork. Everybody in the Botany Department at NMMU with whom I had dealings with is also thanked for their overall friendly and prompt assistance, enlightening conversation and the occassional lighter moments we shared. I would also like to thank all my friends and family who had to tolerate my rebuffs when I was invited for social gatherings or outings. Thanks especially to Gillian MacGregor, Robert Dugtig, Andre Stahmer, Etienne van der Vyver, Jacqueline Brown, Colette Stoltz, Faadil Khan, Margherita Razza, Steve Kalule, Cornel du Toit, Sadia Bundgaard, Luke Hutchinson, Eduard Lewis and Andon van der Merwe for their precious friendship and encouragement.

Table of Contents

Acknowledgements ...... i Table of Contents ...... iii List of Figures ...... iv List of Tables ...... vi Summary ...... 1 Chapter 1 – General Introduction ...... 3 Introduction ...... 3 Restoration and Ecology ...... 4 Spekboom Thicket ...... 6 Restoration of Spekboom Thicket ...... 9 Chapter 2 ...... 11 Restoring biodiversity by planting propagules of canopy species in degraded spekboom thicket is neither ecologically nor economically feasible...... 11 Abstract ...... 11 Introduction ...... 12 Materials and Methods ...... 15 Study Site...... 15 Nursery Propagation and Planting ...... 18 Survival and Growth Assessment ...... 19 Restoration Costs ...... 20 Statistical Analysis ...... 21 Results ...... 22 Survival and Growth Analysis ...... 22 Restoration costs ...... 27 Discussion ...... 27 Chapter 3 ...... 31 Biodiversity Regeneration and Carbon Sequestration on Restored Spekboom Thicket Landscapes – Portulacaria afra Jacq. as an Ecological Engineer...... 31 Abstract ...... 31 Introduction ...... 32

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Methods ...... 37 Site Description ...... 37 Data collection ...... 39 Statistical analysis ...... 44 Results ...... 45 Carbon ...... 45 Floristics ...... 49 Discussion ...... 54 Chapter 4 – General Discussion ...... 58 Research significance of this study ...... 60 Shortcomings and limitations of this study ...... 60 What are the next steps ? ...... 61 Literature cited ...... 66 APPENDIX A...... 75 APPENDIX B...... 76

List of Figures

Figure 1.1: A fenceline contrast in spekboom thicket showing intact thicket (behind) and degraded thicket (front). These scenes are common in P. afra dominated thickets as a result of mismanagement by stock farmers...... 7 Figure 1.2: The distribution of spekboom thicket and the extent of its degradation. Green shows intact spekboom thicket and red shows degraded spekboom thicket areas (Lloyd et al., 2002) ...... 8 Figure 2.1: Location map of the experimental site Krompoort (-33.544690 ; 25.174098) The location of Krompoort is shown in yellow. Green shading represents intact spekboom thicket while red shading represent areas where spekboom thicket landscapes are moderately to heavily degraded (Lloyd et al, 2002)...... 16 Figure 2.2: Google Earth© image of Krompoort showing the Intact (cyan), Degraded (pink), Restored33 (= 33 yr-old post restoration) (red), Restored19 (yellow) and Restored11 (blue) sampling sites...... 17 Figure 2.3: Boxplots showing monthly rainfall measured at the Krompoort restoration site over a period of 40 years (1970-2009). Filled circles indicate medians, boxes indicate the lower and upper quartile (i.e. 25% and 75% of the distribution), whiskers indicate the sample minimum and maximum, unfilled circles show outliers...... 17

Figure 2.4: Results from the conditional inference tree analysis of seedling survival planted in autumn and assessed after 12 months. Rs33, Rs19 and Rs11 are abbreviations for the different aged restoration treatments with appended numbers indicating restoration age and Por_afr, Lyc_fer, Pap_cap, Sea_lon and Rhi_obo are species name abbreviations. The y-axis of the bars shows Y (survived) and N (dead) expressed as a fraction of the total number of propagules...... 26 Figure 2.5 Results from the conditional inference tree analysis of seedling survival planted in spring and assessed after 24 months. Lyc_fer, Pap_cap, Sea_lon and Rhi_obo are species name abbreviations. The y-axis of the bars shows Y (survived) and N (dead) expressed as a fraction of the total number of propagules...... 26 Figure 3.1: Portulacaria afra Jacq. (spekboom) tree. Notice the skirt of rooted branches around the base of the tree...... 34 Figure 3.2: Hypothetical state-and-transition model, adapted from Hobbs & Norton (2004) to describe spekboom thicket degradation (by goat pastoralism) and recovery dynamics. Shown are alternative states (states 1-6) that can manifest along a gradient of browsing pressure intensity and duration (left to right) that is the cause of increasing degradation. States 1 and 2 show possible alternating states of the system where browsing pressure ≤ autogenic recovery under the established browsing regime. The biotic threshold (green bar) symbolizes the point after which no autogenic return to states 2 and 1 are possible without removing the browsing pressure. The system may at this stage exhibit a number of new alternating states (states 3 and 4). Should the browsing pressure continue unabated, the system moves across the abiotic threshold (red bar) where abiotic factors limit any autogenic recovery potential and the system degenerates towards a karoo-like landscape (State 6)...... 35 Figure 3.3: Location of study sites shown in yellow. The green represent spekboom thicket areas still fairly intact, the red represent areas where spekboom thicket landscapes are moderately to heavily degraded (Lloyd et al, 2002)...... 37 Figure 3.4: Krompoort site showing the Intact (cyan), Degraded (pink), Restored33 (red), Restored19 (yellow) and Restored11 (blue) sampling sites. The numbers refer to age of restoration...... 38 Figure 3.5: Rhinosterhoek site. Top shows the spread of the four sampling sites, while bottom left shows the Restored50 (red), Restored35 (yellow) and Degraded (pink) sites, bottom right shows the Intact (cyan) sampling site. The numbers refer to age of restoration...... 40 Figure 3.6: Carbon stocks at Rhinosterhoek and Krompoort a soil depth of 50 cm. Data for the Krompoort (prefix=K.) sites were taken from Mills & Cowling (2006) and data for the intact sites were taken from Mills et al. (2005a). Data from this study (Rinosterhoek) has prefix = R...... 49 Figure 3.7: Multiple correspondence analysis showing results according to (a) treatment (b) status, (c) site and (d) cover components. In (a) R = Rhinosterhoek and K = Krompoort, Dgrd is Degraded, Intc is Intact and Rs is Restored, and the number attached shows number of years under restoration treatment. In (a) and (b) the labels for K.Rs33 /Restored33 (purple) obscures that of K.Rs19 / Restored19 (blue).In (c) labels are placed over the midpoint (barycentre) of the respective point- clouds. Ellipses are so called inertia ellipses and encapsulate 67% of the data points associated with each point-cloud. The ellipse axes shows the principle axes of variation (Chessel & Dufour, 2004)...... 50 Figure 3.8: Multiple correspondence analysis showing (a) results according to treatment (Fig. 3.6a) displaying the total carbon content (TCC) (t C ha-1) of each treatment sampled up to a depth of 50 cm below ground (the label displaying 186 t C ha-1 (Restored33) obscures that of 101.8 t C ha-1 (Restored19) (b) The same display but with overall mean TCC subtracted from observed TCC values and expressed as various sized squares proportional to the calculated deviation from the mean and scaled according to SD=1. The size of the squares thus show each point’s deviation from the mean expressed as a factor of SD. Black shows positive deviation values, white negative...... 51

Figure 3.9: Rao’s diversity coefficient (Quadratic Entropy) using a taxonomic distance metric based on Order – Family – Genus. Prefix=K signifies Krompoort, prefix=R signifies Rhinosterhoek and numbers appended to Rs (Restored) shows the number of years under restoration treatment, Dgrd refers to degraded stands...... 51 Figure 3.10: Boxplots illustrating the pseudo-cover of spekboom thicket canopy recruits across restoration treatments. The width of the boxes increase with the sample size of recruits detected. Recruits included all thicket canopy plant species, ≤ 100 cm in height (excluding P. afra), be they seedlings or sprouts...... 53 Figure 3.11: The pseudo-cover of spekboom thicket canopy constituent recruits per species across treatments with abundance (as opposed to canopy cover) assigned to a cover scale used with adult plants where 1=1, 2=2-4, 3=5-10, 4=11-15, 5=16-24, 6=25-34 individuals ...... 53 APPENDIX B: Figure 1: Results from a standard linear regression model using ordinary least squares for estimating Total Carbon Content (TCC) at a soil depth of 50 cm. The model is based on the formula TCC ~ Depth + Status, and R2 = 0.905 and adjusted R2 = 0.899...... 76 APPENDIX B: Figure 2: A summary of the floristics data set, with species as rows and treatments as columns (tabular part of the figure). Species names with prefix=R signifies recruits (pseudo-species). The Dendrogram is based on a taxonomic distance metric. Square-size shows counts per treatment. The columns of the abundance matrix have been arranged using the scores (coordinates) from the first dimension of a correspondence analysis of the matrix. The arrangement differs little from that that would be got by arranging columns in a more or less ordered progression from degraded sites through increasing time of restoration to intact sites (from right to left). Degraded vs Restored/Intact are well characterized by the presence of some species and the absence of others...... 77

List of Tables

Table 2.1: Number of propagules planted per treatment during two seasons (spring and autumn). Dgrd refers to the degraded stand and Intc to the intact stand (control) . Rs refers to the restored stands where the number appended refers to the no. of years under restoration treatment...... 19 Table 2.2: Cover scale used for determining a measure of shading for each seedling planted...... 20 Table 2.3: Survival and growth by treatment for Pappea capensis. The number appended to the name of the restored sites show restoration age. Survival3 refers to the survival assessment conducted 3 months after planting in September 2008, Survival12 refers to the survival assessment 12 months after the planting in May 2009 and Survival24 refers to the survival assessment 24 months after the September 2008 planting. Grwth_par refers to the total growth parameters of surviving propagules, irrespective of planting times...... 23 Table 2.4: Survival and growth by treatment for Searsia longispina. The nu mber appended to the name of the restored sites show restoration age. Survival3 refers to the survival assessment conducted 3 months after planting in September 2008, Survival12 refers to the survival assessment 12 months after the planting in May 2009 and Survival24 refers to the survival assessment 24 months after the September 2008 planting. Grwth_par refers to the total growth parameters of surviving propagules, irrespective of planting times...... 23 Table 2.5: Survival and growth by treatment for Lycium ferocissimum. The number appended to the name of the restored sites show restoration age. S urvival3 refers to the survival assessment

conducted 3 months after planting in September 2008, Survival12 refers to the survival assessment 12 months after the planting in May 2009 and Survival24 refers to the survival assessment 24 months after the September 2008 planting. Grwth_par refers to the total growth parameters of surviving propagules, irrespective of planting times...... 24 Table 2.6: Survival and growth by treatment for Rhigozum obovatum. The number appended to the name of the restored sites show restoration age. Survival3 refers to the survival assessment conducted 3 months after planting in September 2008, Survival12 refers to the survival assessment 12 months after the planting in May 2009 and Survival24 refers to the survival assessment 24 months after the September 2008 planting. Grwth_par refers to the total growth parameters of surviving propagules, irrespective of planting times...... 24 Table 2.7: Survival and growth by treatment for Portulacaria afra. The number appended to the name of the restored sites show restoration age. Survival12 refers to the survival assessment 12 months after the planting in May 2009...... 25 Table 2.8: Summary of cost estimation for planting an area of 30 ha with nursery propagated cuttings and seedlings (1:1 ratio) by a typical Working for Water contractor team. Capital Build-up refers to the contractor’s profit for completing the contract, and is calculated at 20% of wage cost. The estimation below involves a total number of 4800 propagules planted at a density of 60 plants ha- 1species-1...... 27 Table 3.1: Belowground carbon sample (n) sizes per depth interval at Rhinosterhoek ...... 40 Table 3.2: Allometric regression equations used for aboveground carbon sampling ...... 41 Table 3.3: Species allometric equivalents used...... 42 Table 3.4: Outline of different treatment sites sampled ...... 43 Table 3.5: Cover scale used for Adult plants and Recruits (“pseudo-species”) ...... 43 Table 3.6: Carbon (t .ha-1) data (mean values) from Rhinosterhoek. Below G = below ground, Above G = above ground...... 48 APPENDIX A: Table 1: Restoration costs per canopy species planted at a density of 60 plants per hectare based on the costing of actual large scale Working for Woodlands contracts for planting dense rows of P. afra truncheons. The table shows the difference per species based on propagation method...... 75

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Summary

I investigated the return of plant canopy diversity to degraded spekboom thicket landscapes under restoration treatment. I attempted the reintroduction of five nursery propagated and naturally-occurring plant species in severely degraded Portulacaria afra Jacq. (spekboom) dominated thickets that have been subjected to a restoration method involving the planting of dense rows of P. afra truncheons for various time periods and also in degraded and intact thickets. I also planted nursery propagated P. afra cuttings. An average of 30 propagules of each species, were planted in each of the chosen areas in two distinct seasons that exhibited distinct rainfall peaks. Sixteen propagules of P. afra were also planted in each treatment only once. Propagules of the two thicket woody canopy species (S. longispina and P. capensis) showed a total survival of 1% and 9%, respectively. Survival of L. ferocissimum and R. obovatum was 19% and 70% and all propagules of P. afra survived. Analyses showed that survival is primarily tied to a species effect, with R. obovatum and P. afra showing significantly better survival than the other species. Within the other surviving few species a significant preference for overhanging canopy cover was observed. The results show little significance of restoration treatment for propagule survival, suggesting that a range of conditions is needed for the successful establishment of canopy species that likely involves a microclimate and suitable substrate created by canopy cover and litter fall, combined with an exceptional series of rainfall events. I found that the high costs involved with a biodiversity planting endeavour, and the low survival of propagules of thicket canopy plant species (P. afra excepted), renders the proposed biodiversity planting restoration protocol both ecologically and economically inefficient.

Restoration success involves the autogenic regeneration of key species or functional groups within the degraded ecosystem. Heavily degraded spekboom- dominated thicket does not spontaneously regenerate its former canopy species composition and this state of affairs was interpreted in terms of a state-and- transition conceptual model. Floristic analyses of degraded, intact and a range of stands under restoration treatment for varying time periods at two locations in Sundays Spekboomveld revealed that the stands under restoration are progressively regenerating canopy species biodiversity with increasing restoration age, and that intact sites are still the most diverse. The high total carbon content (TCC) measured within the older restored stands Rhinosterhoek (241 t C ha-1 after 50 years at a depth of 50 cm) rivals that recorded for intact spekboom thickets, and the number of recruits found within older restored sites rivals intact sites sampled.

The changes recorded in the above- and belowground environments potentially identify P. afra as an ecosystem engineer within spekboom dominated thickets that facillitates the build-up of carbon above- and belowground and the accompanying changes in soil quality and the unique microclimate aboveground, which enables the hypothetical threshold of the degraded state to be transcended. This restoration methodology is accordingly considered efficient and autogenic canopy species return was found to be prominent after a period of 35-50 years of restoration treatment.

Keywords: Restoration ecology, Spekboom thicket, Portulacaria afra, Biodiversity, Ecosystem engineer, Carbon sequestration, Subtropical thicket, Canopy species, Ecological restoration.

Chapter 1 – General Introduction

Introduction

The idea that natural resources have a finite capacity to satisfy growing human population growth and consumption habits is not a new one. It is widely held that many forms of territorial behavior displayed by a variety of creatures are rooted in an innate recognition mechanism of a finite resource base that can only sustain a restricted consuming population. In a number of cases it is argued that the societal collapse of once thriving bastions of human culture, organisation and wealth are based on overutilization of natural resources (see Tainter, 1988; Diamond, 1994; 2004). Most experienced farmers will concede that a single fenced piece of land can only support an animal population of a limited size and remain sustainable. Our prevailing global economic outlook is still largely based on the premise that nature has an infinite potential for satisfying our every desire (Wackernagel & Rees, 1997). This tacit assumption and its manifested environmental consequences have led to the current situation where humans have altered global ecosystems to such a degree and at such a pace that no period of similar length in human history can compare to the considerable and largely irrevocable loss of biodiversity and habitat that has occurred in the last 50 years (Millenium Ecosystem Assessment, 2005).

The realization of this trend has led some economists and ecologists to conceptualize in theory and advocate in practice the valuation and quantification of ecosystem services and natural capital (Constanza & Daly, 1994). It is increasingly being realized that natural capital is a limiting factor in economic growth and in human psychological, social and political health (Aronson et al., 2007, Hobbs, 2004). Conservation efforts are crucial, but to stem the tide of degradation and address the monumental transformation of life-giving ecosystems worldwide, the restoration of natural capital has become an increasingly worthwhile enterprise (Aronson et al., 2006). The restoration of natural capital includes the practice of ecological restoration, but also considers the socio-economic interface between humans and nature, thus striving to involve local communities to realize the benefits and value of their natural environment and to allow local economies to benefit and take sustainable ownership thereof (Aronson et al.,2006; 2007).

Restoration and Ecology

A plethora of definitions of restoration ecology and ecological restoration can be found in the literature. The etymology of the term ‗ecological restoration‘ suggests a drive to return a degraded or altered ecosystem to some resemblance of its historical or ideal reference state. ―Ecological restoration is the process of assisting the recovery of an ecosystem that has been degraded, damaged, or destroyed‖ (SER, 2004). It has also been hailed as the difficult, if not impossible, attempt to return a degraded ecological system to some historical state (Palmer et al., 2006). This difficulty refers to the complex and often non-linear workings of ecosystems and its relation to our hypothetical models on which management actions are based (Westoby et al., 1989; Wallington et al., 2005) and a range of other complications attributed to little understanding of the ecosystem dynamics before degradation and a corresponding wrong diagnosis or perception of the cause of the degradation (Hobbs, 2007). Insufficient methodology to reverse or mitigate the degradation and inadequate follow-up monitoring, evaluation and adaptive management have been identified as factors contributing to failures. Exacerbating this are i) the reality of changing climate (Harris et al., 2009), ii) possible interactions between adjoining ecosystems and larger-scale effects (Hobbs, 2007), and in some cases iii) the paucity or lack of historical data and information of significant past species compositions and historical events that influenced the system (Jackson & Hobbs, 2009). Furthermore the uncertainty regarding the parameters that characterize success for restoration attempts (Ruiz-Jaen & Aide, 2005; Cortina et al., 2006), and the intricacies involved with choosing reference sites are added factors to contemplate (White et al., 1997).

To approach the difficulties inherent in ecological restoration responsibly, it is highly recommended to formulate clear and realistic restoration goals that take into account i) the past and current dynamics of the degraded ecosystem (Jackson & Hobbs, 2009), ii) a variety of site-specific sociological and economic factors which includes the high costs of restoration practice and local stakeholder and political involvement (Hobbs, 2007; Aronson et al., 2006; Palmer et al.,1997; Hobbs & Norton, 1996). Appropriate restoration targets and trajectories should reflect both ecological and societal constraints on restoration practice (Hobbs, 2007; Aronson et al., 2007). The necessity of integrating ecological theory and research with restoration practice cannot be overstressed (Halle & Fattorini, 2004; Young et al., 2005; Palmer et al., 2006).

Restoration ecology has been called ―the science behind ecological restoration‖ or ―the scientific process of developing theory to guide restoration and using restoration to advance ecology‖ (Palmer et al., 2006). A paradigmatic shift in

4 ecological thinking has emerged during the past two or three decades. There has been a major shift from the notion of ecosystems as stable entities that are in equilibrium, towards the notion of dynamic and complex systems that display non- linear responses to disturbance regimes and multiple states over time and space (Westoby et al., 1989; Wallington et al., 2005; Young et al., 2005). New thinking on ecosystem regeneration and assembly, with concepts such assembly rules (Belyea & Lancaster, 1999; Temperton et al., 2004; Hobbs et al., 2007), functional diversity (Diaz & Cabido, 2001), thresholds (Rietkerk & Van de Koppel, 1997; Briske et al., 2006), keystone species (Mills et al., 1993; Palmer et al., 2007) and ecological engineers (Jones et al., 1997; Byers et al., 2006; Jones et al., 2010) are increasingly being implemented in ecological restoration methodology (Suding et al., 2004; Cortina et al., 2006). There have been numerous pleas for close networking between restoration practice and ecological theory (Young et al., 2005; Palmer et al., 2006; Hobbs, 2007).

In many cases restoration of degraded landscapes requires an integrated approach drawing on expertise from a range of different disciplines. A general rule of thumb is that the difficulty of restoration is directly related to the degree of transformation of the original ecosystem. Restoration practice may thus range in intensity from a slight management action such as eliminating a single disturbance from the system and allowing the ecosystem to recover, driven by its own innate dynamics of regeneration, to large-scale interventions and physical landscape engineering. Restoration success may be defined as that stage where a recovering ecosystem has regained a natural range of ecosystem composition, structure and dynamics, which translates as regaining its resilience (SER, 2004; Ruiz-Jaen & Aide, 2005). This is rarely achieved for severely transformed landscapes (Palmer et al., 2006; Benayas et al., 2010). The Society for Ecological Restoration International sets a guideline by describing nine attributes of restored ecosystems which, inter alia, entail i) a characteristic species assemblage comparable to a reference ecosystem, ii) the presence of all necessary functional groups for sustained ecosystem stability, or at least the potential exists for natural colonization, and iii) the physical environment of the restored ecosystem is capable of sustaining reproducing populations of the species necessary for its continued stability or development along the desired trajectory (SER, 2004). Benayas et al. (2010) adds to this the return of ecosystem services. Often the integrated approach to restoration allows for mechanisms such as payment for ecosystem services (PES) (Palmer & Filoso, 2009), carbon credit markets (Galatowitsch, 2009) and social upliftment programmes (Marais et al., 2009; Aronson et al., 2007) to be incorporated with ecological restoration. These essentially social components provide socio-political support (including financial resources) to meet the high costs involved with attempting restoration of large-scale degraded landscapes.

Spekboom Thicket

Subtropical thicket vegetation has only relatively recently been recognized as a biome separate from the adjacent karoo, savanna and forest biomes (Vlok et al. 2003; Cowling et al., 2005). It can be characterized as near-impenetrable vegetation consisting of evergreen, often spiny, small trees and shrubs ranging between 2 – 5m in height that are adorned with numerous vines and lianas. Thicket is rich in succulents and bulbs, many of which are locally endemic (Vlok et al., 2003). It occurs as a solid form where pure thicket covers large areas, and as a mosaic form where it appears as small strips and islands within a matrix of other adjacent vegetation types such as fynbos, karoo, forest and savanna. Different major thicket types are associated with a rainfall gradient from xeric (Arid Thicket) via Valley Thicket, to mesic (Mesic Thicket) (Vlok et al., 2003).

Regeneration of canopy constituents, except for arborescent succulents (Cowling et al. 2009), seem to occur more through resprouting from established long-lived plants rather than seedling establishment (Midgley & Cowling, 1993). Seeds of most canopy thicket plant species are bird-dispersed (Cowling et al., 1997; Sigwela et al., 2009), yet seedlings are seldom encountered, especially in Arid forms (Midgley & Cowling, 1993; Sigwela et al., 2009). Spekboom thicket in this study refers to an Arid Thicket form (250-450 mm MAP) where the succulent tree Portulacaria afra Jacq., locally known as spekboom, is a dominant species (Vlok et al., 2003). Spekboom thicket stores carbon in excess of 200 tons per hectare (measured up to a soil depth of 50 cm), a remarkable feat for a xeric ecosystem and comparable to that of mesic forest ecosystems (Mills et al., 2005a).

Portulacaria afra is a small succulent tree with a characteristic growth form and a unique xeric adaptation to switch between C3 and CAM-photosynthetic pathways (Guralnick, 1984b), high leaf litter production (Lechmere-Oertel, 2008), high rainfall interception capabilities (Cowling & Mills, 2010) and soil-binding properties, easy vegetative reproduction capabilities and is very palatable to both game and stock. Intense browsing pressures applied by goat pastoralists since the early 1900‘s and the characteristic feeding behavior of goats as opposed to that of wild browsers (Stuart-Hill,1992) have today resulted in heavy and moderate degradation of 46% and 36% respectively of the total 16,942 km of area occupied by spekboom- dominated thicket (Lloyd et al., 2002). The patterns and causes of this degradation have been well documented (Hoffman & Cowling, 1990; Stuart-Hill, 1992; Moolman & Cowling, 1994; Kerley et al., 1995; Mills & Fey, 2004, Lechmere-Oertel et al., 2005a; Lechmere-Oertel et al., 2008, Sigwela et al., 2009) and involve the gradual loss of particularly P. afra from the system.

Figure 1.1: A fenceline contrast in spekboom thicket showing intact thicket (behind) and degraded thicket (front). These scenes are common in P. afra dominated thickets as a result of mismanagement by stock farmers. Browsing disturbance opens up paths and gaps within the dense canopy of intact spekboom thicket. With sustained goat browsing, these gaps widen, more spekboom disappears as the browsing pressure exceeds the rate of recovery, causing the gradual deterioration of soil carbon and nutrient content (Mills & Fey, 2004; Lechmere-Oertel et al., 2005b; Mills & Cowling, 2010), and the destruction of canopy, rainfall interception (Cowling & Mills, 2010), microclimate and litter provided by P. afra and other canopy species (Sigwela et al., 2009). Should this degradation continue a hypothetical threshold is crossed where the changes in soil properties and destruction of aboveground microclimate creates an abiotic barrier that prevents effective regeneration of P. afra and other thicket canopy constituent species (Mills & Fey, 2004). This transformation greatly reduces ecosystem carbon storage and carbon losses have been estimated at more than 80 t C ha-1 (Mills, 2003; Mills et al., 2005b). These losses are evident from the decrease in above ground biomass (75% biomass carbon loss), but also manifest in a massive

7 reduction in soil organic carbon (35% soil carbon loss) and nutrient content (Mills & Fey, 2004; Mills et al., 2005b). It has been argued that P. afra plays an important role in facilitating carbon integration into the soil, and thus maintains ecosystem functioning (Lechmere-Oertel et al., 2008). At present, large areas degraded in this way appear as a savanna with a conspicuous absence of P. afra, and dotted with remnant long-lived canopy trees in a matrix of bare soil, karoo shrubs and ephemeral grasses and forbs (locally known as ―opslag‖) that produce an occasional temporary green flush after rainfall events. This ―pseudo-savanna‖ is not a stable state as no recruitment of the canopy trees occurs and the remnant trees produce fewer fruits than those in intact vegetation (Sigwela et al., 2009) and a higher mortality rate of the remnant canopy trees exist when compared to those within an intact spekboom matrix (Lechmere-Oertel et al., 2005a). Consequently, these degraded areas are on a downward path towards a treeless karoo-like landscape without any autogenic mechanisms to drive it back towards its thicket- like state.

Figure 1.2: The distribution of spekboom thicket and the extent of its degradation. Green shows intact spekboom thicket and red shows degraded spekboom thicket areas (Lloyd et al., 2002)

Restoration of Spekboom Thicket

The planting of P. afra truncheons en masse in parallel rows on degraded landscapes once dominated by it has been widely suggested (Mills & Cowling, 2006; Mills et al., 2007; Sigwela et al., 2009) and implemented in preliminary trials as a restoration strategy by the Subtropical Thicket Restoration Programme (STRP) which is administered by the Working for Woodlands Project of the national Department of Water and Environmental Affairs (Marais et al., 2009). The proven potential of sequestering high amounts of carbon in this way (Mills & Cowling, 2006) has opened up the possibility of a large-scale restoration effort that can finance itself through selling carbon credits on international markets while at the same time recovering local ecosystem services and uplifting local communities through job-creation and skills development. Stakeholders have raised concern regarding this restoration methodology, implying that it establishes a monoculture of P. afra, without considering the possible regeneration of other plant species. It has been hypothesized that once P.afra is re-established on a degraded landscape, it will provide the conditions necessary for the autogenic regeneration of plant biodiversity, especially that of other canopy constituent species which shows no recruitment under degraded conditions (Mills & Cowling, 2006; Sigwela et al., 2009). This study aims to test the hypothesis that canopy constituent plant species will regenerate within landscapes that have been subjected to this type of restoration treatment.

In chapter 2 I present the results of an experiment to restore populations of canopy species in degraded spekboom thicket that has been restored with spekboom truncheons (ramets) for different time periods. Krompoort, a site that, starting in 1976, had been planted extensively with P. afra in plots of varying ages by a visionary farmer was selected (Graham Slater, pers. comm.; Mills & Cowling, 2006). Initial data on the species composition of restoration plots at Krompoort suggested very little return of canopy plant biodiversity (unpublished data). I attempted an introduction of four nursery-propagated species, two of which are characteristic canopy constituents, one a canopy margin species, and one a karoo shrub that commonly occurs in in the inter-clump areas of some spekboom- dominated thicket types, by planting these into the restoration sites that have been planted with P.afra at different periods in the past. My aim was to investigate survival and growth of these seedlings and assess whether such restoration methodology is viable when considering seedling survival, restoration costs and overall restoration success.

In chapter 3 I aimed to quantify the actual autogenic regeneration of canopy plant species on sites that have been planted extensively with P. afra and to interpret the restoration of spekboom thicket landscapes in the light of current ecological conceptual models. I show the results of an investigation of carbon sequestration at another site, Rhinosterhoek, that was subsequently found to have an area planted with P. afra that is older than 50 years (Warren Rudman, pers.comm.), making it and an adjacent area planted around 35 years ago, the oldest known areas under restoration using this methodology. I measured regeneration of plant species in these older restored stands at Rhinosterhoek and the younger ones at Krompoort by specifically favouring recruits in the sampling methodology, and comparing the findings i) between restoration treatments of different ages ii) between sites and, iii) aligning these with total carbon sequestration at both these sites and perceived site-specific conditions. Considering all the evidence that has so far been gathered regarding the dynamics of spekboom dominated thicket landscapes, and probing the literature for conceptual ecological models to explain the findings, I interpreted the degradation and restoration dynamics of spekboom- dominated thicket landscapes according to non-equilibrium (alternate stable states) theory and applied the concept of an ecological engineer to Portulacaria afra.

In chapter 4 I assessed all my findings in terms of achieving my research goals and discuss the shortcomings to this study. I then proceed to identify further research potential that this study has laid bare.

Chapter 2

Restoring biodiversity by planting propagules of canopy species in degraded spekboom thicket is neither ecologically nor economically feasible.

Abstract

A significant factor contributing to restoration success is cost-effectiveness as related to realistic restoration targets. I attempted the reintroduction of five nursery propagated plant species in severely degraded Portulacaria afra Jacq. (spekboom) dominated thickets that have been subjected to a restoration method involving the planting of dense rows of P. afra truncheons for periods of 11, 19 and 33 years and also in degraded and intact thickets. Two of these species (Pappea capensis and Searsia longispina) are native canopy constituent species of intact spekboom thickets, one (Lycium ferocissimum) is a canopy-margin species whose relative abundance increases with degradation, and one (Rhigozum obovatum) is a species common in karroid vegetation but also found in inter-clump areas in arid thicket. I also planted nursery propagated P. afra cuttings. An average of 30 propagules of each species, excluding P. afra, were planted in each of the chosen areas. Half of these were planted during the spring of 2008 and the other half during the autumn of 2009. An average of 16 rooted cuttings of P. afra were planted in the same way in autumn 2009. Overall, propagules of the two thicket woody canopy species (S. longispina and P. capensis) showed a total survival of 1% and 9%, respectively. Survival of L. ferocissimum and R. obovatum was 19% and 70%, all propagules of P. afra survived. An inference tree analysis found that the survival of propagules planted in spring 2008 is primarily tied to a species effect, with R. obovatum showing significantly higher survival than the other species which showed a significant preference for higher canopy shade conditions. Propagules planted in autumn 2009 confirmed the significant species effect for survival. Portulacaria afra and R. obovatum exhibited significantly higher survival here, L. ferocissimum showed a significant preference for the degraded, 11 year restored and the intact sites, while the other surviving species planted displayed a significant proclivity to overhanging canopy cover. The results show little significance of restoration treatment for propagule survival. This suggests that a range of conditions is needed for the successful establishment of canopy species, which likely includes a microclimate and substrate created by canopy cover, litter fall and an exceptional series of rainfall events. I found that the high costs involved with a biodiversity planting endeavour, and the low survival of propagules from

11 thicket canopy plant species (P. afra excepted), renders this proposed restoration protocol both ecologically and economically inefficient.

Introduction

The Society for Ecological Restoration International Primer on Ecological Restoration (SER, 2004) lists among its attributes of restored ecosystems i) that a restored ecosystem reflects the distinctive species assemblage of a reference intact ecosystem and the corresponding community structure, ii) the functional groups that sustain the resilience and stability of the system are present, or that at least evidence of their return through colonization should be obtained, and iii) ―the physical environment of the restored ecosystem is capable of sustaining reproducing populations of the species necessary for its continued stability or development along the desired trajectory‖. The reality is that restoration projects rarely reproduce the full species diversity, structure or ecosystem services that reference intact ecosystems contain (Benayas et al., 2009; Ruiz-Jaen & Aide, 2005). Ecosystems are multidimensional and often show non-linear regeneration responses to varying disturbance regimes (Hobbs, 2007; Aronson et al., 2006; Suding et al., 2004; Stephenson, 1999). The difficulty involved with ecological restoration is mirrored by the large proportion of restoration projects that have been unsuccessful (Benayas et al., 2009; Palmer et al., 2006). These failures are attributed to a variety of causes, chief of which is inappropriate restoration methodology due to i) not recognizing possible non-linear ecosystem dynamics (Hobbs, 2007) , ii) insufficient understanding of the causes and extent of the degradation (Jackson & Hobbs, 2009), iii) unrealistic goals or restoration targets or the complete absence thereof (Ehrenfield, 2000). It is increasingly being recommended that restoration projects should have i) well defined goals or targets, ii) be clear on the definition and measures of restoration success and iii) be rooted in ecological theory (Hobbs, 2007; Young et al., 2005; Hobbs & Norton, 1996). The high cost of restoration, which generally increases with the extent of the degradation, is another central limiting factor to restoration success. For this reason the utilization of structures such as carbon credit markets and other payments for ecosystem services, and the involvement of local communities through poverty-relief and job-creation projects are factors that must be considered as part of a multi-faceted approach to restoration (Mills et al., 2007).

The subtropical thicket is an ancient biome that dates back at least to the mid Cenozoic (Cowling et al, 2005). It is characterised by dense, evergreen and mostly spiny shrubs and small trees varying in height of between 2 to 5 metres and adorned with lianas. Succulents, especially tree aloes and euphorbias, are 12 common and often dominant. Spekboom thicket (ST) here refers to the more xeric (250-450 mmyr-1) vegetation types found in subtropical thicket, where the dominant plant, Portulacaria afra Jacq., locally known as spekboom, forms the matrix within which other thicket plant species persist. P. afra is a small succulent tree capable of switching between C3 and CAM-photosynthetic pathways (Guralnick, 1984b), has high leaf litter production (Lechmere-Oertel, 2008), high rainfall interception capabilities (Cowling & Mills, 2010) and soil-binding properties, easy vegetative reproduction capabilities and is very palatable to both game and stock. Intense browsing pressures applied by goats since the early 1900‘s and the characteristic, bottom-up browsing behavior of goats as opposed to the top-down foraging of wild browsers (Stuart-Hill, 1992) have today resulted in heavy and moderate degradation of 46% and 36% respectively of the total 16,942 km of area occupied by spekboom thicket (Lloyd et al., 2002). Spekboom thicket ecosystems have been found to store carbon in excess of 200 tons per hectare (measured up to a soil depth of 50 cm), a remarkable feature for a xeric ecosystem and comparable to that of mesic forest ecosystems (Mills et al., 2005).

Goat browsing opens up paths and gaps within the dense P. afra canopy of intact spekboom thicket as livestock initially target mostly P. afra. With intense sustained and localised goat browsing, these gaps widen, more spekboom disappears as the browsing pressure exceeds the rate of recovery, thereby causing the gradual deterioration of soil carbon and soil nutrient content (Mills & Fey, 2004; Lechmere- Oertel et al., 2005a; Mills & Cowling, 2010). With ongoing destruction of canopy, a decrease in rainfall interception (Cowling & Mills, 2010), microclimate and litter provided by P. afra and other canopy species (Sigwela et al., 2009, Lechmere- Oertel et al., 2008) ensues. Should this degradation continue the changes in soil properties and destruction of aboveground microclimate creates a barrier that prevents effective regeneration of P. afra and other thicket canopy constituent species (Mills & Fey, 2004; Lechmere-Oertel et al., 2005). This transformation greatly reduces ecosystem carbon storage.

Comparisons of degraded and intact sites reveal carbon losses of more than 80 t C ha-1 (Mills, 2003; Mills et al., 2005b). These losses are evident from the decrease in above ground biomass, but also manifest in a massive reduction in soil organic carbon content (Mills & Fey, 2004). At present, large areas degraded in this way appear as a savanna with a conspicuous absence of P. afra, and dotted with remnant long-lived canopy trees in a matrix of bare soil, karoo shrubs and ephemeral grasses and forbs (locally known as ―opslag‖) that produce an occasional temporary green flush after rainfall events. This ―pseudo-savanna‖ is on a degradation downhill as no recruitment of the canopy trees occur and the long-lived remnant trees produce fewer seeds (Sigwela et al., 2009) and have a

13 higher mortality rate than those within an intact spekboom matrix (Lechmere-Oertel et al., 2005a). Consequently, these degraded areas are on a downward path towards a treeless karoo-like landscape without any inherent functionality to return to the previous thicket-like state.

Current restoration methodology for spekboom thicket transformed in this way involves the planting in closely packed parallel rows of spekboom truncheons sourced from surrounding areas that are still relatively intact (Mills et al., 2007). Large-scale planting trails are underway through the initiative of the Subtropical Thicket Restoration Programme (STRP) administered by the Working for Woodlands project of the Department of Environmental and Water Affairs (Mills et al., 2007; Marais et al., 2009). To date, some 1630 ha have been restored, mainly in protected areas in the Eastern Cape Province. These restoration projects were initiated in 2004, thus the oldest plantings are only about 6 years old.

However, older trials exist on privately-owned ranchlands where land managers have planted spekboom to protect degraded slopes from soil erosion. One such site is Krompoort, where the landowner planted fenced off plots with parallel rows of spekboom mostly of a genotype sourced from a mountain near Steytlerville (±90 km away) in successive time intervals ranging from 1976 to 1998 (Graham Slater, pers.comm.). Mills & Cowling (2006) found that 112 tons carbon per hectare - at a rate of 4.2 t C ha-1 yr-1 - were sequestered in a 27 year period at Krompoort. It has been calculated that intact spekboom-dominated thicket stores up to 245 tons carbon per hectare (Mills et al., 2005). Thus, the international carbon market provides an excellent opportunity to finance the restoration of degraded spekboom thicket (Mills et al., 2007; Galatowitsch, 2009; Marais et al., 2009; Palmer & Filoso, 2009).

An important criticism of this restoration methodology is the fact that only one species, i.e. P. afra, is planted, which amounts to the establishment of a monoculture, lacking the full complement of species – and associated functioning - of intact (reference) spekboom thicket ecosystems. Research has shown that in spekboom-dominated landscapes almost all recruits (genets and ramets) are associated with beneath-canopy microsites, irrespective of degradation status (Midgley & Cowling, 1993; Sigwela et al., 2009). It has been hypothesized that once spekboom and its associated shaded, organic-rich micro-climate is returned (Sigwela et al., 2009; Mills & Cowling, 2010) other functional groups representative of intact spekboom thicket landscapes will re-establish via the same colonizing processes as that of intact spekboom thicket (Mills & Cowling, 2006). Colonization processes of thicket canopy constituents occurs through dispersal by vertebrates, particularly birds (Holmes & Cowling, 1993; Cowling et al., 1997; Sigwela, 2004), but the majority of long-lived species (such as Pappea capensis) reproduce 14 mostly by producing ramets (i.e. vegetative reproduction) (Midgley & Cowling, 1993; Sigwela et al., 2009). Preliminary observations of the restored plots at Krompoort indicated a very low regeneration of recruits from spekboom thicket canopy constituents.

Here I investigate the feasibility of restoring populations of three locally occurring spekboom thicket canopy constituent plant species and one karoo-scrub species that commonly occurs in spekboom thicket. I did this by i) planting nursery- propagated seedlings into three plots subjected to P. afra restoration treatment over different time periods, ii) assessing survival and growth of seedlings after one and two-year periods, and iii) calculating the costs of this restoration attempt above that for planting only P. afra. Finally, I discuss with reference to my results the concept of restoration success.

Materials and Methods

Study Site

The farm Krompoort is situated in a valley to the west of the town of Kirkwood on the northern footslopes of the Groot Winterhoek mountains. It falls within the Sundays Spekboom Veld (SSV) vegetation type (Vlok et al., 2003), is situated at an elevation of approximately 320 - 400 m above sea level, and receives a mean annual precipitation of 317 mm per annum (measured between 1970 -2009) with peaks in the autumn and spring months. The geology of the site consists of Bokkeveld Group shales, sandstones and siltstones of the Ceres and Traka Subgroups. Weathered bedrock and transported material are generally the parent material for the soils found on site. Soil forms in the larger study area include Calcaric Cambisols, Calcic Luvisols, Rhodic Luvisols, and Calcaric regosols (FAO, 1998)

In 1976 the land manager at Krompoort fenced off camps and planted P.afra cuttings in rows that are spaced between 4.1 – 1.5 m, with a mean space between plants of 0.8 – 1.1 m (Mills & Cowling, 2006). This procedure was done on a degraded north-facing slope opposite a barn (Graham Slater, pers. comm.). The land manager used a certain P. afra genotype (so-called ―bergspekboom‖) which was sourced close to the town of Steytlerville (±90 km to the northwest) for most of the plantings. This genotype has an upright growth form and a general absence of skirt development, a characteristic of other P. afra genotypes. In addition, one row of a locally occurring P. afra variety was also planted near the bottom of the slope.

The land manager established more blocks planted in this way at later intervals, effectively culminating in restoration blocks of varying ages (Figure 2.2).

I obtained rainfall data for the site spanning the period 1970 - 2009. I analysed the data in order to select appropriate times for planting propagules, specifically with regards to seasonal rainfall peaks. The monthly rainfall at Krompoort showed distinct peaks during the spring and autumn months, however, as represented by outliers, heavy rainfall could be expected in almost all months of the year (Fig. 2.3). Yearly totals recorded for 2008 and 2009 amounted to 213.2 and 269.4 mm, respectively, both less than the mean annual precipitation of 317 mm measured over a 40 year period. These totals reflect the protracted drought conditions experienced in the south coastal regions of South Africa during the course of my study.

Figure 2.1: Location map of the experimental site Krompoort (-33.544690 ; 25.174098) The location of Krompoort is shown in yellow. Green shading represents intact spekboom thicket while red shading represent areas where spekboom thicket landscapes are moderately to heavily degraded (Lloyd et al, 2002).

Figure 2.2: Google Earth© image of Krompoort showing the Intact (cyan), Degraded (pink), Restored33 (= 33 yr-old post restoration) (red), Restored19 (yellow) and Restored11 (blue) sampling sites.

Figure 2.3: Boxplots showing monthly rainfall measured at the Krompoort restoration site over a period of 40 years (1970-2009). Filled circles indicate medians, boxes indicate the lower and upper quartile (i.e. 25% and 75% of the distribution), whiskers indicate the sample minimum and maximum, unfilled circles show outliers.

Nursery Propagation and Planting

I planted a total of 693 individual propagules over a period of two years in five fenced-off stands, three of which were under restoration for 11, 19 and 33 years, in one degraded, and in one intact stand. I selected two intact spekboom thicket canopy species, namely Pappea capensis Eckl. & Zeyh. and Searsia longispina Eckl. & Zeyh., and two species commonly found in moderately impacted and regenerating spekboom thicket, namely Lycium ferocissimum Miers and Rhigozum obovatum Burch. Pappea capensis is a long-lived evergreen tree, up to 5 m in height; S. longispina is a dense, much-branched and rounded evergreen shrub to small tree with spiny branches; L. ferocissimum is a large deciduous and spiny shrub, up to 5 m in height; R. obovatum is a much-branched, spiny shrub or small tree, 1.0-4.5 m high with small grey-green leaves that are drought-deciduous. Rhigozum obovatum is wind-dispersed, while the other three species yield fleshy fruits with seeds typically dispersed by birds. These four species were selected particularly as they i) naturally occur within the area of the experimental site, ii) are representative of a range of ―thicket tolerant‖ species where Pappea capensis and Searsia longispina represent spekboom thicket canopy constituents that are characteristic of intact spekboom thicket canopy species composition and less tolerant to karoo environments. Lycium ferocissimum and Rhigozum obovatum are also species occurring within spekboom dominated thicket, but have a wider distribution within the adjacent karoo vegetation and may be regarded as components representative of earlier successional stages of regenerating spekboom thicket. All species were propagated from seed or cuttings under nursery conditions at the Working for Woodlands nursery near Patensie (Fig. 2.1), 155 km by road from Krompoort. From here on the term propagules will refer to both cuttings and seedlings, unless specific reference is made to the propagation method. All propagules were hardened off prior to planting in the field.

The selected species were propagated according to the following method. For seeds, no dormancy treatment was applied and seeds were directly removed from fruits and sown into sterilized seed trays with a soil mix of 4:2:1 parts of bark : vermiculite : sand respectively. Cuttings of diameter ca.1-2 cm and length 20 cm were harvested from plants adjacent to the nursery. They were treated with a growth hormone and the cut edges on the top were sealed. Cuttings were planted in a soil mix of 4:2 parts of compost and sand, respectively. The propagule trays were kept in 40% shade and watered once a day. Seedlings were removed from the shaded propagation tunnels after a month and placed in full sunlight where they were watered once a day. Portulacaria afra, L. ferocissimum and S. longispina propagules were treated in this way for at least 6 months, R. obovatum

18 for at least 8 months and P. capensis for at least 12 months before being planted (Victoria Wilman, pers. comm.).

I planted the first batch of propagules during the last week of September (spring) 2008. This batch comprised 306 plants that included seedlings of P. capensis and R.obovatum and cuttings of S. longispina and L. ferocissimum. I planted the second batch during May (autumn) 2009 comprising 387 propagules, of which 308 were of the same species as the previous planting plus 79 P. afra rooted cuttings that had been propagated the same way. Table 2.1 outlines the number of propagules planted per treatment. Spring and autumn were chosen as planting periods to coincide with the two annual rainfall peaks recorded for the area. I chose to repeat the planting procedure in autumn as propagules planted at this time were expected to show higher survival since they would not have to contend with high evapotranspiration rates during the hot summer months shortly after being planted (Ehrlinger & Sandquist, 2006). I included P. afra propagules in the second batch of planting to evaluate their survival and growth in comparison to the other species.

September 2008 May 2009 Species Totals Dgrd Rs11 Rs19 Rs33 Intc Dgrd Rs11 Rs19 Rs33 Intc Pappea capensis 15 14 17 15 14 16 13 16 15 16 151 Searsia longispina 16 17 16 14 19 15 15 15 15 15 157 Lycium ferocissimum 15 13 15 15 15 15 15 15 15 17 150 Rhigozum obovatum 15 16 15 16 14 17 15 16 16 16 156 Portulacaria afra - - - - - 15 16 16 16 16 79

All propagules were planted to a depth of about 50 mm deeper and wider than the 4 litre container bags. This was done to enable root expansion and create a trough for the accumulation of water. The propagules were placed randomly within each plot with the help of a GPS device and random grid selection using QGIS software (Quantum GIS Development Team, 2010). All propagules were watered up to and including the day of planting but not thereafter.

Survival and Growth Assessment

I assessed for the survival of propagules separately as per planting time and twice for the initial spring planting, which was i) 3 months after the initial planting in September 2008 (Survival3) where only survival was assessed of a slightly smaller

19 subsample of the total number of propagules planted, ii) 24 months after the initial planting in September 2008 (Survival24) where survival and growth was assessed of all propagules planted. The propagules planted in autumn (May 2009) were assessed for both survival and growth 12 months after the second planting in May 2009 (Survival12) of the total number of propagules planted.

Seedling measurements for growth varied according to propagation method (seedling vs cutting), and growth form. I measured all cuttings by adding the sum of the length of all shoots that are equal to or longer than 50 mm. This minimum length was used to exclude the spines that typically grow from the stems of L. ferocissimum and S. longispina and rarely exceed 50 mm. Owing to its slow growth and relatively small size, P. capensis was measured by the number of green leaves per seedling that were equal to or longer than 5 mm, whereas R. obovatum was measured by canopy diameter (mm2) per seedling. I assessed growth by subtracting the initial parameter measured per individual before or shortly after planting from the measurement taken after the specific growth assessment period (i.e. 12 or 24 months). All measurements were taken with standard measurement tape and a vernier caliper. I assessed canopy shade cover by neighbouring trees and shrubs for each plant according to a five-point scale ranging from 0 (no cover) to 5 (full canopy cover) (Table 2.2).

Table 2.2: Cover scale used for determining a measure of shading for each seedling planted.

Scale Shade Cover 0 No shade cover 1 Shade provided from one direction 2 Shade provided from two directions 3 Shade provided from three directions 4 Shade provided from four directions 5 Full canopy cover - shade provided from all sides and above .

Restoration Costs

I estimated the costs for the propagation and planting of nursery-propagated propagules as per Working for Woodlands guidelines. The costing was separated into propagation costs which were already determined for the nursery (M. McConnachie, pers. comm.) by taking into account the time, labour and other expenses for propagating seedlings and cuttings up to the stage where they are ready to be planted in the field. I based the planting costs on the costs incurred by a typical contractor team to plant rows of P. afra cuttings in a transformed area of 30 ha in Addo Elephant Park during October 2009. Costs include wages, camping, site clearing, transport, and administrative and equipment costs. A typical truckload could transport ±5000 propagules from the nursery to restore an area of 30 ha with

4800 propagules at a rate of 240 plants per hectare with a team of 12 people over a period of 20 working days. The typical running cost of a truck and driver for a distance of 200 km was added to the transport costs of the contractor team. A typical contract for large-scale planting of P. afra lasts generally over a period of 20 working days. The density of planting was based on that calculated for the experiment at Krompoort as described above where 15 propagules were planted in various aged restoration plots, reference and intact plots, each with an average size of 0.25 ha. This amounts to 60 plants per hectare. An average planting rate of 60 propagules per day by a team of three people was observed during the same experiment. The chosen contractor team consisted of 12 people, and the average planting rate was adjusted accordingly. Transport, labour, administrative, equipment, camping costs and capital build-up for planting P. afra were used to estimate the planting costs, and the propagation costs were added for 4800 propagules. Capital buid-up refers to the contractors profit for executing the contract, which is to be used to build up his business, and is calculated at 20% of the wage cost. The cost estimation did not include fencing or any personnel training costs, rendering it somewhat of an underestimation.

Statistical Analysis

Differences in seedling survival and growth were analysed across treatments using the Pearson‘s Chi-square test and Kruskal-Wallis tests. The Kruskal-Wallis test is a non-parametric one-way ANOVA for testing equality of population medians among groups and does not assume a normal distribution. The Chi-square statistic is calculated by finding the difference between each observed and theoretical frequency for each possible outcome, squaring them, dividing each by the theoretical frequency, and taking the sum of the results. Both these analyses are provided by the R package Hmisc (Harrel et al., 2010). A conditional inference tree analysis was applied to each of the plantings executed in autumn and spring. Recursive partitioning or ―trees‖ is a class of simple regression models used for explanation and prediction. In essence it involves a two-stage algorithm that i) partitions observations by univariate splits in a recursive way and ii) fits a constant model in each cell of the resulting partition. Early methods were deemed problematic due to overfitting and a selection bias towards covariates with many possible splits. Hothorn et al. (2006) developed an unbiased recursive partitioning called a conditional inference framework that uses the conditional distribution of statistics. The association between responses and covariates is used as the foundation for an even-handed selection among covariates measured at different scales. Where no significant relation between any of the covariates and the response can be detected using multiple test procedures, the recursion stops. This 21 method is thus a suitable, unbiased, non-parametric model with a well-defined theoretical background.

All statistical analysis was performed with base R (R Development Core Team, 2010) and related packages including lattice (Sarkar, 2008), Hmisc (Harrel et al., 2010) and party (ctree) (Hothorn et al., 2006).

Results

Survival and Growth Analysis

Searsia longispina and P. capensis showed very high mortality across all treatments, without any significant difference between treatments in both the autumn and spring plantings (Table 2.3 and 2.4). The few available growth parameters for surviving P. capensis propagules showed negative growth across all treatments with no significant difference. Table 2.5 shows that Lycium ferocissimum propagules planted in autumn displayed a higher survival in the intact (53%) and degraded (47%) treatments but less in the 11 (33%), and 33 (13%) year restoration treatments; no propagules survived in the 19 year old treatment. The difference across treatments was significant (P=0.005), but suggesting rather more survival success in degraded and intact areas than the restored plots. The spring- planted L. ferocissimum propagules showed very poor survival. The growth parameters for surviving propagules also show no difference across treatments and a general negative growth trend (Table 2.5).

Rhigozum obovatum seedlings planted in autumn showed 100% survival in the intact treatment and an overall survival of more than 50% with some differences across treatments (p = 0.032) (Table 2.7). In both the spring and autumn plantings the highest mortality was recorded for the 33 year old restored site (40% and 50% respectively). P. afra propagules showed a 100% survival across treatments (Table 2.8). The growth parameters indicate that the most growth occurred in the intact, degraded and the 19 year old restored sites; however, negative growth was recorded in the 11 year old restoration site.

Table 2.3: Survival and growth by treatment for Pappea capensis. The number appended to the name of the restored sites show restoration age. Survival3 refers to the survival assessment conducted 3 months after planting in September 2008, Survival12 refers to the survival assessment 12 months after the planting in May 2009 and Survival24 refers to the survival assessment 24 months after the September 2008 planting. Grwth_par refers to the total growth parameters of surviving propagules, irrespective of planting times.

Table 2.4: Survival and growth by treatment for Searsia longispina. The number appended to the name of the restored sites show restoration age. Survival3 refers to the survival assessment conducted 3 months after planting in September 2008, Survival12 refers to the survival assessment 12 months after the planting in May 2009 and Survival24 refers to the survival assessment 24 months after the September 2008 planting. Grwth_par refers to the total growth parameters of surviving propagules, irrespective of planting times.

Table 2.5: Survival and growth by treatment for Lycium ferocissimum. The number appended to the name of the restored sites show restoration age. S urvival3 refers to the survival assessment conducted 3 months after planting in September 2008, Survival12 refers to the survival assessment 12 months after the planting in May 2009 and Survival24 refers to the survival assessment 24 months after the September 2008 planting. Grwth_par refers to the total growth parameters of surviving propagules, irrespective of planting times.

Table 2.6: Survival and growth by treatment for Rhigozum obovatum. The number appended to the name of the restored sites show restoration age. Survival3 refers to the survival assessment conducted 3 months after planting in September 2008, Survival12 refers to the survival assessment 12 months after the planting in May 2009 and Survival24 refers to the survival assessment 24 months after the September 2008 planting. Grwth_par refers to the total growth parameters of surviving propagules, irrespective of planting times.

Table 2.7: Survival and growth by treatment for Portulacaria afra. The number appended to the name of the restored sites show restoration age. Survival12 refers to the survival assessment 12 months after the planting in May 2009.

The conditional inference tree analysis of the autumn seedling survival assessed after 12 months (Survival12) showed primarily a species effect (Fig. 2.4). Lycium ferocissimum seedling survival displays a treatment effect, but contrary to expectations, higher survival in the youngest restoration treatment (11 yr restored site). The few surviving propagules of P. capensis and S. longispina showed a significant preference for higher shade cover (p < 0.001). The conditional inference tree analysis of the survival of propagules planted in spring and assessed after 24 months (Fig. 2.5) reaffirms a primary species affect, again separating R. obovatum (P. afra was only planted once in autumn) from L. ferocissimum, P. capensis and S. longispina. A significant (P< 0.002) preference for canopy shade cover conditions was detected for the few surviving propagules of L. ferocissimum, S. longispina and P. capensis.

Figure 2.5 Results from the conditional inference tree analysis of seedling survival planted in spring and assessed after 24 months. Lyc_fer, Pap_cap, Sea_lon and Rhi_obo are species name abbreviations. The y-axis of the bars shows Y (survived) and N (dead) expressed as a fraction of the total number of propagules.

Restoration costs

The total cost for restoring an area of 30 ha with dense rows of P. afra truncheons during October 2009 in Addo Elephant National Park was R 50,559.60, which amounted to R 1,685.32 ha-1. My estimate for planting 4800 individuals of canopy species consisting of seedlings and cuttings at a ratio of 1:1, in the same area of the same size (30 ha) by the same team, was R 72,251.60 at an average rate of R 2,016.68 ha-1 per species planted at a density of 60 propagules ha-1 (Table 2.8). Should the planting of canopy plant species be accepted as part of an effective restoration protocol after planting P. afra truncheons densely over this 30 ha area, the total restoration cost would amount to R 122,811.20 at a rate of R 3,702.00 ha- 1. The total amount for a combined treatment of planting P. afra rows and the four additional species was 243% higher than just planting P. afra alone.

Table 2.8: Summary of cost estimation for planting an area of 30 ha with nursery propagated cuttings and seedlings (1:1 ratio) by a typical Working for Water contractor team. Capital Build-up refers to the contractor’s profit for completing the contract, and is calculated at 20% of wage cost. The estimation below involves a total number of 4800 propagules planted at a density of 60 plants ha-1species-1.

Propagation Cost R 18,792.00 Planting Costs Transport R 10,384.00 Labour R 19,444.60 Equipment R 10,191.80 Admin R 1,230.28 Camping R 8,320.00 Capital Build-up R 3,888.92 Subtotal Planting Costs R 53,459.60 Total R 72,251.60

Discussion

When considering propagule survival, the concept of species tolerance to a variety of biotic and abiotic stressors is key. Abiotic conditions pose ecophysiological challenges to establishing plants that can generally be divided into i) light and energy relations in above-ground processes and ii) water and nutrient relations in belowground processes (Ehleringer & Sandquist, 2006). In this experiment, the introduction of species other than Portulacaria afra into landscapes under restoration treatment of 11, 19 and 33 years failed for three of the four species attempted, whereas zero mortality was recorded for P.afra propagules. Contrary to expectations there was no treatment effect detected and thus survival was not

27 significantly different between the various aged restored sites nor between the restoration sites and the reference (intact) and control (degraded) sites. A possible exception is L. ferocissimum where the surviving few propagules preferred the degraded, intact and younger (11 year-old) restored sites. The low total survival percentages of S. longispina (1%), P. capensis (3%) and L. ferocissimum (19%) propagules strongly suggest that failure may be ascribed to abiotic conditions that are outside the tolerance range of these propagules. Biotic conditions such as root competition may also play a role when considering the significant difference in survival of L. ferocissimum propagules planted in the degraded and intact stands as opposed to those planted in stands under restoration. R. obovatum, which is a wind-dispersed species and not a spekboom thicket canopy constituent, has shown an overall survival success of 70%. This may be attributed to its greater tolerance towards semi-arid karoo-like conditions as its natural distribution testifies. The growth parameters show positive growth that is significantly higher in those propagules that were planted in spring and assessed after 24 months than that of propagules planted in autumn and assessed after 12 months, proving that these propagules survive well even under extreme summer conditions. Lycium ferocissimum, which showed marginal survival results, is also considered a more karoo-tolerant species that is not just associated with spekboom-dominated thickets but also with other arid vegetation types. P. afra, which showed 100% survival across all treatments, is a succulent species which has unique ecophysiological adaptations, such as switching from CAM to C3 photosynthetic pathways as a response to water limitations and extreme climatic conditions (Guralnick & Ting, 1987; Guralnick et al., 1984).

The fact that the majority of propagules that survived after a minimum time of 12 months in the soil showed little affinity to treatment, even in the case of the older restored and intact sites, should be regarded in the light of the nature of the restoration experiment and spekboom-dominated thicket dynamics. The profusion of seeds produced by canopy constituent species under intact conditions allows for widespread dispersal, and should the right conditions develop, propagules will establish and grow. It is an open question how often this happens, even in intact conditions, and if these seedlings stay seedlings for extended periods of time until some form of disturbance allows them to become adult canopy constituents. Considering that ramets rather than seedlings are more responsible for recruitment of woody canopy constituents in spekboom-dominated thickets (Midgley & Cowling, 1993), it is probably not often. From the available evidence, the right conditions for seedling and ramet establishment in spekboom-dominated thicket can be characterized as i) an aboveground microclimate established by plant canopy cover (Sigwela et al., 2009) and a mulch rich substrate provided by leaf litter (Lechmere-Oertel et al., 2008), ii) fertile soils linked with organic carbon

28 content, nutrients and soil water retention (Lechmere-Oertel et al., 2005b; Mills & Cowling, 2010; Mills & Fey, 2004), and iii) periods of sufficient follow-up rainfall events. It is likely that high evapotranspiration rates, especially during the summer months and the generally low precipitation are main limiting factors for seedling survival across treatments, and that species most tolerant to these conditions survived well, e.g. P. afra and R. obovatum. Within a restoration setting where the planting of propagules are involved the stress on individuals through transportation and physical planting is also a significant factor.

Restoration is costly. The idea held by some conservationists and economists of ecological restoration as an expensive diversion and a waste of time and money (Aronson et al., 2006) is reinforced when considering the relative rarity of achieving restoration success and the costs involved in attempting it (Palmer et al., 1997; Benayas et al., 2010). Planting P. afra truncheons in closely spaced rows is increasingly being considered a viable restoration methodology for degraded spekboom thicket landscapes (Mills et al 2007). Funding for restoration projects is generally limited and thus mechanisms for financing restoration through the selling of carbon offset credits on international markets (Galatowitsch, 2009; Palmer & Filoso, 2009) are invaluable for achieving restoration on a large scale (Mills et al 2009). The potential for earning carbon credits from planting P. afra to finance restoration costs and motivate restoration practice on a large scale (Mills et al., 2003; Marais et al., 2009) is promising, but requires strict management of costs to be a viable land-use option for degraded spekboom dominated landscapes. The high survival rate (100%) of P. afra propagules proves that this dominant spekboom thicket canopy species thrives in both degraded and intact landscapes.

To keep restoration costs reasonable, large-scale planting operations of P. afra truncheons is more desirable. The estimated costing of what it would cost to restore an area of 30 ha i) with densely packed rows of P. afra truncheons and ii) following on that with four canopy plant species is 243% higher than to restore the same area with P. afra truncheons in closely spaced rows alone. A cost estimate of a combination of the two treatments over a 30 ha area amounts to R 122,811.20 of which 41% is spent on establishing P. afra truncheons, and the rest on attempting the establishment of only a fraction of the full complement of canopy species occurring in spekboom dominated thickets. At a total survival percentage of 19% (L. ferocissimum), 9% (P. capensis), and 1% (S. longispina) and a general negative growth trend of surviving propagules, this attempt at the re-establishment of biodiversity of canopy species is inefficient and a waste of limited resources.

This restoration approach does not promote restoration success as defined by the Society of Ecological Restoration International that requires, among other conditions, that the physical environment of the restored landscape to be capable 29 of accommodating reproducing populations of key species (SER, 2004). The possibility exists that more propagules would survive should they be watered regularly. It should be realized that with the already high cost of propagation and planting and the high mortality rate of propagules, devising a system to water individual propagules on a regular basis would increase the restoration costs significantly, and better survival would not be guaranteed. The considerably better survival of R. obovatum does not influence this state of affairs as it is not strictly a spekboom thicket canopy constituent. This species may be valuable as a nurse plant to prevent heavy utilization of spekboom truncheons that have been planted in unfenced areas. Similarly, even though L. ferocissimum does not show a high survival rate, its red fleshy fruits may be useful for attracting birds, the prime dispersal agents of woody thicket canopy constituents, to a degraded area under P. afra restoration. The fact remains that the physical environment of the stands under restoration are still incapable of sustaining reproducing populations of woody canopy thicket constituents and adding some of these species artificially does not solve the problem.

It is clear from the literature that the focus of restoration efforts should in many cases be towards overcoming possible thresholds and hysteresis effects by restoring ecological function, rather than the full range of native species (Hobbs, 2007; Byers et al., 2006). In this regard attention should be directed towards taking account of the role of individual species, particularly those that play disproportionately strong roles in altering the physical environment, and those that are strong interactors with other species (Palmer et al., 1997). In the case of degraded spekboom thicket, all restoration efforts should be focused on producing the right environmental conditions, so that when adequate rainfall does occur, spontaneous recruitment allows for the return of woody canopy trees. It is argued that this is best achieved by planting P. afra truncheons to gradually alter above- and belowground environments (Mills & Cowling, 2006). In chapter 3 I investigated landscapes densely planted with P. afra for periods ranging between 11 and 50 years and demonstrate the spontaneous return of canopy species diversity on landscapes restored using only this restoration protocol.

Chapter 3

Biodiversity Regeneration and Carbon Sequestration on Restored Spekboom Thicket Landscapes – Portulacaria afra Jacq. as an Ecological Engineer.

Abstract

Restoration success involves, inter alia, the autogenic regeneration of key species or functional groups within the degraded ecosystem. Heavily degraded spekboom- dominated thicket does not spontaneously regenerate its former canopy species composition and I interpreted this state of affairs in terms of a state-and-transition conceptual model. I investigated the autogenic return of thicket canopy species biodiversity to degraded spekboom thicket under a restoration treatment that entails the planting of densely spaced rows of Portulacaria afra Jacq. (spekboom) truncheons. I sampled the floristics of a degraded, an intact (reference) and a range of stands that have been subjected to this restoration treatment for varying time periods at each of two locations, 25 km‘s apart, in Sundays Spekboomveld. The stands sampled at Krompoort have been planted with an imported P. afra genotype for periods of 11, 19 and 33 years respectively, and carbon content data was obtained for these sites and that of intact sites from existing studies. The stands under the same restoration treatment (with a local P. afra genotype) sampled at Rhinosterhoek were 35 and 50 years old, and I determined the carbon content of the above- and belowground components for these stands. Recruits of canopy species were favoured in the floristic sampling methodology and data analysis, and used as indicators of intact and successfully regenerating spekboom thicket stands. A multivariate correspondence analysis and a diversity index incorporating taxonomic distance revealed that the sites under restoration are progressively regenerating canopy species biodiversity with increasing restoration age, and that intact sites are still the most diverse. The high total carbon content (TCC) measured within the older restored stands Rhinosterhoek (241 t C ha-1 after 50 years up to a depth of 50 cm) rivals that recorded for intact spekboom thickets, and similarly the number of recruits found within older restored sites rivals the intact (reference) sites sampled. The steep increase in ecosystem carbon within sites under this restoration treatment with increasing age and the corresponding functional characteristics of P. afra that effects carbon build-up and other related changes in the above- and belowground environments potentially identify P. afra as an ecosystem engineer within spekboom dominated thickets that allows for the

31 theoretical threshold to be transcended. The return of canopy plant biodiversity was found to be directly related to the build-up of carbon in the soil and the accompanying changes in soil quality and the unique microclimate aboveground and mulch-rich substrate that the canopy of restored P.afra provides. This restoration methodology is accordingly considered efficient and autogenic canopy species return was found to be prominent after a period of 35-50 years of restoration treatment.

Introduction

Depending on the nature and extent of the degradation, ecological restoration targets should ideally incorporate some measure of return of species composition, ecosystem services (such as nutrient cycling and carbon sequestration), and ecological functioning (Hobbs & Harris, 2001; Hobbs, 2007). Owing to a range of challenges involved with ecological restoration, it has been argued that restoration success should not only be measured by the full return of historical species composition to a restored ecosystem (Ehrenfield, 2000), but also by measures mindful of (i) the cause and nature of the degradation (Hobbs & Norton, 1996; Hobbs et al., 2007) (ii) the corresponding ecosystem dynamics at play (Palmer et al., 1997) and (iii) current and future climate change scenarios (Harris et al., 2006). The return of species composition, measurable ecosystem services and ecosystem functioning, when compared with that occurring on intact reference sites (White & Walker, 1997), provide appropriate yardsticks for the evaluation of restoration success (Benayas et al., 2010).

Subtropical thicket vegetation has only relatively recently been recognized as a biome separate from the adjacent karoo, savanna and forest biomes (Vlok et al., 2003; Cowling et al., 2005). It can be characterized as near-impenetrable vegetation consisting of evergreen, often spiny, small trees and shrubs ranging between 2–5 m in height that are adorned with numerous vines and lianas. Thicket is rich in succulents and bulbs, many of which are locally endemic (Vlok et al., 2003). It occurs as a solid form where pure thicket covers large areas, and as a mosaic form where it appears as small strips and clumps within a matrix of other adjacent vegetation types such as fynbos, karoo, forest and savanna. Different major thicket types are associated with a rainfall gradient from xeric (Arid Thicket) via Valley Thicket, to mesic (Mesic Thicket) (Vlok et al., 2003).

Regeneration of canopy constituents, except for arborescent succulents (Cowling et al., 2009), seem to occur more through resprouting from established long-lived plants (ramets) than seedling establishment (genets) (Midgley & Cowling, 1993).

Propagules of most canopy thicket plant species are fleshy and dispersed by vertebrates, mainly birds but also ungulates (Cowling et al., 1997; Sigwela, 2005; Sigwela et al., 2009). Spekboom thicket, the focus of this study, refers to an Arid Thicket form (250-450 mm MAP) where the succulent tree Portulacaria afra Jacq., locally known as spekboom, is a dominant species (Vlok et al., 2003). Spekboom thicket stores carbon in excess of 200 tons per hectare (measured up to a soil depth of 50 cm), a remarkable feature for a xeric ecosystem, and comparable to that of mesic forest ecosystems (Mills et al., 2005).

P. afra is a small, evergreen succulent tree (2-5 m in height and up to 5 m in width) with a dense canopy that forms a skirt of rooted branches at the base.

Physiologically, it is capable of switching between C3 and CAM-photosynthetic pathways with the onset of soil moisture stress (Guralnick, 1984b). Leaf litter production is extraordinarily high and contributes the bulk of the input to soil carbon (Lechmere-Oertel et al., 2008). Other features that contribute to high soil carbon are the extraordinarily high rainfall interception of the canopy (Cowling & Mills, 2010) and low beneath-canopy temperatures relative to ambient temperatures outside (Lechmere-Oertel et al., 2008). Furthermore, P. afra has excellent soil-binding properties, easy vegetative reproduction capabilities and is very palatable to both game and stock.

Intense browsing pressures applied by goat pastoralists since the early 1900‘s, and the characteristic feeding behavior of goats as opposed to that of wild browsers (Stuart-Hill,1992) have today resulted in heavy and moderate degradation of 46% and 36% respectively of the total 16,942 km2 of area occupied by spekboom-dominated thicket (Lloyd et al., 2002). The patterns and causes of this degradation have been well documented (Hoffman & Cowling, 1990; Stuart- Hill, 1992; Moolman & Cowling, 1994; Mills & Fey, 2004, Kerley et al., 1995; Lechmere-Oertel et al., 2005a; Lechmere-Oertel et al., 2005b; Lechmere-Oertel et al., 2008; Sigwela et al., 2009). Browsing disturbance opens up paths and gaps within the dense canopy of intact spekboom thicket. With sustained browsing, these gaps widen, more spekboom disappears as the browsing pressure exceeds the rate of recovery, causing the gradual deterioration of soil carbon and nutrient content (Mills & Fey, 2004; Lechmere-Oertel et al., 2005a; Mills & Cowling, 2010). With the destruction of canopy, rainfall interception is reduced (Cowling & Mills, 2010), the beneath-canopy microclimate is changed and litter production by P. afra and other canopy species is reduced substantially (Lechmere-Oertel et al., 2008).

Along this degradation trajectory, an abiotic threshold is likely crossed where the changes in soil properties and microclimate creates a barrier that prevents effective regeneration of P. afra and other thicket canopy constituent species

(Mills& Fey, 2004). This transformation greatly reduces ecosystem carbon storage and carbon losses of more than 80 t C ha-1 between degraded and intact sites

Figure 3.1: Portulacaria afra Jacq. (spekboom) tree. Notice the skirt of rooted branches around the base of the tree. have been reported (Mills, 2003; Mills et al., 2005). These losses are evident from the decrease in above ground biomass, but also manifest in a massive reduction in soil organic carbon (Mills & Fey, 2004). It has been argued that P. afra plays an important role in facilitating carbon integration into the soil, and thus maintains ecosystem functioning (Lechmere-Oertel et al., 2008). At present, large areas degraded in this way appear as a savanna with a conspicuous absence of P. afra, and dotted with remnant long-lived canopy trees in a matrix of bare soil, karoo shrubs and ephemeral grasses and forbs (locally known as ―opslag‖) that produce an occasional temporary green flush after rainfall events. This ―pseudo-savanna‖ is not a stable state as no recruitment of the canopy trees occurs. Furthermore, remnant trees produce fewer fruits than those in intact vegetation (Sigwela et al., 2009) and many eventually die, starved as they are of the protective spekboom matrix (Lechmere-Oertel et al., 2005a). Consequently, lacking autogenic mechanisms to drive the system back towards a thicket-like state, these degraded

34 areas are on a downward path towards a treeless, karoo-like landscape (Hoffman & Cowling, 1990) (Fig 3.1).

This system behavior suggests dynamics similar to the state-and-transition model based on rangeland ecosystems research (Westoby et al., 1989; Stringham et al., 2003; Briske et al., 2006) or the alternative stable states model (Lewontin, 1969; Rietkerk & Van de Koppel, 1997; Beisner et al., 2003; Suding et al., 2004; Schröder et al., 2005), with which the traditional succession type models for characterizing ecosystem recovery after disturbance events are being revised where applicable (Hobbs & Norton, 2004). An ecological threshold occurrence may

Transition threshold controlled by Transition threshold controlled by Level of system biotic interactions abiotic limitations function/health

Fully State 1 functional

State 2

Sustainability State 3 requires Recovery requires responsible modification of State 4 browsing the physical pressures environment

Recovery requires State 5 the long-term removal of goat- State 6 Non- browsing pressure functional

Intact Degraded Ecosystem state Figure 3.2: Hypothetical state-and-transition model, adapted from Hobbs & Norton (2004) to describe spekboom thicket degradation (by goat pastoralism) and recovery dynamics. Shown are alternative states (states 1-6) that can manifest along a gradient of browsing pressure intensity and duration (left to right) that is the cause of increasing degradation. States 1 and 2 show possible alternating states of the system where browsing pressure ≤ autogenic recovery under the established browsing regime. The biotic threshold (green bar) symbolizes the point after which no autogenic return to states 2 and 1 are possible without removing the browsing pressure. The system may at this stage exhibit a number of new alternating states (states 3 and 4). Should the browsing pressure continue unabated, the system moves across the abiotic threshold (red bar) where abiotic factors limit any autogenic recovery potential and the system degenerates towards a karoo-like landscape (State 6).

35 be described as ―a switch from the dominance of negative feedbacks that maintain ecosystem resilience to the dominance of positive feedbacks that degrade resilience and promote the development of post-threshold states…‖ (Briske et al., 2006). This process characterize the switch from intact spekboom-dominated thickets towards the pseudo-savanna states of the degraded forms where ecosystem resilience had been severely compromised through injudicious past goat-browsing and the inherent mechanisms for regeneration of thicket species have collapsed. The concept of a species as an ecological engineer that ―creates, modifies, and maintains habitat by causing physical state changes in abiotic and biotic materials and thus govern the accessibility of resources to other organisms within the system‖ (Jones et al., 1994; 1997) is increasingly being applied to particular restoration scenarios to transcend perceived thresholds and establish the dominance of negative feedbacks that promote ecosystem resilience (Byers et al., 2006).

It has been suggested that by planting P. afra truncheons close together in multiple rows on heavily degraded spekboom thicket landscapes, some form of recovery may be achieved (Mills & Cowling, 2006; Lechmere-Oertel et al., 2008). Landscapes being restored in this way have sequestered 70 t C ha-1 after 27 years, making this methodology a viable restoration method that can potentially earn carbon credits on international carbon markets to finance large scale restoration initiatives (Mills & Cowling, 2006). Stakeholders have, however, raised concern regarding this restoration methodology, implying that it only establishes a monoculture of P. afra, without considering the possible regeneration of other plant species. It has been hypothesized that once P. afra is re-established on a degraded landscape, it will provide the conditions necessary for the autogenic regeneration of plant biodiversity, especially that of other canopy constituent species which show no recruitment under degraded conditions (Lechmere-Oertel et al., 2005b; Mills & Cowling, 2006; Sigwela et al., 2009).

In this chapter, I test this hypothesis by examining the composition of plant species diversity on a range of spekboom thicket sites under restoration, and compare them with intact and degraded sites. I apply the concept of ecological thresholds on the degradation and restoration dynamics of the system by identifying P. afra as an ecological engineer capable of transcending the post- threshold state and allowing for autogenic regeneration of canopy thicket species. I examine the possible associations of plant species composition and ecosystem function with above- and belowground carbon build-up in a range of degraded, restored and intact sites. I ask if the build-up of carbon within the system that P. afra facilitates allows for a switch across the perceived threshold by establishing i) an aboveground canopy microclimate (Sigwela et al., 2009) and thick litter layer

(Lechmere-Oertel et al., 2008) and ii) a substantial change in organic soil carbon content and associated soil fertility (Mills & Fey, 2004; Mills & Cowling, 2006; Mills & Cowling, 2010), and thus creating a steadily recovering functional ecosystem.

Methods

Site Description

My primary research site is the farm Krompoort, which is situated in a valley to the west of the town of Kirkwood on the foot slopes of the Groot Winterhoek mountains (Fig. 3.2). It falls within the Sundays Spekboom Veld (SSV) vegetation type (Vlok et al., 2003), is situated at an elevation of approximately 320 - 400 m above sea level, and receives a mean annual precipitation of 317 mm per annum (measured between 1970 - 2009) with peaks in the autumn and spring months. The dominant underlying geology of the site consists of Bokkeveld Group shales, sandstones and siltstones of the Ceres and Traka Subgroups of the Cape Supergroup. In 1976

Figure 3.3: Location of study sites shown in yellow. The green represent spekboom thicket areas still fairly intact, the red represent areas where spekboom thicket landscapes are moderately to heavily degraded (Lloyd et al, 2002).

Figure 3.4: Krompoort site showing the Intact (cyan), Degraded (pink), Restored33 (red), Restored19 (yellow) and Restored11 (blue) sampling sites. The numbers refer to age of restoration.

the landowner started planting P. afra cuttings in closely spaced rows on a degraded north-facing slope. For most of the plantings, he used a certain P. afra genotype (so-called ―bergspekboom‖) which was sourced close to the town of Steytlerville, ±90 km to the northwest of the study site. This genotype has an upright growth form and a general absence of skirt development which is characteristic of most mature P. afra plants. In addition one row of a locally occurring P. afra variety was also planted near the bottom of the slope. The landowner planted and fenced-off more blocks in this way at later intervals, effectively culminating in restoration blocks of varying ages (see Table 3.4 and Fig. 3.4). Mills & Cowling (2006) determined total carbon stocks of the different restoration plots, and these data were used in this study.

A second study site, the farm Rhinosterhoek, is situated approximately 25 km to the northwest of Krompoort (Fig. 3.2) and also falls within Sundays Spekboom Veld (Vlok et al., 2003). It receives a mean annual precipitation of 265 mm (measured between 1950 and 2008) with peaks in the spring and autumn months. It is situated at an altitude of between 480 – 490 m above sea level. The dominant underlying geology of the site consists of Bokkeveld Group shales of the Traka Subgroup of the Cape Supergroup. The landowner planted - in degraded thicket -

38 one row of P. afra, approximately 60 m long on a gentle north-facing hill behind the farmhouse, perpendicular to the slope (Fig. 3.3). A photo taken in 1960 already shows the row of planted P. afra, making it at least 50 years old, and effectively the oldest known planted P. afra population. Subsequently, in the mid 1970‘s, more P. afra cuttings were planted in adjacent rows in the space in front of the established older row (Warren Rudman, pers. comm.). This second planting is estimated at approximately 35 years old. For both Krompoort and Rhinosterhoek degraded and intact stands were selected on north-facing slopes nearby to act as control and reference sites, respectively. The intact site selected at Rhinosterhoek was the nearest one available and is situated approximately 3 km to the south-east of the restored sites.

Data collection

Carbon

Sampling for above- and belowground carbon was carried out at Rhinosterhoek on the 35 year-old restored site (Restored35), the 50 year-old restored site (Restored50), and on a degraded (Degraded) site nearby. Belowground sampling involved the excavation of successive soil depth intervals, each 20 cm deep, up to a total depth of 1 meter where possible, depending on the depth of the bedrock (in many instances bedrock was exposed at a depth < 1m)(see Table 3.1 for summary). Consequently only the first three depth intervals were used (up to 60cm) for analysis. Within each 20 cm depth interval three horizontally adjacent samples of soil (20 cm x 10 cm x 10 cm each) were excavated after all litter and above-ground material were carefully removed. Each of these samples was bagged, and the excavated sampling cube was filled with pre-weighed river sand of known bulk density. After all samples (three per depth interval) were air-dried, they were weighed and sieved (<2mm); one of these sieved soil samples was sent to a laboratory for analysis of carbon content using the Walkley-Black method (Walkley, 1947). Another was sorted to determine the total weight and volume (by water displacement) of rock within the sample, and another was used to separate the total root biomass. Extracted root fragments were weighed and then dried in an oven at 60°C until constant mass. The carbon content of root biomass was calculated by multiplying the dry material mass by 0.48 (Lamlom & Savidge, 2003; Mills & Cowling, 2006).

Seven such replicate sample excavations were done within each site sampled. The 50 year-old restored site was completely under canopy cover and the samples were taken at varying distances from the main stems of living trees. Three of the

Table 3.1: Belowground carbon sample (n) sizes per depth interval at Rhinosterhoek

Sample size (n) per Depth Interval Sampling site 1st 2nd 3rd 4th 5th (0:20 cm) (20:40 cm) (40:60 cm) (60:80 cm) (80:100 cm) Degraded 7 7 6 6 4 Restored35 7 7 4 0 0 Restored50 7 7 7 2 0

Restored35

Restored50

Intact

Degraded

Figure 3.5: Rhinosterhoek site. Top shows the spread of the four sampling sites, while bottom left shows the Restored50 (red), Restored35 (yellow) and Degraded (pink) sites, bottom right shows the Intact (cyan) sampling site. The numbers refer to age of restoration. 40

samples taken at the 35 year- old restored site were taken from under spekboom canopy cover and four were taken outside canopy cover. No sampling pits were dug under vegetation cover on the degraded site, which was dominated by low karroid shrubs and ephemerals. The aboveground carbon sampling involved the collection of aboveground litter in a 25cm x 25cm frame. The litter samples were dried in an oven at a temperature of 60°C until it reached a constant mass. Seven such samples were taken within each of the three sampling sites.

The carbon content of the above ground living biomass was estimated using species-specific allometric equations developed by Powell (2009) (Table 3.2). Depending on the species measured, either the canopy area (CA) or cumulative basal stem diameter (CBSA) was measured as explanatory variables for the allometric equations. The species for which no allometric equations were developed was assessed using other species of similar growth form and habit for which regression equations existed (Table 3.3).

Table 3.2: Allometric regression equations used for aboveground carbon sampling

Species Formula (Powell, 2009) 2 Portulacaria afra Log10 y (C(kg) = 1.1043(Log10 CBSA(m )) + 2.4464 * 2  Grewia robusta Log10 y (C(kg) = 1.0044(Log10 CA(m )) - 0.6259 2 Lycium ferocissimum Log10 y (C(kg) = 0.8615(Log10 CBSA* (m )) + 1.7706 * 2  Pappea capensis Log10 y (C(kg) = 1.3355(Log10 CA(m )) + 0.1357 2 Plumbago auriculata Log10 y (C(kg) = 1.0821(Log10 CBSA(m )) + 2.7320 * 2  Pteronia incana Log10 y (C(kg) = 1.4032(Log10 CA(m )) - 0.4224 2 Putterlickia pyracantha Log10 y (C (kg) = 1.0622(Log10 CBSA(m )) + 2.7834 * 2  Searsia longispina Log10 y (C (kg) = 1.1012(Log10 CA(m )) - 0.2938 2 Jatropha capensis Log10 y (C (kg) = 0.9067(Log10 CA(m )) - 0.7349 *CBSA = Cumulative Basal Canopy Diameter CA = Canopy Area

The soil carbon data were analyzed using a regression model (see Fig. 1, Appendix B) to get a reasonable estimate for the total carbon content (TCC) up to a depth of 50 cm in order to compare these data with results gathered from other studies in spekboom thicket (Mills et al., 2005a, Mills & Cowling, 2006). Mills & Cowling (2006) sampled the total carbon content of the restoration blocks at Krompoort and calculated an annual rate of carbon accrual. These data were used to adjust the carbon content of the sites they sampled according to the time lapse between the time of sampling for that study (2003) and the time of sampling (2009) for this study. Data on carbon content were used as explanatory variables in multivariate analyses of the full floristic data for each state (intact, degraded, restored) at both sites.

Table 3.3: Species allometric equivalents used.

Species Guild-based equivalent All Asparagus species Putterlickia pyracantha Lycium cinereum Lycium ferocissimum Helichrysum rosum Pteronia incana Pollichia campestris Pteronia incana Carissa bispinosa Putterlickia pyracantha Tephrosia capensis Pteronia incana Gymnosporia capitata Putterlickia pyracantha Searsia pterota Searsia longispina Sarcostemma viminale Jatropha capensis Azima tetracantha Putterlickia pyracantha Gymnosporia capitata Putterlickia pyracantha Pentzia incana Pteronia incana Rhigozum obovatum Lycium ferocissimum Crassula tetragona Jatropha capensis Cadaba aphylla Grewia robusta Maireana brevifolia Jatropha capensis Felicia species Pteronia incana Duvalia pillansii Jathropa capensis Limeum aethiopicum Pteronia incana

Floristics

At each of the sampling locations (Krompoort and Rhinosterhoek), areas that had been planted with P. afra truncheons for 11, 19, 33, 35 and 50 years at the time of sampling were identified according to Mills & Cowling (2005) and Mr. Warren Rudman (pers. comm.), together with nearby intact and degraded reference sites (see Fig. 3.4 and Fig. 3.5). Table 3.4 outlines the different treatments per sampling site. Within each of these treatments ten 3m x 3m quadrats were randomly placed using QGIS software (Quantum GIS Development Team, 2010) and a GPS device. The 50 year old restored site at Rhinosterhoek posed a problem as the area suitable for sampling is only approximately 280 m2 in size. Here only 7 quadrats, placed adjacent to each other, were used as sampling units. On all sites, all plants rooted within and overlapping each quadrat were recorded in terms of its species name, height and cover. Table 3.5 shows the cover scale used.

Where thicket canopy plants were identified, namely Putterlickia pyracantha, Rhigozum obovatum, Searsia longispina, Searsia pterota, Tarchonanthus minor, Grewia robusta, Gymnosporia capitata, Lycium ferocissimum, Lycium oxycarpum, Pappea capensis, Azima tetracantha, Carissa bispinosa, Crassula ovata, Ehretia rigida and Euclea undulata, an arbitrary distinction was made between plants with

42 a height exceeding 1 m (adult plants) and plants with a height equal or lower than 1 m (recruits). Recruits included individuals established vegetatively from ramets

Table 3.4: Outline of different treatment sites sampled

Site Treatment Abbreviated Year Restoration Approximate name planted Age (in Size (m2) 2009) Krompoort Degraded K.Dgrd n/a 0 2 500 Krompoort Restored11 K.Rs11 1998 11 years 2 500 Krompoort Restored19 K.Rs19 1990 19 years 2 500 Krompoort Restored33 K.Rs33 1976 33 years 2 500 Krompoort Intact K.Intc n/a Intact 2 500 Rhinosterhoek Degraded R.Dgrd n/a 0 2 500 Rhinosterhoek Restored35 R.Rs35 ca 1974 ca 35 years 830 Rhinosterhoek Restored50 R.Rs50 ca 1959 ca 50 years 280 Rhinosterhoek Intact R.Intc n/a Intact 2 500

as well as genets (seedlings) (Midgley & Cowling, 1993). I treated recruits (R) differently in the analysis from adult plants. They were allocated to a different - or pseudo - species. Recruits were also allocated a different cover scale from adults. This provides the recruits with more weight in the statistical analysis - thereby enabling a better assessment of regeneration status in the different treatments. Bare ground (B) was assigned a cover value according to the scheme used for adult plants (S). I assessed the cover estimation for adult plants using a seven- point scale as shown in Table 3.5. The abundance measure used for the recruits was determined by finding the sampling unit (quadrat) with the most recruits (34) of a single species (P. capensis) and then assigning this number as the upper limit of the second highest possible cover scale (6), and dividing it into the lower scales, leaving the highest scale undeterminedly higher (see Table 3.5).

Table 3.5: Cover scale used for Adult plants and Recruits (“pseudo-species”)

Adult plants (S) Recruits (R) Scale (cover) (abundance) 1 0-2% 1 2 2-5% 2-5 3 5-10% 6-10 4 10-25% 11-15 5 25-50% 16-25 6 50-75% 26-34 7 75-100% 35+

Statistical analysis

Carbon analysis

A linear mixed effects model was applied to the belowground carbon data from Rhinosterhoek, in order to investigate treatment and sample depth interval (each individual soil layer sampled up to 60cm), and treatment and sample depth (sum of individual soil layers sampled at increasing depths). In both cases these variables were treated as fixed effects. The response variables were in each case: soil carbon content, root carbon content and belowground carbon content (soil carbon + root carbon). An ANOVA table was generated and t-test contrasts was conducted where necessary. The R packages nlme (Pinheiro et al., 2010), multcomp (Hothorn et al., 2008) and contrast (Kuhn et al., 2010) was used for these analyses.

Multiple correspondence analysis

A multiple correspondence analysis was performed on the floristic data. This multivariate method allows for the interpretation of categorical data by mapping objects and variables and their categories onto a shared metric plane. The R package homals (De Leeuw & Mair, 2009) provides a multiple correspondence analysis (homogeneity analysis) of multivariate data according to techniques developed by Gifi (1990). Homogeneity analysis in a broad sense ―refers to a class of criteria for analyzing multivariate data in general, sharing the characteristic aim of optimizing the homogeneity of variables under various forms of manipulation and simplification‖ (Gifi, 1990). The advantage of this approach is that it allows for a homogeneity analysis which corresponds to a multiple correspondence analysis, but can take account of the scale level of variables and handle missing data. It was applied to the dataset where the recorded variables of cover, average height and total carbon content (TCC) were treated as numeric variables and (restoration) treatment age as ordered factors. The results were plotted with the ade4 package (Chessel et al., 2004).

Rao’s diversity coefficient

I used Rao‘s diversity coefficient, also known as quadratic entropy, to measure plant diversity in the restored , degraded and intact states at both sites. Taxonomic and other differences (such as functional difference, morphology, feeding behavior, ecological guild) among species are not reflected in commonly used diversity

44 indices (e.g. the Shannon-Wiener index and the Gini-Simpson index). Rao (1986, 2010) defined a diversity coefficient and a distribution dissimilarity coefficient and proceeded to propose a decomposition of the diversity coefficient in a manner resembling ANOVA. Rao‘s quadratic entropy takes into account some measure of species richness (e.g. taxonomy). With this approach species assemblages with greater taxonomic distances (based on phylogeny to the degree to which the particular taxonomy is based thereupon) produce greater quadratic entropy. This index was applied to the data, as a simple measure of species richness would perhaps favour a more degraded stand due to the opportunistic species composition of annuals and alien plants likely to occur there, although Lechmere- Oertel et al. (2005a) found no difference in species richness but a significant difference in the Shannon-Wiener diversity indices determined for a range of degraded and intact spekboom thicket sites. I used a taxonomic distance function, based on taxonomic levels of family, order and genus, provided by the ade4 (Chessel et al., 2004) software package. A dendrogram was constructed based on taxonomic distance between the species and coupled to a table of the species count data arranged according to species as rows and treatment as columns. The columns of the abundance matrix have been arranged using the scores (coordinates) from the first dimension of a correspondence analysis of the matrix (see Fig. 2, Appendix B).

All analyses were performed with base R (R Development Core Team, 2010) and associated software packages including lattice (Sarkar, 2008), nlme (Pinheiro et al., 2010), ade4 (Chessel et al., 2004; Dray et al., 2007), homals (De Leeuw & Mair, 2009), Hmisc (Harrell, 2010) and contrast (Kuhn et al., 2010).

Results

Carbon

Soil carbon content between sample depth intervals showed a significant difference across restoration treatments (F = 4.05329, p = 0.0352) and sample intervals (F = 4.47665, p = 0.0196). T-test contrasts revealed that soil carbon content is significantly different in the 50 year (Restored50) treatment (t = 4.06, p = 0.002) between the 0:20 cm (first) and 20:40 cm (second) depth intervals, and between the 20:40 cm (second) and 40:60 cm (third) depth intervals (t = 2.03, p = 0.0481). The 35 year (Restored35) treatment showed a slightly more pronounced difference in soil carbon between the second and third (t = 2.24, p = 0.0296) than between the first and second (t = 1.76, p = 0.0854) sample depth intervals. No significant difference was found between the Degraded and Restored35

45 treatments in all the sample depth intervals. Differences in soil C between the Restored35 and Restored50 sites showed no significant difference in all sample depth intervals. Difference in soil C between the Degraded and Restored50 stands were significant in the first sample depth interval (t = -2.84, p = 0.0066), but not in in the second (t = -0.77, p = 0.4464) or third depth intervals (t = -0.47, p = 0.6432).

An ANOVA comparing soil carbon content between the different sampling depths revealed a significant difference (F = 15.511, p < 0.0001). Contrast tests on soil carbon between sample depths of 0:20 cm and 0:40 cm revealed highly significant differences for the Degraded (t = -3.50, p = 0.001), Restored35 (t = -4.95, p < 0.0001) and Restored50 (t = -4.45, p = 0.0001) treatments. Significant differences between the 40 cm and 60 cm sample depths for the Degraded (t = -2.16, p = 0.0359), Restored35 (t = -2.04, p = 0.0473) and Restored50 (t = -2.44, p = 0.0183) was detected. Soil carbon content between the 50 year (Restored50) and 35 year (Restored35) restored stands does not reveal significant differences measured across all sampling depths. The largest difference in soil carbon content between the Degraded and Restored35 treatments was detected at a soil depth of 60 cm (t = -1.76, p = 0.0854), followed closely by 40 cm (t = -1.73, p = 0.091). At a depth of 20 cm this difference is not found significant (t = -0.9, p = 0.3728). Between the Degraded and Restored50 stands the differences in soil carbon were significant at a depth of 60 cm (t = -2.16, p = 0.0361) and 40 cm (t = -2.11, p = 0.0401) but not at 20 cm (t = -1.66, p = 0.1031).

An ANOVA comparing root carbon content at the different soil sample intervals showed a significant difference between treatments (F = 3.697, p = 0.0452) but not between sampling depth intervals (F = 0.002113, p = 0.9979). Between the Degraded and the Restored50 treatments a difference in root carbon content was detected at 60 cm depth (t = -2.12, p = 0.0393), but less significant variation at a depth of 40 cm (t = -1.83, p = 0.0729) and 20 cm (t = -1.51, p = 0.1368). A significant difference (t = -3.29, p = 0.0019) in root C for the Restored50 stand between the sample depths of 20 cm and 40 cm was detected, and also for the Restored35 stand (t = -2.08, p = 0.0426) between the same depths. For the Degraded stand no significant root C difference was detected between the different sample depths. The Restored35 stand also showed no significant difference in root C between the sample depths of 40 cm and 60 cm (t = 0.29, p = 0.7738), but The Restored50 stand did show a significant difference in root C between the sample depths of 40 cm and 60 cm (t = -2.63, p = 0.0116). An ANOVA comparing root carbon content between soil sample depths and treatments showed a significant difference (F = 4.1627, p = 0.008), but no significant differences were detected between sample depths or between treatments.

An ANOVA of total belowground soil carbon measured per soil sample depth interval detected significant differences between sample intervals (F = 3.765, p = 0.0344), between treatments (F = 9.544, p = 0.0015) and also between treatments and sample intervals (F = 2.908, p = 0.0375). T-test model contrasts revealed a significant difference in belowground carbon content between the first and second depth intervals only in the 50 year restored site (t = 49.608, p < 0.0001). Between the second and third depth intervals a significant difference in belowground carbon content were detected for the 35 year (t = 2.28, p =0.0272) and marginally for the 50 year (t = 1.94, p = 0.0582) restored stands.

An ANOVA of total belowground soil carbon measured per soil sample depth detected significant differences between sample depths (F = 16.324, p < 0.0001), and between treatments (F = 3.756, p = 0.0433). T-test model contrasts showed significant differences between the 20 cm and 40 cm sample depths for the degraded (t = -3.61, p < 0.0001), 35 year (t = -5.35, p < 0.0001) and 50 year (t = - 4.97, p < 0.0001) restored sites. Similarly significant differences were detected between the 40 cm and 60cm sample depths for the degraded (t = -2.20, p = 0.0329), 35 year (t = -2.03, p = 0.0478) and 50 year (t = --2.84, p = 0.0067) restored sites. Between the degraded and 35 year restoration treatments a significant difference was detected for the 40 cm (t = -2.14, p = 0.0377) and 60 cm (t = -2.12, p = 0.0394) sample depths. No significant differences were detected in belowground carbon content between the 35 and 50 year restored sites at any of the sampling depths.

An ANOVA of aboveground carbon showed a significant difference (F = 31.442, p < 0.0001) between treatments, which was also reflected in both its components namely litter carbon (F = 44.958, p < 0.0001), and biomass carbon (F=20.325, p < 0.0001). The mean carbon content values recorded is shown in Table 3.6. Assuming that the sites under restoration treatment resembled the Degraded site sampled prior to the P. afra truncheon planting, aboveground carbon sequestration amounts to 91.7 t C ha-1 after 50 years and 29.5 t C ha-1 after 35 years of restoration treatment. Belowground C sequestration amounts to 44.4 t C ha-1 after 50 years and 27.3 t C ha -1 after 35 years of P. afra restoration treatment measured up to a soil depth of 60 cm. Total carbon sequestration of sites under P. afra restoration at Rhinosterhoek amounted to 161.4 t C ha-1 after 50 years and 69.6 t C ha-1 after 35 years, measured up to a soil depth of 60 cm. Very high rates of carbon sequestration per annum were detected, especially for the period between 35 and 50 years of restoration treatment across all soil depth levels (Table 3.6).

Table 3.6: Carbon (t .ha-1) data (mean values) from Rhinosterhoek. Below G = below ground, Above G = above ground.

Rate/ Soil Depth Treatment Soil SE Root SE Below G Biomass SE Litter SE Above G Totals yr Restored50 100.5 10.9 17.3 9.6 117.9 107.6 10.8 25.3 2.2 132.9 250.8 3.2 0-60 cm Restored35 96.7 21.0 4.1 1.3 100.8 45.4 10.6 12.8 2.4 58.2 159.0 2.0 Degraded 73.1 9.3 0.4 0.1 73.5 15.9 9.8 0.0 0.0 15.9 89.4 Restored50 84.5 7.4 15.1 8.4 99.6 107.6 10.8 25.3 2.2 132.9 232.6 3.2 0-40 cm Restored35 82.1 12.0 4.3 1.0 86.4 45.4 10.6 12.8 2.4 58.2 144.6 2.1 Degraded 55.4 9.1 0.7 0.3 56.0 15.9 9.8 0.0 0.0 15.9 72.0 Restored50 55.4 5.7 12.3 8.6 67.7 107.6 10.8 25.3 2.2 132.9 200.6 3.0 0-20 cm Restored35 44.9 7.8 2.2 0.9 47.1 45.4 10.6 12.8 2.4 58.2 105.3 1.6 Degraded 32.5 5.2 0.4 0.2 32.9 15.9 9.8 0.0 0.0 15.9 48.8

A standard linear regression model was used to predict total soil carbon to a depth of 50 cm (see Fig. 1 in Appendix B) in order to compare these data with other studies conducted in spekboom thicket, which all present data at 50 cm depth. This yielded values of a TCC of 80.5 t C ha-1 for the degraded, 151.5 t C ha-1 for the 35 year restored and 241 t C ha-1 for the 50 year restored treatments at Rhinosterhoek. This amounts to 160.5 t C ha-1 sequestered at Rhinosterhoek after 50 years of restoration treatment and 71 t C ha-1 sequestered after 35 years.

The data obtained from the carbon stocks sampled at Rhinosterhoek were compared with data from Krompoort collected by Mills and Cowling (2005) (Fig. 3.6). There is a general decrease in TCC from intact through the differently aged restoration sites (decreasing age), to degraded sites. A generally higher aboveground carbon stock was detected at the Rhinosterhoek sites, especially that of the 50 year old restored site when compared to the intact and other restored sites sampled. The soil carbon stock of the 50 and 35 year-old restored sites showed little difference and is lower than that recorded for the 27 year old restored site at Krompoort. The root carbon stocks recorded at the Krompoort 13 and 27 years restored stands were higher than that detected for the 35 year old restored stand at Rhinosterhoek.

Figure 3.6: Carbon stocks at Rhinosterhoek and Krompoort a soil depth of 50 cm. Data for the Krompoort (prefix=K.) sites were taken from Mills & Cowling (2006) and data for the intact sites were taken from Mills et al. (2005a). Data from this study (Rinosterhoek) has prefix = R.

Floristics

The multiple correspondence analysis based on treatment (intact, 11-, 19-, 33-, 35-, 50-year restored and degraded) (Fig. 3.7a) revealed three distinct groupings of data points in ordination space, which corresponds with i) the two degraded sites at Krompoort and Rhinosterhoek, ii) the three restored sites at Krompoort (Restored11, Restored19 and Restored33), and iii) the intact sites at Krompoort and Rhinosterhoek, together with the Restored35 and Restored50 sites at Rhinosterhoek. There is a clear horizontal procession of point clouds representing different restoration treatment age from left to right, with the degraded sites far left

(a) (b)

Figure 3.7: Multiple correspondence analysis showing results according to (a) treatment (b) status, (c) site and (d) cover components. In (a) R = Rhinosterhoek and K = Krompoort, Dgrd is Degraded, Intc is Intact and Rs is Restored, and the number attached shows number of years under restoration treatment. In (a) and (b) the labels for K.Rs33 /Restored33 (purple) obscures that of K.Rs19 / Restored19 (blue).In (c) labels are placed over the midpoint (barycentre) of the respective point-clouds. Ellipses are so called inertia ellipses and encapsulate 67% of the data points associated with each point-cloud. The ellipse axes shows the principle axes of variation (Chessel & Dufour, 2004).

(a) (b) Figure 3.8: Multiple correspondence analysis showing (a) results according to treatment (Fig. 3.6a) displaying the total carbon content (TCC) (t C ha-1) of each treatment sampled up to a depth of 50 cm below ground (the label displaying 186 t C ha-1 (Restored33) obscures that of 101.8 t C ha-1 (Restored19) (b) The same display but with overall mean TCC subtracted from observed TCC values and expressed as various sized squares proportional to the calculated deviation from the mean and scaled according to SD=1. The size of the squares thus show each point’s deviation from the mean expressed as a factor of SD. Black shows positive deviation values, white negative.

followed by sites of increasing restoration age and culminating in the Intact and and the older Restored50 sites on the far right. A similar pattern emerged when the analysis was based on degradation status, irrespective of site (Fig. 3.6b). Figure 3.6 (c) shows a clear separation vertically between the two sampling locations at Krompoort and Rhinosterhoek. In Fig. 3.6d, bare ground is strongly associated with the ordination space occupied by degraded sites, and recruits with that occupied by intact and older restored sites. Overall, these results very clearly show differences in composition among treatments and a trend for the older restored sites to converge in ordination space with the intact sites, despite compositional differences between Krompoort and Rhinosterhoek.

Figure 3.7 (a) shows the TCC attributed to the respective treatments. These values reveal a general increase in TCC in the horizontal plane from left to right in ordination space in line with restoration age. The only discrepancy is the high carbon content value of the Restored33 site at Krompoort (186 t C ha-1) in relation to that of the Restored35 site at Rhinosterhoek (152 t C ha-1). Figure 3.14(b) is the same figure with the observed values of TCC transformed to z-scores (i.e. overall mean is subtracted from the observed TCC values and the results expressed as various sized squares proportional to this deviation, and scaled to where the SD=1). Negative deviation values are expressed as white and positive as black. This figure displays the increase in TCC from left to right more vividly.

There was a clear trend of increasing diversity from degraded, via restored to intact treatments at both Rhinosterhoek and Krompoort (Fig. 3.9). However, there was substantial variation within the restored sites: at the Krompoort site the 11 year-old restored site had a higher diversity coefficient than the stands under 33 and 19 years of restoration treatment, rivaling that of the 50 year old restored site at Rhinosterhoek. The degraded stand at Krompoort was the least taxonomically diverse, followed by that of the 19 year old restored site and the degraded site at Rhinosterhoek. The 35 year old restored site at Rhinosterhoek showed a surprisingly high taxonomic diversity, even higher than that of the intact stand at Krompoort.

The most recruits recorded were in the Restored50 treatment, followed by the two intact treatments (Fig. 3.10). When depicted in terms of individual canopy species (Fig. 3.11), it becomes clear that recruits are mainly associated with intact and older restoration treatments. In terms of abundance, the most commonly recorded recruiting species was Pappea capensis, the canopy dominant in Sundays Spekboomveld.

Figure 3.10: Boxplots illustrating the pseudo-cover of spekboom thicket canopy recruits across restoration treatments. The width of the boxes increase with the sample size of recruits detected. Recruits included all thicket canopy plant species, ≤ 100 cm in height (excluding P. afra), be they seedlings or sprouts.

Figure 3.11: The pseudo-cover of spekboom thicket canopy constituent recruits per species across treatments with abundance (as opposed to canopy cover) assigned to a cover scale used with adult plants where 1=1, 2=2-4, 3=5-10, 4=11-15, 5=16-24, 6=25-34 individuals

Discussion

The planting of P. afra truncheons on heavily degraded spekboom thicket sites clearly enables the sequestration of carbon up to levels comparable with intact sites, but also enables the autogenic return of spekboom thicket canopy plant recruits after some time. Especially the high number of Pappea capensis recruits, a dominant and long-lived canopy constituent in this vegetation type, recorded in the Restored50 treatment is an indication of an autogenic regeneration of thicket canopy plants. It is estimated that after a period of 35 to 50 years of restoration treatment, the return of canopy plant biodiversity may be observed, whereas carbon build-up in the soil may take even longer. This suggests that this restoration methodology enables the restored system to transcend the abiotic threshold and that the return of canopy plant biodiversity is directly related to i) the build-up of carbon in the soil and the accompanying changes in soil quality (Mills & Cowling, 2010; Mills & Fey, 2004) and ii) the unique microclimate aboveground and mulch-rich substrate that the canopy of restored P.afra provides (Cowling et al., 1997; Mills and Fey, 2004; Mills et al., 2005a; Lechmere-Oertel et al., 2008; Sigwela et al., 2009).

The multiple correspondence analysis results of the floristic data show a clear procession of sites horizontally from left to right across the ordination space (see Fig. 3.7a&b) for degraded, various restored and intact sites at both Rhinosterhoek and Krompoort, despite the compositional differences between these two site localities that are located 25 km apart. The separate clustering (vertical plane) in ordination space (Fig. 3.7c) of the two site localities in the multiple correspondence analysis above is likely the result of a compositional difference due to the effect of geographic distance or unspecified environmental gradients between the two sites. The associated shift in total carbon content (TCC) with increasing restoration age (Fig. 3.8a&b) reiterates the strong association of increasing carbon content with increasing restoration age (Mills & Cowling, 2006) and strongly suggesting the appearance of spekboom canopy recruits (i.e. autogenic regeneration) (Fig. 3.7d) at the higher end of ecosystem TCC levels.

The results from Rao‘s diversity coefficient indicates a general increase in taxonomic diversity (measure of taxonomic distance between species within each treatment assemblage) with increasing restoration age, with intact sites being the most, and degraded sites being the least taxonomically diverse. Some anomalities are detected in the general trend of increased taxonomic diversity with increasing restoration age which may be attributed to i) remnant long-lived thicket species were included in sampling of the younger restored sites and ii) smaller sample size of the 50 year old restored site at Rhinosterhoek. The higher diversity coefficient

54 values of the 11 year old restored site at Krompoort and the degraded site at Rhinosterhoek may be explained by the more eclectic mix of species in these assemblages, containing i) species of ephemeral and perennial grasses and (opslag), iii) ephemeral and perennial forbs, iv) karoo-shrubs together with v) long- lived remnant thicket tree species and vi) planted P. afra individuals in the restored sites. Such assemblages are likely to be more taxonomically distant than those of older, undisturbed restored sites shaded with significant P. afra canopy (e.g. 50 year restored site). The high diversity coefficient value for the 35 year restored site at Rhinosterhoek rivals that of the intact sites at Krompoort and Rhinosterhoek, suggesting similar plant diversity and recruit population (see Fig. 3.10 and 3.11), although still displaying some structural and floristic differences. Lechmere-Oertel et al. (2005a) compared plant species richness and Shannon‘s Diversity Index H‘ between a range of intact and degraded spekboom thicket sites and found significantly higher species richness and diversity in intact sites. This result is mirrored in the results from Rao‘s diversity coefficient above that show the same general trend, with coefficient values of degraded sites and intact sites at either sides of a spectrum of diversity coefficient values for degraded, various aged restored and intact spekboom thicket sites.

The poor representation of recruits on the 33 year old restored site at Krompoort compared to the 35 year old restored site at Rhinosterhoek is puzzling. However, it may be explained by one of the following reasons: i) the so-called ―bergspekboom‖ genotype of P.afra planted here displays a characteristic upward growth form and a general absence of a skirt and thus does not provide as good a canopy cover as other genotypes, ii) the steep slope on which this restoration block was established, together with the growth form of this P. afra genotype may have allowed for a faster migration of litter downslope with heavy rainfall events and thus an inability to trap as much organic material aboveground for seedlings to establish in, iii) browsing pressure by wild ungulates such as kudu (Tragelaphus strepsiceros), and the occasional mishap allowing goats and cattle to enter the restoration exclosure plots. It is significant to note that although not formally quantified here, the single row of a local variety of P. afra planted at Krompoort 33 years ago with the other genotype (see Mills & Cowling, 2006), yielded seven recruits of spekboom thicket canopy plants, including Euclea undulata and Pappea capensis, as opposed to the only two recruits recorded from sampling the whole 33 year-old restored block of ―bergspekboom‖.

The relationship between biodiversity and ecosystem functioning have been fervently discussed in the ecological literature (Chapin, 1997, Tilman et al., 1997, Loreau et al., 2001, Naeem, 2002, Naeem & Wright, 2003, Worm & Duffy, 2003, Hooper et al., 2005) and the consensus is that ecosystem characteristics are

55 fundamentally dependent upon the diversity of organisms present in the system (biodiversity) in terms of their functional attributes, and the abundance and distribution of these species in space and time. One such functional attribute is that of an ecological engineer (Jones et al., 1994, 1997; Wright & Jones, 2006; Hastings et al., 2007). The concept has been debated much and is defined broadly as organisms that physically create, maintain or modify habitats by causing physical state changes in abiotic and biotic materials and thus govern the accessibility of resources to other organisms within the system (Jones et al. 1994, 1997). Jones et al. (2010) proposed a four-point framework for predicting the effects of ecosystem engineers on the ecosystems with which they interact, namely i) an engineer causes structural change, ii) structural change causes abiotic change, iii) structural and abiotic change cause biotic change, and iv) structural, abiotic and biotic change can feedback to the engineer.

The role of ecosystem engineer species in overcoming abiotic thresholds within restoration contexts (Byers et al., 2006) are of interest here. In spekboom thicket it has been established that P. afra, due to its exceptional eco-physiological characteristics to adapt to seasonal soil moisture conditions (Guralnick et al., 1984; Guralnick & Ting, 1987) and high productivity, effects structural changes in the above and belowground environments by creating a canopy that bestows a litter- rich substrate and a favourable below-canopy microclimate (Cowling et al., 1997; Mills and Fey, 2004; Mills et al., 2005a; Lechmere-Oertel et al., 2008, Sigwela et al., 2009). It is held that the exceptionally high soil organic carbon accumulation can be attributed to i) high leaf litter production by canopy trees (Lechmere-Oertel et al., 2008) that provide ii) a relatively cool below-canopy temperatures accompanied by iii) high levels of rainfall interception (mean=56.4%) (Cowling & Mills, 2010). It is hypothesized that the high rainfall interception mechanism contributes to reduced moisture of organic material (litter) and reduced mineralization of soil organic carbon stocks as a result of reduced wet-dry cycles (Birch-effect) and reduced microbial activity which is linked with reduced soil moisture. The high soil organic carbon accumulation is in turn related to aspects of soil fertility (see Lechmere-Oertel, 2005a and Mills & Cowling, 2010). All these factors enables an abiotic change (transcending the hypothetical threshold), which leads to a biotic change, which, in turn, manifests as a change in plant species composition. This biotic change involves the regeneration of species representative of reference intact spekboom thicket landscapes. It is still unclear how these changes feed back to the engineer at a later stage, and if they are positive or negative in overall effect (Jones et al., 1997), but it is evident from the definitions in the literature, and the evidence in hand that P.afra may be regarded as an ecosystem engineer in spekboom thicket ecosystems.

A restoration protocol focused primarily on planting only P. afra extensively on degraded landscapes is an effective way to overcome the theoretical abiotic threshold that has been established due to severe degradation. It enables the autogenic regeneration of spekboom thicket biodiversity after a period of approximately 35-50 years. In terms of large-scale restoration of degraded spekboom thicket landscapes, it is deemed unnecessary to consider introducing other plant species as part of the restoration protocol in an effort to return biodiversity to these sites. Such introductions may only be necessary in isolated cases where intact landscapes or source species populations of previous intact spekboom thickets have disappeared from the larger area.

Chapter 4 – General Discussion

Summary of major findings

Given the potential large scale restoration opportunities within degraded spekboom-dominated thicket, and the drive towards developing restoration protocols that i) are scientifically informed, ii) are directed according to realistic goals, and iii) have appropriate measures for monitoring and measuring restoration success, it was necessary to investigate concerns noted by stakeholders regarding the lack of restoration of other species that occur in intact spekboom thicket reference sites. The objective of this study was to examine the return of biodiversity to degraded spekboom-dominated landscapes after it has been extensively planted with Portulacaria afra. More specifically, the focus was on the restoration potential of canopy plant species composition in spekboom dominated thicket. I approached the problem from two angles.

In Chapter 2 I tested whether the actual planting of nursery-propagated plants was a viable restoration method to implement alongside the planting of P. afra. The idea was to approach the restoration of spekboom thicket according to the prevalent equilibrium or succession-type models of rangeland dynamics (Westoby et al., 1989; Briske et al., 2003). Here it was hypothesized that once canopy constituent woody species are artificially introduced, they will start to thrive and reproduce and, in time, resemble the species assemblage and species population structures characteristic of intact reference sites. The results were unambiguous and show that i) survival of nursery propagated canopy species planted in degraded landscapes is marginal. In fact, even in intact spekboom seedling survival was exceptionally low, and ii) the restoration costs associated with such a strategy are disproportionate, especially when assessed in terms of seedling survival. This affirmed the observations of previous research on recruitment dynamics that it is ramets rather than seedlings that are mostly responsible for recruitment within spekboom dominated thickets (Midgley & Cowling, 1993), and that seedlings are almost always associated with below-canopy micro-sites (Sigwela et al., 2009), and probably only establish when a suite of favourable conditions coincide.

In Chapter 3 I tested the hypothesis that canopy guild plant biodiversity will return to degraded spekboom-dominated landscapes once P.afra is established through intensive planting of truncheons (Lechmere-Oertel et al., 2005a). The rationale behind this hypothesis is that once canopy conditions are established with sufficient litter build-up (Lechmere-Oertel et al., 2008) and concomitant build-up of soil organic carbon content and associated fertility (Mills & Fey, 2004; Lechmere-

Oertel et al., 2005b; Mills & Cowling, 2010) the woody canopy species assemblage characteristic of intact reference sites will regenerate spontaneously from ramets and genets, provided that source populations and dispersal agents are within suitable proximity (Sigwela et al., 2009; Cowling & Mills, 2010). The results of the multivariate analysis showed a clear partitioning of sites within ordination space according to restoration treatment age and a general procession of restored stands with increasing age from degraded towards intact stands. The number of juveniles recorded in the older restored sites testifies to this, with the 50 year old restored stand showing even more recruitment than the reference intact stands. This reveals that woody plant canopy species are regenerating in this environment, and that some change occurred (a hypothetical threshold has been transcended) within the older restored sites that allowed for this to happen. The results from the carbon content analyses shows an increasing build-up of carbon in the above- and belowground environments with increasing restoration age, suggesting that total carbon content, as measured within each of its above- and belowground components, may be a reliable indicator of the health of the system. P. afra can, in fact, be regarded as a carbon pump, continuously injecting carbon into the system and thus affecting and maintaining thicket health.

The fact that P. afra facilitates these changes testifies to its crucial role in maintaining ecosystem structure and function. In this context, it is thus appropriate to consider P. afra as an ecological engineer (Jones et al., 2010). The goat- browsing degradation and restoration dynamics of spekboom-dominated thicket appear to fit the state and transition or non-equilibrium conceptual models, where the degraded stands typically represent the post-threshold state. In this state the return of thicket canopy constituent species is impossible via any inherent mechanisms within the system (Briske et al., 2003; 2006; Stringham et al., 2003), or through attempts to replenish missing canopy species (other than P.afra) artificially. The autogenic regeneration of canopy constituent species can be achieved, but only after some significant alterations in the soil and above-ground micro-climates have been engineered. By planting closely-packed spekboom truncheons in parallel rows and leaving the stand undisturbed for a period of between 35 and 50 years, the return of canopy plant biodiversity within restored spekboom-dominated thickets is achieved. In terms of achieving restoration success as defined by the Society for Ecological Restoration (SER, 2004), this restoration methodology is considered effective.

Research significance of this study

The current spekboom thicket restoration protocol of establishing P. afra cuttings on degraded landscapes is fully supported by my research results. By planting only P. afra cuttings, above- and below-ground carbon restoration is engineered that facilitates the autogenic return of compositional and functional integrity to degraded spekboom thicket. The results accentuates the futility of approaching spekboom thicket restoration through attempts at superficially recreating reference intact species assemblages without considering the processes that allowed these species to establish and persist within intact reference sites in the first place. This research should send a clear message to concerned stakeholders that even though biodiversity restoration does not seem to be the direct aim of the P. afra planting methodology, it indirectly facilitates plant biodiversity regeneration through effecting the appropriate changes within the above- and below-ground environments.

Secondly, the importance of this study lies within an academic context by highlighting the importance of applying different ecological conceptual models to restoration approaches. Even though it is proven that not all systems behave according to the non-equilibrium or alternate states conceptual models (Briske et al., 2003; Stringham et al.,2003; Suding et al., 2004), the potential of these models to inform restoration methodology should not be underestimated. This study has highlighted the role of P. afra as a species that is capable of establishing and maintaining biophysical conditions that allows for other organisms to establish and persist within a larger abiotic environment that is not conducive to their survival otherwise. This feature puts it in a league of other species identified in the ecological literature as ecosystem engineers (Jones et al., 1994; 2010). This study demonstrates the value of ecological engineers as the focal point of restoration efforts (Byers et al., 2004).

Shortcomings and limitations of this study

A key shortcoming of this study involved experimental design. Firstly, the restoration stands at Krompoort were not established with the principles of experimental design in mind and, therefore, were not all homogeneously planted with P. afra in terms of density. Furthermore, the two restoration sites at Rhinosterhoek were not of similar extent and they also differed in size and density from those at Krompooort. These difficulties are common in ecology where researchers often need to exploit experiments over which they have had little

60 control. In this study these restored stands are the oldest known available and thus provided the only means with which to implement a study of this nature.

The initial spekboom planting at Krompoort in 1976 also involved the planting of a row of locally occurring spekboom in addition to the ―Bergspekboom‖ genotype sourced from a mountain approximately 90 km away. ―Bergspekboom‖ does not develop the characteristic skirt common to most spekboom genotypes, or at least not when planted at these densities. Considering the fact that casual observations indicated more juveniles of canopy species in this row than what was detected in the sampling of the ―Bergspekboom‖ genotype of the same age, it would have been interesting to see if some effect preventing seedling establishment due to the characteristics of the specific P. afra genotype could be established. Mills & Cowling (2006) recorded a much higher root and litter carbon content for the locally occurring P. afra stand than for the ―Bergspekboom‖ genotype, although the imported genotype showed much higher soil and aboveground carbon content.

What are the next steps ?

Research on the restoration of degraded spekboom thicket should focus on refining the right planting methods for P. afra in order to i) ensure maximum survival, and ii) be on par with intact reference sites in terms of carbon sequestration and biodiversity regeneration potential. Such planting trials have already begun, and the results of these need to be assessed. The thicket–wide plots project established by the Subtropical Thicket Restoration Programme (STRP) consists of over 300 plots planted across spekboom-dominated thickets. Many questions remain as to the specific dynamics of spekboom-dominated ecosystems. Important aspects still to be investigated within a restoration context include the variety of landscapes and species assemblages that occur within spekboom dominated thickets as well as the possible feedbacks and interrelations between spekboom-dominated thickets and adjacent vegetation types such as the noorsveld and bontveld vegetation types.

Studies on carbon stocks of different intact spekboom dominated vegetation types are becoming increasingly important. Mills & Cowling (2010) found significantly less soil carbon content in intact Baviaans Spekboom Thicket than in Sundays Spekboomveld and Sundays Spekboom Thicket, even though the carbon sequestration potential of degraded Baviaans Spekboom Thicket is still on par with that of degraded mesic forest ecosystems. This variation implies an inherent difference in carbon cycling processes between these vegetation types that are probably related to soil fertility in terms of concentrations of extractable cations and

P concentrations in the topsoil. This should affect total carbon sequestration potential and yearly carbon sequestration rates. Variation in below-ground carbon stocks of intact spekboom-dominated vegetation types is likely to be mirrored in aboveground carbon stocks. Intact Baviaans Spekboom Thicket had a total aboveground carbon stock of 34.5 t C ha-1 (Powell, 2009) as opposed to the 51.6 t C ha-1 recorded for Sundays Spekboomveld by Mills et al.( 2005a). It is possible that differences in soil characteristics will manifest in differences in P. afra truncheon survival, growth rates and maximum carbon sequestration potential. This in turn will affect carbon credit earning potential in different P.afra dominated vegetation types. To arrive at more precise projections of carbon accrual potential and biodiversity return within degraded counterparts of the same vegetation type, it is worth investigating the natural variation of carbon stocks and biodiversity within a given intact spekboom-dominated vegetation type with regards to i) specific soil characteristics derived from underlying geology, ii) soil depth, iii) terrain slope and iv) aspect.

The continuous monitoring and evaluation of restoration progress is an essential requirement for achieving restoration targets (Hobbs, 2007; Hobbs & Norton, 1996). Strict requirements for measuring carbon accrual in both the below- and above-ground environments according to tested and standardized methods, need to be developed and implemented for spekboom-dominated vegetation types. These requirements will need to be in line with international standards as developed for the carbon markets (Pearson et al., 2005)

Vlok et al. (2003) were the pioneers in delineating and mapping major vegetation types and identifying patterns and process within the larger thicket biome. Focusing on a larger scale, they mapped broad vegetation types with the help of LANDSAT Thematic Mapper (TM) imagery and direct field observation and using expert knowledge of landowners, fence-line contrasts, physiographic correlates and remnant vegetation to identify historical occurrence of mapping units in transformed areas. With the potential of large-scale restoration within spekboom- dominated thicket vegetation types, it is becoming necessary to at least revise the boundaries of these mapping units based on more recent data, including that of total carbon stocks and intact species assemblages and possibly also map these vegetation units on a smaller scale.

Typically, spekboom-dominated thicket thrives on drier north-facing slopes and is generally accompanied by an adjacent bontveld vegetation unit on footslopes and valley floors. Almost no work has been done on the impact of destructive goat- browsing on bontveld vegetation types or on south-facing slope vegetation units where P. afra is also present but does not dominate. Possible interactions between P. afra dominated thickets and these adjacent mapping units need investigation. 62

Should large-scale restoration action become a reality, considerable rainfall interception by P.afra (Cowling & Mills, 2010) would affect the amount of runoff reaching adjacent bontveld and riverine thornveld vegetation units and possibly changing some ecosystem dynamics that have been established during the transformed state.

More research is needed with regards to recruitment dynamics within spekboom- dominated thicket. Sigwela et al. (2009) showed that in spekboom-dominated thicket almost all canopy seedlings occurs within beneath-canopy microsites, yet it is still not clear how long it takes or under what conditions recruits become fully- fledged canopy components. It is possible that some form of disturbance of the canopy, particularly the dense P. afra canopy, may be needed at some stage for recruits of canopy species to break through the established canopy under which they became established. This scenario is plausible given the fact that a major historical presence of large mammals such as the African elephant (Loxodonta africana) and Black rhinoceros (Diceros bicornis) would have impacted P. afra dominated thickets (Cowling et al., 2009; Cowling & Kerley, 2002), albeit differently from the impact caused by injudicious goat browsing (Stuart-Hill, 1990). Should this be the case, restoration protocol would need to be suitably amended to allow for established recruits of canopy species a chance to reach the canopy once a dense canopy of P. afra has been established through some sort of disturbance. Large mammals have only been eradicated from the majority of spekboom-thicket landscapes for about 150 years (Cowling et al., 2009). It is safe to assume that current intact spekboom thickets have established and been maintained under conditions where large mammals were present within these areas in reasonably large numbers, although being dependent on surface water availability the influence of these animals would have been restricted during periods of drought. The degree to which the absence of these mammals affects spekboom-dominated thicket ecosystem process and function, and the degree to which goats and cattle or other alternatives can simulate these effects warrants investigation.

Sigwela (2004) showed that large herbivores such as African elephant and Black Rhinoceros together with animals such as Kudu (Tragelaphus strepsiceros) and even tortoises (Mason, 1997) each rivals birds in the number of thicket plant species dispersed. The role of primate species such as the Chacma Baboon (Papio ursinus) and Vervet Monkey (Chlorocebus pygerythrus) as dispersal agents of spekboom thicket species is still unknown and deserves further study, as these species are, together with tortoises, birds and kudu, in many cases still present (albeit in low densities) in intact spekboom thicket areas.

Valuable work is being done regarding the changes P. afra effects in the soil of transformed landscapes. New insights with respect to rainfall interception by P. 63 afra, and the effects of this on above- and belowground beneath-canopy environments (Cowling & Mills, 2010; Mills & Cowling, 2010) are uncovering some of the dynamics of soil carbon accrual in spekboom-dominated thickets. Data supplied by the landowner at Krompoort reveal precipitation that occurs as several consecutive light rainfall events ranging from a fraction of a millimeter and lasting a few minutes to longer events exceeding 200 mm lasting over a few days. Single erratic events of heavy rain exceeding 75 mm in a few hours also occur. Analyses of separate rainfall events may give valuable information on the frequency of extreme and erratic rainfall events and the effects of these on rainfall interception by P. afra. Within the context of looming climate change, a general consensus is that extreme rainfall events and droughts are to become more prevalent, which could have a significant influence on carbon accrual rates of transformed spekboom thicket sites under restoration. Intensive research is needed on possible climate-change scenarios within the thicket biome and how it will affect species distributions and restoration protocols.

Despite the large-scale degradation of spekboom-dominated thickets, goat pastoralism remains an important economic activity within the area. It is possible that heavy browsing pressure can transform spekboom thickets to open savannas over a time-span as short as a decade (Hoffman and Cowling, 1990; Lechmere- Oertel et al., 2005a). Degradation of remaining intact stands are continuing in many areas. There is a need for a continuous and rapid monitoring system relying on the recent advances in remote sensing technology to detect land cover changes. Other applications of emerging remote sensing technologies such as hyperspectral imaging and lidar capabilities need to be investigated. Despite work done by Stuart-Hill & Aucamp (1993), more research is needed in terms of carrying capacity determination of goats and other stock animals and game in spekboom- dominated thickets in different stages of degradation and practical veld management procedures to avoid reaching a threshold where autogenic recovery is impossible without the planting of P. afra truncheons.

To date, no studies are available on the variety of P. afra genotypes and their distribution (Mills & Cowling, 2006), which is unfortunate. The observed difference in two genotypes are documented by Mills & Cowling (2006) and involves differences in root carbon stocks, soil carbon stocks, litter carbon stocks and aboveground biomass carbon stocks. A significant difference in growth form is also detected. It is possible that a variety of genotypes exist that are native to their area of occurrence, and which presumably are best adapted to that specific habitat. Within a restoration context it is thus necessary to source truncheons from local P. afra stocks. In some areas there may not be enough intact thicket bushclumps harbouring P. afra plants to harvest cuttings from. A study on the possible

64 maximum intensity of harvesting P. afra cuttings and associated recovery rates of harvested bushclumps may provide useful guidelines for sustainable harvesting of P. afra truncheons to plant.

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APPENDIX A

Table 1: Restoration costs per canopy species planted at a density of 60 plants per hectare based on the costing of actual large scale Working for Woodlands contracts for planting dense rows of P. afra truncheons. The table shows the difference per species based on propagation method.

Propagation Planting Propagation Planting Propagation Totals method cost ha-1 cost per density cost ha-1 per (species-1 (per species) plant (plant ha-1) species ha-1 at 60 plants ha-1) Estimated restoration costs of large-scale planting of canopy species per hectare Cutting R 1,781.99 R 3.81 60 R 228.42 R 2,010.41 Seedling R 1,781.99 R 4.02 60 R 240.97 R 2,022.96

APPENDIX B

Figure 1: Results from a standard linear regression model using ordinary least squares for estimating Total Carbon Content (TCC) at a soil depth of 50 cm. The model is based on the formula TCC ~ Depth + Status, and R2 = 0.905 and adjusted R2 = 0.899.

Figure 2: A summary of the floristics data set, with species as rows and treatments as columns (tabular part of the figure). Species names with prefix=R signifies recruits (pseudo-species). The Dendrogram is based on a taxonomic distance metric. Square-size shows counts per treatment. The columns of the abundance matrix have been arranged using the scores (coordinates) from the first dimension of a correspondence analysis of the matrix. The arrangement differs little from that that would be got by arranging columns in a more or less ordered progression from degraded sites through increasing time of restoration to intact sites (from right to left). Degraded vs Restored/Intact are well characterized by the presence of some species and the absence of others.