Ecology of Three Apex Predators in Majete Wildlife Reserve, Malawi

Home , Liwonde National Park, Majete Wildlife Reserve, Sabi Sand Game Reserve

Ecology of three apex predators in Majete Wildlife Reserve, Malawi

Willem Daniel Briers-Louw

Thesis presented in partial fulfilment of the requirements for the Degree of Master of Science, Department of Conservation Ecology & Entomology, Stellenbosch University

Supervisor: Dr Alison J. Leslie

Faculty of AgriSciences

Department of Conservation Ecology & Entomology

December 2017 Stellenbosch University https://scholar.sun.ac.za

Declaration

By submitting this thesis electronically, I declare that the entirety of the work contained therein is my own, original work, that I am the sole author thereof (save to the extent explicitly otherwise stated), that reproduction and publication thereof by Stellenbosch University will not infringe any third party rights and that I have not previously in its entirety or in part submitted it for obtaining any qualification.

Willem Daniel Briers-Louw

December 2017

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Abstract

Apex carnivores play an important role in the ecosystem by regulating prey via predation. Anthropogenic influences have resulted in rapid range and population reductions of large carnivores across the African continent. These carnivores are often reintroduced into protected areas to compensate for human-induced losses, restore ecosystem functioning and promote eco-tourism.

Majete Wildlife Reserve in Malawi is a prime example, as human persecution resulted in the extirpation of large carnivores, with the exception of a small spotted hyena (Crocuta crocuta; hereafter hyena) population. As from 2003, African Parks attempted to rectify this problem by restoring and developing the reserve. Between 2011 and 2012, three lions (Panthera leo) and six leopards (Panthera pardus) were reintroduced. The aim of this study was to describe the ecology of the apex predators and to determine whether the felid reintroduction was successful or not.

Lion and leopard movements and home ranges were determined using GPS collars. The reintroduction of felids was considered successful. This was based on: (1) reduced post-release movements; (2) lack of homing tendencies; (3) breeding success; and (4) population persistence. Mean home ranges of lion (380.45 ± 117.70 km2 [SE]) and leopard (495.08 ± 80.99 km2), were the largest on record for any reintroduced felid in Africa, which was likely due to a low competitor density. Thus, we expect home range sizes to decrease with an increase in conspecific density.

Population abundances and densities were estimated with the use of camera traps. The known lion population increased to eleven individuals in five years, while the leopard population was estimated at 11 (range = 9–17). This indicates population persistence and growth. Both founder populations were small and require additional translocations to maintain genetic diversity. Hyena density (2.62 hyenas/100 km2) and clan size (5.33 ± 0.67) were the lowest estimates in any woodland habitat and comparable to arid areas. This may be explained by decades of direct persecution and poaching of their prey, or a naturally low density.

Predator diets were described and compared by means of scat analysis. Lion and hyena exhibited a high dietary overlap of medium to large herbivores. Using Jacobs’ preference index, both species preferred warthog (Phacocoerus africanus) and waterbuck (Kobus ellipsiprymnus). Hyenas selected a broader range of prey, likely reducing competition with lions (which almost exclusively selected only four species). In contrast, leopards occupied a lower dietary niche, which consisted mainly of small-to medium-sized ungulates. These findings indicate that the three apex predators use resource partitioning to reduce competition.

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This study suggests that reintroduction is a viable tool for re-populating large carnivores in protected areas in Malawi. The current predator population appeared to have a minimal impact on prey populations due to their small population size. We recommend long-term monitoring of predator- prey dynamics as the predator populations increase to prevent major ecological imbalances. Finally, we encourage management to focus energy and resources on the formation of a managed carnivore metapopulation to establish a genetically viable carnivore population within Malawi.

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Opsomming

Top roofdiere speel ‘n belangrike rol in die regulering van ekosisteme deur predasie. Menslike invloede het gelei tot ‘n vinnige afname in karnivoor populasiegetalle en habitatverliese in Afrika. Hierdie karnivore word dikwels hervestig in bewaarde areas om te kompenseer vir menslik ge- induseerde verliese, funksionering van ekosisteme te herstel en eko-tourisme te bevorder.

Majete Wildreservaat is ‘n prima voorbeeld waar menslike vervolging groot karnivore uitgeroei het, behalwe vir die gevlekte hiëna (Crocuta crocuta; hierna hiëna) populasie. Vanaf 2003 het African Parks gepoog om hierdie probleem reg te stel deur restourasie en ontwikkeling van die reservaat. Tussen 2011 en 2012 was drie leeus (Panthera leo) en ses luiperds (Panthera pardus) hervestig. Die doel van hierdie studie was om die ekologie van die top roofdiere te beskryf en te bepaal of die hervestiging van leeus en luiperds suksesvol was.

Bewegings en gebiede van leeus en luiperds is bepaal met hulp van GPS-halsbande. Die hervestiging van hierdie spesies was beskou as suksesvol. Na vrystelling was daar: (1) ‘n afname in beweging; (2) geen behoefte om huiswaarts (herkoms) te keer nie; (3) sukses met teling; en (4) populasie oorlewing. Die gemiddelde gebied van leeus (380.45 ± 117.70 km2)[SE]) en luiperds (495.08 ± 80.99 km2) was die grootste op rekord vir enige hervestigde karnivoor in Afrika, moontlik as gevolg van lae kompetisie digtheid. Dus verwag ons ‘n verkleining van gebiede namate die kompetisie digtheid toeneem.

Populasiegetalle- en digthede is geskat met behulp van kamerastrikke. In vyf jaar het die erkende leeu populasie tot elf individue vermeerder, terwyl die luiperd populasie op 11 (omvang = 9–17) staan, wat populasie oorlewing en groei aandui. Beide stigter populasies was klein, daarom is addisionele translokasie nodig om genetiese diversiteit te behou. Hiëna digtheid (2.62 hiënas/100 km2) en stamgroep grootte (5.33 ± 0.67) was die kleinste in soortgelyke habitat. Hierdie is moontlik as gevolg van direkte vervolging en stroping van hulle prooi oor dekades of ‘n natuurlike lae digtheid.

Roofdierdiëte is beskryf en vergelyk deur misanalise. Leeus en hiënas se diëte van medium tot groot prooi het tot ‘n groot mate oorvleuel. Gebaseer op Jacobs’ indeks, het beide spesies vlakvark (Phacocoerus africanus) en waterbok (Kobus ellipsiprymnus) verkies. Hiënas het egter ‘n wye verskeidenheid prooi geëet, moontlik weens mededinging met leeus (wat amper uitsluitlik net vier spesies geëet het). In kontras met leeus en hiënas, het luiperds ‘n kenmerkende dieet, wat hoofsaaklik uit klein tot medium grootte prooi bestaan. Hierdie resultate bewys dat die drie karnivore kompetitise met mekaar verminder deur verskillende diëte te volg.

Hierdie studie dui daarop dat hervestiging van groot karnivore in bewaarde areas in Malawi ‘n vatbare tegniek is. Tot dusver het die roofdierpopulasie ‘n klein impak op die prooipopulasies gehad weens

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die klein roofdier populasiegetalle. Ons beveel aan dat roofdier-prooi dinamika gemonitor word soos wat die roofdiere toeneem om ekologiese wanbalanse te vermy. Ten slotte, moedig ons die bestuur van African Parks aan om ‘n karnivoor meta-populasie in Malawi te stig, wat ‘n genetiese vatbare populasie kan volhou binne die landsgrense.

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Acknowledgements

Thank you to African Parks (Majete) Pty Ltd. for allowing me to witness and be part of the amazing success story that is Majete Wildlife Reserve. Thank you to everyone who made this study possible, particularly those who were involved in the reintroduction and collaring of the lions and leopards. Thank you to Majete management and the entire game capture team, for allowing me to be part of the lion re-collaring, this was truly a once-in-a-lifetime experience. Next, I would like to thank Craig Hay and Gervaz Tamala for their supervision and willingness to assist me with my research. To the rest of the management team, thank you for making me feel part of the Majete family. Thank you to the law enforcement team, especially Tizola Moyo, Martin Awazi and John Jiya for organising scouts, providing valuable information and allowing access to reports. To all the scouts who assisted me in the field I am greatly appreciative. Your knowledge about Majete really aided my project and helped me learn so much about the bush. I would also like to thank all the stores-men and mechanics at the workshop, particularly Africa’s best bush-mechanic, Isaac Mlilo, who kept our vehicles going despite numerous setbacks and was always there when we needed him.

Thank you to Dr Alison Leslie who provided me the opportunity to experience the “warmest” heart of Africa. Majete has become my home and I will never forget the stories and friends that have been made along this incredible journey. I am overjoyed to have been part of this amazing team and to have experienced this with you. I would like to thank all the Earthwatch volunteers for their contribution.

Thank you to Prof Arthur Rodgers for your correspondence and help with operating Home Range Tools in ArcGIS. Thank you to Dr Matt Hayward for providing equations for predator carrying capacities and Prof Martin Kidd for help with statistics. I would also like to thank Dr Dan Parker and Rhodes University for providing access to their reference of animal hair cross-sections. A special thanks to Prof Antoinette Malan for her generosity with laboratory equipment which proved invaluable for the scat analysis.

Thank you to the Hay family for your support throughout my stay in Majete and being my role-models. To all my fellow researchers, Anel Olivier, Claire Gordon, Frances Forrer and Kayla Geenen, thank you for all your help and for the good times we shared in Majete. Thank you to my colleague and friend, Charli de Vos, for assisting me with my challenging fieldwork. I am so privileged to have had the opportunity to work with you as we battled through some difficult times together, but more importantly, experienced the utter joy and satisfaction of life in the African bush. Thank you to my family for allowing me to pursue my dream of wildlife conservation and providing financial support. I am forever grateful. Finally, I thank my King for showing me the breath-taking beauty of His creation.

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Contents

Declaration ...... i Abstract ...... ii Opsomming ...... iv Acknowledgements...... vi Contents ...... vii List of Tables ...... xi List of Figures ...... xiv List of Appendices ...... xvi Chapter 1 - Introduction ...... 1 1.1 Role of apex carnivores in ecosystems ...... 1 1.2 Conservation status of apex carnivores ...... 2 1.2.1 Lion (Panthera leo; Linnaeus, 1758)...... 3 1.2.2 Leopard (Panthera pardus; Linnaeus, 1758) ...... 6 1.2.3 Spotted hyena (Crocuta crocuta; Erxleben, 1777) ...... 9 1.3 Conservation status of apex carnivores in Malawi ...... 12 1.4 The importance of reintroduction ...... 14 1.5 Aim and objectives ...... 16 1.5.1 Aim ...... 16 1.5.2 Objectives ...... 16 1.6 Thesis outline ...... 17 1.7 References ...... 18 Chapter 2 - Study area and reintroduction of study species ...... 34 2.1 Study area ...... 34 2.1.1 Location and history ...... 34 2.1.2 Climate ...... 35 2.1.3 Topography and altitude ...... 35 2.1.4 Geology and soils ...... 35 2.1.5 Watercourses ...... 36 2.1.6 Vegetation ...... 36 2.2 Large carnivore reintroduction into MWR ...... 39 2.2.1 Pre-release management ...... 41 2.4 References ...... 42 2.5 Appendices ...... 44

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Chapter 3 - Early post-release monitoring and post-release breeding of two reintroduced felids in Majete Wildlife Reserve, Malawi ...... 45 3.1 Abstract ...... 45 3.2 Introduction ...... 46 3.3 Methods ...... 48 3.3.1 Study site ...... 48 3.3.2 Methods and materials ...... 49 3.3.3 Data analysis ...... 51 3.4 Results ...... 53 3.4 Discussion ...... 62 3.6 Conclusions ...... 68 3.7 Acknowledgements ...... 68 3.8 References ...... 69 3.9 Appendices ...... 75 Chapter 4 - Home range and habitat selection of reintroduced lion (Panthera leo) and leopard (Panthera pardus) in Majete Wildlife Reserve, Malawi ...... 79 4.1 Abstract ...... 79 4.2 Introduction ...... 80 4.3 Methods ...... 82 4.3.1 Study site ...... 82 4.3.2 Immobilisation and collaring ...... 83 4.3.3 Statistical analysis ...... 84 4.4 Results ...... 87 4.5 Discussion ...... 103 4.6 Conclusion ...... 106 4.7 Acknowledgements ...... 107 4.8 References ...... 107 Chapter 5 - Dietary ecology of three apex predators in Majete Wildlife Reserve, Malawi ...... 116 5.1 Abstract ...... 116 5.2 Introduction ...... 117 5.3 Methods ...... 118 5.3.1 Study area ...... 118 5.3.2 Data collection ...... 119 3.3.3 Data analysis ...... 122 3.4 Results ...... 124

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3.5 Discussion ...... 137 5.5 Conclusion ...... 143 5.6 Acknowledgements ...... 143 5.7 References ...... 144 5.7 Appendices ...... 153 Chapter 6 - Population dynamics and carrying capacity of two reintroduced felids and a resident hyenid in Majete Wildlife Reserve, Malawi ...... 156 6.1 Abstract ...... 156 6.2 Introduction ...... 157 6.3 Methods ...... 159 6.3.1 Study site ...... 159 6.3.2 Fieldwork ...... 159 6.3.3 Statistical analysis ...... 161 6.4 Results ...... 164 6.5 Discussion ...... 170 6.6 Conclusion ...... 174 6.7 Acknowledgements ...... 174 6.8 References ...... 174 6.9 Appendices ...... 185 Chapter 7 - Research findings and implications for African Parks Majete management ...... 191 7.1 Overview ...... 191 7.2 Research findings ...... 192 7.2.1 Reintroduction success – importance of the pre- and post-release phase ...... 192 7.2.2 A few big cats in a small reserve – insights from home range and habitat use analyses .. 192 7.2.3 Population persistence of apex predators – a look at population growth and density .... 193 7.2.4 Predator-prey interactions – are prey limited by predators? ...... 194 7.2.5 Competition for resources – evidence of dietary separation ...... 195 7.3 Management implications ...... 196 7.3.1 Managing large carnivores in MWR ...... 196 7.3.2 Conservation status of lion in Malawi...... 199 7.3.3 The managed metapopulation approach for carnivores in Malawi ...... 200 7.2.4 Which cats go where? – considering the reintroduction of cheetah into MWR ...... 202 7.4 Conclusion ...... 204 7.5 References ...... 204

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7.5 Appendix ...... 211

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List of Tables

Table 2.1. Biological and technical information of reintroduced lion and leopard in Majete Wildlife Reserve, Malawi...... 40 Table 3.1. Biological and practical details of reintroduced lions and leopards in Majete Wildlife Reserve, Malawi...... 50 Table 3.2. Mean ± SD and Kruskal-Wallis ANOVA result of the distance (km) travelled by reintroduced lions and leopards from the release site during the first three months after their release into Majete Wildlife Reserve, Malawi. Due to collar failure, no data were recorded for LEF3 in the third month. 54 Table 3.3. Mean ± SD and ANOVA result of the daily distance (km) travelled by reintroduced lions and leopards during the first three months after their release into Majete Wildlife Reserve, Malawi. Due to collar failure, no data were recorded for LEF3 in the third month...... 55 Table 3.4. Homing tendencies of released lions in Majete Wildlife Reserve, based on Kilian (2003). Home directions falling within the 95% confidence intervals of mean angles of movement indicated homing behaviour...... 56 Table 3.5. Homing tendencies of released leopards in Majete Wildlife Reserve, based on Kilian (2003). Home directions falling within the 95% confidence intervals of mean angles of movement indicated homing behaviour. No data were available in third month for LEM3 due to collar failure...... 56 Table 3.6. Details of the reintroduction, biology and breeding success of released cats in Majete Wildlife Reserve, Malawi...... 63 Table 4.1. Home range size (km2) of reintroduced lions and leopards in Majete Wildlife Reserve, Malawi, as calculated from the kernel utilisation distribution (UD) and minimum convex polygon (MCP) methods during the study period from August 2012 to May 2017. Time taken for individuals to establish home ranges in the reserve and distance between the home range centroid of an individual and their release site (km) are indicated...... 88 Table 4.2. Range size (km2) change of lioness LIF1 for the first three months following the birth of her cubs. This was compared with the total range size for the same lioness during the entire study period from 2012 to May 2017 in MWR...... 89 Table 4.3. Seasonal home range sizes (km2) of reintroduced lions in Majete Wildlife Reserve using kernel utilisation distribution (UD) method. Values in parentheses indicate number of location fixes...... 90 Table 4.4. Seasonal home range sizes (km2) of reintroduced leopards in Majete Wildlife Reserve using kernel utilisation distribution (UD) method. Values in parentheses indicate the number of location fixes...... 94 Table 4.5. Area of overlap (km2) for the home range (95%) and core area (50%), using kernel utilisation distribution (UD), for reintroduced lions and leopards in Majete Wildlife Reserve. Values above the diagonal represent home range overlap, while values below the diagonal represent core area overlap. Percentage spatial overlap is indicated in parentheses. Overlap of lion (LIM1 & LIF1) range with those of leopard (LEM1–3 and LEF1–2) are highlighted in grey...... 98 Table 4.6. Area of overlap (km2) for the home range (95%) and core area (50%), using minimum convex polygon (MCP), for reintroduced lions and leopards in Majete Wildlife Reserve. Values above the diagonal represent home range overlap, while values below the diagonal represent core area overlap.

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Percentage spatial overlap is indicated in parentheses. Overlap of lion (LIM1 & LIF1) range with those of leopard (LEM1–3 and LEF1–2) are highlighted in grey...... 99 Table 4.7. Habitat use by reintroduced lions in Majete Wildlife Reserve, Malawi. Values in bold are significant...... 100 Table 4.8. Habitat use by reintroduced leopards in Majete Wildlife Reserve, Malawi. Values in bold are significant...... 101 Table 5.1. Prey species recorded in lion diet by means of scat analysis and kills found at GPS cluster locations in Majete Wildlife Reserve, Malawi. Frequency of occurrence (FO) was calculated as the percentage of each prey item relative to the total number of prey items recorded (n = 75). Corrected frequency of occurrence (CFO) shows the percentage of occurrences (per scat) relative to the total number of scats collected (n = 50). Number (n = 61) and percentage of kills represented kill site analysis...... 125 Table 5.2. Biomass consumed and total biomass consumed determined from lion scat (n = 50) collected in Majete Wildlife Reserve, Malawi...... 126 Table 5.3. Proportions (%) of age classes and sex of the main prey killed by lions in Majete Wildlife Reserve, Malawi...... 128 Table 5.4. Prey species recorded in leopard scat collected in Majete Wildlife Reserve, Malawi. Frequency of occurrence (FO) was calculated as the percentage of each prey item relative to the total number of prey items recorded (n = 69). Corrected frequency of occurrence (CFO) shows the percentage of occurrences (per scat) relative to the total number of scats collected (n = 42)...... 129 Table 5.5. Biomass consumed and total biomass consumed determined from leopard scat (n = 42) collected in Majete Wildlife Reserve, Malawi...... 130 Table 5.6. Prey species recorded in hyena scat collected in Majete Wildlife Reserve, Malawi. Frequency of occurrence (FO) was calculated as the percentage of each prey item relative to the total number of prey items recorded (n = 205). Corrected frequency of occurrence (CFO) shows the percentage of occurrences (per scat) relative to the total number of scats collected (n = 128)...... 132 Table 5.7. Biomass consumed and total biomass consumed determined from hyena scat (n = 128) collected in Majete Wildlife Reserve, Malawi...... 133 Table 5.8. Dietary overlap and dietary niche breadth of the three apex predators in Majete Wildlife Reserve, Malawi. Values above the diagonal indicate percentage overlap and values below the diagonal represent Pianka’s dietary overlap index (Pianka, 1973). Values are based on actual prey species, while Jacobs’ index of prey preference are indicated in parentheses. Levins’ niche breadth and standardised niche breadth (Levins, 1968) are also indicated...... 135 Table 6.1. Capture details of leopard and hyena during the three-month camera trap survey in Majete Wildlife Reserve, Malawi...... 165

Table 6.2. Results of population abundance estimates using Mo and Mh models in CAPTURE for leopard and hyena in Majete Wildlife Reserve, Malawi. Capture probabilities (p-hat) and population closure assumption tests are also presented...... 166 Table 6.3. Results of population abundance and density estimates (individuals/100 km2) using SPACECAP for leopard and hyena in Majete Wildlife Reserve, Malawi...... 167 Table 6.4. Density estimates (number of individuals/100 km2) calculated for leopard and hyena using the programs CAPTURE and SPACECAP with the half mean maximum distance moved (½ MMDM)

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buffer method based on camera trapping. Density is presented with standard error for CAPTURE estimates and standard deviation for SPACECAP estimates...... 168 Table 6.5. Estimated carrying capacities of leopard, lion and hyena in Majete Wildlife Reserve, Malawi, using range-wide and site-specific preferred prey species and preferred prey weight ranges determined from Jacobs’ indices. Values indicate population sizes and values in parentheses represent densities. Calculations were based on aerial surveys conducted in 2012 and 2015 (see Chapter 5). 169 Table 7.1. Estimated lion population sizes in protected areas in Malawi……………………………..………. 200

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List of Figures

Figure 2.1. National parks, wildlfie reserves and forest reserves representing the protected areas in Malawi; the red block indicating Majete Wildlife Reserve (insert). Map of habitat types, rivers and artificial waterholes in Majete Wildlife Reserve (Sherry, 1989)...... 38 Figure 2.2. Proposed map of habitat types in Majete Wildlife Reserve, Malawi (African Parks Majete (Pty) Ltd., personal communication, January 15, 2017)...... 38 Figure 3.1. Boma and release sites of reintroduced lions and leopards in Majete Wildlife Reserve, Malawi (Shapefiles, personal communication, African Parks (Pty) Ltd.)...... 49 Figure 3.2. Range size (95%, 75% and 50% UD) change of LIM1 during the first three months after his release in Majete Wildlife Reserve, Malawi...... 58 Figure 3.3. Range size (95%, 75% and 50% UD) change of LIF1 during the first three months after her release in Majete Wildlife Reserve, Malawi...... 58 Figure 3.4. Range size (95%, 75% and 50% UD) change of LEM1 during the first three months after his release in Majete Wildlife Reserve, Malawi...... 59 Figure 3.5. Range size (95%, 75% and 50% UD) change of LEM2 during the first three months after his release in Majete Wildlife Reserve, Malawi...... 59 Figure 3.6. Range size (95%, 75% and 50% UD) change of LEM3 during the first two months after his release in Majete Wildlife Reserve, Malawi. Data were not sufficient to produce range sizes for month 3...... 60 Figure 3.7. Range size (95%, 75% and 50% UD) change of LEF1 during the first three months after her release in Majete Wildlife Reserve, Malawi...... 60 Figure 3.8. Range size (95%, 75% and 50% UD) change of LEF2 during the first three months after her release in Majete Wildlife Reserve, Malawi...... 61 Figure 3.9. Range size (95%, 75% and 50% UD) change of LEF3 during the first two months after her release in Majete Wildlife Reserve, Malawi. Data were not sufficient to produce range sizes for month 3...... 61 Figure 4.1. Map of the habitat types and rivers in Majete Wildlife Reserve, Malawi (Shapefiles, personal communication, African Parks (Pty) Ltd.)...... 83 Figure 4.2. Kernel utilisation distribution (95%, 90%, 75% and 50%) area estimation for LIM1 in Majete Wildlife Reserve, Malawi...... 87 Figure 4.3. Kernel utilisation distribution (95%, 90%, 75% and 50%) area estimation for LIF1 in Majete Wildlife Reserve, Malawi...... 87 Figure 4.4. Kernel utilisation distribution (95%, 90%, 75% and 50%) area estimation for LEM1 in Majete Wildlife Reserve, Malawi...... 91 Figure 4.5. Kernel utilisation distribution (95%, 90%, 75% and 50%) area estimation for LEM2 in Majete Wildlife Reserve, Malawi...... 92 Figure 4.6. Kernel utilisation distribution (95%, 90%, 75% and 50%) area estimation for LEM3 in Majete Wildlife Reserve, Malawi...... 92 Figure 4.7. Kernel utilisation distribution (95%, 90%, 75% and 50%) area estimation for LEF1 in Majete Wildlife Reserve, Malawi...... 93

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Figure 4.8. Kernel utilisation distribution (95%, 90%, 75% and 50%) area estimation for LEF2 in Majete Wildlife Reserve, Malawi...... 93 Figure 4.9. Roaming behaviour of LEM1 outside Majete Wildlife Reserve, Malawi. Dots indicate location points and arrow lines indicate the directional movement outside of the reserve...... 96 Figure 4.10. Roaming behaviour of LEM2 outside Majete Wildlife Reserve, Malawi. Dots indicate location points and arrow lines indicate the directional movement outside of the reserve...... 96 Figure 4.11. Roaming behaviour of LEM3 outside Majete Wildlife Reserve, Malawi. Dots indicate location points and arrow lines indicate the directional movement outside of the reserve...... 97 Figure 4.12. Roaming behaviour of LEF2 outside Majete Wildlife Reserve, Malawi. Dots indicate location points and arrow lines indicate the directional movement outside of the reserve...... 97 Figure 5.1. Jacobs’ index (D) indicating preference (+1) and avoidance (–1) of prey species by lion in Majete Wildlife Reserve, Malawi. D-values are derived from the corrected frequency of occurrence and total biomass consumed of prey species from lion scat...... 127 Figure 5.2. Percentage (%) of different prey age classes and sex killed by lions (n = 61) in Majete Wildlife Reserve, Malawi...... 128 Figure 5.3. Jacobs’ index (D) indicating preference (+1) and avoidance (–1) of prey species by leopards in Majete Wildlife Reserve, Malawi. D-values are derived from the corrected frequency of occurrence and total biomass consumed of prey species from leopard scat...... 131 Figure 5.4. Jacobs’ index (D) indicating preference (+1) and avoidance (–1) of prey species by hyenas in Majete Wildlife Reserve, Malawi. D-values are derived from the corrected frequency of occurrence and total biomass consumed of prey species from hyena scat...... 134 Figure 5.5. Mean preferred prey weight (kg), using CFO and biomass calculations, for the three apex predators in Majete Wildlife Reserve, Malawi. Jacobs’ index values were used to determine preferred prey species. Standard error bars are presented above and below means...... 135 Figure 5.6. Frequency of occurrence (%) of prey weight classes (kg) in the diet of three apex predators in Majete Wildlife Reserve, Malawi. Prey class 0–5 kg was excluded as it only contributed to leopard diet (<5%)...... 136 Figure 6.1. Camera sites (n = 96) were divided into four equal and consecutively sampled grids within Majete Wildlife Reserve. The ½ MMDM buffers of leopards and hyenas used to create effectively sampled areas...... 162 Figure 6.2. Camera traps situated within Majete Wildlife Reserve with potential home range centres for leopard and hyena. Suitable habitat is demarcated in green and unsuitable habitat in red. State space boundaries were created using the ½ MMDM buffer for both species as well as a 15 km buffer for analysis in SPACECAP...... 164 Figure 7.1. Regional node within Malawi consisting of Majete Wildlife Reserve, Liwonde National Park and Nkhotakota Wildlife Reserve (Shapefiles, personal communication, African Park (Pty) Ltd.). ... 203

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List of Appendices

Appendix 2A. Lion boma located in the north-eastern section of MWR...... 44 Appendix 2B. Leopard boma situated in the north-eastern section of MWR...... 44 Appendix 2C. Crush boxes were designed as an access point for providing carcasses to leopards in the bomas...... 44 Appendix 3A. Comparison of three methods used to determine homing tendencies of lions in Majete Wildlife Reserve. Home directions falling within the 95% confidence intervals of mean angles of movement indicated homing behaviour (method 1 and 2), whereas mean angles of direction falling within a 22.5° range of the home direction indicated homing (method 3)...... 75 Appendix 3B. Comparison of three methods used to determine homing tendencies of leopards in Majete Wildlife Reserve. Home directions falling within the 95% confidence intervals of mean angles of movement indicated homing behaviour (method 1 and 2), whereas mean angles of direction falling within a 22.5° range of the home direction indicated homing (method 3)...... 76 Appendix 3C. Range establishment (km2) of reintroduced lions during the first three months after their release into Majete Wildlife Reserve, Malawi...... 77 Appendix 3D. Range establishment (km2) of reintroduced leopards during the first three months after their release into Majete Wildlife Reserve, Malawi...... 78 Appendix 5A. List of all mammalian prey species occurring within Majete Wildlife Reserve. Figures include reintroduced animals and population estimates from aerial surveys conducted in 2012 and 2015...... 153 Appendix 5B. Biomass and biomass consumed based on lion kills (n = 61) collected at GPS cluster sites in Majete Wildlife Reserve...... 154 Appendix 5C. Jacobs’ index (D) indicating preference (+1) and avoidance (–1) of prey species in lion diet using biomass calculations, based on kill site analysis in Majete Wildlife Reserve, Malawi...... 155 Appendix 5D. Jacobs' preference index using CFO and biomass calculations from the scat of the three apex predators in Majete Wildlife Reserve, Malawi...... 155 Appendix 6A. Examples of photographs used to identify individuals based on unique pelage patterns. The top inserts (a & b) depict two left side photographs of the same male, while the bottom inserts (c & d) show clear differences in left side photographs of an adult male and younger male respectively. The same method was used to identify spotted hyenas...... 185 Appendix 6B. Example of the input file 'Animal Capture Details' required by SPACECAP. A sample of the capture details are represented for leopard on the left (a) and hyena on the right (b)...... 186 Appendix 6C. Sample of the input file ‘Trap Deployment Details’ required by SPACECAP. Binary values indicate whether camera traps were functioning (1) or not (0)...... 187 Appendix 6D. Sample of the input file 'Potential Home-Range Centres' required by SPACECAP to determine potential home range centres. Binary values indicate suitable (1) and unsuitable habitat (0) for both leopards and hyenas...... 188 Appendix 6E. Camera traps and capture locations of leopards and hyenas in Majete Wildlife Reserve, Malawi…………………………………………………………………………………………………………………………………………..189

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Appendix 6F. Jacobs’ index-based preferred prey species and preferred prey weight range for leopard, lion and spotted hyena, based on range-wide studies and a site-specific study in Majete Wildlife Reserve, Malawi...... 190 Appendix 7A. Guidelines for future reintroductions or translocations of large carnivores...... 211

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Chapter 1 Introduction

1.1 Role of apex carnivores in ecosystems

Apex mammalian carnivores are defined as species occupying the highest trophic level in a community (Ordiz, Bischof & Swenson, 2013; Ritchie & Johnson, 2009). These carnivores play a pivotal role in maintaining trophic balance (Beschta & Ripple, 2009; Estes et al., 2011; Miller et al., 2001; Ripple & Beschta, 2012; Terborgh et al., 1999), as they are able of controlling lower trophic levels such as herbivore and mesopredator populations through predation and competition (Miller et al., 2001; Polis, Myers & Holt, 1989). These ecosystem effects are known as “top-down” processes and filter downward through trophic levels, maintaining diversity (Estes, Crooks, & Holt, 2001; Paine, 1966). Therefore, apex carnivores drive food-web pathways and ultimately structure ecosystems (Ripple et al., 2014; Terborgh, 1988).

Carnivore populations are also naturally regulated by “bottom-up” processes (e.g. resource limitation; Kissui & Packer, 2004). Essentially, a system dominated by bottom-up processes, controls the upward movement of energy from lower to higher trophic levels (Miller et al., 2001). However, bottom-up processes do not merely act in isolation, but can also occur concomitantly with top-down processes in the same system (Terborgh et al., 1999). The decline of prey in a system will likely result in a decline in predator abundance through bottom-up limitation (Karanth, Nichols, Kumar, Link & Hines, 2004), whereas the removal of an apex predator (or entire predator guild) will invariably lead to appreciable increases in prey populations (Terborgh et al., 1999).

One of the most well-known examples of the impact that apex predators have on an ecosystem is the effect of the reintroduction of grey wolf (Canis lupus) to Yellowstone National Park. The extirpation of wolves from Yellowstone National Park, USA in the 1990s resulted in significant vegetation damage due to the release of herbivore populations from predation pressures (Beschta, 2005; Ripple & Larsen, 2000). Due to the long generation times of large carnivores, it may take several years or decades to observe the responses of lower trophic levels to their removal, often when apex predators have lost their ability to restore ecosystem functioning (Estes et al., 2011). Removing apex predators from the environment clearly affects a multitude of floral and faunal species in a “ripple-like” effect, even those seemingly distant from the predator, thus highlighting their importance in the ecosystem (Miller et al., 2001; Terborgh, 1988).

Large carnivores compete for similar limited resources which may lead to interference or exploitative competition (Frame, 1986; Linnell & Strand, 2000). Interference competition is the direct negative interaction resulting from the utilisation of a shared resource, where the ability of a species to use a 1

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resource is altered by other species (Case & Gilpin, 1974; Vance, 1984). This form of competition includes kleptoparasitism (Cooper, 1991), interspecific territoriality and aggression (Durant, 1998; Kruuk, 1972; Mills, 1984), and intra-guild predation (Palomares & Caro, 1999). Exploitative competition is the indirect negative interaction, where a species gains food resources faster than others (Case & Gilpin, 1974; Vance, 1984). A limitation of food and extensive dietary overlap between large carnivores may enhance this kind of competition among species (Hayward & Kerley, 2008). However, sub- dominant carnivores may adopt a strategy known as resource partitioning to reduce potential conflict with dominant carnivores (Schoener, 1974) and promote coexistence (Kitchen, Gese & Schauster, 1999). Resource partitioning is achieved by a division in selection for prey species and size (Karanth & Sunquist, 1995, 2000; Hayward & Kerley, 2008), feeding behaviours (Stander, Haden, Kaqece & Ghau, 1997; Stein, Bourquin & McNutt, 2014), activity patterns and space use (Cristescu, Bernard & Krause, 2013; Steinmetz, Seuaturien & Chutipong, 2013). Thus the coexistence of multiple large carnivores is dependent on a trade-off between competitive interactions and partitioning of resources.

Large carnivores may naturally occur in multi-species assemblages or guilds (Simberloff & Dayan, 1991; Wilson, 1999). Large carnivore guilds were once widely distributed across all five continents; however extinctions during the Late Pleistocene and the modern era diminished many of these guilds (Koch & Barnosky, 2006; Turvey & Fritz, 2011). Today, the African large carnivore guild is the only relatively complete guild in the world (Dalerum, Cameron, Kunkel & Somers, 2009; Valkenburgh, 1988) and includes the lion (Panthera leo), spotted hyena (Crocuta crocuta), leopard (Panthera pardus), African wild dog (Lycaon pictus) and cheetah (Acinonyx jubatus). Despite this, all members are currently suffering population declines (Ripple et al., 2014). Additionally, intact guilds are required for these species to fulfil their role as keystone species (i.e. species with a greater than expected impact on the ecosystem relative to their biomass; Dalerum, Somers, Krunkel & Cameron, 2008; Paine, 1966). Therefore, systems that host a diversity of apex carnivores present unique opportunities to study their ecology and to deepen our understanding of their role in the ecosystem.

1.2 Conservation status of apex carnivores

From the late Pleistocene to the present era, large-bodied animals, in particular large felids, canids, and ursids have been disappearing (Weber & Rabinowitz, 1996) in what is being called the “sixth mass extinction” (Barnosky et al., 2011; Wake & Vredenburg, 2008). Although extinctions have occurred naturally throughout evolutionary history (Raup & Sepkoski, 1982), this extinction is unique as mankind is largely responsible for the large-scale decline of these animals (Wake & Vredenburg, 2008). Since large carnivores are often keystone species the implications of their loss are not limited to a single species, but rather affect entire ecosystems (Estes et al., 2011).

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The worldwide decline of large predators is a major conservation concern (Weber & Rabinowitz, 1996), as large carnivores have suffered dramatic population declines, geographic range reductions and habitat degradation over the last 200 years (Ceballos & Ehrlich, 2002; Morrison, Sechrest, Dinerstein, Wilcove & Lamoreux, 2007; Ripple et al., 2014; Woodroffe, 2000). According to the International Union for the Conservation of Nature (IUCN), 61% of the world’s largest carnivores are listed as threatened, and 77% are suffering population declines (Ripple et al., 2014). One of the most dramatic examples is the accelerated loss of Asia’s largest carnivore, the tiger (Panthera tigris; Dinerstein et al., 2007). Tigers have suffered a 42% range decline since 2006 and inhabit <6% of their former range (Walston et al., 2010), with only 2 154–3 159 mature individuals estimated in the wild (Goodrich et al., 2015).

Like the rest of the world, Africa has experienced declines in large carnivore species. While these reductions are not as drastic, all members of the large carnivore guild have suffered range and population reductions, some more extreme than others (Morrison et al., 2007; Ray, Hunter & Zigouris, 2005; Ripple et al., 2014). Conclusively, further population and range reductions may lead to local population extinctions and eventually possible species extinctions.

1.2.1 Lion (Panthera leo; Linnaeus, 1758)

Biology

African lions are the largest terrestrial predators on the continent (Sunquist & Sunquist, 2009). They are extremely powerful cats, with females weighing 126 kg on average and males weighing up to 225 kg (Smuts, 1982). This weight difference and the males’ characteristic mane, makes this felid a sexually dimorphic species (Sunquist & Sunquist, 2009). Lions are extinct from North Africa and are thus only found south of the Sahara Desert. They have a wide habitat tolerance and occur in various environments ranging from forests to deserts (Nowell & Jackson, 1996), and have been recorded at elevations of up to 4 240 m in the Bale Mountains in Ethiopia (Yalden, Largen & Kock, 1980). However, lions are savannah specialists, favouring more open habitat such as grasslands, scrub and open- to closed-woodlands (Nowell & Jackson, 1996). Lions require sufficient vegetation cover for hunting (Nowell & Jackson, 1996) and drink water if available, but can obtain moisture requirements from prey and plants (Eloff, 1973a). They are intolerant to human modified habitats and are thus largely restricted to protected areas in most countries (Riggio et al., 2013).

Lions are the most social felid and live in complex social groups called “prides” containing 2–18 individuals (Mosser & Packer, 2009; Packer, Gilbert, Pusey & O’Brien, 1991). A pride consists of related females (none dominant) and their cubs (Mosser & Packer, 2009; West & Packer, 2013); and the size of a pride is measured by the number of adult females (Nowell & Jackson, 1996). Pride membership is stable, although pride members may be subdivided into smaller subgroups when members scatter within the pride’s range. These matrilocal societies are therefore known as fission-fusion groups

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(Mosser & Packer, 2009; West & Packer, 2013). Sociality in lions appears to have multiple benefits such as cooperative hunting (Schaller, 1972), nursing one another’s cubs (“allosuckling”: Pusey & Packer, 1994) and defence of pride cubs from infanticide and predation (Packer, Scheel & Pusey, 1990).

Lions have an aseasonal breeding pattern (Bertam, 1975), although seasonal peaks have been recorded in some areas (e.g. Smuts, Hanks & Whyte, 1978). Females only mate with unrelated resident males and may leave the pride if resident males are related, to avoid inbreeding (Packer et al., 1988). After a gestation period of 110 days (Cooper, 1942), females give birth to between one and four cubs (Pusey & Packer, 1987), each weighing 1.2–2.1 kg (Schaller, 1972). Cubs are highly vulnerable to predation and infanticide, which leads to a cub mortality of up to 73% (van Orsol et al., 1985). Females which lose entire litters will resume sexual activity soon after, while females that raise dependent young will only start mating between 18 and 24 months after the previous mating event (Betram, 1975; Schaller, 1972). Females become reproductively mature at 24 months, while spermatogenesis only begins at 30 months in males (Smuts et al., 1978). Individuals are fully mature between three and four years of age (Bertram, 1975), with males and females living up to 14 and 18 years, respectively (Packer et al., 1988).

Male lions play an important role in population dynamics. For instance, newcomer males kill dependent cubs and evict older cubs after a pride take-over, to induce estrus in females (Bertram, 1975; Packer & Pusey, 1983). These evicted males become nomadic and remain solitary or form small, highly social groups called coalitions, consisting of one to six related or unrelated individual males (Bygott, Bertram & Hanby, 1979; Pusey & Packer, 1987). Coalitions start challenging at four to six years old and the larger the coalition size, the greater the likelihood of successfully evicting resident males (Pusey & Packer, 1987; Schaller, 1972) with a resultant higher reproductive success (Bygott et al., 1979; Packer et al., 1988). If the new coalition is successful, it will hold tenure (territory and breeding rights) over one or more female prides within its territory (Mosser & Packer, 2009; Packer et al., 1991), which is vigorously defended against competitors (Schaller, 1972). Average tenure lasts two years (Packer et al., 1988), although tenure length is largely determined by coalition size, as solitary males or two-male coalitions will hold tenure for half as long as coalitions of four to six males (Bygott et al., 1979). As in pre-tenure periods, coalitions remain highly social and hunt and scavenge cooperatively during tenure periods (Bygott et al., 1979; Hanby & Bygott, 1987).

Lions often scavenge and kleptoparasitise food from other predators, but they are efficient hunters (Schaller, 1972). Lions typically stalk their quarry and run short distances, less than 100m, at speeds of up 45–60 km/h (Guggisberg, 1961; Schaller, 1972). They may hunt alone although success rate increases from 17–19% when hunting singly to 30% when cooperating in a group (Schaller, 1972; Stander, 1992). Females typically do most of the hunting, cooperating in groups with a complex division of labour (Stander, 1992); however, males are just as effective in killing their own prey (Schaller, 1972). 4

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Lions are generalist predators that take a broad range of prey from birds, reptiles and rodents (Eloff, 1973b; Nowell & Jackson, 1996; Sunquist & Sunquist, 2009), to young black rhinoceros (Diceros bicornis; Brain, Forge & Erb, 1999) and elephant (Ruggiero, 1991). However, the bulk of their diet usually consists of no more than five medium-to large-sized antelope species (Hayward & Kerley, 2005; Schaller, 1972).

Lions are nocturnal predators, intensely active for three or four hours per day (Schaller 1972). They are highly inactive for the other 20–21 hours of the day (Schaller, 1972), particularly in the heat of the day when they sleep and rest in shaded areas. Home range size varies from 26 km² in high prey density areas such as the Serengeti-complex (Rudnai, 1973) to over 4 500 km² in low prey density areas such as the Kgalagadi Transfrontier Park, South Africa/Botswana (Funston, 2011). Males and females defend territories against alien conspecifics of the same sex (West & Packer, 2013). This is typically done by scent marking (leaving urine and/or scat), scraping the ground (Schaller, 1972) and using vocalisations (Grinnell, Packer & Pusey, 1995), which can be heard up to 4 km (Schaller, 1972). Males normally patrol the boundaries while females defend the core area (West & Packer, 2013).

Conservation status

The lion has been listed as Vulnerable since its induction into the Red Data List in 1996 till present (Bauer, Packer, Funston, Henschel & Nowell, 2016). This species was traditionally divided into two subspecies, the Asian subpopulation (P. l. persica) and African subpopulation (P. l. leo), though there is debate over the accuracy of this classification (Bauer et al., 2016). More recently, Barnett et al. (2014) proposed a single origin model in southern-East Africa and recognise five major phylogeographical groups (North African/Asian, West African, Central African, southern African, and East/southern African) of the modern lion. However, these results were based entirely on mtDNA and further genetic sampling is recommended to provide an accurate taxonomic classification (Bauer et al., 2016). The species has been provisionally split into two subspecies: P. l. leo of Asia as well as Central, West and North Africa and P. l. melanochaita from southern and East Africa (Bauer et al., 2016).

Historically, lions were widely distributed across Africa, extending into Eurasia and the southern United States (Barnett et al., 2009). Today, lions are only present in Africa, with the exception of a single population remaining in the Gir Forest National Park and Wildlife Sanctuary (1 400 km²) in India (Nowell & Jackson, 1996). The African lion persists in only 8% of their former habitat, with an estimated 23 000–39 000 individuals remaining in the wild (Bauer et al., 2016). Lions are now extinct in North Africa and so their stronghold lies in sub-Saharan Africa. However, lions south of the Sahara are not exempt from extinction as they are now absent from 12 of these countries, while several are on the verge of extinction (Bauer et al., 2015). The importance of subdividing this species into geographical context, by assigning Evolutionary Significant Units (ESUs), cannot be emphasised enough. For 5

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example, lions in southern Africa (which account for a quarter to a third of lions in Africa) appear to be relatively stable and have increased by 12% in four countries, namely Botswana, Namibia, South Africa and Zimbabwe (Bauer et al., 2016). In contrast, lion populations in Central Africa have declined to fewer than 900 individuals (Bauer, de Iongh, Princée & Ngantou, 2003) and the genetically distinct lions of West Africa are now classified as Critically Endangered with only 500 individuals remaining (Riggio et al., 2013; Henschel et al., 2014).

In East Africa, initial population estimates ranged from 11 167 (Bauer & van Der Merwe, 2004) to 20 485 (Chardonnet, 2002) and more recently, populations are estimated at 19 000 (Riggio et al., 2013). This suggests that East Africa represents an important stronghold for African lion populations (Bauer et al., 2016). Regardless of these relatively large population estimates, the decline of lions in East Africa is now apparent and data suggest that lions are endangered in this region with populations likely to decline by 50% over the next two decades (Bauer et al., 2015). These declines are largely due to habitat degradation, prey reduction (Ray et al., 2005) and retaliatory killing of lions in response to the potential threat to livestock and/or humans (Patterson, Kasiki, Selempo & Kays, 2004; Woodroffe & Frank, 2005), with unsustainable trophy hunting (Packer et al., 2009, 2011) and susceptibility to disease due to genetic impoverishment (Packer, Pusey & Rowley, 1991) considered localised or less significant threats.

1.2.2 Leopard (Panthera pardus; Linnaeus, 1758)

Biology

Leopards are large (21–90 kg) carnivores with the widest geographic distribution of all wild felids, occurring throughout the majority of sub-Saharan Africa (Myers, 1976; Nowell & Jackson, 1996). They are considered the most successful large carnivore in Africa, largely due to their generalist feeding behaviour (Nowell & Jackson, 1996) and broad habitat tolerance (Skinner & Chimimba, 2005). Leopards occupy woodland, grassland-savanna complexes and rainforests (Ray, Hunter & Zigouris, 2005), with highest densities occurring in riparian habitats (Bailey, 1993; Turnbull-Kemp, 1967). They may also be found in desert habitats, coastal scrub at sea level and alpine areas (Ray et al., 2005), with an upper elevation limit of 5 638 m (Guggisberg, 1975). They do not require access to drinking water as they obtain sufficient moisture requirements from their prey or vegetation (Bothma & le Riche, 1986). Leopards are relatively tolerant to anthropogenic disturbance, often persisting in human- modified areas and in close proximity to human cities (Kuhn, 2014).

Leopards are typically solitary cats, except when females raise cubs or when males and females mate (Myers, 1976). They have no particular breeding season, although leopard births may be synchronised with the birth season of impala Aepyceros melampus, one of their main prey items (Nowell & Jackson, 1996). After a gestation period of 96 days on average (Owen, Niemann & Slotow, 2010), between one

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and four (mean = 1.7) cubs are born in dense vegetation or in holes in the ground or in caves (Eaton, 1977; Owen et al., 2010; Turnbull-Kemp, 1967). Bailey (1993) found that cubs suffer from a high mortality rate (>50%) in the Kruger National Park, South Africa, though this rate decreases into sub- adulthood (32%) and adulthood (19%). Cubs are led by their mother to kills until they are nine months old and generally start hunting at 11 months (Skinner & Chimimba, 2005). A 40-year dataset from the Sabi Sand Game Reserve, South Africa, revealed that average parental care to litters was 18 months, although this varied substantially from 9 to 35 months (Balme, Robinson, Pitman & Hunter, 2017). In addition, Balme et al. (2017) found that female leopards displayed extended parental care when prey were scarce, while providing prolonged care for sons. Dispersal may also be delayed by several months if food is abundant or if conspecifics occur at high densities (Bailey, 1993). After separation, males will cover a larger distance than females as they aim to find greater home ranges, which can influence gene flow in leopard populations (Martins, 2006; Fattebert et al., 2016). Leopards are sexually mature from 24 months and can live up to 18 years of age (Balme et al., 2013; Hunter, Henschel & Ray, 2013).

Leopards are secretive and stealthy predators that rely on an opportunistic hunting behaviour (Bothma & le Riche, 1984; Myers, 1976). Most hunts occur at night in open habitat (Bailey, 1993), although hunting occurs during the daylight hours in tropical rainforests (Henschel & Ray, 2003; Jenny & Zuberbühler, 2005). Leopards use camouflage and vegetation to stalk very close to their prey before sprinting, up to a speed of 60 km/h within 120 m and pouncing on their prey (Betram, 1979; Stander, Haden, Kaqece & Ghau, 1997). Hunting success varies from 5% in the Serengeti (Bertram, 1979) to 38% in Kaudom (Stander et al., 1997) and captured prey are sometime hauled into trees out of reach of other large carnivores, or cached beneath vegetation (Stander et al., 1997; Balme, Miller, Pitman & Hunter, 2017). Leopards have the most varied diet of any other African large carnivore, with a total of 92 prey species recorded in their diet in sub-Saharan Africa (Bailey, 1993). These include dung beetles, fish, birds, reptiles and small-, medium- and large-bodied ungulates (Fey, 1964; Hirst, 1969; Kingdon, 1977; Mitchell, Shenton & Uys, 1965; Ott, Kerley & Boschoff, 2007; Scheepers & Gilchrist, 1991). They also have the ability to subsist entirely on smaller-sized prey such as rodents in areas where larger and medium prey species are absent (Brown, 1971).

Leopard home range size varies from 10 km² to several hundred kilometres (Stuart & Stuart, 2007). The size of their home ranges is dependent on food availability, habitat quality for females and access to females for males (Bailey, 1993). Males have larger home ranges than females and multiple females may occur within the range of a single male (Bailey, 1993; Fattebert et al., 2016). Range overlap varies within and between seasons (Stander et al., 1997) and the ranges of females more so than those of males (Bothma & Coertze, 2004; Stander et al., 1997). Males also hold larger territories than females (Stander et al., 1997) which are defended by scent marking and vocalisations (Hunter et al., 2013). Leopards are mostly active between sunset and sunrise (Bailey, 1993; Nowell & Jackson, 1996) and

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rest on the ground or in trees during the day (Bailey, 1993). Human activity may affect leopard activity patterns (Henschel & Ray, 2003; Hunter et al., 2013). Leopard density varies considerably across Africa and is dependent on the number of conspecifics, as well as other carnivores, prey density and energetic requirements (Hunter et al., 2013; Stander et al., 1997).

Conservation status

The conservation status of the leopard has changed on numerous occasions since its introduction into the Red List of Threatened Species as Vulnerable in 1986 (Stein et al., 2016). In 1996 and 2002 leopards were listed as Lower Risk and Least Concern respectively, as population estimates at the time were greatly overestimated (Stein et al., 2016). In 2008, leopards were subsequently listed as Near Threatened as more data became available to provide more reliable population estimates (Stein et al., 2016). Today leopards remain widespread throughout Africa and Asia although major range reductions have resulted in reduced population sizes and in some cases isolation or extirpation of populations (Ray et al., 2005). Due to the anthropogenic impacts opposing the leopard, causing an estimated range reduction of more than 30% since its previous assessment in 2008, leopards were restored to their initial listing as Vulnerable in 2016 (Stein et al., 2016).

The IUCN recognises eleven leopard subspecies within P. pardus (Miththapala, Seidensticker & O’Brien, 1996; Uphyrkina et al., 2001). The Eurasian populations are divided into ten subspecies (Khorozyan, Gennady, Baryshnikov & Abramov, 2006) and are thought to be genetically distinct from Africa’s P. p. pardus (Linnaeus, 1758), which is the only subspecies in Africa (Miththapala et al., 1996; Uphyrkina et al., 2001). However, Khorozyan et al. (2006) recommends additional data for the accurate taxonomic classification of leopard subpopulations.

African leopards have lost 48–67% of their former range and populations have declined noticeably within their range (Jacobson et al., 2016). While the genetics of the African leopard indicates a single subspecies, geographically it is subject to different pressures that affect populations at different levels. For example, southern Africa probably has the healthiest leopard population across their range with continuous distributions in several countries (Stein et al., 2016), while leopards in North Africa have suffered a 99% range reduction with only small, isolated populations remaining (Jacobson et al., 2016). The distribution of leopards in the tropical belt of West and Central Africa has also markedly decreased due to increased human activities, such as construction of road networks in forested areas for logging companies; making these areas more accessible for bushmeat poachers and illegal trophy hunters (Henschel & Ray, 2003). Leopard populations are expected to continue declining across Africa, with extinction likely in both North and West Africa (Stein et al., 2016).

Leopard populations in East Africa are relatively stable compared to other parts of Africa, yet the general population trend in this region is continually decreasing, which can be seen by their absence

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from countries such as Somalia and Eritrea (Stein et al., 2016). Across the region, on-going anthropogenic pressure, for example habitat fragmentation and land conversion, is likely to further reduce populations (Stein et al., 2016). Leopards are highly dependent on prey availability within their range (Marker & Dickman, 2005) and the estimated loss of 52% of potential prey may result in a 50% decline of leopard populations in East Africa (Craigie et al., 2010). Leopards were formerly regarded as ‘vermin’ due to livestock depredation and although they are legally protected today, direct persecution remains a significant threat to populations (Ray et al., 2005). Since a large proportion of leopards occur outside of protected areas (Stein et al., 2016), conflict with humans remains a major conservation concern for this species in East Africa, as well as the rest of their range. Furthermore, threats such as poorly managed trophy hunting (Balme, Slotow & Hunter, 2009) and illegal hunting of leopards for the trade in their skins (for traditional ceremonies) and various body parts (for medicinal purposes) can have detrimental impacts on local leopard populations (Stein et al., 2016).

1.2.3 Spotted hyena (Crocuta crocuta; Erxleben, 1777)

Biology

Spotted hyenas (hereafter hyena) are the largest member of the Hyaenidae family, weighing more than 70kg and are second only to lion with regards to body size (East & Hofer, 2013; Mills, 1990). This gregarious species is one the most abundant large carnivores in sub-Saharan Africa and inhabits much of its former range (Estes, 1991; Holekamp & Dloniak, 2010). Hyenas have a wide habitat tolerance including woodland, savanna, semi-deserts, wetlands and mountainous areas up to 4 000 m (Kruuk, 1972). The highest densities of hyena occur in the savanna plains of Kenya and Tanzania as well as in the montane forests of Kenya (East & Hofer, 2013; Young & Evans, 1993). They are absent or rare in extreme desert conditions, high altitude alpine habitats, dense forest habitats and tropical rainforests (Mills & Hofer, 1998; Henschel & Ray, 2003). Hyenas need regular access to drinking water in most areas, except in dry regions where they apparently rely on the moisture from their prey (Green, Anderson & Whateley, 1984).

Hyenas live in complex family groups called ‘clans’ (Holekamp, Smale, Berg & Cooper, 1997; Kruuk, 1972) containing 5–90 individuals (Holekamp & Smale, 1998). The fundamental unit of a hyena clan comprises of one or more matrilineal kin groups of related adult females and their young (Frank, 1986a; Mills, 1990). Additional clan members include one to several immigrant males which are not part of the stable core group and these males move between clans (Frank, 1986a). The hierarchical structure in a hyena clan is strict linear dominance among each sex (Kruuk, 1972), which is unique to large carnivore species and comparable to old-world primates, whereby rank determines access to essential resources (Frank, 1986b; Kruuk, 1972; Tilson & Hamilton, 1984). In these matriarchal societies, females dominate males both in body size and rank (Drea & Frank, 2003). Female hyenas

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have an elongated hypertrophied clitoris called a pseudo-penis, which resembles the shape and size of the male’s penis (East, Hofer & Wickler, 1993). Females also have a false scrotum from the fusion of the labia majora, hence their masculinity (Glickman, Cunha, Drea, Conley & Place, 2006). Females are also philopatric (they remain in their natal clan), thus retaining their social rank (Kruuk, 1972), while their offspring occupy rank directly below them (Holekamp & Smale, 1993). Males usually emigrate from their natal clan to a nearby clan after reaching reproductive maturity where they attain the lowest possible rank, even below the cubs (Frank, 1986b; Drea & Frank, 2003). Clans are relatively stable, but are characterised by fission-fusion dynamics because travelling, resting and foraging of clan members occurs in subgroups that regularly change composition (Kruuk, 1972; Mills, 1990).

Hyenas use dens previously occupied by species such as porcupine (Hysterix africaeaustralis), aardvark (Orycteropus afer) or warthog (Phacochoerus aethiopicus; Pokines & Peterhans, 2007). The first type of den is a natal (or nursing) den which provides cubs with warmth during the first few weeks of their development (Holekamp & Smale, 1998) and the small entrances to the den prevent predation by other large carnivores such as lion (Holekamp & Dloniak, 2010). Natal dens are located at varying distances from the clan’s communal den (Pokines & Peterhans, 2007), which is typically larger and consists of multiple entrances and an extensive network of underground chambers and tunnels (East, Hofer & Türk, 1989; Kruuk, 1972; Mills, 1990). As with natal dens, communal dens serve as a refuge for cubs when threats are detected (East et al., 1989). Communal dens are the centre for social learning and development and are used in a fluid style with periodic occupation (Pokines & Peterhans, 2007).

The hyena mating system is polygynandrous as both males and females mate with several partners (Holekamp & Smale, 1998). Hyenas are not seasonal breeders (Drea & Frank, 2003), although seasonal peaks have been documented in some areas (Stuart & Stuart, 2007). When a female is ready to give birth after a gestation period of 110 days (Kruuk, 1972), she leaves the clan’s communal den and gives birth to one, two or rarely three cubs in a nursing den (East et al., 1989). Within the first four weeks at the natal den, the mother carries her young individually, in her mouth, to the clan’s communal den (Holekamp & Dloniak, 2010). Cubs are introduced to all clan members at the communal den and rank- relationships are established (Cooper, 1993). Cubs leave the communal den for the first time from eight months of age and accompany their mother on exploratory walks to acquire information on the physical environment, clan boundaries and clan members (Holekamp & Smale, 1998). A major challenge for cubs and subadults is to acquire sufficient solid food to sustain them until they are able to capture their own prey (Kruuk, 1972). Hyenas are reproductively mature at 24 months and males disperse from their natal clans to investigate resource availability of neighbouring clans to assess possibility of integration, while females remain in their natal clan where they will have their own offspring (Holekamp & Smale, 1998). The lifespan of a hyena is 18 years which permits up to five generations within a clan (Holekamp & Dloniak, 2010; Holekamp & Smale, 1998).

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Hyenas are successful predators with flexible foraging strategies. They regularly scavenge and kleptoparasitise kills abandoned by other carnivores, especially when densities of other large carnivores are high (Kruuk, 1972), but hunt alone or in groups when densities are low (Hayward, 2006; Kruuk, 1966). Hyenas are cursorial predators that search for the weakest individual in the herd and run them down over several kilometres at speeds of up to 65km/h (Kruuk, 1972; Mills, 1990). On average one third of their hunts are successful (Holekamp et al., 1997), although hunting success largely depends on the hunting technique as well as the prey species (Kruuk, 1972; Mills, 1990). They require 3.8–4.0 kg of meat per day in order to maintain body condition (Henschel & Tilson, 1988) and do so by feeding on virtually any bird, reptile, fish or mammal available (Henschel & Skinner, 1990; Mills, 1990; Sillero-Zubiri & Gottelli, 1992). Hyenas are capable of killing prey as large as adult African buffalo (Pienaar, 1969), black rhinoceros calves (Berger & Cunningham, 1994) and new-born elephants (Salnicki, Teichmann, Murindagomo & Wilson, 2001), but medium- to large-sized antelope species are more common prey items (Hayward, 2006). Hyenas often kill domestic livestock (Abay, Bauer, Gebrihiwot & Deckers, 2011) and can adapt to foraging on anthropogenic food resources in human- dominated areas when their natural prey base has been depleted (Yirga et al., 2012).

Hyena activity varies much across the African continent, although nocturnal and crepuscular peaks are most common (Kolowski, Katan, Theis & Holekamp, 2007; Kruuk, 1972; Mills, 1990). They spend most of the day resting and periods of activity are interrupted by periods of resting (Holekamp & Dloniak, 2010; Kruuk, 1972). Hyenas can move up to 70 km in a single day (Hofer & East, 1993) and males usually move more than females (Kolowski et al., 2007). Territory sizes range from nine to more than 1 000 km2 (East & Hofer, 2013) and size is probably related to prey availability (Hayward, Hayward, Druce & Kerley, 2009). Territories serve as breeding and foraging zones (Hayward et al., 2009) and clans advertise and defend territories through vocalisations, scent marking and border patrols (East & Hofer, 2013; Kruuk, 1972).

Conservation status

The IUCN first listed the spotted hyena (hereafter hyena) as Lower Concern/Conservation Dependent in 1996 and was subsequently down-listed to Least Concern in 2008, where it remained for the following re-assessment in 2015 (Bohm & Höner, 2015). The current classification of hyenas as Least Concern is due to viable populations exceeding 10 000 mature individuals in the wild (Bohm & Höner, 2015). Despite morphological and genetic variation within the species, no region-specific characteristics have been identified to distinguish subspecies (East & Hofer, 2013).

Hyenas have a high degree of ecological plasticity which allows them to inhabit much of their former range and maintain relatively stable populations (Kolowski & Holekamp, 2006). However, mounting threats may further fragment their range and reduce populations, making them more dependent on the continued existence of protected areas (Mills & Hofer, 1998). The classification of hyenas varies 11

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from ‘vermin’ in Ethiopia to ‘protected’ in other parts of Africa (Mills & Hofer, 1998). Therefore, while some populations remain protected, others may be totally extirpated in the near future (Bohm & Höner, 2015).

The historic range of hyena extended from South Africa to Western Europe and Asia (Stiner, 2004). However, today the range of hyenas is confined to the African continent, south of the Sahara (Bohm & Höner, 2015). Hyenas are relatively widespread across their current range, although their distribution is patchy, particularly in West and Central Africa (Bohm & Höner, 2015; Ray et al., 2005). Hyena populations are tentatively estimated between 27 000 and 47 000 individuals (Mills & Hofer, 1998). Populations are relatively stable in parts of several protected areas in southern Africa such as Kruger National Park with an estimated 1 300–3 900 individuals (Bohm & Höner, 2015). However, populations in Central and West Africa are rapidly declining due to anthropogenic factors, with regional extinction likely in certain countries such as Algeria and Togo (Bohm & Höner, 2015).

Apart from southern Africa, East Africa is the only other stronghold for hyena populations (Bohm & Höner, 2015). Populations in the Serengeti ecosystem complex are estimated at 7 200–7 700 individuals in the Tanzanian region and 500–1 000 individuals in the Kenyan region (Bohm & Höner, 2015). Although hyena populations are stable in this part of Africa, various anthropogenic factors threaten the survival of this species (Mills &, Hofer 1998). The most significant threats being habitat loss (due to human encroachment and land conversion) and persecution (Bohm & Höner, 2015). Hyenas are often killed out of fear or due to the real or perceived threat to livestock, especially where hyenas live in close proximity to pastoralists (Mills & Hofer, 1998). Direct (e.g. shooting) and indirect (e.g. poisoning) persecution accounts for the highest percentage of mortalities outside of conservation areas (Henschel, 1986; Mills & Hofer, 1998) and incidental snaring and poisoning can contribute to adult mortality inside protected areas such as Serengeti National Park in Kenya, which affects local populations (Hofer, Campbell, East & Huish, 1996). The decline in their natural prey base (due to poaching) is less of a threat, although this compels hyenas to search for food in human developed areas often leading to conflict with humans (Ray et al., 2005).

1.3 Conservation status of apex carnivores in Malawi

During the early colonial period in Nyasaland (now Malawi), crop raiding animals and tsetse fly were the two major environmental issues throughout the country (Morris, 2006). The Department of Game, Fish and Tsetse Control was established in 1949 and a few officers were assigned to tsetse control and crop protection, which included shooting raiding animals such as elephants (Loxodonta africana) and yellow baboons (Papio cynoecephalus; Morris, 1996). At that time, large carnivores were not protected by any game law in Malawi and were allowed to be shot anywhere in the country, including within game reserves (Morris, 1996). Large carnivores often moved through agricultural areas, causing terror

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in many villages and control officers were often tasked with shooting and killing ‘problem animals’, more specifically man-eating lions (Morris, 2000). For example, during the 1940s, Mozambican lions frequently moved through the already well-populated Mangochi District and the scarcity of their natural prey led to occasional incidences of man-eating lions in the area (Morris, 2000). Several other areas including Chikwawa District had reports of man-eating lions and control officers would search for these individuals and kill them (Carr, 1969). Although leopards and hyenas are not typically seen as man-eaters, a single leopard reportedly killed 37 people in the Kasungu District and hyenas were also blamed for a number of deaths in the country (Morris, 2000). By 1945, Malawi had seen a noticeable loss of large mammalian species, including carnivores, which were once common in the area (Morris, 2006). However, due to the controlling forces of these carnivores on yellow baboons and bushpigs (Potamochoerus larvatus), two major crop raiders, officers from the crop protection unit recommended that only troublesome lions be shot and the others be left alone (Morris, 1996).

From 1945, George Dudley ‘G.D.’ Hayes, a pioneer conservationist in Malawi, inspired a game preservation approach rather than one of control (Morris, 2006). Due to the high levels of human- wildlife conflict he suggested that the preservation of Malawian wildlife was only possible by restricting animals to protected areas (Morris, 2000). Despite his efforts the governmental eradication policy accounted for the known deaths of 556 carnivores (lion and leopard) between 1948 and 1962 (Morris, 1996). Moreover, the increase in subsistence poaching often went undetected both inside and outside protected areas. A more stringent game law in 1953 meant that a license was required to shoot a maximum of two lions or two leopards per annum, and in 1971 a new Game Act was declared (Laws of Malawi Chapter 66:03) which stated that a special permit was required to shoot a lion or leopard (Morris, 1996). Despite this, large carnivores were still frequently killed as a further 115 carnivores (48 lions, 21 leopards and 46 hyenas) were killed between 1977 and 1982 (Morris, 1996).

By the end of the twentieth century, lion populations were largely restricted to protected areas (Nowell & Jackson, 1996). In 2002, the continental population survey predicted only 25 resident individuals in Malawi (Chardonnet, 2002). Lion populations were re-assessed in 2006 and estimated at fewer than 70 individuals (IUCN/SSC Cat Specialist Group, 2006). In 2008, 50 lions were estimated and in 2010 only 35 lions remained in Malawi (Mésochina et al. 2010), although the population size may have declined since then. In 2010, lions were present within five reserves: Kasungu National Park, Liwonde National Park, Nyika National Park, Vwaza National Park and Nkhotakota Wildlife Reserve. Today, lions are also absent from Nyika and Liwonde and only persist in Vwaza (n = 5), Kasungu (n = 6) and Nkhotakota (n = 18) (Riggio et al., 2013). Thus the conservation status of Malawi's lions is of great concern and extinction is likely if local populations are not supplemented and intensively managed.

Leopards remained relatively widespread throughout the country until the early 2000s (Morris, 2000). However much of their former range was reduced due to the expansion of agriculture and 13

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developmental pressures from the growing human population, which inevitably increased human- wildlife conflict (Myers, 1976). Since much of their prey base was depleted (Ray et al., 2005), leopards occasionally took livestock and were therefore killed due to the real or perceived threat to livestock (Myers, 1976). To this day, leopards suffer from direct (e.g. shooting) and indirect (e.g. accidental snaring) persecution whether within protected areas or non-gazetted areas. The largest population occurs in the natural forests and conifer plantations of Nyika National Park in the north of Malawi (Myers, 1976) and is likely the only genetically viable population left in the country. Despite their apparent widespread distribution, the lack of accurate historic and current estimates is worrying for the conservation of leopards in Malawi.

There are no accurate historic estimates of hyenas in Malawi and the only survey that was conducted in 1998 tentatively estimated 100–1 000 individuals in the country, though data appear deficient (Mills & Hofer, 1998). Hyenas are largely confined to protected areas (Ansell & Dowsett, 1988) and are believed to be present in all protected areas in Malawi, including Majete Wildlife Reserve (Mills & Hofer, 1998). However, hyenas still persist outside protected areas, living in close proximity to human- dominated landscapes and are even found in Lilongwe, the capital city of Malawi (Carnivore Research Malawi, 2015). Mills and Hofer (1998) stated that human population growth, habitat loss and direct and indirect persecution were the main reasons for the disappearance of this species from several parts of the country. In the southern region of Malawi, retaliatory killing of hyena resulting from livestock predation contributes a large proportion to annual hyena mortality, despite being protected by the wildlife protection act (Mills & Hofer, 1998; T. Moyo, personal communication, May 28, 2016). The future of hyenas in Malawi is largely uncertain and data on remaining hyena populations and their range are lacking.

In conclusion, the three apex predators of Malawi have suffered major habitat loss and population declines. The remaining populations are genetically isolated in only a few protected areas and populations are likely to go extinct in the future if they are not intensively managed. The reintroduction of apex predators into protected areas is probably the only chance of restoring apex predators in the country.

1.4 The importance of reintroduction

Reintroduction is defined as an attempt to re-establish species within their native range by release of wild or captive-bred individuals after the extirpation or extinction of wild populations (IUCN, 2013). Reintroduction is a valuable tool in wildlife conservation projects and aims at reducing the negative population trends of a species (Griffith, Scott, Carpenter & Reed, 1989). Efforts to reintroduce wildlife into reserves, especially charismatic species, have increased global awareness of the current mass species extinction caused by anthropogenic effects and the importance of conserving biological

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diversity (Seddon, Armstrong & Maloney, 2007). Reintroduction is a quick and proactive approach to repopulate animals in an area, whereas natural recolonization is slow and often restricted in countries such as Malawi where suitable protected areas are isolated by a human-dominated landscape matrix. The main reasons for translocating species to areas where they were previously removed include species conservation (Hayward et al., 2007a), replenishing hunted populations (Fischer & Lindenmayer, 2000), restoring ecosystem structure and function (Terborgh et al., 1999) and supporting ecotourism (Hayward et al., 2007b). Reintroduction is one of the only effective methods available for restoring species within their natural range and ultimately increasing viable populations of threatened species.

Top-order carnivores are of the most frequently reintroduced group of animals (Seddon, Soorae & Launay, 2005), however carnivore reintroductions are extremely difficult and many attempts fail (Breitenmoser, Breitenmoser-Wursten, Carbyn & Funk, 2001; Griffith et al., 1989; Hunter, 1998; Mills, 1991; van der Meulen, 1977). Problems met during carnivore reintroductions are summarized by Breitenmoser et al. (2001): (1) conflict with humans (e.g. livestock predation) was often the reason for the extirpation of a carnivore species and was not addressed prior to reintroduction, (2) large carnivores occur at low densities and thus require vast areas to maintain viable populations, (3) since carnivore species are naturally elusive animals, monitoring individual progress and assessing success may be challenging. However, since the late 1990s long-term monitoring has improved (Seddon et al., 2007) and numerous peer-reviewed articles are now available to guide researchers and managers towards successful carnivore reintroductions (e.g. Hayward et al., 2007b). For example, the success of carnivore reintroduction programmes depends on the habitat quality of the release site, location of release site in relation to former distribution, fencing and the number of individuals released (Griffith et al., 1989; Hayward et al., 2007b; Wolf, Griffith, Reed & Temple, 1996; Wolf, Garland & Griffith, 1998).

Defining the objectives of a large carnivore reintroduction is exceptionally important as it will determine whether a project succeeds or fails (Seddon et al., 2007). Initial planning and long-term monitoring are considered the most effective methods for improving reintroduction success and assessing reintroduction programme outcomes respectively (Hunter, 1998; Mills, 1991; Sarrazin & Barbault, 1996). The planning phase can be used to determine project specific objectives, while long- term monitoring can help to determine how individuals orient themselves with regards to conspecifics, location of resources and location of females to bear young (Hunter, 1998). The reintroduction of lions to KwaZulu-Natal, South Africa between 1992 and 1999 (Hunter, 1998), is considered one of the most successful large carnivore reintroductions due to careful planning and post-release monitoring (Breitenmoser et al., 2001) and serves as an example for future reintroduction projects. The success of top predator reintroductions is typically defined as a population size with a threshold value of >500

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individuals (Griffith et al., 1989), although very few systems (only mega reserves or Transfrontier Parks) are large enough to sustain such a large population of predators (Hayward & Somers, 2009). As a result the success in smaller reserves should be defined as the first wild-born generation, or a three-year breeding population with a natural birth rate exceeding mortality (Griffith et al., 1989). Therefore, to ensure population growth and stability (i.e. self-sustaining populations, post-reintroduction), a combination of long term monitoring (Breitenmoser et al., 2001; Hunter, 1998; Mills, 1991; Weber & Rabinowitz, 1996) and on-going management is needed (Seddon et al., 2007).

1.5 Aim and objectives

1.5.1 Aim

The aim of this study was to understand the ecology of reintroduced lion and leopard as well as resident hyena in Majete Wildlife Reserve and to propose recommendations for the future management of all three apex predators.

1.5.2 Objectives

The main objectives for this study were to:

1) Assess early post-release movement patterns of lion and leopard in Majete Wildlife Reserve using GPS collar data

a) What were the distances moved by lion and leopard post-release?

b) How long did it take reintroduced felids to establish their ranges?

c) Did lion and leopard exhibit homing tendencies when reintroduced, over a long distance, into a small, enclosed environment?

d) Did lions and leopards display signs of breeding success post-reintroduction?

2) Determine home range, range overlap and habitat use of reintroduced lion and leopard in Majete Wildlife Reserve using GPS collar data

a) What were the distances between release site and home range centroid for each reintroduced felid?

b) What were the home range sizes of lions and leopards and did these animals display range overlap?

c) Did lions and leopards prefer or avoid any habitat types?

d) Was there seasonal variation in home range size?

3) Compare the diet of lion, leopard and hyena in Majete Wildlife Reserve by means of scat and GPS cluster visitation analyses 16

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a) What was the diet of each predator and did they prefer any prey species?

b) To what degree did predator diet overlap?

c) Did these predators reduce direct competition by selecting different prey?

4) Determine the population and density estimates of leopard and hyena using camera trapping, within Majete Wildlife Reserve

a) What was the population size of leopard five years after their reintroduction?

b) What was the resident hyena population size and density?

5) Determine the number of predators Majete can sustain a) What was the carrying capacities of lions, leopards and hyenas?

1.6 Thesis outline

This thesis is composed of seven chapters. Chapters One and Two provide background to the study species and study site. Chapters Three, Four, Five and Six are written as stand-alone manuscripts to assist publication in peer-reviewed journals. There is therefore some repetition between chapters. Chapter Seven is prepared for African Parks Majete (Pty) Ltd. to present major research findings and provide recommendations for the management of lion, leopard and hyena in Majete Wildlife Reserve.

Chapter 1 – Introduction: This chapter provides an introduction to the role of apex carnivores in ecosystems, their global decline, conservation status and importance of their reintroduction.

Chapter 2 – Study site: This chapter describes the study site, its history and provides technical and biological details of the felid reintroduction.

Chapter 3 – Early post-release monitoring and post-release breeding of reintroduced felids: This chapter investigates the distance and direction of movement, range establishment, as well as post-release breeding of felids to determine the success of lion and leopard reintroductions.

Chapter 4 – Home range and habitat selection of two reintroduced felids: This chapter evaluates home range, habitat selection and roaming behaviour of reintroduced felids.

Chapter 5 – Dietary ecology of three apex predators: This chapter describes the diet of lion, leopard and hyena using a combination of scat collection (for all three predators) and GPS cluster site analysis (for lion only). Preferred prey, dietary overlap and prey weight range are compared among these three carnivores.

Chapter 6 – Population dynamics and carrying capacity of lion, leopard and hyena: This chapter reports on estimated population sizes and densities of leopard and hyena using spatially explicit capture- recapture analyses. Estimated carrying capacities are also provided for lion, leopard and hyena.

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Chapter 7 – Research findings and management recommendations: The thesis concludes with the major findings and provides recommendations for the future management of lion, leopard and hyena in Majete Wildlife Reserve and the rest of Malawi.

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Chapter 2

Study area and reintroduction of study species

2.1 Study area

2.1.1 Location and history

Majete Wildlife Reserve (MWR) is a 700 km² reserve situated in the Lower Shire Valley, forming the southern part of the Great Rift Valley in the south of Malawi. The history of MWR started in 1951 when the area around Majete Hill, a prominent hill in the western region of the reserve, was declared a non- hunting area due to human encroachment and declines in large mammalian populations such as elephant (Loxodonta africana) and buffalo (Syncerus caffer). In 1955, MWR was declared a game reserve covering 500 km², with the objective of restricting elephants to the reserve (Morris, 2006). The northward and eastward extension of MWR in 1969 allowed the reserve to include the Mkulumadzi River and an additional kilometre east of the Shire River along the eastern border (Morris, 2006; Sherry, 1989), which would provide wildlife with water during the dry season. According to Bell (1984), by the early 1980s MWR had “respectable” numbers of large mammalian species including sable (Hippotragus niger), kudu (Tragelaphus strepsiceros), waterbuck (Kobus ellipsiprymnus) and elephants. However, between the late 1980s and 1990s, large-scale poaching led to the depletion and extirpation of many mammalian species. Game scouts and law enforcement officials were unable to control poaching within the reserve, largely due to insufficient finances, inadequate resources and poor management (Morris, 2006).

The future course of MWR changed dramatically on 28th March 2003, as African Parks Ltd. (AP) and the Malawian Department of National Parks and Wildlife (DNPW) signed a 25-year public-private partnership to restore, manage and develop MWR. Funding was allocated to developing infrastructure, equipping and training law enforcement, improving tourism and establishing community relations. The construction of the sanctuary fence (140 km²) in the north-east of the reserve marked the beginning of MWR’s recovery. The next phase was the construction of artificial water points within the reserve to provide drinking access for animals and to aid tourism in the form of improved game viewing. The final phase was the reintroduction of more than 2 550 animals representing 12 species which took place between 2003 and 2009. Animals were sourced from within Malawi (Lengwe National Park and Liwonde National Park) as well as from South Africa and Zambia. Herbivore populations were given time to establish and increase, before the reintroduction of predators. By 2011 the (predator-proof) perimeter fence was completed and the sanctuary fence was removed, allowing animals access to the

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rest of the 700 km2 reserve. Between 2011 and 2013, three lions (Panthera leo) and six leopards (Panthera pardus) were reintroduced into the reserve.

2.1.2 Climate

MWR is situated in a tropical climate zone with semi-arid conditions (Staub, Binford & Stevens, 2013). Hall-Martin (1972) described the climate for Lengwe National Park, which lies about 5 km south of MWR and thus relevant to the area:

 Hot wet season (December – March)

 Cool dry season (April – August)

 Hot dry season (September – November)

June and July are the coldest months with a mean temperature of 19.3°C and November is the hottest month with a mean temperature 26.8°C (Wienand, 2013). The minimum and maximum temperatures are 11°C and 45°C respectively (Sherry, 1989). Average annual precipitation is 680–800 mm in the east and 700–1 000 mm in the west (Hall-Martin, 1972; Wienand, 2013). The wet season contributes a significant proportion of the annual rainfall, which occurs most days in January. During the cool dry season, south-easterly winds blow moist air from the Mozambican coastline over the highlands of the Great Rift, which condenses over Mount Chiperone, forming clouds and occasional drizzle over the Lower Shire Valley (Morris, 2006; Sherry, 1989).

2.1.3 Topography and altitude

The topography of MWR is relatively flat with few rocky outcrops in the east and undulating with many rocky outcrops and hills in the west (Wienand, 2013). The low lying east and upland west are divided by a distinct N-S line that runs through the middle of the reserve. Altitude decreases towards the east and is lowest at the Kapichira Falls along the Shire River at 100 m (Bell, 1984). The highest and most prominent peak is the conical Majete Hill (766 m), which occurs in the west of the reserve.

2.1.4 Geology and soils

According to the Geological Survey Bulletins, the majority of the reserve has a Precambrian Basement Complex Horneblende-Biotite Gneiss, with long bands of quartz-schists and granulites in the escarpment and short bands of biotite-gneiss and psammite-gneiss across the reserve (Bell, 1984). Relatively recent layers of alluvial deposits are found above the basement complex and dolerite formations (dykes or sills) are found throughout, although mostly in the south-east, the most well- known being the sill that formed the Kapichira Falls (Bell, 1984; Sherry, 1989). The soils are lithosols, shallow, stony, ferruginous and of poor nutrient status, with more fertile alluvial soil occurring only

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along a few river beds (Sherry, 1989). The soils are mostly shallow and thus un-suitable for agriculture (Bell, 1984; Sherry, 1989).

2.1.5 Watercourses

The Shire River is a major perennial river which drains Lake Malawi, flowing southwards through MWR and into the Zambezi River. This river is an important source of water in the Lower Shire Valley and forms a section of the eastern boundary (12 km) of the reserve, neighbouring with settlements and a hydroelectric power station (completed in 1996) near the Kapichira Falls. The Mkulumadzi River is the other perennial river in MWR and has its confluence with the Shire River in the north east of the reserve. The other rivers (which find their source inside or outside of the reserve) are non-perennial and only flow in the form of storm drains in the wet season. MWR has eleven natural springs, both perennial and seasonal and numerous seasonal pans or ‘dambos’ distributed throughout the reserve which fill up in wet season (Wienand, 2013). Additionally, ten borehole-fed artificial waterholes were constructed in several key areas of the reserve to supplement animals with drinking water particularly in the dry season, and to aid tourism, which is an important source of revenue for MWR.

2.1.6 Vegetation

Sherry (1989) described the vegetation as tropical dry woodland/miombo savanna woodland and divided MWR’s vegetation into six types (Figure 2.1). Vegetation types and associated habitat features are listed below:

(i) Riverine associations

This vegetation type is only found along river systems and includes riverine and alluvial associations below 230 m. This type is dominated by tree species such as Acacia tortilis, Kigelia africana and Lonchocarpus capassa. Common shrubs are Allophylus spp., Cardiogyne africana and Grewia spp. and dominant grasses are Cynodon, Digitaria and Phragmites.

(ii) Low-altitude (205-280 m) mixed tall deciduous woodland

This category is the second largest (40.93%) in MWR. The tree layer is dense with a higher average canopy height compared to the riparian zone, although this type is still relatively open woodland. Major tree species include Acacia nigrscens, Combretum imberbe and Sclerocarya caffra. Perennial grasses such as Heteropogon occur in this type and the shrub layer is sparser than in the riparian zone.

(iii) Ridge-top (220-300 m) mixed short deciduous woodland

This vegetation type supports the lowest overall biomass of all communities in the reserve and is confined to the ridge-tops and upper ground along the tributaries of the Shire River. This short, medium density woodland is dominated by Diospyros kirkii, Terminalia sericea and Diplorhynchus condylocarpon. 36

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(iv) Medium-altitude (230–410 m) mixed tall deciduous woodland

This eco-tonal association covers a strip in the central region of MWR that splits the low altitude eastern vegetation from the higher altitude western vegetation. As a result this transitional zone includes vegetation from the eastern low-lying matrix (iii) and the western high lying matrix (v) such as Brachystegia boehmii, D. kirkii and Combretum spp. Therefore, this transitional zone represents the highest tree biomass in the reserve.

(v) High-altitude (410–770 m) tall miombo woodland

This type is the largest (30.34%) in the reserve and is located in the higher lying west of the reserve. It supports a high biomass of low browse quality and the dominant trees are Brachystegia boehmii and Julbernardia globiflora, with Burkea africana, D. condactylcarpon, and Pterocarpus angolensis also present. The shrub layer’s key representatives are Acacia torrei, Acacia erubescens, Bridelia cathartica, Bauhinia petersiana and Ormocarpum kirkii. Grasses are mainly represented by Andropogon, Diheteropogon, Heteropogon and Hyparrhenia.

(vi) Riparian thicket

This vegetation type represents 7.45% of the reserve and is largely confined to the rivers and river junctions, particularly the flat areas (below 240 m), in the eastern part of the reserve. The major tree species are Adansonia digitata, Albizia anthelmintica and Euphorbia ingens. The shrub component includes Bauhinia tomentosa, Diospyros senensis, Ehretia spp. and Grewia bicolor, while grasses are largely represented by Brachiara spp. and Leptochloa spp.

Another survey was conducted in 2015 and classified four vegetation types (see Figure 2.2): woodland above 400 m (Brachystegia boehmii and Julbernardia globiflora), woodland between 250 and 400 m (Brachystegia boehmii, Pterocarpus rotundifolius and Combretum spp.), woodland below 250 m (Acacia spp. and Steculia) and savanna (Combretum spp., Acacia spp. and Panicum spp.) (African Parks Majete (Pty) Ltd., personal communication, January 15, 2017). However, this new classification still needs to be ground-truthed (C. Hay, personal communication, November 5, 2016) and so the previously mentioned vegetation types from Sherry (1989) were used in this study.

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Figure 2.2. Proposed map of habitat types in Majete Wildlife Reserve, Malawi (African Parks Majete (Pty) Ltd., personal communication, January 15, 2017). 38

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2.2 Large carnivore reintroduction into MWR

Large carnivores were once common in the Lower Shire Valley (Hayes, 1979), but they were already rare by the early 1960s (Mésochina et al., 2010). This was largely due to poaching and problem animal control, which decimated lion and leopard populations (Morris, 1996). In MWR, game patrols observed one lion per 100 patrol days between 1959 and 1962 (Morris, 2006). According to Bell (1984), lions were occasionally sighted by head game scout Katema in MWR between 1972 and 1976, although no signs of resident lion were recorded after 1976. Leopards and spotted hyenas (Crocuta crocuta) were considered relatively common in MWR until the 1980s, as their presence was indicated by frequent tracks, droppings and vocalisations (Bell, 1984). However, the last known sighting of leopard within MWR was by game scout Moses in 1992, close to the Shire River (N. Moses, personal communication, November 12, 2016; see Figure 2.1), and by the beginning of the twenty-first century, leopards were believed to be extirpated (Dowsett & Dowsett-Lemaire, 2005). Spotted hyenas were always considered common in MWR (Bell, 1984; Dowsett & Dowsett-Lemaire, 2005), but today only a small population persists inside the reserve, due to pressure from surrounding villages (T. Moyo, personal communication, May 28, 2016).

Between October 2011 and November 2012, four lions and six leopards were translocated from South Africa to MWR (Table 2.1). After the capture in South Africa, all four lions were held together in a quarantine enclosure to ensure social bonds between individuals prior to their reintroduction into MWR. The relatedness of the two male lions is unknown, although both females were unrelated. One female lion died en route due to hypertension and therefore only three lions (two males and one female) were reintroduced. All leopards were unrelated. The animals were photographed to identify individuals based on spot and rosette markings (leopards) and whisker spots (lions). Each release was a separate event, which were between three and six months apart (Table 2.1). According to Hunter (1998), sufficient delays in successive release events may allow individuals to establish home ranges and subsequent releases of new individuals might minimise potential encounters with established animals for the first few weeks or months. The releases took place at four sites within the north- eastern section of MWR. This predator reintroduction marked the end of a large wildlife reintroduction programme (2003–2011) into MWR.

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Table 2.1. Biological and technical information of reintroduced lion and leopard in Majete Wildlife Reserve, Malawi.

Release Release Group Estimated Translocation Boma Species Weight Origin Comment Event Date Composition age (years) distance (km) (days)

One male 2–3 - Kruger National Park 1 041 7 Both animals in conflict with 1 10/2011 Leopard human activity One female 2 - Ohrigstad 1 078 15

One male 3–4 - Ohrigstad 1 084 23 - 2 01/2012 Leopard Thaba Ingwe Nature One female 2–3 - 1 235 24 Reserve

Two males 3 & 3.5 - Pilanesberg National Park 1 346 26 Males from same coalition 3 07/2012 Lion One female 3 - Madikwe Game Reserve 1 324 26 -

One male 7 62 kg Loskop Dam 1 204 16 Livestock raider; fence breaker 4 11/2012 Leopard Malelane, southern Kruger One female 5 40 kg 1 117 15 Problem animal National Park

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2.2.1 Pre-release management

During the reintroduction period individuals were held in specially designed acclimation bomas (or enclosures) in the reserve. This is known as a ‘soft-release’ and is highly recommended for large carnivores (Caughley & Gunn, 1996; Hayward et al., 2007), especially if they are to adapt to collars, conspecifics within a social group and an entirely new environment (Hayward et al., 2007). Individuals were released within four weeks, although Hayward et al. (2007) stated that leopards may require more time in the enclosure to reduce aggression levels and become less secretive. The lion boma measured 50 m x 50 m, with a 5 m x 7 m double gated area and was constructed of 2.4 m high diamond mesh and electric fencing, which discharged approximately 7 000 volts (see Appendix 2A). Vegetation inside the boma provided lions with sufficient shelter. The leopard boma was divided into two sections, each measured 6m x 6m x 4m and was constructed of diamond mesh fencing (see Appendix 2B). To prevent animals from escaping, the fence was buried up to a depth of 1 m. Each boma had two smaller crushes measuring 2 m x 2 m x 1 m (see Appendix 2C), which is where animal carcasses were placed every three to four days. Only the carcasses of wild prey species were given to lions and leopards to reduce the likelihood of ‘imprinting’ on livestock (Hunter, 1998). Management monitored the level of aggression and habituation to vehicles and individuals were released once management believed they were relatively well acclimatised. Upon their release into the reserve, individuals were provided with a carcass, for the final time, outside the enclosure.

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2.4 References

Bell, R.H.V. (1984). Majete Game Reserve: report of an ulendo and suggestions for management and utilisation (Malawi Government Report WRU/1//50/5/3//1). Kasungu, Malawi: Department of National Parks and Wildlife.

Caughley, G. & Gunn, A. (1996). Conservation biology in theory and practice. Victoria, Australia: Blackwell Science.

Dowsett, R.J. & Dowsett-Lemaire, F. (2005). The mammals of the Lower Shire Valley wildlife reserves (Lengwe, Majete and Mwabvi), Malawi. Nyala, 23, 3–14.

Hall-Martin, A. (1972). Classification and ordination of forest and thicket vegetation of the Lengwe National Park, Malawi. Kirkia, 10, 131–144.

Hayes, G.D. (1979). Lions – man-eaters and other. Nyala, 5, 6–11.

Hayward, M.W., Adendorff, J., O’Brien, J., Sholto-Douglas, A., Bissett, C., Moolman, L.C., Bean, P., Fogarty, A., Howarth, D., Slater, R. & Kerley, G.I.H. (2007). Practical considerations for the reintroduction of large, terrestrial, mammalian predators based on reintroductions to South Africa’s Eastern Cape Province. The Open Conservation Biology Journal, 1, 1–11. DOI: 10.2174/1874–8392/07

Hunter, L.T.B. (1998). The behavioural ecology of reintroduced lions and cheetahs in the Phinda Resource Reserve, KwaZulu-Natal, South Africa. (Unpublished Ph.D. dissertation). Pretoria, South Africa: University of Pretoria.

Morris, B. (1996). A short history of wildlife conservation in Malawi (Occasional Papers No. 64). Edinburgh, U.K.: Centre of African Studies, University of Edinburgh.

Morris, B. (2006). The history and conservation of mammals in Malawi (Monograph No. 21). Zomba, Malawi: Kachere Series.

Sherry, B.Y. (1989). Aspects of the ecology of the elephant Loxodonta africana (Blumenbach, 1797) in the Middle Shire Valley, southern Malawi. (Unpublished M.Sc. thesis). Zomba, Malawi: University of Malawi.

Staub, C.G., Binford, M.W. & Stevens, F.R. (2013). Elephant herbivory in Majete Wildlife Reserve, Malawi. African Journal of Ecology, 51(4), 536–543. DOI: 10.1111/aje.12064

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Wienand, J.J. (2013). Woody vegetation change and elephant water point use in Majete Wildlife Reserve: implications for water management strategies. (Unpublished M.Sc. thesis). Stellenbosch, South Africa: Stellenbosch University.

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2.5 Appendices

Appendix 2A. Lion boma located in the north-eastern section of MWR.

Appendix 2B. Leopard boma situated in the north-eastern section of MWR.

Appendix 2C. Crush boxes were designed as an access point for providing carcasses to leopards in the bomas.

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Chapter 3

Early post-release monitoring and post-release breeding of two reintroduced felids in Majete Wildlife Reserve, Malawi

W.D. Briers-Louw1, A.J. Leslie1

1Department of Conservation Ecology and Entomology, Stellenbosch University, Matieland, Western Cape, 7602, South Africa

3.1 Abstract

Large carnivores are frequently reintroduced into reserves to compensate for anthropogenic-driven losses. The lack of post-release monitoring has impeded our knowledge on how these carnivores adapt to their new environment, which often results in uncertainty of whether reintroductions were successful or not. Between 2011 and 2012, three lions (Panthera leo) and six leopards (Panthera pardus) were reintroduced into Majete Wildlife Reserve and each animal was fitted with a GPS collar to monitor their early post-release movements after the first three months. We considered the reintroduction successful if felids exhibited: (1) reduced daily distances; (2) stabilised distances from the release site; (3) reduced range expansion; and (4) no homing behaviour (i.e. no movement towards the capture location). Felids generally showed initial movements away from the release site and increase daily distance were recorded for lions (mean = 2.04 km) and leopards (mean = 0.36 km) between month one and two. For lions, daily distance moved decreased significantly (Tukey HSD test, p = 0.004), while distance travelled from release site stabilised from 5.09 ± 2.76 [SD] km to 5.08 ± 1.37 km, between the second and third month. Of the six leopards, four showed stabilised daily movement rates (3.62 ± 0.19 km to 3.58 ± 0.25 km) between month two and three, LEF3 displayed significantly reduced movements (F1,37 = 7.20, p = 0.01), while LEM3 showed increased movements from 3.25 ± 3.11 km (month two) to 4.12 ± 2.12 km (month three). Distance travelled from the release site also stabilised for leopards from 6.92 ± 5.65 km (month two) to 6.63 ± 5.43 km (month three). Lion range size decreased significantly by the third month (Tukey HSD test, p = 0.02), while leopard home ranges showed a general reduction, although this was not significant (F2,13 = 0.21, p = 0.81). Released cats exhibited no consistent movement towards their capture location and post-release breeding provided encouraging signs for population growth six years post-reintroduction. Based on these results the felid reintroduction was successful. This study highlights the importance of post-release monitoring to determine the success or failure of carnivore reintroductions. We encourage collaboration between researchers and managers to apply scientific knowledge to aid reserve management.

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3.2 Introduction

Large carnivores are the most frequently reintroduced group of animals (Seddon, Soorae & Launay, 2005). This is largely due to their extirpation on a local and global scale (Weber & Rabinowitz, 1996; Woodroffe, Frank, Lindsey, Ranah & Romanach, 2007) as well as the important role that carnivores play in conservation and eco-tourism (Hayward, Hayward, Druce & Kerley, 2009). The main aim of a reintroduction is to re-establish species within their historic range from which they were extirpated (IUCN/SSC, 2013). Reintroduction is also an important conservation tool which aims to promote ecosystem restoration and even the recovery of endangered carnivore species to prevent potential extinction (Kleiman, 1989; Ripple & Beschta, 2012; Sarrazin & Barbault, 1996).

The reintroduction of large carnivores can be divided into two phases: pre- and post-release (Somers & Gusset, 2009). The pre-release stage refers to the period prior to the release of individuals into a new environment. It is well-known that releasing large carnivores directly into the release area (hard- release) has little success, whereas keeping individuals in enclosures (or bomas) for a short period before release (soft-release) improves success (Hunter, 1998). The period in a boma allows individuals to acclimatise to their new environment and break homing tendencies, as large carnivores often return to the capture site (Linnell, Aanes, Swenson, Odden & Smith, 1997; Massei, Quy, Gurney & Cowan, 2010; Miller, Ralls, Reading, Scott & Estes, 1999; Somers & Gusset, 2009). Other factors to consider are the number of individuals released and the degree of sociality of the reintroduced species (Breitenmoser, Breitenmoser-Wursten, Carbyn & Funk, 2001). For example, highly social felids such as lions (Panthera leo), are likely to form enduring social relationships when kept together in a holding facility prior to their release (Hunter et al., 2007). The establishment of these cohesive social bonds may improve breeding success and reduce extensive post-release ranging behaviour (Somers & Gusset, 2009), ultimately increasing the likelihood of reintroduction success.

The post-release phase involves monitoring the response of carnivores after their release, by assessing movement patterns of individuals from their release site as well as behaviour in their new environment (Hunter, 1998). Essentially, this monitoring phase aims to determine whether a reintroduction was successful or not (Hunter et al., 2007). In the past, large carnivore reintroductions rarely incorporated post-release monitoring, which was largely due to a lack of funding, difficulty in monitoring elusive animals and disparity between goals of managers and researchers (Breitenmoser et al., 2001; Hayward et al., 2007a, 2007b). More recently research efforts have increased and peer-reviewed publications have developed guidelines to bridge these gaps (e.g. Hayward et al., 2007a). Irrespective of these current advances, only a few studies have documented the early post-release movements of reintroduced felids in Africa (Hunter, 1998; Killian, 2003; Yiu, Keith, Karezmarski & Parrini, 2015).

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Large felids are particularly difficult to reintroduce and there are a number of reasons that may cause failed establishment of populations after release (Hayward et al., 2007b; Hunter, 1998; Mills, 1991; van der Meulen, 1977) For instance, these animals may display high post-release dispersal as they recover from translocation stress and acclimatise to the new environment, often with a tendency to return to the capture site (Somers & Gusset, 2009; Weilenmann, Gusset, Mills, Gabanapelo & Schiess- Meier, 2010). Large felids are extremely territorial (Bailey, 1993; Bothma & Walker, 1999; Hunter, 1998; Schaller, 1972) and the presence of resident conspecifics may significantly alter the movements of newly translocated individuals, especially if the population is at carrying capacity (Griffith, Scott, Carpenter & Reed, 1989; Hunter, 1998). Consequently, these individuals often roam outside of protected areas to establish a suitable home range, which may result in deleterious anthropogenic edge effects (Somers & Gusset, 2009; Weilenman et al., 2010; Woodroffe & Ginsberg, 1998). Several studies have documented the release of livestock raiding felids from conflict areas into protected areas (Hamilton, 1981; van der Meulen, 1977), and these ‘problem animals’ often resume livestock killing outside of protected areas, increasing the likelihood of being killed by landowners (e.g. Weilenmann et al., 2010).

Despite the challenges of large felid reintroductions, a number of studies have documented successful reintroductions. For example, lions are the most successfully reintroduced African felid with reintroduction efforts starting in 1965 in the Umfolozi Game Reserve (now Hluhluwe-iMfolozi Park), South Africa (Anderson, 1981). Between 1992 and 2006, lions were reintroduced into more than 27 reserves in South Africa and most reintroductions were considered successful (Funston, 2008). In contrast, leopard (Panthera pardus) reintroductions are few in number and success rates are generally low. For example, of the 13 leopards reintroduced into five small reserves in the Eastern Cape Province, South Africa since 2001, only seven individuals were accounted for in 2005 (Hayward et al., 2007b). The reintroduction of leopards appears to be more challenging than lions and may be attributed to extensive post-release ranging behaviour or a lack of post-release monitoring. Consequently, little is known about leopard reintroductions.

The objective of this study was to determine whether the reintroduction of lions and leopards in Majete Wildlife Reserve, located in Malawi, was successful. Based on previous studies (Hunter et al., 2007; Yiu et al., 2015), we classified early-post release success according to the following criteria: (1) reduction in felid movement rates; (2) range establishment i.e. reduction in range expansion; and (3) no tendencies for felids to return to the capture site i.e. homing. We also monitored post-release breeding to determine whether felids successfully raised offspring to the point of independence and/or sexual maturity. This was the first reintroduction of large predators in Malawi and this study aimed to provide guidelines to aid future management and reintroduction of large felids.

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3.3 Methods

3.3.1 Study site

Majete Wildlife Reserve (MWR) covers 700 km² in the Lower Shire Valley in the southern region of Malawi. MWR is characterised by two distinct seasons: a wet season from December to May and a dry season from June to November. The average annual precipitation is between 680–800 mm in the eastern lowlands and 700–1 000 mm in the western uplands (Hall-Martin, 1972). Mean daily temperatures range from 23.3°C in the cool winter months to 28.4°C in the hot summer months (Wienand, 2013). The Shire and Mkulumadzi Rivers are the only perennial rivers in MWR, while non- perennial rivers flow on account of flash floods in the wet season (Figure 3.1). Several natural springs and ten artificial water points are located within the reserve, which serve as an important source of water for animals during the dry season. The eastern region is relatively low and flat, with a mixed woodland below 250 m, dominated by Acacia spp. and Steculia spp. Vegetation transitions into mixed woodland between 250 m and 400 m towards the centre of the reserve and is dominated by Brachystegia boehmii, Pterocarpus rotundifolius and Combretum spp. Altitudinal gradient further increases towards the rugged, hilly western region and is covered by tall Miombo woodland above 400 m dominated by Brachystegia boehmii and Julbernardia globiflora (Staub, Binford & Stevens, 2013).

Historical context

Apex predators were once common throughout the Lower Shire Valley, including MWR (Bell, 1984; Hayes, 1979; Morris, 2006). However, agriculture and economic development resulted in large-scale habitat loss (Sherry, 1989) and human-predator conflict and as a result, predators were largely restricted to reserves. Bell (1984) stated that lion, leopard and spotted hyena (Crocuta crocuta) were regularly sighted in MWR, until the beginning of the 1970s. The effects of direct persecution and poaching of suitable prey, coupled with the regular removal by Problem Animal Control (PAC), were noticeable, as sightings of these predators became rare (Morris, 2006). By 1976, lions were absent from MWR (Bell, 1984) and leopard tracks and vocalisations were occasionally reported, but by the early 1990s they had also disappeared from the reserve (N. Moses, personal communication, November 12, 2016). Spotted hyenas were the only large predator remaining in MWR, due to pressure from surrounding villages (T. Moyo, personal communication, May 28, 2016).

In 2003, a public private partnership (PPP) was signed between African Parks (AP) and the Malawi Department of National Parks and Wildlife (DNPW) to restore and develop MWR. A road network was developed to improve accessibility in the reserve and law enforcement was trained and equipped to deal with poaching. A 140 km2 sanctuary fence was constructed in the north-eastern section of the reserve and 2550 animals from 12 different herbivore species were reintroduced between 2003 and 2011. The perimeter fence line (142 km) was completed in 2011 and the sanctuary fence was removed. 48

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By this time, herbivore populations were large enough to support large predators and thus between 2011 and 2012, three lions and six leopards were reintroduced into MWR.

3.3.2 Methods and materials

Collaring

Two lions (one male and one female) and six leopards (three males and three females) were fitted with global positioning system (GPS) satellite collars (n = 8, African Wildlife Tracking, Pretoria, South Africa) following their reintroduction. The third lion (a male) was fitted with a very high frequency (VHF) tracking collar (n = 1, African Wildlife Tracking, Pretoria, South Africa), although due to several challenges, data were insufficient for analyses (see Table 3.1). All individuals were immobilised by veterinarians using standard drug combinations (Kreeger, 1996). Collars weighed approximately 650 g for leopard (1.5–2 % of leopard body mass) and 900 g for lion (less than 1 % of lion body mass). All collars were made from thick leather belting and fitted by pop-rivets. Frequencies of collars were in the 148–149 MHz wavelength. They were programmed to record a GPS location every four to five hours (to conserve battery life) and location data were downloaded from the African Wildlife Tracking website (www.awt.co.za).

Figure 3.1. Boma and release sites of reintroduced lions and leopards in Majete Wildlife Reserve, Malawi (Shapefiles, personal communication, African Parks (Pty) Ltd.). 49

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Boma

During the pre-release phase, all reintroduced individuals were kept in acclimation bomas (leopards adjacent and lions together) for between one and four weeks, prior to their release (Table 3.1, Figure 3.1). This was done to reduce aggression levels of individuals, habituate animals to vehicles and familiarise individuals to conspecifics within a social group (Hayward et al., 2007a). This is termed a “soft-release” and has been recommended to improve the success of a reintroduction event (Fischer & Lindenmeyer, 2000). Individuals were provided with sufficient water and fresh carcasses from wild animals in the reserve, every three to four days. The first set of reintroduced leopards (LEM1 and LEF1) were released at the boma site, while the last female (LEF3) escaped from the boma prior to her released. All lions (LIM1, LIM2 and LIF1) and the second set of leopards (LEM2 and LEF2) were released along the Mkulumadzi River in the north of the reserve. Finally, the last male leopard (LEM3) was released along the Shire River (Figure 3.1).

Table 3.1. Biological and practical details of reintroduced lions and leopards in Majete Wildlife Reserve, Malawi.

Estimated Period in Collar Release Monitoring ID Code Sex age boma Comment type date period (days) (months) (days) Lion LIM1 M 36 26 GPS 07/2012 90 - LIM2 M 42 26 VHF 07/2012 - No data for analyses LIF1 F 36 26 GPS 07/2012 90 - Leopard LEM1 M 24–36 7 GPS 10/2011 90 - LEM2 M 36–48 24 GPS 01/2012 90 - Habitual livestock raider LEM3 M 84 16 GPS 11/2012 90 and fence breaker LEF1 F 24 15 GPS 10/2011 90 - LEF2 F 24–36 23 GPS 01/2012 90 - Considered a problem LEF3 F 60 15 GPS 11/2012 63 animal; Collar failed at the start of month three

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3.3.3 Data analysis

Early post-release monitoring

The first three months of the post-release phase were selected as the sampling period for early post- release monitoring. Methods for early post-release movements were based on Hunter (1998) who monitored reintroduced lion and cheetah in Phinda Game Reserve, South Africa. Animals remained close to the release site for a day or two as they recovered from the anaesthesia (personal observation). Monitoring started when the individual moved more than one kilometre from the release site. Animal movement (distance from release site, daily distance and direction of travel) and range establishment were determined using Home Range Tools (HRT) for ArcGIS (Rodgers, Kie, Wright, Beyer & Carr, 2015) and ArcToolbox in ArcGIS 10.5 (Environmental Systems Research Institute (ESRI), Redlands, California, U.S.A.).

Daily distance and distance from release site were determined using a straight-line method (Hunter, 1998). This method represents the minimum distance moved by felids between locations. A one-way ANOVA with Tukey post hoc test was used to compare daily distance travelled during the three-month monitoring period. The Kruskal-Wallis test by ranks was used to test for monthly differences in distance travelled from the release site. Statstica version 13.2 (Dell Software, 2016) was used to conduct all statistical analyses.

The direction of daily movement from the release site was determined to test whether reintroduced lions and leopards exhibited strong tendencies to return to the capture site i.e. home. Capture locations were documented for each released individual and thus measurements of homing tendencies were compared to their direction home. All directional movements were analysed, but only the first location was included when individuals were stationary at a site, due to feeding or mating for several hours or days.

Three methods were compared to determine homing behaviour. Method one was based on Hunter (1998), whereby mean angles of movement were calculated from the direction of each movement from the release site. Method two and three were based on Kilian (2003) and Fies, Martin and Blank (1987) respectively. Both methods computed mean angles from consecutive movements, rather than from the release site. The Rayleigh test (one-sample test for mean angles) was used to test for even distribution around a circle (Zar, 1984). This test provided the mean angle with a 95% confidence interval and determines whether the direction of the capture location falls within the interval (Zar, 1984), which would indicate whether reintroduced lions and leopards moved consistently in the direction of the capture site (Hunter, 1998; Kilian, 2003; Yiu et al., 2015). Instead of generating a 95% confidence interval for the mean angle of movement, method three (Fies et al., 1987) included 22.5° on either side of the actual capture location and mean angles falling within this range were considered

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homing. All circular statistics were analysed in Oriana 4.0 (Kovach Computing Services, Anglesey, Wales).

To determine whether lions and leopards established permanent home ranges within the reserve, we computed range sizes for each month. Home ranges and core areas were estimated using minimum convex polygon (MCP), which connects the outermost location points and kernel utilisation distribution (UD) which estimates a kernel (or probability density) over each location points and removes outlying location points (Harris et al., 1990; Rodgers et al., 2015). Released cats were defined to have settled down and established home ranges when they displayed a reduction in range size and began to occupy a particular area. The one-way ANOVA test with Tukey HSD post hoc test was used to test for differences in monthly range sizes and a t-test was used to conduct comparisons between overall range sizes within the monitoring period.

Post-release breeding

Post-release breeding of lions and leopards was monitored from the initial reintroduction between 2011 and 2012 to September 2017. Information was provided by the park manager and operations manager. Experienced guides and scouts provided regular information of lion sightings, which included number of individuals, cub sightings and individuals involved in mating events). Information was also provided for leopards: age, sex and number of individuals, presence of cubs or subadults and whether individuals were collared or not. However, leopard sightings were relatively rare. Photographs from guides and tourists were also helpful in determining the leopard individual sighted. The presence of ‘new individuals’ was aided with camera trap grids (see Chapter 6) and individuals were aged as follows: cubs/juveniles (<2 years old), sub-adults (2–3 years old) and adults (≥3 years old; Balme, Hunter & Braczowski, 2012). Males were identified by having a heavier build, thick neck, dewlap and orange external scrotal sac (which was generally the only way of identifying juvenile and subadult males), whereas females were relatively smaller and lacked external scrotum (Balme et al., 2012).

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3.4 Results

Lions and leopards generally moved one kilometre from the release site (initiating the monitoring period) after one or two days. Lions established social bonds during the time spent in the boma and this enduring relationship continued throughout the study period. The pride typically travelled together; however the males were occasionally sighted separated from the pride, moving together as a coalition. Leopards (released in male-female pairs) showed variable tendencies to remain together. Most noticeably was the three-month association of LEM2 and LEF2 in the north-eastern sector of the reserve.

Distance from release site

The mean distance moved from the release site ranged from 4.75 km (± 1.17 [SD]) to 5.42 km (± 1.47) for lions and 1.50 km (± 1.09) to 19.03 km (± 3.10) for leopards (Table 3.2). LIM1 showed an initial increase in movements away from the release site, although his movements decreased by the third month (Kruskal-Wallis test, H = 0.84, d.f. = 2, p = 0.66). Interestingly, LIF1 showed a significant increase in movements from the release site (H = 9.26, d.f. = 2, p = 0.01), despite being part of the same pride and showing home range reduction during month three (Appendix 3C). This female also recorded the maximum distance travelled for both lions (12.41 km).

Of the six leopards, three individuals (LEM2, LEF2 & LEF3) showed reduced movements by the second month (Table 3.2). For LEF3, we excluded month three due to insufficient data (Table 3.1), although a significant decrease in distance moved from release site was recorded (H = 5.87, d.f. = 1, p = 0.02) (Mann-Whitney U test produced matching results). LEM1 showed an initial increase away from the release site, but by the third month his distances stabilised. For the other two leopards (LEM3 & LEF1), both showed significant increases in movements away from their release sites (H = 53.17, d.f. = 2, p < 0.001; H = 24.63, d.f. = 2, p < 0.001). LEM3 also recorded the greatest distance travelled from the release site (24.93 km).

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Table 3.2. Mean ± SD and Kruskal-Wallis ANOVA result of the distance (km) travelled by reintroduced lions and leopards from the release site during the first three months after their release into Majete Wildlife Reserve, Malawi. Due to collar failure, no data were recorded for LEF3 in the third month. Distance moved Kruskal-Wallis ANOVA ID Code Month 1 Month 2 Month 3 H-value d.f. p-value

Lions

LIM1 4.76 ± 1.94 5.08 ± 2.96 4.75 ± 1.17 0.84 2 0.66

LIF1 4.84 ± 2.28 5.09 ± 2.57 5.42 ± 1.47 9.26 2 0.01

Leopards

LEM1 10.87 ± 6.10 13.10 ± 2.28 12.60 ± 3.73 7.13 2 0.03

LEM2 2.79 ± 0.83 2.51 ± 0.70 2.44 ± 0.87 6.93 2 0.03

LEM3 1.50 ± 1.09 16.82 ± 5.76 19.03 ± 3.10 53.17 2 < 0.001

LEF1 5.37 ± 1.59 6.56 ± 2.84 6.99 ± 2.42 24.63 2 < 0.001

LEF2 2.21 ± 0.82 1.87 ± 0.87 1.78 ± 0.92 17.19 2 < 0.001

LEF3 6.34 ± 4.15 3.93 ± 2.27 - 5.87 1 0.02

Daily distance

There was a general increase in the daily distance moved by felids between the first to the second month after release (Table 3.3). This increase was found to be significant for lions (one-way ANOVA,

F2,172 = 9.72, p < 0.001; Tukey HSD test, p < 0.001). However, between the second and third month, daily distance travelled decreased significantly for lions (Tukey HSD test, p = 0.004). Of the six leopards, only one (LEM3) exhibited significant increases in daily movements (F2,46 = 3.44, p = 0.04). For one female leopard (LEF3), we excluded month three due to collar failure (Table 3.1), yet, a significant decrease was observed after the first month (F1,37 = 7.20, p = 0.01) (T-test produced matching results). The other four leopards displayed reduced (or at least stabilised) daily movements by the third month (Table 3.3). The mean daily distance travelled during the first three months ranged from 1.86 km (± 1.30) to 5.09 km (± 1.94) for leopards and 3.10 km (± 2.53) to 6.03 km (± 2.85) for lions. The maximum distance recorded in a single day for each species was 12.62 km (lions) and 13.40 km (leopards).

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Table 3.3. Mean ± SD and ANOVA result of the daily distance (km) travelled by reintroduced lions and leopards during the first three months after their release into Majete Wildlife Reserve, Malawi. Due to collar failure, no data were recorded for LEF3 in the third month. Distance moved ANOVA ID Code Month 1 Month 2 Month 3 F-value d.f. p-value

Lions

LIM1 4.49 ± 2.83 6.03 ± 2.85 4.02 ± 2.38 4.38 2 0.02 LIF1 3.10 ± 2.53 5.61 ± 2.69 4.55 ± 2.19 7.75 2 0.001

Leopards

LEM1 4.58 ± 2.67 5.09 ± 1.94 4.93 ± 3.71 0.26 2 0.77 LEM2 2.16 ± 1.46 2.72 ± 1.72 2.67 ± 1.97 0.84 2 0.43 LEM3 1.86 ± 1.30 3.25 ± 3.11 4.12 ± 2.12 3.44 2 0.04

LEF1 2.47 ± 1.79 3.52 ± 2.05 3.56 ± 2.19 2.78 2 0.07

LEF2 2.42 ± 1.53 2.93 ± 1.64 3.01 ± 1.60 1.18 2 0.31 LEF3 4.92 ± 3.13 2.56 ± 2.19 - 7.20 1 0.01

Homing behaviour

Mean angles of movement and 95% confidence intervals are presented relative to capture locations to determine homing for lions (Table 3.4) and leopards (Table 3.5). Directional movements for lions and leopards were not different from random (Rayleigh test of uniformity, p > 0.05) and all three methods showed weak homing behaviour (see Appendices 3A & 3B). This suggests that the direction of movement for reintroduced felids was not consistently towards the capture location and therefore homing tendencies were not apparent, although, movements of one male leopard (LEM3) during the second month revealed evidence of homing behaviour, as the direction of home fell within the 95% confidence interval of the mean directional movement.

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Table 3.4. Homing tendencies of released lions in Majete Wildlife Reserve, based on Kilian (2003). Home directions falling within the 95% confidence intervals of mean angles of movement indicated homing behaviour.

Homing Number of Mean angle (µ) ± Rayleigh's ID Code Month Home behaviour angles 95% CI test (z)

LIM1 1 No 26 232° 45° ± 54° 0.09

2 No 31 232° 189° ± 43° 0.29

3 No 27 232° 141° ± 50° 0.15 LIF1 1 No 25 227° 156° ± 50° 0.24

2 No 27 227° 40° ± 52° 0.17

3 No 30 227° 141° ± 49° 0.16

Table 3.5. Homing tendencies of released leopards in Majete Wildlife Reserve, based on Kilian (2003). Home directions falling within the 95% confidence intervals of mean angles of movement indicated homing behaviour. No data were available in third month for LEM3 due to collar failure.

Homing Number of Mean angle (µ) ± Rayleigh's ID Code Month Home behaviour angles 95% CI test (z) LEM1 1 No 27 251° 189° ± 35° 2.09

2 No 24 251° 1° ± 38° 1.57

3 No 24 251° 114° ± 53° 0.11 LEM2 1 No 28 245° 305° ± 44° 0.42

2 No 25 245° 26° ± 52° 0.12

3 No 26 245° 52° ± 50° 0.14 LEM3 1 No 27 241° 161° ± 39° 1.04

2 Yes 25 241° 227° ± 38° 0.88

3 No 15 241° 344° ± 51° 0.66 LEF1 1 No 28 245° 2° ± 54° 0.15

2 No 27 245° 109° ± 39° 1.04

3 No 26 245° 360° ± 51° 0.11 LEF2 1 No 27 230° 56° ± 43° 0.48

2 No 25 230° 53° ± 49° 0.31

3 No 29 230° 316° ± 54° 0.48 LEF3 1 No 28 252° 314° ± 44° 0.13

2 No 29 252° 341° ± 37° 1.29

3 - - 252° - -

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Range establishment

The establishment of lion range sizes during the first three months after their release into the reserve are presented in Figures 3.2 and 3.3. A detailed report on range establishment is given in Appendix 3C. Range size (based on 100% MCP) increased significantly for both lions from month one to two (one- way ANOVA, F1,3 = 22.11, p = 0.02; Tukey HSD test, p = 0.02). However, their ranges decreased significantly between month two and three (Tukey HSD test, p = 0.02). Interestingly, LIF1 had slightly larger ranges in month one and three compared to LIM1, despite them being members of the same, solitary pride.

Leopard ranges during the first three months after release varied substantially (see Figures 3.4 – 3.9). Overall, leopards showed a reduction in home range towards the third month, although this difference

2 was not significant (F2,13 = 0.21, p = 0.81). LEM1 decreased his home range from 115.36 km in month one to 48.49 km2 in month 3. The ranges of LEM2 remained stable throughout the monitoring period and were significantly smaller compared to those of LEM1, based on 100% MCP (T-test, t = –4.38, d.f. = 4, p = 0.01). The UD method was unable to estimate range size during the third month, so range establishment for LEM3 was based on the MCP method (to maintain consistency for each individual; see Appendix 3D). Range size for this animal increased substantially between month one (16.88 km2) and month two (104.71 km2), but decreased towards month three (52.60 km2), as he moved into a vacant area in the south-east of the reserve (Figure 3.6). However, this individual showed further wide- ranging behaviour after the initial monitoring period (Chapter 4), thus did not appear to stabilise his home range.

For the female leopards, overall range size of LEF2 was significantly smaller compared to those of LEF1 (t = 3.08, d.f. = 4, p = 0.04). Her monthly ranges remained stable, similar to LEM2, and both clearly showed release site fidelity (see Figure 3.1, 3.5 & 3.8). LEF3 had an extremely large range during month one (91.31 km2), which was likely influenced by her escape from the boma prior to her scheduled release a few days later. Despite this, LEF3 showed little roaming behaviour and established her range in the east of the reserve, following a range reduction in the second month (17.85 km2; Figure 3.9, personal observation). Unlike the other leopards, LEF1 did not show reduced ranges during the first three month, as her range size increased from 10.21 km2 in month one to 61.47 km2 in month three. However, upon further investigation her range size eventually decreased to 57.37 km2 in the fourth month.

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Month 1 Month 2 Month 3

Figure 3.2. Range size (95%, 75% and 50% UD) change of LIM1 during the first three months after his release in Majete Wildlife Reserve, Malawi. Star indicates release site.

Month 1 Month 2 Month 3

Figure 3.3. Range size (95%, 75% and 50% UD) change of LIF1 during the first three months after her release in Majete Wildlife Reserve, Malawi. Star indicates release site.

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Month 1 Month 2 Month 3

Figure 3.4. Range size (95%, 75% and 50% UD) change of LEM1 during the first three months after his release in Majete Wildlife Reserve, Malawi. Star indicates release site.

Month 1 Month 2 Month 3

Figure 3.5. Range size (95%, 75% and 50% UD) change of LEM2 during the first three months after his release in Majete Wildlife Reserve, Malawi. Star indicates release site. 59

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Month 1 Month 2

Figure 3.6. Range size (95%, 75% and 50% UD) change of LEM3 during the first two months after his release in Majete Wildlife Reserve, Malawi. Data were not sufficient to produce range sizes for month 3. Star indicates release site.

Month 1 Month 2 Month 3

Figure 3.7. Range size (95%, 75% and 50% UD) change of LEF1 during the first three months after her release in Majete Wildlife Reserve, Malawi. Star indicates release site.

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Month 1 Month 2 Month 3

Figure 3.8. Range size (95%, 75% and 50% UD) change of LEF2 during the first three months after her release in Majete Wildlife Reserve, Malawi. Star indicates release site.

Month 1 Month 2

Figure 3.9. Range size (95%, 75% and 50% UD) change of LEF3 during the first two months after her release in Majete Wildlife Reserve, Malawi. Data were not sufficient to produce range sizes for month 3. Star indicates release site.

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Post-release breeding

Lions

The three lions, a single female (LIF1) and a male coalition (LIM1 and LIM2), reintroduced into Majete, formed an enduring pride. Both males were observed mating with the lioness, and in 2013 two cubs (one female and one male) were born (Table 3.6). In 2015, the male cub (by then a young adult) was evicted from the pride and moved to a vacant area in the south of the reserve. LIF1 had another litter in 2015, and again two cubs (both females) were born. The male coalition then mated with their first daughter, and in 2016 a single male lion was born. LIF1 had a third litter in 2017, which consisted of three cubs (sex unknown). Therefore, the lion population had grown from three to eleven individuals between 2012 and 2017, with a sex ratio of 1:1 (M:F).

Leopards

Sightings of reintroduced leopards were infrequent and thus camera trapping proved helpful in monitoring post-release breeding. Two reintroduced females (LEF1 and LEF3) had three litters between them (Table 3.6). LEF1 and her daughter were observed together on two separate occasions, while LEF3 and her second litter of two cubs were observed with a subadult male (2-3 years old) for several days, presumably from her first litter.

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Table 3.6. Details of the reintroduction, biology and post-release breeding of reintroduced cats in Majete Wildlife Reserve, Malawi.

Release Age at release ID Code Sex Origin Breeding success Comments date (years)

Lion Unknown relation to LIM2; found together on LIM1 M 07/2012 3 Pilanesberg National Park Mated with LIF1 a carcass at the capture site Unknown relation to LIM1; found together on LIM2 M 07/2012 3.5 Pilanesberg National Park Mated with LIF1 a carcass at the capture site Successfully raised two litters LIF1 F 07/2012 3 Madikwe Game Reserve (2013 & 2015); four cubs Leopard

LEM1 M 10/2011 2–3 Kruger National Park - Located throughout Majete

LEM2 M 01/2012 3–4 Ohrigstad - Last seen: 07/2014

LEM3 M 11/2012 7 Loskop Dam - Located in the south of Majete One successful litter (2014); LEF1 F 10/2011 2 Ohrigstad Mother and daughter observed together twice raised one cub Thaba Ingwe Nature LEF2 F 01/2012 2–3 - Last seen: 09/2015 Reserve Malelane, Southern Two successful litters (2013 & Observed with second litter and another male LEF3 F 11/2012 5 Kruger National Park 2016); raised four cubs (presumably from her previous litter)

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3.4 Discussion

The reintroduction of lions and leopards into MWR was successful by our definition, as animals soon decreased movement rates, established permanent home ranges and showed weak tendencies to return home. Our findings suggest that reintroduced lions and leopards did not encounter the historic difficulties related to felid translocation, which corresponded with other studies (Hayward, Adendorff, Moolman, Hayward & Kerley, 2007c; Hunter, 1998; Kilian, 2003; Yiu et al., 2015).

Distances moved

Lion and leopard displayed varied movements from the release site. An initial increase in movement rates was found, with two male leopards (LEM1 and LEM3) travelling the largest distances, which seemed to stabilise with time. Felids are expected to extend their movements from the release site until they find an empty ‘patch’ to establish their home range, while others may remain near or even return to the release site to establish their range (see Hunter, 1998; Kilian, 2003; Yiu et al., 2015). This potentially explains the variation recorded in the post-release movements of lion and leopard in MWR. The distance travelled from the release site broadly corresponded with the daily distances for each animal. Movements of released felids found in this study were similar to those recorded for lion and cheetah (Hunter, 1998; Kilian, 2003).

In the Dinokeng Game Reserve, South Africa, Yiu et al. (2015) found that post-release movements of lions released first were generally higher, due to a lack of intraspecific competition, than those released later. Equally, leopards reintroduced first into MWR (LEM1 and LEF1) could explore and establish territories without intraspecific competition, resulting in greater distances travelled than subsequently released individuals. In our study, this trend applied to daily distances moved, rather than distances from release site. This could be explained by increased competition in the north of the reserve, forcing the third set of leopards (LEM3 and LEF3) to the southern section to find a vacant area. However, individual history carried over from their natal environment, in this case livestock raiding, may also influence dispersal and range establishment decisions (van der Meulen, 1977; Weilenmann et al., 2010).

Direction of movement

None of the felids released into MWR showed consistent homing behaviour. However, LEM3 exhibited apparent homing tendencies during the second month after release. This could be attributed to his movement to the southern section of the reserve (likely to avoid intraspecific competition), where he established his range, which was in a similar direction to his original capture site.

Homing calculations were based on three methods. Method one recorded homing in two male leopards. Method two and three yielded similar results, each with only one incident of homing, which

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was similar to the lack of homing behaviour recorded for reintroduced lions (Kilian, 2003) and translocated leopards (Weise et al., 2015). However, these methods must be interpreted with caution, as there are several factors that could influence initial movements of released cats from the release site. For example, fences may confound homing measurements by obstructing intended movements of individuals, while the position of the release site relative to specific habitat types and areas of high prey densities may also affect movement decisions (Hunter, 1998; Kilian, 2003). In MWR, lions and leopards were released in the north-eastern section of the reserve, which had a high prey concentration, while two leopards were released next to the Shire River, which formed a barrier to initial movement decisions. Ultimately, if released felids did exhibit consistent homing tendencies, one would have expected them to attempt to leave the reserve boundaries after their release or at least remain close to the fence line for prolonged periods (Kilian, 2003).

Weise et al. (2015) suggested that long distance translocations (>200 km) of leopards could prevent homing behaviour and proposed a minimum recipient reserve size of 875 km2 to enhance release site fidelity. However, Kilian (2003) and Yiu et al. (2015) recorded release site fidelity and an absence of homing for lions translocated <200 km, into small reserves (<500 km2). In our study, animals were translocated >1 000 km and the recipient reserve size (700 km2) appeared sufficiently large for reintroduced individuals. Consequently, we can assume that translocation distance and reserve size (given that it is large enough to sustain more than one individual) probably do not influence post- release movements of lions, which may be explained by their confinement to fences, when the population is below carrying capacity (Hayward et al., 2007a). In contrast, leopards have the ability to move relatively freely between fenced reserves and adjacent landscapes (Balme & Hunter, 2004), which explains why leopards would need a large-enough reserve size to contain their home ranges, while being translocated far-enough to prevent them from returning to their capture site. Additionally, the presence of resident individuals influences ranging behaviour of newcomers, often leading to escapes and subsequent edge effects (Weilenmann et al., 2010). Based on this and other studies (Hamilton, 1981; Hayward et al., 2007c, 2009; Hunter, 1998; Stander, 1990; Weilenmann et al., 2010; Weise et al., 2015), we suggest that animal history and conspecific density are likely to influence post- release movements and ranges, while translocation distance and reserve size likely influences leopards more than lions.

Range establishment

The establishment of lion ranges within MWR corresponded with that of reintroduced lions in other small, enclosed reserves (Hunter, 1998; Kilian, 2003). Both lions settled down soon after their release and established permanent home ranges in the north-eastern part of the reserve. High prey densities were found in this area due to the presence of readily available water in the form of two perennial rivers, four artificial waterholes and a few natural springs. Given that lion ranges are driven by resource 65

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availability (Loveridge et al., 2009) and inter-pride competition (Schaller, 1972; Mosser & Packer, 2009), we deduce that the range selection for lions was based on optimal resource availability, although the selection could also be influenced by the release site.

Leopards ranges varied considerably, although a general range reduction was observed within the first three months after their release into MWR. To our knowledge, this is the first-time range establishment has been investigated for reintroduced leopards. Previous studies have either conducted research on already established leopards (Cristescu et al., 2013) or have failed to locate the released animal during the range establishment period (Hayward et al., 2007c). One study recorded the seasonal home range of a single male leopard reintroduced into Addo Elephant National Park, South Africa, but no information was given on how he established his range (Hayward et al., 2009).

Our findings suggest that reintroduced leopards require at least three or four months to establish permanent ranges after reintroduction into an enclosed reserve with low leopard density. Purchase and du Toit (2000) recorded home range reduction of fourteen cheetahs in Matusadona National Park, Zimbabwe, five years after their reintroduction, which they attributed to range establishment. However, establishment periods may be influenced by conspecific density, especially at the release site (Weise et al., 2015). For example, LEM1 and LEF1 were released into a leopard-free environment which may explain the lengthy range establishment periods and distances travelled after release; LEM2 and LEF2 had smaller ranges with rapid range reduction periods; and LEM3 and LEF3 displayed extensive movements initially, probably to locate a vacant patch, although their capture site history (livestock raiders) could also explain their movement patterns (Weilenmann et al., 2010).

Factors influencing success of carnivore reintroductions

In the past, large carnivores were released directly (hard-release) into the environment which often resulted in extensive roaming behaviour with tendencies to return to the original capture site (Linnell et al., 1997). Today, researchers and managers alike are aware of the factors (e.g. biological considerations and technical elements) influencing project success and can thus increase the chances of success by implementing relevant guidelines (Hayward & Somers, 2009; Hunter et al., 2007; Hayward et al., 2007a). For instance, keeping carnivores in a boma for a temporary period (soft- release), is designed to allow translocated animals to recover from the capture and transport, acclimatise in their new environment and familiarise themselves with additionally translocated individuals (Hunter et al., 2007; Somers & Gusset, 2009; van Dyk, 1997). The holding period varies across studies and species. For example, Hayward et al. (2007a) recorded that leopards required the most time in the boma (more than four months); in MWR however, leopards spent an average of 17 days in the boma. Similarly, lions were held for about four weeks, despite the average being six to eight weeks (Hunter et al., 2007). The longer the period spent in the boma, the greater the chance of predators reducing their elusiveness (especially for leopards) and fear of humans (Hayward et al., 66

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2007a). Therefore, the primary reason for the reintroduction (e.g. eco-tourism or predator re- population) should dictate the length of time spent in the boma.

It is also crucial to expose felids to electric fencing in the boma so they may become accustomed to the predator-proof perimeter fence and avoid escapes (Hunter et al., 2007; Hayward et al., 2007a; van Dyk, 1997). This proved effective for both lions and leopards, excluding the several brief escapes of leopards, which occurred after the three-month study period. The location of the release is also important and should be situated in an area with suitable habitat, a high prey concentration and low conspecific density (Weise et al., 2015). The findings from this and other studies (Hunter, 1998; Kilian, 2003) showed that reintroduced lions and leopards remained relatively close to the release site for the first month, suggesting that they acclimatised to their new environment.

Lions established a single pride during the captivity period, which endured beyond the study period (Hunter, 1998). This is often an additional benefit of a holding facility when reintroducing social carnivores such as lions (Hayward et al., 2007b; Hunter et al., 2007), African wild dog (Lycaon pictus; Gusset, Slotow & Somers, 2006) and grey wolf (Canis lupus; Fritz et al., 1997). Leopards, on the other hand, are solitary animals and all individuals were therefore kept in separate enclosures. Interestingly, the second set of leopards (LEM2 & LEF2) remained in the same area for three months after their release. This suggests social interaction between these individuals, likely due to an affiliation established in their adjacent bomas prior to release and their release at the same location.

It is believed that age of released animals can influence whether they will return home or not (Kilian, 2003). Felids translocated at a young age (≤24 months) are expected to establish themselves within the recipient environment, as this is the age when they typically disperse and search for an open area, whereas adults are already established in an area (at the capture site) and will likely return to this area. Of the released cats in MWR, all three lions were adults (36 to 42 months old), while leopards ranged from young to older adults (24 to 84 months old). However, as previously mentioned, none showed consistent homing tendencies and therefore it is unlikely the age influences homing tendencies, as previously believed.

Post-release breeding

The lion pride was extensively monitored since their release into Majete. The population grew from three to eleven individuals, with four litters documented over five years, one of which included a cub from the initial offspring of the reintroduced female. This provides evidence for breeding success and encouraging signs of populations growth five years post-release. The male coalition maintained tenure over the single pride throughout the study period, which is almost double the average natural tenure of male lions (Nowell & Jackson, 1996; Packer et al., 1988), but in line with tenures recorded for other reintroduced lions (e.g. Trinkel et al., 2010). Due to the small founder population and risk of inbreeding,

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immediate management action is required. Management plan to translocate the male coalition to Liwonde National Park, Malawi to establish a lion population; while translocations into Majete include a new male coalition (n = 2) and four females from Zimbabwe in February 2018 (C. Hay, personal communication, March 10, 2017).

Monitoring post-release breeding of leopards was challenging. Infrequent sightings revealed that two reintroduced females raised three litters between them and several subadults were found within the reserve. This provides encouraging signs for leopard population growth post-reintroduction. Additionally, one cub and one subadult each accompanied a female, likely offspring of reintroduced individuals. Interestingly, two females (LEF1 and her daughter aged 2.5–3 years old) were observed together on two different occasions, while on another occasion, an adult female and her two cubs were observed with a subadult male for several days, which we presume was from her previous litter. These groupings could suggest that female leopards exhibit prolonged parental care of up to 30–36 months in Majete (Balme, Robinson, Pitman & Hunter, 2017), however, this data are based on brief observational sightings.

3.6 Conclusions

The successful reintroduction of lions and leopards in MWR and elsewhere, demonstrates the efficacy and viability of reintroductions for re-populating large carnivores in areas where they were extirpated. The success of large carnivore reintroductions is largely influenced by pre-release management. For example, temporarily keeping reintroduced carnivores in a boma prior to their release may reduce or even prevent translocation-associated problems such as homing and ultimately enhance the likelihood of individual establishment and persistence within a reserve. Furthermore, post-release monitoring is essential for understanding the movements and behaviour of released animals, which will ultimately help to determine whether reintroductions were successful or not. From this and other studies, it appears that release site, habitat quality, high prey concentrations and intraspecific competition plays an important role in the post-release movements and range establishment of released cats. Our findings, which illustrate the importance of monitoring felids post-release, will be utilised by management for future translocations or reintroductions. We emphasise the importance of collaboration between research and management and encourage the application of scientific knowledge to reserve management. Finally, it is important to broaden our understanding of reintroduction biology, especially since the continued decline of large felids in Africa suggests that reintroductions are likely to continue into the future (Somers & Gusset, 2009).

3.7 Acknowledgements

Thank you to the Earthwatch Institute for funding the Animals of Malawi Project and for funding this study. I would like to thank Craig Hay and Gervaz Tamala for providing access to collar data. 68

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3.9 Appendices

Appendix 3A. Comparison of three methods used to determine homing tendencies of lions in Majete Wildlife Reserve. Home directions falling within the 95% confidence intervals of mean angles of movement indicated homing behaviour (method 1 and 2), whereas mean angles of direction falling within a 22.5° range of the home direction indicated homing (method 3).

Method 1a Method 2b Method 3c

Home Mean angle (µ) Homing Mean angle (µ) Homing Home direction Mean angle Homing ID Code Month n direction ± 95% CI behaviour ± 95% CI behaviour ± 22.5° (µ) behaviour

LIM1 1 26 232° 162° ± 4° No 45° ± 54° No 232° ± 22.5° 45° No

2 31 232° 170° ± 6° No 189° ± 43° No 232° ± 22.5° 189° No 3 27 232° 172° ± 5° No 141° ± 50° No 232° ± 22.5° 141° No

LIF1 1 25 227° 154° ± 3° No 156° ± 50° No 227° ± 22.5° 156° No

2 27 227° 168° ± 6° No 40° ± 52° No 227° ± 22.5° 40° No 3 30 227° 171° ± 5° No 141° ± 49° No 227° ± 22.5° 141° No aFrom Hunter (1998); mean direction of movements based on movements from the release site with confidence intervals, relative to home direction bFrom Kilian (2003); mean direction of movement based on consecutive movements with confidence intervals, relative to home direction cFrom Fies et al. (1987); mean direction of movement based on consecutive movements relative to a home direction range of 22.5°

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Appendix 3B. Comparison of three methods used to determine homing tendencies of leopards in Majete Wildlife Reserve. Home directions falling within the 95% confidence intervals of mean angles of movement indicated homing behaviour (method 1 and 2), whereas mean angles of direction falling within a 22.5° range of the home direction indicated homing (method 3).

Method 1a Method 2b Method 3c Home Mean angle (µ) Homing Mean angle (µ) Homing Home direction Mean angle Homing ID Code Month n direction ± 95% CI behaviour ± 95% CI behaviour ± 22.5° (µ) behaviour LEM1 1 27 251° 146° ± 9° No 189° ± 35° No 251° ± 22.5° 189° No 2 24 251° 231° ± 5° No 1° ± 38° No 251° ± 22.5° 1° No 3 24 251° 249° ± 3° Yes 114° ± 53° No 251° ± 22.5° 114° No LEM2 1 28 245° 99° ± 13° No 305° ± 44° No 245° ± 22.5° 305° No 2 25 245° 108° ± 11° No 26° ± 52° No 245° ± 22.5° 26° No 3 26 245° 98° ± 10° No 52° ± 50° No 245° ± 22.5° 52° No LEM3 1 27 241° 324° ± 36° No 161° ± 39° No 241° ± 22.5° 161° No 2 32 241° 230° ± 5° No 227° ± 38° Yes 241° ± 22.5° 227° Yes 3 15 241° 241° ± 4° Yes 344° ± 51° No 241° ± 22.5° 344° No LEF1 1 22 245° 190° ± 5° No 2° ± 54° No 245° ± 22.5° 2° No 2 27 245° 201° ± 10° No 109° ± 39° No 245° ± 22.5° 109° No 3 26 245° 188° ± 4° No 360° ± 51° No 245° ± 22.5° 360° No LEF2 1 27 230° 111° ± 9° No 56° ± 43° No 230° ± 22.5° 56° No 2 23 230° 83° ± 11° No 53° ± 49° No 230° ± 22.5° 53° No 3 23 230° 79° ± 14° No 316° ± 54° No 230° ± 22.5° 316° No LEF3 1 36 252° 197° ± 11° No 314° ± 44° No 252° ± 22.5° 314° No 2 29 252° 217° ± 44° No 341° ± 37° No 252° ± 22.5° 341° No

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3 7 252° 230° ± 6° - 335° ± 72° - 252° ± 22.5° 335° -

aFrom Hunter (1998); mean direction of movements based on movements from the release site with confidence intervals, relative to home direction bFrom Kilian (2003); mean direction of movement based on consecutive movements with confidence intervals, relative to home direction cFrom Fies et al. (1987); mean direction of movement based on consecutive movements relative to a home direction range of 22.5°

Appendix 3C. Range establishment (km2) of reintroduced lions during the first three months after their release into Majete Wildlife Reserve, Malawi.

MCP UD ID Code Month Fixes 100% 95% 50% 95% 75% 50% LIM1 1 151 22.18 15.31 6.11 10.72 5.23 2.33 2 169 57.51 51.40 5.79 17.86 6.16 1.94 3 177 22.21 14.22 2.11 6.34 2.67 0.98 LIF1 1 170 30.34 26.96 7.21 15.14 6.08 1.81 2 159 53.49 46.00 7.20 16.92 6.52 1.66 3 159 30.64 23.63 2.31 8.60 3.63 1.58

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Appendix 3D. Range establishment (km2) of reintroduced leopards during the first three months after their release into Majete Wildlife Reserve, Malawi.

MCP UD ID Code Month Fixes 100% 95% 50% 95% 75% 50% LEM1 1 157 177.82 159.22 89.37 115.36 46.33 14.89 2 160 87.48 58.38 20.61 42.45 17.79 8.16 3 173 116.64 86.15 15.29 48.49 18.92 5.26 LEM2 1 93 10.10 9.69 7.00 8.70 3.78 1.23 2 140 10.09 9.55 4.79 6.87 3.61 1.28 3 180 11.96 10.94 4.55 9.68 3.72 1.10 LEM3 1 30 16.88 10.62 3.02 5.33 3.18 1.22 2 33 104.71 100.75 19.38 46.77 22.35 13.13 3 18 52.60 52.60 15.55 - - - LEF1 1 153 33.89 20.16 3.50 10.21 3.74 0.93 2 155 90.16 77.76 11.23 36.68 13.19 5.15 3 176 100.59 89.56 34.40 61.47 28.95 13.43 LEF2 1 171 10.20 9.19 2.99 6.49 3.57 1.51 2 169 12.41 10.13 4.73 7.05 3.93 1.57 3 177 10.77 9.09 3.84 7.57 3.07 0.72 LEF3 1 56 142.33 133.61 45.38 91.31 33.30 22.00 2 41 30.23 26.92 11.77 17.85 10.37 3.91 3 3 ------

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Chapter 4

Home range and habitat selection of reintroduced lion (Panthera leo) and leopard (Panthera pardus) in Majete Wildlife Reserve, Malawi

W.D. Briers-Louw1, A.J. Leslie1

1Department of Conservation Ecology and Entomology, Stellenbosch University, Matieland, Western Cape, 7602, South Africa

4.1 Abstract

Home range studies provide vital information with regards to ecological, behavioural and demographic factors influencing the spatial distribution of an animal within a given area. When large carnivores are translocated into an entirely new environment, they must adapt to a new suite of resources and establish home ranges, while reducing competition with conspecifics or more dominant carnivores. Between 2012 and 2017 GPS collars were used to determine the home range and habitat selection of two lions (Panthera leo) and six leopards (Panthera pardus) that were reintroduced into Majete Wildlife Reserve. Mean ± SE home range using 100% minimum convex polygon (MCP) was 380.45 ± 117.70 km2 for lions and 495.08 ± 80.99 km2 for leopards, while using 95% utilisation distribution (UD), ranges were 134.00 ± 6.66 km2 for lions and 257.94 ± 52.51 km2 for leopards. These results are the largest on record for any reintroduced felid. Based on Jacobs’ indices, all felids preferred riverine habitat and avoided high-altitude miombo woodland and extensive range overlap was found within and between species. We suggest that the coexistence of leopards with lions is likely due to the leopard’s adaptability in terms of behaviour and diet. Despite a few incidences of roaming behaviour, we advocate the use of reintroductions to re-populate large predators in areas of their former distribution, provided that the cause of the extirpation is resolved. Prior to the translocation of additional individuals, we recommend that managers use home range estimates of established individuals to carefully select release sites, preferably in areas with high habitat quality and low intraspecific competition, to increase the likelihood of establishment within the reserve.

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4.2 Introduction

An animal’s spatial utilisation reveals its relationship to the resources within a given environment (Grimbeek, 1992). Quantifying an animal’s home range may provide valuable insight into the ecological, behavioural and demographic factors influencing their spatial distribution. Home range is defined as the area normally traversed by an animal or group of animals during activities of food gathering, shelter-seeking, reproduction and caring for young (Burt, 1943). This definition excludes exploratory “sallies” and therefore does not include the total area moved by an individual (Burt, 1943).

Carnivores have the largest ranges of any mammalian group (Harestad & Bunnell, 1979). Home range size varies by several orders of magnitude across carnivore species (e.g. Nielsen, Herfindal & Linnell, 2005). This interspecific variation in home range size is linked to metabolic requirements based on body mass (Gittleman & Harvey, 1982; Harestad & Bunnell, 1979; Lindstedt, Miller & Buskirk, 1986; McNab, 1963). As body mass increases, home range will also increase to satisfy the species energetic needs (Carbone, Cowlishaw, Isaac & Rowcliffe, 2005). This would explain why large carnivorous species, that regularly consume large-bodied prey species (Carbone, Mace, Roberts & MacDonald, 1999), have the largest home ranges and the lowest densities relative to their energetic requirements (Gittleman & Harvey, 1982; Kelt & van Vuren, 2001; Swihart, Slade & Bergstrom, 1988).

Home range size also differs within species and is typically influenced by resource availability (Hunter, 1998; Mizutani & Jewell, 1998; Sandell, 1989). For example, the home ranges of lion (Panthera leo) and leopard (Panthera pardus) vary substantially by region due to a variation in abundance of prey between regions. In arid regions, both felids adopt larger home ranges to locate prey, which normally occur at low densities. Stander (1991) recorded lion prides with home ranges of up to 2 075 km2 in Etosha National Park, Namibia, and Funston (2011) documented lion home ranges over 4 500 km2 in the Kgalagadi Transfrontier Park, South Africa/Botswana. Similarly, leopards in the Kalahari Desert, South Africa, were found to occupy home ranges as large as 2 182 km2 (Bothma, Knight, le Riche & van Hensbergen, 1997). In more mesic habitats, higher prey densities result in smaller home ranges. Home ranges as small as 45 km2 were recorded for lions in the Ngorongoro Crater, Tanzania (Hanby, Bygott & Packer, 1995) and leopard ranges were at a minimum of 10 km2 in both the Matopos National Park, Zimbabwe (Smith, 1977) and Tsavo National Park, Kenya (Hamilton, 1976).

Apart from resource availability, sex and social dominance hierarchy can also influence intraspecific home range size (Hayward, Hayward, Druce & Kerley, 2009). The home range of a female carnivore is determined by food supply and habitat quality needed to raise cubs successfully, whereas a male’s home range is closely tied to an ability to find and successfully mate with one or more females without disturbance from other males (Bailey, 1993; Mizutani & Jewell, 1998; Sandell, 1989). Therefore, a

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male’s home range size is largely dictated by the distribution of females in the area (Loveridge et al., 2009).

Lions are highly social felids, living in family groups called ‘prides’ with between 2–21 individuals (Mosser & Packer, 2009; Packer & Pusey, 1982). Related females and their young form the fundamental unit of a pride and typically maintain the same home range for generations (Hanby et al., 1995; Packer, Scheel & Pusey, 1990). Males form coalitions of one to six individuals and hold tenure over one or more prides for an average of two (Packer et al., 1988) to three (Stander, 1991) years. Prides are characterised by fission-fusion dynamics as members frequently subdivide into smaller groups when they scatter within the pride’s home range (Mosser & Packer, 2009; Packer et al., 1990; West & Packer, 2013). Lions defend their territories by scent marking, roaring and aggressive encounters (Grinnell, Packer & Pusey, 1995; Mosser & Packer, 2009; Schaller, 1972), with males patrolling the perimeter of the pride’s home range and females defending the core area against conspecifics of the same sex (West & Packer, 2013).

In contrast, leopards are solitary felids (Bailey, 1993). The only exception of sociality in leopards is when females raise cubs to sub-adulthood or when males and females briefly associate during mating (Bailey, 1993; Stander, Hayden, Kaqece & Ghau, 1997). However, “spatial groupings” are known to exist when home ranges of individual’s overlap (MacDonald, Mosser & Gittleman, 2010). The degree of overlap varies between sex and age (Bailey, 1993; MacDonald et al., 2010). For instance, there is usually little home range overlap between resident neighbouring adult males, however the home range of a single male may encompass one to several female home ranges (Bothma & Coertze 2004; Stander et al., 1997). Home ranges of transient leopards (subadults or old adults) are variable in size and typically superimposed on both adult male and female home ranges (Bailey, 1993; MacDonald et al., 2010). Ultimately, the socio-spatial distribution of leopards is influenced by an individual’s territory, which is defended by scent-marking, tree scratching and vocalisations (Hunter, Henschel & Ray, 2013).

Despite this knowledge, there are few detailed studies on range use of reintroduced felids in Africa (Cristescu, Bernard & Krause, 2013; Druce et al., 2004; Hayward et al., 2009; Hunter, 1998; Kilian, 2003). Of these studies, only two compare range use of lion and leopard. Hayward et al. (2009) recorded home ranges of six lions but only one male leopard. Similarly, Cristescu et al. (2013) compared home ranges of a single lioness with three cubs and a solitary female leopard, with only 19 locations recorded for the female leopard.

Since these felids compete for space, inter- and intraspecific competition can influence the home ranges of reintroduced felids in closed systems (Hunter, 1998), as high predator densities can result in high mortality rates of inferior species (Cristescu et al., 2013). Therefore, evaluating the impact of established individuals on the home range of newly released individuals can help us to understand the spatial response of felids to conspecifics and other species within small, enclosed reserves. The 81

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continued decline of Africa’s big cats suggests that reintroductions are also likely to continue into the future (Somers & Gusset, 2009). Monitoring individuals post-release has become an essential practice for researchers and reserve managers to understand the underlying factors driving spatial distributions of reintroduced felids in small reserves. This eventually helps to determine whether a reintroduction project was successful or not.

Majete Wildlife Reserve (MWR), in the south of Malawi, supported a great variety of mammals in the past (Bell, 1984; Hayes, 1979; Morris, 2006; Sherry, 1989). However, wide-scale poaching, as well as a lack of law-enforcement and management resulted in the local extinction of the majority of the reserve’s wildlife by the early 2000s, including the extirpation of lions and leopards. Between 2003 and 2011, over 2550 animals from 12 different species were reintroduced into MWR with the aim of restoring the reserve to its previous state. Between late 2011 and 2012 three lions and six leopards were also reintroduced into MWR, which provided a unique opportunity to study the post-release spatial movements of these two felids. The main aim of this study was to determine home range, habitat selection and range overlap of reintroduced lion and leopard in MWR.

4.3 Methods

4.3.1 Study site

Majete Wildlife Reserve covers 700 km2 of the Lower Shire Valley region in southern Malawi. MWR has a distinct wet season from December to May and a dry season from June to November. Mean annual precipitation varies from 680–800 mm in the east to 700–1 000 mm in the west (Wienand, 2013). The Shire and Mkulumadzi Rivers are the only two perennial rivers, with several natural springs and ten artificial waterholes also found within the reserve. Altitude is highest in the western region (Diwa and Namitsempha), where steeply undulating hills are dissected by rivers. The terrain gradually flattens towards the eastern region and altitude is lowest along the Shire River.

Based on Sherry (1989), the habitat (i.e. vegetation type, elevation and terrain) is divided into six fine- scale types. High-altitude miombo woodland occurs in the western hilly region between 410–770 m. This supports a high biomass of low browse quality, with Brachystegia boehmii and Julbernardia globiflora representing the dominant trees. Medium-altitude mixed woodland is the ecotonal region between the higher-lying west and lower-lying east, between 230–410 m. This is dominated by Brachystegia boehmii, Pterocarpus rotundifolius and Combretum spp. Low-altitude mixed woodland occurs between 205–280 m. This type is described as relatively open woodland with a dense canopy layer and dominant trees include Acacia spp. and Steculia spp. Ridgetop mixed woodland (220–300 m) is a low biomass, intermediate density woodland dominated by Diospyros kirkii, Terminalia sericea and Diplorhynchus condylocarpon. Riverine habitat (or associations) are found along rivers (below 230 m) and are dominated by species such as Kigelia africana and Acacia tortilis. Riparian thicket (closely 82

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associated with riverine habitat) is relatively small and dense, with Adansonia digitata, Albizia anthelmintica and Euphorbia ingens representing the dominant species (Sherry, 1989).

Figure 4.1. Map of the habitat types and rivers in Majete Wildlife Reserve, Malawi (Shapefiles, personal communication, African Parks (Pty) Ltd.).

4.3.2 Immobilisation and collaring

All reintroduced lions and leopards were collared during the pre-release stage. Individuals were immobilised by veterinarians using standard drug combinations (Kreeger, 1996). Two lions (one male and one female) and all six leopards (three males and three females) were fitted with global positioning system (GPS) satellite collars (n = 8, African Wildlife Tracking, Pretoria, South Africa). Lion collars were made from thick industrial belting and weighed approximately 900 g (<1% of body weight). Leopard collars were made from leather belting and weighed approximately 600 g (<2% of body weight). All collars were fitted to animals using pop-rivets and were programmed to record GPS locations at four- hour intervals to maximise data output, while conserving battery life. Satellite collars stored location data which were downloaded from the African Wildlife Tracking website (www.awt.co.za). Collars transmitted data for over one year for all leopard individuals, except for one collar (LEF3; Table 4.1), 83

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which only transmitted data for three months. Three leopards were captured in September 2013 to remove their collars, but the other three individuals could not be captured after their collars had expired. Lion collars transmitted data for over two years and were subsequently replaced in 2014. In 2016, lions were collared with GPS/global system for mobile communication (GSM) units (i.e. GPS/GSM, GPS accessible via mobile phone).

4.3.3 Statistical analysis

Home range estimation

Home range analyses of reintroduced felids were conducted in ArcGIS 10.5 (Environmental Systems Research Institute (ESRI), Redlands, California, U.S.A.). Location points (or fixes) from GPS data recorded between 2011 and 2017 were used for home range analyses. Latitude and longitude positions were converted to the Universal Transverse Mercator (UTM) co-ordinate system by using WGS84 36 S datum for compatibility in ArcGIS. The minimum number of location points required to estimate animal home ranges, while obtaining statistically relevant results, is 30–50 location points (Kernohan, Gitzen & Millspaugh, 2001; Marzluff, Millspaugh, Hurvitz & Handcock, 2004; Seaman et al., 1999). Location points were initially limited to one point per day to minimise autocorrelation i.e. the lack of independence of sequential data locations (Swihart & Slade, 1985), however, this tends to reduce the representativeness of the relative amount of time spent in various areas of a reserve (Otis & White, 1999). Instead, equally-spaced intervals of location points can reduce autocorrelation and provide more precise estimates (de Solla, Bonduriansky & Brook, 1999; Finberg, 2006). Therefore, to reduce autocorrelation and enhance representativeness of data, collars were set to report location data at four-hour intervals, which resulted in six fixes per day (Grant, 2012).

Home Range Tools (HRT) extension (Rodgers, Kie, Wright, Beyer & Carr, 2015) for ArcGIS 10.5 was used to construct the home range of each collared individual. Home range estimations were determined by using minimum convex polygons (MCPs), which were generated by linking the outermost location points, and utilisation distributions (UDs), which estimated the probability distributions from a set of location points (Rodgers et al., 2015). The latter removes outlying fixes and is considered a closer estimation of total home range size, although only MCPs can be used for comparisons between study sites (Harris et al., 1990). Therefore, we report the 95% and 100% MCPs and the 90% and 95% UDs to determine the home range of lions and leopards. To estimate the core area of monitored individuals, both 50% MCPs and 50% and 75% UDs were used. The 50% UD is generally considered the most robust estimator of an individual’s core area (Harris et al., 1990; Mizutani & Jewell, 1998).

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Factors that influence range

Seasonal range size

Environmental factors are known to affect seasonal range sizes in large predators, particularly in large, open systems, where predators track migratory herds (Loveridge et al., 2009; Tuqa et al., 2014). In small, enclosed reserves, Hunter (1998) and Kilian (2003) noted that large felids tended to decrease their range sizes by congregating around waterpoints during the dry season, as prey become increasingly dependent on water. MWR is a small, fenced reserve with two perennial rivers, natural springs, artificial waterholes and a high prey density. Hence, we assume that seasonal ranges of lions and leopards are likely to follow similar trends. In our study, we divided seasonal ranges into a wet (December – May) and dry season (June – November) and t-tests were used to determine if there were significant differences between seasons. All statistical analyses were performed using Statistica version 13.2 (Dell Software, 2016).

Female with cubs

During lactation, female lions may decrease range sizes, as they tend to their cubs (Sandell, 1989). Cubs typically remain nearby their den sites, for the first three months, thereafter they start following their mother (Hunter, 1998; Kilian, 2003). Thus, the home range and core area of the collared female lion was determined within the first three months after giving birth to her cubs.

Roaming

Large felids are known to roam widely after translocation (van der Meulen, 1997; Weilenmann, Gusset, Mills, Gapanapelo & Schiess-Meier, 2010) and may occasionally escape through reserve fences. To determine whether lions or leopards moved outside the reserve, all location points for each felid were plotted in ArcGIS 10.5. Location points falling outside the reserve boundary were given a directional path, using a straight line between points, which was the minimum estimate (Merrill & Mech, 2003). The total distance for each path was calculated in HRT extension for ArcGIS.

Habitat selection

Habitat selection was based on data from Sherry (1989), which included fine-scale habitat types that were likely to be used by large felids (e.g. Spong, 2002). A 100% MCP was constructed to determine habitat availability within each animal’s home range. This method included all location points so that individual animals had at least some knowledge of the habitat types that were not utilised (Hunter, 1998). A chi-square goodness-of-fit analysis was used to test if the observed values (the number of location points within each habitat type) differed significantly from the expected values, which was the proportional area of a habitat within an animal’s home range (Byers & Steinhorst, 1984; Hunter, 1998; Kilian, 2003; Neu et al., 1974). This test was conducted to determine whether reintroduced felids

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were observed more often than expected in a habitat type and data were tested for significance at the 5% level (α = 0.05).

Bonferroni simultaneous 95% confidence intervals were computed using the equation adapted from Neu et al. (1974):

(1 − 푃푖) (1 − 푃푖) 푃푖 − 푍√푃푖 ≤ 푃푖 ≤ 푃푖 + 푍√푃푖 푛 푛

Where Pi is the proportional use of a habitat type, Z (1-α/2k) is the upper standard distribution figure consistent with the tail area of α/2k, n is the total number of location points, α is the significance level at 5% and k is the number of available habitats types. The proportions of each habitat type were compared to the 95% confidence intervals to determine preference, avoidance or no selection of habitat types (Byers & Steinhorst, 1984). Proportions below the lower limit values of a particular habitat type indicated preference, proportions above the upper limit values indicated avoidance and proportions within this confidence interval values showed no selection of a particular habitat type.

A Jacobs’ index (Jacobs, 1974) was also used to determine habitat selection for lions and leopards: 푟 − 푝 퐷 = − 2푟푝 푟 + 푝

Where D is the Jacobs’ index, r is the observed use and p is the expected use. Values range from +1 to –1, with +1 indicating maximum selection of a particular habitat type in proportion to its availability and –1 showing maximum avoidance of a particular habitat use in proportion to its availability. In addition, values greater than 0.5 were considered strongly preferred and values smaller than –0.5 were considered strongly avoided, while values close to 0 were defined as “no selection”.

Range overlap

Home range overlap was calculated within and between species. Spatial overlap was calculated as the percentage overlap between two individuals (A and B) using the equation (Cristescu et al., 2013):

100 × (2 × 퐴푟푒푎 표푓 표푣푒푟푙푎푝)⁄ 푂푣푒푟푙푎푝 (%) = 퐴푟푒푎 퐴 + 퐴푟푒푎 퐵

We used HRT for ArcGIS and ArcToolbox in ArcGIS 10.5 to compare overlap of home range (95%) and cores area (50%) for both MCP and UD methods.

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4.4 Results

Lion

Home range

Home range size estimates for monitored lions are displayed in Figures 4.2 and 4.3. A detailed report of lion home ranges is provided in Table 4.1. Mean ± SE home range of lions was 380.45 ± 117.70 km2 using 100% MCP and 134.00 ± 6.66 km2 using 95% UD. The male lion (LIM1) had a home range size (95% UD) of 140.65 km2 and a core area size (50% UD) of 32.27 km2. The home range size (95% UD) of the female lion (LIF1) was 127.34 km2 and her core area size (50% UD) was 25.13 km2.

Figure 4.2. Kernel utilisation distribution (95 %, 90 %, 75% and 50%) area estimation for LIM1 in Majete Wildlife Reserve, Malawi.

Figure 4.3. Kernel utilisation distribution (95 %, 90 %, 75% and 50%) area estimation for LIF1 in Majete Wildlife Reserve, Malawi.

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Table 4.1. Home range size (km2) of reintroduced lions and leopards in Majete Wildlife Reserve, Malawi, as calculated from the kernel utilisation distribution (UD) and minimum convex polygon (MCP) methods during the study period from August 2012 to May 2017. Time taken for individuals to establish home ranges in the reserve and distance between the home range centroid of an individual and their release site (km) are indicated.

MCP (km2) UD (km2) Fixes (HR HR distance to ID Code Sex Study period estimation) release site (km) 100% 95% 50% 95% 90% 75% 50%

Lion LIM1 M Aug 2012 – May 2017 2791 4.42 498.15 155.64 54.32 140.65 113.22 70.69 32.27 LIF1 F Aug 2012 – May 2017 2913 4.26 262.75 149.17 46.04 127.34 104.83 60.22 25.13 4.34 380.45 152.41 50.18 133.96 109.03 65.46 57.40 Leopard LEM1 M Oct 2011 – Nov 2012 2200 7.51 679.35 516.17 185.30 349.96 247.52 155.46 64.81 LEM2 M Jan 2012 – Jan 2013 1762 12.00 431.14 365.07 222.47 213.48 183.65 73.19 17.48 LEM3 M Dec 2012 – Dec 2013 594 9.27 684.10 589.48 219.08 413.12 413.12 189.03 83.66 9.59 598.20 490.24 208.95 325.52 281.43 139.23 55.32 LEF1 F Oct 2011 – Nov 2012 2150 6.12 271.25 206.64 65.74 162.52 124.90 67.78 26.39 LEF2 F Jan 2012 – Dec 2012 1793 5.24 409.55 253.77 83.58 150.61 113.91 52.10 19.25 LEF3 F Dec 2012 – Feb 2013a ------5.68 340.40 230.21 74.66 156.57 119.41 59.94 22.82 8.03 495.08 386.23 155.23 257.94 216.62 107.51 42.32

aData were insufficient for analyses

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Factors influencing range

Females with cubs

Details for the range size of LIF1 with her young cubs is presented in Table 4.2. The female had her first litter in July 2013, and her home range (95% UD) was 24.61 km2, with a core range (50% UD) of 2.98 km2. During her second litter born September 2015, her home range was 21.43 km2, with a small core area of 2.18 km2. The mean home range was 23.02 ± 1.59 km2 and mean core area was 2.58 ± 0.40 km2. Data from her third litter were not recorded here as it fell outside of the study period.

Table 4.2. Range size (km2) change of lioness LIF1 for the first three months following the birth of her cubs. This was compared with the total range size for the same lioness during the entire study period from 2012 to May 2017 in MWR. Number of Number of UD Date Fixes litters cubs 95% 75% 50%

Total Range - - 2663 130.78 64.52 30.22 LIF1 with cubs 1 2 276 24.61 9.59 2.98 LIF1 with cubs 1 2 248 21.43 7.08 2.18

Seasonal range size

Seasonal home ranges for reintroduced lions in MWR appear in Table 4.3. No seasonal home ranges were provided for the dry season of 2012 as individuals were yet to establish home ranges in the reserve. In addition, no ranges were presented for the wet season of 2013/14 or the dry season of 2014 due to persistent location failure. Seasonal home range (95%) size of LIM1 decreased from 102.58 ± 14.58 km2 in the wet season to 81.76 ± 2.17 km2 in the dry season, however this difference was not significant (T-test, t = 1.20, d.f. = 5, p = 0.28). Core area (50%) for LIM1 shrank from 24.05 ± 4.09 km2 in the wet season to 15.87 ± 0.97 km2 in the dry season (t = 1.67, d.f. = 5, p = 0.16). Similarly, the seasonal home range (95%) of LIF1 decreased from 92.92 ± 18.53 km2 in the wet season to 69.89 ± 23.94 km2 in the dry season (t = 0.08, d.f. = 5, p = 0.47). Seasonal core area (50%) size was 23.62 ± 5.97 km2 in the wet season and 10.96 ± 4.16 km2 in the dry season, which was not significantly different (t = 1.61, d.f. = 5, p = 0.17).

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Table 4.3. Seasonal home range sizes (km2) of reintroduced lions in Majete Wildlife Reserve using kernel utilisation distribution (UD) method. Values in parentheses indicate number of location fixes. Overall home range

ID Code Wet 2012/13 Dry 2013 Wet 2014/15 Dry 2015 Wet 2015/16 Dry 2016 Wet 2016/17 (95% UD) Wet Dry

Male LIM1 95% 82.55 (581) 78.19 (520) 120.63 (198) 81.41 (179) 73.38 (179) 85.67 (432) 133.74 (502) 75% 41.03 (581) 33.95 (520) 60.73 (198) 40.41 (179) 39.06 (179) 36.79 (432) 75.71 (502) 50% 18.12 (581) 14.16 (520) 26.51 (198) 17.50 (179) 17.02 (179) 15.95 (432) 34.53 (502) 102.58 81.76 Female LIF1 95% 75.30 (492) 29.72 (324) 115.45 (250) 112.55 (419) 49.90 (311) 67.41 (395) 131.03 (472) 75% 37.61 (492) 12.07 (324) 63.43 (250) 47.74 (419) 24.89 (311) 30.52 (395) 72.41 (472) 50% 17.22 (492) 4.20 (324) 29.86 (250) 18.55 (419) 10.50 (311) 10.12 (395) 36.89 (472) 92.92 69.89

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Leopard

Home range

Home range size estimates for all monitored leopards (except LEF3, see Table 4.1) are presented in Figures 4.4 – 4.8. Leopard home range size (95% UD) varied from 150.61 km2 for LEF2 to 413.12 km2 for LEM3. Mean home ranges were 495.08 ± 80.99 km2 using 100% MCP and 257.94 ± 52.51 km2 using 95% UD. Male range sizes were larger than those of females, but only significantly larger using 50% MCP (t = 8.04, d.f. = 3, p = 0.004). Home range size averaged 156.57 ± 5.96 km2 (n = 2, range = 150.61– 162.52 km2) for females and 325.52 ± 58.91 km2 (n = 3, range = 213.48–413.12 km2) for males. Mean core area (50% UD) was 22.82 ± 3.57 km2 (n = 2) for females and 55.32 ± 19.69 km2 (n = 3) for males.

Figure 4.4. Kernel utilisation distribution (95%, 90%, 75% and 50%) area estimation for LEM1 in Majete Wildlife Reserve, Malawi.

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Figure 4.5. Kernel utilisation distribution (95%, 90%, 75% and 50%) area estimation for LEM2 in Majete Wildlife Reserve, Malawi.

Figure 4.6. Kernel utilisation distribution (95%, 90%, 75% and 50%) area estimation for LEM3 in Majete Wildlife Reserve, Malawi.

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Figure 4.7. Kernel utilisation distribution (95%, 90%, 75% and 50%) area estimation for LEF1 in Majete Wildlife Reserve, Malawi.

Figure 4.8. Kernel utilisation distribution (95%, 90%, 75% and 50%) area estimation for LEF2 in Majete Wildlife Reserve, Malawi.

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Factors influencing range

Seasonal range size

Seasonal home ranges for reintroduced leopards in MWR appear in Table 4.4. No seasonal home ranges were provided for individuals for the first season after their release, as individuals were yet to establish themselves in the reserve. Mean seasonal home range for male leopards was 116.02 ± 11.81 km2 in the wet season compared to 276.89 ± 72.88 km2 in the dry season, which was not significantly different (t = –2.18, d.f. = 4, p = 0.09). For male leopards, mean core area ranges increased from 16.01 ± 1.20 km2 in the wet season to 50.53 ± 21.69 km2 in the dry season (t = –1.59, d.f. = 4, p = 0.19). For female leopards, mean home range increased from 74.15 ± 53.07 km2 in the wet season to 143.28 ± 17.43 km2 in the dry season (t = –1.24, d.f. = 2, p = 0.34). Mean core area size for females was 8.07 ± 4.15 km2 in the wet season compared to 18.70 ± 0.30 km2 in the dry season, however this was not significantly different (t = –2.56, d.f. = 2, p = 0.13).

Table 4.4. Seasonal home range sizes (km2) of reintroduced leopards in Majete Wildlife Reserve using kernel utilisation distribution (UD) method. Values in parentheses indicate the number of location fixes.

Overall home range

ID Code Wet 2011/12 Dry 2012 Wet 2012/13 Dry 2013 (95% UD) Wet Dry Males LEM1 95% 125.85 (797) 212.26 (918) 75% 41.85 (797) 87.99 (918) 50% 16.80 (797) 33.70 (918) LEM2 95% 196.06 (958) 129.69 (117) 75% 82.69 (958) 48.81 (117) 50% 24.33 (958) 17.59 (117) LEM3 95% 92.51 (92) 422.36 (422) 75% 52.98 (92) 179.22 (422) 50% 13.65 (92) 93.57 (422) 116.02 276.89

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Table 4.4. Continued.

Females LEF1 95% 127.22 (739) 160.70 (930) 75% 46.86 (739) 56.33 (980) 50% 12.21 (739) 18.99 (980) LEF2 95% 125.85 (977) 21.08 (34) 75% 47.62 (977) 10.77 (34) 50% 18.40 (977) 3.92 (34) 74.15 143.28

Roaming behaviour

Movements of four leopards outside of MWR are presented in Figures 4.9 – 4.12. LEM1 escaped the reserve on two occasions, once in the north (13.95 km) for three days, and once in the south (17.85 km) for two days while the predator-proof fencing was being repaired. LEM2 also escaped twice. The first distance totalled 14.45 km, which occurred directly after LEM1 escaped in the same area and the second totalled 17.85 km towards the west. LEM3 roamed the furthest of all reintroduced felids. During his first escape, the male moved south, through a hole in the fence into Lengwe National Park, Malawi, and returned after four days, travelling a total of 33.20 km. The male then escaped two months later in the north-west and travelled for 37 days (197.55 km), traversing villages, agricultural areas and relatively close to the city of Blantyre, before returning to the reserve in the north-east. LEF2 was the only female to escape, travelling a mere 5.84 km and returning to the reserve after two days.

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Figure 4.9. Roaming behaviour of LEM1 outside Majete Wildlife Reserve, Malawi. Dots indicate location points and arrow lines indicate the directional movement outside of the reserve.

Figure 4.10. Roaming behaviour of LEM2 outside Majete Wildlife Reserve, Malawi. Dots indicate location points and arrow lines indicate the directional movement outside of the reserve.

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Figure 4.11. Roaming behaviour of LEM3 outside Majete Wildlife Reserve, Malawi. Dots indicate location points and arrow lines indicate the directional movement outside of the reserve.

Figure 4.12. Roaming behaviour of LEF2 outside Majete Wildlife Reserve, Malawi. Dots indicate location points and arrow lines indicate the directional movement outside of the reserve.

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Lion and leopard

Range overlap

Home range overlap of lion and leopard individuals is displayed in Tables 4.5 (UD) and 4.6 (MCP). Overall, both methods revealed noticeable range overlap for home range and core areas, although MCPs were generally higher than UDs. As expected, the two lions had the highest overlap of home range (95.58%) and core area (89.99%) of all reintroduced felids. Lion and leopard ranges had a mean overlap of 42.25 ± 7.12% (95% UD) and 8.38 ± 2.48% (50% UD). For leopards, core area overlap was largest for LEF1 and LEF2 (73.81%) of the 50% UDs. Male and female overlap averaged 35.35 ± 6.43% for 95% UD and 17.83 ± 10.65% for 50% UD. Male leopard home range overlap averaged 38.05 ± 8.39%, although only the core area of LEM1 and LEM3 overlapped (10.40%).

Table 4.5. Area of overlap (km2) for the home range (95 %) and core area (50 %), using kernel utilisation distribution (UD), for reintroduced lions and leopards in Majete Wildlife Reserve. Values above the diagonal represent home range overlap, while values below the diagonal represent core area overlap. Percentage spatial overlap is indicated in parentheses. Overlap of lion (LIM1 & LIF1) range with those of leopard (LEM1–3 and LEF1–2) are highlighted in grey.

ID Code LIM1 LIF1 LEM1 LEM2 LEM3 LEF1 LEF2

123.03 124.24 24.83 55.39 98.88 88.25 LIM1 - (91.93) (51.04) (14.17) (20.14) (66.05) (61.39) 22.91 118.73 25.64 49.81 94.97 87.67 LIF1 - (73.77) (49.39) (14.90) (18.32) (64.76) (62.31) 3.25 4.48 85.13 209.11 138.33 141.47 LEM1 - (6.72) (9.43) (30.22) (54.81) (53.98) (56.52) 0.84 0.72 0.00 91.22 54.82 50.67 LEM2 - (3.40) (3.02) (0.00) (29.12) (29.16) (27.83) 0.00 0.00 7.72 0.00 71.17 56.06 LEM3 - (0.00) (0.00) (10.40) (0.00) (24.73) (19.89) 5.99 6.44 13.37 0.01 0.00 96.68 LEF1 - (20.56) (22.75) (29.32) (0.05) (0.00) (61.75) 2.18 2.33 5.03 12.05 0.00 4.40 LEF2 - (8.53) (9.42) (11.97) (65.61) (0.00) (19.28)

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Table 4.6. Area of overlap (km2) for the home range (95 %) and core area (50 %), using minimum convex polygon (MCP), for reintroduced lions and leopards in Majete Wildlife Reserve. Values above the diagonal represent home range overlap, while values below the diagonal represent core area overlap. Percentage spatial overlap is indicated in parentheses. Overlap of lion (LIM1 & LIF1) range with those of leopard (LEM1–3 and LEF1–2) are highlighted in grey.

ID Code LIM1 LIF1 LEM1 LEM2 LEM3 LEF1 LEF2

145.67 155.64 113.60 155.64 139.21 153.29 LIM1 - (95.58) (44.34) (40.63) (10.59) (75.24) (72.20)

45.16 149.15 112.32 149.17 136.76 147.30 LIF1 - (89.99) (42.88) (40.64) (10.17) (75.23) (70.45)

20.89 15.45 335.35 546.43 208.69 257.08 LEM1 - (17.06) (13.06) (70.60) (32.82) (54.86) (63.05)

31.50 30.44 79.46 403.52 185.62 206.61 LEM2 - (22.63) (22.54) (38.33) (25.33) (60.08) (61.44)

54.32 46.04 190.60 212.96 214.40 269.01 LEM3 - (17.48) (15.01) (50.30) (53.83) (14.31) (17.63)

29.14 28.12 25.82 63.72 65.74 193.18 LEF1 - (48.55) (50.32) (20.14) (43.98) (20.77) (79.92)

49.39 45.67 28.34 65.53 84.01 55.27 LEF2 - (71.41) (70.24) (20.64) (42.55) (25.80) (73.81)

Habitat selection

Habitat selection for reintroduced felids appear in Tables 4.7 (lions) and 4.8 (leopards). All felids showed preferences for particular habitat types. The lion pride selected riverine associations (D = 0.51; D = 0.49) and riparian thicket (D = 0.44; D = 0.64), while avoiding high-altitude mixed woodland (D = – 0.91; D = –0.95). Similarly, all leopards selected riverine associations (mean D = 0.54) and avoided high- altitude mixed woodland (mean D = –0.46). Leopards also generally displayed selection for low-altitude mixed woodland, which made up 39.08% of the habitats in the reserve.

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Table 4.7. Habitat use by reintroduced lions in Majete Wildlife Reserve, Malawi. Values in bold are significant.

Percentage Jacobs’ Bonferroni Habitat types per individual Observeda Expectedb Preferencee χ2 resultsf habitat available indexc 95% CId LIM1 High-altitude miombo woodland 28.97 18 340.41 -0.91 0.00−0.01 − Medium-altitude mixed woodland 14.33 603 845.16 -0.20 0.14−0.17 − χ2 = 2137.94, Low-altitude mixed woodland 39.08 2 041 2038.09 0.00 0.50−0.55 0 d.f. = 5, p < Ridgetop mixed woodland 5.30 209 277.89 -0.15 0.04−0.06 − 0.001 Riparian thicket 0.69 139 55.29 0.44 0.03−0.04 + Riverine 11.63 882 335.17 0.51 0.21−0.24 + LIF1

High-altitude miombo woodland 28.97 27 795.68 -0.95 0.00−0.01 − Medium-altitude mixed woodland 14.33 539 822.85 -0.25 0.13−0.16 − 2 Low-altitude mixed woodland 39.08 2 024 1654.98 0.19 0.51−0.55 + χ = 1410.72, d.f. = 5, p < Ridgetop mixed woodland 5.30 271 211.60 0.13 0.06−0.08 + 0.001 Riparian thicket 0.69 159 35.60 0.64 0.03−0.05 + Riverine 11.63 826 325.29 0.49 0.20−0.23 +

aThe observed value indicates the actual number of location fixes for each habitat type bThe expected value was determined by using the percentage of each habitat type available, derived from the 100% minimum convex polygon of each individual cJacobs’ preference index (Jacobs, 1974) dBonferroni 95% simultaneous confidence intervals (CI) corrections (α = 0.05, k = 6, Z = 2.64) e Significant preference (+), avoidance (−) and no significant selection (0) of habitat type fChi-square goodness-of-fit: significant results in bold (5% significance level)

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Table 4.8. Habitat use by reintroduced leopards in Majete Wildlife Reserve, Malawi. Values in bold are significant.

Percentage Jacobs’ Bonferonni Habitat types per individual Observed Expected Preference χ2 results habitat available index 95% CI LEM1 High-altitude miombo woodland 28.97 281 936.91 -0.63 0.08−0.11 − Medium-altitude mixed woodland 14.33 477 470.38 0.01 0.14−0.18 0 χ2 = 1372.15, Low-altitude mixed woodland 39.08 1 425 1194.22 0.15 0.45−0.50 + d.f. = 5, p < Ridgetop mixed woodland 5.30 155 173.62 -0.06 0.04−0.06 0 0.001 Riparian thicket 0.69 24 22.63 0.03 0.00−0.01 0 Riverine 11.63 655 219.23 0.56 0.20−0.24 + LEM2 High-altitude miombo woodland 28.97 855 1083.11 -0.21 0.37−0.42 − Medium-altitude mixed woodland 14.33 67 407.57 -0.76 0.02−0.04 − χ2 = 1075.93, Low-altitude mixed woodland 39.08 823 466.06 0.38 0.35−0.41 + d.f. = 5, p < Ridgetop mixed woodland 5.30 13 45.88 -0.56 0.00−0.01 − 0.001 Riparian thicket 0.69 9 20.44 -0.39 0.00−0.01 − Riverine 11.63 405 148.98 0.51 0.16−0.21 + LEM3 High-altitude miombo woodland 28.97 114 228.39 -0.42 0.11−0.18 − χ2 = 193.04, p Medium-altitude mixed woodland 14.33 64 121.15 -0.35 0.06−0.11 − < 0.000, d.f. = Low-altitude mixed woodland 39.08 411 317.48 0.24 0.48−0.58 + 5

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Ridgetop mixed woodland 5.30 61 44.27 0.17 0.05−0.10 0 Riparian thicket 0.69 3 5.84 -0.32 0.00−0.01 0 Riverine 11.63 120 55.89 0.40 0.12−0.19 + LEF1 High-altitude miombo woodland 28.97 398 630.08 -0.27 0.12−0.15 − Medium-altitude mixed woodland 14.33 659 562.19 0.10 0.21−0.25 + χ2 = 1684.26, Low-altitude mixed woodland 39.08 1 076 1318.55 -0.17 0.35−0.39 − d.f. = 5, p < Ridgetop mixed woodland 5.30 49 186.53 -0.60 0.01−0.02 − 0.001 Riparian thicket 0.69 54 41.26 0.14 0.01−0.03 0 Riverine 11.63 679 176.39 0.65 0.21−0.25 + LEF2 High-altitude miombo woodland 28.97 126 719.75 -0.77 0.04−0.06 − Medium-altitude mixed woodland 14.33 518 552.71 -0.04 0.18−0.22 − χ2 = 1634.46, Low-altitude mixed woodland 39.08 1 139 964.73 0.14 0.41−0.46 + d.f. = 5, p < Ridgetop mixed woodland 5.30 70 124.78 -0.29 0.02−0.04 0 0.001 Riparian thicket 0.69 54 32.52 0.25 0.01−0.03 0 Riverine 11.63 709 221.52 0.60 0.25−0.29 + aThe observed value indicates the actual number of location fixes for each habitat type bThe expected value was determined by using the percentage of each habitat type available, derived from the 100% minimum convex polygon of each individual cJacobs’ preference index (Jacobs, 1974) d Bonferroni 95% simultaneous confidence intervals (CI) corrections (α = 0.05, k = 6, Z = 2.64) eSignificant preference (+), avoidance (−) and no significant selection (0) of habitat type fChi-square goodness-of-fit: significant results in bold (5% significance level)

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4.5 Discussion

Home range use

Lions had relatively large range sizes, comparable to those found in Savuti, Botswana (McBride, 1990; Viljoen, 1993) and Hwange National Park, Zimbabwe (see Hemson, 2003). These range sizes remained stable from their release in 2012 until the end of the monitoring period in 2017. Home range (95% UD) of LIM1 (140.65 km2) was slightly larger than that of LIF1 (127.34 km2), as both individuals maintained occupancy in the north of the reserve. Reintroduced female lions have been found to establish small, exclusive ranges in four South African reserves (Druce et al., 2004; Hayward et al., 2009; Hunter, 1998; Kilian, 2003; Lehmann, 2007). Based on a single lioness, the home range size of LIF1 is the largest on record for any reintroduced lioness. MWR is larger in comparison to the above-mentioned reserves, which suggests that reserve size may influence range size of reintroduced lions. In addition, the single pride in MWR lacked interspecific competition experienced in other enclosed reserves (Druce et al., 2004; Hayward et al., 2009; Hunter, 1998; Kilian, 2003), which could also explain the large ranges observed in our study. Therefore, to determine whether established lions respond to increased conspecific density, future studies should continue monitoring this pride after the translocation of additional lions in the reserve.

As expected, male leopards had larger home ranges than those of females, which is consistent with other studies (Bothma et al., 1997; Bailey, 1993; Mizutani & Jewell 1998; Stander et al., 1997). Leopards in MWR had the largest home ranges of any reintroduced leopards on record, even comparable to those from the semi-arid Kalahari Gemsbok National Park, South Africa (Bothma et al., 1997) and Kaudom Game Reserve in Namibia (Stander et al., 1997). The larger leopard ranges found in more arid regions are due to lower conspecific and prey densities. MWR is a relatively prey-rich reserve with few conspecifics, thus we assume that intraspecific density was the factor causing the large ranges in our study. If this the case, we could expect similar findings of spatial organisation to those documented in a recovering leopard population in Phinda Game Reserve. Fattebert et al. (2016) found that an increase in leopard density resulted in the contraction of home ranges and formation of matrilineal kin clusters (evident from higher spatial overlap) for females, whereas males maintained large home ranges (not tracking female range contractions) and showed lower levels of spatial overlap (Fattebert et al., 2016).

Very little data exist on home ranges of reintroduced leopards. Leopards are notoriously difficult to monitor after their release (Hayward, Adendorff, Moolman, Hayward & Kerley, 2006), especially when VHF collars (which require active tracking of the animal) are used instead of GPS/GSM collars (which transmit location data directly via the GSM network) in dense or mountainous environments (Martins, Horsnell, Titus, Rautenbach & Harris, 2011; personal observation). Hayward et al. (2009) estimated the

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range of a single reintroduced male leopard at 32 km2 in Addo Elephant National Park, South Africa, while Cristescu et al. (2013) estimated a range of 11.5 km2 for a single female brought into Shamwari Game Reserve, South Africa. Although, both range estimates are noticeably smaller than those from this study, caution must be used when comparisons are based on such small sample sizes. Unfortunately, research on reintroduced predators has typically been fragmented and ad hoc (Seddon, Armstrong & Maloney, 2007), which may be attributed to a host of factors (see Hayward et al., 2007). Therefore, this study presents valuable information from a research and management point of view of the home range sizes of reintroduced leopards in a small, enclosed environment.

Factors influencing range

Seasonal ranges

Seasonal variation in lion range size has been documented throughout Africa. In large, open systems range expansion may occur during the dry season when resources are limited and lions increase their ranges in search for food (Viljoen, 1993) or the wet season when lions track migratory herds that move outside or away from their resident reserve (Tuqa et al., 2014; Tumenta et al., 2011). However, in small, enclosed reserves, sedentary prey usually congregate around rivers and artificial waterholes during the dry season, and lions subsequently decrease their ranges by concentrating around these waterpoints (de Boer et al., 2010; Druce et al., 2004; Hunter, 1998). Similarly, lions in MWR displayed smaller ranges during the dry season and tended to congregate around available water sources (personal observation). In addition, the supplementation of artificial waterholes has been shown to affect ungulate movement (Ephaphras et al., 2007; Owen-Smith, 1996), which in turn influences lion range. Therefore, these waterpoints can be manipulated by management to enhance the persistence of both predator and prey populations or increase the impact of predators in specific areas of a reserve.

Roaming

Four leopards ‘escaped’ from the reserve, three of which were males. The furthest linear distance recorded in this study was 198 km travelled within 37 days by LEM3 (a stock raider and fence breaker). Although greater distances have been recorded (Weilenmann et al., 2010), these individuals were still likely to encounter human-induced mortality beyond the reserves’ boundary (i.e. edge effects; Balme et al., 2010; Woodroffe & Ginsberg, 1998). Roaming movements, particularly those of LEM3, were traced through agricultural lands and villages, and we assume livestock depredation occurred. ‘Problem animals’ are not recommended for translocation, largely due to their wide roaming behaviour outside reserve boundaries and continuation of stock predation at the site of release (Athreya, Odden, Linnell & Karanth, 2011; Stander, 1990; Weilenmann et al., 2010). Clearly, it is important to consider an animal’s history prior to translocation and although LEM3 did establish a

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range within MWR, he also displayed extensive roaming outside the reserve. Therefore, we recommend that problem leopards should not be translocated into MWR.

It seems improbable that roaming behaviour of leopards was to avoid interspecific competition with lion and spotted hyena (Crocuta crocuta). When leopards are translocated into areas with resident conspecifics, the intruders are likely to show negligible release site fidelity and roam extensively in search of vacant areas (Weilenman et al., 2010; Weise et al., 2015). Releasing leopards into areas with a high density of conspecifics is not recommended (Athreya et al., 2011; Hamilton, 1981; Weilenmann et al., 2010; Weise et al., 2015) as it may lead to escapes (see Hayward et al., 2007). However, when densities are far below carrying capacity, additional translocations may be necessary to assist population growth and increase the genetic pool (Armstrong & Seddon, 2008). Weise et al. (2015) suggested that additional translocations should mimic the natural sub-adult dispersion of leopards, with a conservative 18-month inter-release interval (Fattebert, Dickerson, Balme, Slotow & Hunter, 2013), to provide an adjustment period of resident individuals to the newcomer (similar for lions, see Ferreira & Hofmeyr, 2014). Therefore, management of MWR should carefully consider the ranges of resident individuals and locate suitable release sites, with low conspecific density, to minimise roaming behaviour and improve the chances of translocation success (Hayward et al., 2007; Weise et al., 2015).

Range overlap

Both the UD and MCP methods recorded range overlap within and across reintroduced felids, which was expected based on fence boundaries. Lions overlapped extensively with both male and female leopards within MWR, suggesting some degree of spatial co-existence between guild members. However, lions are more dominant that leopards and both compete for similar resources, thus we expect leopards to exhibit strategies to reduce competition with lions. Leopards are highly adaptable predators (by means of their diet and behaviour) and are thus capable of competing for resources despite spatial overlap with dominant predators (Karanth & Sunquist, 2000). Leopard avoidance strategies of larger predators include variation in temporal patterns, habitat selection, prey specialisation and caching kills in trees to minimise kleptoparasitism (Karanth & Sunquist, 2000; Stein, Bourquin & Krause, 2014).

The inter- and intrasexual area overlaps for reintroduced leopards exceeded the 10% threshold thought to show exclusivity (Sandell, 1989), except for the core area of half of the pairings (Table 4.7). The relatively high levels of range overlap suggest that reintroduced leopards rarely utilise active territorial defence to exclude conspecifics from their ranges, due to their large home range sizes (Marker & Dickman, 2005); spatial competition can also be relaxed when resources are relatively abundant (Maher & Lott, 2000). However, this do not imply intraspecific tolerance. For example, previous studies have found strong mutual avoidance between leopards and a combination of scent-

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marking and vocalisation serve as indirect signals for temporal segregation of conspecifics (Bailey, 1993; Bothma & le Riche 1984; Hamilton, 1981; Mizutani & Jewell, 1998).

Habitat selection

Habitat use of large predators is typically defined by their requirements for suitable prey, adequate water supply, mating opportunities and sufficient cover for successful hunting (Bailey, 1993; Hanby et al., 1995; Schaller, 1972; Sandell, 1989). Lions strongly selected riparian thicket and riverine areas, despite constituting only 0.69% (riparian thicket) and 11.63% (riverine) of the available habitat in MWR. In Phinda Game Reserve, South Africa, lions selected riparian forest more than expected when it occurred in their range (Hunter, 1998). Riparian thicket is a relatively dense vegetation which might provide sufficient cover while hunting, as well as shade during the heat of the day. Similarly, Spong (2002) found that lions utilised riverine areas in the Selous Game Reserve, Tanzania. Lions are ambush predators and their selection for riverine habitats may be that water provides foreseeable sites for encountering prey, particularly during the dry season when prey congregate around water sources. As a result, Hopcraft et al. (2005) found that lions select habitats with a high prey ‘catchability’ i.e. habitats where prey are easiest to capture, and not necessarily habitats with the highest prey densities. Since lions typically select high-quality habitats (Mosser & Packer, 2009), we assume that riverine habitat and riparian thicket are optimal habitat for the lions in MWR.

Previous studies found that spatial distribution of female leopards is largely driven by food supply, as well as locating optimal habitat in which to successfully hunt and raise their cubs (Bailey, 1993; Mizutani & Jewell, 1998). This study found that female leopards strongly favoured riverine association, which suggests optimal habitat selection. However, all reintroduced leopards showed a negative selection for high-altitude miombo woodland, which is situated in the rugged western region of the reserve, with a low prey density (based on the aerial survey data of 2015) and few water sources. However, two female leopards (one adult and one sub-adult) were observed scent-marking a tree in this habitat after the study period, suggesting that ranges were established in this habitat type. Like lions, Balme, Hunter and Slotow (2007) also found that leopards prefer habitats with high levels of prey catchability, above those with more abundant prey; however, instead of selecting the densest vegetation (which likely could increase detection), leopards were found to hunt more successfully in intermediate vegetation cover (cf. Pitman, Kilian, Ramsay & Swanepoel, 2013).

4.6 Conclusion

The home ranges recorded in this study are the largest for any reintroduced felid, which were comparable to ranges in more arid areas. Felids showed no significant changes in seasonal ranges, although ranges were smaller during the dry season for both lions, presumably to ambush prey that were congregated around artificial waterpoints (de Boer et al., 2010).

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The general inter-and intra-specific range overlap may be attributed to the low density and large home ranges of felids, resulting in lower levels of active territorial defence. Furthermore, all felids showed clear selection for the riverine habitat. This could be explained by the concentration of prey around available water sources during the dry season as well as prey catchability in this habitat type (Balme et al., 20007; Hopcraft et al., 2005). Clearly, leopards can co-exist with lions at a spatial scale, although we expect leopards to minimise encounters with more dominant lions, by adapting their behaviour and diet accordingly (Karanth & Sunquist, 2000).

Despite the few incidences of roaming outside the reserve, our findings indicate that reintroductions are a potentially viable tool for re-populating large predators in areas of their former distribution, provided that the reason for the initial extirpation is resolved. We advise managers to consider the home ranges of established individuals prior to translocating additional individuals, to prevent aggressive encounters close to the release site and reduce roaming behaviour, which can be managed by selecting release sites in areas with high habitat quality and low competitor densities (Weise et al., 2015).

4.7 Acknowledgements

I would like to thank Earthwatch Institute for funding this study. Thank you to Craig Hay and Gervaz Tamala for providing the access to collar data. I would also like to thank Prof Arthur Rodgers for help with operating Home Range Tools for ArcGIS and Prof Martin Kidd for help with statistics.

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Chapter 5

Dietary ecology of three apex predators in Majete Wildlife Reserve, Malawi

W.D. Briers-Louw1, A.J. Leslie1

1Department of Conservation Ecology and Entomology, Stellenbosch University, Matieland, Western Cape, 7602, South Africa

5.1 Abstract

Between 2011 and 2012, the carnivore guild in Majete Wildlife Reserve, Malawi, was restored following the reintroduction of lion (Panthera leo) and leopard (Panthera pardus). Our aim was to describe and compare the diet of lion, leopard and resident spotted hyena (Crocuta crocuta) using scat analysis (all three species) and kill site analysis derived from GPS collar data (only lion) between March 2016 and May 2017. Results indicated that lions and hyenas selected medium- to large-bodied prey species and displayed the greatest dietary overlap. Lions had a narrow actual dietary breadth and showed preference for warthog (Phacochoerus africanus), waterbuck (Kobus ellipsiprymnus), kudu (Tragelaphus strepsiceros) and Lichtenstein’s hartebeest (Alcelaphus lichtensteinii) (mean = 120.25 kg, range = 43–188 kg). Hyenas also selected warthog and waterbuck, but had the greatest preference for plains zebra (Equus quagga; mean = 149.67 kg, range = 43–218 kg), within a broad actual dietary range. In contrast, leopards occupied a dietary niche below that of lions and hyenas, by selecting small-to medium-sized prey (mean = 27.50 kg, range = 16–47 kg). Our results indicate resource partitioning among the apex predators in the reserve. It appears that leopards avoid lions, which are the most dominant apex predator, by selecting relatively smaller prey; whereas hyenas avoid lions by consuming a wider range of prey and different age classes of overlapping prey. Management is advised to continue monitoring predator-prey relationships, to prevent imbalances leading to higher interspecific competition, through a reduction in prey biomass, or rapid growth in prey populations, due to a reduction in predators.

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5.2 Introduction

Dietary studies are necessary to understand the carnivore’s role in the ecosystem (Mills, 1992) and the influence that they have on their prey populations (Owen-Smith, 2008; Owen-Smith & Mason, 2005; Radloff & du Toit, 2004). Due to their body size, large carnivores typically have high daily energy requirements (Carbone, Teacher & Rowcliffe, 2007) and will place the greatest pressure on herbivore prey populations (Grange & Duncan 2006; Grange et al., 2004; Kissui & Packer, 2004). Where prey populations are sedentary, high predator densities can potentially limit prey populations (Hirst, 1969; Mills & Shenk, 1992; Peel & Montagu, 1999; Whyte & Joubert, 1988). In contrast, fenced reserves with little or no predators may result in rapidly growing prey populations, which may negatively alter the habitat and species composition (Terborgh & Estes, 2010). Therefore, large carnivores may need to be reintroduced or supplemented to maintain ecological stability and functioning.

Africa’s large carnivore guild occurs at the top of the food chain and includes three apex predators, namely lion (Panthera leo), spotted hyena (Crocuta crocuta; hereafter hyena) and leopard (Panthera pardus). These predators exceed the threshold of 21.5kg body mass that necessitates obligate vertebrate carnivory (Carbone, Mace, Roberts & MacDonald, 1999). Consequently, all three species compete for similar food resources. However, the extent of competition is thought to be reduced by the morphological partitioning of predator body mass (Hayward & Kerley, 2008), as predators typically select large vertebrate prey near or above their own body size (Sunquist & Sunquist, 1997).

Of these three apex predators, lion and hyena are considered the fiercest rivals (Mills & Harvey, 2001) and display high levels of aggression over food when living in sympatry (Hayward & Kerley, 2008; Kruuk, 1972; Mills, 1984). They steal from each other (i.e. kleptoparasitism) and scavenge off each other’s kills (Kruuk, 1972; Schaller, 1972) because their prey is relatively difficult to capture and represents a large food resource (Wang, Tedford, van Valkenburgh & Wayne, 2004). If lion density is high, hyenas readily scavenge from lions, but when lion density is low, hyenas hunt more frequently (Cooper, 1991; Höner, Wachter, East & Hofer, 2002).

Both species are highly social and cooperative hunting strategies can improve capture success rate and allow for selection of larger prey species (Kruuk, 1972; Schaller, 1972; West & Packer 2013). Lions and hyenas typically prey on the most locally abundant medium- to large-bodied ungulates (Hayward, 2006). However, the degree of dietary overlap may be reduced by the large body size of lions and the flexible feeding strategies of hyenas (Hayward & Kerley, 2008). This explains why lions prefer prey in the 190–550 kg range (Hayward & Kerley, 2005) and hyenas prefer relatively smaller prey in the 56– 182 kg range (Hayward, 2006).

Unlike lions and hyenas, leopards are mostly solitary and typically hunt alone. They have a remarkable adaptability, both in terms of their behaviour and diet (Karanth & Sunquist, 2000). For example, Bailey

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(1993) reported 92 species in leopard diet in sub-Saharan Africa, ranging in size from arthropods (Fey, 1964) to adult male eland (Kingdon, 1977). They require 1.6–4.9kg of meat per day to maintain body condition (Bailey, 1993; Bothma & le Riche, 1986; Stander, Haden, Kaqece & Ghau, 1997) and achieve this by killing between 40 and 60 prey items annually (Bailey, 1993; le Roux & Skinner, 1989; Schaller, 1972).

Across sub-Saharan Africa there is some degree of dietary overlap between leopards, lions and hyenas (Hayward & Kerley, 2008) and leopards may lose 5–10% of their kills to these more dominant species (Bertram, 1979; Schaller, 1972; Stander et al., 1997). However, leopards may reduce potential conflict with lions and hyenas through mechanisms of resource partitioning (Bertram, 1979). This is achieved by specialising on specific prey (Stein, Bourquin & McNutt, 2014), selecting relatively smaller prey (Bertram, 1982; Hayward et al., 2006) and caching prey in trees (Balme, Robinson, Pitman & Hunter, 2017). As a result, leopards can live in sympatry with more dominant competitors and often thrive in these multi-predator systems (Bailey, 1993; Hayward & Kerley, 2008).

The feeding ecology of lion, leopard and hyena has been studied extensively in southern and East Africa and provides a good understanding of their preferred prey species and weight range (Hayward, 2006; Hayward et al., 2006; Hayward & Kerley, 2005). However, local studies are still required to understand the impacts of regional environmental factors and prey availability on predator diet. Thus, the aim of this study was to compare the dietary ecology of two reintroduced felids and a resident hyenid in Majete Wildlife Reserve, Malawi, using scat analysis, while global positioning system (GPS) cluster analysis was also used for lion diet. This is the first comparative study conducted on the diet of lion, leopard and hyena in Malawi and this information will contribute towards the management of predator and prey populations within the reserve and ensure their future persistence.

5.3 Methods

5.3.1 Study area

Majete Wildlife Reserve (700 km2) is situated at the southern tip of the Great Rift Valley in the Lower Shire Valley region of southern Malawi. The eastern boundary of the reserve is defined by the Shire River which is the largest perennial river in Majete. The Mkulumadzi River is the only other perennial river which transects Majete and finds its confluence with the Shire River in the north-east of the reserve. Majete has two distinct seasons: a wet season from December to May and a dry season from June to November. Mean annual precipitation is 680–1 000 mm, with the average temperatures increasing from 23.3°C in the cooler winter months to 28.4°C in the warmer summer months (Wienand, 2013). The rugged western region is covered by tall Miombo woodland above 400 m (Brachystegia and Julbernadia spp.). Altitude decreases eastwards, and the vegetation transitions into medium altitude (250–400 m) mixed woodland, (Brachystegia boehmii, Combretum spp. and Diiospyrus kirkii) in the 118

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centre of the reserve. The eastern region is relatively flat with a few rocky outcrops and covered by mixed woodland below 250 m (Acacia spp., Combretum spp. and Sclerocarya caffra) (Staub, Binford & Stevens, 2013).

Majete was declared a game reserve in 1955, although by the early 2000s most large mammalian stocks were depleted. Only a few herbivore species (each less than 20 individuals) and a small hyena population remained in the reserve. In 2003, African Parks, Majete (Pty) Ltd. and the Malawian Department of National Parks and Wildlife (DNPW) signed a 25-year private public partnership (PPP) agreement to restore Majete to its former state. The restoration started in 2003 with the construction of a fence line around the sanctuary (140 km2) in the north-eastern section of the reserve, which served as the area for wildlife reintroductions between 2003 and 2011. In total, 2 550 herbivores from 12 different species were reintroduced during this period (see Appendix 5A). A predator-proof perimeter fence line was completed in 2008 and by 2011, the sanctuary fence line was removed and animals could utilise the entire reserve. Between 2011 and 2012, three lions and six leopards were reintroduced into Majete to assist hyenas in regulating herbivore populations, thus facilitating the ecological restoration process.

5.3.2 Data collection

Diet determined from scat collection

Lion, leopard and hyena scats were collected opportunistically along roads and trails from March 2016 to May 2017. Hyena scats were also sampled from dens and latrines and to avoid pseudo-replication, only one scat per prey item was selected from samples on the same day at a specific site (Périquet et al., 2015a). Lion scats were additionally sampled at kill sites, although this was infrequent (n = 4) due to the dense vegetation. Scat were only collected when fresh and only half of each scat was sampled, as all focal species use scat as a means of territorial marking (Bailey, 1993; Kruuk, 1972; Schaller 1972). Predator scat samples were collected in brown envelopes and labelled with the species name, date, GPS coordinates and site in relation to access (on the road, game trail or no trail). Previous studies recommend a sample size of at least 50 scats to infer reliable results, particularly for opportunistic carnivores (Williams, Goodenough & Stafford, 2012; Trites & Joy, 2005).

Each scat was distinguished using size, segmentation, colour, shape, and size of bone shards, if present (Walker, 1996). Lion scats are large and sausage-shaped, with a dark colouration indicating consumption of meat and a light colouration indicating consumption of bones (Walker, 1996). Hyena scats are also large, but have an irregular shape and colouration is green when wet and distinctively white when dry (Murray, 2011). Leopard scats are segmented with tapered ends and the presence of large amounts of hair is characteristic (Walker, 1996). The presence of hair from the focal species in

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scat samples was considered a result of grooming, which may occur in felid species (Martins, Horsnell, Titus, Rautenbach & Harris, 2011).

Scats were dried in the sun for two to three days. All dried scat samples were first soaked in warm water for two days and then in 95% ethyl alcohol for 24 hours, to kill all potential parasites and to further soften the scat. A pestle and mortar were used to gently help fragment and grind scats to ease the washing procedure. Separated faecal matter was placed in a wire sieve (1.5 mm) and thoroughly rinsed with water until only the undigested hairs and bones remained. These samples were again sun- dried in petri dishes and stored in small brown envelopes until further analyses.

Macroscopic identification of ungulate hooves and hair were used where possible. However, all samples were identified microscopically to the species-level using hair samples (except reptiles and livestock). Hair samples remain relatively undamaged during digestion and deterioration is minimal over time, therefore predator scats are an effective way to examine feeding habits (Keogh, 1979; Keogh, 1983). For each scat sample, a minimum of 15 hairs were required to make the analyses possible (Forbes, 2011). Cross-sections of hairs were prepared using the techniques proposed by Douglas (1989). Cross-sections were prepared by selecting hairs with a pair of forceps and positioning each one longitudinally in a 3 mm plastic Pasteur pipette (Trites & Joy, 2005). Molten wax (Paraplast Plus®, Leica Biosystems) was then taken up in the pipette and cooled in an ice container to set the wax. Thin cross sections were cut using a minora blade and mounted on glass slides. A LeicaTM DM 2000 light microscope was used to examine slides at 20–40x magnification. Cross-sections were used to identify prey species based on the cortex, medulla colouration, shape and thickness of the hair sample (Keogh, 1983). Photographs of cross sections were taken and measurements were made using Leica Application Suite (LAS) Core V4.0 software (Leica Microsystems, 2011). Hair cross-sections were compared with several reference collections (Buys & Keogh, 1984; Keogh, 1979; Keogh, 1983; Rhodes University; Stellenbosch University).

Diet determination from GPS cluster visitation

Like other large felids, lions usually remain in close proximity to their kills until the kill is entirely consumed (Schaller, 1972). Therefore, if a lion is fitted with a GPS/global system for mobile communications (GSM) collar (i.e. GPS with mobile phone capabilities), a cluster of data points may indicate a potential kill site i.e. presence at a site for multiple hours or days. In August 2016, two lions were fitted with GPS/GSM collars from African Wildlife Tracking (http://www.awt.co.za). Collars transmitted data remotely from the GSM network to a server every hour, although fixes were dependent on GSM coverage.

Between August 2016 and May 2017, GPS fixes were monitored between one and three days for potential clusters. GPS cluster (or kill) sites were manually identified by two or more locations within

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50 m of each other (Martins et al., 2011). Once potential kill sites were identified, clusters were visited with an armed scout. Sites were visited within one to nine days after identifying clusters, which depended on the presence of lions at the site and scout availability. A handheld Garmin eTrex® 30 (Garmin International, Olathe, KS, USA) was used to navigate to the central point of the cluster. Each site was thoroughly searched by the field team (two to eight individuals) in zig-zag transect lines covering a 100 m radius from the cluster centroid. Field teams searched for any prey remains such as bones, horns, hooves, hairs and rumen. The smell of decaying meat was helpful in determining the presence and location of a kill. Prey remains were photographed and representative material was collected for identification. When prey could not be identified, hair was sampled and analysed under a microscope at a later stage (Keogh, 1979; Keogh, 1983). If no prey remains were found after an intensive search, the cluster was termed as a non-kill cluster (Martins et al., 2011). However, the lack of a kill could also be attributed to small prey items being entirely consumed by lions (see Power, 2002), thus highlighting some degree of bias towards large prey. In addition, scavengers could eliminate signs of a feeding site (Tambling et al., 2012).

Prey species were categorised into the following groups based on Mbizah, Marino and Groom (2012); large mammals (>100 kg), medium mammals (25–100 kg), small mammals (5–25 kg), very small mammals (<5 kg). Reptiles and livestock were also recorded, but not categorised to a lower taxonomic level. If possible, prey age classes were assigned according to Hirst (1969): adult (>24 months), subadult (6–24 months) and juvenile (<6 months). Sex of prey was determined where possible by the presence of horns in males (if applicable) as well as genitalia.

Aerial census

An aerial census was conducted during the dry season between 21–23 September 2015 to determine the population abundance of large mammal species in Majete. Due to the reserve’s size, it was impossible to sample the entire reserve in a single day and therefore Majete was divided into three blocks and sampled over three consecutive days. The census was carried out in a Bell 407 helicopter (Bell Helicopters, 2014) and the team consisted of one pilot and three observers. Transects were flown from east to west and spaced 500 m apart according to the recommendations by Jachmann (2002). The calibrated height above sea level was between 106–139 m and a speed of less than 100 knots was maintained throughout the sampling period. Upon sighting an animal, the observers recorded the GPS track log of all transects, way point, species sighted and number of individuals in the group. Aerial census data were provided by African Parks (Pty) Ltd. and were used for calculations of prey preferences in predator diet.

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5.3.3 Data analysis

Diet composition

Diet composition of each predator species was determined by the frequency of occurrence (FO) for each prey item. FO was calculated by dividing the number of times a specific prey item occurred by the total number of all prey items identified and converted to a percentage by multiplying by 100 (Klare, Kamler & MacDonald, 2011). This method tends to overestimate smaller prey species (Klare et al., 2011) and was therefore only used for comparative purposes. The favoured frequency-based method is corrected frequency of occurrence (CFO), which assigns a weighting of 1 per scat (Karanth & Sunquist, 1995). If two prey items are present in a scat, each item receives a weighting of 0.5 and if three items are recorded, each receives a weighting of 0.33.

The most accurate approximation of carnivore diet is represented by linear regression models which convert frequency of occurrence into relative biomass (Klare et al., 2011). These linear models were pioneered by Floyd, Mech and Jordan (1978) for grey wolf (Canis lupus) and Ackerman, Lindsey and Hemker (1984) for cougar (Puma concolor). The latter model has since been applied on similar body- sized carnivores with similar digestive systems such as tigers (Panthera tigris) and leopards (e.g. Karanth & Sunquist, 1995; Lyngdoh et al., 2014). Thus, to calculate the average biomass consumed by lions, leopards and hyenas, the following regression was used (Ackerman et al., 1984):

y = 1.98 + 0.035x

Where y is the weight of the prey consumed per scat (kg/scat) and x is the mean body mass of the prey item (kg) (Mann, 2014). The value of x was determined using 75% of the female body weight, which was based on Stuart and Stuart (2015), which accounted for juveniles and subadults (Jooste, Hayward, Pitman & Swanepoel, 2013). Differences in the proportion of prey age class and sex (based on kills) were tested using a Kruskal-Wallis ANOVA test by rank and Mann-Whitney U test respectively, due to assumption of normality not being met. All statistical analyses were conducted in Statistica version 13.2 (Dell Software, 2016).

Prey preference

Prey preference was calculated using Jacobs’ index because it reduces problems related to electivity indices (Jacobs, 1974). 푟 − 푝 퐷 = 푟 + 푝 − 2푟푝

Jacob’s index compares the relationship between the relative proportion that each species is preyed upon r and relative abundance of prey p. The relative abundance was derived from game count data conducted in 2015 (Appendix 5A). The index values range from +1 (maximum preference) to -1

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(maximum avoidance) and 0 indicates no preference or avoidance (Jacobs, 1974). This is clearly a simplification, as Jacobs’ index might reflect not only prey preference, but also ease of capture and energy invested in hunting specific prey species (Hayward et al., 2006; Schaller, 1972).

Diet comparison

Comparison of diet for the three apex predators was based only on results from scat analyses. The Kruskal-Wallis test by ranks was used to test for differences in mean preferred prey weight, calculated using Jacobs’ index.

Dietary overlap and breadth

Diet overlap was determined using Pianka’s (1973) index:

∑푛 푃푖푎푃푖푏 푂푎푏 = 2 2 1/2 (∑ 푃푖푎 푃푖푏) where Oab is the degree of dietary overlap between species a and b; Pia is the relative frequency of the prey item i found in the scat of species a; Pib is the relative frequency of the prey item i found in the scat of species b; and n is the total number of prey items in a predator scat. The resulting values range from 0 (no overlap) to 1 (complete overlap; Breuer, 2005; Woodward & Hildrew, 2002).

2 Niche breadths were calculated using Levins’ index 퐵 = ∑푛 1/푝푖 and Levins’ standardised niche breadth 퐵푎 = (퐵 − 1)/(푛 − 1), where B is the degree of niche breadth, p is the proportion of occurrence of each prey item in predator diet; Ba is the standardised niche breadth and n is the number of prey items in predator diet (MacArthur & Levins, 1967). The index values of niche breadth (B) range from 1 to n, while standardised niche breadth (Ba) range from 0 to 1. Lower values show specialist carnivore diets and higher values show generalist carnivore diets.

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5.4 Results

Lion diet

Scat analysis

A total of 50 lion scats were collected during the sampling period and analysed to determine lion diet. Lion diet consisted of 13 prey species, all of which all were mammals. Warthog (Phacochoerus africanus; 38.33%), waterbuck (Kobus ellipsiprymnus; 26.00%) and kudu (Tragelaphus strepsiceros; 12.00%) were the most frequently recorded prey items in lion scat (Table 5.1). The total biomass consumed was 8 686 kg, with waterbuck (38.96%), kudu (14.97%) and warthog (11.88%) contributing the most to the relative biomass ingested, whereas sable antelope (Hippotragus niger; 9.73%), African buffalo (Syncerus caffer; 9.51%) and plains zebra (Equus quagga; 5.02%) also contributed, but to a lesser degree (Table 5.2). Naïve biomass consumed was converted to corrected (or actual) biomass consumed using the linear regression equation developed by Ackerman et al. (1984). Consequently, the biomass consumed was substantially lower at 603.35 kg. Corrected biomass consumed increased the relative importance of smaller species, but decreased those of larger species. Despite this, waterbuck (34.05%), warthog (18.48%) and kudu (14.43%) remained the three most important species based on lion scats.

Kill site analysis

A total of 61 lion kills were found using GPS kill site analysis. The percentage of kills found was highest for warthog (32.79%), waterbuck (24.59%) and kudu (11.48%), which corresponds with results from the scat analysis (Table 5.1). The total biomass consumed was 7 448 kg, while the corrected biomass consumed was 625.34 kg. Warthog, waterbuck and kudu were again the most important prey species in lion diet (Appendix 5B).

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Table 5.1. Prey species recorded in lion diet by means of scat analysis and kills found at GPS cluster locations in Majete Wildlife Reserve, Malawi. Frequency of occurrence (FO) was calculated as the percentage of each prey item relative to the total number of prey items recorded (n = 75). Corrected frequency of occurrence (CFO) shows the percentage of occurrences (per scat) relative to the total number of scats collected (n = 50). Number (n = 61) and percentage of kills represented kill site analysis.

Scat analysis Kill site analysis Number of occurrences Number of occurrences Prey species FO (%) CFO (%) Number of kills Kills (%) (prey species) (per scat) African buffalo 2 2.67 1.50 3.00 1 1.64 Bushbuck 3 4.00 1.50 3.00 2 3.28 Bushpig 1 1.33 0.33 0.67 - - Eland 1 1.33 0.33 0.67 2 3.28 Impala 3 4.00 2.00 4.00 1 1.64 Kudu 10 13.33 6.00 12.00 7 11.48 Lichtenstein's hartebeest 1 1.33 0.50 1.00 5 8.20 Nyala 3 4.00 1.50 3.00 3 4.92 Plains zebra 2 2.67 1.00 2.00 2 3.28 Reedbuck 2 2.67 0.50 1.00 - - Sable antelope 5 6.67 2.67 5.33 3 4.92 Warthog 24 32.00 19.17 38.33 20 32.79 Waterbuck 18 24.00 13.00 26.00 15 24.59

75 100.00 50.00 100.00 61 100.00

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Table 5.2. Biomass consumed and total biomass consumed determined from lion scat (n = 50) collected in Majete Wildlife Reserve, Malawi.

Correction Number of Biomass Biomass Corrected Corrected Prey mass Occurrence of Prey species factor occurrences (prey consumed consumed biomass biomass (kg)a prey species (%) (kg/scat)b species) (kg)c (%) consumed (kg)d consumed (%) African buffalo 413 16.44 2 2.67 826 9.51 43.83 7.26 Bushbuck 23 2.79 3 4.00 69 0.79 11.14 1.85 Bushpig 45 3.56 1 1.33 45 0.52 4.74 0.79 Eland 338 13.81 1 1.33 338 3.89 18.41 3.05 Impala 34 3.17 3 4.00 102 1.17 12.68 2.10 Kudu 130 6.53 10 13.33 1 300 14.20 87.07 14.43 Lichtenstein's hartebeest 120 6.18 1 1.33 120 1.38 8.24 1.37 Nyala 47 3.63 3 4.00 141 1.62 14.50 2.40 Plains zebra 218 9.61 2 2.67 436 5.02 25.63 4.25 Reedbuck 24 2.82 2 2.67 48 0.55 7.52 1.25 Sable antelope 169 7.90 5 6.67 845 9.73 52.63 8.72 Warthog 43 3.49 24 32.00 1 032 11.88 111.52 18.48 Waterbuck 188 8.56 18 24.00 3 384 38.96 205.44 34.05

1 792 88.46 75 100.00 8 686 100.00 603.35 100.00

a From Stuart and Stuart (2015) b Based on Ackerman et al. (1984), y = 1.98 + 0.035x, only for prey >1 kg c Prey mass x Number of occurrences d Correction factor x Occurrence of prey species

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Prey preferences

The Jacobs’ index values for the scat analysis (based on CFO and biomass methods) revealed a strong preference for warthog (D = 0.78) and waterbuck (D = 0.53) respectively (Figure 5.2). Kudu were also utilised in lion diet, but to a lesser extent (D = 0.33). Nyala were consumed according to their availability due to their relatively low, positive D-values (see Appendix 5D). Both methods indicated avoidance of bushpig (Potamochoerus larvatus) and impala (Aepyceros melampus), while African buffalo, bushbuck (Tragelaphus scriptus), plains zebra, reedbuck (Redunca arundinum) and sable antelope were not selected or consumed relative to their proportional abundance (D < 0).

Warthog

Waterbuck

Lichtenstein's hartebeest

Kudu

Nyala

Bushbuck

Sable

Prey species Plains zebra

Reedbuck

African buffalo

Eland

Bushpig

Impala

-1.00 -0.80 -0.60 -0.40 -0.20 0.00 0.20 0.40 0.60 0.80 1.00 Jacobs' Index (D)

Biomass CFO

Figure 5.1. Jacobs’ index (D) indicating preference (+1) and avoidance (–1) of prey species by lion in Majete Wildlife Reserve, Malawi. D-values are derived from the corrected frequency of occurrence and total biomass consumed of prey species from lion scat.

The Jacobs’ index for the kill site analysis revealed that lions preferred Lichtenstein’s hartebeest (D = 0.87) and warthog (D = 0.59), while waterbuck, eland (Taurotragus oryx), nyala (Tragelaphus angasii) and kudu were also utilised, but to a lesser extent (Appendix 5C). However, eland and nyala occurred in <5% of lion kills. Furthermore, lions avoided African buffalo (D = –0.58) and impala (D = –0.90).

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Prey age class and sex

Based on kill site analysis, adult animals were preyed upon significantly more than sub-adults or juveniles (Kruskal-Wallis test, H = 10.20, d.f. = 2, p = 0.006; Figure 5.2). Overall, lion kills consisted mostly of adult males (>60.00%). A large proportion of males were killed compared to females; however, the difference was not significant (Mann-Whitney U test, U = 15.50, p = 0.08).

80 70 60 50 40 30 20 Percentage (%) kills (%) Percentage 10 0 Adult Sub-adult Juvenile

Prey age classes

Male Female

Figure 5.2. Percentage (%) of different prey age classes and sex killed by lions (n = 61) in Majete Wildlife Reserve, Malawi.

The most frequently killed prey i.e. warthog, waterbuck, kudu and hartebeest were separated into different age classes and sex, but due to the small sample size, no statistical differences could be detected (Table 5.3). For these four species, adult prey appeared more important in lion diet than lower age classes. Males were more prominent in lion diet than females for Lichtenstein’s hartebeest, waterbuck and warthog, while for kudu, females were killed marginally more than males.

Table 5.3. Proportions (%) of age classes and sex of the main prey killed by lions in Majete Wildlife Reserve, Malawi.

Age class Sex Prey species n Adult Sub-adult Juvenile Unknown Male Female Kudu 6 50.00 50.00 0.00 0.00 42.86 57.14

Lichtenstein’s hartebeest 5 100.00 0.00 0.00 0.00 100.00 0.00

Warthog 20 75.00 10.00 10.00 5.00 75.00 25.00

Waterbuck 14 71.43 14.29 7.14 7.14 76.92 23.08

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Leopard diet

A total of 42 leopard scats were collected and analysed to determine the dietary ecology of leopards. Leopard diet included a total of 18 prey species, consisting mostly of mammalian prey (97.10%) with a small percentage of reptiles (2.90%). Common duiker (Sylvicapra grimmia; 24.21%), impala (21.43%) and bushbuck (12.31%) occurred most frequently in leopard scats, while nyala (9.12%) and reedbuck (6.75%) also contributed, but to a lesser extent. (Table 5.4). The total biomass consumed based on leopard scats was 2 496.35 kg (Table 5.5). Impala (16.34%), kudu (15.62%), waterbuck (15.06%), nyala (13.18%) and common duiker (10.90%) made up the majority of relative biomass consumed. Biomass consumed was converted to corrected biomass consumed, which resulted in a substantially lower total of 328.41 kg. This conversion reduced the importance of larger prey species such as waterbuck and kudu and increased the contribution of smaller prey to the corrected biomass consumed. As a result, common duiker (19.62%), impala (17.29%), nyala (11.53%), bushbuck (11.39%) were the main contributors.

Table 5.4. Prey species recorded in leopard scat collected in Majete Wildlife Reserve, Malawi. Frequency of occurrence (FO) was calculated as the percentage of each prey item relative to the total number of prey items recorded (n = 69). Corrected frequency of occurrence (CFO) shows the percentage of occurrences (per scat) relative to the total number of scats collected (n = 42). Number of occurrences Number of occurrences Prey species FO (%) CFO (%) (prey species) (per scat)

Baboon 1 1.45 0.50 1.19 Bushbuck 9 13.04 5.17 12.31 Bushpig 1 1.45 1.00 2.38 Common duiker 17 24.64 10.17 24.21 Elephant shrew 1 1.45 0.50 1.19 Impala 12 17.39 9.00 21.43 Klipspringer 1 1.45 0.50 1.19 Kudu 3 4.35 2.00 4.76 Lizard 1 1.45 0.33 0.79 Nyala 7 10.14 3.83 9.12 Porcupine 1 1.45 0.33 0.79 Reedbuck 5 7.25 2.83 6.75 Sable antelope 1 1.45 0.33 0.79 Savanna hare 2 2.90 1.33 3.17 Sharpe's grysbok 1 1.45 0.50 1.19 Snake 1 1.45 0.50 1.19 Warthog 3 4.35 1.83 4.37 Waterbuck 2 2.90 1.33 3.17

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Table 5.5. Biomass consumed and total biomass consumed determined from leopard scat (n = 42) collected in Majete Wildlife Reserve, Malawi. Number of Corrected Corrected Prey mass Correction Occurrence of Biomass Biomass Prey species occurrences biomass biomass (kg)a factor (kg/scat)b prey species (%) consumed (kg)c consumed (%) (prey species) consumed (kg)d consumed (%) Baboon 12.00 2.40 1 1.49 12.00 0.48 3.58 1.10 Bushbuck 23.00 2.79 9 13.43 207.00 8.29 37.47 11.51 Bushpig 45.00 3.56 1 1.49 45.00 1.80 5.30 1.63 Common duiker 16.00 2.54 17 25.37 272.00 10.90 64.44 19.79 Elephant shrew 0.35 - 1 1.49 0.35 0.01 - - Impala 34.00 3.17 12 17.91 408.00 16.43 56.77 17.44 Klipspringer 10.00 2.33 1 1.49 10.00 0.40 3.47 1.07 Kudu 130.00 6.53 3 4.48 390.00 15.62 29.25 8.98 Nyala 47.00 3.63 7 10.45 329.00 13.18 37.93 11.65 Porcupine 13.00 2.44 1 1.49 13.00 0.52 3.64 1.12 Reedbuck 24.00 2.82 5 7.46 120.00 4.81 21.04 6.46 Sable antelope 169.00 7.90 1 1.49 169.00 6.77 11.77 3.61 Savanna hare 5.00 2.16 2 2.99 10.00 0.40 6.46 1.98 Sharpe's grysbok 6.00 2.19 1 1.49 6.00 0.24 3.26 1.00 Warthog 43.00 3.49 3 4.48 129.00 5.17 15.64 4.80 Waterbuck 188.00 8.56 2 2.99 376.00 15.06 25.59 7.86

765.35 56.51 67 100 2 496.35 100.00 325.62 100.00 a From Stuart and Stuart (2015) b Based on Ackerman et al. (1984), y = 1.98 + 0.035x, only for prey >1 kg c Prey mass x Number of occurrences 130 d Correction factor x Occurrence of prey species Stellenbosch University https://scholar.sun.ac.za

Prey preference

Jacobs’ index revealed a strong preference for nyala (D = 0.66), common duiker (D = 0.62) and bushbuck (D = 0.62) in leopard diet (Figure 5.3). D-values suggested that reedbuck, Sharpe’s grysbok (Raphicerus sharpei) and klipspringer (Oreotragus oreotragus) were utilised; however, grysbok and klipspringer each only occurred once in leopard scat. Despite having the second highest CFO proportion (21.43%) and corrected biomass consumed (17.29%), impala had Jacobs’ index values close to 0, indicating that they were not preferred, but rather consumed in proportion to their relative abundance. Leopards strongly avoided sable antelope (D = –0.52), while bushpig, porcupine (Hystrix africaeaustralis), warthog and waterbuck also had negative D-values for both CFO and biomass calculations (Appendix 5D).

Duiker Bushbuck Nyala Klipspringer Grysbok Reedbuck Impala Bushpig Prey species Kudu Warthog Waterbuck Sable Porcupine

-1.00 -0.80 -0.60 -0.40 -0.20 0.00 0.20 0.40 0.60 0.80 1.00 Jacobs' Index (D)

Biomass CFO

Figure 5.3. Jacobs’ index (D) indicating preference (+1) and avoidance (–1) of prey species by leopards in Majete Wildlife Reserve, Malawi. D-values are derived from the corrected frequency of occurrence and total biomass consumed of prey species from leopard scat.

Hyena diet

A total of 128 hyena scats were collected throughout the sampling period and analysed to determine hyena diet. We recorded a total of 19 prey species in hyena diet, which comprised largely of wild prey (97.56%), with livestock (cattle and goat) contributing the other 2.44% (Table 5.6). Waterbuck (21.10%), warthog (17.97%) and impala (12.11%) occurred most frequently, while common duiker (9.25%) and plains zebra (7.81%) contributed marginally less. The total biomass consumed based on

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hyena scat was 21 651 kg (Table 5.7). Waterbuck (30.39%), plains zebra (17.12%), sable antelope (10.15%) and kudu (9.61%) made up the bulk of biomass consumed. Biomass consumed was converted to corrected biomass consumed and totalled 567.65 kg. This conversion increased the importance of smaller prey species and decreased the contribution of larger prey species; however, waterbuck (25.75%) and plains zebra (14.04%) remained the two most important species in hyena diet.

Table 5.6. Prey species recorded in hyena scat collected in Majete Wildlife Reserve, Malawi. Frequency of occurrence (FO) was calculated as the percentage of each prey item relative to the total number of prey items recorded (n = 205). Corrected frequency of occurrence (CFO) shows the percentage of occurrences (per scat) relative to the total number of scats collected (n = 128).

Number of Number of Prey species occurrences FO (%) occurrences CFO (%) (prey species) (per scat) Aardvark 1 0.49 0.50 0.39 African buffalo 4 1.95 2.50 1.95 Bushbuck 6 2.93 3.33 2.60 Bushpig 3 1.46 1.50 1.17 Cattle 3 1.46 1.17 0.91 Common duiker 14 6.83 11.83 9.25 Eland 3 1.46 2.33 1.82 Goat 2 0.98 1.00 0.78 Impala 28 13.66 15.50 12.11 Kudu 16 7.80 10.50 8.20 Lichtenstein's hartebeest 2 0.98 1.00 0.78 Nyala 9 4.39 4.83 3.77 Plains zebra 17 8.29 10.00 7.81 Porcupine 3 1.46 1.33 1.04 Reedbuck 8 3.90 4.33 3.38 Sable antelope 13 6.34 6.00 4.69 Sharpe's grysbok 1 0.49 0.33 0.26 Warthog 37 18.05 23.00 17.97 Waterbuck 35 17.07 27.00 21.10 205 100.00 128.00 100.00

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Table 5.7. Biomass consumed and total biomass consumed determined from hyena scat (n = 128) collected in Majete Wildlife Reserve, Malawi.

Correction Number of Biomass Biomass Corrected Corrected Prey mass Occurrence of Prey species factor occurrences consumed consumed biomass biomass (kg) prey species (%) (kg/scat) (prey species) (kg) (%) consumed (kg) consumed (%) Aardvark 41 3.42 1 0.49 41 0.19 1.67 0.29 African buffalo 413 16.44 4 1.95 1 652 7.63 32.07 5.65 Bushbuck 23 2.79 6 2.93 138 0.64 8.15 1.44 Bushpig 45 3.56 3 1.46 135 0.62 5.20 0.92 Cattle 131 6.57 3 1.46 393 1.82 9.61 1.69 Common duiker 16 2.54 14 6.83 224 1.03 17.35 3.06 Eland 338 13.81 3 1.46 1 014 4.68 20.21 3.56 Goat 24 2.82 2 0.98 48 0.22 2.75 0.48 Impala 34 3.17 28 13.66 952 4.40 43.30 7.63 Kudu 130 6.53 16 7.80 2 080 9.61 50.97 8.98 Lichtenstein's hartebeest 120 6.18 2 0.98 240 1.11 6.03 1.06 Nyala 47 3.63 9 4.39 423 1.95 15.91 2.80 Plains zebra 218 9.61 17 8.29 3 706 17.12 79.69 14.04 Porcupine 13 2.44 3 1.46 39 0.18 3.56 0.63 Reedbuck 24 2.82 8 3.90 192 0.89 11.00 1.94 Sable antelope 169 7.90 13 6.34 2 197 10.15 50.07 8.82 Sharpe's grysbok 6 2.19 1 0.49 6 0.03 1.07 0.19 Warthog 43 3.49 37 18.05 1 591 7.35 62.90 11.08 Waterbuck 188 8.56 35 17.07 6 580 30.39 146.15 25.75 2 023 108.43 205 100.00 21 651 100.00 567.65 100.00 a From Stuart and Stuart (2015) b Based on Ackerman et al. (1984), y = 1.98 + 0.035x, only for prey >1 kg c Prey mass x Number of occurrences d Correction factor x Occurrence of prey species 133

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Prey preference

Jacobs’ index value showed that hyenas preferred plains zebra (D = 0.56), based on biomass calculations (Figure 5.4). Overall D-values for warthog (0.32) and waterbuck (0.31) indicate that these species were also selected, but to a lesser degree. Lichtenstein’s hartebeest and nyala also had positive D-values, although each made up <5% of hyena diet. Low overall D-values indicate that hyenas consumed kudu and eland relative to their proportional abundance in the reserve (Appendix 5D). Based on the combination of CFO and biomass calculations, African buffalo and bushpig were strongly avoided in hyena diet (D < –0.5).

Warthog Plains zebra Waterbuck Nyala Common duiker Lichtenstein's hartebeest Reedbuck Kudu Bushbuck Eland Prey species Impala Porcupine Sharpe's grysbok Sable Bushpig African buffalo

-1.00 -0.80 -0.60 -0.40 -0.20 0.00 0.20 0.40 0.60 0.80 1.00

Jacobs' Index (D)

Biomass CFO

Figure 5.4. Jacobs’ index (D) indicating preference (+1) and avoidance (–1) of prey species by hyenas in Majete Wildlife Reserve, Malawi. D-values are derived from the corrected frequency of occurrence and total biomass consumed of prey species from hyena scat. Diet comparison

Dietary overlap and niche breadth

Based on Pianka’s dietary overlap index, lion and hyena had the highest overlap of actual prey species (0.88), followed by leopard and hyena (0.61; Table 5.8). The least amount of overlap was found between lions and leopard (0.29). Percentage overlap of actual prey species revealed a 68.42% overlap for lion and hyena, 48.00% for hyena and leopard and 40.91% for lion and leopard. Lion and hyena had a 60.00% overlap of preferred prey species (calculated using Jacobs’ indices), while leopard had zero overlap of preferred prey with lion and hyena. Niche breadth and standardised niche breadth indices were highest for hyena (B = 9.27, Ba = 0.46), second highest for leopards (B = 7.64, Ba = 0.39) and lowest for lions (B = 5.27, Ba = 0.36). 134

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Table 5.8. Dietary overlap and dietary niche breadth of the three apex predators in Majete Wildlife Reserve, Malawi. Values above the diagonal indicate percentage overlap and values below the diagonal represent Pianka’s dietary overlap index (Pianka, 1973). Values are based on actual prey species, while Jacobs’ index of prey preference are indicated in parentheses. Levins’ niche breadth and standardised niche breadth (Levins, 1968) are also indicated.

Dietary overlap Dietary niche Standardised dietary

Lion Leopard Hyena breadth (B) niche breadth (Ba) Lion - 40.91 (0.00) 68.42 (60.00) 5.27 0.36 Leopard 0.29 - 48.00 (0.00) 7.64 0.39 Hyena 0.88 0.61 - 9.27 0.46

Prey weight

Jacobs’ indices were used to determine preferred prey species. Mean preferred prey weight was highest for hyenas (149.67 ± 54.03 [SE] kg) and lions (120.25 ± 29.79 kg; Figure 5.5). Leopards had a mean preferred weight of 27.50 ± 6.74 kg, which was lower than that of hyenas and lions, but these differences were not significant (Kruskal-Wallis ANOVA, H = 5.37, d.f. = 2, p = 0.07). Lions and hyenas had a high frequency of occurrence of medium (25–100 kg) and large (>100 kg) prey in their diet, while small (5–25 kg) prey were less important (Figure 5.6). In contrast, small prey occurred most frequently (50.72%) in leopard diet, while large prey contributed only 8.70% to the total occurrence. Medium (25–100 kg) prey class appeared to be important for all three apex predators (range = 33.33–41.33%). Very small prey (0–5 kg) were excluded from these analyses due to the low occurrence (<5%) in predator diet.

250

200

150

100 Prey weight (kg) 50

0 Lion Leopard Spotted hyena

CFO Biomass

Figure 5.5. Mean preferred prey weight (kg), using CFO and biomass calculations, for the three apex predators in Majete Wildlife Reserve, Malawi. Jacobs’ index values were used to determine preferred prey species. Standard error bars are presented above and below means.

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20 Frequency of occurrence (%) occurrence of Frequency

0 Lion Leopard Hyena

5-25 25-100 >100

Figure 5.6. Frequency of occurrence (%) of prey weight classes (kg) in the diet of three apex predators in Majete Wildlife Reserve, Malawi. Prey class 0–5 kg was excluded as it only contributed to leopard diet (<5%).

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5.5 Discussion

Lion diet

Lions utilised prey with a mean body weight of 120.25 kg (range = 43–188 kg). Based on Jacobs’ index values from both scat and kill site analysis, strongly preferred prey included warthog, waterbuck and Lichtenstein’s hartebeest. Kudu were also selected, but to a lesser degree, and while nyala and eland were perceived to be selected, each made up <5% of the total scats or kills and were thus excluded (Klare et al., 2011). Across their range, lions prefer medium-to large-bodied prey in the 190–550 kg weight range (Hayward & Kerley, 2005), and in the Serengeti, lions have a preferred prey weight range of 170–250 kg (Sinclair, Mduma & Brashares, 2003). Despite the high abundance of large-bodied prey species in Majete (see Appendix 5A), our findings were well below those from broader scale studies, but were comparable to findings in small, enclosed reserves (Druce et al., 2004; Hunter, 1998; Kilian, 2003; Lehmann, Funston, Owen & Slotow, 2008; Power, 2002).

Lions reintroduced into small, enclosed reserves typically prefer warthog (Druce et al., 2004; Hunter, 1998; Kilian, 2003; Lehmann et al., 2008) or at least select them relative to their abundance (Power, 2002). The preferential selection of warthogs can be explained by 1) a high encounter rate, due to their abundance within the reserve or similar habitat usage; 2) high hunting success, due to the relative ease of capture (e.g. reduced vigilance due to small group size, incautious movements or lack of height advantage in tall grass; Hayward, Hayward, Tambling & Kerley, 2011); and 3) predictable movements to and from their burrows creating ambush sites or digging sites at the entrance of the burrow for lions (van Orsdol, 1984).

Jacobs’ index revealed that waterbuck were also a preferred prey item, and their body weight (188 kg) puts them at the upper end of the utilised weight range. Waterbuck were the second most abundant prey item in Majete, which could explain their occurrence in lion diet as well their apparent preference. Their selection was also likely to be influenced by their requirement of habitats near water, as lions exhibited preference for riverine habitat (Chapter 4), suggesting a regular encounter rate with lions (Hirst, 1969). The selection for waterbuck contradicts the theory that lions have a taste aversion for waterbuck due to the musky odour released by glands to protect the skin (see Hayward & Kerley, 2005).

Sable antelope were not selected, despite being one of the most abundant prey species and falling within the utilised prey weight range of lions. Hayward and Kerley (2005) noticed similar trends and ascribed this to either morphological (e.g. horns) or behavioural (e.g. enhanced vigilance through large herd sizes) strategies to avoid predation. Sable are considered an ‘edge’ species i.e. frequenting the ecotone/boundary between two habitat types (IUCN/SSC Antelope Specialist Group, 2017; Ries & Sisk, 2010) as they typically tend to move into woodland areas in the wet season and open grassland areas

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in the dry season in search of suitable forage (Estes, 2013). Although sable had a tendency to congregate around waterpoints in the dry season (based on aerial survey data), which increases vulnerability to predation by lions, their movement between habitat types (and occupying edge zones) likely reduces encounters with lions, which showed strong selection for dense wooded areas.

The avoidance of African buffalo can be explained by the small pride of lions in Majete, as successfully subduing larger prey typically requires larger pride sizes (Packer, Scheel & Pusey, 1990). Smaller prey species such as impala and reedbuck were also not utilised, although their importance in lion diet cannot be disregarded (Table 5.1). These smaller prey species could also have been missed when searching for kills as lions (and scavengers) often consume the entire animal leaving no remains. This bias in kill site analysis can be reduced by using this technique in combination with scat analysis to provide a more accurate and comprehensive representation of predator diet (Bacon, Becic, Epp & Boyce, 2011; Martins et al., 2011; Tambling et al., 2012). As expected, GPS cluster analysis slightly underestimated smaller prey items (3.28%), while scat analysis did not appear to show typical biases in lion diet, especially after correction factors were applied (Karanth & Sunquist, 1995). However, the complementary use of scat and GPS cluster analysis proved helpful in identifying both preferred and avoided prey species of lions in Majete. We encourage using a combination of scat and GPS cluster analysis to provide the most complete reconstruction of predator diet.

The selection of prey species by lions in Majete is probably due to prey abundance within lion habitat. For example, Hayward et al. (2011) found that a higher prey abundance can lead to higher encounter rates (by choosing to forage in areas rich in preferred prey), which ultimately increases hunting frequencies. Although this does not translate directly into hunting success, it does suggest that lions actively decide which prey to hunt (Hayward et al., 2011). Furthermore, hunting skills associated with small pride size might also influence prey selection in the reserve (Bissett, Bernard & Parker, 2012; Packer et al., 1990; Funston, 2001).

Leopard diet

Leopards have a broad diet, which is evident from their ability to survive on small locally abundant prey in difficult times and their morphological adaptations allowing them to successfully capture and subdue large-bodied prey (Hayward et al., 2006; Nowell & Jackson, 1996). Leopards may also target smaller age classes of large species (Bailey, 1993; Karanth & Sunquist, 1995) or cache carcasses of large prey in trees to avoid kleptoparasitism from lions and hyenas (Nowell & Jackson, 1996; Schaller, 1972). Despite this, leopards frequently show selection for small-to medium-sized ungulates (e.g. Hart et al., 1996).

In Majete, leopards had a wide range of actual prey species, ranging from rodents (0.35 kg) to waterbuck (188 kg). However, leopards selected small- to medium-sized prey, including nyala,

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common duiker, bushbuck and reedbuck, within a weight range of 16–47 kg (mean = 27.5 kg). Impala are often a major prey item in leopard diet (e.g. le Roux & Skinner, 1989), however, in Majete they were taken according to their availability. Our results were comparable to weight ranges in the Democratic Republic of Congo (Hart, Katembo & Punga, 1996), South Africa (Bailey, 1993; Bothma, van Rooyen & le Riche, 1997; Norton et al., 1986), Kenya/Tanzania (Schaller, 1972) and Zambia (Mitchell, Shenton & Uys, 1985). Additionally, our findings were similar to a range-wide study, which found a preferential prey weight range of 10–40 kg, with a mean of 23 kg (Hayward et al., 2006). Prey availability Jacobs’ preference index is

This selection may be explained by actual preference for certain prey species (Hayward et al., 2006). However, aerial survey data may underestimate the abundance of more cryptic species such as duiker and bushbuck, which means that prey preference may be an artefact of the game census metholodgy (depending on the accuracy of this methodology), rather than actual preference. An alternative hypothesis is that leopards select prey based on the optimal foraging theory, which predicts that carnivore foraging decisions are driven by maximising energy gain, while minimising energy expenditure and risk of injury (due to larger or more dangerous prey) to the carnivore (Krebs & Davies, 1993). This is expressed in their hunting tactics as leopards invest more effort in hunting medium- bodied prey, as opposed to considerably smaller or larger prey (Bailey, 1993; Bothma et al., 1997). For example, hunting bushbuck, common duiker, nyala and reedbuck also involves less risk and are generally easier to subdue (Hayward et al., 2006). However, Balme et al. (2017) argue that leopards could select even smaller-than-expected prey species, as predicted by the optimal foraging theory, to balance trade-offs between losses (encountered through kleptoparasistism) and energetic gains (obtained by killing larger prey).

Furthermore, several studies have suggested that leopards select these species due to their ideal size, solitary nature (e.g. bushbuck, common duiker and reedbuck) and preference for dense habitats (Bailey 1993; Mitchell et al., 1985; Pitman, Kilian, Ramsay & Swanepoel, 2013; Wilson, 1966). However, Balme, Hunter and Slotow (2007) found that leopards in Phinda Private Game Reserve, South Africa, hunted in habitats (with intermediate cover) where prey were easier to catch, rather than hunting the most abundant prey in the densest vegetation, which was likely due to lower detection of prey or increased detection of predator. Despite falling within the utilised weight range of leopards, both bushpig and warthog were avoided. These species are capable of inflicting considerable injury (Hayward et al., 2006) and thus the risks involved in subduing these and other suids (Ramakrishnan, Coss & Pelkey, 1999) may explain their avoidance.

To infer reliable results about predator diets, at least 50 scat samples should be collected and analysed (Williams et al., 2012; Trites & Joy, 2005). Locating leopard scats proved challenging and only 42 samples were collected and analysed during this study. Our results did however coincide with other 139

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studies, although we recommend further investigation to provide a more comprehensive analysis of leopard diet in Majete.

Hyena diet

The highly catholic nature of hyena predation was displayed in their dietary range from Sharpe’s grysbok (6 kg) to African buffalo (413 kg). This suggests that very few species are exempt from hyena predation. Hyenas are morphologically and behaviourally adapted to killing anything from a springhare to a giraffe (Cooper, 1990). Hyenas are also excellent, cursorial hunters and typically target smaller prey when solitary, but may take much larger prey when hunting in groups (Kruuk, 1972). Evidently, their extensive dietary range contributes to their status as ‘Least Concern’ by the International Union for Conservation of Nature (Bohm & Höner, 2015).

Although hyenas are opportunistic predators, they frequently exhibit selection for certain species, investing more energy in hunting abundant medium- to large-sized prey (Cooper, 1990; Cooper, Holekamp & Smale, 1999; Holekamp, Smale, Berg & Cooper, 1997). Our study showed that hyenas had site-specific prey preferences, as recorded in Moremi Game Reserve (Cooper, 1990), Ngorongoro Crater (Höner et al., 2002) and Addo Elephant National Park (Wentworth, Tambling & Kerley, 2011).

Hyenas showed the greatest preference for plains zebra, despite having an intermediate abundance (see Appendix 5A). This contrasts with a range-wide study which revealed a significant avoidance of zebra in hyena diet (Hayward, 2006).

Pains zebra typically form herds, which is a predator-avoidance strategy (Skogland, 1991). In Majete, plains zebra did not form herds (de Vos, 2017), which could explain why they were selected by hyena. However, the dense vegetation of Majete and encounter rate (due to similar habitat use) could also influence their selection. Furthermore, Jacobs’ index values revealed that hyenas also preferred warthog and waterbuck (which were of the most abundant species in the reserve, while nyala and Lichtenstein’s hartebeest had positive Jacobs’ index values, but were excluded as each contributed <5% to hyena diet (Klare et al., 2011).

A previous study conducted between 2014 and 2015, found that hyenas utilised common duiker, bushbuck, reedbuck and plains zebra (Retief, 2016). Our results indicate a change in hyena diet, except for the preference of plains zebra. This change in prey selection may reflect: (1) prey switching to more abundant prey (Owen-Smith & Mills, 2008; see Appendix 5A); (2) adapting foraging strategy from active hunting to scavenging from reintroduced lions and leopards (Höner et al., 2002; Periquet, Fritz & Revilla, 2015); or (3) changes in hyena population size and composition between 2014 and 2017 (Cooper, 1990; Trinkel, 2010; Wentworth et al., 2011). Since the resident hyena population appeared to remain stable since 2014 (see Chapter 6), we suggest that the shift in diet was either due to prey switching or increased scavenging, as hyena showed a high dietary overlap with lion for both actual

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and preferred prey (Table 5.7). Despite this, scat analysis is an indirect sampling method and therefore we acknowledge potential biases between studies (see Klare et al., 2011).

Hayward (2006) suggests that prey selection is driven by their abundance and whether they fall within the preferred weight range (56–182 kg). In contrast, Wentworth et al. (2011) found that the most abundant prey were avoided and the least abundant prey species were preferred by hyenas in Addo Elephant National Park and two preferred species fell outside of the preferred prey weight range based on Hayward (2006). In our study, hyenas appeared to select abundant prey which were outside of this utilised weight range. This highlights the importance of fine scale studies and caution that should be used when applying broad scale patterns to infer site-specific predator prey relationships (Hopcraft, Sinclair & Packer, 2005; Rapson & Bernard, 2007).

Dietary overlap

Leopard had a zero overlap of utilised prey with both lion and hyena. Pianka’s dietary overlap index was 0.58 for leopard and hyena and 0.27 for leopard and lion, which was considerably lower compared to a study in Bubye Valley Conservancy, Zimbabwe (du Preez et al., 2017). Ultimately, the adaptability of leopards, their use of denser habitats to hunt and their preference for smaller prey species compared to more dominant guild members, could explain why this species remains largely unaffected with changes in competitor levels (Hayward & Kerley, 2008).

Pianka’s dietary overlap index revealed a high overlap between lion and hyena in Majete (0.88). Their percentage actual and utilised prey overlapped 68.42% and 60.00% respectively, which was almost identical to other studies (Hayward & Kerley, 2008; Periquet et al., 2015b). Sympatric carnivores with similar diets may suggest exploitative or interference competition (Breuer, 2005; Kruuk & Turner, 1967), although this only applies when food is a limiting factor (MacDonald & Thom, 2011; Melero, Palazon, Bonesi & Gosalbez, 2008). The high dietary overlap between lions and hyenas could also be attributed to scavenging and kleptoparasitism, especially since waterbuck and warthog were preferred prey items in both lion and hyena diet, based on Jacobs’ index.

Each member of Africa’s large carnivore guild have long-term evolutionary adaptions allowing them to take prey within their capabilities (e.g. Hayward, 2006; Hayward et al., 2006; Hayward & Kerley, 2005). For instance, predator body size allows each predator to utilise a specific prey weight range, which, in theory, should reduce interspecific competition. Even within these parameters predators may select for specific species, further reducing competition (e.g. du Preez, Purdon, Trethowan, MacDonald & Loveridge, 2017). However, variation in feeding strategies, group hunting, the opportunist nature of predators and other factors may lead to dietary overlap (Hayward & Kerley, 2008). For example, the overlap found between lion and hyena in our study may be explained by opportunism, variation in foraging tactics or the availability of medium to large-sized prey such as warthog and waterbuck.

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Hyenas had the highest Levin’s niche breadth index which corresponds with findings from a range- wide study (Periquet et al., 2015b). The eclectic nature of their predation and opportunistic feeding behaviour clearly contribute to their relatively secure conservation status (Hayward, 2006). This may also allow for niche partitioning, thus minimising resource competition with lions (du Preez et al., 2017).

Relative to their body size, hyenas are expected to take smaller prey compared to lions (Hayward, 2006); however, hyenas took both small and large prey more so than lions in this and another study (Periquet et al., 2015b). This may be attributed to their ability to hunt in groups, allowing predation of much larger prey species, while taking smaller prey when hunting alone (Kruuk, 1972). Another hypothesis is that hyenas access larger-than expected prey by driving lions off large kills (Cooper, 1991); however, lions could respond by avoiding larger prey items due to the greater effort required to defend these large kills from hyenas. Furthermore, hyenas typically select the weakest individuals, often juveniles, within a herd of medium and larger prey species (Kruuk, 1972), which may be a confounding factor in hyena diet when using scat analysis. However, selection of different age classes within medium to large-sized prey species potentially further reduces competition with lions.

Scat analysis is an indirect method that often carries inherent biases (Klare et al., 2011) or undesired additional features. For example, during the dry season, various ungulate species suffer due to the heat and lack of forage, particularly buffalo, with some individuals dying. This presents the opportunity for scavengers to feed off animals that were not killed by themselves, although this probably occurs infrequently and is limited to a few months during the dry season. Another example, is the inclusion of younger animals (calves or juveniles) of large-bodied prey items (e.g. buffalo) which could misrepresent the diet of predators. However, to minimise this influence, we used three-quarters of the female prey body weight in the biomass calculations, which represents a ‘middle’ value to account for various age classes and sex (Jooste et al., 2013).

Scavenging and kleptoparasitism

Scavenging and kleptoparasitism occur in both directions for lion and hyena and the outcome of these interactions depends on the ecosystem. For example, if only lionesses and large cubs are present at a kill, hyenas can displace them if numerically dominant by a factor of 4 (Cooper, 1991), whereas the presence of at least one adult male lion at a kill almost always makes hyenas subordinate (e.g. Höner, Wachter, East, Runyoro & Hofer, 2005). Hyenas generally lose more kills to lions than vice versa, but hyenas do not lose greater quantities, in terms of biomass, than they gain from lions (Périquet et al., 2015b). Hyenas possibly compensate for the loss of their kills to lions by scavenging more food from the remains of lion kills (Périquet et al., 2015b). Their strong jaws and efficient digestive system allow them to utilise elements such as the skin and bones that lions cannot; lions only utilise carcasses with meat (Kruuk, 1972). In Majete, hyena clan and foraging group sizes were relatively small and the two 142

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male lions were almost always with the pride (Chapter 6). Therefore, it is unlikely that hyenas could regularly displace lions from a kill, but we may expect lions to drive hyenas off kills. In addition, lions had a relatively low risk of encountering hyenas (or least large groups of hyenas) as the pride’s range was in the north-east of the reserve (Chapter 4), with no hyena dens within this area. In Etosha National Park, Namibia, hyenas were scattered over large territories and were unable to steal lion kills or defend their own kills (Trinkel & Kastberger, 2005). However, hyenas would certainly benefit from scavenging kills left by lions. Thus in the context of Majete, where predators are sparsely distributed with large ranges and have small group sizes, we expect lions to benefit more from kleptoparasitism and hyenas more from scavenging. Interestingly, both obtain similar quantities of food from each other. Hyenas obtain 5–20% from lions (Gasaway, Mossestad & Standers, 1991; Henschel & Skinner, 1990; Höner et al., 2002; Kruuk, 1972) and lions obtain 0.6–24% from hyenas (Kruuk, 1972; Cooper, 1991; Cooper et al., 1999; Périquet et al., 2015b; Watts & Holekamp, 2008). Ultimately, the net loss or gain (from scavenging and kleptoparasitsm) contributes only a fraction of their overall daily requirements (Périquet et al., 2015b).

Hyenas and lions also regularly steal kills from leopards (Schaller, 1972; Stander et al., 1997). In the Sabi Sand Game Reserve, South Africa, leopards had 21% of their kills kleptoparasitised, mostly by hyenas (Balme et al., 2017). Kills had a higher probability of being kleptoparasitised if prey were large, not cached and in areas with high risk of encountering hyenas (Balme et al., 2017). This may influence prey selection by leopards in Majete, by selecting prey species that can be rapidly consumed or cached with relative ease to avoid kleptoparasitism from competitors.

5.6 Conclusion

Lions and hyenas had the greatest dietary overlap as both selected medium-to large-bodied prey species. Hyenas had the highest dietary niche breadth, which was likely due to their opportunistic nature and their ability to hunt alone or in groups. Leopards occupied a vacant dietary niche below that of lions and hyenas, and thus showed lower dietary overlap. Majete is a prey rich reserve and the overall impact of predators on prey populations appeared minimal. However, we advise management to carefully monitor predator-prey interactions, as a decline in total prey biomass can increase intraguild competition (especially between lion and hyena), while fewer predators will cause a rapid boom of prey in the reserve.

5.7 Acknowledgements

I would like to thank Earthwatch Institute for funding this research project. Thank you to African Parks (Majete) Pty Ltd., Andre Uys, Kester Vickery and all others involved in the lion re-collaring. Thank you to Craig Hay and Gervaz Tamala for access to GPS collar data. I would like to thank Prof Antoinette Malan for providing access to laboratory equipment and Dr Dan Parker and Rhodes University for 143

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providing reference images to help in the cross-section identification process. Thank you to Charli de Vos, Anel Olivier, Kayla Geenen, Claire Gordon and Frances Forrer for helping me collect scat samples and all the scouts for locating carcasses.

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5.9 Appendices

Appendix 5A. List of all mammalian prey species occurring within Majete Wildlife Reserve. Figures include reintroduced animals and population estimates from aerial surveys conducted in 2012 and 2015.

Aerial survey estimations Species Scientific name Introduced 2012 2015 African buffalo Syncerus caffer 306 915 1 319 African elephant Loxodonta africana 217 260 389 Black rhino Diceros bicornis 8 11 16 Bushbuck Tragelaphus scriptus - 150 400 Bushpig Potamochoerus larvatus 1 70 400 Common duiker Sylvicapra grimmia - 120 800 Eland Taurotragus oryx 77 180 320 Hippopotamus Hippopotamus amphibius - 80 85 Impala Aepyceros melampus 737 1 200 2 000 Klipspringer Oreotragus oreotragus - 0 50 Kudu Tragelaphus strepsiceros - 200 1 022 Lichtenstein's hartebeest Alcelaphus lichtensteinii 59 80 80 Nyala Tragelaphus angasii 59 100 300 Plains zebra Equus quagga 174 270 571 Porcupine Hystrix africaeaustralis - 0 200 Reedbuck Redunca arundinum - 150 400 Sable antelope Hippotragus niger 352 650 1 337 Sharpe's grysbok Raphicerus sharpei - 0 200 Warthog Phacochoerus africanus 158 500 1 500 Waterbuck Kobus ellipsiprymnus 402 700 1 782

2 376 5 636 13 171

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Appendix 5B. Biomass and biomass consumed based on lion kills (n = 61) collected at GPS cluster sites in Majete Wildlife Reserve.

Correction Number of Biomass Biomass Corrected Corrected Prey mass Occurrence of Prey species factor occurrences (prey consumed consumed biomass biomass (kg)a prey species (%) (kg/kill)b species) (kg)c (%) consumed (kg)d consumed (%) African buffalo 413 16.61 1 1.64 413 5.55 26.94 4.31 Bushbuck 23 2.79 2 3.28 46 0.62 9.13 1.46 Eland 338 13.81 2 3.28 676 9.08 45.28 7.24 Impala 34 3.17 1 1.64 34 0.46 5.20 0.83 Kudu 130 6.53 7 11.48 910 12.23 74.93 11.99 Lichtenstein's 120 6.18 5 8.20 600 8.06 50.66 8.10 hartebeest Nyala 47 3.63 3 4.92 141 1.89 17.83 2.85 Plains zebra 218 9.61 2 3.28 436 5.86 31.51 5.04 Sable antelope 169 7.90 3 4.92 507 6.81 38.83 6.21 Warthog 43 3.49 20 32.79 860 11.55 114.26 18.28 Waterbuck 188 8.56 15 24.59 2820 37.89 210.49 33.68

1 723 82.28 61 100.00 7 443 100.00 625.06 100.00

a From Stuart and Stuart (2015) b Based on Ackerman et al. (1984), y = 1.98 + 0.035x, only for prey >1 kg c Prey mass x Number of occurrences d Correction factor x Occurrence of prey species

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Lichtenstein's hartebeest Warthog Nyala Waterbuck Kudu Eland Bushbuck Prey species Plains zebra Sable African buffalo Impala

-1 -0.8 -0.6 -0.4 -0.2 0 0.2 0.4 0.6 0.8 1 Jacobs' Index (D)

Appendix 5C. Jacobs’ index (D) indicating preference (+1) and avoidance (–1) of prey species in lion diet using biomass calculations, based on kill site analysis in Majete Wildlife Reserve, Malawi.

Appendix 5D. Jacobs' preference index using CFO and biomass calculations from the scat of the three apex predators in Majete Wildlife Reserve, Malawi.

Lion Leopard Hyena Prey species CFO Biomass CFO Biomass CFO Biomass African buffalo -0.57 -0.18 - - -0.70 -0.31 Bushbuck -0.01 -0.26 0.63 0.61 -0.08 -0.37 Bushpig -0.65 -0.60 -0.13 -0.32 -0.46 -0.55 Common duiker - - 0.66 0.58 0.22 -0.35 Eland -0.58 0.11 - - -0.15 0.19 Impala -0.70 -0.83 0.07 -0.06 -0.19 -0.48 Klipspringer - - 0.52 0.47 - - Kudu 0.23 0.33 -0.26 0.07 0.02 0.07 Lichtenstein's hartebeest 0.24 0.38 - - 0.12 0.27 Nyala 0.13 0.02 0.62 0.69 0.25 0.10 Plains zebra -0.38 -0.02 - - 0.30 0.56 Porcupine - - -0.99 -0.16 -0.19 -0.42 Reedbuck -0.52 -0.43 0.39 0.37 0.05 -0.23 Sable antelope -0.34 -0.09 -0.87 -0.51 -0.40 -0.08 Sharpe's grysbok - - 0.52 0.45 -0.27 -0.34 Warthog 0.76 0.46 -0.29 -0.25 0.45 0.20 Waterbuck 0.38 0.53 -0.66 -0.30 0.26 0.37

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Chapter 6

Population dynamics and carrying capacity of two reintroduced felids and a resident hyenid in Majete Wildlife Reserve, Malawi

W.D. Briers-Louw1, A.J. Leslie1

1Department of Conservation Ecology and Entomology, Stellenbosch University, Matieland, Western Cape, 7602, South Africa

6.1 Abstract

Large carnivores are declining across the African continent due to habitat loss, prey base depletion and direct persecution. As a result, they are often reintroduced into small, enclosed reserves to compensate for human-induced losses and restore ecosystem functioning. In Majete Wildlife Reserve, all large carnivores, except spotted hyenas (Crocuta crocuta), were extirpated, but between 2011 and 2012, three lions (Panthera leo) and six leopards (Panthera pardus) were reintroduced into the reserve. We conducted a three-month, intensive camera trap survey to determine population and density estimates for reintroduced leopard and resident spotted hyena, whereas the lion population was regularly monitored and the population size known (n = 8). Capture-recapture and spatially explicit capture-recapture analyses generated similar population abundance and density estimates. The leopard population was estimated at 11.43 (± 2.72 [SD]), which almost doubled since the initial reintroduction six years prior. The spotted hyena population abundance was 18.36 (± 2.55), which highlights their behavioural plasticity to persist despite anthropogenic change. Spotted hyena density (2.62 hyenas/100 km2) and mean clan size (5.33 ± 0.67 [SE]) recorded in this study was the lowest to date in woodland habitats. Managing large carnivores in small, enclosed systems has questionable conservation value and therefore we encourage management to establish a regional scale managed metapopulation of large predators in Malawi to enhance their genetic viability in the future.

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6.2 Introduction

Apex predators are declining at an unprecedented rate across the African continent (Ray, Hunter & Zigouris, 2005; Weber & Rabinowitz, 1996; Woodroffe & Frank, 2005). Increasing human population and development has resulted in habitat fragmentation, with large carnivores suffering the most dramatic population and geographic range contractions (Estes et al., 2011; Ripple et al., 2014). These anthropogenic effects often surround protected areas, which is of particular concern for wide-ranging carnivores that move beyond the boundaries of protected areas, resulting in detrimental edge effects and ultimately extinction (Balme, Slotow & Hunter, 2010; Brashares, Arcese & Sam, 2001; Woodroffe & Frank, 2005; Woodroffe & Ginsberg, 1998).

Recently, Packer et al. (2013) stated that the future of large carnivore conservation, especially for lion (Panthera leo), requires fenced reserves. However, Creel et al. (2013) argued that conservation efforts should rather be directed at maintaining large, unfenced populations i.e. stronghold populations, as they hold the greatest conservation value. Unfortunately, only a few of these strongholds exist today (Riggio et al., 2013).

In some cases, the lack of fencing can cause rapid extirpation of remaining carnivore populations due to human-wildlife conflict (e.g. Bauer et al., 2015). Consequently, fencing reduces conflict with humans which typically occurs outside of unfenced protected areas (Cushman, Elliot, Macdonald & Loveridge, 2015; Ogada, Woodroffe, Oguge & Frank, 2003). However, one of the major disadvantages of enclosed reserves is restricted movement of wide-ranging species (Hayward & Kerley 2009). The lack of immigration and emigration often results in genetic isolation which causes inbreeding depression (Dubach et al., 2005; Packer, 1996; Björklund, 2003; Trinkel, Cooper, Packer & Slotow, 2011). This was observed in two lion populations in South Africa (Trinkel et al., 2008; Trinkel et al., 2010) and is likely occurring in many more fenced populations. Therefore, maintaining genetic integrity of large carnivore populations requires intensive management, which involves considerable costs (Hunter et al., 2007; Hayward et al., 2007a; Slotow & Hunter, 2009; Trinkel et al., 2008).

Confinement may also result in rapid population growth, as predator populations in these reserves are often above carrying capacity (Creel et al., 2013; Hayward, O’Brien & Kerley, 2007c). This is likely due to high densities of suitable prey and low levels of competition (Hayward et al., 2007b; Lehmann, Funston, Owen & Slotow, 2008). High predator densities in enclosed systems can rapidly deplete predator-naïve prey populations through top-down processes (Hayward et al., 2007b; Peel & Montagu, 1999; Slotow & Hunter, 2009; Tambling & du Toit 2005). To reduce these ecological impacts, active management is required, either by supplementing prey populations or controlling predator numbers.

There are several pleas in scientific literature for small, fenced reserves to move away from isolated reserve management towards a metapopulation approach to improve the conservation value at a

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regional scale (Ferreira & Hofmeyr, 2014; Funston, 2008; Hayward & Kerley, 2009; Slotow & Hunter, 2009; Trinkel et al., 2010). However, small reserves are often isolated and surrounded by land use types that are not suitable for removing fences to ‘physically’ connect reserves (Miller et al., 2013). Ideally, this form of management should involve a network of reserves which can be established ‘virtually’ by translocating carnivores to relatively nearby (unconnected) reserves. To maintain genetic diversity of carnivores in such a network requires reserve managers to simulate natural processes targeted at the reproductive, survival and dispersal stages of large carnivores (Ferreira & Hofmeyr, 2014).

Lions, leopards and hyenas were once common throughout Malawi in East Africa (Hayes, 1979; Morris, 2006). By 2010, lions were restricted to only a few protected areas totalling 35 resident individuals (Mésochina et al., 2010). Similarly, leopards are largely restricted to isolated patches and the only ‘viable’ population is believed to be in Nyika National Park, yet no data exist for leopards in Malawi. Hyenas are relatively widespread throughout the country (Morris, 2006), however populations appear to be declining, especially outside of protected areas.

In 2003, African Parks (AP) and the Malawi Department of National Parks and Wildlife (DNPW) collaborated in a venture to restore and develop Majete Wildlife Reserve. Majete was previously poorly managed and heavily poached, with only a small hyena population representing the remanants of large carnivores, despite on-going conflict with humans surrounding the reserve. Between 2003 and 2011, a total of 2550 herbivores (12 different species) were reintroduced into the reserve. This was followed by the reintroduction of three lions and six leopards between late 2011 and 2012. Majete is a small, fenced reserve that will require intensive management and research of its apex predators.

Effective and sustainable management of large carnivores in small, enclosed reserves requires an understanding of large carnivore ecology and how large carnivores respond to different management decisions (Druce et al., 2004). However, very little data exist on the ecology of reintroduced large carnivores within these enclosed systems, especially that of population dynamics and persistence (Funston, 2008; Hunter, 1998). Therefore, the aim of this study was to determine the population sizes and density estimates of reintroduced leopard and resident hyena in Majete. Carrying capacity was also calculated for lion, leopard and hyena, and recommendations for the future management of predator populations were provided.

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6.3 Methods

6.3.1 Study site

Located in the Lower Shire Valley of southern Malawi, Majete Wildlife Reserve covers 700 km2. The Shire and Mkulumadzi Rivers are the only two perennial rivers in the reserve and define the eastern and northern boundaries of the reserve. Majete has a semi-arid climate with two distinct seasons; a wet season (December to May) and a dry season (June to November). Mean annual rainfall varies from 680–800 mm in the east to 700–1 000 mm in the west, with a significant proportion of rainfall falling in the wet season (Staub, Binford & Stevens, 2013; Wienand, 2013). Average daily temperature is 28.4°C in summer and 23.3°C in winter (Wienand, 2013). An altitudinal gradient exists in the reserve, which decreases from the rugged and hilly western region to the flatter eastern region. Miombo woodland covers the western region (Brachystegia boehmii and Julbernardia globiflora), while vegetation transitions into mixed woodland towards the east (dominated by Acacia spp., Cleistochlamys kirkii and Steculia spp.) (Staub et al., 2013; Forrer, 2016).

6.3.2 Fieldwork

Survey design

A camera trapping survey was conducted from 8 September to 30 November 2016. The study area was divided into four grids and each grid was sampled sequentially for 20 days using 24 Cuddeback™ (Attack©, Ambush© and CE models©) camera traps (Balme, Hunter & Slotow, 2009; Karanth & Nichols, 2002; Silver et al., 2004) (Figure 6.1). This resulted in a total of 96 sites sampled during the 88- day camera trapping survey, which is within the recommended time period to satisfy the assumption of population closure (Karanth & Nichols, 2002).

A sampling occasion was defined as two consecutive trap-nights, resulting in ten sampling occasions. Although four grids were sampled, the total number of captures for occasion one was the total number of captures from the first sampling occasion of each grid. The total number of captures for occasion two included the sum of captures and recaptures from the second occasion in each grid, and so on (Balme et al., 2009; Chapman & Balme, 2010; Karanth & Nichols, 2002; Silver et al., 2004).

Site selection

Camera sites were selected based on the assumption that no animal had zero probability of being captured (Balme et al., 2009; Karanth & Nichols, 1998; Silver et al., 2004), meaning that at least one camera location was situated within an individual’s home range. The minimum home range of female leopards recorded in similar habitat is 10 km² (Smith, 1977), which is smaller than the minimum home range size recorded for hyenas in Majete (Retief, 2016). Therefore, by overlaying 10 km² circles within each of the four grids, camera sites were selected within each grid cell on a topographical map of the 159

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study area. The aim of a carnivore-based camera trap survey is to optimise capture probabilities of target species; therefore camera sites were selected in order to maximise target species capture (Balme et al., 2009; Chapman & Balme, 2010).

Camera placement & servicing

Sites were primarily selected in areas of known leopard and hyena activity, based on tourist sightings and field signs (Balme et al., 2009; Karanth & Nichols, 1998; Martins, 2010; Silver et al., 2004). If no signs of predator activity were found within a grid cell, sites were selected along expected paths of movement, which were occasionally located outside of target grid cells (Mann, 2014). We encountered small gaps in the centre of our survey owing to the dense vegetation, although these gaps probably had minimal impact on estimates, given the large home ranges of leopard (Chapter 4) and hyena (Retief, 2016). Due to the limited number of camera traps (24) and large study area (700 km2), each station comprised of only a single camera trap (Chapman & Balme, 2010). Cameras were positioned perpendicularly along active game trails, roads, and river beds (to ensure full body photographs). Camera traps were fastened to a suitable tree at 50–60 cm above the ground (Grant, 2012; Silver et al., 2004). Cameras were fitted inside a protective steel casing to prevent damage from animals such as elephant (Loxodonta africana) and hyena. A one-minute delay was used for all cameras in order to conserve battery life in highly active areas. Each camera was set to record the date and time and the GPS coordinates were taken from each station.

Cameras were serviced every seven to fourteen days to prevent data loss due to camera damage or malfunction. Servicing involved replacing memory cards and batteries if necessary. At the end of each grid phase, cameras were collected from the field, cleaned and deployed in the following grid within three days. Data memory cards were downloaded and files were stored.

Leopard and hyena identification

Leopards and hyenas were individually identified by their unique pelage patterns (Henschel & Ray, 2003; Miththapala et al., 1989; Holekamp, Smith, Strelioff, van Horn & Watts, 2012). Identikits were established over time using photographs of the flanks of individuals (see Appendix 6A) from previous camera trap studies as well as those obtained from tourists and management staff (Chapman & Balme, 2010). All identified individuals were verified by several independent researchers to avoid observer bias. Sexing male leopards was relatively easy due to their heavier-set, large neck and shoulders, and a prominent external orange scrotal sac (Balme, Hunter & Braczowski, 2012). Distinguishing females and subadults was more challenging. However, supplementary photographs allowed all individuals to be sexed by the end of the survey. The age of each individual leopard was classified as follows: juveniles or cubs (<2 years old), subadult (2–3 years old), adults (≥4 years old; Balme et al. (2012). Hyenas were aged based on Höner, Wachter, East, Runyoro and Hofer (2005) and sexed (where possible) based on

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the dimorphic morphology of their genitalia i.e. scrotum and erect phallus (Frank, 1986; Frank, Glickman & Powch, 1990).

6.3.3 Statistical analysis

We used two different approaches to estimate animal population abundance and density from our camera trap survey. The conventional capture-recapture (C-R) model requires capture-frequency data to estimate population abundance (Otis, Burnham, White & Anderson, 1978; Karanth & Nichols, 1998). A buffer, based on animal distances between recapture sites, is then used to create an effective sampled area to estimate density. However, the recent development of spatially explicit capture- recapture (SECR) models, which integrate spatial information directly into the models, has surpassed C-R models in terms of statistical robustness (Gopalaswamy et al., 2012; Royle, Karanth, Gopalaswamy & Kumar, 2009). Despite this, for comparison and consistency, both C-R and SECR models were used to estimate population abundance and density in this study (Athreya, Odden, Linnell, Krishnaswamy & Karanth, 2013; Braczkowski et al., 2016; Grant, 2012).

CAPTURE

For the C-R model we used the program CAPTURE. For this program, capture histories of leopard and hyena were created in a standard X-matrix format using binary values (1 = capture, 0 = no capture) in the defined sampling occasions (Otis et al., 1978). Data were analysed and estimates were generated using different probabilistic models (Otis et al., 1978; Rexstad & Burnham, 1991). These include the null Mo model (constant capture probabilities), Mh model (capture heterogeneity), Mb model

(behavioural response) and Mt model (time variation; Karanth & Nichols, 1998). CAPTURE determined the appropriateness of each model by calculating several goodness-of-fit and model-comparison test statistics for each of these models. Model selection is based on a discriminant function procedure which assigns a criterion score to each model, with the highest score representing the best-fitting model for the data (Rexstad & Burnham 1991). It also provides a closure test to determine whether the population remained demographically and geographically closed for the duration of the survey period (Otis et al., 1978; Rexstad & Burnham, 1991). CloseTest was also used to test for population closure, as it produces a more robust test (Stanley & Richards, 2005).

Estimating animal density using conventional C-R models depends largely on a subjective evaluation of a suitable effectively sampled area (Alexander, Gopalaswamy, Shi & Riordan, 2015). Regardless, density was defined as D = N/A(W), where N is the population abundance estimate computed in CAPTURE and A(W) is the effectively sampled area (Karanth & Nichols 1998). The effectively sampled area was calculated by adding a buffer to the camera trap polygon in ArcGIS 10.5 (Environmental Systems Research Institute (ESRI), Redlands, California, U.S.A.). Balme et al. (2009) evaluated the accuracy of buffer methods and found that the half mean maximum distance moved (½ MMDM) by

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recaptured animals was the most accurate method based on camera trapping. Therefore, we used ½ MMDM as the buffer width which was applied around each camera trap site, rather than around the entire polygon (Balme et al. 2009; Figure 6.1).

Figure 6.1. Camera sites (n = 96) were divided into four equal and consecutively sampled grids within Majete Wildlife Reserve. The ½ MMDM buffers of leopards and hyenas used to create effectively sampled areas.

SPACECAP

Leopard and hyena densities were also estimated using an SECR model, namely SPACECAP (Gopalaswamy et al., 2012) version 1.1.0 (Gopalaswamy et al., 2014) in R version 3.4.0 (R Development Core Team, 2014). SPACECAP directly estimates species density by explicitly using capture frequency data coupled with spatial information of camera trap sites under a Bayesian SECR framework and Markov chain Monte Carlo (MCMC) simulation. Additionally, this program generates non-asymptotic inferences that are suitable for small sample sizes (Royle et al., 2009).

Three data files were created following Gopalaswamy et al. (2012), which consisted of an animal capture file, trap deployment file and potential home range centres file (see Appendix 6B – 6D). The input files were uploaded into SPACECAP and data were run using the following model definitions: Half-normal model, Bernoulli (binary) encounter model, Spatial capture-recapture and Trap response 162

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absent. The number of iterations were set at 50 000 (leopard) and 100 000 (hyena), burn-in at 1 000 (leopard) and 10 000 (hyena), no thinning rate (i.e. default of 1) and data augmentation at ten times the number of captures for each species (see Gopalaswamy et al., 2012 for details of method selection).

Using ArcGIS 10.5, the outermost camera traps were connected to form a minimum area rectangle. The ½ MMDM buffer was initially applied to the rectangular camera trap polygon to create the ‘state- space’ area which displayed potential home range centres of leopards and hyenas (Figure 6.2; Royle et al., 2009). We then applied a larger buffer of 15 km with a grid of equally spaced points (5 385) each with a size of 0.56 km2 to determine whether density changed with a greater state-space (see Gopalaswamy et al., 2012 for methodological details). Areas outside of protected areas were considered unsuitable animal habitat due to the high human population density and extensive agriculture, while the Shire River was also excluded (Athreya et al., 2013; Braczkowski et al., 2016; Grant, 2012). Therefore, only Majete, Lengwe National Park and Thambani Forest Reserve were considered suitable habitat for home range centres.

The Bayesian p-value, based on individual encounter frequencies, was used to assess model suitability. A p-value close to 0.5 indicates an adequate model, while values close to 0 or 1 indicate an inadequate model. To determine whether MCMC simulations have converged around a solution, the set of Geweke diagnostic statistics should fall within the range –1.6 to 1.6 (Gopalaswamy et al., 2014).

Carrying capacity estimation

Carrying capacities for leopard, lion and hyena in Majete were determined using equations from Hayward et al. (2007c), where carnivore densities are related to prey densities within a given area. We compared utilised prey species and utilised prey weight ranges found in range-wide studies (Hayward, 2006; Hayward & Kerley, 2005; Hayward et al., 2006) and our site-specific study (see Chapter 5).

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6.4 Result

Figure 6.2. Camera traps situated within Majete Wildlife Reserve with potential home range centres for leopard and hyena. Suitable habitat is demarcated in green and unsuitable habitat in red. State space boundaries were created using the ½ MMDM buffer for both species as well as a 15 km buffer for analysis in SPACECAP.

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6.4 Results

Capture details

A total of 1920 trap nights resulted in a trap success of 0.99 leopards/100 trap nights (Table 6.1). Leopard capture locations are presented in Appendix 6E. In total, 19 leopard captures were recorded throughout the sampling period. This consisted of nine individuals of which five were recaptured. The leopard sex ratio was 1:3.5. In total, seven adults and two sub-adults were captured.

A total of 39 hyena captures were recorded throughout the camera trapping survey, which resulted in a trap success of 2.03 hyenas/100 trap nights (Table 6.1). We recorded a total of 16 individuals, including ten recaptured individuals, at 19 camera trap locations (Appendix 6E). The sex ratio of hyenas was 1:1, although four individuals could not be sexed. Three communal dens and two natal dens were found between 2014 and 2016 (Retief, 2016; this study), which includes three clans. Mean clan size was 5.33 (± 0.67 [SE]) members with two transient males.

Table 6.1. Capture details of leopard and hyena during the three-month camera trap survey in Majete Wildlife Reserve, Malawi.

Sex Total Individuals Individuals Captures/100 Species M F U ratio Subadults captures captured recaptured trap nights (M:F)

Leopard 19 9 5 0.99 2 7 - 1:3.5 2

Hyena 39 16 10 2.03 6 6 4 1:1 5

Population abundance

Leopard

The C-R analysis (program CAPTURE) estimated Mo (selection criterion 1.0) as the best-fitting model for leopard, which is based on constant capture probabilities (Otis et al., 1978; Rexstad & Burnham, 1991). The estimated leopard population size using this model was 10 (± 1.30 [SE]), with a 95% confidence interval of 10–17 individuals (Table 6.2). However, this model is sensitive to violations of model assumptions (e.g. individual heterogeneity) of capture likelihoods (Otis et al., 1978). The jackknife estimator associated with model Mh also had a high selection (0.86). This model is usually recommended for carnivore studies (e.g. Karanth & Nichols, 1998) as it assumes individual capture heterogeneity, thus improving robustness against violation of model assumptions (Otis et al., 1978). 165

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This model resulted in a population size of 11 (± 3.75) with a 95% confidence interval of 10–31 individuals, which appeared extremely large compared to that estimated by the Mo model. Capture probabilities (p-hat) were relatively low, but consistent for both models, (0.17–0.21). Population closure assumption was confirmed by CAPTURE (z = –0.15, p = 0.44), but not by CloseTest (Chi-square: χ2 = 14.71, d.f. = 7, p = 0.04).

Closure tests Model Abundance Species (Selection 95% CI p-hat CAPTUREa CloseTestb ± SE criterion) z-value p-value χ2 p-value

Mo (1.00) 10 ± 1.30 10–17 0.20 Leopard -0.15 0.44 14.71 0.04

Mh (0.86) 11 ± 3.75 10–31 0.17

Mo (1.00) 17 ± 1.27 17–23 0.23 Hyena -0.07 0.47 8.07 0.53

Mh (0.83) 19 ± 3.55 17–34 0.21 aCAPTURE – Closure test based on Otis et al. (1978). bCloseTest – Closure test based on Stanley & Richards (2005).

The SECR analysis (program SPACECAP) estimated leopard density, which was used to determine actual leopard population abundance within Majete. Leopard population size was 11.43 (± 2.72 [SD]) individuals, with a range of 9–17 (Table 6.3). Bayesian probability determines whether a model adequately describes the data, with intermediate values (near 0.5) being adequate and extreme values (near 0 or 1) being inadequate (Royle et al., 2009). Here the probability was 0.61 which indicated that the model adequately described the leopard data. Additionally, the Geweke diagnostic test showed that model parameters converged with z-values falling within the range –1.6 to 1.6 (sigma = 0.28, lam0 = 0.35, psi = –0.85, N = –1.09).

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Table 6.3. Results of population abundance and density estimates (individuals/100 km2) using SPACECAP for leopard and hyena in Majete Wildlife Reserve, Malawi.

95% Lower 95% Upper Bayesian Variable Mean SD HPD Level HPD Level probability

Leopard sigma 6965.20 6558.87 2712.97 14116.28 Buffer = 2.5 km lam0 0.01 0.006 0.00 0.02 State space = 995 km2 psi 0.13 0.044 0.06 0.22 0.61 N 12.20 2.94 9 18

density 1.63 0.39 1.20 2.40

sigma 5573.81 1704.78 3270.98 8584.82 lam0 0.01 0.01 0.00 0.02 Buffer = 15 km psi 0.18 0.06 0.08 0.30 0.62 State space = 3 020 km2 N 17.21 4.82 9 26 density 1.63 0.46 0.85 2.47 Hyena sigma 3692.92 506 2777.14 4694.15 Buffer = 3.9 km lam0 0.02 0.01 0.01 0.04 State space = 1 173 km2 psi 0.12 0.03 0.07 0.18 0.69 N 20.39 2.84 16 26 density 2.62 0.36 2.06 3.34 sigma 3743.94 522.62 2832.27 4812.65 Buffer = 15 km lam0 0.02 0.01 0.01 0.04 State space = 3020 km2 psi 0.17 0.04 0.09 0.24 0.69 N 28.37 5.09 19 38 density 2.69 0.48 1.80 3.60

aVariable: sigma – individual range parameter; lam0 – expected rate of encounter; psi – ratio of the number of individuals within state-space to total allowable number of individuals in the model; N – actual number of individuals in state-space

Hyena

The Mo model also performed best (selection criterion 1.0) in CAPTURE for hyena estimates (Table 6.2). The hyena population size using this model was 19 (± 3.55 [SE]) with a confidence interval of 17–34.

CAPTURE estimated Mh as the next most appropriate model (0.83). This model resulted in a population estimate of 17 (± 1.27), with a confidence interval of 17–23. Capture probabilities were slightly higher compared to those of leopards and the population closure assumption was met by CAPTURE (z = – 0.07, p = 0.47) and CloseTest (Chi square: χ2 = 8.07, d.f. = 9, p = 0.53).

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Using SPACECAP results we found a hyena population abundance of 18.36 (± 2.55 [SD]) with a confidence interval of 16–23 (Table 6.3). The Bayesian probability (0.69) and Geweke diagnostic statistics (sigma = 0.75, lam0 = –0.61, psi = 0.10, N = 0.01) suggested that the model was adequate in describing hyena data (Royle et al., 2009).

Density estimates

Leopard

Densities were calculated using effectively sampled areas (C-R) and state space areas (SECR), based on the ½ MMDM buffer method (2.5 km). Using CAPTURE, leopard density estimates were 1.52

2 2 leopards/100 km for the Mo model and 1.67 leopards/100 km for the Mh model, while the 95% confidence interval for both models ranged from 1.52–4.71 leopards/100 km2 (Table 6.4). SPACECAP estimated leopard densities at 1.63 leopards/100 km2, with a 95% minimum of 1.20 and a maximum of 2.40 leopards/100 km2. Furthermore, density estimates showed little deviation from those obtained from the 2.5 km buffer after the 15 km buffer was added (Table 6.3).

Hyena

Densities were higher for hyenas than leopards in Majete. Using the Mh model in CAPTURE, we found a density estimate of 2.62 hyenas/100 km2, with a 95% minimum of 2.34 and a maximum of 4.68

2 2 hyenas/100 km (Table 6.4). The Mo model estimated a lower density of 2.34 hyenas/100 km . Interestingly, the SECR analysis estimated hyena density at 2.62 ± 0.36 hyenas/100 km2 which was almost identical to estimates using the Mh model. Like our findings for leopards, hyena density remained relatively stable using the different buffer methods (Table 6.3).

Table 6.4. Density estimates (number of individuals/100 km2) calculated for leopard and hyena using the programs CAPTURE and SPACECAP with the half mean maximum distance moved (½ MMDM) buffer method based on camera trapping. Density is presented with standard error for CAPTURE estimates and standard deviation for SPACECAP estimates.

Buffer Effectively sampled Density Species Model 95% CI width (km) area/ State area (km2) (individuals/100 km2)

CAPTURE (Mo) 2.5 658 1.52 ± 0.20 1.52–2.58

CAPTURE (Mh) 2.5 658 1.67 ± 0.57 1.52–4.71 Leopard SPACECAP (Trap 2.5 995 1.63 ± 0.39 1.20–2.40 response absent)

CAPTURE (Mo) 3.9 727 2.34 ± 0.17 2.34–3.17

CAPTURE (Mh) 3.9 727 2.62 ± 0.49 2.34–4.68 Hyena SPACECAP (Trap 3.9 1173 2.62 ± 0.36 2.06–3.34 response absent)

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Predator carrying capacity

Carrying capacity estimates were based on Hayward et al. (2007c) and we compared range-wide and site-specific prey preferences (see Appendix 6F). Carrying capacity increased for all three carnivores from 2012 to 2015 and range-wide estimates were larger compared to site-specific estimates (Table 6.5). Estimated population sizes and densities were high. Leopards had lower carrying capacity estimates compared to lions and hyenas. Jacobs’ index was used to determine preferred prey species. Using preferred prey species in 2015, carrying capacity averaged 25 (range = 22–29) for leopards, 56 (range = 53–59) for lions and 96 (range = 94–98) for hyenas. Across range-wide and site-specific estimates, we calculated an average of 47 (range = 22–71) leopards, 72 (range = 59–90) lions and 90 (range = 80–98) hyenas.

Table 6.5. Estimated carrying capacities of leopard, lion and hyena in Majete Wildlife Reserve, Malawi, using range-wide and site-specific preferred prey species and preferred prey weight ranges determined from Jacobs’ indices. Values indicate population sizes and values in parentheses represent densities (per km2). Calculations were based on aerial surveys conducted in 2012 and 2015 (see Chapter 5). Range-widea Majete Wildlife Reserveb Carnivore Preferred prey Preferred prey Preferred prey Preferred prey species species weight range species weight range 2012 2015 2012 2015 2012 2015 2012 2015 21 29 58 66 14 22 62 71 Leopard (0.03) (0.04) (0.08) (0.09) (0.02) (0.03) (0.09) (0.10) 40 59 78 90 38 53 77 88 Lion (0.06) (0.08) (0.11) (0.13) (0.05) (0.08) (0.11) (0.13) 60 98 69 80 63 94 78 90 Hyena (0.09) (0.14) (0.10) (0.12) (0.09) (0.14) (0.11) (0.13) aSee Hayward et al. (2007c) for further details bSee Chapter 5 for further details

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6.5 Discussion

Population size and density estimates

Leopard

The SECR analysis estimated 11.43 (± 2.72) leopards (range = 9–17) in Majete. Our results suggest that the population has probably doubled six years post-reintroduction which provides encouraging signs of population persistence and growth within the reserve. To our knowledge this is the first documented case of leopard population increases with evidence of reproduction following reintroduction into a reserve with little or no intraspecific competition (this study; Chapter 3).

We captured more leopard females than males (ratio = 1:3.5) in our survey. This high capture rate of females characterises natural spatial patterns, as male home ranges normally encompass several female ranges (Bailey, 1993; Mizutani & Jewell, 1998). Our camera trap grid was based on the minimum home range size in similar habitat (Smith, 1977), although leopards in Majete had larger than expected home ranges, due to low intraspecific density (see Chapter 4). This resulted in an intensive camera trap design, with several camera traps within each female home range. Additionally, cameras were not only placed along favoured male paths (i.e. roads), but also on trails and riverine paths which females frequently use as they criss-cross their range (Balme et al., 2009). Cameras placed on trails and riverine paths (28%) accounted for 38% of female captures and zero male captures.

We recorded a density of 1.63 (± 0.39) leopards/100 km2. Densities were comparable to those from more arid areas (Bothma & le Riche, 1984; Stander, Haden, Kaqece & Ghau, 1997; Stein, Fuller, DeStefano & Marker, 2011). However, leopards were only recently reintroduced and since higher leopard densities occur in areas with higher prey densities (Bailey, 1993; Hayward et al., 2006; Hayward et al., 2007c; le Roux & Skinner, 1989), we expect the leopard density to increase due to the availability of prey in Majete (see Chapter 5). Leopards are extremely adaptable in terms of their behavioural and dietary flexibility, and their ability to cache prey items enables coexistence with high densities of sympatric carnivores (Balme, Miller, Pitman & Hunter, 2017).

Hyena

Hyena population size in Majete was estimated at 18 (± 2.55) using SECR analysis. Mean clan size was 5.33 (±0.67), ranging from four to six members, which was consistent with findings from a previous study (Retief, 2016). Hyena clan size varies considerably across Africa and is influenced by local prey abundance and water availability, with smaller clans (similar to those found in Majete) occurring in more arid areas with low prey densities (Mills, 1990; Tilson & Henschel, 1986; Trinkel, Fleischmann & Kastberger, 2006). Another factor affecting the number of hyenas in a clan is habitat type. For example, average clan size in woodland habitat is 12, while in savanna habitat mean clan size is 47 (see Holekamp

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& Dloniak, 2010). However, considering that Majete is a predominantly woodland habitat with abundant prey and readily available water, average hyena clan size appears lower compared to both woodland habitat recorded elsewhere.

Carnivore group size may strongly influence the outcome of competition (Palomares & Caro, 1999) and ultimately on survival (Holekamp et al., 2012; Watts & Holekamp, 2009). There are several trade-offs associated with hyena clan size. Food acquisition, arguably the most important factor, is more effective in larger groups, as such groups tend to have numerical dominance and better resource defence against lions (Kruuk, 1972). For example, smaller groups sizes suffer more losses to lions by kleptoparasitism, due to the inability to recruit sufficient numbers to displace lions from kills (Trinkel & Kastberger, 2005). Larger groups also provide benefits of reduced predation and improved cub survival rates (Watts & Holekamp, 2009). However, increased clan size may also result in increased feeding competition; more individuals are present at larger kills, which reduces food provisioning for lower ranked females and their offspring (Kolowski, Katan, Theis & Holekamp, 2007), although this does not appear to influence juvenile survival (Watts & Holekamp, 2009). Another potential cost of larger group sizes is disease transmission within such clans (Watts & Holekamp, 2009). For these, smaller groups would be beneficial. As a result, benefits of large group size are not equally distributed in all hyenas, as seen in other social carnviores, such as African wild dogs (Lycaon pictus, Creel & Creel, 2015), as selective pressures vary across different ecosystems which influence cost-benefit trade-offs.

We hypothesise that hyenas in Majete maintain small clans to minimise resource competition and maximise fitness (see Honer et al., 2005). However, hyenas are well below carrying capacity and clan size is typically limited by prey density within a clan’s communal territory and territory size (Mills & Hofer, 1998). Therefore, with a further increase in both prey and lion density, we expect hyena clan size (and density) to increase, to compete with lions by kleptoparasitising lion kills and defending their own kills (Périquet, Fritz & Revilla, 2015). We could also expect a switch in foraging strategies from active hunting to scavenging due to the increased competition and risk of kleptoparasitism from lions (Périquet et al., 2015).

Similar to clan size, density was also low. Using SECR analysis, hyena density was estimated at 2.62 (± 0.36) hyenas/100 km2. Densities are usually lowest (0.9 hyenas/100 km2) in deserts such as the Kalahari and Namib Deserts (Mills, 1990) and highest (154–165 hyenas/100 km2) in the prey-rich savannas of Kenya and Tanzania (Kruuk, 1972; Watts & Holekamp, 2008). Our estimates were lower than those from a study conducted in 2014 (Retief, 2016); however, the study in 2014 used a poorly defined camera trap grid, small buffer width and lacked robust SECR analysis. Therefore, we attribute the lower density estimate to methodological inaccuracies in the previous study by Retief (2016), rather than an actual decrease in density over three years. Our findings were the lowest recorded for hyenas in woodland habitats and comparable to those from more arid areas (see Holekamp & Dloniak, 2010). 171

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It is plausible that decades of human-induced mortality may have limited hyena clan size and density directly or indirectly by depleting their prey base, since food availability influences hyena clan size and density (Holekamp et al., 2012). For example, Frank, Holekamp and Smale (1995) found that an increase in the Masai population (and their cattle) resulted in greater opportunities for hyenas to depredate cattle, which subsequently increased retaliatory killing of hyenas. Pangle and Holekamp (2010) recorded an increase in the number of hyena mortalities since 1988 in Kenya. In the Serengeti, Kenya, snaring and poisoning contributed approximately 8% to the mortality of breeding female hyenas (Hofer & East, 1995), whereas in Etosha, Namibia, >10% of the entire Etosha hyena population was killed on farmlands adjacent to Etosha (Trinkel, 2009).

However, it is now 14 years since prey species were reintroduced into Majete and these populations have increased drastically throughout the reserve. In addition, the law enforcement has also made a significant contribution to the reduction of poaching and snaring incidences within the reserve (T. Moyo, personal communication, May 28, 2016). Thus, we would also have expected the hyena population to display signs of growth. For example, in the Serengeti hyena populations doubled over a 25-year period, likely due to a marked increase in prey numbers (Hofer & East, 1995); similarly, the Ngorongoro Crate hyena population increase was linked to the increase of prey (Honer et al., 2005). In contrast to lions, hyena populations tend to recover slowly following local extinction, as observed in Kruger National Park, South Africa, which may be due to the philopatry exhibited by female hyenas (Henschel, 1986); however, evidence suggests that previously occupied clan-ranges may be rapidly reoccupied by lower ranked females if hyena density is high (Holekamp, Ogutu, Frank, Dublin & Smale, 1993). Hyenas show remarkable behavioural plasticity in response to anthropogenic change, evidently contributing to their ‘Least Concern’ conservation status (Holekamp & Dloniak, 2010; Pangle & Holekamp, 2010; Bohm & Höner, 2015). Interestingly, hyenas appeared more elusive than leopards, suggesting that they could be modifying their behaviour to minimise encounters with humans (Boydston, Kapheim, Watts, Szychman & Holekamp, 2003; Kolowski et al., 2007; Pangle & Holekamp, 2010), although further scientific investigation is required.

Management of large predators in Majete

To prevent ecological imbalances, managers frequently ask ‘how many predators can our reserve sustain?’ and subsequently manage populations according to these numbers (Miller et al., 2013). Based on the available biomass of preferred prey (Hayward et al., 2007c), all three apex predators in Majete were well below carrying capacity. However, where prey are both abundant and naïve, predator numbers, especially those of lions, can rapidly increase and exceed carrying capacity within a short period (Kettles & Slotow, 2009; Miller & Funston, 2014). This may also result in genetic inbreeding as males have longer tenure periods in reintroduced lions and inevitably mate with their relatives (Trinkel et al., 2010). 172

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Rather than managing the entire population, recent studies have suggested simulating natural population-regulatory events (Ferreira & Hofmeyr, 2014; Miller et al., 2013). For example, managers could manipulate male take-over events in lions by replacing pride males to mimic natural tenure periods, which will remove younger cubs and improve genetic variation. However, genetic viability of predators in small, enclosed reserves is questionable, particularly since a minimum of 50 prides with unrestricted dispersal is recommended to maintain genetic viability (Björklund, 2003). Therefore, for these reserves, this would only be possible if a network of reserves was created on a regional scale i.e. managed metapopulation approach (Ferreira & Hofmeyr, 2014; Hayward & Kerley, 2009; Trinkel et al., 2010).

The Malawian DNPW and AP currently manage three parks in Malawi, namely Majete, Liwonde National Park and Nkhotakota Wildlife Reserve, thus creating a network of managed reserves in the country. A carnivore management plan is currently being devised for these three reserves. We encourage the establishment of a managed carnivore metapopulation for lions, leopards and cheetahs, by initially sourcing animals from countries such as South Africa and Zimbabwe (C. Hay, personal communication, March 11, 2017) and subsequently managing these predator populations within Malawi. We recommend sourcing predators from adjacent countries (Mozambique, Zambia and Zimbabwe), as this would simulate more natural dispersal events, but only if these sourced animals are in ‘excess’ (as in many small reserves in South Africa; Miller et al., 2013) with no effect on free- roaming populations. This will establish a regional node (see Ferreira & Hofmeyr, 2014 for lions and Marnewick et al., 2007 for cheetahs) based on geographic genetic structure. Ferreira and Hofmeyr (2014) suggested that integration occur predominantly within a node and occasionally between nodes, which would mimic natural dispersal and gene flow, enhancing population viability (Miller et al., 2013). Additionally, Nkhotakota has the largest resident lion population in Malawi (n = 18) (Mésochina et al., 2010), and so this managed metapopulation approach may incorporate genes from some of the last remaining lions in Malawi. Apart from the reintroduced predators in Majete, cheetahs (n = 6) have already been reintroduced into Liwonde.

Large carnivores are known to disperse over great distances in search of suitable habitat, prey and mates (MacDonald, 1983; Sandell, 1989). We already know that lions can move into the unfenced Lengwe National Park (C. Hay, personal communication, March 11, 2017; Mésochina et al., 2010), presumably from the free roaming population in Mozambique (e.g. Jacobson, Cattau, Riggio, Petracca & Fedak, 2013). Given that Majete has a high prey abundance, suitable habitat and presence of potential mates, coupled with the ability of leopard and hyena to move in an out of the reserve (Chapter 4), we deduce that carnivores could immigrate into the reserve. Therefore, it is possible that leopard and hyena occurring in Lengwe or outside protected areas (in Malawi or Mozambique) could move into Majete.

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In theory, linking Majete and Lengwe via an ecological corridor will undoubtedly prove effective for large carnivore management (e.g. Palomares, 2001; Walker & Craighead, 1997). However, Lengwe is currently unfenced, mismanaged and poaching is common. Linking these reserves will thus require substantial management, law enforcement and funding, coupled with relocation of villages. The future connection of Majete and Lengwe is uncertain at this stage. However, the use of various reserves (in Malawi and Mozambique) as ecological stepping stones by larger predators, is exciting for gene flow into Majete predator populations, which could enhance population viability (Fattebert, Dickerson, Balme, Slotow & Hunter, 2013). Regardless, we know very little about the population size and density of carnivores outside of protected areas in Malawi and adjacent Mozambique (Bauer et al., 2015; Jacobson et al., 2016), including the distances that they can traverse through an unsuitable matrix of human-dominated landscapes without mortality.

6.6 Conclusion

The reintroduction of lions and leopards into Majete was successful from a breeding perspective (Griffith et al., 1989). Population sizes increased for both reintroduced felids. Lions in Majete now represent the second largest lion subpopulation in the country. Female leopards could exhibit prolonged parental care of up to 30–36 months in Majete, although this requires a more thorough investigation. The resident hyena population appears to have remained stable since the reintroduction of lion and leopard. Hyena clan size and density was very low considering that the reserve is mostly woodland habitat with a high prey abundance. These findings could be explained by decades of human persecution. Managing large carnivores in small, enclosed reserves has its limitations, and thus we identify the potential of connecting reserves either directly (e.g. stepping stone or corridor) or ‘virtually’ (e.g. translocation). We support decisions made by management to establish a managed metapopulation of large predators within Malawi (including Majete, Liwonde and Nkhotakota) and simulate natural dispersal processes to maintain genetic viability.

6.7 Acknowledgements

Thank you to the Earthwatch Institute for funding this research study. Thank you to Dr Matt Hayward for providing me with formulae to calculate predator carrying capacity. Thank you to Charli de Vos, Erik Nyman and all the game scouts for your assistance in the field.

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Stander, P.E., Haden, P.J., Kaqece, I.I. & Ghau, I.I. (1997). The ecology of asociality in Namibian leopards. Journal of Zoology, 242(2), 343–364. DOI: 10.1111/j.1469-7998.1997.tb05806.x

Stanley, T.R. & Richards, J.D. (2005). Software review: a program for testing capture-recapture data for closure. Wildlife Society Bulletin, 33(2), 782–785. DOI: 10.2193/0091- 7648(2005)33[782:SRAPFT]2.0.CO;2

Staub, C.G., Binford, M.W. & Stevens, F.R. (2013). Elephant herbivory in Majete Wildlife Reserve, Malawi. African Journal of Ecology, 51(4), 536–543. DOI: 10.1111/aje.12064

Stein, A.B., Fuller, T.K., DeStefano, S. & Marker, L.L. (2011). Leopard population and home range estimates in north-central Namibia. African Journal of Ecology, 49(3), 383–387. DOI: 10.1111/j.1365-2028.2011.01267.x

Tambling, C.J. & du Toit, J. (2005). Modelling wildebeest population dynamics: implications of predation and harvesting in a closed system. Journal of Applied Ecology, 42(3), 431–441. DOI: 10.1111/j.1365-2664.2005.01039.x

Tilson, R.L. & Henschel, J.R. (1986). Spatial arrangement of spotted hyaena groups in a desert environment, Namibia. African Journal of Ecology, 24(3), 173–180. DOI: 10.1111/j.1365- 2028.1986.tb00358.x

Trinkel, M. (2009). A keystone predator at risk? Density and distribution of the spotted hyena (Crocuta crocuta) in the Etosha National Park, Namibia. Canadian Journal of Zoology, 87(10), 941–947. DOI: 10.1139/Z09-082

Trinkel, M., Cooper, D., Packer, C. & Slotow, R. (2011). Inbreeding depression increases susceptibility to bovine tuberculosis in lions: an experimental test using an inbred-outbred contrast through translocation. Journal of Wildlife Diseases, 47(3), 494–500. DOI: 10.7589/0090-3558-47.3.494

Trinkel, M., Ferguson, N., Reid, A., Reid, C., Somers, M., Turelli, L., Graf, J., Szykman, M., Cooper, D., Haverman, P., Kastberger, G., Packer, C. & Slotow, R. (2008). Translocating lions into an inbred lion population in the Hluhluwe-iMfolozi Park, South Africa. Animal Conservation, 11(2), 138–143. DOI: 10.1111/j.1469-1795.2008.00163.x

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Trinkel, M., Fleischmann, P.H. & Kastberger, G. (2006). Comparison of land-use strategies of spotted hyenas (Crocuta crocuta, Erxleben) in different ecosystems. African Journal of Ecology, 44(4), 537– 539. DOI: 10.1111/j.1365-2028.2006.00658.x

Walker, R. & Craighead, L. (1997, July). Analyzing wildlife movement corridors in Montana using GIS. Paper presented at the Proceedings of the 1997 ESRI User Conference, California, U.S.A.

Watts, H.E. & Holekamp, K.E. (2008). Interspecific competition influences reproduction in spotted hyenas. Journal of Zoology, 276(4), 402–410. DOI: 10.1111/j.1469-7998.2008.00506.x

Watts, H.E. & Holekamp, K.E. (2009). Ecological determinants of survival and reproduction in the spotted hyaena. Journal of Mammalogy, 90(2), 461–471. DOI: 10.1644/08-MAMM-A-136.1

Weber, W. & Rabinowitz, A.R. (1996). A global perspective on large carnivore conservation. Conservation Biology, 10(4), 1046–1054. DOI: 10.1046/j.1523-1739.1996.10041046.x

Woodroffe, R. & Frank, L.G. (2005). Lethal control of African lions (Panthera leo): local and regional population impacts. Animal Conservation, 8(1), 91–98. DOI: 10.1017/S1367943004001829

Woodroffe, R. & Ginsberg, J.R. (1998). Edge effects and the extinction of populations inside protected areas. Science, 280(5372), 2126–2128. DOI: 10.1126/science.280.5372.2126

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6.9 Appendices

Appendix 6A. Examples of photographs used to identify individuals based on unique pelage patterns. The top inserts (a & b) depict two left side photographs of the same male, while the bottom inserts (c & d) show clear differences in left side photographs of an adult male and younger male respectively. The same method was used to identify spotted hyenas.

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Appendix 6B. Example of the input file 'Animal Capture Details' required by SPACECAP. A sample of the capture details are represented for leopard on the left (a) and hyena on the right (b). a b

LOC_ID ANIMAL_ID SO LOC_ID ANIMAL_ID SO 10 1 1 13 1 1 8 2 5 5 2 3 8 2 6 20 2 7 8 2 9 20 2 8 23 2 10 46 2 7 41 1 10 4 3 4 27 3 2 22 3 9 31 4 10 23 4 1 47 4 8 77 4 2 47 4 9 7 4 4 40 5 4 20 5 8 62 6 2 22 5 9 62 7 1 41 5 2 62 7 2 22 6 9 76 8 1 50 7 4 76 1 6 50 8 7 76 8 9 53 8 1 81 8 3 59 8 2 81 9 1 59 8 4

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Appendix 6C. Sample of the input file ‘Trap Deployment Details’ required by SPACECAP. Binary values indicate whether camera traps were functioning (1) or not (0).

LOC_ID X_Coord Y_Coord 1 2 3 4 5 6 7 8 9 10

1 677098 8243531 1 1 1 1 1 1 1 1 1 1

2 679120 8249901 1 1 1 1 1 1 1 1 1 1

3 681016 8250202 1 0 0 0 0 0 0 0 0 0

4 686887 8242848 1 1 1 1 1 1 1 1 1 1

5 686401 8241148 1 1 1 1 1 1 1 1 1 1

6 683009 8243549 1 1 1 1 1 1 1 1 1 1

7 683866 8250622 1 1 1 1 1 1 1 1 1 1

8 681458 8252630 1 1 1 1 1 1 1 1 1 1

9 678872 8253318 1 1 1 1 1 1 1 1 1 1

10 686173 8248919 1 1 1 1 1 1 1 1 1 1

11 682724 8245400 1 1 1 1 1 1 1 1 1 1

12 682111 8242789 1 1 1 1 1 1 1 1 1 1

13 677832 8246310 1 1 1 1 1 1 1 1 1 1

14 683573 8249037 1 1 1 1 1 1 1 1 1 1

15 687161 8246181 1 1 1 1 1 1 1 1 1 1

16 676983 8252336 1 1 1 1 1 1 1 1 1 1

17 686756 8243810 1 1 1 1 1 1 1 1 1 1

18 684722 8246219 1 1 1 1 1 1 1 1 1 1

19 681025 8246771 1 1 1 1 1 1 1 1 1 1

20 682283 8239939 1 1 1 1 1 1 1 1 1 1

21 679322 8239681 1 1 1 1 1 1 1 1 1 1

22 680609 8242024 1 1 1 1 1 1 1 1 1 1

23 678524 8247553 1 1 1 1 1 1 1 1 1 1

24 677058 8248723 1 1 1 1 1 1 1 1 1 1

25 679294 8229468 1 1 1 1 1 1 1 1 1 1

26 676305 8233382 1 1 1 1 1 1 1 1 1 1

27 675923 8236919 1 1 1 1 1 1 1 1 1 1

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Appendix 6D. Sample of the input file 'Potential Home-Range Centres' required by SPACECAP to determine potential home range centres. Binary values indicate suitable (1) and unsuitable habitat (0) for both leopards and hyenas.

X_COORD Y_COORD HABITAT

671098.9 8240639 1

671848.9 8240639 1

672598.9 8240639 1

673348.9 8240639 1

674098.9 8240639 1

674848.9 8240639 1

675598.9 8240639 1

676348.9 8240639 1

677098.9 8240639 1

677848.9 8240639 1

678598.9 8240639 1

679348.9 8240639 1

680098.9 8240639 1

680848.9 8240639 1

681598.9 8240639 1

682348.9 8240639 1

683098.9 8240639 1

683848.9 8240639 1

684598.9 8240639 1

685348.9 8240639 1

686098.9 8240639 1

686848.9 8240639 0

687598.9 8240639 1

688348.9 8240639 0

689098.9 8240639 0

689848.9 8240639 0

657598.9 8241389 1

658348.9 8241389 1

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Appendix 6E. Camera traps and capture locations of leopards and hyenas in Majete Wildlife Reserve, Malawi.

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Range-widea Majete Wildlife Reserveb Carnivore Preferred Preferred species Preferred prey species prey weight Reference Preferred prey species prey weight range (kg) range (kg) Leopard Bushbuck (Tragelaphus scriptus) 10–40 Hayward et al. Bushbuck 10–47

Common duiker (Sylvicapra grimmia) (2006) Common duiker

Impala (Aepyceros melampus) Nyala (Tragelaphus angasii)

Reedbuck (Redunca arundinum)

Lion Blue wildebeest (Connochaetes taurinus) 190–550 Hayward & Kerley Warthog (Phacochoerus africanus) 43–188

Buffalo (Syncerus caffer) (2005) Waterbuck (Kobus ellipsiprymnus) Lichtenstein's hartebeest (Alcelaphus Gemsbok (Oryx gazelle) lichtensteinii) Giraffe (Giraffa camelopardalis)

Plains zebra (Equus quagga)

Spotted hyena No preferred species 56–182 Hayward (2006) Warthog 43–218

Waterbuck

Plains zebra

aFrom Hayward et al. (2007) bFrom Chapter 5

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Chapter 7

Research findings and implications for African Parks Majete management

7.1 Overview

The successful reintroduction of lion (Panthera leo) and leopard (Panthera pardus) in Majete Wildlife Reserve (hereafter MWR), suggests that reintroduction is currently the most viable tool for re- populating large carnivores in well-managed protected areas in Malawi. Currently, lion, leopard and spotted hyena (Crocuta crocuta; hereafter hyena) populations are relatively small and well below carrying capacity. They displayed preferences for specific prey species, but their overall impact on prey numbers was low. Given the abundance of prey in the reserve, we expect predators (especially lions) to increase rapidly within a short period. It is important to therefore continue monitoring predator- prey dynamics within the reserve to avoid major ecological imbalances. The small founder population of lion requires immediate attention, and additional translocations should be prioritised to maintain genetic diversity. It is plausible that leopards from outside have now entered the reserve, facilitating population growth and gene flow, although supplementing new individuals should also be considered. The resident hyena population appears resilient to anthropogenic change and does not currently warrant translocation, however continuous monitoring is recommended. Finally, we strongly encourage the establishment of a managed carnivore metapopulation within Malawi, to enhance genetic integrity and thus maintain population viability. In this report we describe, in detail, the major findings from our study, guidelines for future translocations and implications for carnivore management in MWR and Malawi.

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7.2 Research findings

7.2.1 Reintroduction success – importance of the pre- and post-release phase

Previous studies found that carnivores were difficult to translocate successfully, largely due to their tendencies to return to their capture site i.e. homing behaviour (see Linnell, Annes, Swenson, Odden & Smith, 1997). However, it is now clear that keeping large carnivores in a boma prior to their release (pre-release phase/soft-release), increases the likelihood of reintroduction success (Hunter et al., 2007). In addition, post-release monitoring has often been neglected in carnivore reintroduction projects, which is essential to determine whether a reintroduction was successful or not (Hunter, 1998). Here, we report on the pre- and post-release phase of reintroduced lion (Panthera leo) and leopard (Panthera pardus) in MWR. Based on previous studies (Griffith, Scott, Carpenter & Reed, 1989; Weise et al., 2015; Yiu, Keith, Karczmarski & Parrini, 2015), we defined reintroduction success according to the following criteria: (1) reduced movement rates; (2) reduced home range expansion; (3) no homing tendencies towards the capture site; (4) post-release survival for at least one year; and (5) defined breeding events (i.e. first wild-born generation or a three-year breeding programme).

Lions and leopards showed a general stabilisation of daily movements and reduction of home range sizes during the early post-release period. Reintroduced felids did not display consistent movements towards their home direction. Consequently, all Individuals showed release area fidelity and established permanent ranges within the reserve. All individuals survived for at least one year. After six years in MWR, the single lion pride had a total of four litters, consisting of eight cubs, one of which was a cub from the offspring of the reintroduced female. For leopards, a total of five litters were recorded, of which three were from reintroduced females. Therefore, our findings indicated that the reintroduction of lions and leopards into MWR was successful. We recommend the use of pre-release techniques (e.g. boma training, see Appendix 7A), as it clearly contributed to the successful establishment and persistence of reintroduced felids within the reserve.

7.2.2 A few big cats in a small reserve – insights from home range and habitat use analyses

Using GPS collar data we estimated home range size and habitat use of reintroduced lions and leopards in MWR. Mean ± SE home range for lions was 380.45 ± 117.70 km2 using 100% minimum convex polygon (MCP) and 134.00 ± 6.66 km2 using 95% kernel utilisation distribution (UD). Mean home range for leopards was 495.08 ± 80.99 km2 using 100% MCP and 257.94 ± 52.51 km2 using UD. To our knowledge, home ranges recorded for lion and leopard in MWR are the largest on record for reintroduced felids in Africa. Ranges for leopard were comparable to more arid areas such as the Kalahari Gemsbok National Park in South Africa/Botswana (Bothma, Knight, le Riche & van Hensbergen, 1997) and Kaudom Game Reserve in Namibia (Stander, Haden, Kaqece & Ghau, 1997),

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whereas lion ranges were similar to those found in Savuti in Botswana (McBride, 1990; Viljoen, 1993) and Hwange National Park in Zimbabwe (see Hemson, 2003). These large ranges were probably a result of low competitor density in MWR.

Both lions and leopards showed a general preference for riverine habitat. This habitat likely provided cover for hunting (Pitman, Kilian, Ramsay & Swanepoel, 2013) and predictable sites for encountering prey due to the presence of water, especially during the dry season (Hopcraft, Sinclair & Packer, 2005). Almost all individuals avoided high-altitude miombo woodland in the west of the reserve. However, miombo woodland was the furthest habitat away from the release sites of felids in MWR, which suggests that felids might not necessarily avoid this habitat.

7.2.3 Population persistence of apex predators – a look at population growth and density

We conducted an intensive camera trap survey between 8 September and 30 November 2016 within MWR. We used a combination of conventional capture-recapture (C-R, Rexstad & Burnham, 1991) and spatially-explicit capture-recapture (SECR, Gopalaswamy et al., 2012) analyses to estimate population size and density of leopards and hyenas (Crocuta crocuta) in MWR. Lions were excluded from this analysis as the population was regularly monitored and the number of individuals was known. Both C- R and SECR analyses generated similar results, although we used estimates based on the more statically robust SECR method.

We captured a total of nine leopards, including seven adults and two subadults, with a sex ratio of 1:3.5 (M:F). The SECR analysis estimated the leopard population size at 11 (range = 9–17). Sightings of cubs born after the camera trap study suggests that the population is now probably closer to the upper limit of the estimated population size, indicating that the population has at least doubled since their reintroduction. Interestingly, overlapping ranges of females and their subadult and adult offspring were recorded in this survey. This suggests either delayed dispersal, which usually occurs in prey-rich areas (Bailey, 1993), or the formation of matrilineal kin clusters, where females share a portion of their range with philopatric daughters (Fattebert et al., 2016).

A total of 16 hyenas were captured, including 11 adults and five subadults. A sex ratio of 1:1 was recorded, although four individuals could not be sexed. The hyena population was estimated at 18 (range = 16–23). Density (2.62 hyenas/100 km2) and clan size (mean = 5.33 ± 0.67) recorded in this study, are the lowest estimates in woodland habitats to date and comparable to those from arid areas (see Holekamp & Dloniak, 2010). Our results were slightly lower than a previous study conducted in 2014 (Retief, 2016), although we identify potential methodological inaccuracies in the latter. The hyena population has probably remained stable within the last two years. We recommend continued monitoring of the population, to determine whether their density is naturally low or whether it has

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been limited by decades of anthropogenic influences such as direct persecution or prey reduction due to poaching.

7.2.4 Predator-prey interactions – are prey limited by predators?

We reconstructed predator diets using scat analysis, while kill site analysis (using GPS location data) was used as a complementary method for lion diet. The dietary analysis revealed that lions and hyenas selected medium- to large-sized prey species. Based on Jacobs’ index values, lions preferred waterbuck (Kobus ellipsiprymnus; despite their apparent taste aversion), warthog (Phacocoerus africanus), kudu (Tragelaphus strepsiceros) and Lichtenstein’s hartebeest (Alcelaphas lichtensteinii). Hyenas also preferred waterbuck and warthog, although they displayed the greatest preference for plains zebra (Equus quagga), based on Jacobs’ index values. We suggest that the selection of waterbuck and warthog was due to a high encounter rate (based on similar habitat usage), and high hunting success (based on relative ease of capture; Hayward, Hayward, Tambling & Kerley, 2011). Interestingly, lions and hyenas avoided sable antelope (Hippotragus niger) and buffalo (Syncerus caffer), despite being of the most abundant species, and falling within their utilised prey weight range (sable antelope). This could be due to morphological (e.g. horns), behavioural (e.g. increased vigilance through large herd sizes) or ecological (e.g. different habitat selection) predation-avoidance strategies (Hayward & Kerley, 2005), although further research is required. Leopards utilised small-to medium- sized prey which included nyala (Tragelaphus angasii), common duiker (Sylvicapra grimmia), bushbuck and reedbuck (Redunca arundinum).

The daily nutritional requirements and consumption rate of the three apex predators are as follows: (1) lion normally require 4.6–7.6 kg of meat/day (Viljoen, 1993) and the pride in MWR consume 65– 75 prey/annum (see Chapter 5); (2) leopard require 1.6–4.9 kg of meat/day (Bailey, 1993; Bothma & le Riche, 1986; Stander et al., 1997) and consume 40–60 prey per year (Bailey, 1993; le Roux & Skinner, 1989; Schaller, 1972); and (3) hyena require 3.8–4 kg of meat/day (Henschel & Tilson, 1988), but no data are available on the number of prey they consume per annum.

Based on the substantial increase in prey between 2012 and 2015, it appears that predator populations (including the resident hyena population) in MWR have had a minimal impact on prey thus far, despite their energetic requirements and consumption rates. Given the small founder population of reintroduced lion and leopard, it may take several years (or decades) for these predator populations to increase to a level where they begin to drastically influence prey populations in the reserve. However, lions recover rapidly following reintroduction into small, enclosed reserves. Their prolific breeding can result in them exceeding carrying capacity in a short period, drastically reducing available prey populations (Kettles & Slotow, 2009; Hayward, O’Brien & Kerley, 2007). Additionally,

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the resident hyena population occurs at a low density (2.62 hyenas/100 km2), possibly due to decades of direct and indirect human influences. However, due to their behavioural plasticity in response to anthropogenic change (Pangle & Holekamp, 2010), coupled with improved law enforcement in the reserve, we expect the hyena population to recover.

Based on current prey populations, predator numbers in MWR are expected to increase, but this increase will be relatively slow, due to the small founder populations. Consequently, prey populations will continue growing rapidly which may negatively modify the habitat and plant species composition (Terborgh & Estes, 2010). The reserve should aim to maintain a healthy balance between the numbers of predators and prey, with neither exceeding carrying capacity. One way of dealing with rapidly growing prey populations is to translocate more predators into the reserve, especially those that specialise on prey species that are above carrying capacity. Another option is to translocate prey species to over reserves, which has already started in MWR, as elephant, zebra and varous antelope species have been moved to Liwonde Natiional Park and Nkhotakota Wildlife Reserve (African Parks (Pty) Ltd., 2017). Furthermore, stochastic events such as drought and fire could also regulate prey populations (e.g. Duncan, Aliénor, Chauvenet, McRae & Pettorelli, 2012). During a typical dry season in MWR, vegetation dies back significantly and surface water dries up with only a few natural and artificial water sources in the reserve. The decrease in forage and water impact prey populations to a variable degree during this period, although aerial census data show increases for all prey species between 2012 and 2015 (both conducted during the dry season). However, if conditions become unfavourable (e.g. more erratic or lower than normal annual rainfall), prey populations may be negatively impacted.

7.2.5 Competition for resources – evidence of dietary separation

Large carnivores compete for similar resources, which often leads to avoidance strategies to minimise direct competition (Durant, 1998). Based on GPS collar data, we recorded a high degree of spatial overlap between reintroduced leopards and lions (cf. Cristescu, Bernard & Krause, 2013) and similar habitat preferences. Although no collar data were available for hyena, camera trap photographs and infrequent sightings indicated that there was evidence for spatial overlap with reintroduced leopard and lion.

In contrast, the dietary analysis revealed a clear dietary separation by leopards. Leopards occupied a dietary niche substantially lower than that of lions or hyenas. This suggests that leopards could actively be avoiding more dominant predators by selecting smaller prey species. However, leopards are smaller, solitary carnivores and will typically select prey in the 10–40 kg weight range, that are easy to catch and present low risk of injury (Hayward et al., 2006a). Therefore, their selection could also be

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attributed to the availability of suitable prey in MWR. Conversely, we also found a strong dietary overlap between lion and hyena (68.42%), but the broad dietary range of hyenas potentially reduces competition with lions, which feed almost exclusively on four species.

7.3 Management implications

7.3.1 Managing large carnivores in MWR

Lion

In MWR, the lion population grew from three to eleven individuals between 2012 and 2017. The size and genetic diversity of the founder population are considered the most important factors in mitigating genetic inbreeding, with subsequent translocations also being important (Trinkel et al., 2010). However, in MWR, the founder population was small (n = 3) and therefore management will need to focus energy and resources on translocating additional lions with diverse genetic origins and mimicking natural processes. Management plan to replace the male coalition and introduce four lionesses from Zimbabwe in February 2018, to establish an additional pride within the reserve (C. Hay, personal communication, March 10, 2017).

MWR is a small, isolated reserve and it is thus important to identify and address potential problems which may arise. For example, in various small, enclosed reserves in South Africa, lion populations increase rapidly after reintroduction and their high reproductive potential often results in excess lions i.e. individuals exceeding the carrying capacity (Kettles & Slotow, 2009; Miller & Funston, 2014). In response, researchers have recommended that managers mimic natural social dynamics at the reproductive (e.g. contraception of females), survival (e.g. removal of specific individuals) or dispersal (e.g. male takeover events) stages of lions (Ferreira & Hofmeyr, 2014; Miller et al., 2013).

However, MWR is a non-hunting/culling reserve and contraception is still a foreign concept in Malawi (C. Hay, personal communication, March 10, 2017). Therefore, targeting specific dispersal/survival stages, by using translocation to nearby reserves, is currently the most viable option in regulating the lion population. Based on Ferreira and Hofmeyr (2014), we describe the potential options for management.

Management options at the dispersal/survival stage:

1. Mimic subadult dispersal: Simulate natural dispersal of subadult males or females by removing and introducing subadults. Based on Kettles and Slotow (2009), subadults between 18–22 months old are ideal for these translocations because they:

 are fully weaned

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 are able to adapt to new environments and survive by themselves

 (males) are usually evicted from their natal pride and are in search of a vacant patch to establish a range or join a pride

2. Mimic high death rate of older individuals: This involves the removal of older males or females. Males can live to 15 years and females up to 19 years, with noticeable declines in reproductive output by 14 years (Packer, 2010). Therefore, to reduce population sizes in the reserve, management could remove males and females >10 years, depending on the situation.

3. Mimic male takeover events: Manipulating male takeovers involves the translocation of male coalitions within or between reserves. This is an effective way of reducing the potential for inbreeding with close relatives (Kettles & Slotow, 2009).

 Firstly, management should dictate the length of male tenure. Tenure length in large reserves typically lasts for two to three years (Nowell & Jackson, 1996; Packer et al., 1988). However, in small, enclosed reserves, tenure length extends up to nine years, inevitably leading to inbreeding (Trinkel et al., 2010). However, translocating lions every second or third year, is not feasible from a financial perspective (Hayward et al., 2007; Packer et al., 2013). Thus, considering that inbreeding usually occurs after five years (Trinkel et al., 2010), we advise management to replace male coalitions every four to five years to maintain higher levels of genetic variation (Miller & Funston, 2014).  Secondly, management should consider the timing of male takeovers. Newcomer males usually kill cubs <12 months i.e. dependent on lionesses (Miller & Funston, 2014). Therefore, if lion cubs are inbred or the population size is growing rapidly, new males should be translocated when cubs are younger than 12 months.

Given the present state of MWR’s lions population, we strongly advise the immediate replacement of the current male coalition with another coalition of two males. We advise management to source lions from nearby countries such as Mozambique to mimic natural dispersal events. However, ideally lions should be sourced from fenced reserves which already have ‘excess’ lions (Kettles & Slotow, 2009) that can be supplied with no negative impact on the source population. Thus, we also recommend sourcing these additional lions from the Lowveld/KwaZulu or Frontier (Cape)/Kalahari node to enhance genetic diversity, since the current, small pride was sourced from the North West node. This often provides the luxury of selecting for specific age, sex, group size and level of habituation (see Appendix 7A). However, when sourcing these lions, we recommend disease screening prior to their

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translocation to prevent the introduction of pathogens into MWR (Hunter et al., 2012). For example, lion populations in the latter reserves were severely affected by bovine tuberculosis (bTB) caused by Mycobacterium bovis, and although this bacterial infection has now essentially been bred out, it highlights the potential effects of pathogens on lion populations (Ferreira & Funston, 2010; Trinkel, Cooper, Packer & Slotow, 2011).

Leopard

The leopard population in MWR has doubled since their reintroduction and is now probably close to 17 individuals. The founder population was relatively small (n = 6), and we are unable to confirm whether leopards have entered the reserve from surrounding areas, although the distance travelled by one male leopard (LEM3) out of the reserve suggests that leopards can move through the human- dominated landscape and thus immigrate into MWR. We recommend that a study be conducted to assess the rates natural dispersal in and out of MWR (i.e. immigration and emigration), in order to provide an assessment of the genetic status of leopards in and around the reserve. This knowledge should aid management to determine whether additional leopard translocations are required. However, if supplementation of leopards was to be considered in the future, we recommend the following adaptations from Weise et al. (2015) and Ferreira and Hofmeyr (2014) targeted at the dispersal level.

Management options at the dispersal/survival stage:

1. Simulate dispersal of subadults: Management can mimic subadult dispersal by removing and introducing males and females. The inter-release interval of these individuals should be at least 18 months, to allow resident individuals to assimilate these newcomers. At this age, subadults usually disperse in search of an open patch, and given that leopards (especially males) have large ranges, we expect subadults to establish ranges either partly or entirely outside of the reserve. However, leopards are below carrying capacity and results from Chapter 4 suggest that leopard ranges (both males and females) may overlap substantially.

2. Simulate male takeovers: Mimic male takeover events by introducing and removing (especially dominant) males. Male tenure length is between four and five years (Balme et al., 2013). Therefore, we suggest male replacement to simulate natural male tenure length.

3. Simulate high death rate of older individuals: Mimic natural deaths of older male and female leopards by translocation out of the reserve. The lifespan of female leopards is 18.6 years (Balme et al., 2013), but males generally have a shorter lifespan than females. In addition, female reproductive output remains relatively constant throughout their lifespan and only at 16 years do they show noticeable declines (Balme et al., 2013). Therefore, if management

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need to control the leopard population and older individuals (especially females) could be targeted for translocation to other reserves.

Hyena

Hyenas were the last free-roaming large carnivore in the MWR area, with hyenas still occurring outside of the reserve, despite their conflict with humans (T. Moyo, personal communication, May 28, 2016). The resident hyena population in the reserve appears to be stable at 18 individuals (range = 16–23), and gene flow is probably facilitated by male dispersal (Watts, Scribner, Garcia & Holekamp, 2008). Hyenas display notable behavioural plasticity in response to human influence, and their elusiveness in MWR (personal observation) suggests that they have adapted their behaviour to minimise encounters with humans (e.g. Boydston, Kapheim, Watts, Szychman & Holekamp, 2003). Therefore, we do not currently recommend removal or supplementation of hyenas. However, we recommend that the be continually monitored to determine the impacts of reintroduced predators and human persecution outside of the reserve. Future studies should aim to establish immigration and emigration rates of hyenas in and around MWR to determine how gene flow may be influencing the hyena population.

7.3.2 Conservation status of lion in Malawi

Lions were once abundant throughout Malawi (Hayes, 1979). But by the 1950s, lions had declined drastically due to the high human density, coupled with habitat loss and prey base depletion (Mésochina et al., 2010). A survey conducted in 2010 estimated a total of 29 resident lions (excluding transient males) in Malawi (Table 7.1), with the greatest population found in Nkhotakota Wildlife Reserve (n = 18; Mésochina et al., 2010). These figures are of great concern for Malawian lions, as extinction seems inevitable, especially since the country-wide population trend continues to decline. However, these estimates excluded the population of lions in MWR (n = 11), which, based on Table 7.1 is currently the second largest population of lions in the country (assuming all populations remained the same).

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Table 7.1. Estimated lion population sizes in protected areas in Malawi.

Chardonnet IUCN CSG Mesochina et al. Region Protected area (2002) (2006) (2010) Northern Nyika National Park - <10 0 Vwaza Marsh - <10 5 Central Kasungu National Park 5 <10 6 Nkhotakota Wildlife Reserve 10 <10 18 Southern Liwonde Ecosystema 5 - 6b Liwonde National Park - <10 0 Mangochi Forest Reserve - <10 - Namizimu Forest Reserve - <10 - Majete Wildlife Reserve - - - Total 25 <70 29 (35c)

aLiwonde National Park, Mangochi Forest Reserve and Namizimu Forest Reserve bTransient lions from Mozambique cValue includes transient lions

The IUCN/SSC Cat Specialist Group (2006) established potential Lion Conservation Units (LCUs) in Malawi based on the remaining isolated lion populations. Mésochina et al. (2010) also documented cases of lions crossing international borders between Malawi and neighbouring countries. For example, lions from Niassa in Mozambique used a corridor to move through Liwonde National Park, Mangochi Forest Reserve and Namizimu Forest Reserve, Malawi, although these events are now rare (Mésochina et al., 2010). The reasons for the decline of lions in these areas have also clearly not been addressed. Furthermore, Björklund (2003) estimated that a minimum of 50 lion prides is required to maintain genetic variation. Therefore, we suggest that creating a regional node with nearby reserves (i.e. metapopulation approach), under effective management with links to other nodes (e.g. South African nodes, see Ferreira & Hofmeyr, 2014), will establish a potentially viable lion population in Malawi.

7.3.3 The managed metapopulation approach for carnivores in Malawi

In an attempt to protect wildlife, Liwonde National Park (548 km2) and Nkhotakota Wildlife Reserve (1 800 km2) in Malawi were recently included on the list of reserves managed cooperatively by African Parks (Pty) Ltd. and local government, making it the second and third reserve co-managed by these two parties in Malawi. For these two reserves, management have set comparable ecological objectives to those already achieved in MWR, starting with the reintroduction of elephant, zebra and various antelope species. Finally, when the prey abundance increases to a respectable level, predators will be

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reintroduced. This has already started in Liwonde, with the reintroduction of cheetah (Acinonyx jubatus) in May 2017.

From a large carnivore perspective, managing relatively small reserves in isolation has little conservation value, because these reserves are usually surrounded by a landscape matrix that is not conducive to removing fences to link reserves (Miller et al., 2013). However, a network of reserves does not necessarily have to be directly connected. A metapopulation is defined as the network of isolated areas with either natural or facilitated (via translocation) dispersal (Armstrong & Seddon, 2008). Therefore, MWR, Liwonde and Nkhotakota can be ‘virtually’ connected via the translocation of predators, thus a managed metapopulation.

The aim of this approach is to create a regional node within which translocations occur regularly (Figure 7.1), with infrequent translocations to other nodes (see Ferreira & Hofmeyr, 2014). The genetic origin of predators is an important consideration as a wider genetic diversity could minimise the risk of inbreeding (Miller et al., 2013). For example, lions reintroduced into MWR were sourced from Pilanesberg National Park and Madikwe Game Reserve in South Africa (North West node), although these lions originated from Etosha National Park in Namibia (Ferreira & Hofmeyr, 2014). Predators may be sourced from various areas in South Africa, Namibia and Botswana (see Appendix 7A). Despite this, the additional lions to be translocated to Malawi will probably be sourced from Zimbabwe (adjacent to Malawi). Sourcing animals from neighbouring countries such as Mozambique, Zambia and Zimbabwe seems attractive (especially as this would mimic more natural dispersal patterns), although this should only be considered if targeted individuals are considered ‘excess’ within the selected source-reserve. From this perspective, sourcing animals from South Africa is justifiable, as lions, for example, are superfluous in many small reserves and managers are willing to facilitate the translocation process. In addition, managers often have the luxury of selecting for particular individuals, groups sizes and level of habituation.

These translocations are costly, but not fundamentally more expensive that other methods to manage predators, with an estimated translocation cost per individual at $2 108 for leopard, $2 760 for cheetah and $ 1 672 for hyena (Weise et al., 2014). Cooperation between reserve managers is essential to establish procedures and schedules for simulated dispersal events to maximise gene flow and ultimately ensure population viability of large carnivores in Malawi.

To guide carnivore management in Malawi, we suggest that the proposed Malawi node be linked with already existing managed carnivore metapopulations in South Africa e.g. Lion Management Forum (LiMF) for lions, Endangered Wildlife Trust (EWT) metapopulation for cheetah and Wild Dog Advisory Group South Africa (WAG) for wild dogs. These managed metapopulations are well organised and

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could provide structure to the management of carnivores in Malawi and even act as a source population.

7.2.4 Which cats go where? – considering the reintroduction of cheetah into MWR

In Malawi, historic records show that certain large predators, such as lion, leopard and hyena were abundant (Hayes, 1979; Morris, 1996), while others such as cheetah were never considered common (Myers, 1975). However, humans had a major impact on wildlife in Malawi. Unrestricted hunting already started during the early colonial period and hundreds of predators were killed annually (Morris, 1996). Thus, making it difficult to identify the distribution and abundance of predators throughout the country prior to the massive reduction of wildlife. We know is that the reintroduction of lion and leopard into MWR was successful and it appears that cheetah have also established themselves within Liwonde National Park. However, if the managed metapopulation approach materialises, as researchers and managers we must ask: ‘which predator should go in which reserves?’ and ‘what is the likelihood of success for each reintroduction?’

For example, management are currently considering the reintroduction of cheetah into MWR (C. Hay, personal communication, March 11, 2017), which may have occurred historically in the reserve (Purchase, Mateke & Purchase, 2007). Firstly, we must consider the order of predator reintroduction. Since reintroduction is essentially a recolonisation event, Ferreira and Hofmeyr (2014) and Hayward et al. (2007) suggested that smaller carnivores (in this case cheetah) be reintroduced first to allow them to locate refugia within the given area prior to the reintroduction of larger, more dominant predators. Thus, simulating natural successional theory (Young, Chase & Huddleston, 2001).

Another factor to consider is habitat. The dense vegetation in MWR would not typically be defined as ‘ideal’ vegetation for cheetah, lacking grasslands and savannas (Myers, 1975). However, a recent survey conducted by the Endangered Wildlife Trust identified sufficient suitable habitat for cheetah in MWR (C. Hay, personal communication, March 10, 2017). In addition, cheetahs are adaptable and may utilise relatively dense woodland habitats (Broomhall, Mills & du Toit, 2003), despite the availability of more open areas, especially for resting and movement through a given area (Purchase & du Toit, 2000). Cheetah, like most carnivores, have a varied diet, but prefer prey within a weight range of 23– 56 kg (Hayward, Hofmeyr, O’Brien & Kerley, 2006b). MWR has sufficient prey within the preferred parameters of cheetah diet, although they would be competing with leopards for similar prey (see Chapter 5). To reduce competition with dominant predators, cheetah display facultative avoidance behaviour both spatially and temporally (Durant, 1998). Despite this ‘competition refuge’, cheetahs are often killed by these predators. Interestingly, Hayward et al. (2007) found that reintroduced, competitor-naïve cheetahs had a higher mortality rate than reintroduced cheetahs that lived in

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sympatry with lions prior to their reintroduction. Thus, interspecific competition, with established leopard, lion and hyena is likely to have the greatest impact on cheetah reintroduction success in MWR. Ultimately, we do not discourage the reintroduction of cheetah into MWR, but rather recommend intensive post-release monitoring to identify the success of the reintroduction and to manage the population accordingly.

Figure 7.1. Regional node within Malawi consisting of Majete Wildlife Reserve, Liwonde National Park and Nkhotakota Wildlife Reserve (Shapefiles, personal communication, African Park (Pty) Ltd.).

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7.4 Conclusion

African Parks and the Malawi Department of National Parks and Wildlife currently co-manage three reserves in Malawi, namely Majete Wildlife Reserve, Liwonde National Park and Nkhotakota Wildlife Reserve. The success of carnivore reintroductions in Majete and Liwonde is encouraging for the re- establishment of large carnivores in Malawi. We advise management to follow guidelines provided in this chapter for future carnivore reintroductions and/or translocations. We also recommend that carnivores in these reserves be managed as a metapopulation to maintain genetic diversity and ensure long-term population viability within the country.

7.5 References

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Bailey, T.N. (1993). The African leopard: ecology and behaviour of a solitary felid. New York, U.S.A.: Columbia University Press.

Balme, G.A., Batchelor, A., de Woronin Britz, N., Seymour, G., Grover, M., Hes, L., MacDonald, D.W. & Hunter, L.T.B. (2013). Reproductive success of female leopards Panthera pardus: the importance of top-down processes. Mammal Review, 43(3), 221–237. DOI: 10.1111/j.1365- 2907.2012.00219.x

Björklund, M. (2003). The risk of inbreeding due to habitat loss in the lion (Panthera leo). Conservation Genetics, 4(4), 515–523. DOI: 10.1023/A:1024737317695

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Boydston, E.E., Kapheim, K.M., Watts, H.E., Szykman, M. & Holekamp, K.E. (2003). Altered behaviour in a large African carnivore associated with increased human activity. Animal Conservation, 6, 207– 219. DOI:10.1017/S1367943003003263

Broomhall, L.S., Mills, M.G.L. & du Toit, J.T. (2003). Home range and habitat use by cheetahs (Acinonyx jubatus) in the Kruger National Park. Journal of Zoology, 261(2), 119–128. DOI: 10.1017/S0952836903004059

Duncan, C., Chauvenet, A.L.M., McRae, L.M. & Pettorelli, N. (2012). Predicting the future impact of droughts on ungulate populations in arid and semi-arid environments. PLOS ONE, 7(12), e51490. DOI: 10.1371/journal.pone.0051490

Durant, S.M. (1998). Competition refuges and coexistence: an example from Serengeti carnivores. Journal of Animal Ecology, 67(3), 370–386. DOI: 10.1046/j.1365-2656.1998.00202.x

Ferreira, S.M. & Funston, P.J. (2010). Estimating lion population variables: prey and disease effects in Kruger National Park, South Africa. Wildlife Research, 37(3), 194–206. DOI: 10.1071/WR09030

Griffith, B., Scott, J.M., Carpenter, J.W. & Reed, C. (1989). Translocation as a species conservation tool: status and strategy. Science, 245(4917), 477–480. DOI: 10.1126/science.245.4917.477

Hamilton, P.H. (1981). The leopard Panthera pardus and the cheetah Acinonyx jubatus in Kenya. Retrieved from Cat Specialist Group website: http://www.catsg.org/cheetah/05library/5_3_publications/H/Hamilton_1981_Leopard_and_Che etah_in_Kenya.pdf

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Hayes, G.D. (1979). Lions – man-eaters and others. Nyala, 5, 6–11.

Hayward, M.W., Henschel, P., O’Brien, J., Hofmeyr, M., Balme, G. & Kerley, G.I.H. (2006a). Prey preferences of the leopard (Panthera pardus). Journal of Zoology, 270(2), 298–313. DOI: 10.1111/j.1469-7998.2006.00139.x

Hayward, M.W., Hofmeyr, M., O’Brien, J. & Kerley, G.I.H. (2006b). Prey preferences of the cheetah (Acinonyx jubatus) (Felidae: Carnivora): morphological limitations or the need to capture rapidly consumable prey before kleptoparasites arrive? Journal of Zoology, 270(4), 615–627. DOI: 10.1111/j.1469-7998.2006.00184.x

Hayward, M.W. & Kerley, G.I.H. (2005). Prey preferences of the lion (Panthera leo). Journal of Zoology, 267(3), 309–322. DOI: 10.1017/S0952836905007508

Hayward, M.W., O’Brien, J. & Kerley, G.I.H. (2007). Carrying capacity of large African predators: Predictions and tests. Biological Conservation, 139(1), 219–229. DOI: 10.1016/j.biocon.2007.06.018

Hemson, G. (2003). The ecology and conservation of lions: human-wildlife conflict in semi-arid Botswana. (Unpublished Ph.D. dissertation). Oxford, U.K.: Oxford University.

Henschel, J.R. & Tilson, R.L. (1988). How much does a spotted hyaena eat? Perspective from the Namib Desert. African Journal of Ecology, 26(4), 247–255. DOI: 10.1111/j.1365-2028.1988.tb00977.x

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Hunter, L.T.B., White P., Henschel, P., Frank, L., Burton, C., Loveridge, A., Balme, G., Breitenmoser, C. & Breitenmoser, U. (2012). Walking with lions: why there is no role for captive-origin lions Panthera leo in species restoration. Oryx, 47(1), 19–24. DOI: 10.1017/S0030605312000695

Kettles, R. & Slotow, R. (2009). Management of free-ranging lions on an enclosed game reserve. South African Journal of Wildlife Research, 39(1), 23–33. DOI: 10.3957/056.039.0103 le Roux, P.G. & Skinner, J.D. (1989). A note on the ecology of the leopard (Panthera pardus Linnaeus) in the Londolozi Game Reserve, South Africa. African Journal of Ecology, 27(2), 167–171. DOI: 10.1111/j.1365-2028.1989.tb00940.x

McBride, C.J. (1990). Liontide. Johannesburg, South Africa: Jonathan Ball.

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Morris, B. (1996). A short history of wildlife conservation in Malawi (Occasional Papers No. 64). Edinburgh, U.K.: Centre of African Studies, Edinburgh University.

Myers, N. (1975). The cheetah Acinonyx jubatus in Africa (Monograph No. 4). Morges, Switzerland: International Union for Conservation of Nature and Natural Resources.

Nowell, K. & Jackson, P. (1996). Wild cats. Status survey and conservation action plan. Gland, Switzerland: IUCN/SSC Cat Specialist Group.

Packer, C. (2010). Lions. Current Biology, 20(14), 590–591. DOI: 10.1371/journal.pone.0005941.West

Purchase, G.K. & du Toit, J.T. (2000). The use of space and prey by cheetahs in Matusadona National Park, Zimbabwe. South African Journal of Wildlife Research, 30(4), 139–144.

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Purchase, G.K., Mateke, C. & Purchase, D. (2007). A review of the status and distribution of carnivores, and levels of human carnivore conflict, in the protected areas and surrounds of the Zambezi Basin. Bulawayo, Zimbabwe: The Zambezi Society.

Retief, F. (2016). The ecology of spotted hyena, Crocuta crocuta, in Majete Wildlife Reserve, Malawi. (Unpublished M.Sc. thesis). Stellenbosch, South Africa: Stellenbosch University.

Rexstad, E., & Burnham, K.P. (1991). User’s guide for interactive program CAPTURE. Abundance estimation of closed animal populations. Colorado, U.S.A.: Colorado State University.

Schaller, G.B. (1972). The Serengeti lion. Chicago, U.S.A.: University of Chicago Press.

Slotow, R. & Hunter, L.T.B. (2009). Reintroduction decisions taken at the incorrect social scale devalue their conservation contribution: the African lion in South Africa. In: M.W. Hayward & M.J. Somers (Eds.), Reintroduction of top-order predators (pp. 41-73). Oxford, U.K.: Wiley-Blackwell.

Stander, P.E., Haden, P.J., Kaqece, I.I. & Ghau, I.I. (1997). The ecology of asociality in Namibian leopards. Journal of Zoology, 242(2), 343–364. DOI: 10.1111/j.1469-7998.1997.tb05806.x

Stewart, A-M.E. (2006). The influence of predator-prey relationships and managerial actions on ungulate population dynamics in Madikwe Game Reserve. (Unpublished M.Tech. thesis). Tshwane, South Africa: Tshwane University of Technology.

Terborgh, J. & Estes, J.A. (2010). Trophic cascades: predators, prey and the changing dynamics of nature. Washington, U.S.A.: Island Press.

Trinkel, M., Cooper, D., Packer, C. & Slotow, R. (2011). Inbreeding depression increases susceptibility to bovine tuberculosis in lions: an experimental test using an inbred–outbred contrast through translocation. Journal of Wildlife Diseases, 47(3), 494–500. DOI: 10.7589/0090-3558-47.3.494

Trinkel, M., Funston, P., Hofmeyr, M., Hofmeyr, D., Dell, S., Packer, C. & Slotow, R. (2010). Inbreeding and density-dependent population growth in a small, isolated lion population. Animal Conservation, 13(4), 374–382. DOI: 10.1111/j.1469-1795.2009.00344.x van Dyk, G. (1997). Reintroduction techniques for lion (Panthera leo). In: J. van Heerden (Ed.), Proceedings of a symposium on lions and leopards as game ranch animals (pp. 82-91). Onderstepoort, South Africa: The Wildlife Group of the South African Veterinary Association.

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Watts, H.E., Scribner, K.T., Garcia, H.A. & Holekamp, K.E. (2011). Genetic diversity and structure in two spotted hyena populations reflects social organization and male dispersal. Journal of Zoology, 285(4), 281–291. DOI: 10.1111/j.1469-7998.2011.00842.x

Weise, F.J., Lemeris, J., Stratford, K.J., van Vuuren, R.J., Munro, S.J., Crawford, S.J., Marker, L.L. & Stein, A.B. (2015). A home away from home: insights from successful leopard (Panthera pardus) translocations. Biodiversity and Conservation, 24(7), 1755–1774. DOI: 10.1007/s10531-015-0895- 7

Weise, F.J., Stratford, K.J. & van Vuuren, R.J. (2014). Financial costs of large carnivore translocations – accounting for conservation. PLOS ONE, 9(8), e105042. DOI: 10.1371/journal.pone.0105042

Young, T.P., Chase, J.M. & Huddleston, R.T. (2001). Community succession and assembly: comparing, constrasting and combining paradigms in the context of ecological restoration. Restoration Ecology, 19(1), 5–18.

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7.5 Appendix

Appendix 7A. Guidelines for future reintroductions or translocations of large carnivores.

1. Planning

1.1 Identify suitable protected areas for reintroduction/translocation 1.2 Clearly define the project aim and objectives and analyse outcomes after project completion to determine project success or failure

2. Technical details

2.1 Pre-release stage: Keep translocated animals in bomas to allow acclimatize to their new environment  Boma size: o Cheetah = 100 m x 100m (minimum of 100 m x 50 m) and 2.4 m high, with a minimum of a 50 cm inner overhang (https://www.ewt.org.za/) o Leopard = 6 m x 6 m and 4 m high, fully enclosed (see Chapter 2) o Lion = 60 m x 80 m and 2.4 m high (van Dyk, 1997) o Hyena = Unknown, but see dimension for cheetah and leopard  Fencing: o Bonnox or diamond mesh (van Dyk, 1997) o Inner electric fencing is recommended with three live strands at 30 cm, 1 m and 2.5 m for lion and cheetah (Hunter, 1998). Note: The reintroduction of leopard in MWR did not have inner electric fencing. o Outer electric fencing required if reserve has resident predators  Time in boma: Depends on the reason for translocation, for example: o Ecological purposes: . Leopard = 7–24 days . Lion = 26 days (Chapter 2) o Eco-tourism purposes (Hayward et al., 2007): . Cheetah = 1–1.5 months . Leopard = 4–5 months . Lion = 1–1.5 months . Hyena = 4 months

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 Habitat in boma: Sufficient vegetation should be available to provide predators with sufficient shade and cover (van Dyk, 1997).  Food: Provide full carcasses twice per week (only game, no livestock). Carcasses should be within the preferred weight range of each predator and from within the recipient reserve. Here is a list of dietary requirements for each predator: o Cheetah = 2.5 kg of meat/cheetah/day (Stewart, 2006) o Leopard = 1.6–4.9 kg of meat/leopard/day (Bailey, 1993; Bothma & le Riche, 1986; Stander et al., 1997) o Lion = 4.6–7.6 kg of meat/lion/day (Viljoen, 1993) o Hyena = 3.8–4 kg of meat/hyena/day (Henschel & Tilson, 1988)  Water: Should always be available (e.g. a 1 m x 1m shallow trough with slightly higher edges to avoid soil and vegetation fouling; van Dyk, 1997)  Collar type: GPS/GSM collar type (for animals in MWR, we strongly recommend GPS/GSM collars rather than VHF collars, as tracking is in MWR is challenging and yields low quality data) 3. Biological considerations

3.1 Release site:  Select area with low conspecific (i.e. same species) density (Weise et al., 2015)  Select vacant areas based on home ranges of resident individuals  High prey concentration  Readily available water  Consider distance to the fence line

3.2 Consider release-site sex ratios (M:F):  Cheetah = 1:1 in Liwonde National Park  Leopards = 1:3.5 in MWR  Lions = 1:1 in MWR  Hyenas = 1:1 in MWR

3.3 Genetic origin and potential sources:  Cheetah o See EWT’s Cheetah Metapopulation Project. Available at: https://www.ewt.org.za/  Leopard

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o Ideally should be sourced from Mozambique, Zimbabwe or Zambia to mimic more natural dispersal  Lion o To mimic more natural dispersal, lions should ideally be sourced from adjacent countries (Mozambique, Zambia and Zimbabwe), although only ‘excess’ individuals should be taken o However, the lion metapopulation in South Africa is well organised (Ferreira & Hofmeyr, 2014; Slotow & Hunter, 2009) and lions are superfluous and could have the luxury of selecting specific animals: 1) Lowveld/KwaZulu node . Recommended: E.g. Phinda Private Game Reserve & Greater Makalali Private Game Reserve . Source population: Kruger National Park (wild/free-roaming) 2) North West node . Recommended: E.g. Pilanesberg National Park & Welgevonden Game Reserve . Source population: Etosha National Park (wild/free-roaming) 3) Frontier (Cape)/Kalahari node . Recommended: E.g. Addo Elephant National Park & Tswalu Desert Reserve . Source population: Kgalagadi Transfrontier Park (wild/free-roaming)

3.4 Translocation/Reintroduction management options: To manage carnivore populations in a naturally regulating system, we should mimic natural processes (Weise et al., 2015 for leopard; Ferreira & Hofmeyr, 2014 for lion and cheetah)  Cheetah 1) Simulate male dispersal by removing and introducing subadults 2) Simulate territory tenure 3) Simulate high rate death of older males through removal of old males  Leopard 1) Mimic dispersal by removing or introducing subadults = 18 months, for residents must have time to assimilate newcomers (i.e. inter-release interval) 2) Simulate male takeover by replacing a dominant male (to supplement new genes). Note: tenure length in was 4–5 years (Balme et al., 2013)  Lion

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1) Simulate dispersal of male or female by removing and introducing subadults 2) Simulate male takeovers = every 4–5 years 3) Simulate high rate of death of older males or females by removing old individuals  Hyena Does not currently require translocation 3.5 Thing to avoid:  Translocation of persistent livestock raiders (occasional livestock raiders are acceptable)  Translocation of Individuals habituated in the presence of humans  Translocation of young individuals (<18 months)  Continuous releases of animals (Hamilton, 1981; Athreya et al., 2011)  Short distance dispersal, especially for animals that move in and out fenced reserves (Weise et al., 2015)

4. Post-release management

4.1 Monitor translocated animals regularly: With GPS collars, animal movements are recorded one to several times per day (depending on the collar setttings). This allows management to determine:  Kills: Clusters of location points can be used to identify prey consumed by an animal  Births: A female with location clusters can also indicate potential dens for raising cubs  Escapes: Location points recorded outside reserve borders indicate that an animal has escaped  Deaths: Consecutive data points in the exact same location could mean the animal is dead or lost his collar 4.2 Establish a conflict resolution strategy 4.3 Formulate a compensatory agreement with surrounding villages for losses or damages

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