SOUTH AFRICA January – May 2004

SOUTH AFRICA

January – May 2004

Deedra McClearn Julie Coetzee Laurence Kruger Mike Smith Kinesh Chetty

Organization for Tropical Studies ACKNOWLEDGMENTS

The time and efforts of so many people went into the planning and

implementation of this OTS Kruger UCT Wits Duke program. To Everyone Involved: a metaphorical giant toast with a fine sparkling wine from the Cape. Duke University Bill Robertson at the Andrew W. Mellon Foundation provided vision, patience, and funding.

Kruger National Park

University of Cape Town

University of the Witwatersrand

ii STUDENTS

Blanchie Asberry Tammy Baudains Scott Briscoe P.O.Box 311374 [email protected] 1239 Vermont Ave Birmingham, Al [email protected] NW # 407 35231 Mobile: 083 483 9316 Washington, DC, 20005 [email protected] [email protected] [email protected] Tel: (202) 408 4932 Tel: (205) 798 1801 Laura Buckley Eric Caldera Michael Chazan 62 Maple Str 5522 Brook Hill 35 Greenbank Ave West Boylston, MA, 01583 San Antonio, TS, 78228 Piedmont, CA, 94611 [email protected] [email protected] [email protected] Tel: (508) 8352029 Tel: (210) 435 3347 Tel: (530) 304 1562 Fahiema Daniels Megan Eastwood Kyle Harris [email protected] 2417 N. Fremont Blvd [email protected] Mobile: 083 739 0409 Flagstaff, AZ, 86001 Tel: (011) 704 1446 [email protected] Tel: (928) 774 1124 Shannon Hatmaker Gareth Hempson Stephanie Johnson 1035 Lower Mill rd [email protected] 1017 Millard Rd Hixson, TN, 37343 Mobile: 072 2232 486 Stone Mountain, GA, 30088 [email protected] [email protected] Tel: (423) 877 0190 Tel: (770) 498 8533 Sally Koerner Zoë Layton Taryn Morris 3130 Bee Tree Ln 670 Keyser Run Rd [email protected] Signal Mountain, TN, 37377 Washington, VA, 22747 [email protected] [email protected] [email protected] Mobile: 082 334 4167 Tel: (423) 886 5730 Tel: (540) 987 9544 Tel: (011) 485 2361 Justine Norman Govan Pahad Jasper Slingsby [email protected] [email protected] [email protected] Mobile: 083 680 6772 Mobile: 072 156 2286 Mobile: 083 4060 581 Tel: (011) 706 7331 Tel: (011) 447 0992 Carla Staver Simon Thomson Ben Wigley 7477 Wise Ave [email protected] [email protected] St Louis, MO, 63117 Mobile: 082 798 0314 Mobile: 083 481 8829 [email protected] Tel: (011) 452 3370 Tel: (314) 647 5991

FACULTY Kinesh Chetty Julie Coetzee Laurence Kruger WITS/OTS WITS/OTS UCT/OTS School of A.P.E.S. School of A.P.E.S. Botany Dept, UCT Wits University Wits University P Bag, Rondebosch, 7701. P Bag 3, WITS, 2050 P Bag 3, WITS, 2050 [email protected] [email protected] [email protected] Mobile: 082 4226225 Mobile: 083 793 3032 Mobile: 083 784 4272 0000000 Deedra McClearn Mike Smith OTS Kruger National Parks/OTS US: P.O. Box 598 P.O. Box 33 Lemont, PA, 16851 Skukuza, 1350 RSA: P.O. Box 33 [email protected] Skukuza, 1350 Tel: (013) 735 4379 [email protected] Mobile: 083 447 6869 Mobile: 072 630 3369

iii

SCHEDULE FOR OTS KRUGER NATIONAL PARK SEMESTER 2004

WEEK DATES SITE AND COMMENTS COMMENTS 1 22 January Johannesburg (Thursday) -Students arrive and go immediately to Nylsvley 22-29 January Nylsvley -Orientation activities -Plant, insect, and savanna workshops -Faculty-led projects -Introduction to history of South Africa 2 30 January – 4 Wits Rural Facility February -Rural development projects -Visit to cheetah rehabilitation center 3-5 5-25 February Kruger National Park: Skukuza -Orientation to Park and wildlife (including game drives) -Introduction to conservation issues -Faculty-led projects -First independent research project -Day trips to Nelspruit on weekends (movies, shopping) -Cultural activities -Quizes and presentations 6-7 26-28 February Ladysmith -Anglo-Boer War Museum and battle sites -Zulu women’s craft workshop 29 February – 10 Kruger National Park: Shingwedzi March -Primary emphasis on conservation issues in Park -Faculty-led projects -Drafts for first independent research project due -Midterms and presentations 11 – 14 March Semester Break in Drakensberg Mountains (hiking and outdoor pursuits) 8-9 15 – 27 March Kruger National Park: Punda Maria -Archaeology and expedition to Thulamela -Second independent research project -Cultural activities in Venda community -Visit to elephant museum at Letaba on trip back to Skukuza 10 - 12 28 March – 3 Kruger National Park: Skukuza April -Science Networking Meeting in Skukuza (students participate) -Drafts for second independent research project due -Make preparations for Cape trip 4 – 20 April Train to Cape Town from Johannesburg -Table Mountain, Kirstenbosch Gardens, Robben Island, Boulders Beach, Cape Point De Hoop Nature Reserve -Fynbos vegetation and Indian Ocean intertidal zone -Faculty-led projects (sharks, mice, bats, serotiny, birds) 13-15 21 April – 5 Kruger National Park: Skukuza May -Finish final projects -Screening of two documentary films (Jill Kruger) -Symposium of final research presentations (29 April) -Final exams -Evaluations; course debriefing

iv OTS KRUGER PROGRAM VISITING AND CONSULTING FACULTY JANUARY–MAY 2004

Science Courses Kevin Balkwill, Wits (curator of Moss Herbarium, head of school APES*) Duan Biggs, UCT (bird conservation) Harry Biggs, KNP (adaptive management) William Bond, UCT (botanist / ecologist) Peter Buss, KNP (veterinarian, game capture specialist) Marcus Byrne, Wits (dung beetles) Vincent Carruthers (herpetologist and general wildlife specialist) Andrew Deacon, KNP (small vertebrate specialist) Abri duBuys, KNP (Shingwedzi research) Edmund February, UCT (botanist / ecologist / archaeologist) Angela Gaylard, KNP (Rivers Boundaries Project) Navashni Govender, KNP (fire ecology specialist) Rina Grant, KNP (northern plains, waterholes) David Jacobs, UCT (mammalogist / animal behaviorist) Christo Marais, Working for Water (scientific director) Jeremy Midgley, UCT (botanist / ecologist / head of Botany school) Mike Picker, UCT (entomologist) Danie Pienaar, KNP (director of scientific services) Kevin Rogers, Wits (rivers & savannas / water management) Mark Rountree, Wits (river assessment specialist) Mary Scholes, Wits (savanna ecosystems) Robert Scholes, CSIR (savanna ecosystems) Corrie Schoeman, UCT (bats) Justin Smith (environmental law) Robert Timm, U. Kansas (mammalogist, OTS education committee) Freek Venter, KNP (geology and soil specialist) Ian Whyte, KNP (elephant specialist) Eleanor Yeld (sharks and oceanography)

History and Culture Course Lara Allen, Wits (ethnomusicologist) David Bunn, Wits (cultural theorist, Dean of Arts School) Jane Carruthers, UNISA (historian, historian of science) Mark Collinson, Wits Rural Facility (infectious diseases) Claudia Ford, Wits (sociologist / rural development specialist) Jill Kruger, University of Natal (documentary film maker) Tsepo Mamatu (television personality) Sharon Pollard, Wits Rural Facility (Sabie Sands Rivers Program) Tara Polzer, Wits Rural Facility (immigrants) Paul Pronyk, Wits Rural Facility (Aids development program) Fiona Rogers (consultant on South African educational system) Zweli Sibiya (Wits, Zulu praise singer) Wayne Twine, Wits (director of Wits Rural Facility)

Consultants Kevin Balkwill, Wits (Head of Department, APES) Marcus Byrne, Wits (Head of Honours Program for APES) Jenny Day, UCT (Head of Department, Zoology) Ed February, UCT (Head of Honours Program, Botany) Jeremy Midgley, UCT (Head of Department, Botany) Justin O’Riain, UCT (Head of Honours Program, Zoology)

v MAP OF AFRICA

10°W 10°E 30°E 50°E MEDITERRANEAN SEA o Tunisia cc ro Mo Algeria West Libya Egypt Sahara

20°N 20°N Mauritania Mali Niger Sen Eritrea ega The Gambia l a Chad rkin Sudan Guinea-Bissau Bu o Fas Guinea G Benin Togo h Sierra Leone a Nigeria Ivory n Ethiopia a Central African Liberia Coast lia Republic a m Cameroon o Equatorial Uganda S Guinea o 0° g Kenya 0° Gabon n Democratic o Rwanda C Republic Burundi INDIAN of Congo Tanzania ATLANTIC OCEAN

OCEAN M Angola a l a Zambia w i e u r iq a b c m s Zimbabwe a a z g Namibia o a 20°S 20°S d 10°W Botswana M a 1000 0 1000 2000 Kruger M Swaziland National Kilometers Park South Lesotho Lambert Azimuthal Equal Area Projection centered on 0°N, 20°E Africa

Data: ESRI Cartography: Kruger National Park GIS Lab June 2001 10°E 30°E 50°E

vi MAP OF KRUGER NATIONAL PARK

vii TABLE OF CONTENTS

FRONT OF THE BOOK

Acknowledgments………………………………………………………………….... ii Students……………………………………………………………………………… iii Course schedule……………………………………………………………………... iv Visiting and consulting faculty……………………………………………………… v Maps………...……………………………………………………………………….. vi-vii Table of contents…………………………………………………………………….. viii-x Keyword index………………………………………………………………………. xi-xii

FACULTY FIELD PROBLEMS

Nylsvley

Nylsvley Tree Guide. Scott Briscoe and Kyle Harris (editors), Kevin Balkwill (resource person)…………………………………………………………..... 2-6 Variation in grasshopper species (Acridoidea) content across plant communities in Nylsvley Nature Reserve. Eric Caldera and Stephanie Johnson (editors), Mike Picker and Johnathan Colville (resource people)………….…………. 7-10

Skukuza

Measuring biodiversity using habitat as a surrogate indicator—An analysis of the GRADSEC methodology. Laura Buckley and Fahiema Daniels (editors), Andrew Deacon and Rina Grant (resource people)………………………… 12-19 The effects of fire on vegetation structure and habitat in broadleaf savannas. Carla Staver and Taryn Morris (editors), William Bond and Edmund February (resource people)……………………………………………………………. 20-33 The effects of the 2000 floods on channel type heterogeneity in the Sabie River, Kruger National Park, South Africa. Megan Eastwood and Gareth Hempson (editors), Mark Rountree (resource person)………………...….... 34-40 The habitat- based relationship between peak frequency and wing morphology in Microchiropterans of the southern Kruger National Park. Tammy Baudains and Govan Pahad (editors), David Jacobs and Corrie Schoeman (resource people)………………………………………………………………………. 41-45 Frog species diversity in three pans located in the Kruger National Park. Shannon Hatmaker, Sally Koerner, and Zoë Layton (editors), Vincent Carruthers (resource person)………………………………………………………...….. 46-51

Punda Maria

The use of matrix models for calculating sustainable utilization rates of natural resources. Justine Norman and Benjamin Wigley (editors) and Christo Marais (resource person)…………………………………………………... 53-60

De Hoop

Rhabdomys pumilio in the fynbos. Blanchie Asberry and Michael Chazan (editors), Deedra McClearn, Robert Timm, and Julie Coetzee (resource people)……. 62-64

viii An investigation into home range size, home range overlap, and use of runways in the striped mouse, Rhabdomys pumilio (Sparrmann) in the South Western Cape, South Africa. Jasper Slingsby and Simon Thomson (editors), Deedra McClearn, Bob Timm, Julie Coetzee, and Laurence Kruger (resource people)………………………………………………………………………. 65-69

INDEPENDENT PROJECTS

Skukuza

Mating strategies of the dung beetle Kheper nigroaeneus. Blanchie Asberry ……… 71-74 Dung beetle fidelity in Kheper nigroaneus. Zoe Layton and Simon Thomson……... 75-79 Male competition in pollinating and parasitizing fig wasps in Ficus sycomorus synconia. Stephanie Johnson……………………………………………….. 80-82 Navigation of nocturnal flying insects. Michael Chazan……………………………. 83-84 Dispersal ability of the neonate instars of Dactylopius opuntiae (Homoptera: Dactylopiidae), a biological control agent of Opuntia stricta (Cactaceae) and the implications for biocontrol in the Kruger National Park. Scott Briscoe, Kyle Harris, and Shannon Hatmaker……..……………………….. 85-88 The effect of varying fire regimes on ant diversity in a savanna ecosystem. Eric Caldera and Jasper Slingsby………………………………………………... 89-98 Termites and fire: The burning question. The effect of different fire regimes on termite activity. Fahiema Daniels…………………………………………... 99-106 Fire and bush nucleation in broadleaf savannas. Carla Staver……………………… 107-117 Biological nitrogen fixation nodules in legumes in three different sites in Kruger National Park. Sally Koerner, Taryn Morris, and Justine Norman………… 118-127 Ligno-tubers, obligate or facultative? Benjamin Wigley…………………………… 128-132 Root suckering dynamics of Dichrostachys cinerea. Gareth Hempson…………….. 133-139 Waterhole preferences among large mammalian herbivores in the Skukuza region of the Kruger National Park, South Africa, based on track evidence. Tammy Baudains, Megan Eastwood, and Govan Pahad…………..……….. 140-146 Spatial distribution of Ploceus velatus (Greater Masked Weaver) nests within the canopy. Laura Buckley……………………………………………………... 147-150

Punda Maria

The revenue generation of ecotourism and ecosystem services: two primary practices by South African rural communities. Scott Briscoe……………... 152-154 Structural defence of Acacias: hooks, spines and architecture. Gareth Hempson…... 155-165 An investigation into the fine-scaled variation in tree diversity and the varied architecture of Colophospermum mopane in the Mopaniveld. Benjamin Wigley and Jasper Slingsby.………………………………………………... 166-178 Biotic Determinants of soil characteristics in the Northern Plains of Kruger National Park. Carla Staver………………………………………………… 179-195 Assessment of the proportion and extent of elephant damage on Acacia nigrescens and Sclerocarya birrea in the Punda Maria area in the Kruger National Park. Laura Buckley, Shannon Hatmaker, Justine Norman, and Simon Thomson……………………………………………………………………. 196-207 The effects of elephant damage on Adansonia digitata distributions along a slope in the Kruger National Park, South Africa. Michael Chazan and Kyle Harris... 208-213

ix Herbivore density, impala group size, and vigilance of herbivores while feeding. Blanchie Asberry…………………………………………………………… 214-217 The effect of overlapping piospheres on landscape heterogeneity. Sally Koerner and Zoe Layton……………………………………………………………... 218-228 The impact of vegetation disturbance around water holes on rodent abundance and diversity in the north of Kruger National Park, South Africa. Tammy Baudains, Taryn Morris, and Govan Pahad……………………………….... 229-238 Reduced kin recognition in the success of a widespread ant: Lepisiota capensis. Eric Caldera………………………………………………………………… 239-244 Differences in dung beetle diversity on granite and basalt-derived soils and between periods of diel activity in Kruger National Park, South Africa. Stephanie Johnson, Fahiema Daniels, and Megan Eastwood………………. 245-261

BACK OF THE BOOK

South African terminology………………………………………………………….. 263 OTS Goldsworthy images…………………………………………………………… 264-289 Images from Nylsvley……………………………………………………………….. 290-293 Images from Wits Rural Facility…………………………………………………….. 294-295 Images from Kruger National Park………………………………………………….. 296-299 Images from Skukuza……………………………………………………………….. 300-310 Images from Ladysmith…………………………………………………………...... 311-313 Images from Shingwedzi……………………………………………………………. 314-317 Images from Punda Maria…………………………………………………………… 318-324 Images from the Train Ride…………………………………………………………. 325-326 Images from Cape Town…………………………………………………………….. 327 Images from Robben Island…………………………………………………………. 328 Images from De Hoop Nature Reserve……………………………………………… 329-331 Images from Skukuza (2)……………………………………………………………. 332-336

x KEYWORD INDEX abiotic factors 166 fire treatments 99 Acacia 155 flood effects 34 Acacia cyclops 62 Formicidae 89 accessibility 208 frogging 46 Acridoidea 7 fynbos 62 active partner 71 grass species 218 Agaonidae 80 grasshoppers 7 angle of light 83 group size 214 ants 89 habitat 41 Argentine ant 239 herbivore 214 aspect ratio 41 herbivory 20, 179 Baobab 208 heterogeneity 20, 218 biocontrol 85 home range size 65 biodiversity 12, 89, 245 Hymenoptera 89 biodiversity indicators 12 impala 214 broad-leaf 118 increaser II 218 broad-leafed savanna species 128 index 155 bush clumping 107 intraspecific aggression 239 Cape Floristic Region 65 invasive alien plants 62 celestial cue 75 Kambeni experimental burn plots 99 channel type 34 Kheper nigroaeneus 71, 75 colonial nesting 147 kin recognition 239 Colophospermum mopane 166 Kruger Park 229 competition 166 large herbivores 140 Dactylopius opuntiae 85 large infrequent disturbances 34 decreaser 218 legumes 118 defense 155 Lepisiota capensis 239 Demographic Bottleneck Model 133 lignotubers 128 density 196 local male competition 80 Dichrostachys cinerea 133 lunar navigation 83 diel activity 245 management policy 208 dispersal 85 mating call 46 disturbance 133, 166, 229 matrix modeling 53 diversity 166, 229 microchiroptera 41 dung 245 mixed feeders 214 dung beetle 71 monitoring 229 dung pat 75 Mopaniveld 166 ecosystem services 152 natural resource management 152 ecotourism 152 natural resources 53 elephant damage 208 nesting site selection 147 elephant impact 196 nitrogen fixation 118 facultative 128 nocturnal insects 83 feeding simulation 155 nodules 118 fig wasps 80 Northern Plains 179 fighting 80 nucleation 107 fine-leaf 118 obligate 128 fire 20, 107, 118 Opuntia stricta 85 fire regimes 89 pans 46

xi peak frequency 41 Shabeni experimental burn plots 99 piosphere 218 Sherman traps 229 piospheres 229 similarity 7 plant communities 7 small vertebrates 12 Ploceus velatus 147 soil nutrients 179 polarized light 75 soil type 245 population structure 196 South African frogs 46 predator avoidance 140 spatial distribution 147 preferences 140 species content 89 pronotum width 71 species richness 166 recover 196 stocking rates 53 reptiles 12 surrogate 12 resprouting 128 sustainable harvesting 53 revenue generation 152 synconia 80 rhabdom 75 termite activity 99 Rhabdomys pumilio 62 thorns 155 riverine ecosystems 34 tree damage 196 rodents 229 vegetation 179 root sucker 133 vegetation cover 140 runway utilization 65 vigilance 214 Sabie River 34 water availability 140 Scarabaeidae 71, 245 water hole 140, 218 scatter-hoarding 65 wing loading 41 sedimentation 34 wing morphology 41 seed predation 65

xii

Nylsvley Faculty Field Problems

Nylsvley Tree

Burkea Site (broad leaf) Sodic Site

Maroelakop Termetaria

Terminalia Burkea

Burkea Site (broad leaf)

• a. Leaves simple ...... 2 b. Leaves compound ...... 3

2. a. Leaves alternate ...... T. sericia b. Leaves opposite . . . . . C. apiculatum

3. a. Leaves bipinnately compound . . B. africanum

Common Names

T. sericia ...... Silver Cluster- Leaf B. africanum ...... Wild Seringa C. apiculatum ...... Red Bushwillow

Combretum

Diplorhynchus Lannea

Maroelakop

• a. Leaves simple ...... 2 b. Leaves compound ...... 3

• a. Leaves alternate ...... G. falvescens b. Leaves opposite ...... D. condylocarpon

• a. Leaves pinnatrly compound . . . L. discolor

Common Names

L. discolor ...... Live-long D. condylocarpon ...... Horn-pod tree G flavescens Sandpaper

Grewia

Peltaphorum Acacia nilotica africanum

Termiteria Site

• a. Leaves simple ...... Z. mucronata b. Leaves compound ...... 2

• a. Bipinnately compound . . . . . 3

• a. Spines absent ...... P. africanum b. Spines present ...... A. nilotica

Common Names

Ziziphus

Acacia Acacia

Sodic Site

1. a. Leaves simple ...... 2. b. Leaves compound ...... 3.

2. a. Leaves opposite ...... C. bisponsa

3. a. Spines hooked, mixed with some some straight. Straight ones often inflated . . A. luederitzii b. Thorns are in pairs at nodes, short and hooked, and mixed with thin long ones . . . A. tortilis

Common Names

Boscia albitrunca

6 Variation in grasshopper species (Acridoidea) content across plant communities in Nylsvley Nature Reserve

Category: Faculty Field Project Participants: Eric Caldera (editor), Stephanie Johnson (editor), Mike Picker (faculty resource person), Johnathan Colville (faculty resource person) Site: Nylsvley Nature Reserve

Key Words: Acridoidea, grasshoppers, plant communities, similarity

Abstract: Grasshoppers are the dominant invertebrate herbivore in South African savannas. Studies in multiple savanna ecosystems have indicated that plant species composition and plant morphological characteristics are two major factors affecting grasshopper habitat choice and assemblages. The purpose of this study was to determine if grasshopper assemblages mirror plant communities in Nylsvley Nature Reserve. Ten samples of 100 sweep-net sweeps were taken from six sites representing the major plant communities found in the reserve. Samples were counted, sorted into morphospecies, and diversity indices were calculated. The six sites did not differ in abundance or diversity; however, we found that no two sites shared more than 55% similarity in species makeup.

Introduction Grasshoppers are ecologically important parts of South African rangelands. Their functions include being primary consumers and generating and transporting nutrients. They constitute a large percentage of the biomass of above-ground phytophagous insects. Grasshoppers have been used as indicators of threatened habitats and of land use (Gebeyehu and Samways 2002). Most grasshopper species are highly mobile and can travel to and inhabit a variety of plant communities. Plant species composition and plant morphological characteristics are two of the major factors affecting grasshopper habitat choice (Prendini et al. 1996). In the Mountain Zebra National Park, Gebeyehu and Samways found that vegetation composition and structure, particularly grass height and percentage cover determine grasshopper assemblages (Gebeyehu and Samways, 2002). Grazing pressure and environmental characteristics, such as the rockiness of an area, are also important in determining grasshopper communities (Gebeyehu and Samways, 2002; Predini et al., 1996). Grasshoppers are members of the superfamily Acridoidea. Acrididae is the largest family within the super family Acridoidea, suborder Caelifera, order Orthoptera. Acrididae is a large family with a cosmopolitan distribution. Thirteen subfamilies, 146 genera, and 356 species have been recorded for Southern Africa. The purpose of this study was to determine if grasshopper assemblages mirror plant communities in Nylsvley Nature Reserve.

Methods Study Area The project took place from 22-25 January, 2004 at Nylsvley Nature Reserve. Nylsvely Nature Reserve covers 3,120 ha in mixed bushveld type habitat. A survey conducted by the South African National Programme for Environmental Sciences characterized the Nylsvley savanna ecosystem study area floristically and inventoried the plant communities. The survey floristically divided the vegetation into four major groups of communities: communities of the elevated sandstone and felsite areas; communities of termite mounds and of the flat bottomlands; a bottomland community on self-mulching, vertic soils; and communities of abandoned settlements. These four major groups were then further divided based on additional variation (Coetzee et al. 1976). Sampling Methods Six sites were selected representing major vegetation types and features. Samples were taken from a Burkea (B) broadleaf savanna, Rhus leptodictya and Combretum apiculatum broadleaf savanna (S1), and Cymbopogon plurinodis and Combretum apiculatum broadleaf savanna (S2), Erogrostis pallens grassland (G1), Tristachya rehmannii and Digitaria monodactyla grassland (G2), and an Acacia dominated sodic site (S). Sampling was conducted using sweep nets. Ten samples of 100 sweep-net sweeps were taken from each site, for a total of 1000 sweeps per site. A species accumulation curve was constructed using the ten sweeps from the Cymbopogon plurinodis and Combretum apiculatum broadleaf savanna site in order to verify that ten samples containing 100 sweep-net sweeps is an adequate sample size. Each 100 sweep-net sample was taken by running in a

7 line and continuously making sweeps in grass covered areas. The samples were placed into medium sized white garbage bags. The bags were securely tied close and frozen overnight to kill all of the insects. The Acridid grasshoppers were removed from each of the samples, sorted to morphospecies based on morphological characteristics, and counted. A reference collection of the grasshoppers was made in order to aid in identifying the different morphospecies. Both adults and nymphs were included in the study because there were insufficient numbers of adult grasshoppers. This was probably a result of the late summer rains. Analysis The total number of grasshopper species collected from each of the six sites sampled represents the total species richness, S, for each site. The total species richness, S, and the Shannon- Wiener diversity index, H’(loge), were calculated for each site. A cluster analysis (CA) and a multi-dimensional scaling (MDS) ordination plot were constructed from a similarity matrix to show the associations between collection sites and grasshopper species. In CA the location of each node on the similarity axis shows the level of similarity in grasshopper species content between the sites that join at that node. The Bray-Curtis coefficient was used to determine similarity. Coefficient values were then converted to percentages by dividing each entry by the total sampled abundance and multiplying by 100. MDS is an ordination technique that uses data from a similarity matrix such that species content and sites sampled are arranged in two- dimensional space. As distance between two sites increases, the similarity in species content between those sites decreases. All analyses were performed with Primer™5.12 software. Results A total of 64 Acridoidea species were identified from the six sites sampled at Nylsvley Nature Reserve. The average species richness for the six sites was 22.67 ± 4.46. Whereas sample site S1, Rhus lebtodictya and Combretum apiculatum, had the lowest richness, S=14, while site B, Burkea, contained the highest richness, S=26 (Table 1). The average abundance of individuals, N, collected for the six sites was 113.50 ± 46.10. Abundance was greatest at Savanna site S2, Cymbopogon plurinodis and Combretum apiculatum, N=165, and lowest at the Sodic site, N=42. Site S1 also had a low abundance, N=73, (Table 1). Values for the Shannon-Wiener diversity index, H’, ranged from H’=1.89, to H’=2.88. By the Shannon-Wiener index, the Sodic site was the most diverse, followed by the broad leaf savanna site (B) and grasslands, G1&G2, respectively. The lowest diversity occurred in the savanna sites, S1&S2 (Table 1). No two sites were more than 55 % similar in grasshopper species content. The sites labeled grassland and savanna, G1&G2 and S1&S2, overall, contained 45% similarity (Figure 1); however, the two grassland sites were not the most similar to each other, nor was the case for the two savanna sites. The two most similar sites in grasshopper species content were G2 and S2 (similarity=52%). Similarity of grasshopper species content was lowest for the sodic site with a similarity of 22% between the sodic site and all other sites. Similarity between site B and grassland and savanna sites was 30% (Figures 1& 2). Discussion Grazing intensity affects grasshopper guild structure in grasslands (Gebeyehu and Samways, 2002; Prendini et al., 1996). Here we show that grasshopper species content is also determined by vegetation communities. Of the six sites where acridoids were sampled, none shared more than 55% similarity in species content. One interesting aspect of the CA is that it showed a grassland (G2) and a savanna site (S2) as having the greatest similarity (Fig. 1). The fact that both the grassland and savanna sites fell into one clade indicates that grassland and savanna type communities may have similar acridoid species content. This may be explained by the fact that, while the savanna communities contain some trees, mainly of the genus Acacia, trees do not dominate the community and many of the grasses found in grassland can still persist in the savanna community. Another explanation is that grasshoppers with morphological traits that are adaptive against predation can persist in both of these communities. For example, any acridoidea species mimic vegetation. This characteristic, which reduces the risk of predation, is advantageous in the savanna and grassland communities. The sodic site displays the least similarity with the other sites (Fig. 1). In the sodic site there are fewer grasses and more shrubs and fine leafed trees. This may explain the high degree of dissimilarity between the sodic site and all other sites surveyed. From general observations, species collected in the sodic were less slender and less green than other species collected in grassy areas. Many of the species collected in the sodic site mimicked the soil and rock in the area. This morphological character may also reduce the risk of predation. Thus, the risk of predation may be a significant factor in determining grasshopper species content across vegetation communities. Shannon-Wiener diversity was similar across all of the sites sampled (Table 1); however, there were some outliers in abundance, N, and richness, S. The sodic site had the least abundance, N, of

8 individuals (Table 1). This is perhaps due to reduced plant biomass in the sodic site and consequently less available habitat for shelter and feeding. The lowest richness was found in site S1 (Table 1). Site S1 contained large amounts of turpentine grass relative to the other sites. Turpentine grass is know to have low palatability, low nutrient value, and is toxic to many grazers. The high density of turpentine grass in S1 may severely limit grasshopper richness to those species that are able to cope with this relatively toxic environment. Because these outliers in abundance and richness are directly linked to changes in vegetation structure, they support our hypothesis that grasshopper species content is determined by vegetation communities.

Acknowledgments We greatly thank Deedra McClearn, Laurence Kruger, and Julie Coetzee for their assistance and support in the field, with data anaysis, and in the classroom. Thank you to Blanchie Asberry, Tammy Baudains, Scott Briscoe, Laura Buckley, Michael Chazan, Fahiema Daniels, Megan Eastwood, Kyle Harris, Shannon Hatmaker, Gareth Hempson, Sarah Koerner, Zoe Layton, Taryn Morris, Justine Norman, Govan Phad, Jasper Slingsby, Carla Staver, Simon Thomson, and Ben Wigley for carrying out the field research.

Literature Cited Coetzee, B. J., F. Meulen, S. Zwanziger, P. Gonsalves, and P. J. Weisser. 1976. A phytosociological classification of the Nylsvley Nature Reserve. Bothalia 12: 137-160. Gebeyehu, S. and M. Samways. 2002. Grasshopper assemblage response to a restored national park (Mountain Zebra National Park, South Africa). Biodiversity and Conservation 11: 283-304. Prendini, L, L. Theron, K. Merwe, and N. Owen Smith. 1996. Abundance and guild structurea of grasshioppers (Orthoptera: Acridoidea) in communally grazed and protected savanna. South African Jounal of Zoology 31: 120-130.

9 Table 1. Species richness (S), abundance (N), and Shannon-Wiener diversity (H’), for grasshoppers (Acridoidea) in four localities differing in vegetation structure: two grasslands, (G1&G2), a sodic site (Sodic), two savanna sites (S1&S2), and a Burkae dominated site.

S N H'

G1 25 134 2.67

G2 25 133 2.45

Sodic 22 42 2.88

S1 14 73 2.06

S2 24 165 1.89

B 26 134 2.76

Figure 1. CA of similarity of Acridoidea species content in four localities differing in vegetation structure: two grasslands, (G1&G2), a sodic site (Sodic), two savanna sites (S1&S2), and a Burkae dominated site (B).

Figure 2. Ordination plot showing similarity of Acridoidea species content in four localities differing in vegetation structure: two grasslands, (G1&G2), a sodic site (Sodic), two savanna sites (S1&S2), and a Burkae dominated site (B). Among six sites the species richness, S, is represented by the size of the circle at the location of the site on the plot. As distance between two sites increases, the similarity in species content between those sites decreases.

Skukuza Faculty Field Projects

11 Measuring biodiversity using habitat as a surrogate indicator: an analysis of the GRADSEC methodology

Category: Faculty Field Problem Participants: Laura Buckley (co-editor), Fahiema Daniels (co-editor), Andrew Deacon (resource person), Rina Grant (resource person) Site: Skukuza, Kruger National Park, Mpumalanga Province, South Africa Key words: biodiversity, biodiversity indicators, surrogate, small vertebrates, reptiles

Abstract: The aim of this study is to determine the effectiveness of using habitats as a surrogate for species diversity of small vertebrates in the Kruger National Park. Vegetation structure and abundance of various habitats were assessed in five different catenal units. A scoring system was used to predict which reptile species would be found in these areas and a list of species found in each area and their relative abundances was generated. The next phase of this study will be to ground-truth the results and thereby determine the effectiveness of using this surrogate. We concluded by suggesting modifications to the methodology.

Introduction Measuring biodiversity in all its facets and fluxes poses many challenges to researchers. Extensive and thorough counts of species richness and abundance can be time and labour- intensive, costly, and impractical in the case of certain elusive species. Solutions to these problems include generating models, following indicator species, and using surrogate indicators to measure biodiversity in an area. The use of surrogates for measuring overall biodiversity in a particular area is an appealing method in that it reduces the cost and increases the feasibility of the study. Despite its attractiveness, several concerns have been voiced over this method. Firstly, assessment techniques tend to influence the amount of support given to indicator taxa as valid surrogates (Reyers and Van Jaarsveld, 2000). Secondly, many studies have found that the use of surrogates when selecting sites to be conserved results in a loss in overall biodiversity (Pantzer and Schwartz,1998; Pharo et al., 2000; Reyers et al., 2000; Virolainen et al., 2000; Juutinen et al., 2003). Also, particular surrogate selection can have a major effect on extent of the results. For example, Juutinen et al. (2003) used dead wood as a surrogate for biodiversity in boral forests and found that it may only be a suitable indicator for species associated with dead wood and not for overall biodiversity. They concluded that there is little alternative for a complete inventory of biodiversity if the goal is to maintain populations of all species in the area. Similar studies came to comparable conclusions, stating that there were trade-offs between the somewhat inconclusive results of surrogate use and the ease, reduced cost, and feasibility of the method. They maintain that the selection of an indicator depends on the goals and constraints of each particular study (Juutinen et al. 2003, Caro and O’Doherty, 1999; Reyers et al., 2000; Reyers and Van Jaarsveld, 2000). This study comprises the initial steps of a larger study aimed at using small vertebrate habitat frequencies as a surrogate for biodiversity along a catenal sequence in the Kruger National Park. The objectives are to test the methodology and collect data that will be added to the data pool for the larger study. Various habitats utilized by small vertebrates can be seen more easily than the species themselves, hence the benefit of using this surrogate. Habitats range from the landscape and geomorphological level down to the level of specific substrates. For example, certain species utilize only rocky outcrops while others are restricted to riffles in perennial rivers. And many small vertebrates utilize medium-sized holes while others are specific to the undersides of dead prostrate logs. In this phase of the study, the habitats focused upon were those preferred by the species of reptiles found in Southern Africa. Along with contributing to the data

12 pool for this study of biodiversity using habitats as surrogates, the second aim is to investigate the effectiveness of the methodology and of the surrogate itself and to make recommendations for improvement.

Methods The data for this experiment were collected along a catenal sequence in the Skukuza area of the Kruger National Park, Mpumalanga Province South Africa on the ninth and tenth of February 2004. Five different catena units were sampled for this experiment, i.e. the valley bottom, sodic site (this unit was sampled twice), seep-zone, mid-slope and the crest. The data collection in each of the catenal units was the same. A 100m transect was laid down to assess the general structure and the number of different habitats along the transect. The general structure includes recording the aspect, slope, soil, animal activity, signs of small vertebrates, mammal dung and termite activity. A habitat survey was done within a five meter band on both sides of the 100m transect line. Holes, rocks, standing logs, prostrate logs, live stems, brush habitats, and litter habitats were counted as potential habitats in the survey. A point survey was done at the end of the transect where the percentage grass, forbs and tree cover was recorded, number of small trees in a 15m radius (trees with a stem diameter of less than 15m are classified as small) and number of large trees in a 40m radius was counted. The height classes of the woody vegetation and the five most common tree species were recorded for each catenal unit. Another 100m transect was laid down and the process was replicated. After the 200m transect was completed we walked for another 100m to look for specialized habitats and sample them, if there were any. A scoring system designed by Andrew Deacon was used to determine the likelihood of finding a certain species in the sampled area. The scoring system works by assigning a score for each type of habitat which is determined by the frequency of the habitat along the transect. The scoring system then gives probability scores for species that are likely to be found in each of the habitats. The habitat score is then multiplied by the species probability score. The final score for species probability was compiled and a cut off score of five was chosen, we chose this number because we wanted to be conservative and prevent an underestimation of species richness. Therefore in each catenal unit, a list of species with a score higher than five was counted as present in the habitat. The species list will be used to investigate differences between the various catenal units sampled. Primer 5 was used to get the Shannon-Weiner diversity indices for each catenal unit.

Results General structure Vegetation structure was measured by looking at grasses as trees size classes. The grass communities in both transects of the sodic site, the seep zone and the mid-slope were dominated by ankle length grasses (Table 3). The valley bottom and crest grass communities comprised mainly of taller grasses (Table 3). The trees size class distributions of the point surveys were different in each catenal unit, the most commonly occurring size classes were the one to two meter size classes and the three to five meter size class (Table 3). Indications of the presence of other animals were greatest in the two sodic sites and the crest (Figure1). The Valley bottom had the highest number of trees, both large and small, and sodic site two had the least amount of both large and small trees in the point survey (Figure 1). The most common trees in all sites changed in each site (Figure1). Habitats The various potential habitats for reptiles that were sampled changed along the catenal sequence. Ground holes were absent in both transects of the sodic site and was variable, but present in the other sites (Table 4).

13 The most frequently occurring habitat in all six of the sampling areas was the live tree habitat. Most of the dead standing trees were either sold or had flaking bark, all of the catenal units except for sodic site two and the crest had dead standing trees (Table 4). Dead prostrate trees (all categories) were found in all of the sites except the sodic site ( Table 4). Our analysis produced 28 reptile species in total with variable distributions; eight of the 28 species are common in all of the catenal unitis (table1). All of the catenal units had similar Shannon Wiener diversity indices (table 2).

Discussion The data for which we had scoring guidelines were used to generate measures of predicted species richness and abundance in each of the catenal units. Using this technique, the similarities and differences in habitat frequencies between areas directly account for the eight species shared across all units and the variable distribution of the remaining twenty. Many of the data collected but not used in the scoring procedure will be utilized as the study proceeds. The results concerning common tree species will be useful in determining presence of habitat for other small vertebrates, such as birds, specific to certain tree habitats. It will also be worthwhile to examine the turnover between different terrain zones to search for correlations in the turnover of species. Analysis of the vegetation structure data gathered (i.e. grass cover, tree and shrub height structure) will help to determine if there is a correlation between structure and various small vertebrate communities. This structure affects their habitat in terms of predator avoidance and prey detection. Since we found evidence of larger mammal activities in these areas, it is necessary to further examine the impact they have on small vertebrate communities. For example, the abundances of species that act as predators and competitors for resources will affect the numbers and types of small vertebrates in an area. It would also be useful to examine the rate at which large mammals such as elephants and other herbivores create and destroy habitats by repeating these habitat surveys over many years. The scoring technique designed for this study takes into consideration the frequency of habitats within each terrain zone and the likelihood that each species will utilize a particular habitat. For instance, a species that frequently utilizes a habitat rarely found in a certain zone would score similarly to a species that rarely utilizes a habitat found frequently in a given zone. In this fashion, the final probability score reflects the probable relative abundance of a species within a zone. Through this method we were able to generate a list of predicted reptile species and abundances in the various catenal zones we sampled, although the accuracy of these predictions can only truly be assessed once the results are compared to a ground-truthing study. At this point, it is important to keep two things in mind. Firstly, the indices returned give values that are relative between zones assessed with this technique; however, they do not return true measures of biodiversity since the values representing abundance are relative probability scores and not actual measurements. Secondly, the cut-off point at 5 is a rather arbitrary number. A different cut-off point would drastically change the results. Hence we concur with Reyers and Van Jaarsveld (2000) that the technique used to assess the data can strongly influence the degree of confidence and weight given to indicators and recommend that a ground-truthing study is the only way to set a justifiable value. This study is the first step in a larger examination of the biodiversity of small vertebrates along a catenal sequence. The next step will be to ground-truth the results gathered from the habitat assessment and determine the effectiveness of the chosen surrogate. If the results are satisfactory, assessments of other small vertebrate habitats will be performed using the same scoring system as used with the reptiles. A more extensive sampling of all possible terrain zones will be performed at that time. In addition to these steps, we suggest that certain modifications be made to the methodology. Other indicators of habitat might be included, such as a vegetation structure profile and a densitometer measure of canopy cover. These methods will give a more standardized, objective assessment of habitat structure. Once assessed, these data will give a

14 clearer picture of the ideal habitat for various species as they pertain to predator avoidance and prey spotting from both an aerial and ground-level perspective. Pending the results from the ground-truthing study, we believe that the sensitivity of the scoring system may need adjustments. It would be useful to perform a survey of the range of abundances of various habitats across terrain zones to ensure that the scoring system accurately represents the true range of habitat abundances. For example, the scoring system for multi-stem brushwood gives the highest score (5) to areas containing greater than 10 individuals. However, we found several areas containing between 10 and 25 individuals, all of which received the same score as those with only 10, while areas with fewer were separated into narrow ranges. In this circumstance, the scoring gradations do not accurately represent the range of habitat abundance. We had similar concerns for the scoring of termite mounds, dead trees (standing and prostrate), and ground holes. Other than these suggested modifications, we believe the methodology to be effective for carrying out the next phases of the study.

Acknowledgments We would like to thank Andrew Deacon, Rina Grant, Mary Scholes, Laurence Kruger and the OTS students in the 2004 program.

Literature Cited Caro, T.M. and O’Doherty, G. 1999. On the Use of Surrogate Species in Conservation Biology. Conserv. Biol. 13: 805-814. Pantzer, R. and Schwartz, M.W. 1998. Effectiveness of a Vegetation-Based Approach to Insect Conservation. Conserv. Biol. 12:693-702. Pharo, E.J., Beattie, A.J. and Pressey, R. 2000. Effectiveness of using vascular plants to select reserves for bryophytes and lichens. Biol. Conserv. 96:371-378. Reyers, B. and van Jaarsveld, A.S. (2000). Assessment techniques for biodiversity surrogates. South African Journal of Science 96: 406-408. Reyers, B. and van Jaarsveld, A.S., and Krüger, M. (2000). Complementarity as a biodiversity indicator strategy. Proceedings of the Royal Society, London (B) 267, 505 513.

15 Table 1: List of common names and the probability scores for finding a species in each catenal unit. Seep common names Valley Zone Sodic I sodic2 Crest Midslope

1 Yellow throated plated lizard 6 6 2 Rufous beaked snake 6 9 9 9 9 15 3 Black headed centipede eater 21 32 30 22 30 36 4 Variable skink 8 20 15 25 15 23 5 Jones' girdled snake 16 10 10 6 10 16 6 Herald snake 6 0 6 7 Cape wolf snake 9 0 8 Two toed burrowing skink 9 0 9 Sundeval's writhing snake 15 12 12 12 18 10 Wahlberg's snake eyed skink 16 28 19 27 25 25 11 Striped skink 25 0 25 5 5 25 12 Variegated bush snake 10 0 10 10 13 Eastern tiger snake 20 8 18 8 18 14 Boomslang 25 8 23 23 15 Black mamba 25 26 23 12 26 35 16 Tree agama 10 0 10 10 17 Southern vine snake 15 0 15 15 18 Flap necked chameleon 25 12 22 12 8 22 19 Rock leguaan 15 0 15 15 20 Bushveld lizard 20 16 16 16 16 21 Lesser worm snake 15 24 21 15 21 21 22 Three-lined grass snake 15 20 18 10 18 18 23 South-eastern egyptian cobra 15 9 15 9 24 Brown house snake 12 12 12 12 25 Mfesi 15 9 15 9 26 Spotted shovel snout 9 9 9 9 9 27 Half banded garter snake 6 6 6 6 6 28 Bibron's thicktoed gecko 12 12 12

Table 2: Shannon-Wiener {H'(loge)} indices and species richness (S) for each site. Sample S H'(loge) Valley 22 2.998 Seep Zone 19 2.838 Sodic I 19 2.87 Sodic II 17 2.72 Crest 19 2.836 Midslope 25 3.106

16 Table 3. Comparison of the general structure of the area sampled for the Gradsec transects. Catenal unit Grass community structure (% grass) Tree size classes (height in meters) Signs of other animals Common Tree Species

ankle Combretum hereroense <1 Valley Bottom knee Impala, duiker, kudu and Albizia forbesii i 1- elephant Peltophorum africanum hip 2- Spirostachys Africana 3- Acacia nigrecens >5

>hip

ankle <1 Spirostachys Africana 1- Sodic site 1 knee Impala, giraffe and Acacia nilotica 2- hip 3- elephant Dichrostachys cinerea <5 Euclea divinorum > hip Acacia tortilis

ankle <1m Acacia nilotica Sodic site 2 knee 1-2m Giraffe zebra steenbok Acacia tortilis 2-3m hip impala, rodent Acacia burkeii 3-5m >hip Ziziphus mucronata <5m

Albizia forbesii ankle <1m Seep zone 1-2m impala, zebra, elephant, Peltophorum africanum knee 2-3m giraffe Spirostachys Africana 3-5m Acacia robusta > hip <5m Grewia flavescens

hip Balanites maughamii

ankle ankle knee knee Albizia forbesii i i Midslope hip hip elephant , impala, zebra Peltophorum africanum > hip > hip Spirostachys Africana Ziziphus mucronata Balanites maughamii

ankle <1m Albizia forbesii 1-2m Crest knee Elephant Balanites maughamii 2-3m hip Combretum apiculatum 3-5m Marula > hip <5m Dalbergia melanoxylon Terminalia sericea

17 Table 4. Different available habitats in the various catenal units. Catenal unit Ground holes Dead standing Dead Prostrate

3 1.2 3 2.5 1 2. Valley Bottom 2 1.5 0.8 2 1 0.6 0.5 1. 0 0.4 1 0.2 0. Smal Mediu Large 0 5x5c 10x10c 10x10c 0 Solid Holes Flaking Hollow Soli Hole Flakin Rotte bar 2. 2.5

2 Sodic site one none 2 1.

1 1.5

0. 1

0 0.5

Solid Holes Flaking Rotten 0 bark Solid Holes Flaking Rotten bark 1.6 Sodic site two none 1.4 1.2 1 none 0.8 0.6 0.4 0.2 0 Solid Holes Flaking Rotten bark 4. Seep-zone 4 none none 3. 3 2. 2 1. 1 0. 0 Small Mediu Large ( 5x5cm) (10x10cm) (10x10cm 3 ) 1.6 3 1.4 2.5 Midslope 1.2 2.

2 1.0 2 0.8 1.5 1. 0.6 1 0.4 1

0.5 0.2 0. 0 0 0 Small Medium Large > Solid Holes Flaking Hollow Soli Holes Flaking Rotten 5x5cm 10x10cm 10x10cm bark bark

3 Crest 3 none 2.5 2.5

2 2

1.5 1.5

1 1

0.5 0.5 0 0 Small Medium Large > Solid Holes Flaking Rotten 5x5cm 10x10cm 10x10cm bark

20 small trees large trees 15 number of trees of number

0 Valley Sodic site1 Sodic site 2 Seepline Midslope Crest Bottom catenal unit

Figure 1: Number of small and large trees in each catenal unit. The valley bottom has a higher number of both small and large trees.

19 The effects of fire on vegetation structure and habitat in broadleaf savannas

Category: Faculty Field Problem Participants: Carla Staver and Taryn Morris Site: Skukuza, Kruger National Park, Mpumalanga Province, South Africa

Key words: fire, herbivory, heterogeneity

Abstract: Fire in savannas is instrumental in creating heterogeneous vegetation structures and a patchwork landscape in savannas. The frequency and season of fire have a direct effect on plant community composition and vegetation structure. In turn this has implications for faunal composition and diversity, since vegetation provides both habitat and food for herbivores. This study found that burn regime has a marked effect on both the structure and composition of the floral components. Infrequently burnt sites have a higher proportion of larger trees and denser tree and grass cover. These sites were dominated by climax, decreaser and mostly palatable grasses. More frequently burned sites have fewer large trees and more small trees. Frequent fires continually knock back woody vegetation to small sizes. This fire trap results in a high proportion of seedlings and gullivers. Grass communities mainly comprise pioneer increasers of moderate to low palatability. Ant species richness increased with increasing canopy cover, possibly because more canopy results in a greater diversity of thermal niches. Bird communities closely followed variation in vegetation due to differing burn regiments. Bird communities in more frequently burned sites were more closely related to each other than less frequently burned sites. These findings support the policy of managers of Kruger National Park with respect to promoting patch heterogeneity. Heterogeneity of ecosystems does indeed promote biodiversity of fauna in all its natural facets and fluxes. Fire is a powerful tool for managers hoping to drive this heterogeneity.

Introduction The principle drivers of savannas have been a topic of debate among savanna ecologists for decades. Some maintain that the abiotic template, including rainfall and geologic substrate, regulates the occurrence of savannas and the coexistence of trees and grasses (). Others suggest that fire plays the foremost role in maintaining the balance between trees and grasses and preventing the dominance of one over the other (Bond and van Wilgen 1996). However, fire certainly does play a role in altering the shape of savannas, whether it is vital in maintaining the savanna itself or not (Van Wilgen et al. 2003). In Kruger National Park, where heterogeneity and its role in the resilience of complex adaptive systems have become of primary importance to management, fire has become a powerful tool for achieving a diversity of landscapes in the park (Van Wilgen et al. 2003). The frequency and season of fire have a direct effect on plant community composition and vegetation structure. The effects of these have been and are still being

20 well studied, both in Kruger National Park and elsewhere in savannas. Increasing fire frequency and intensity (correlated with burn season – winter, dry season burns are generally hotter and more intense than summer, wet season burns) are known to lead to decreasing density of woody vegetation, although the occurrence of gullivers significantly increases due to the demographic bottleneck (Bond and van Wilgen 1996). Furthermore, disturbance has marked effects on grass composition. We would expect annual burn plots, which are more disturbed, to have a higher proportion of pioneer and subclimax species, which colonize disturbed areas more successfully than climax species (Van Wyk and van Oudtshoorn 1999). Moreover, we would expect the proportion of palatable grasses to decrease, and the proportion of increaser II grasses, which predominate in overgrazed veld, to increase in areas of more frequent burn (Van Wyk and van Oudtshoorn 1999). Previous studies have shown that grazing intensity increases in areas of annual burn (Van Wilgen et al. 2003), which would result in a change in grass species composition. We may be able to use grass species composition as an indicator of grazing intensity, and thus determine the extent to which the various burn plots are used by grazers. However, the effects of fire frequency and season on faunal composition are less well understood. It seems logical that a marked effect on plant species composition and vegetation structure would have serious implications for faunal composition and diversity, since vegetation provides both habitat and food for resident herbivores. The effects of vegetation structure on large herbivores has been studied to some extent in Kruger, although not exhaustively. Grass species composition, abundance, and height have a well documented effect on the occurrence of roan antelope and other type II grazers in the Northern Plains area around Shingwedzi. However, the preferences of large herbivores for varying vegetation types in more southern areas of the park are not as well known. Solitary herbivores and browsers such as kudu are known to prefer areas with more dense trees and generally thicker vegetation, where decreased visibility makes predator avoidance easier and where forage is readily available. Large herd antelope, such as impala, are thought to prefer more open areas, where they can more clearly see predators approaching. The dynamics of traveling alone versus in a large herd change predator avoidance tactics. Invertebrate populations have been less well studied in the Kruger than the more charismatic mega fauna. In tropical forests, temperate forests, and within individual trees in Australia, invertebrate (mostly insect) populations respond significantly to disturbance and to varying degrees with the intensity of disturbance; different insect groups responded differently, indicating a response to changes in habitat and food availability. Lepidoptera showed a decrease with decreasing canopy cover, while sap-sucking phytophages, including some hemipterans, showed an increase with decreasing canopy cover (Schowalter 1994; Schowalter 1989; Majer and Recher 1988). Data on other faunal diversity are not readily available but we expect to find distinct bird communities in the different burn treatments due to the differences in structure and species composition. We hoped to be able to define the effects of varying fire frequency and season on faunal communities. Vegetation data, including community composition of trees and grasses, foliage profiles, which give an estimate of biomass at various heights, and canopy cover, not only indicate the types of habitat available, but also may indicate the

21 types and intensity of usage in the various burn plots. However, we will also examine direct evidence of animal presence of various guilds, including mega-herbivores, birds, ants and lepidopterans.

Methods Kruger National Park management established a series of experimental burn plots (EBPs) in 1954. These burn plots are found in four areas of the park and within each of the four areas includes four burn strings. Each string consists of twelve burn treatments of various burn frequency and season (Van Wilgen et al. 2003). Our study was conducted in three of the four strings in the Pretoriuskop EBPs. Where possible we sampled the same burn season and frequency treatments across the three strings but due to herbivore exclusion experiments, access to some of the plots was restricted (Table 1). In each plot we ran one transect of 150 m, collecting a variety of vegetation structure and animal presence data. Within two meters of the transect, we the height and species of each tree. Every two meters along the transect, we noted the closest species of grass to the transect, with species, from which we calculated relative species abundance. Every fifteen meters along the transect, we calculated canopy cover and determined foliage profiles using boards 2m high and 10cm wide, divided into height classes of 25cm. We measured the distance from the point at which 50% of each height class (25 cm, 50 cm, 75 cm, 1 m, 1.5 m, and 2 m) on the board was obscured. The reciprocal of this distance is directly proportional to foliage density and is thus an approximation of the vertical distribution of biomass. Evidence of large herbivores (dung) was also assessed within 2m2 of the transect line. To determine butterfly species richness, all butterfly species within sight of the transect line were counted to yield a count of butterfly species richness. Bird species richness and ant diversity were determined independent of the transect. Bird species seen or heard were recorded for each site visited over a period of 1.5 hrs. Only the data from the first three samples were used, as activity of birds decreased in the warmer periods later in the morning. Ant diversity was determined using pitfall traps set out overnight in each of four burn plots (Numbi and Shabeni no burn and August annual burn plots). We set out two pitfall traps every 10 m along a 40 m transect, giving 8 replicates for each site. Data analysis Foliage profiles, size class distributions and grass characteristics were graphed using Microsoft Excel 2002. Multi-dimensional scaling community analyses were generated using Primer 5. Primer 5 was also used to determine the Shannon wiener indices. The relationship between time of day and bird diversity was generated in Statistica 6.1.

Results Vegetation As burning frequency decreases the proportion of trees in bigger size classes increases. In the frequently burned plots (annual and biannual), trees in height size class 1 (<1m) are most abundant with the other size classes having relatively low numbers (figures 1,2 and 3). This may be due to mortality as a consequence of frequent burning preventing recruits from establishing and maintaining the woody vegetation in a short multistemmed state.

22 No burn and low intensity February triannual burn plots have similar patterns of woody vegetation structure and compositional patterns. More frequently burned (annual, biannual and high intensity October triannual burns) sites were also closely related in composition (figure 4). Each burn treatment is compromises of different grass species in varying abundances burn treatments in different strings display closely related species compositions (figure 5). This demonstrates that grass communities have a distinct response to burning regiments. The control burn plots are dominated by decreaser grasses, which according to van Wyk and van Oudtshoorn (1999) are abundant in good veld and decrease when the veld is under or over grazed. The fact these decreasers are present indicates that they must be subject to a certain level of grazing in order to be maintained (Table 2).Intermediately burnt plots display a combination of grass responses but nevertheless have a high percentage of increaser I grasses which are usually unpalatable (Table 2). Annual burn sites are dominated by increaser II grasses. These increase due to disturbing effects and mostly include pioneer species which are generally unpalatable (Table 2) As burn frequency increases the grasses become predominantly unpalatable (Table 3). As expected, annual burns, which are more disturbed, have a higher proportion of pioneer and subclimax species (Table 4), which colonize disturbed areas more successfully than climax species. A trend seen in the foliage profiles indicates that as burn frequency decreases vegetation becomes more dense at the lower heights (figures 9, 10 and 11). As burn frequency decreases the density of vegetation at taller heights increases (figure 9, 10 and 11). In addition, the August triannual and February triannual burns show the greatest variation in vegetation density at all heights, indicating that regardless of burn season, a three year burn cycle may be most suitable for creating patch heterogeneity in vegetation at a very small scale. Fauna Although only four plots were sampled for ant diversity a pattern is still evident within these plots. Ant species diversity decreases as percent canopy open increases (figure 12). The number of species of birds sighted decreased as time progressed (figure 13). Although the data used was minimal it still showed that the bird communities in different burn treatments varied. The annual and biannual burn sites were only approximately 50% related to each other with regard to species composition, and were only approximately 20% related to the control plot (figure 14). Unfortunately, faunal data from biannual and triannual burn cycles yielded no results. Thus, it is impossible to determine whether the increased patch heterogeneity of habitats at a small scale evident in the vegetation structure of triannual burn plots has led to higher species diversity and a different faunal communities.

Discussion Results of the vegetation sampling indicate that burn regime has a marked effect on both the structure and composition of the floral components. Size class distributions for trees have shown that annually burnt sites generally have a much larger proportion of smaller individuals and many fewer large individuals, although there is some inter-site variation in size class distributions. Unfortunately, much of our data on faunal

23 composition yielded inconsistent and unreliable results, which makes comparison between expected animals present based on vegetation structures with animals actually present impossible. Based on the higher density of trees for browsing and the larger amount of biomass at taller heights, we conclude that the no burn and triannual burn areas are more appropriate for browsers than for grazers, but have no way of corroborating these findings. Grass communities in no burn areas were dominated by palatable, climax, decreasers. This indicates that the area is suitable for grazers, due to high nutritional content of the grasses available. However, predator avoidance may be more of a consideration in these areas. The climax vegetation indicates that the area undergoes very little disturbance, while the abundance of decreasers would suggest that the area probably undergoes limited grazing pressure. Decreaser grasses decrease in number in over and undergrazed areas (Van Wyk and van Oudtshoorn 1999), suggesting that areas where these predominate experience moderate levels of grazing pressure. As fire frequency increased, we found a marked increase in the proportion of increaser II grasses of medium to low palatability. Increaser IIs generally indicate higher levels of grazing or disturbance. Proportions of subclimax grasses also increased as fire frequency increased, which would tend to suggest that disturbance is greater on annually burned sites, where subclimax grasses dominate. However, from this data there is no way to determine if this disturbance takes the form of grazing or of fire. Both probably play a role, but we have no way of determining the extent of this interaction. Decreasing palatability would suggest that more frequently burned areas are less suitable for grazers, but here again, predator avoidance may play a major role in the habitat selections of different types of herbivores. The vegetation profile certainly decreases in height in more frequently burned areas. In fact, our surrogate for biomass at various heights was visibility range, making an analysis of animal visibility entirely plausible. Thus, when vegetation profile decreases in height, animals can see farther closer to the ground. Grazers would be less likely to be found in very dense vegetation as their vigilance capacity would be low at the heights at which they feed (0-0.5m) thus making them susceptible to attack by undetected predators. As browsers feed at different heights to grazers they may be better suited to inhabit more densely vegetated areas as less time is spent with vision near to the ground. Thus, large herd grazers should predominate on more frequently burned sites, while solitary grazers and browsers should predominate on control and triannual burn sites. However, here again, we have no data for the occurrence of herbivores at each of these sites, making confirmation impossible. A higher percentage of canopy should result in a higher diversity in invertebrate fauna. Covered areas have a much greater range of thermal niches throughout the day, which creates a heterogeneous microenvironment for invertebrates. Ants are particularly controlled by thermal conditions, and different species are adapted to tolerate different thermal ranges (Agosti et al. 2000). Thus a habitat with a wider range of thermal niches, which occurs when canopy cover is greater, should have a higher diversity of ants. In the Pretoriuskop EBPs, this is indeed the case. As canopy cover decreases, ant diversity increases. Defining communities based on species present, however, proved to be unfruitful. Sites within a string were more closely related than no burn or annual burn

24 sites, perhaps because ants have a fairly large range and can cross over from one plot to another within adjacent areas. We were also able to show relationships between bird communities and vegetation structure and composition to a limited extent. Vegetation structure is commonly believed to be a primary factor in defining bird communities, although floristic composition also plays a vital role, especially with respect to fruiting and seeding (Skowno 2000). Thus changes in the composition and structure of the vegetation due to differing burning regiments should influence the bird communities in that area. Although we were only able to analyze a limited number of sites, annual burn areas seem to have more closely related bird communities than sites within a string. Thus, bird composition does seem to be defined by vegetation structure. However, we have thus far been unable to relate bird species present and their habitat preferences to vegetation structure present at the various sites. It is clear that burning regime plays a large role in determining the vegetation structure and plant species composition and this in turn influences the diversity and abundance of fauna found in areas of differing burn regimes. Managers of Kruger National Park have long known that fire is one of the most important tools they have available to them to create this patch heterogeneity that is a characteristic of thriving complex adaptive systems. It is far more than just the vegetation that changes, but rather also the mammal, bird, and insect species that live there.

Acknowledgements Thanks to William Bond and Ed February for assistance in the field, OTS students and staff for collecting data and our esteemed game guards for ensuring our safety.

Literature Cited Agosti, D., J. D. Majer, L.E. Alonso and T. R. Schultz. 2000.Ants: Standard methods for measuring and monitoring biodiversity. Smithsonian institution press. Washington , USA. Bond, W. J. and B. W. van Wilgen. 1996. Fire and Plants. Population and Community Biology Series 14. London: Chapman and Hall. Majer, J.D. and H.F. Recher. 1988. Invertebrate communities on Western Australian eucalypts – a comparison of branch clipping and chemical knockdown procedures. Australian Journal of Ecology 13: 269-278. Microsoft Excel 2002. Primer for Windows. 5.1.2 2000 Schowalter, T.D. 1989. Canopy arthropod community structure and herbivory in old- growth and regenerating forests in western Oregon. Can. J. For. Res. 19: 318-322. Schowalter, T. 1994. Invertebrate community structure and herbivory in a tropical rain forest canopy in Puerto Rico following Hurricane Hugo. Skwono 2000 => Laurence, you have the reference for this one, I already gave it back to you… Statistica 6.1 Stasoft Inc 1984-2004 Van Wilgen, BW, W Trollope, HC Biggs, A Potgieter, and BH Brockett. 2003. Fire as a Driver of Ecosystem Variability. Pgs 149-170 in Du Toit, JT, KH Rogers, and HC

25 Biggs, (eds.), The Kruger Experience:Ecology and Management of Savanna Heterogeneity. Washington: Island Press. Van Wyk, E. and F. van Oudtshoorn. 1999. Guide to Grasses of Southern Africa. Pretoria: Briza.

Table 1. EBP included in the study with corresponding codes. Season and frequency of Plot Name String burn N7C Numbi No Burn N10AB3 Numbi August triannual N1OB2 Numbi October biannual N6AB1 Numbi August annual S7C Shabeni No burn S12FB3 Shabeni February triannual S10AB3 Shabeni August triannual S3AB1 Shabeni August annual K1C Kambeni No burn K8FB3 Kambeni February triannual K3OB2 Kambeni October biannual K7AB1 Kambeni August annual

26 Table 2. Proportion of decreaser, increaser I, and increaser II grasses in each of the burn sites and treatments. No increaser III grasses were found. Feb triannual Aug triannual Oct biannual Aug annual No Burn burn burn burn burn

Numbi

Shabeni

Kambeni

Legend: decreasers increaser I increaser II

Table 3. Proportion of grasses of high, medium, and low palatability in each of the burn sites and treatments. Feb triannual Aug triannual Oct biannual Aug annual No Burn burn burn burn burn

Numbi

Shabeni

Kambeni

Legend: high medium low

27 Table 4. Proportions of climax, subclimax, and pioneer grasses in each of the burn sites and treatments. Feb triannual Aug triannual Oct biannual Aug annual No Burn burn burn burn burn

Numbi

Shabeni

Kambeni

Legend: climax subclimax pioneer

1 1 50 3 s 1 40

30 2 1 20 2 Abundance of individual 10 2 4 4 2 3 3 3 4 4 0 No burn August triannual burn October biannual burn August annual burn Site

Figure 1. Size class distributions for the Numbi EBP string. Size class 1: height ≤ 1m; size class 2: 1m < height ≤ 2m; size class 3: 2m < height ≤ 4 m; size class 4: height > 4m.

60 1

s 2 40

3 30 1 3 1

20 2

Number of individual 4

10 4 22 44 1 3 3 0 No Burn February triannual burn August triannual burn August annual burn Site

Figure 2. Size class distributions for the Shabeni EBP string. Size class 1: height ≤ 1m; size class 2: 1m < height ≤ 2m; size class 3: 2m < height ≤ 4 m; size class 4: height > 4m.

70 2

60 1 1 s 50 1 40 3 2 1 2 30

Number of individual of Number 20 2 3 3 10 4 3 4 4 0 No burn February triannual burn October biannual burn August annual burn Site

Figure 3. Size class distributions for the Kambeni EBP string. Size class 1: height ≤ 1m; size class 2: 1m < height ≤ 2m; size class 3: 2m < height ≤ 4 m; size class 4: height > 4m.

Figure 4. Multi-dimensional scaling community analysis of tree species in the various burn plots.

Figure 5. Multi-dimensional scaling community analysis of grass species in the various burn plots.

2.0 m 1.5 m 1.0 m August annual 0.75 m 0.5 m 0.25 m

2.0 m 1.5 m

October biannual 1.0 m 0.75 m 0.5 m 0.25 m

2.0 m 1.5 m 1.0 m August triannual 0.75 m 0.5 m 0.25 m

2.0 m 1.5 m 1.0 m No burn 0.75 m 0.5 m

0 5 10 15 20 25 1/distance (m^-1)

Figure 9. Foliage profiles of vegetation at the Numbi EBP

2.0 m 1.5 m 1.0 m August annual 0.75 m 0.5 m 0.25 m

2.0 m 1.5 m 1.0 m August triannual 0.75 m 0.5 m 0.25 m

2.0 m 1.5 m 1.0 m February triannual 0.75 m 0.5 m 0.25 m

2.0 m 1.5 m 1.0 m No burn 0.75 m 0.5 m

0123456789 1/distance (m^-1)

Figure 10. Foliage profiles of vegetation at the Shabeni EBP

31 2.0 m 1.5 m 1.0 m August annual 0.75 m 0.5 m 0.25 m

2.0 m 1.5 m 1.0 m October biannual 0.75 m 0.5 m 0.25 m

2.0 m 1.5 m 1.0 m February triannual 0.75 m 0.5 m 0.25 m

2.0 m 1.5 m 1.0 m No burn 0.75 m 0.5 m 025

0123456789 1/distance (m^-1)

Figure 11. Foliage profiles of vegetation at the Kambeni EPB.

100 1. 7

1. 6

80 1. 5 1. 4

1. 3 60 1. 2

1. 1 40 1. 0 S hannon Diveristy index

Percentage canopy open (%) open canopy Percentage 0. 9

20 0. 8

0. 7

0 0. 6 Numbi Numbi Shab Nu S e C m

Figure 12. Graph indicating the percentage of aerial space open as well as the Shannon diversity indices of ant communities at four different burn plots. █ Shannon-Weiner Diversity Index; █ Percent of canopy open (= 100% – canopy cover).

32 12

Bird species richness 2

0 0123456789 Order of Sampling

Figure 13. Bird species richness v. order of sampling (r2=0.7121, p=0.0084).

No Burn Augu October st biannua

Figure 14. Cluster analysis of bird communities in three burn plots in the Numbi string.

33 The effects of the 2000 floods on channel type heterogeneity in the Sabie River, Kruger National Park, South Africa

Category: Faculty Field Project Participants: Megan Eastwood and Gareth Hempson (secretaries), Mark Roundtree (resource person) Site: Sabie River, Kruger National Park, Mpumalanga Province, South Africa

Key words: channel type, flood effects, large infrequent disturbances, riverine ecosystems, Sabie River, sedimentation

Abstract: The naturally dynamic flow regime of a river promotes heterogeneity in channel types found within the river and therefore also promotes diversity in the riverine flora and fauna. Increased sedimentation in the Sabie River caused by logging and farming in the catchment area seemed to be shifting all channel types towards an alluvial-dominated state, causing a loss in species dependent upon bedrock-dominated areas. However, the 2000 floods on the Sabie River drastically changed the template of the river, removing a great deal of vegetation and redistributing sediment. Examining an aerial photo history of sites along the Sabie River reveals that the flood returned the river system to a state similar to the river’s condition after the last large flood in 1921, suggesting that the 1:100 year floods periodically reset the river system in an episodic model of change.

Introduction The natural flow regime of a river is dynamic, which promotes heterogeneity in the different channel types and morphological units found within the river (Poff et al 1997). This in turn creates a variety of habitats and promotes diversity in the riverine vegetation and fauna. As a result, riverine ecosystems are acknowledged throughout the world as “biodiversity hotspots” (O’Keeffe and Rogers 2003). The Sabie River in the Kruger National Park, in addition to providing water to many of the animals in the park, supports 45 fish communities, 134 taxa of invertebrates, and a diverse community of riverine vegetation (O’Keeffe and Rogers 2003). However, the catchment of the Sabie lies outside the KNP, and human activities in the catchment area have impacted the river considerably. Although the pollution in the Sabie has been greatly reduced in recent years, the fruit farms and commercial forestry industry in the catchment area still consume approximately 30% of the Sabie’s mean annual runoff while greatly increasing the sediment load carried by the river (O’Keeffe and Rogers 2003). The combination of a reduced flow and an increased sediment load leads to a great deal of sediment accumulation within the river channel. (Van Coller et al 1997, O’Keeffe and Rogers 2003). Sedimentation in bedrock-dominated areas of the river has been of particular concern to KNP managers as it shifts all channel types in the river towards an alluvial-dominated state (Van Coller et al 1997). The Sabie River contains four main channel types: bedrock anastomosing, mixed anastomosing, mixed pool-rapid and braided. Bedrock anastomosing channel types have a steep slope and contain multiple channels cutting through outcrops of resistant rock such as dolerite. Sedimentation in a bedrock anastomosing channel creates a mixed anastomosing channel type, which has a shallow slope and multiple channels running through a mixture of bedrock and alluvium. Mixed pool-rapid channel types contain low- energy pools intermixed with short rapids caused by bedrock deposits. Braided channel types, which result from extensive sediment accumulation, contain many sandbars and low-flow alluvial channels. Sedimentation in the first three channel types ultimately transforms them into a braided channel type, leading to a loss of channel type heterogeneity and therefore a decrease in the diversity of flora and fauna the river can support (Rogers and O’Keeffe 2003). Species like Breonadia salicina, which recruits only on bedrock, are greatly reduced, while species that thrive on alluvial units, such as Phyllanthus reticulatus and Conbretum erythrophyllum, spread throughout the river (Van Coller et al 1997). Despite management’s fears about sedimentation, the 1996 flood on the Oliphants River cleared out a great deal of sediment and vegetation, leading to speculation that large infrequent disturbances (LIDs) such as this extensive flood periodically reset the river system in an episodic stripping model (Rogers and O’Keefe 2003). However, while this may have been true in the past, would the human-caused increase in sediment load disrupt the natural system? For example, the 1996 flood on the Sabie River was of intermediate size and had only moderate effects, and the vegetation removed seemed to recover rapidly.

34 In 2000, the Sabie River experienced extensive flooding of a magnitude not seen there since 1921. Like the 1996 Oliphants flood, the 2000 floods visibly changed the face of the Sabie, removing large amounts of vegetation and redistributing sediment. Kruger’s management is interested in examining the river template before and after the 2000 floods to understand how the river geomorphology has changed in recent years at different spatial and temporal scales. It also presents an opportunity to further study LIDs and the possibility of an episodic stripping model of change. With these goals in mind, we set out to examine the effects of the 2000 floods on the Sabie River in the context of the long-term dynamics of the river.

Methods We examined four sites along the Sabie River in the Kruger National Park, South Africa using an aerial photo history, channel profiles and present-day field observations. Site four is a bedrock anastomosing channel, site five is a mixed pool-rapid channel, site seven is a mixed anastomosing channel, and site ten is a bedrock anastomosing channel. Photos of the sites from key years in the Sabie’s flooding history were examined: photos from 1940 (just after the large 1921 flood and an intermediate flood in 1936), 1996 (just after an intermediate flood), 1999 (just before the 2000 floods), 2000 (just after the large floods), and 2002 (to see how the river was recovering from the LID). The percentage of water, vegetation cover, sand, and bedrock at each site was determined using the photos from each year. We covered the photos with transparencies and colored in the different features, then used graph paper to estimate the area covered by each feature as a proportion of the total area measured. Channel profiles created by Mark Roundtree for site five, seven and ten in 1995 and 2000 also helped us to investigate how the 2000 floods altered the river template. Field observations made on 22/2/04 helped us to see changes in the sites since the channel profiles and the most recent photos (2002) and enabled us to distinguish particular features seen in the photographs.

Results The effects of the 2000 floods were clearly visible at all study sites on the Sabie River, with the most obvious change in river morphology at all sites occurring between the 1999 aerial photos and the 2000 aerial photos. The impacts of the flood appear to have been more severe at the upstream sites 4, 5 and 7 than further downstream at site 10. Site 4 As seen in figure 1, the period from 1940 to 1996 saw a notable increase in the percentage of vegetation cover (27.7% to 74.7%). This increase in plant cover appears to mainly be woody growth on bedrock in the channel but also takes the form of woody vegetation on alluvium. This trend continued from 1996 to 1999. Comparing the 1999 and 2000 aerial photographs reveals a substantial decrease in vegetation cover from 83.6% to 26.3%. Most of the vegetation removed was woody material growing on bedrock substratum within the river channel, resulting in a much higher percentage of bedrock being exposed in 2000 than in 1999 (34.6% from 2.8%). Revegetation of Site 4 from 2000 to 2002 mainly took the form of reed and bushy growth on sandbars and the bank. There appears to be no recovery of woody vegetation on the in-channel bedrock. Examination of the site in 2004 revealed reed growth in the channel, and reed, bush and tree regrowth on the banks. The channel also appears to have widened, and evidence of sand deposition and channel braiding was observed. Site 5 An increase in vegetation cover was also observed at Site 5 from 1940 to 1999 (74.8% to 83.4%), although this was less pronounced than at Site 4. This increase would chiefly appear to be woody vegetation on the banks. The intermediate floods of 1996 exposed slightly higher percentages of bedrock and alluvial substratum, but this had been largely revegetated by 1999. However, the floods of 2000 had a much larger impact, dramatically increasing the amount of exposed alluvium from 4.8% to 21.1%, and bedrock from 1.9% to 4.2%. Vegetation cover decreased from 83.4% to 48%, with the highest losses in trees growing along the banks. (Figure 2) The 2002 aerial photo shows vegetation recovery to largely be reedy growth on alluvial substrata. River channel profiles from 1995 and 2000 show the channel bed to have flattened substantially as a result of the flood (Figure 3). The 2002 aerial photo further suggests that the channel has become more braided after the 2000 flood, largely as a result of alluvial deposits. Site 7 The percentage of vegetation increased from 73.5% to 79.3% from 1940 to 1999. This increase was primarily due to woody growth on alluvium but was also due to woody growth on bedrock. The 2000

35 floods cleared a large proportion of this woody growth from both bedrock and alluvial substrates (Figure 4). The comparison of river channel profiles from 1995 and 2000 reveals a significant flattening of the channel (Figure 5). Similar to Site 5, the river channel is more braided in the 2002 aerial photograph, again appearing to be a result of in-channel alluvial deposits. That vegetation recovery after the 2000 flood appears to be primarily reedy growth on alluvial substrata. Vegetation recovery is very limited on the bedrock substratum. Site 10 Site 10 differed from the other three sites in that no alluvial deposits were in evidence on this stretch of river. A substantial increase in vegetation cover was observed from 1940 to 1999 (34.7% to 76.4%), largely in the form of woody material on the in- stream bedrock. The 2000 floods dramatically reduced this vegetation cover (76.4% to 25.8%), exposing much bedrock. (Figure 6). The river channel profile comparison between 1995 and 2000 shows a slight smoothing of the channel (Figure 7), but the effect is less dramatic here than at sites 5 and 7. The observation of a downed marula (Sclerocarya birrea) tree in the river channel suggests that it has widened somewhat since the 2000 floods.

Discussion The general trend in evidence along this stretch of the Sabie River from 1940 to 2004 is consistent with that of an episodic stripping model of geomorphic change (Rogers and O’Keeffe 2003). The consistent observation of a build up of vegetation at all sites suggests that sediment accumulated between the large, infrequent flood events of 1921 and 2000. This sediment would provide a substratum for the establishment of vegetation, the presence of which further contributes to sediment storage by slowing the river’s flow. Woody vegetation was the dominant recruiting growth form within the macro channel. This would suggest that enough time had elapsed between the two floods for the two biophysical succession trajectories of O’Keeffe and Rogers (2003) to develop to their end phases. A flood event of the magnitude seen in 2000 on the Sabie River has been estimated to have a return frequency of one in one hundred years, and is referred to as a large, infrequent disturbance (LID) event (Rogers and O’Keeffe 2003). The effects of this flood event were significantly more dramatic at all sites included in this study than that of the 1996 one-in-fifty year flood. Comparing the 1940, 1996 and 2000 photos suggests that the intermediate 1996 flood had less impact on the Sabie River system than the 1936 flood of similar magnitude. Rogers and O’Keeffe’s hypothesis (2003) that as succession stabilizes and becomes directional after a LID event, the magnitude of the flood required to reset the river system increases, appears to be supported. The 2000 flood was of a magnitude large enough to reset the river system. This is evident in the mass removal of woody vegetation from within the macro channel at the bedrock anastomosing sites 4 and 10, leaving large areas of exposed bedrock. The alluvium accumulated here during the period since the last LID event was scoured away and redistributed in the river channel. This redistribution of sediment is likely to be responsible for the build up of sediment in evidence at the mixed pool-rapid Site 5. Sediment deposition was also observed in certain areas at sites 4 and 7. The smoothing of the river channel from 1995 to 2000 at sites 5 and 7, as shown in the channel profiles (Figures 3 and 5), is further evidence of the resetting capacity of floods, although this does not distinguish between the effects of the 1996 flood and the 2000 flood. Mixed pool-rapid systems were found to be the most dynamic in terms of shifts between more alluvial and more bedrock dominated states in the Olifants River (Rountree et al 2002 in Rogers and O’Keeffe 2003). The response at site 5 to the 2000 flood appears to be a shift to a more alluvial state. Site 7 is a mixed anastomosing site and also showed a shift towards a more alluvial-dominated state. The bedrock anastomosing sites 4 and 10 responded to the 2000 flood by shifting to a more bedrock-dominated state. Vegetation recovery at sites 4, 5 and 7 on alluvial substrata took the form of reed growth. These areas are likely to develop according to the first trajectory predicted by O’Keeffe and Rogers (2003), eventually developing into patchy woody communities dominated by Combretum erythrophyllum, Ficus sycomorus and Nuxia oppositifolia. Breonadia salicina recruitment on the exposed bedrock at sites 4 and 10 is likely the initial phase of the second biophysical succession trajectory that gives rise to a B. salicina, Syzigium cordatum and S. guineense community (O’Keeffe and Rogers 2003). It should be kept in mind that vegetation recovery after a LID event could follow a slightly different developmental pathway to that of events of lesser magnitude. The LID event of 2000 on the Sabie River appears to have had the effect of resetting the river system to a state similar to that caused by the flood of 1921. The resulting template is still however highly

36 heterogeneous and would thus conceivably indicate that it has the ability to regain and sustain the high levels of biodiversity it supported before 2000. Current understanding of recovery from LID events such as this is limited by their scarcity, making the years ahead an exciting opportunity to enhance the understanding of their effects on these systems.

Acknowledgements We would like to thank Mark Roundtree for his teaching and assistance in this project, and all of the OTS group for helping us collect data and draw many, many dots.

Literature Cited O’Keeffe, J. and K. Rogers. 2003. Heterogeneity and management of the lowveld rivers. Pages 447-468 in Dutoit, J., K. Rogers, and H. Biggs, eds. The Kruger Experience: ecology and management of savanna heterogeneity. Island Press, Washington DC, USA. Poff, N., J. Allan, M. Bain, J. Karr, K. Prestegaard, B. Richter, R. Sparks, and J. Stromberg. 1997. The natural flow regime: a paradigm for river conservation and restoration. Bioscience 47: 769-784. Rogers, K. and J. O’Keeffe. 2003. River heterogeneity: ecosystem structure, function, and management. Pages 189-218 in Dutoit, J., K. Rogers, and H. Biggs, eds. The Kruger Experience: ecology and management of savanna heterogeneity. Island Press, Washington DC, USA. VanColler, A., K. Rogers, and G. Heritage. 1997. Linking riparian vegetation types and fluvial geomorphology along the Sabie River within the Kruger National Park, South Africa. African Journal of Ecology 35: 194-212.

45 90 42.2 83.6

40 80 74.7 35 70 30 30 60 25 50 20 20 40 15.3 28.8 15 30 27.7 26.3 Percentage of Water 11.1

Percentage of Vegetation 10 20 5 10 0 0 1940 1996 1999 2000 2002 1940 1996 1999 2000 2002

45 42.8 12 10.7 40 10 34.6 9.1 35 8.4 m k 30 8

25 6 19.3 20 15 4 3.4 Percentage of Bedroc Percentage of Alluviu 2.47 10 6.5 2 5 2.8 0 0 1940 1996 1999 2000 2002 1940 1996 1999 2000 2002

Figure 1: Percentage cover of water, vegetation, bedrock and alluvium at Site 4 on the Sabie River in 1940, 1996, 1999, 2000 and 2002.

30 90 83.4 26.8 81.9 80 74.8 25 70 65.5

20 60 18.1 16.6 50 48 15 40 9.9 10 30 Percentage of Water of Percentage

PercentageVegetation of 6.1 20 5 10 0 0 1940 1996 1999 2000 2002 1940 1996 1999 2000 2002

5 25 4.4 4.5 4.2 21.1 4 20 3.4 3.5 3.2 3 15 12 2.5 1.9 2 10 8.6

Percentage of Bedrock of Percentage 1.5 Alluvium of Percentage 5.4 4.8 1 5 0.5 0 0 1940 1996 1999 2000 2002 1940 1996 1999 2000 2002

Figure 2: Percentage cover of water, vegetation, bedrock and alluvium at Site 5 on the Sabie River in 1940, 1996, 1999, 2000 and 2002.

Channel Profile Site 5

346

344

) 342

2000 340 1995

Elevation (masl 338

336

334 0 20 40 60 80 100 120 140 160 Distance (m)

Figure 3: Channel profile comparison at Site 5 on the Sabie River for 1995 and 2000.

30 90 27.4 79.3 80 73.5 74.4 25 70

58.3 20 60 17.7 51.6 50 15 40 11.3 10 8.8 30 Percentage Water of 8.1

Percentage Vegetation of 20 5 10 0 0 1940 1996 1999 2000 2002 1940 1996 1999 2000 2002

7 20 18.3 18 6 5.8 5.7 16 15.1 14.6 5 14 11.6 12 4 3.6 9.6 10 2.9 3 8 2.3

Percentage of Bedrock of Percentage 6 2 Alluvium of Percentage 4 1 2 0 0 1940 1996 1999 2000 2002 1940 1996 1999 2000 2002

Figure 4: Percentage cover of water, vegetation, bedrock and alluvium at Site 7 on the Sabie River in 1940, 1996, 1999, 2000 and 2002.

Channel Profile Site 7

326 325

324

) 323

322 2000 321 1995 320

Elevation (masl 319

318 317

316 0 40 80 120 160 200 Distance (m)

Figure 5: Channel profile comparison at Site 7 on the Sabie River for 1995 and 2000.

35 90

30.3 80 76.7 76.4 30

70 25 60 19.7 20 50

14.8 37.4 15 40 34.7 30

Percentage ofWater Percentage 25.8

10 Percentage of Vegetation 20 5 3.8 2.8 10 0 0 1940 1996 1999 2000 2002 1940 1996 1999 2000 2002

60 54.5 50 47.8 k 40 34.9 30

20.5 19.8 20 Bedroc of Percentage

0 1940 1996 1999 2000 2002

Figure 6: Percentage cover of water, vegetation, bedrock and alluvium at Site 10 on the Sabie River in 1940, 1996, 1999, 2000 and 2002.

Channel Profiles Site 10

299

298

297

) 296

295 2000 294 1995 293

Elevation (masl 292

291 290

289 0 50 100 150 200 250 300 350 400 Distance (m)

Figure 7: Channel profile comparison at Site 10 on the Sabie River for 1995 and 2000.

40 The habitat- based relationship between peak frequency and wing morphology in Microchiropterans of the southern Kruger National Park

Category: Faculty Field Project Participants: Tammy Baudains (editor), Govan Pahad (editor), David Jacobs (resource person) and Corrie Schoeman (resource person) Site: Skukuza, Kruger National Park, Mpumalanga Province, South Africa

Key Words: aspect ratio, habitat, peak frequency, microchiroptera, wing loading, wing morphology

Abstract: Our study is an analysis of the aspect ratio, wing loading and call peak frequency of five species of microchiropterans in Skukuza, Kruger National Park. Bats were caught on four consecutive nights in February using mist nets erected over and around water sources. Species of bats with higher peak frequencies grouped together in our analysis and were also characterized by low wingloadings and aspect ratios as compared with those bats with lower peak frequencies. These characteristics are indicative of already observed foraging strategies. Our results therefore supported the notion that bats flying in open areas tend to have high aspect ratios and wingloadings but low peak frequencies of calls and that bats which forage in clutter have low aspect ratios and wingloadings but high peak frequencies of calls.

Introduction Microchiropterans display a wide variety of wing morphologies and utilize many different echolocation calls which vary in frequency, bandwidth and duration. These differences are largely a reflection of the foraging strategy and foraging habitat of the bat (Altringham 1996). The shape of a bats wing operates as an idealized aerofoil by producing the forces responsible for keeping the bat in the air and moving forward. Air moves faster over the convex upper surface of the wing thereby creating a lower pressure than that of air below the wing. As a result the net aerodynamic force pushes upwards to compensate for this unequal pressure distribution. In this way the bat is lifted. The forward and downward movement of the wings acts relative to the airflow lift to produce forward thrust (Altringham, 1996). Two important parameters of wing morphology are aspect ratio and wingloading. Aspect ratio is calculated by dividing wingspan2 by wing area. Short broad wings (low aspect ratio) are less aerodynamically efficient than long narrow wings (high aspect ratio). Broad wings are found in all slow flying maneuverable bats. The wing membrane found anterior to the humerus and radius is called the propatagium and is large in broad winged bats. This membrane can be lowered to increase the camber of the wing thereby increasing drag on the wing. This compromise is worthwhile in slow flying bats when drag is low (Altringham 1996). Narrow wings have a lower surface area than broad ones and therefore the bat needs to fly faster to maintain as much lift. Increased flight speed requires less induced power to generate thrust and lift (Altringham 1996). Wingloading is calculated by dividing body mass, multiplied by acceleration due to gravity (9.81), by wing area2. Because wingloading is derived from mass and surface area, it is effectively an estimate of the ratio of the inertia of the bat to the force its wings can generate. A low wingloading therefore has a similar effect to a low aspect ratio: reducing the minimum speed at which a bat can remain airborne and increasing its maneuverability. Echolocation calls are immensely variable. Calls can be of a high or low frequency, be constant frequency or frequency modulated, have a narrow or broad band width and have a high or low duty cycle. In general, high frequency calls pick up more detail but attenuate faster with distance (making it best for cluttered environments but poor in the open), whereas low frequency calls travel further, but pick up less detail (good in open areas but poor in clutter) (Aldridge & Rautenbach 1987). Peak frequency (the frequency at which a bat puts the most energy into its call) can be used to represent the dominant frequency of a call. It is therefore an indicator of the habitat type for which the call is most suitable. Aldridge and Rautenbach (1987), in a study in Pafuri, found significant correlations between morphological parameters that improve maneuverability (wingloading, aspect ratio and wingtip shape index) and echolocation calls that are resistant to acoustic clutter. Our study, which took place around Skukuza in the south of the park, aimed to investigate the relationship among wing loading, aspect ratio, peak frequency and foraging strategy as stated by Norberg

41 and Rayner, 1987: Bats foraging in open areas tend to be fast flyers with high aspect ratio wings, high wingloading and typically use low frequency calls while bats foraging in and around vegetation have broad wings, low wing loading and usually make use of high frequency calls.

Methods Our study period consisted of four nights in February, 2004. Mist nets of 6 and 12 meters in length were erected at three sites in Skukuza camp in the southern part of Kruger National Park. At all three sites the nets were erected over water. An artificial waterhole, a pond and a public swimming pool proved to be good sites. Bats drink about twice nightly from known water sources, usually permanent ones, and therefore one would expect a better catch rate at places where they converge to drink (Taylor 2000). Between four and six nets were put up each night. Bats were kept overnight. Whilst in hand the following measurements were taken from each bat; left and right forearm, head length, tail length, and body mass. The right wing was stretched out over graph paper and was photographed with a digital camera. The image was calibrated to actual size and analysed using SigmaScan Pro 5. Wing span and wing area were measured using this software. Bats were hand-released the night following that of their capture and echolocation calls were recorded using an Anabat 2 recorder and analysed in BatSound Pro 3.3 to obtain the peak frequency of each call. Aspect ratio and wingloading were calculated from the acquired measurements. These two indices of wing morphology were graphed against peak frequency using Microsoft Excel 2000.

Results Five species of microchiropterans were caught: one Pipistrellus zuluensis, two P. nanus, thirteen Scotophilus dinganii, six Mops condylurus and three Chaerephon pumilus. Of these one Mops condylurus and one Scotophilus dinganii were eliminated from the analysis because of an incorrect call recording and a recording from a room release (which affects the call frequency) respectively. The results suggest that the species can be lumped into two groups. (Figures 1, 2 & 3.) The MDS plot revealed a low stress value of 0.03. In both graphs the groups consisted of the same species, with Scotophilus dinganii, and the two Pipistrellus species (P. nanus and P. zuluensis) in group 1 and Mops condylurus and Chaerephon pumilus in group 2. (Figures 1, 2 & 3.) Group 1 individuals were characterized by a higher peak frequency, lower aspect ratios and lower wingloadings. Group 2 individuals were characterized by a lower peak frequency, higher aspect ratios and higher wingloadings. Another trend noticed was that the aspect ratios of group 1 individuals were dispersed while their wingloadings were clustered. In contrast the aspect ratios of group 2 individuals were clustered while their wingloadings were dispersed. (Figures 1 & 2.) Our results suggest that there is an inverse relationship between peak frequency and both aspect ratio (gradient of -6.0046) and wingloading (gradient of -6.2576) (Figures 1 & 2). There were a few individuals whose characteristics did not fit with those of the others of their species. One of the two Pipistrellus nanus individuals had a much higher peak frequency of its call (65.7 kHz) than the other (32.1 kHz). One Chaerephon pumilus had a much lower aspect ratio (6.075) than did the other two (7.132 and 7.108).

Discussion The results from the study provide convincing support for the hypothesis that wing structure and echolocation calls of microchiropterans are related in a predictable manner. Group 1 consists of species from the family Vespertilionidae which, according to Norberg and Rayner (1987), contain bats that usually forage in and around vegetation. Group 2 consists of species from the family Molossidae, most of which are open space, fast flying bats (Norberg and Rayner 1987). The individuals within the two groups showed close associations with each other with the exception of the two outliers (MDS stress value of 0.03 which indicates a highly parsimonious ordination) (Figure 3). Given that we have only sampled a few species from the two families, we do not imply that the patterns found represent all species within these families. We took species specific information concerning the foraging behavior of the bats from Taylor’s Bats of Southern Africa and found that their foraging behavior matches what we would have predicted based on the aspect ratios, wingloadings and peak call frequencies of the bats. Group 1 (low aspect ratios, low wingloadings and high peak call frequencies) consisted of woodland edge feeders (Pipistrellus nanus and P. zuluensis) and intermediate clutter feeders (Scotophilus dinganii). Group 2 (high aspect ratios, high wingloadings and low peak call frequencies) consisted of high flying aerial feeders (Chaerephon pumilus

42 and Mops condylurus). It is important to remember that those species in group 1 are midway on the continuum between open area foragers and dense vegetation foragers, rather than at the dense vegetation extreme. Those species adapted for flight in dense vegetation most likely had calls which provided sufficiently high resolution for them to detect and avoid our mist nets and are therefore absent from our study. Bats are very flexible in their use of echolocation. They overcome restrictions imposed by particular combinations of echolocation call characteristics by varying these to suit a particular foraging situation. (Aldridge and Rautenbach 1987.) This could explain the outlying Pipistrellus nanus that had a peak call frequency far higher than the other of its species. The one individual of Pipistrellus zuluensis had a higher peak frequency than we would have expected based on its morphology. Unfortunately, the lack of other individuals of its species makes it impossible to tell whether it is an outlier or a true representative of its species. On of the Chaerephon pumilus individuals had a much lower aspect ratio than the other two. We suspect this is an unusual individual as is grouped more with group 1 than with group 2 (in terms of aspect ratio) but again, the small sample size makes it difficult to be sure. The trend that we noticed of increased variation of aspect ratio in group 1 and increased variation of wingloading in group 2 could possible be attributed to an increased variation in wing shape for those species in group 1 and an increased variation in body mass for species in group 2. These variations in clustering could merely be attributable to poor sample size or they could be based on the foraging strategies or relatedness of the two groups. The time constraints on the project meant that the sample size of species used was very small and limited our ability to determine patterns in a species when only 1 to 3 individuals were caught. It would be very interesting to do a similar study at the same site for a longer period and also over a different season. One could also test the same hypothesis for bats in the north of the Park. We would like to have examined more fully the trend we noticed of clustering and dispersion in individuals on the graph with respect to aspect ratio and wingloading. This could be done by trying to catch bats from different families and with different foraging strategies. This would show us whether or not it was just our two families which show this pattern. Our study has clearly shown that there are distinct differences separating the species of two families that were sampled with respect to peak frequency, aspect ratio and wingloading. These differences were found to be directly related to the foraging habitat for the species as described in Taylor, 2000. We conclude that patterns of bat ecology in Skukuza match those made by Norberg and Rayner, 1987, in that wing morphology and peak frequency are indicative of foraging strategy.

Acknowledgements We wish to thank David Jacobs and Corrie Schoeman of the University of Cape Town for their insights into this project and for assisting in the capture of all the bats; the Organisation for Tropical Studies class 2004 for helping with the captures and Deedra McLearn and Laurence Kruger for assisting with data analysis.

References Aldridge, H. D. J. N. and I. L. Rautenbach. 1987. Morphology, echolocation and resource partitioning in insectivorous bats. Journal of animal ecology. 56: 763-778. Altringham, J.D. 1996. Bats- biology and behaviour. Oxford University press. New York. Norberg, U. M. and J. M. V. Rayner. 1987. Ecological morphology and flight in bats (Mammalia; Chiroptera): wing adaptations, flight performance, foraging strategy and echolocation. Philosophical transactions of the Royal society (London). B 316: 335- 427. Taylor, P. J. 2000. Bats of Southern Africa. University of Natal Press. South Africa.

60 Sd & Pn/z

50 Species: PipistrellusPn nanus Cp 40 ○ Pipistrellus zuluensis Mc & Cp ■ ChaerephonMc pumilus ▲MopsSd condylurus Pz 30 ° Scotophilus dinganii Peak Frequency (kHz) Frequency Peak 20 y= -6.0046x + 70.706 10

0 55.566.577.58 Aspect Ratio

Figure 1. The relationship between peak frequency and aspect ratio for 21 individuals of 5 species of microchiropterans in Skukuza (trendline and equation to show inverse relationship).

Sd & Pn/z 60

50 Species: Pipistrellus nanus Pn 40 Mc & Cp ○ PipistrellusCp zuluensis ■ ChaerephonMc pumilus 30 ▲MopsSd condylurus ° ScotophilusPz dinganii Peak Frequency (kHz) Frequency Peak 20

10 y= -6.2576x + 38.746

0 00.511.522.5 Wingloading (N.m-2)

Figure 2. The relationship between peak frequency and wing loading for 21 individuals of 5 species of microchiropterans in Skukuza (trendline and equation to show inverse relationship).

Appendix 1. Raw data used in construction of graphs

wingloading (N.m- bat code wing span (mm) area (m) peak freq (kHz) mass (g) aspect ratio 2) 17.02.04Sd1 329.636 18.906038 37 26 5.747364545 0.713577754 17.02.04Sd2 333.1 18.181498 34.9 26 6.102666018 0.771583675 17.02.04Sd3 301.942 17.568676 34.5 26 5.18929095 0.826350488 17.02.04Sd4 292.748 16.406004 41.5 23 5.22378219 0.838284078 18.02.04Cp1 229.27 8.651426 27.6 12 6.075846098 1.572804919 19.02.04Cp1 240.092 8.085564 25.1 11.5 7.129269951 1.725624148 19.02.04Cp2 260.216 9.525132 27.6 12 7.108811369 1.297502633 19.02.04Mc2 305.46 13.41584 25.1 23 6.954898955 1.253606118 19.02.04Mc3 307.924 13.041508 28.7 31 7.270416103 1.788031135 19.02.04Mc4 323.14 14.11318 30.6 31 7.398719466 1.526795742 19.02.04Mc5 305.04 12.579808 27.6 33 7.396726691 2.045666918 19.02.04Mc6 323.158 13.878894 23.2 23 7.524453531 1.171351197 19.02.04Pz1 211.743908 7.2599898 47.7 6 6.175694982 1.116730137 20.02.04Mc1 327.204 15.050126 72.5 30 7.113725002 1.299301651 20.02.04Sd1 326.112 18.148268 32.4 22 5.860010252 0.655271562 20.02.04Sd2 335.63 19.357092 30.9 26 5.819443174 0.680709994 20.02.04Sd3 332.348 19.674528 34.9 26.5 5.614121625 0.671593117 21.02.04Pn1 204.856 7.181138 32.1 4 5.843917877 0.760926052 21.02.04Pn2 191.786 6.782578 65.7 4 5.422992525 0.852981093 21.02.04Sd1 332.076 19.36948 32.1 24 5.69320755 0.627544208 21.02.04Sd2 316.464 19.047438 33.9 26 5.257896799 0.703022489 21.02.04Sd3 325.358 19.959396 33.4 26 5.303658896 0.64024702 21.02.04Sd4 298.356 15.91546 46.7 21.5 5.593071312 0.832662635 21.02.04Sd5 339.748 19.525634 33.4 23 5.91164945 0.591815799 21.02.04Sd6 313.626 18.03383 36.8 20.5 5.454263896 0.618367884

45 Frog species diversity in three pans located in the Kruger National Park

Category: Faculty Field Problem Participants: Shannon Hatmaker, Sally Koerner, and Zoë Layton (secretaries), Vincent Carruthers (resource person), Blanchie Asberry, Tammy Baudains, Scott Briscoe, Laura Buckley, Eric Caldera, Jane Carruthers, Michael Chazan, Kinesh Chetty, Julie Coetzee, Fahiema Daniels, Abri de Buys, Marcus Byrne, Megan Eastwood, Kyle Harris, Gareth Hempson, Stephanie Johnson, Laurence Kruger, Tarryn Morris, Justine Norman, Govan Pahad, Jasper Slingsby, Mike Smith, Carla Staver, Simon Thomson, and Ben Wigley Sites: Skukuza, Kruger National Park, Mpumalanga Province, South Africa—Shingwedzi & Punda Maria, Kruger National Park, Northern Province, South Africa

Key words: South African frogs, pans, mating call, frogging

Abstract: Approximately thirty species of frogs are found in the Kruger National Park. Throughout the park there are three distinguishable types of pans - permanent, semi-permanent, and temporary, which vary substantially in different areas, creating patches that are dynamic in space and time. There are also different zones in the pans where distinct frog species choose to dwell. The purpose of this study was to examine the species composition of frogs in and around pans in three areas of the Kruger National Park; the Skukuza area, the Shingwedzi area, and the Punda Maria area. We also looked at the different species of frogs found in the different zones of the pan. We went out after sunset at each of the three sites and captured the frogs after noting their behavior and where they were calling from. In Skukuza, we caught seven species of frogs and identified the calls of two additional species, in Shingwedzi we caught four species of frogs and identified the calls of five additional species, and in Punda Maria we caught two species of frogs. We found that the three main factors affecting species diversity in Kruger National Park were pan type and size, location in KNP, and weather patterns (both long term and daily fluctuations).

Second language abstract: Ongeveer dertig padda spesies word in die Nasionale Krugerwildtuin aangetref. Panne in die park kan geklassifiseer word in drie groepe naamlik—permanent, nie-standhoudend en tydelik. Hierdie panne varieer ook in grootte en gevolglik dra by tot die tyd en ruimtelike variasie in die landskap. Verskillende sones wat paddas kan benut kan in elke pan identifiseer word. Die doel van hierdie projek was om die spesiesamestelling van paddas in die Skukuza, Shingwedzi en Punda Maria areas te ondersoek. Sone voorkeure van paddaspesies is ook ondersoek. Al drie studiegebiede is na sononder besoek en paddas is gevang nadat hulle geobserveer en aantekening oor hul gedrag gemaak is. Die sone waarvan af roepe gehoor is, is ook aangeteken. Sewe spesies is gevang en twee ekstra deur roepe identifiseer in Skukuza. In Shingwedzi is vier spesies gevang en vyf addisionele spesies met behulp van roepe uitgeken. Slegs twee spesies is in Punda Maria gevang. Die gevolgtrekking was dat pantipe, pangrootte, lokaliteit in die Wildtuin en weerpatrone (beide langtermyn en daagliks) ‘n rol speel in die spesiesamestelling wat aangetref word.

Introduction There are nearly 130 species of frogs in South Africa. Thirty of these species are found in the Kruger National Park and are attracted to the park because of its rainfall, temperature and high humidity. In the savanna, there are many pans where frogs gather to breed. A pan is a depression filled with rainwater and depleted by evaporation or absorption into the substrate (Carruthers 2001). Throughout the park, there are three distinguishable types of pans—permanent, semi-permanent, and temporary. These three types of pans vary substantially in different areas, which creates patches in the ecosystem. In the west, the pans are shallow and circular with muddy banks that become exposed due to the fluctuating water level. In the east, where it is wetter, the pans are surrounded by dense reeds and inundated grass with some trees (Carruthers 2001).

46 In the rainy season, up to 20 species of frogs can gather in a single pan to breed. Because of competition for space, different species call from, and live in, various zones of the pan. The first zone is the bank where frogs such as the Banded Rubber Frog (Phrynomantis bifasciatus), the Mottled Shovel- nosed Frog (Hemisus marmoratus), and the Ornate Frog (Hildebrandtia ornata) gather. The Banded Rubber Frog enjoys the bank because it calls from partially concealed positions while the Mottled Shovel- nosed Frog uses the bank for tunneling. The second zone is the open water and this zone is usually inhabited by the African Bull Frog (Pyxicephalus edulis) and the Red-legged Kassina (Kassina senegalensis). The Red-legged Kassina favors the open water because it attaches to water plants leaving most of its body submerged in the water. The third zone is inundated grass and is inhavited by the Snoring Puddle Frog (Phrynobatrachus natalensis), the Common Caco (Cacosternum boettgeri), the Broad-banded Grass Frog (Ptychadena mossambica), the Plain Grass Frog (Ptychadena anchietae), and the Banded Rubber Frog (Phrynomantis bifasciatus). The Grass Frogs breed in shallow flooded grasslands because they are well camouflaged in these areas. The fourth zone is reed beds. In this zone, Waterlily Reed Frogs (Hyperolius pusillus) and Painted Reed Frogs (Hyperolius marmoratus) are often found. The reed zone is favored by these two types of frogs because the Painted Reed Frogs climb reeds to do their calling and the Waterlily Reed Frogs lay their eggs in the overlapping layer of waterlily pads. The final zone is the tree zone that is inhabited by the Foam Nest Frog (Chiromantis xerampelina). The tree zone is inhabited by this type of frog because it lays its nest in overhanging branches (Carruthers 2001). Ultimately, frogs settle in certain areas of the pan because it is the ideal position for the males to call their mates. Frogs have several different types of calls including mating calls, heard most frequently, territorial calls, release calls and distress calls (Carruthers 2001). Mating calls vary widely between species and are used for the male to call females to the breeding site. Frogs are stimulated to call by many different factors, the most common stimulant being rain. The purpose of this study was to examine the species composition of frogs in three areas of the Kruger National Park; the Skukuza area, the Shingwedzi area and the Punda Maria area. This will allow us to examine species composition in pans that vary in size, seasonality and their location in Kruger National Park. Based on this dynamic variance we hypothesized that the species composition would be different in the three study sites. Moreover, we expect to find different species of frogs in each of the different zones of the pan.

Methods Species composition of frogs in pans was evaluated in three different areas of the Kruger National Park. We performed our study in Skukuza on 14 February 2004, in Shingwedzi on 4 March 2004 and in Punda Maria on 24 March 2004. On all three occasions, we went frogging after sunset at approximately 20h00. We listened attentively to the calls of the frogs, and if we were able to spot them with our torches, they were collected into bags. We recorded behavior of the frogs, and where they were located at the time of capture. We then put the frogs into plastic containers and brought them back to our research lab for species identification. In order to identify the frogs, we examined five characteristics. The first characteristic was locality—whether or not a certain species was found in the Kruger National Park. Secondly, we examined the frogs size and distinguishing characteristics. We made sure not to distinguish based on color because there is great color variance within single species. The next characteristic was habitat, more specifically, which zone of the pan the frog was found in. Finally, we matched the mating call with recorded calls (Carruthers 2001). All frogs were released the following day. During all frogging sessions, we also recorded species present in the pans based on identification of their calls, even if these species were not collected.

Results The pan near Skukuza was a fairly large permanent pan. It was approximately 0.5m deep with reeds in the middle region of the pan. The vegetation around the pan consisted of short grasses and about 5m away from the ponds edge, a ring of bushes and trees was evident. Prior to collecting, it rained for two days (Table 1). Weather conditions on the night were hot and humid with no wind. Seven species of frogs were caught at this site (Table 2). Adult and newly metamorphosed Mottled Shovel-nosed frogs were found along the shallow edges of the pan. The Bubbling Kassinas were found calling on the vegetated banks, and newly metamorphosed specimens were also found. The Painted Reed Frogs had a wide color variation, and were found calling from low on the reeds. Waterlily Frogs were calling from lily pads, and two pairs were found in amplexus. They also laid eggs in captivity using

47 grass stems. The Banded Rubber Frogs were calling in great numbers from the water’s edge. One pair was found in amplexus. Only one subadult Ornate Frog was found. The Foam Nest Frogs were found in a group reconstructing a nest in an emergent shrub, around 30 cm above the water. The Broad-banded Grass Frog was seen and heard in the inundated grasses and on the edges of the pans but was never caught. The Tropical Platanna Frog was found, but only in the tadpole form. The pan in Shingwedzi was dry, with seasonal flooding. Although there were not many reeds in this pan, there was an abundance of woody vegetation and fallen limbs from the trees. The weather before the frogging consisted of raining a substantial amount every day the week before (Table 1) and then getting slightly warmer the day of the frogging. We went out from 8 p.m. to 10 p.m., and it was very windy during that time. In Shingwedzi, we caught four species of frogs (Table 2). The Banded Rubber Frog was found on the bank of the pan. Within this zone it was located at the base of a tree, approximately 2 m from the pan’s edge and was not calling. Two Foam Nest Frogs were found sitting on a branch 1 to 1.5 m above the water. The Plain Grass Frog was calling from the grass on the edge of the pan. The African Bull Frog was not found in the pan, but in a puddle on the side of the road. Calls from five additional species were also recorded (Table 2). The pan at Punda Maria was a temporary pan. It had steep sloping edges that were covered in dense tall grasses and small trees that continued into the water. Forty-one mm of rain fell the day before and 16 mm on the day of frogging (Table 1). The night of frogging was characterized by warm weather and high winds. Two species of frogs were caught at this site (Table 2). Many individuals of Waterlily Reed Frogs were found on the surface of the water calling from the waterlily pads, and the Painted Reed Frogs were found calling from the reeds. It is important to note that all of the frogs were found in their described zone. The Mottled Shovel-nosed Frog, Bubbling Kassinas and Banded Rubber Frogs were found on the banks of the pans. The Painted Reed Frogs and Waterlily Reed Frogs were all found in the reed bed zone. The foam nest frogs were found in the tree zone on overhanging branches. Finally, the Broad-banded Grass Frog and Plain Grass Frog were found in the inundated grass zone. The African Bull Frog was not found in the pan but was found in a rain-filled pool on the side of the road which is the typical location for African Bull Frog breeding (Carruthers 1995).

Discussion Our data suggest that the three main factors affecting frog species diversity in Kruger National Park are pan type and size, pan location in KNP, and weather conditions (both long term and daily). These three factors represent the patchiness of the system and it is difficult to pinpoint which type of variance influenced our data the most. For example, Punda Maria and Skukuza had very different results, but since the pans varied in terms of type and size, location, and weather it is difficult to determine which factor was the greatest controller in species variation. Punda Maria has higher rainfall than Skukuza, and therefore should have more species and a higher numbers of frogs, but this was not the case. There were few frog species in the pan at Punda Maria, of which we only captured two. This could be because we went frogging in Punda Maria on 24 March 2004, more than a month later than in Skukuza. The timing during the rainy season could have greatly affected this difference. Long term rain patterns for Punda Maria show that the rainfall in October 2003 was high while in November and December 2003 the rainfall was low. The rainfall levels rose again in January 2004 (Table 2). This pattern could have caused the frogs to begin their mating and egg laying at the beginning of October, suggesting that the frog level would have been higher in earlier months in response to the high rains. The Skukuza pan also could have had higher species diversity because it is a more permanent, bigger pan. Furthermore, the vegetation structures around the two pans were quite different. Another reason could also be that when we went frogging in Skukuza we had weather conditions that are typically conducive with breeding behaviors. It rained on day one, five and six and we sampled on day seven, a day that was hot. The night air was still and humid – perfect for frogging. When we were in Punda Maria, it did rain prior to collecting and got warmer the day of frogging, but still it was extremely windy while we were out on the pan. Shingwedzi’s low catching numbers could have been caused by the false start of the rainy season (Table 2). It is hard to assess accurately the state of herpetological fauna in South Africa due to the inaccuracy of inventories. Numerous inaccuracies have been found in museum data collection including outdated taxonomy and missing specimens (Baard et al). It is known, however, that twenty-one (23%) of South Africa’s frog species are in the red category of endangered species. One example is the Table

48 Mountain Ghost Frog (Heleophryne rosei) which is being threatened by habitat degradation due to the damming of streams, alien vegetation, and reduced stream flow. It is recommended that all sites of this threatened taxon be included in the conservation action plan (Baard et al). This project could be of great value for monitoring reasons, as well as valuable learning experiences for future groups. With frog numbers decreasing, this project could help from year to year to monitor the changes in the frog population. However, we would like to make a few recommendations. The first would be to pick one variable to test for its effects on frog species diversity. For example, frogs could be collected from the same pan on multiple nights throughout the rainy season in order to analyze frog species diversity in relation to timing in the rainy season. Collections could also be made every night for a week in order to analyze the importance of daily rain patterns. It would also be effective to collect on a single at several different types of pans in the same location of Kruger National Park. This would make it possible to test the influence of varying pan size or type while keeping weather and location within the park constant. In order to test the variance in species diversity based on location within the park, three similar size/type pans could be tested on the same night. This would allow weather and pan size/type to be kept constant. It is crucial that sites be surveyed over a narrower time span with larger collection frequency. No matter how this project is approached next year, a larger sample size is needed (frogging must take place on more than just one night at each sight). Another factor that would help next year is to have a consistent resource person with the frogging group. Vincent Carruthers was an extremely valuable resource to have along in Skukuza as was Abri de Buys at Shingwedzi and with their help we were able to identify many of the frogs just from their call. This helps to identify rare species that may be in danger thus helping the groups efforts towards conservation.

Acknowledgements We would like to thank first of all Vincent Carruthers for everything (including writing the book on frogs). We would also like to thank all the students and faculty who came out with us into the field to help catch the frogs. We would also like to thank the game guards for keeping us safe. Finally, the Broad

Literature cited Baard, E. and de Villilers, A. 2000. State of Biodiversity: Western Cape Province, SouthAfrica Amphibians and Reptiles. Western Cape Nature Conservation Board. Stellenbosch, South Africa. Carruthers, V. and Passmore, N. 1995. South African Frogs: A Complete Guide. Witwatersrand University Press: Johannesburg. Carruthers, V. 2001. Frogs and Frogging in Southern Africa. Struik Publishers: Cape Town.

Table 1. Daily rainfall for a week before collection at each of the three sites Skukuza 8-Feb 9-Feb 10-Feb 11-Feb 12-Feb 13-Feb 14-Feb Daily rainfall (mm) 2.3 0 0 0 25 0.1 0 Shingwedzi 27-Feb 28-Feb 29-Feb 1-Mar 2-Mar 3-Mar 4-Mar Daily rainfall (mm) 14.8 27.6 trace trace trace 13 22.7 Punda Maria 18-Mar 19-Mar 20-Mar 21-Mar 22-Mar 23-Mar 24-Mar Daily rainfall (mm) 0.2 0.1 0.3 0.1 0 41 16.2

Table 2. Frog species found in three areas of the Kruger National Park Frogs of KNP SkukuzaShingwedzi Punda Maria 1 Shovel-footed Squeaker Arthroleptis stenodactylus 2 Northern Pigmy Toad Bufo fenoulheti 3 Olive Toad Bufo garmani 4 Guttural Toad Bufo gutturalis 5 Flat-backed Toad Bufo maculatus 6 Red Toad Schismaderma carens 7 Mottled Shovel-nosed Frog Hemisus marmoratus sp* call** 8 Painted Reed Frog Hyperolius marmoratus sp sp 9 Waterlily Frog Hyperolius pusillus sp sp 10 Golden Leaf-folding Frog Afrixalus aureus 11 Red-legged Kassina Kassina maculata 12 Bubbling Kassina Kassina senegalensis sp 13 Brown-backed Tree Frog Leptopelis mossambicus 14 Bushveld Rain Frog Breviceps adspersus 15 Banded Rubber Frog Phrynomantis bifasciatus sp sp 16 Common Platanna Xenopus laevis 17 Tropical Platanna Xenopus muelleri sp 18 Common Caco Cacosternum boettgeri call 19 Dwarf Puddle Frog Phrynobatrachus mababiensis call 20 Snoring Puddle Frog Phrynobatrachus natalensis call 21 Ornate Frog Hildebrandtia ornata sp 22 Plain Grass Frog Ptychadena anchietae sp 23 Broad-banded Grass Frog Ptychadena mossambica call call 24 Sharp-nosed Grass Frog Ptychadena oxyrhynchus 25 Striped Grass Frog Ptychadena porosissima 26 African Bullfrog Pyxicephalus edulis sp 27 Common River Frog Rana angolensis 28 Clicking Stream Frog Strongylopus grayii 29 Tremolo Sand Frog Tomopterna cryptotis 30 Knocking Sand Frog Tomopterna krugerensis 31 Russet-backed Sand Frog Tomopterna marmorata 32 Foam Nest Frog Chiromantis xerampelina sp sp *sp=species was collected **call=no specimen was collected but the call of this species was identified

Table 3. Monthly rainfall data for three sites in Kruger National Park Year 2003-2004 Location July Aug Sept Oct Nov Dec Jan Feb March Skukuza 0 0 31.7 18.6 24.6 22.5 208.2 153.7 81.1 Shingwedzi 0 1.1 9.5 83.6 24.8 141.7 0.9 98.1 80.2 Punda Maria 0.4 0 7.8 105.8 52.9 30.5 146 140.3 172.1 Averages from 1961-1990 Location July Aug Sept Oct Nov Dec Jan Feb March April May June Skukuza 10 6 26 35 76 84 93 87 73 33 14 10 Shingwedzi 3 5 1235 5184 59 71 38 29 9 4 Punda Maria 6 5 28 48 77 65 64 84 23 32 15 3

Punda Maria Faculty Field Projects

52 The use of matrix models for calculating sustainable utilization rates of natural resources

Category: Independent Project Participants: Justine Norman, Benjamin Wigley, Laurence Kruger (editor) and Christo Marais (Working for Water) Site: Punda Maria, Kruger National Park, Mpumalanga Province, South Africa

Key words: Natural resources, matrix modeling, sustainable harvesting and stocking rates

Abstract: Many rural communities living in savannas on the outskirts of the Kruger National Park suffer from poverty and unemployment. These communities are surrounded by large areas of communal lands which often offer a vast array of natural resources. The Italian funded project aimed to address these differences and come up with ways of increasing the social, economic and environmental benefits from the area. The communal area surrounding Malumalele was visited and the main floral components were assessed to determine potentially utilizable resources. A spreadsheet model was constructed to determine sustainable stocking rates for livestock based on the grass productivity. Matrix models were constructed for three woody species; Acacia nigrescens, Peltophorum africanum and Dichrostachys cineria. These were then used to ascertain the most suitable size classes for harvesting and sustainable harvesting rates. Our preliminary models suggest that current livestock stocking rates are appropriate and could possibly even be increased. The grass model predicts that approximately R550 000/annum from livestock alone could be generated from 5000 hectares. The matrix models predict that approximately 10% or seven adult A. nigrescens trees could be harvested per hectare. While 15% of all juvenile P. africanum and D. cinerea plants (approximately 40 plants per hectare) could also be sustainably harvested. However more data would need to be collected to determine the necessary time interval between harvesting events for each species.

Introduction The history of the Kruger Park is a long and controversial one (Carruthers 1995). The implementation of the park has had different meaning and consequences to the different racial groups living in the area. In the past 100 years many of the indigenous people previously living within the parks borders have been forcibly removed and resettled in areas adjacent to the park (Carruthers 1995). Many of these people have not benefited from the improved infrastructure to the area and income generated from the park (Carruthers 1995). In many of the rural communities surrounding the park unemployment is rife and poverty proliferates. Many households are thus forced to supplement the low incomes generated from employment by utilizing any potential natural resources found in their vicinity. A long term Italian funded project has been set up to address these issues and aims to enhance the productive potential of the area (Project Draft 2001). Savannas are extremely productive ecosystems with a vast array of resources which can be utilized by humans. Humans have in fact been utilizing savannas for millions of years for food and other resources (Mary Scholes pers comm.). The floral component of savannas consists of three main layers including the grass, shrub and tree layers all which have the potential to be sustainably utilized by humans. We propose to construct a number of models which will act as guidelines for a sustainable harvesting program for the communal lands surrounding the park. Each one of the three layers will be considered independently as they represent different exploitable resources. Many of the tree species growing in savannas are slow growing and highly resistant to insect damage, resulting in extremely hard wood which makes them a very valuable resource which can be used for timber, carving and building (Cunningham and Davis 1997). A major component of savannas consists of the grass layer which can be utilized for grazing domestic livestock. These livestock may be used to provide an array of products such as meat, milk and leather to the household, alternatively the livestock may be sold to generate income. The third layer we hope to investigate is the shrub layer which may be utilized for fence posts, medicinal uses or for browsing by goats. Although many other natural resources exist in savannas the three above mentioned layers will be the focus of this study. Aims To visit one of the communal areas found on the border of KNP where will ascertain what resources are best suited to sustainable utilization. Once we know what we can use we hope to determine how much can be used using modeling to predict long term harvesting rates. Matrix modeling can be used as a powerful tool to determine the life history strategy of a woody species. Once the life history strategy of a species is known it can be determined which stages are most suitable to harvesting. We aim to determine which of the tree and shrub species will be most suitable for sustainable

53 utilization. For each of these species we will construct a matrix model which will be used to determine the major demographic hurdles limiting population expansion and the maximum sustainable harvesting levels and the most appropriate size class/classes to be harvested.

Methods Aim 1 We constructed a grass model based on the understanding of the abiotic template in the area and suitable management strategies. Aim 2 We ascertained the present condition of the veld in this area and determined what potential tree species were utilizable on a sustainable basis. For these species we constructed models which could then be used as guidelines for sustainable utilization by the local communities. Study area The study was undertaken in a communal area surrounding Malumalele, a rural village found on the north west of the Kruger Park. Data collection In the area of study five random quadrats of 400m2 were sampled. In each quadrat all woody plants were identified, the height estimated and DBH (diameter at breast height) was measured. Each plant was then divided into one of five size classes according to height, the size class divisions can be seen in Figure 1. For the grasses a transect was walked for 100m whereby every ten meters a radius of 1m was sampled and all species present were identified. For each species the palatability, abundance and increaser/decreaser status was determined. Data analysis The grass data was then used to construct a model which incorporated season, rainfall, variability in rainfall, the preceding biomass, stocking densities and fire regime. This model was used to estimate the productivity of grasses which can be used as a surrogate for potential livestock biomass generated by the communal lands. The model was then run for a number of iterations which allowed us to predict long-term sustainable stocking rates. For the tree species we constructed matrix models by first constructing a transition probability matrix using the five size class categories. Due to the fact that we only sampled the plant populations once we had to estimate fecundity, mortality and recruitment for each of the models we constructed. These transition probability matrix (TPM) models were then used to predict future changes in the populations for each size class after each iteration (in this case a fire). This was done by multiplying the TPM by the state vectors, which are the total numbers of each size class of the population at the time of sampling (t1). This allows the population growth rate lambda (λ) to be calculated according to the following equation: N(t2)/N(t1). The model was run for twenty iterations for each species to model the population dynamics over time. We then graphed the data generated by the matrix to observe the different changes in each size class and for the total population. The TPM’s were then used to estimate sustainable harvesting rates by readjusting the appropriate components of the model according to harvesting levels and ensuring that λ remains above one.

Results We ran the grass model using mean summer rainfall of 361mm and mean winter rainfall of 89mm; we added variability to this by adding and subtracting 35% to both summer and winter rainfall to simulate the natural rainfall pattern. The model then incorporated this by using a random rainfall figure falling between the two extremes for summer; this was used as an indication of whether it was a good or bad growing season. The rainfall was used to estimate current grass biomass based on up to the previous three years rainfall cycles. Biomass was estimated as 7000kg/Ha after a wet year, if this was followed by another wet season it would increase to 9000kg/Ha and if followed by a third wet year would increase to 10000kg/Ha. When biomass reached 10000 kg/Ha the modeled biomass returned to zero to incorporate fire into the model. The model was run for an area of 5000 Ha, using R250 as an estimate of the value of a large stocking unit (LSU). This information could then be used to calculate the annual income generated from grazing livestock. Table 1 is an example of an outcome of the model showing the annul income for fifteen years. In the fiftenn years the model predicts a total income of R8250 000 which works out to an average of R550 000 per year. The species composition, abundance and palatability of the communal grazing lands in Malumalele were estimated from a transect (Table 2). Seven of the nine species that were found along the transect can be seen to be increaser species, with nearly 60% of all grass in the area being palatable. TPM’s were constructed for three of the most common species found in the study area, these being Acacia nigrescens, Peltophorum africanum and Dichrostachys cineria. The number of individuals of Acacia nigrescens, Peltophorum africanum and Dichrostachys cineria in each size class at (t1) are given (Figure 1). Acacia nigrescens’s TPM is shown in Table 3 and the modeled numbers of each size class run for twenty

54 iterations is shown in Figure 2. Using the model we predicted that 10% of size classes four and five could be harvested before the population crashed. Based on current numbers this works out to approximately seven adult trees per hectare. Once the growth rate for A. nigrescens was known the length between harvesting events could be calculated. The TPM for Peltophorum africanum is shown in Table 4. There are only four size classes as the sampled area didn’t have any plants in size class 5. The TPM also shows the majority of plants to be in the smaller size classes. Figure 3 shows the modeled growth and λ of P. africanum for twenty iterations. The TPM for D. cineria is shown in Table 5 and the predicted population growth rate and λ are shown in Figure 4. The matrix model predicted that 10% of size class 1, and 5% of size class 2 could be harvested before the population crashed. This equates to approximately 40 plants per hectare, because these are in the small size classes the time between harvesting is likely to be short.

Discussion The model we constructed to estimate potential income generated from grass productivity (Table 1) was fairly complex. It managed to incorporate rainfall and the variability in rainfall which is so typical of these parts of the world. However although we managed to incorporate some of this variability into the model it must be realized that rainfall cycles also operate on long term cycles. Rainfall data for the Kruger National Park as shown in Venter et al. 2003 shows there to be a ten year cycle of above average and below average rainfall periods. It would be possible to include this in the model however that would be above the scope of this study. The other main shortcoming of the model we constructed was that it assumed the whole area’s biomass returned to zero when biomass reached 10 000 kg/Ha simulating a fire. If this were to happen there would be no forage for the livestock in the entire area. To overcome this our model would ideally break the communal area into say four separate zones which would be burned separately, thus every year only a quarter of the area would be burned ensuring that the livestock would still have enough forage at all times. However it needs to be realized that this would require an effective management strategy which would be extremely difficult to implement in a communal area such as Malumalele where no one person or authority is in charge of managing the land. The occurrences of uncontrolled fires in such areas are common and would be almost impossible to prevent. If it were possible to implement a controlled fire burning regime we could construct four different models, one for each zone with each one incorporating previous biomass and likelihood of fire. The income from the four zones could then be added to arrive at total income generated by the communal lands. We need to stress that the model we constructed was just a preliminary one to give a rough estimate of potential stocking rates. With more research into rainfall, current ecological state of the area and available area it would be possible to construct a potentially accurate model to simulate income. However no matter how good the model we need to realize that there will always be some unpredictable components in the model due to the sheer stochasticity of nature. The communal lands surrounding Malumalele appear to be in a fairly disturbed state which is further emphasized by the high abundance of increaser grass species which are indicative of disturbance. However of the nine grass species found along the transect, seven were increaser species (Table 2) which increase with increased disturbance. The majority of grass (nearly 60%, Table 2) in the area is palatable making the area more suitable for livestock. Due to the way we sampled the woody vegetation in the area we were unable to separate out the shrub species from the tree species. Instead we combined the two layers and looked for the three most commonly occurring and potentially utilizable species. The first species we constructed a matrix model for was Acacia nigrescens (Table 3). The TPM shows that A. nigrescens relies more on the larger size classes for recruitment into the small size classes and that the survivorship and recruitment into larger size classes by these small size classes is fairly low. Thus it would be unsuitable to harvest seedlings and saplings. The survivorship of the larger size classes is much higher and therefore suitable for being harvested. However overharvesting the large classes would prevent new seeds from recruiting into seedlings and could lead to the collapse of the population. We found that up to 10% of adults could be sustainably harvested. Thus it would appear that the most suitable use for this species would be to harvest timber from the adult trees. Another possible use for the species would be to harvest the bark which is rich in tannins (Schmidt et al. 2002). The model could be used to estimate the potential income for the entire area as we could work out the value of the timber and calculate the volume of harvestable timber. The TPM’s for Peltophorum africanum and Dichrostachys cinerea as shown in Table 4 & 5 show that these species have a different life history strategy to A. nigrescens. They seem to rely more heavily on resprouting then on recruitment from seeds. This is suggested by the high proportion of plants that remain in size classes one and two, with both species having some plants moving from size class two back to size class one. Furthermore when we collected the data it was found that the majority of plants for both species were from size classes one and two with very few individuals in the larger size classes (Figure 1). Thus by using the TPM’s we could estimate sustainable harvesting rates for both species by reducing the number of surviving plants in the small size classes and ensuring that λ remains above one (Figures 3 & 4).

55 The bark and roots of Peltophorum africanum can be used to make decoctions which are used in Zimbabwe as a panacea. Root decoctions and infusions can be taken for ascites, abdominal, nausea and a number of other ailments (Van Wyk & Gericke 2000). The small plants could therefore be harvested and sold for medicinal uses. Dichrostachys cinerea is an important medicinal plant in southern Africa, the leaves, bark and roots all have a number of medicinal uses used to treat an array of ailments (Van Wyk & Gericke 2000). The flexible branches may also be used to make hunting bows, the bark yields a tough fibre used for ropes and tying. Straight branches are used for durable fence posts, and the wood is a valuable source of high quality firewood which can be sold commercially in South Africa (Van Wyk & Gericke 2000). Dichrostachys cinerea therefore has huge potential to be harvested to generate income; it is also a fast growing highly resilient species. Furthermore the smaller size classes of Dichrostachys cinerea would be suitable for most of the above mentioned uses The main aim of this study was to introduce us to the concept of using models to estimate sustainable harvesting rates. To construct more realistic models we would need to collect more substantial data. We would also need a better idea of the life history strategies of the species for which we constructed matrix models. This would be used to advise the community on how to harvest and utilize the communal areas surrounding their villages in a sustainable way. This in itself would not be a simple matter as it would raise a whole suite of socio-economic issues. Conflict would be hard to avoid as the communities are often bound by traditions and may not be open to new suggestions from outsiders. Problems like who benefits and how the generated income would be distributed fairly among the community would need to be resolved. Thus it would be necessary to perform a socio-economic study in the area to resolve these issues before we could implement our advice on natural resource utilization. Many of the ecosystem services and natural products are currently being utilized as indicated by the high number of increaser grass species. While collecting data it was also clear that many of the tree species are being utilized for a number of uses. However our models indicate that the area, which many would consider to be highly disturbed still, has the potential to be further exploited. If this were to be done in a sustainable way according to our models the communities could benefit financially from the available ecosystem services in the area.

Literature Cited Carruthers, C. 1995. The Kruger National Park; a Social and Political History. University of Natal Press, Pietermaritzburg. Cunningham, A. B. and G. W. Davis. 1997. Human use of plants. Pages 475-506 in Cowling, R. M., D. M. Richardson and S. M. 1997. Vegetation of Southern Africa. Cambridge University Press. Proposal. 2001. Optimising Sustainable Development Opportunities for Rural Communities adjacent to the Limpopo Trans-Frontier Park. Implementing agency, Working for Water (DEAT). Schmidt, E., M. Lotter and W. McCleland. 2002. Trees and shrubs of Mpumalanga and Kruger National Park. Jacana, Johannesburg. Van Wyk, B.E. and Gericke, N. 2002. People’s plants; a guide to useful plants of Southern Africa. Briza, Pretoria. Venter, F.J., R.J. Scholes and H.C. Eckhardt. 2003. The Abiotic Template and Its Associated Vegetation Pattern. Pages 83-129 in Du Toit, J.T., K.H. Rogers and H.C. Biggs. 2003. The Kruger Experience; Ecology and Management of Savanna Heterogeneity. Island Press, Washington.

56 Table 1. An outcome of the grass model run for 15 years, the model incorporates variation in rainfall and previous biomass to provide present biomass which can be converted to stocking rates to provide an indication of annual income.

good/bad Annual Years Months Summer/Winter year Biomass Stock/ha Stock/ha/yr income 1 0 1 1 0 0 - 1 6 0 0 0 0 0 - 2 12 1 0 0 0 0 - 2 18 0 0 0 0 0 - 3 24 1 1 7 0.25 0 - 3 30 0 0 7 0.25 0.5 625,000 4 36 1 0 7 0.25 - - 4 42 0 0 7 0.25 0.5 625,000 5 48 1 0 7 0.25 - - 5 54 0 0 7 0.25 0.5 625,000 6 60 1 1 9 0.4 - - 6 66 0 0 9 0.4 0.8 1,000,000 7 72 1 1 10 0.5 - - 7 78 0 0 0 0 0.5 625,000 8 84 1 1 7 0.25 - - 8 90 0 0 7 0.25 0.5 625,000 9 96 1 0 7 0.25 - - 9 102 0 0 7 0.25 0.5 625,000 10 108 1 0 7 0.25 - - 10 114 0 0 7 0.25 0.5 625,000 11 120 1 1 9 0.4 - - 11 126 0 0 9 0.4 0.8 1,000,000 12 132 1 1 10 0.5 - - 12 138 0 0 0 0 0.5 625,000 13 144 1 0 0 0 - - 13 150 0 0 0 0 0 - 14 156 1 1 7 0.25 - - 14 162 0 0 7 0.25 0.5 625,000 15 168 1 0 7 0.25 - - 15 174 0 0 7 0.25 0.5 625,000 Total 8,250,000 Ave/yr 550000

Table 2. The grass species, abundance and palatability as estimated from the transect at Malumalele. The total biomass was estimated to be 60% of the grass biomass in Kruger at the time of sampling. Spp Inc / Dec % Abundance Palatabity % Palatable Urochloa mosambicensis Inc 2 34.48 palatable 34.48 Aristida congesta barbicolas Inc 2 13.79 unpalatable 0.00 Eragrostis superba Inc 2 13.79 palatable 13.79 Bothriochloa radicans Inc 2 10.34 unpalatable 0.00 Cymbopogon sp Inc 1 10.34 unpalatable 0.00 Heteropogon contortis Inc 2 3.45 unpalatable 0.00 Panicum maximum Dec 3.45 palatable 3.45 Themeda triandra Dec 3.45 palatable 3.45 Tragus berteronianus Inc 2 3.45 unpalatable 0.00 Total 55.17

57 Table 3. The transition probability matrix model calculated for Acacia nigrescens, 1 2 3 4 5 1 0.15 0 0 0.1 0.3 2 0.05 0.35 0 0 0 3 0 0.5 0.9 0 0 4 0 0 0.1 0.98 0 5 0 0 0 0.02 0.99

Table 4. The transition probability matrix for Peltophorum africanum. 1 2 3 4 1 0.4 0.2 0.2 0.3 2 0.5 0.4 0 0 3 0 0.2 0.85 0 4 0 0 0.15 0.95

Table 5. The transition probability matrix model for Dichrostachys cineria 1 2 3 4 1 0.85 0.1 0.2 0.1 2 0.1 0.8 0 0.1 3 0 0.1 0.9 0 4 0 0 0.05 0.99

30 Acacia nigrescens

N 25 Peltophorum africanum Dichrostachys cinerea 20

0 0 -1m 1 - 2m 2 - 4m 4 - 6m >6m Size class

Figure 1. The numbers of individuals in each size class for the three species at the time of sampling.

140

120

100

5 80 4 3 2 60 1 no of individuals

1 3 5 7 9 3 5 9 1 11 1 1 17 1 2 23 iterations

Figure 2. The number of individuals of Acacia nigrescens in each size class run for twenty iterations using the TPM model.

160 1.08

1.06 140 1.04 120 1.02

100 1

0.98 (N)

N 80 Peltophorum 0.96 lambda africanum Lambda 60 0.94

0.92 40 0.9 20 0.88

0 0.86 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 iteration

Figure 3. Total population size and λ for Peltophorum africanum, run for twenty iterations using the TPM shown in Table 4.

160 1.04

140 1.035

1.03 120

1.025 100 1.02 Population 80 Lambda 1.015 Population

60 Lambda value 1.01

40 1.005

20 1

0 0.995 1234567891011121314151617181920 Iterations

Figure 4. Total population growth and λ for Dichrostachys cineria over twenty iterations using the TPM shown in Table 5.

De Hoop Faculty Field Projects

61 Rhabdomys pumilio in the fynbos

Category: Faculty Field Project Participants: Blanchie Asberry, Michael Chazan, (editor), (editor), Deedra McClearn, Bob Timm, Julie Coetzee (resource people) Site: De Hoop, Western Cape Province, South Africa

Key Words: Acacia cyclops, fynbos, invasive alien plants, Rhabdomys pumilio

Abstract: The fynbos is a biome that is at great risk to invasive alien plants such as Acacia cyclops. The rooikrans seed weevil had been introduced as a form of biocontrol and appears to be effective. Rhabdomys pumilio is a seed-eating rodent that is abundant in the fynbos and preys on seeds from the A. cyclops. One concern that arises is, whether reduction in the A. cyclops (through effective insect biocontrol) will lead to reduced populations of R. pumilio and consequent reductions in birds of prey that may rely on the rodents as a food source. We discovered that there were no significant differences in habitat choice, weight, gender, and trap success of R. pumilio in adjacent areas of the acacia and fynbos vegetations. Furthermore, rodent densities were high. Effective control of A. cyclops would not appear to be a threat to R. pumilio populations.

Second Language Abstract: Fynbos i biome leyi yinga nghozi kuya emahlweni na swimilana leswi swinga tumbulukangiki eka yona ku fana na Acacia cyclops. Mbewu ya rooikrans weevil yi tivisiwe ku fana na biocontrol naku tlhela yi humelela kahle. Rhabdomys pumilio ixi hatyana lexi xi tyaka mbewu na swona swihadyana leswi switele endzhawini ya fynbos na kona swidya mbewu ya A. cylcops. Lexi xingha ha humelelaka i ku, R. pumilio nhloyo ya swona loko kuri na swinyenyana leswi swiswi dyaka yinga ha andza njhana loko kungari na biocontrol. Hiku lavisisa hikume leswaku aku ngari naku hambana ka kahle eka ndzhawa leyi ku tshamaka swihadyana, ntiko rimbewu na switlhaka leswi aswi tirha eka R. pumilio e xikarhi ka swimilani swa acacia na fynbos.

Introduction The fynbos biome is famous for the richness and diversity of its plants, many which are locally endemic. Fynbos communities are highly susceptible to invasion by alien trees and shrubs (Homes et al, 1987). A. cyclops (from Australia) was introduced and used with A. saligna in the second half of the century to stabilize the shifting of sands on the Cape Flats (Bromilow 2001). The seeds of this species are black with bright red seed stalks and the plant produces large amount of seeds that form an enormous seed bank in the soil. The rooikrans seed weevil (Melanterius servulus) was released in 1994 and seems to be effective in controlling rooikrans (Acacia cyclops). Rhabdomys pumilio (striped mouse) is abundant in the fynbos community. This seed-eating rodent plays a major role in controlling the regeneration of many shrub species by consuming a large portion of the seeds that are released after fire (Cowling and Richardson, 1995). Because R. pumilio feeds on the seeds of the fynbos, such as A. cyclops, and because the weevil appears to be effective in controlling A. cyclops, it is important to determine if elimination of A. cyclops would have an affect on the R. pumilio population. In order to evaluate the R. pumilio populations in the fynbos, we asked the following three questions: Are the R. pumilio populations high? Are there more R. pumilio in the acacia or fynbos vegetation? Do individuals use both habitats?

Methods This study was conducted in the fynbos community of De Hoop, Western Cape Province, South Africa on 13-15 April 2004. Two trap lines were set, one on the side of the road with the fynbos vegetation and the other with the A. cyclops vegetation. The distance between the two trap lines was approximately 50 meters apart. On 13 April traps were set ten meters apart from each other on both trap lines. Cotton nest material and a mixed seed bait were placed within each trap. On 14 April there were two trap checks, one in the morning and one in the afternoon. The data collected included the following: trap number, species name, weight gender, age, and reproductive condition. Mice that were caught in the fynbos

62 vegetation received a spot of black hair dye on the upper region of their bodies near the chest and those caught in the Acacia vegetation received a spot of black hair dye on the lower region of the bodies near the thigh area and were then released. On 15 April a final trap check was conducted in the morning and again data were collected. Finally a Chi Square test was performed to analyze the weight of R. pumilio between the two habitats. Also, a Contingency Table was created to analyze gender of R. pumilio by site.

Results We had an extremely high trap success rate (Table 1). The average trap success for the fynbos was 57.7% and 59.4 for the A. Cyclops. The total number of mice captured was 62 in the A. cyclops and 60 in the fynbos. Excluding recaptures the total numbers of mice known alive were 50 for the A. cyclops and 53 for the fynbos. Only one mouse was recaptured in the different habitat. This mouse was marked on the fynbos in the morning and recaptured in the A. cyclops. Male R. pumilio are heavier than males (Kruskal-Wallis P<.0019; Table 2) but weights of the animals captured in the two habitats are not different (Kruskal-Wallis P<.4654; Table 2). There is no significant difference between the numbers of females and males (G-test P<.54; Table 3) as well no significant difference between the number of mice in the two habitats (G-test P<.54; Table 3).

Discussion The trap succession in De Hoop National reserve was extraordinarily high for both the A. cyclops and fynbos habitats. While the average trap success for many prior trials in North America hovered around 5% the average for these trials was 58.4% (McClearn 2004). The high trap success and number of mice captured suggests that there is a large R. pumilio population and with equal numbers of mice captured on both sides of the road the population is homogenous. Compared to the surrounding fynbos the A. cyclops is a relatively small habitat so the removal of these invasive species would not have a huge impact on the total area of rodent habitat. The mice did not prefer the fynbos or the A. cyclops even with the large quantity of A. cyclops seeds as a food source. Removing the A. cyclops would not destroy the R. pumilio’s primary habitat or food source and a significant drop in population would not be expected. Considering the already large population already present a slight decrease in numbers would not have a detrimental effect on the species or on their predators. In addition to their habitat, the R. pumilio do not seem to be dependant on the seeds of the A. cyclops for sustenance. Mice in the A. cyclops and the Fynbos show no significant difference in their weight (table 2). Availability of food in both habitats does not seem to be a factor in the growth of mice so the dietary value on each side of the road should not be a major issue if the A. cyclops were removed. If the alien vegetation was removed there is evidence that the mice would be able to move and adapt to the Fynbos environment. One of the mice captured in the Fynbos was able to cross the road and was recaptured in the A. cyclops hence it is possible for the mice to simply switch their habitats.

Acknowledgements We would like to thank Deedra McClearn for help with organizing this faculty field project, driving everyone to and from the fynbos and assisting with analyzing data. Also, we would like to thank Bob Timm, Julie Coetzee, and Laurence for their assistant with organizing this faculty field project, driving everyone to and from the fynbos and their knowledge. Next, we would like to thank the entire OTS Spring 2004 students for their valuable help with setting traps and collecting data. Finally, we would like to thank Godfrey Sekhula for translating our second language abstract into Shangaan.

Literature Cited Bromilow, C. 2001. Problem plants of South Africa: a guide to the identification and control of more than 300 invasive plants and other weeds. Briza Publications. Pretoria, SA. Cowling, Richard and Richardson, Dave. 1995. Fynbos: South Africa’s unique floral kingdom. Fernwood Press. Vlaeberg, SA. Holmes, P. M., MacDonald, I. A.W., and Juritz, J. 1987. Effects of clearing treatment on seed banks of the alien invasive shrubs acacia saligna and acacia cyclops in the southern and south-western Cape. Journal of Applied Ecology 24: 1045-1051.

Table 1. Trapping success by site. Date & Time Acacia Cyclops Fynbos 14th AM 32.4% 44.1% 14th PM 74.3% 60% 15th AM 71.4% 68.6%

Table 2. Average weight of capturea by site. Number Average Weight (g) Standard Deviation Acacia 49 39.9 12.55 Fynbos 47 42.8 10.71

Table 3. Sex of captures by site. Female Male Total Acacia 21 27 48 Fynbos 23 23 46 Total 44 50 94

64 An investigation into home range size, home range overlap, and use of runways in the striped mouse, Rhabdomys pumilio (Sparrmann) in the South Western Cape, South Africa.

Category: Faculty Field Problem Participants: Jasper Slingsby and Simon Thomson (secretaries), Deedra McClearn (resource person and editor), Bob Timm (resource person), Julie Coetzee (resource person) and Laurence Kruger (resource person). Site: De Hoop Nature Reserve, Western Cape, South Africa

Key words: Cape Floristic Region, home range size, runway utilization, seed predation, scatter- hoarding

Abstract: Rhabdomys pumilio is a widespread, common and predominantly granivorous rodent species in the CFR, yet little is known about the ecology, habitat use, and its role in the Cape Floristic Region. It is thought to be potentially important as a seed predator, a seed disperser, and as a prey species for carnivores. This study investigates the area used in 24 hours, the overlap in these areas, and use of runways by R. pumilio, in an attempt to learn more about the microecology of the species in the De Hoop Nature Reserve. The spool and line tracking technique was used to asses the size of the area utilized in 24 hours and microhabitat use. There was a trend for males to have greater mean home range area, maximum polygon dimension, estimated thread used, and proportion of distance moved along runways, but only the mean estimated thread used was significantly different between males and females (Z=2.12132, P=.05). It was suggested that it is possible that territoriality amongst males, and limited key resource areas could limit population sizes and densities.

Introduction The Cape Floristic region (CFR) is an extremely botanically diverse area in which there are many potentially sensitive, fine-scaled interactions between organisms (Bond 1994). Invasive Australian Acacias such as Acacia cyclops and Acacia saligna pose threats to both native biodiversity and water availability in the South Western Cape (Holmes et al. 1987). The effectiveness of mechanical clearing programs has been limited by regeneration of populations from large seed banks (Holmes et al. 1987). Seed predation by vertebrates, and Rhabdomys pumilio in particular, could be an important factor in limiting these seed banks. It has been hypothesized that the presence of extensive stands of Acacia cyclops in the De Hoop Nature Reserve has resulted in elevated numbers of Rhabdomys pumilio individuals, and that these large populations of mice support large raptor populations in the reserve. It is feared that the release of recent biocontrol agents to control Acacia cyclops may result in reduced R. pumilio numbers, and in turn affect the raptor population in the reserve. High densities of granivorous vertebrates have significant effects on seed survival in the region (Bond and Breytenbach 1985, Pierce and Cowling 1991). This should have important effects on the populations of plant species, and the vegetation community structure as a whole. Recent research has found scatter-hoarding by rodents to be an important form of seed dispersal in the CFR (Midgley et al. 2002). It is as yet unclear exactly which species are responsible for this phenomenon, but much attention is being paid to the granivorous rodents in the area (Jeremy Midgley pers comm.). R. pumilio is a widespread, common, predominantly granivorous rodent species in the CFR (Skinner and Smithers 1990). The species as been well studied throughout its distribution range, with particular focus on agricultural areas and areas with dense grass cover (Skinner and Smithers 1990). Studies have been performed on the demography and population dynamics of R. pumilio in the South Western Cape, with particular reference to invasive Australian Acacias (Holmes et al. 1987). There seems, however, to be little literature on the ecology of individuals of the species in the CFR and how they influence on the vegetation at a finer scale. This study aims to investigate some aspects of the ecology and habitat use of Rhabdomys pumilio so as to provide a basis for future investigation into the role of the species in the De Hoop Nature Reserve, and the CFR in general. The size of areas utilized by Rhabdomys pumilio individuals in 24 hours, variations in the size of these areas associated with gender and body weight, and runway utilization are investigated.

65 Methods Study site The study was performed in the Strandveld vegetation type in the De Hoop nature reserve in the South Western Cape, South Africa. The vegetation is short and comprised predominantly of Restionacea and ericoid shrubs. Study species A good description of Rhabdomys pumilio and its basic ecology is given in Skinner and Smithers (1990). They are predominantly diurnal rodents with peaks in activity between 0500 hrs and 0830 hrs in the morning and 1430 hrs and 1730 hrs in the afternoon. Data collection Thirty-five Sherman traps were set, each ten metres apart, in a straight transect parallel to the coastline. Of the captures four male and four female individuals were kept and weighed. The spool and line tracking technique was used to assess home range size and microhabitat use (Ryan et al. 1993, Stokes and Slade 1994). Spools of 1.5 grams (less than five percent of each mouse’s body weight), approximately three by one centimetres, of 100 metres of thread each, were glued to the backs of the mice between their scapulae. The ends of the threads were secured, and the mice released at their individual sites of capture at approximately 1600 hrs in the afternoon the day after they were captured. The trails were assessed approximately 24 hrs after the mice were released. The outermost points of the trails were marked and measured so that a convex polygon could be constructed, allowing the area the area they utilized in 24 hours to be calculated. The actual length of thread laid out by each individual before the thread was broken or the spool was groomed off was estimated. To give an indication of how the mice utilized runways the proportion of the trail that was on runway, off runway and in clearings was noted for twenty continuous metres of the thread. This was done after the first five metres of the thread as they would be expected to be in flight for at least this initial distance. Several qualitative observations were made about the pathways taken by the animals. Data analysis The convex polygons were used to calculate the area used by each individual. These areas were then compared between sexes, and overlap between individuals assessed. The longest dimension of each polygon was calculated to provide an alternative indication of home range size. Mann-Whitney U tests were performed, using STATISTICA 6.0, to investigate differences in home range area, longest polygon dimension and length of thread used between sexes.

Results The differences of the mean estimated thread used was shown to be significant, with the male R. pumilio individuals utilizing more thread than the females (Z=2.12132, P<.05, Figure 1). The males appeared to utilize larger areas, have longer maximum polygon dimensions and spend proportionately more time on runways than females, but the differences were not shown to be significant (Figures 2, 3 and 4 respectively). Utilized areas seemed to be orientated lengthways up slopes. Generally there was little overlap between the areas utilized, but this may have been because most of the mice were released relatively far apart. The utilized areas of a male and a female that were released ten metres apart overlapped to a large extent. The home ranges of two males that were released ten metres apart, however, did not overlap at all. R. pumilio individuals appear to move predominantly along runways inside their home range (Figure 4). Droppings and various other plant detritus was found on, or alongside sheltered sections of the runways. Roughly five centimeter long sections of restio culm were also found on the runways, these are typically cut by the Vlei rat Otomys laminatus, but a recent rodent community study in the area did not trap any O. laminatus individuals (Asberry and Chazan unpublished data).

Discussion Male and female utilized areas appear to overlap. Further study is needed to be sure that males are territorial. Larger sample sizes and a more directed approach is required to evaluate territoriality and home range overlap satisfactorily. If R. pumilio individuals are territorial then their home ranges would be mutually exclusive. This would limit the population size and density. Population densities have important effects on seed predation pressure, or availability of scatter- hoarding dispersers. Male-female overlap in home ranges is also important as there is often divergence in diets between sexes, often varying with season or reproductive status (Skinner and Smithers 1990).

66 The utilized areas could have been orientated length ways up slopes so that different habitat types were included in the home ranges. The species composition and structure of the vegetation changed very rapidly with small changes in altitude. It would be interesting to investigate whether each home range contains a similar combination of habitat types. It is possible that some habitat types are key resources required by each individual. If this is the case then limits on this resource could limit population size and density. Runways are thought to form as the result of regular movement of mice along the same route (Skinner and Smithers 1990). Elephant shrews (Macroscelididae) use their runways as escape routes along which they can move very rapidly, and actively maintain them so that there are no obstructions that could hinder their movement. The high proportion of thread that lay on runways could be an indication that the mice usually move predominantly along runways. It could, however, have merely been that the mice were trying to escape their captors or remove the spool, and were fleeing along their runways in an attempt to do so. The use of other tracking methods which would have less influence on the behaviour of the mice, such as dusting them with florescent powder and following the trails at night with an ultra violet light, should be investigated. Differences between the methods could be compared to investigate the effect of the spool and string method on the mice. The way in which they use their runways should be further investigated. This could be done by placing obstructions on the runways and seeing whether or not the mice remove them so as to maintain an unobstructed escape route. The possibility that different species share the same runways should be investigated by performing similar tracking studies on other species caught in the same area. The implications for runway use on the vegetation should be investigated. Seed predation on and off runways could be studied by placing seeds on and off runways and assessing if there is a difference in predation pressure. If R. pumilio is found to be important in scatter-hoarding of Fynbos seeds then the location of hoards in relation to the runways could be investigated by tagging seeds using the method described in Midgley et al. 2002. It is clear that there is much room, and need, for further research into the fine-scaled behaviour and ecology of Rhabdomys pumilio, and indeed other rodent species in the De Hoop Nature Reserve and the CFR, in order to assess their influence on the vegetation and predator populations. Our intention with this paper is to highlight the areas of particular interest, and has provided some insight into the pros and cons of the methods that can be used to do so.

Acknowledgements Special thanks to Deedra McClearn and Bob Timm for guidance when designing the methods for, and assistance in executing this experiment. Thanks to the OTS South Africa class of 2004 for assisting in data collection.

Literature cited Bond. W. J. 1994. Do mutualisms matter? Assessing the impact of pollinator and disperser disruption on plant extinction. Phil. Trans. R. Soc. Lond. 344: 83-90. Bond. W. J. and Breytenbach. G. J. 1985. Ants, rodents and seed predation in Proteaceae. S. Afr. J. Zool. 20(3): 150-154. Holmes. P. M., MacDonald. I. A. W. and Juritz. J. 1987. Effects of clearing treatment on seed banks of the alien invasive shrubs Acacia saligna and Acacia cyclops in the Southern and South Western Cape, South Africa. Journal of Applied Ecology 24: 1045-1051. Midgley. J., Anderson. B., Bok. A. and Fleming. T. 2002. Scatter-hoarding of Cape Proteaceae nuts by rodents. Evolutionary Ecology Research 4: 623-626. Pierce. S. M. and Cowling. R. M. 1991. Dynamics of soil stored seed banks of six shrubs in fire- prone dune Fynbos. The Journal of Ecology 79(3): 731-747. Ryan. J. M., Creighton. G. K. and Emmons. L. H. 1993. Activity patterns of two species of Nesomys (Muridae: Nesomyinae) in a Madagascar rain forest. Journal of Tropical Ecology 9:101-107. Skinner, J. D. and Smithers, R. H. N. 1990. The mammals of the Southern African subregion. Second edition. University of Pretoria, Pretoria. Stokes. M. K. and Slade. N. A. 1994. Drought-induced cracks in soil as refuges for small mammals: an unforeseen consequence of climate change. Conservation Biology, 8(2):577-580

67 70

40 String distance estimate (m)

25 Mean MF ±SE Gender

Figure 1. A whisker plot showing the means and standard errors for the estimated amount of thread used by three male (M) and four female (F) Rhabdomys pumilio individuals. A Mann-Whitney U test found the difference between sexes to be significant (Z=-2.12132, P<.05).

320

300

280

260

240

220

200 ) 2 180

160 Area(m 140

120

100

20 Mean MF ±SE Gender

Figure 2. A whisker plot showing means and standard errors of utilized areas for three male (M) and four female (F) Rhabdomys pumilio individuals. A Mann-Whitney U test showed a near significant difference between the sexes (Z=-1.76777, P=.077).

68 36

20 Maximum distance across polygon (m) polygon across distance Maximum 18

14 Mean MF ±SE Gender

Figure 3. A whisker plot showing the means and standard errors of the maximum dimensions of the polygons for three male (M) and four female (F) Rhabdomys pumilio individuals. A Mann-Whitney U test found there to be no significant difference between the sexes (Z=-0.707107, P>.05).

1.1

1.0

0.9

0.8

Proportion on runway 0.7

0.6

0.5 Mean MF ±SE Gender

Figure 4. A whisker plot showing the means and standard errors for the proportion of a 20 metre section of the threads that three male (M) and four female (F) Rhabdomys pumilio individuals moved along runways. A Mann-Whitney U test found there to be no significant difference between the sexes (Z=-1.106066, P>.05).

Skukuza Independent Projects

70 Mating strategies of the dung beetle Kheper nigroaeneus

Category: Independent Project Participants: Blanchie Asberry, Eric Caldera (editor), Zoë Layton (editor), Julie Coetzee (resource person) Site: Skukuza, Kruger National Park, Mpumalanaga Province, South Africa

Key words: “active” partner, dung beetle, Kheper nigroaeneus, pronotum width, Scarabaeidae

Abstract: Male dung beetles within the tribe Scarabaeini are most often the “active” partners in making and rolling brood balls. Here I observed the hypotheses of who’s making and rolling dung balls in the dung beetle Kheper nigroaeneus. It has been shown that ball size is correlated with beetle size and that under competition, ball construction time and ball volume are both reduced. I tested these studies by measuring and weighing dung beetles and their dung balls. My finding is that pronotum width is correlated with ball size and is the most important determining factor of dung ball size.

Second Language Abstract: Manzemu wa xinunu e nthlhambini wa Scarabaeini hiyena anga mahlweni hiku endla tibola naku ti khunguluxa. Laha ndzi kambele hiku vhumba ka dyondzo leyi ndzingana yona leswaku hiwihi loyi a endlaka naku khunguluxa ti bola entlhambini i Kheper nigraeneus. Switi komba kahle leswaku manzemu un’wana na un’wana u endla bolo yaku n’wi ringana. Sweswo ndzi swi kume hiku kala ntiko wava manzemu, ku kula ka vona na tibolo ta vona. Ndzi kume leswaku ku anama ka nhlana wa manzemu un’wana na un’wana ku ringanana na bolo leyi a yi kunguluxaka, hikuva hlana wa mansemu i xirho xa nkoka loko swi fika kaku endla naku khunguluxa bolo.

Introduction Dung beetles are classified into the family Scarabaeidae. This family contains about 5000 species and many subfamilies. In Africa, south of the Sahara, there are more than 2000 species of dung beetles in the family Scarabaeidae (Hanski and Cambefort 1991). Though there are many species, this study focused specifically on the diurnal Large Copper Dung Beetle (Kheper nigroaenues) of South Africa. There are three types of behavioral or functional groups that dung beetles are classified into: the dwellers, the tunnelers, and the rollers or telocoprides, paracoprides, and endocoprides, respectively. The rollers are the most sophisticated of the three. These beetles make dung balls and subsequently roll them away from the central food source to avoid competition from other beetles (Hanski and Cambefort 1991). Dung is very important to these beetles for the following reasons: it is the primary food source for larval, nutrient content and size of the ball determines adult body size, and dung ball size is also used by females in sexual selection. There are usually two types of dung balls made: food balls, which are also known as nuptial balls, and brood balls. In some species, the male offers a food ball (nuptial ball) to the female. Scarabaeidae larvae develop in brood balls prepared by females or by a pair of beetles (Hanski and Cambefort 1991). An “active partner” is the individual that initiates the making of the brood ball. The ball that is being made or completed may act as a sexual display for the “passive partner.” In the Scarabaeini and Canthonini tribes, most often the active partner is the male, but in the Gymnopleurini and Sisyphini tribes it is the female. The Kheper nigroaeneus falls into the Scarabaeini. The brood ball is often rolled by two partners, but in some cases the female climbs on top of the ball and is rolled off with it by the male. Competition in dung beetles is severe for rollers; however, competition for space in the soil is eliminated by the transport of the dung ball away from the food source. It is paradoxical that competition appears to be most frequent and most severe in the rollers, where only one form of competition- adult competition for food is possible, and least frequent and intense in the dwellers, in which both adults and larvae may compete for both space and food (Hanski and Cambefort 1991). Under competition, ball construction time and ball volume are both reduced in K. nigroaeneus. This could be a terrible disadvantage to smaller dung beetles because of the relationship between beetle size and ball size; therefore smaller beetles creating smaller dung balls than larger beetles. Although ball

71 construction time is reduced under competition, larger beetles still have an advantage over smaller beetles because of their size; thus, still enabling them to create bigger dung balls. There is also innumerable and direct interference competition among rollers in attempts to steal dung balls from one another. Fighting for a dung ball by rollers is the most conspicuous kind of interference in dung beetles (Hanski and Camefort 1991). The objective of this study was to test the following hypotheses: males and females participate in making and rolling dung balls; males make and roll dung balls more than females. This will be done by measuring and weighing dung beetles and their dung balls.

Methods This study was carried out at the Skukuza golf course, in Kruger National Park, Mpumalanaga Province, South Africa. The experiment was conducted from 23-25 February 2004; however due to rain on 24 February 2004 no dung beetles were available to be surveyed. To ensure that a reasonable sampling size was taken, 21 Large Copper Dung Beetles were analyzed. While walking around the golf course, beetles were observed and data was collected. Once a dung beetle had completely made a ball, I determined the gender of the beetle. Males were identified by a dark patch of little hairs on the hind tibiae. Females, on the other hand, lack these dark patches of little hairs. Beetles and their dung balls were measured. The diameter of the dung balls and one of the morphometric measurements of the beetles (pronotum width) was made using a calliper. Following that, beetles and their dung balls were separately weighed in grams using a hanging balance. Then, the beetles were observed while rolling the dung to see if another beetle would fly in to steal it. Finally, linear regressions were made between prontoum width and ball size, pronotum width and ball weight, ball size and ball weight, pronotum width and beetle weight, and beetle weight and ball size.

Results Upon observing dung beetles and determining their gender, only males were observed making and rolling the balls. Data were statistically analyzed to see if there was a significant relationship between the diameter of the dung ball and pronotum width of the beetle. As a result, there was a positive, significant relationship between the two variables (r2 = 0.606, P<.05, F = 29.286) (Figure 1). Between pronotum width and ball weight (r2= 0.179), ball size and ball weight (r2= 0.030), pronotum width and beetle weight (r2= 0.238), and beetle weight and ball size (r2= 0.282) there was no significant relationship. No beetles were observed stealing and fighting for the balls.

Discussion Due to the fact that only males were observed, my data did not determine whether females make and roll dung balls; however, according to Tomkins et al. (1999), females also construct and roll dung balls alone. My hypothesis about males making and rolling dung balls more than females was also supported. The fact that Large Copper Dung Beetles belong to the tribe Scarabaeini explains why only males were observed making and rolling the dung balls. The active partner is most often the males in the tribes Scarabaeini and Canthonini (Hanski and Camefort 1991). Males construct large dung balls and wait for female beetles to arrive, rolling the ball away without assistance from the female who clings to the rolling ball (Tomkins et al. 1999). The results from the linear regression between pronotum width and ball size demonstrate that in the ball rolling scarab Kheper nigroaeneus, pronotum width is positively correlated with the size of the dung ball (Figure1). The size of the dung ball is related to the size of the beetle that has prepared it (Hanski and Camefort 1991) (Table 1). The relation between body size and dung ball diameter in the K. nigroaeneus may arise as a consequence of competition for dung balls: dung ball size is reduced under competition in K. nigroaeneus and ball diameter may therefore reflect the competitive ability of the beetle (Tomkins et al. 1999). In the remaining linear regressions analyzed there was no significant relationship. Although one could expect a significant correlation between beetle weight and ball weight, there was none. This might be due to two reasons: new beetles are lighter than old beetles; therefore, it might have affected the ball weight regarding the ability of the beetle to push a heavier ball; and because the hanging balance used to weigh the beetles and their balls was not precise enough. This could have skewed the data. In the future, a hanging balance with a more precise measurement should be used. Also, this study should not be conducted when it is raining. During the rolling process, other beetles often attempt to steal the ball. The attacker is most often a male (Hanski and Cambefort 1991). In fights over dung balls, the larger beetles tend to win. Larger beetles, with less chance of loosing their ball to another beetle, might therefore construct larger dung balls than smaller beetles” (Tomkins et al. 1999).

72 In conclusion, the most important determinant of dung ball size is pronotum width in this study. This study supports the notion that body size is a condition independent trait in dung beetles.

Acknowledgements Firstly, I would like to thank Julie Coetzee for assisting me throughout this study with suggestions and collecting data. Secondly, I would like to thank Laurence Kruger for his suggestions in the revision of my study before it began. Next, I would like to thank Fahiema Daniels for assisting me in the process of saving photographs. Then, I would like to thank Eric Caldera for assisting me with the procedure of linear regression and analysis of data. Also, I would like to thank Lucas Masinga for assisting me with looking for dung beetles as well as being an excellent game guard. I would also like to thank Godfrey Sekula for translating my second language abstract into Shangaan. Finally, I would like to thank Mike Smith and Deedra McClearn for driving me and picking me up from sites.

Literature Cited Hanski, I. and Cambefort, Y. (1991). Dung beetle ecology (Chapters 1, 4, 8, 9, 17). Princeton University Press: Princeton, NJ. Tomkins, J. L., Simmons, L. W., Knell, R. J., and Norris, K. A. (1999). Correlates of ball size and rolling speed in the dung beetle Kheper nigroaeneus (Coleopetra: Scarabaeidae). J. Zool., Lond. 248:483-487.

Table 1: Relationship between body size and ball size in rollers (fresh weight in g) (Hanski and Cambefort 1991).

y = 0.0305+0.357*x 2.2

2.0

1.8

1.6

1.4 Pronotum Width

1.2

1.0 3.0 3.2 3.4 3.6 3.8 4.0 4.2 4.4 4.6 4.8 5.0 5.2 5.4 5.6 Ball Size

Figure 1. The significant positive relationship between pronotum width and the diameter of the dung ball that the beetle was rolling (r2 = 0.6065, P< .05).

74 Dung beetle fidelity in Kheper nigroaneus

Category: Independent Project Participants: Zoe Layton, Simon Thomson Site: Skukuza, Kruger National Park, Mpumalanga Province, South Africa

Key words: celestial cue, dung pat, Kheper nigroaneus, polarized light, rhabdom

Abstract Ball rolling dung beetles often roll their dung balls towards certain celestial cues including the sun, the moon and polarized light. These cues are utilized by detecting the concentric circles in the sky with receptors located on the dorsal part of the eye. It is of interest to examine whether or not, due to these celestial cues and receptors, dung beetles have fidelity to specific angles on a species level and an individual level. In order to test the fidelity of K. nigroaneus to a single angle of orientation, a sample of individual beetles were subjected to repeated trials where their angle of orientation was measured. It was found that K. nigroaneus individuals have fidelity to a single angle of orientation and this was displayed over two consecutive days.

Samevatting: Bal rolende miskruiers rol dikwels hul misballe in die rigting van sekere hëmellike aanwysings wat dieson, maan en poliseerde lig insluit. Hierdie aanwysings word verbruik deur ontdekking van konsentries kringe in die lug met opneemings wat in die dorsaal deel van die oog gelokaliseer is. Dit is belangrik om te ondersoek as miskruiers spesifieke hoeke verkies op ŉ spesie en individueel vlak in respons na hierdie aanwysings. Om te toets of Kepher nigroaneus ŉ enkelle hoek verkies was ŉ monster van individueel miskruiers verbruik en hulle was ondergeskik na reeks proefneemings waar hulle hoek van orienteering gemeet was. Ons vindings dui aan dat K. nigroaneus idividueelle slegs ŉ enkele hoek van oriëntering verkies en dit was oor twee daë bewys.

Introduction Dung is a valuable but patchy resource. Certain dung beetles construct a ball as a transportable resource unit that can be rolled away for a long or short distance. Most other beetles, however, do not roll balls, but rather take dung into tunnels that they dig directly beneath the dung pat, without ever exposing them to the soil surface (Bernon 1981, Cambefort and Hanski 1991 in Byrne et al. 2003). Ball rolling is considered to be a derived behaviour that evolved in a few species in response to competition for food and space within the dung pat and below it (Bernon 1981, Cambefort and Hanski in Byrne et al. 2003). Once a patch of dung is found, ball-rolling dung beetles form their ball of dung and roll it away as straight as the terrain will allow (Dacke et al. 2003). There are three factors that the ball-rolling dung beetle must consider when rolling the ball away, namely; the slope of the terrain, the wind direction and the sun. Ball rolling dung beetles will generally roll uphill as it is hard to maintain control of the ball rolling downhill or perpendicular to the slope. They also roll with the wind for similar reasons, but this is not an overriding factor. The last important factor in choosing a direction is the sun, which is important in maintaining a straight line course away from the dung pat (Matthews 1963). In addition to the sun, dung beetles use other celestial cues such as lunar cues and polarized light cues (Byrne et al. 2001). Polarized light is invisible to humans as it is in the UV-light range and is created by the scattering of light in the atmosphere overhead. The pattern of polarized light will encircle the sun, surrounding it like a ring of concentric circles (Gould and Gould 1995). If the dung beetle can associate images in the sky with corresponding directions in space, it can use the sky as a reference point (Whener 1984). The receptors that are believed to be used for polarized light analysis are restricted to the dorsal part of the dorsal eye in the dorsal rim area (DRA) (Dacke et al. 2003). In Scarabaes zambesianus (Coleopter: Scarabaeidae), there are receptors used to detect polarized light in the ommatidia of the DRA of the eye, which differ structurally from the rest of the eye (Dacke et al. 2003). The microvilli of the retinula cells form light absorbing rhabdoms which, in polarization sensitive insects, are arranged at only two orthogonal orientations (Dacke et al. 2003). Dung beetles are thus able to detect polarized light and orient themselves in a straight direction for ball rolling with respect to the sun. The hypothesis that individual dung beetles have fidelity to a certain angle of orientation will be tested by quantitatively analyzing their behaviour when rolling dung balls.

Methods To determine the rolling fidelity of K. nigroaneus, we collected 20 individuals using dung baited pit-fall traps, as well as by hand on the Skukuza golf-coarse in the Kruger National Park. Each individual was labelled with a number, one through twenty. Sex of the beetles was not identified; both males and females were used. Buffalo dung was used throughout the study, as it was easily available for collection from the buffalo enclosure at the Kruger National Park. Our study site was in the Skukuza campsite of the Kruger National Park. The trials were conducted on a circular arena, 2m in diameter. The arena was levelled, and made flat so as to eliminate the effects of slope and obstacles altering orientation angles. The dung was placed in the centre of the arena and moulded into a circular shape so as not to affect the choice of angular orientation. Each individual beetle, male or female, was placed in the centre of the dung pat and allowed to roll to the edge of the arena. The beetles were not rolled consecutively but placed on the dung pat in a random fashion. Each time a beetle rolled it was placed back on the dung pat and forced to construct a new ball, thereby forcing the beetle to make a new choice of orientation each trial. Each beetle was forced to roll three times. The position of the beetle’s ball was measured by pulling a string taut from the centre of the dung pat to the position of the ball. A compass was used to measure the bearing of the string. The bearing was recorded onto a data sheet. The study was performed over two consecutive afternoons between 16h00 and 18h00. The weather conditions, approximate percent cloud cover and position of the sun were also recorded. Circular statistics were used to analyze our data, more specifically, the Hotelling test for paired samples of angles. The Hotelling test was first used to compare the three trials for nine individuals on day one. In order to compare the angles of the individuals who rolled on both day one and two (n=4) we took the average for each individual on each day. We then used the Hotelling test to analyze the presence or absence of significantly different rolling angles by the same individual on a different day.

Results The experiment was conducted over two consecutive days and we found that nine individuals completed three trials on day one (Table 1) and seven individuals completed three trials on day two (Table 2). Of these individuals, four of the individuals rolled on both day one and day two. Using the data from day one we compared the angle of roll one to the angle of roll two and found that there was no significant difference between the two angles (F=0.68,<4.47). We then compared the angle of roll one to the angle of roll three which also showed that there was no significant difference between the two angles (F=0.73<4.47). Finally, we compared roll angle two to roll angle three showing that none of the rolling angles had a significant difference from each other (F=1.69<4.47)(Figure 1). We then compared the rolling angles of the four individuals who rolled on both afternoons. Using the mean angle for each of these four individuals (Table 1) we found no significant difference in the rolling angles of individuals on separate days (F=1.02<19.0) (Figure 2). It is important to note that on the first day there was partial cloud cover but the sun was directly visible. On the second day there was 100% cover and so the sun was not directly visible.

Discussion This study demonstrates that beetles do, in fact, have fidelity to certain angles. Structurally, this is due to the fact that the structure of the eye is designed to pick up celestial cues. Research has found that rhabdoms in beetles do not twist long their lengths which suggests that the photoreceptors are polarization sensitive (Wehner et al. 1975 in Dacke et al. 2003). These polarized light analyzers are also found in spiders and the dorsal rim area in the compound eye of many other insects (Dacke et al. 2001). To account for angle fidelity on the individual level it is interesting to consider the possibility of genetic variance in eye structure determination. If the genes that make up the rhabdoms are slightly varied between the individuals in the population, this may cause individuals to line up slightly differently with polarized light. Therefore, on the individual level, not on the species level, individuals would have a small margin of angle choice. On the species level, the margin would be larger but still in reference to a particular cue, usually the sun or the moon. Our data shows strong conviction for this suggestion as individual beetles had a margin of angle orientation as small as twelve degrees as in the case with beetle number five’s rolls on day one. On the species level, the margin of angle distribution was much greater. Reasons that dung beetles may have adapted this unique polarization analyzer on their dorsal eye could be both predation and competition. For many animals it is advantageous, due to the risk of

76 predators and dehydration, to find their way back home as quickly as possible (Dacke et al. 2002). Although ball rolling dung beetles do not have a specific destination, it is presumably advantageous for the dung beetle to roll away from the dung pat in a rapid manner. Since the shortest distance between two points is a straight line, it is advantageous for the dung beetle to immediately establish a direction in which to roll, in order to decrease the amount of time spent around the dung pat. This enables the dung beetles to be less vulnerable to predators. Since our data strongly supported our hypothesis, showing no significant difference between roll one and roll two, roll one and roll three, or roll two and roll three, it would be interesting to examine a larger sample size. It is believed that with a greater number of beetles, and a greater number of trials our data would strongly support an important aspect of dung beetle behaviour. It would also be fascinating to compare K. nigroaneus to another species of dung beetle.

Acknowledgements Thanks to Marcus Byrne for help with circular statistics.

Literature Cited Byrne, M., Dacke, M., Nordstrom, P., Scholtz, C. and Warrant, E. 2003. Visual cues used by Ball- rolling dung beetles for orientation. Journal of Comparative Physiology. Dacke, M., Nordstrom, P. and Scholtz, C. 2003. Twighlight orientation to polarized light in the crepuscular dung beetle Scarabaeus zambesianus. The Journal of experimental biology. 206:1535- 1543. Gould, J. and Gould, C. 1995. The Honey Bee. U. S. A.: Scientific American Library. Matthews, E.G. 1963. Observations of the ball-rolling behaviour of Canthon pilularius. Psyche 70:75- 93.

77 Table 1. Three rolling angles and the mean rolling angle of nine individuals of Kheper nigroaneus on day one. Beetle # Roll 1 Roll 2 Roll 3 Mean 10 300 310 66 225 4 302 282 300 295 14 48 62 64 58 5 268 280 276 275 12 66 0 8 25 8 238 248 180 222 16 284 266 270 273 15 40 60 14 38 3 180 184 280 215

Table 2. Three rolling angles and the mean rolling angle of nine individuals of Kheper nigroaneus on day two. Beetle # Roll 1 Roll 2 Roll 3 Mean 12 48 310 292 217 10 290 282 26 199 17 48 308 280 212 18 42 38 20 33 3 140 50 32 74 15 34 20 52 35 6 180 218 234 211

Table 3. Average rolling angles of four individuals on two consecutive days. Beetle # Day One Day Two 10 335 314 12 24 331 15 38 35 3 210 70

78 0º

Figure 1. Three rolling angles of nine individuals of Kheper nigroaneus on day one.

0 º

Figure 2. Average rolling angles of four individuals on two consecutive days.

79 Male competition in pollinating and parasitizing fig wasps in Ficus sycomorus synconia.

Category: Independent Project Participants: Stephanie Johnson Site: Skukuza, Kruger National Park, Mpumalanga Province, South Africa

Key words: Agaonidae, fighting, fig wasps, local male competition, synconia

Abstract: Some pollinating fig wasp males fight each other for access to females and there is a suite of morphological characters that can be used to predict whether a male fig wasp will fight or not. This experiment tested out the validity of using morphological characters to predict fighting behavior by examining the morphology and observing the behavior of male fig wasps in Ficus sycomorus synconia. Males of each species present in the sycamore figs were examined and their fighting behavior was predicted and then observed. The presence or absence of fighting behavior was correctly predicted for three out of the four species found in the figs, suggesting that fighting syndrome morphology is a good indicator of wasp fighting behavior.

Samevatting: Party betstuiwingde vy-perdeby mannetjies baklei met mekaar vir toegang aan vy- perdeby wyfies en daar is ŉ reeks morfologiese karakters wat verbriuk kan word om te voorspel as ŉ mannetjie sal baklei of nie. Hierdie eksperiment toets die geldigheid van die verbruik vanaf morfologiese karakters om geveg gedrag te voorspel deur die ondersoek van morfologie en waarneeming van die gedrag van die mannetjie vy- perdeby in Ficus sycomorus synconia. Mannetjies van elke spesies wat in die vye gevind is, was geondersoek, hul geveg gedrag was gevoorspel en waargeneem. Die afwesigheid of teenwoordigheid van die geveg gedrag was korrek voorgespel vir drie van die vier spesies van vy-perdebye species wat deel van die eksperiment was. Dit suggereer dat geveg morfologie ŉ goeie aanwyser van perdeby geveg gedrag is.

Introduction The fig wasp and fig tree system is one of the most fascinating examples of co-evolution between a plant and a pollinator. Fig trees belong to the genus Ficus in the family Moraceae. Ficus is one of the largest plant genera in the world with over 750 species (Cook and Rasplus 2003). Ficus species are characterized by synconia, or enclosed inflorescences of male and female flowers (Cook and Rasplus 2003; Otero and Ackerman 2002). Fig wasps belong to the family Agaonidae. Female agaonid wasps enter figs through the ostiole and oviposit in some ovaries after piercing the style, pollinating the flower in the process. Some agaonid wasps pollinate passively, carrying pollen scattered on their bodies and without any specific pollination behavior. Others actively pollinate the flowers using coxal combs on their legs to accumulate pollen in pollen pockets on their thorax and later deposit it into receptive flowers (Cook and Rasplus 2003). Occasionally, multiple foundresses or females will colonize the same fig. Wasp larvae develop inside the ovary thus preventing the flower from producing a seed. Male wasps are wingless, emerge early, and mate with the females while they are still inside their galls (Greeff et al. 2003; Otero and Ackerman 2003). In some species, males fight for access to females (Greeff et al. 2003). The males then use their mandibles to chew a hole out of the fig and usually fall out of the fig to their death or die inside of the fig (Greeff et al. 2003; Otero and Ackerman 2003). However, some male wasps have been observed to leave their natal fig and enter another fig on the same tree, sometimes over 50 cm away (Greeff et al. 2003). Female wasps escape through the male excavated hole and carry pollen from the mature flowers inside her birth fig to the fig that she pollinates and oviposits in (Otero and Ackerman 2002). Ficus sycamorus L. (Moraceae) is a fig tree species commonly found in riparian forests in East Africa from Sudan and Ethiopia to South Africa (Galil et al. 1968). The fig fruits are located on leafless clusters of branchlets arising from the major branches and trunk. The fruits are large and slightly flattened and edible. Young fruits are green and become yellowish or reddish when ripe (Schmidt et al. 2002). Ficus sycamorus has a suite of associated Agaonid fig wasps. Ceratosolen arabicus is the legitimate pollinator. Sycophaga sycomori, Apocrypta longitarsus, and other fig wasp species oviposit in the ovaries of female flowers but do not pollinate the flowers in the synconia (Galil et al. 1968). It was once believed that males either die in their birth fig or they fall to their death after chewing a hole in the fig (Greeff et al. 2003). Consequently, one would predict extreme local mate competition between brothers for access to females because they do not disperse. As a result, female agaonid wasp broods have a very female biased sex ratio to reduce useless competition between

80 brothers (Greeff 2002). The combination of high sex ratios and the relatedness of males are believed to prevent fighting in pollinating agaonid species (Greeff et al. 2003). On the other hand, non-pollinating males often fight. Several theories have been espoused to explain this phenomenon. Non-pollinating males are not related, roughly the same number of males and females are produced resulting in an equal sex ratio, and this increases the potential for conflict between males for access to females (Greeff et al. 2003). Also, the location of where the mating takes place also appears to favor competition between males. However, relatedness does not totally prohibit fighting because brother compete locally mating opportunities, thus negating any effects relatedness may have in limiting conflict (Greeff et al. 2003). In this experiment, I attempted to predict which species of male fig wasps found in F. sycomorus figs do and do not fight based solely upon their anatomy.

Methods This experiment was conducted from 23-25 February 2004 at the Skukuza camp in Kruger National Park, Mpumalanga Province South Africa. Ficus sycomorus trees in and around the Skukuza camp were located, the stage of development of the chalcidoid wasps inside the figs was determined, and two trees were found that contained wasps at the fourth or male development stage where mature males emerge from their galls, mate with females, and chew an exit hole out of synconia (Galil et al. 1968). Ten figs were collected from each of the trees. The synconia were bisected with a scalpel and five figs in which males had already emerged or were beginning to emerge in large numbers were selected. The female wasps had not yet begun to emerge from their galls. One specimen of each species of male found inside the fig was removed and examined to determine if the males had any of the traits that are associated with fighting males, including: sickle-shaped mandibles, large head, long antennal scape, antennae not projecting forward, pronotum broader than long, mesonotum, metanotum and propodeum strongly fused (Greeff et al. 2003) (Table 1). After assessing the presence or absence of these traits, I predicted whether the males from the different species would fight. Tentative male species identifications were made using the Galil and Eisikowitch 1968 paper on the biology of the wasps found in the synconia of F. sycomorus trees in East Africa. One half of the synconia was placed in a small petri dish, covered with plastic wrap to prevent desiccation, and allowed to sit undisturbed for 30 minutes to one hour. The males were then observed for any displays of fighting behavior such as pushing a male away from a female, fighting with clear exchanges of bites and displacement of the resident male, biting another’s head, picking up a male and throwing him to the side, attacking a male’s thorax from the bottom, or antennae grabbing and pushing (Greeff et al. 2003). Each fig was observed for 30 minutes and the presence and type of fighting behavior displayed were recorded. Species that did not show any fighting behavior during the observation period were considered to not be fighters.

Results Four different species of males were found in the synconia, C. arabicus, S. sycomori, and A. longitarsus, and a fourth unknown species. Based upon their morphology, it was predicted that S. sycomori, A. longitarsus, and species four males should all show fighting behavior because they each have six or more of the morphological characters associated with males that fight (Table 2). Ceratosolen arabicus only had one of the characters of the fighting syndrome, a long antennal scape, and thus was predicted to not fight. Greeff et al. (2003) found that most male wasps fall into one of two distinct behavior categories; they are either totally unaware of each other and show no aggressive interactions or they fight every time that males strive to mate with the same female. Apocrypta longitarsus and species four males showed aggressive behavior such as biting, pushing, or biting and lifting their opponent almost every time they encountered another conspecific male. No fighting or aggressive interactions were observed in C. arabicus or S. sycomori, indicating that these two species are not fighters.

Discussion Greeff et al. (2003) were able to predict whether several species of Agaonid wasps males would fight based upon their anatomy. They found that there is a fighting syndrome or series of morphological traits associated with fighting males that one can use to predict whether an Agaonid wasp species will have fighting males. Using this fighting syndrome, the fighting behavior of three wasp species, C. arabicus, A. longitarsus, and species four, was correctly prediced based upon their morphological characteristics. However, S. sycomori males were not observed to fight although it had been predicted that they would. When there were only a few males present in the fig being observed,

81 A. longitarsus and species four males would not fight. This suggests that there may be a threshold density above which males will fight and below which males will not interact aggressively. S. sycomori males were only found in two of the ten figs. In both figs, there were relatively few S. sycomori males. Thus, S. sycomori males may actually fight each other, but it was not observed because a fig with a high density of S. sycomori males was not examined. These results support the validity of using the fighting syndrome morphological characteristics to predict fighting behavior in Agaonid wasps.

Acknowledgements I would like to thank Deedra McClearn, Laurence Kruger and Julie Coetzee for providing me with scientific papers on fig wasps.

Literature Cited Cook, J., and J. Rasplus. 2003. Mutualists with attitude: coevolving fig wasps and figs. Trends in Ecology and Evolution. 18:241-248. Galil, J., and D. Eisikowitch. 1968. On the pollination ecology of Ficus sycomorus in East Africa. Ecology. 49: 259-269. Greeff, J. M. 2002. Mating system and sex ratios of a pollinating fig wasp with dispersing males. Proceedings of the Royal Society of London B. 269: 2317-2323. Greeff, J. M., S. van Noort, J. Rasplus, and F. Kjellberg. 2003. Dispersal and fighting in male pollinating fig wasps. Comptes Rendus Biologies. 326: 121-130. Otero, T., and J. D. Ackerman. 2002. Flower style length and seed production in two speciesof Ficus (Moraceae) in Puerto Rico. Caribbean Journal of Science. 38:249-251. Schmidt, E., M. Lotter, and W. McCleland. 2002. Trees of Mpumalanga and Kruger National Park. Jacana Earth. Johannesburg, South Africa.

Table 1: Fighting syndrome morphological traits in the males collected from the Ficus sycamorus synconia.

Ceratosolen Sycophaga Apocrypta Morphological Trait arabicus sycomori longitarsis other male falcate mandibles 0 1 1 1

large head 0 1 1 1

long antennal scape 1 1 1 0 antennae not projecting forward 0 1 1 1 elongate legs 0 1 1 1 narrower tibia and femur 0 1 1 1 pronotum broader than long 0 0 0 1 mesonotum, metanotum, and propedeum strongly fused 0 0 0 1

Table 2: Fig wasp species and fighting behavior observed Species Fighting (expected) Fighting (observed) Ceratosolen arabicus no no Sycophaga sycomori yes no Apocrypta longitarsis yes yes other male yes yes

82 Navigation of nocturnal flying Insects

Category: Independent Project Participants: Michael Chazan Site: Skukuza & Shingwetzi camp, Kruger National Park

Key words: angle of light, lunar navigation, nocturnal insects

Abstract: Nocturnal flying insects fly in a straight line by maintaining a constant angle with the brightest light in the sky, usually the moon. The moon is always above the horizon and therefore moonlight always strikes insects from the top and never from below. When observing whether insects would fly to a light source shining from below them in comparison to a light source shining from above, no insects flew to the light shining from below, while several flew to the lights source shining from above.

Samevatting: Nagtelike vliënde insekte vlieg in ŉ reguit lyn deur ŉ konstante hoek te bly hou met die helderste lig in die hệmel, wat gewoontlik die maan is. Die maan is altyd bo die horison en daarvoor raak dit altyd die insekte van bo af en nooit van onder af nie. Terwyl my ondersoek om uit te vind of insekte verkies om na ŉ lig vanaf onder of bo te kom het ek gevind dat daar min was wat na die lig vanaf onder gekom het. Die insekte het dikwels na die lig van bo gekom.

Introduction Over the last few hundred millions years, nocturnal insects have evolved a method for navigation by using the moon as a guide. In order to fly in a strait line insects maintain a constant angle to the brightest object in the sky (Damus 1996). Celestial bodies can be considered infinitely far away considering no matter how far the insects fly the angle will not significantly change. However, when an artificial light is brighter than the moon, the angle of the light changes as the insect moves in relation to the artificial light. As the angle of the light changes, the insect adjusts its path accordingly and spirals inwards, usually to its demise (Damus 1996). The moon is always above the horizon and therefore the moonlight rays always strike the insect from above and never from below. My hypothesis is that insects would never approach a light source shining from below them but would approach a light source shining down from above. The reason being there is an advantage to their survival to ignore light from below them in order to navigate.

Methods My study sites were the Skukuza and Shingwetzi camps in the Kruger National Park. I conducted two trials in Skukuza and one trial in Shingwetzi. I cut out a piece of cardboard one meter in length and .4 meters in width. In the middle of the rectangle two holes 1.5 cm in diameter were cut out. In each hole I placed a flashlight pointing in opposite directs. I then taped the barrier to a wall with equal amount of wall area above and below the barrier. For each trial I observed and tallied the insects that flew towards the light from above and below. An event would only be considered if the insects flew towards the light and either touched the light source or the barrier. Each trial was one hour long and began at 20h00. The trials were completed 23 February, 24 February, and 2 March respectively. The statistics performed was a χ2 – test.

Results As shown in Table 1 no insects approached the light shining up from below them. There were a number of insects that approached the light shining down from above them and were tallied as moths, beetles, or unidentified. The unidentified insects were tiny and could only be seen by their silhouettes so they could not be captured. For the χ2–test the number of each insect that approached from the top and the bottom were added and divided by 2 to calculate the expected value. My χ2 value was <.0000001 and with 41 degrees of freedom (P<.05). My results show an obvious distinction in that no insects approached the light shining up from below, while many did approach the light shining down from above (Figure 1). Nevertheless insects were attracted to the light shining up from below as I could see them in the light beam. However, these insects never flew down towards the light source and flew straight right through the light beam. The insects that were attracted to the light source shining down flew up until they touched either the flashlight or the cardboard barrier.

83 Discussion The biggest issue conducting this experiment was sample size. Due to rainy nights, battery power, and light intensity only three trials were managed to be completed during the time available with a sample size of 42 individuals. For further study a very powerful light source is suggested. Although the numbers may not be statistically viable there does seem to be a distinct trend from the observations, and if so the question still remains why. There is an evolutionary advantage for insects to not use light shining from below. Only celestial bodies are far enough away for their light rays to maintain a constant angle with an insect. Artificial light or light reflection can alter insects’ navigation, causing them to spiral into the water or ground disoriented. Natural selection would favor a method to only use a light source above the horizon for navigation. The way this method may be implemented is a topic for further study. The next question to be asked is whether the light is being analyzed or merely not being sensed. It may be the case that insects don’t sense light shining from below, or that they do sense the light but have some form of neurological analysis to distinguish between the two directs of light. My observation however showed that the insects do in fact sense the light. As noted in table 1 insects were attracted to the light source shining up, however the insects flew through the light and did not fly down and approach the light. There were other sources of error that also need to be considered. As mentioned before, my experiment did not test whether it was gravity and not the angle of the light that was causing the behavior of the insects. Insects may be able to sense the direction of gravity and therefore will know whether they are flying with or against the force of gravity. Insects also fly at many different altitudes while my experiment was a meter and a half off the ground. The concentration of insects at different altitudes may have had an effect on my data.

Literature cited Damus, Christian. 1996. The Evolution of Flight. http://hannover.park.org/Canada/Museum/insects/evolution/deforming.html

Table 1: Insects approaching light traps. Trial Location Light Shining Moths Beetles Non-Identified 1 Skukuza Up 0 0 0 1 Skukuza Down 5 1 13 2 Skukuza Up 0 0 0 2 Skukuza Down 4 0 10 3 Shingwetzi Up 0 0 0 3 Shingwetzi Down 2 0 8

Figure 1:

8 moths beetles 6 non-identified insects# of 4

0 up down up down up down

112233 Trial

Figure 1. Number of insects trapped at light traps according to lamp orientation.

84 Dispersal ability of the neonate instars of Dactylopius opuntiae (Homoptera: Dactylopiidae), a biological control agent of Opuntia stricta (Cactaceae) and the implicationsfor biocontrol in the Kruger National Park

Category: Independent Project Participants: Scott Briscoe, Kyle Harris, Shannon Hatmaker Site: Skukuza, Kruger National Park, Mpumalanga Province, South Africa

Key words: Biocontrol, Dactylopius opuntiae, dispersal, Opuntia stricta

Abstract: Opuntia stricta is an invasive plant species that Kruger National Park (KNP) is seeking to control by using the cochineal insect Dactylopius opuntiae. We studied the dispersal ability of D. opuntiae during the crawler stage by collecting instars and observing their angle and direction of movement towards or away from O. stricta. Results show that the D. opuntiae instars disperse randomly and do not move towards O. stricta cladodes. We conclude that the current method used by KNP for controlling O. stricta is adequate, but the introduction and implementation of an Integrated Pest Management plan (IPM) would further enhance the control of O. stricta.

Samevatting: Opuntia stricta is? invallende plant specie wat die Kruger Nasionale Wildtuin (KNW) wil beheer deur die verbruik van? cochenille insek Dactyloppius opuntia . Ons het die verstrooiheid vermo van D. opuntia gedurende die kruiper fase ondersoek, deur die versamelling van instars, en die hoek en rigting van hulle beweging jeen en weg van O. stricta was opgemerk. Uitslae bewys dat die D. opuntiae instars ewekansig verstrooi en hulle beweeg nie na die O. stricta cladodes nie. Ons gevolgtrekking was dat die korent manier wat gebruik is deur KNW vir die beheer van O. stricta voldoene is, maar dat die inleiding en implementasie van ? integreerde pes bestuur plan die beheer van O. stricta verder sal verbeter.

Introduction Currently there are 15 species of Cactaceae that are regarded as weeds of varying importance within South Africa (Moran and Zimmerman 1991), including Opuntia stricta. Until recently, O. stricta was considered to be a minor problem, but has now proliferated into many parts of South Africa. Although O. stricta is widespread, it is most prevalent in the Northwest Province and the Kruger National Park (KNP) within the Mpumalanga and Limpopo Provinces, where it is found in approximately 30 000 ha of conserved land (Foxcroft and Hoffmann 2000). Opuntia stricta is problematic due to its rapid spread and dense, impenetrable thickets. It is an aggressive invader and heavily infested areas are so severe that land is rendered useless. Chemical and mechanical control of O. stricta has been attempted, however due to the high rate of germination and seed dispersal by elephants (Loxodonta africana) and baboons (Papio ursinus), it is labor intensive and often, is not cost effective (Hoffmann et al. 1998). In addition, chemical treatments are ineffective as a result of smaller O. stricta in undergrowth escaping the treatment and easily replenishing. As a result of the labor intensity and cost ineffectiveness of mechanical and chemical control, biological control methods have been introduced. In 1987 the phycitid moth, Cactoblastis cactorum, was released; in 1993 the population peaked, but has since declined because the immature stages of the moth are highly susceptible to predation. A second control agent, Dactylopius opuntiae, the cochineal insect, has been established in South Africa since 1938 (Petty 1948). However, it was found to have limited effect on O. stricta. Further research into additional forms of biological control found that there were different biotypes of D. opuntiae which had a greater impact on O. stricta. In 1997 one of these more specific biotypes of D. opuntiae was imported from Australia and released in KNP (Lotter and Hoffmann 1998). The new biotype established quickly and destroyed a large number of cactus infested plots (Foxcroft and Hoffmann 2000). Despite establishment of D. opuntiae, dispersal is limited because once adult females are settled they become sessile (Foxcroft and Hoffmann 2000). Dispersal success is therefore dependent on the first instar nymphal stages of the insect. The nymphs either hatch on the plant and remain there, or make use of long wax filaments, which enables them to be lifted by air currents. They are then deposited at random and make their way to the plant (Foxcroft and Hoffmann 2000). Currently, dispersal in KNP is aided by deploying O. stricta cladodes infested with populations of D. opuntiae close to stands of O. stricta. This allows O. stricta to be brought under control more quickly.

85 Nevertheless, this is a time consuming exercise and more information is needed in order to form a more cohesive, integrated management plan for the control of O. stricta in KNP. Therefore the main aim of the experiment is to further understand the direction and dispersal in D. opuntiae instars by testing the following hypothesis: D. Opuntiae instars will disperse towards an O. stricta cladode in a non-random fashion.

Methods The study was conducted at Skukuza, Kruger National Park, Mpumalanga Province over a period of 2 days (24-25 February 2004). Approximately 2400 D. opuntiae first instar larvae were collected from infected O. stricta, which was taken from the Alien Biota Greenhouse. The instars were removed from the plants using pipette aspirators. Twelve paper sheets were used for the study. On each sheet, 500 cm diameter circles were drawn and divided into eight 45 segments. The sheets were then divided in half with four uninfected O. stricta cladodes, collected from outside of the Skukuza camp, placed in one half of the circle and at the edge. The positioning of each equally sized cladode corresponded with four of the segments. The half of the circle with cladodes was the cladode experiment and the half without was the control. The angles and directions at which the instars moved, were measured by placing 200 instars in the center of the circle. The total number of instars in each segment was counted every 10 minutes for one hour. This was repeated 12 times on 12 different sheets, using 200 new instars each time. Once all the data had been collected, statistical analysis was conducted using the Statistica program. A t-test was performed to compare the total number of instars found on each half of the diagramed sheets (cladode experiment and control). In addition, a Kruskal-Wallis test was performed to compare the number of instars found in each of the eight segments.

Results Instars appeared to be concentrated more on the control half of the divided sheets (Figure 1). However, when comparing the total numbers of the instars on both sides of the sheet we found that their distribution was random (t=1.4195, P=.1697) and neither side was preferred. The instars appear to have no preference for any of the segments (Figure 2). However, there is a higher total number of instars found on the control half of the segmented sheet (Figure 2). Further analysis using Kruskal-Wallis ANOVA showed that D. opuntiae instars had no preference for particular segments (X ²=8.7196, P=.2734).

Discussion The dispersal direction of the D. opuntiae instars is random. The instars do not move in any specific direction and do not show any preference for a particular segment. The instars also did not move towards the O. stricta cladodes. This may indicate that the dispersal of the instars is not based on sensory cues, but is rather based on physical dispersal techniques such as wind. Due to the slow and random dispersal qualities of the crawlers, colonization of O. stricta will be more successful if infestations of the plant are found in close proximity to one another. Long-range dispersal has been shown to be limited and is dependent on wind dispersal of the instar, which is also random (Foxcroft and Hoffmann, 2000). The current method of aided dispersal employed by the Alien Biota Unit in the KNP is probably the best way to maximize the colonization of O. stricta by D. opuntiae. However, this method is time consuming and is reliant on a database indicating the position of populations of O. stricta. Presently, control of O. stricta in KNP is based on the two biocontrol agents; C. cactorum and D. opuntiae. Control of O. stricta may be further enhanced by the implementation of an integrated pest management plan, involving not only biocontrol agents, but mechanical methods of removal as well as the use of pesticides. In addition to this, an in depth post-release analysis of the effectiveness of the biocontrol agents should be conducted. A follow up study should also be done in order to gauge the effectiveness of pesticides and mechanical control methods.

Literature Cited Foxcroft. L. C. and J. H. Hoffmann. 2000. Dispersal of Dactylopius opuntiae Cockerell (Homoptera: Dactylopiidae), a biological control agent of Opuntiae stricta (Haworth.) Haworth (Cactaceae) in the Kruger National Park. Koedoe. 43(2): 1-5. Hoffmann, J. H., V. C. Moran, H. G. Zimmermann 1999. Integrated management of Opuntia stricta (Haworth) Haworth (Cactaceae) in South Africa: an enhanced role for two, renowned, insect agents. African Entomology Memoir No. 1 (1999): 3-14 Hoffmann. J. H., V. C. Moran and D. A. Zeller. 1998. Evaluation of Cactoblastis cactorum

86 (Lepidoptera: Phycitidae) as a Biological Control Agent of Opuntia stricta in the Kruger National Park, South Africa. Biological Control. 12: 20-24. Lotter. W. D. and J. H. Hoffmann. 1998. An integrated management plan for the control of Opuntia stricta (Cactaceae) in the Kruger National Park, South Africa. Koedoe. 41(1): 63-68. Moran, V. C. and H. G. Zimmermann 1984. The biological control of cactus weeds: achievements and prospects. Biocontrol News and Information, Commonwealth Agriculture Bureaux 5: 297- 320. Petty, F. W. 1948. The Biological Control of Prickly Pears in South Africa. Union of South Africa, Department of Agricultural Science Bulletin, Entomology Series No. 22.

Figure 1. Histogram showing the total amount D. opuntiae instars within the experiment and the control groups.

Figure 2. Historgram showing the total number of D. opuntiae instars with in each 45 degree segment, for the experimental and control groups.

88 The effect of varying fire regimes on ant diversity in a savanna ecosystem

Category: Independent Project Participants: Eric Caldera and Jasper Slingsby (Secretaries), Shannon Hatmaker and Kyle Harris (Field assistants) Site: Numbi and Shabeni experimental burn plots, Pretoriouskop, Kruger National Park, South Africa

Key words: ants, biodiversity, fire regimes, Formicidae, Hymenoptera, species content

Abstract: Fire regimes are thought to have an important influence on biodiversity (in terms of savanna ecosystems). This study aimed to investigate the effects of fire frequency on faunal structure and composition, using ants as indicators. The plots in one site had greater species richness and diversity (S=19, H’=2.327 and S=17, H’=1.818) than the other (S=15, H’=0.6016 and S=15, H’=1.31). The overlap in species between the plots at the more diverse site, and the proximity of the two plots, suggest that immigration of species between the two plots could have resulted in elevated diversity. The other site may have been less diverse as there was no refugia from fire for species that prefer the structure of annually burnt plots, but are intolerant of burning events. It is also possible that this site did not conform to our predictions as a result of the unusually high density of a single species (Lankyblackass spp.) in the annual plot samples, which could have caused the exclusion of other species. Heterogeneity in terms of patchiness of different fire frequencies was deemed important as it allows migration of species between patches, and creates refugia for some species from either competition or disturbance.

Introduction Fire is an important phenomenon that drives ecological processes and influences evolutionary trajectories in many ecosystems. Fire frequency influences many aspects of savanna ecology such as vegetation structure and composition, and soil type and nutrient content (Bond and van Wilgen 1996, Van Wilgen et al. 2003). Consequently, these factors influence the overall biodiversity and species richness of savanna fauna. It is thought that very high fire frequencies could only have been caused by human influence because “natural” ignition sources, such as lightning, are comparatively rare (Scholes et al. 2003, Van Wilgen et al., 2003). Low fire frequencies, such as those that would occur in the absence of humans as an ignition source, usually result in very high fire intensities that often have very negative effects on the vegetation. This implies that the vegetation is not adapted for low fire frequencies. For this reason they are thought to be far less important as drivers for ecosystem function when considered in isolation from higher fore frequencies (W. Bond pers. comm.). Preservation of biodiversity is a high priority for global conservation efforts, and thus it is important to understand how it is affected by fire, especially with respect to the development of fire management policies in protected areas. Only a few long term fire regime experiments that allow investigation of biodiversity exist. One such experiment is based in Kruger National Park (KNP), South Africa. It is neither practical nor possible to survey the entire diversity of an ecosystem, so indicator taxa are often used to estimate overall biodiversity. Good indicator taxa meet the basic requirements of being highly speciose and easily sampled (Agosti et al. 2000). Ants (Hymenoptera: Formicidae) are one such taxon that meet these requirements, and were used in this study to represent general arthropod fauna in the fire plots. Many correlation coefficients have already been established between ant diversity and other taxa for many habitats, including savannas (Alonso 2002). The current hypotheses describing the influence of fire on ant diversity predict that higher diversity should be expected in areas with higher fire frequencies, when compared to areas with lower fire frequencies. This may be the case for two reasons. Firstly, fire events cause great stochasticity within the frequently burnt plots, preventing competitive interactions from running their course and causing competitive exclusion. If competitive exclusion were to take place, the result would be a reduction in species richness and/or evenness (Parr et al. 2002). Secondly, the less dense canopy cover in the frequently burnt plots results in less shading and greater fluctuations in temperatures throughout the day. It has been well documented that thermal envelopes greatly influence the distribution of ant species (Retana and Cerdá 1999, Kaspari 2000). For this reason greater temperature fluctuations within a single environment should provide greater niche space, allowing more species to co-occur. It is also possible, however, that some species will be intolerant of very frequent fires, causing a loss of species in the plots with higher fire frequency.

89 We predicted that there should be a significant difference in the species diversity across different fire regimes and in the species composition of ants occurring across fire regimes. We undertook a survey to investigate the ant diversity across plots of savanna hat have been treated with different fire regimes for the last 50 years to test whether high fire frequencies promote ant diversity. As high fire frequencies are thought to be an artifact of human influence. This study also investigates the effects of humans on the diversity of ants in savanna ecosystems.

Methods Site selection and sampling We surveyed ant diversity in the Shabeni and Numbi experimental burn plot (EBP) strings in the Kruger National Park (KNP), South Africa. These EBP’s are located in the Pretoriuskop region of KNP. Each EBP is approximately 6-8ha. At each string we surveyed two plots. A plot that had been artificially burned annually for 50 years, and a control plot that had not burned for the same time period. Annual burn plots were burned in August, in the middle of the dry season, and were thus of relatively high intensity. Control plots in KNP have not burned in over 50 years with the exception of the Shabeni string, which experienced a natural partial burn in 2002. To compensate for this we sampled an unburnt part of the Shabeni control plot. The control and annual burn plots in the Numbi string were directly adjacent to each other, while the two plots sampled in the Shabeni string were separated by four plots with deferring fire regimes. We used standard pitfall-trapping methods to sample ant diversity in the EBPs. In each plot we laid 20 traps in a 5x4 grid with each trap spaced 5m apart. Traps were left out for 4 days during February of 2004 and the ants collected were sorted to morpho-species. Analysis We calculated the total ant abundance (N), total species richness (S) and two diversity indices for each EBP. To calculate diversity, we used Pielou’s evenness index (J’) and the Shannon-Weiner diversity index (H’), which places bias on rare species. Multivariate analyses of absolute ant abundance data were done using PRIMER 5.0 (Clarke and Gorley 2001). Cluster anayses (CA) using group averaging and Bray-Curtis similarity measures were used to examine the relationship between samples at burn plots. Data were square-root transformed prior to analyses to reduce the weight of common species. Analysis of similarity (ANOSIM) was used to establish if there were significant differences in the ant assemblages on plots with different fire frequencies. This is a non-parametric matrices underlying sample ordinations (Clarke and Warwick 1994), in which global R-statistic provides an absolute measure of how separated groups are; a value close to one indicates significant differences between the assemblages compared, whereas close to zero indicates that they are barely separable.

Results The highest ant species richness was recorded for the Numbi annual (NA) plot, while the Numbi control (NC) plot contained slightly fewer (Table 1). Eleven of the species at this site occurred in both plots (Appendix 1). The control and annual plots at the Shabeni string (SC and SA respectively) both contained 15 species (Table 1) of which only five were common to both plots (Appendix 1). Abundance of individuals, N, was greatest for SA. NA had the second highest abundance, followed by NC and SC respectively. The two diversity indices yielded similar results. Rankings of values, from highest to lowest, for diversity indices H’ and J’ were: NA, NC, SC, and SA (Table 1). Control plots and annual plots at both Shabeni and Numbi were significantly different in ant species composition, R=0.489, P<.001. Pairwise tests for dissimilarity in ant species composition between all combinations of EBPs yielded significant dissimilarity between all combinations with the exception of NC-SC (Table 2). There was significant dissimilarity between plots sampled at Shabeni and Numbi, R=0.067, P<.032. An MDS ordination plot and a cluster analysis of all samples, using Bray-Curtis similarity values, showed all samples falling into two main clades with less than 20% similarity between them (Figure 2). One cluster contained 28 samples, 22 of which were controls while six were annual burns. The second clade contained 20 samples, 18 of which were from annual burn plots and two were from control plots. This clade split once again into the samples from the Numbi annual (six of eight) and Shabeni annual plots (ten of thirteen).

Discussion The significant difference between the samples from the annual and control plots reaffirmed the results of previous studies, that differing fire frequencies have an important effect on the diversity of ants in savannas (Parr 2002). Despite the significant difference between the Numbi and Shabeni

90 plots as a whole, the control samples were not significantly split between these sites (Table 2, Figures 1 and 2). This indicated that similar processes were affecting diversity in the two control plots. The annual samples were significantly split between the Numbi and Shabeni plots (Figure 1), indicating that different processes were affecting diversity in the two plots. The Numbi annual burn plot had greater species richness and diversity than the Numbi control plot as predicted (Table 1). This, however, was not the case at the Shabeni plot. The Numbi site had greater species richness and diversity than the Shabeni site (Table 1). It had a greater number of species shared between the two fire treatments too (Appendix 1). This could be because the Numbi plots are adjacent, whereas the Shabeni plots are separated by three other EPBs. The close proximity of the Numbi plots would have allowed immigration of species between the two treatments, resulting in higher species richness in both plots. The frequent fires in the annual plot would create habitat or refugia for species that cannot survive or compete in dense, unburnt vegetation, and the control site would provide refugia for species which can survive in both vegetation structures, but are intolerant of burning events. It is possible that there was low species richness and diversity in the Shabeni annual plot because there was no such refuge from fire (the neighbouring plots were burnt biannually). Previous studies have found ant habitat specificity to confound results and suggest that this is an aspect of ant biology that requires further attention when diversity studies are performed (Parr 2002). It could also have been caused by the dominant number of Lankyblackass spp. individuals found in the annual plot samples (Appendix 1), causing low evenness (Table 1). It is possible that the trapping grid was laid out in an area which had an unusually high density of Lankyblackass spp. when compared to the rest of the annual burn plot. Many ant species are known to have erratic patchy distributions (Kaspari 2000). The high density of Lankyblackass spp. may have resulted in the competitive exclusion of many other species from that area (Bond and Slingsby 1984, Christian 2001). Although sampling was performed according to a widely used and accepted method that has shown good results in the past (Parr 2002), the results gained may be an indication that this study was performed at an inappropriate scale. In future either more trapping grids should be laid out in each plot, or the distance between traps in the trapping grid should be greater. This study hints at the importance of adjacent patches with highly variable fire frequency i.e. heterogeneity and patch dynamics, in promoting biodiversity in savanna ecosystems. Heterogeneity has been the focus of much interest of late (Pickett et. al. 2003). The implication for management practices would be that a fine scale patchwork of adjacent areas with highly varied fire frequencies, such as that seen in the experimental burn plots, is required to maintain and promote ant, and perhaps other invertebrate diversity, as ants are known to be good indicators of diversity (Alonso 2000), in savanna ecosystems. An important point to consider, however, is that fine scale patch heterogeneity could be promoting elevated alpha diversity at smaller scales through ecotone principles. There is a turnover in species from patch to patch. A greater number of patches in an area would result in a large turnover in species between these patches, and thus a greater number of species within the defined area. This may, however, result in reduced beta diversity at larger scales, and thus reduced diversity overall (Crawley 1997). In other words, should an area be managed to show at a finer scale all the variation that would usually only be seen at a large scale, then the area would show less variation at a large scale, and thus be relatively homogeneous. This lack of heterogeneity would result in a smaller turnover of species across the area at the larger scale. Future research into the effects of fire frequency on ant diversity should be performed at a larger scale, or should incorporate an island biogeographic or edge effect component that would investigate the effects of the fire frequencies and community composition of adjacent areas on the focal area. The biology of the species under study should also be taken into account. For example aggressive ant species can often exclude other species.

Literature Cited Alonso, L. E. 2000, Ants as indicators of diversity. Pages 80-88 in D. Agosti, J. D. Mayer, L. E. Alonso and T. Schultz, editors. Ants: Standard methods for measuring and monitoring biodiversity. Smithsonian Institution Press. Washington and London. Bond, W. J. and Slingsby, P. W. O. 1984, Collapse of an ant-plant mutualism: the argentine ant (Iridomyrmex humilis) and myrmecochorous Proteaceae. Ecology 65(4): 1031-1037 Bond, W. J. and B. W. van Wilgen. 1996. Fire and the evolutionary ecology of plants, and Surviving fires – vegetation and reproductive response. Pages 123-147 and 34-50 in Fire and plants, population and community biology series 14. Chapman and Hall, London, UK. Christian, C. 2001, Consequences of a biological invasion reveal the importance of mutualism for plant communities. Nature 413: 635-639

91 Clarke, K. R. and Warwick, R. M. 1994. Change in marine communities: an approach to statistical analysis and interpretation. Plymouth Marine Laboratory, Plymouth. Crawley, M. 1997, Structure of plant communities. Pages 475-531 in M.J. Crawley, editor, Plant Ecology. Oxford: Blackwell Science. Kaspari, M. 2000, A primer in ant ecology. Pages 9-24 in Agosti, D., Mayer, J. D., Alonso, L. E. and Schultz, T. editors. Ants: Standard methods for measuring and monitoring biodiversity. Smithsonian Institution Press. Washington and London. Parr, C., Bond, W. J. and Robertson, H. G., 2002. A preliminary study of the effect of fire on ants (Formicidae) in South African savanna. African Entomology 10(1): 101-111. Pickett, S. T. A., Cadenasso, M. L. and Benning, T. A. 2003, Biotic and abiotic variability as key determinants of savanna heterogeneity at multiple spatiotemporal scales. Pages 22-40 in J. T. Du Toit, K. H. Rogers and H. C. Biggs editors. The Kruger Experience. Island Press. Washington, Clovelo and London. Retana, J. and Cerdá, X. 2000. Patterns of diversity and composition of Mediterranean ground ant communities tracking spatial and temporal variability in the thermal environment. Oecologia 123: 436. Scholes, R. J., Bond, W. J. and Eckhardt, H. C. 2003, Vegetation dynamics in the Kruger ecosystem. Pages 242-262 in J. T. Du Toit, K. H. Rogers and H. C. Biggs editors. The Kruger Experience. Island Press. Washington, Clovelo and London. Van Wilgen, B. W., W. S. W. Trollope, H. C. Biggs, A. L. F. Potgieter, and B. H. Brockett. 2003. Pages 150-153 in J. T. Du Toit, K. H. Rogers, and H. C. Biggs, editors, The Kruger Experience. Island Press. Washington, Clovelo and London.

Table 1. Diversity indices for experimental burn plots (NC=Numbi control (has not burned in 50 years), NA= Numbi annual, (burns annually), SC= Shabeni control, SA=Shabeni annual)

Plot S N H' (loge) J' NA 19 469 2.327 0.7902 NC 17 323 1.818 0.6418 SA 15 989 0.6016 0.2222 SC 15 325 1.31 0.4837

Table 2. Pairwise tests of similarity, in species composition and abundance, between experimental burn plots (NC=Numbi control (has not burned in 50 years), NA= Numbi annual, (burns annually), SC= Shabeni control, SA=Shabeni annual)

Plots R-statistic P-value NA, NC 0.44 P<.001 NA, SA 0.372 P<.002 NA, SC 0.575 P<.001 NC, SA 0.517 P<.001 NC, SC 0.034 P<.218 SA, SC 0.73 P<.001

Figure 1. Multi dimensional scaling ordination plot of ant species content, based on Bray Curtis similarity values, in four experimental burn plots: NC=Numbi control (has not burned in 50 years), NA= Numbi annual, (burns annually), SC= Shabeni control, SA=Shabeni annual. The number following plot site represents sample number.

Figure 2. Cluster Analysis of similarity of ant species content, based on Bray Curtis similarity values, in four experimental burn plots: NC=Numbi control (has not burned in 50 years), NA= Numbi annual, (burns annually), SC= Shabeni control, SA=Shabeni annual. The number following plot site represents sample number.

94 Appendix 1. Species NA1 NA2 NA3 NA4 NA5 NA6 NA7 NA8 NA9 NA10 NA11 NA12 baby gold 3 1 16 0 0 0 3 0 0 0 0 1 baby gold 2 0 0 0 0 0 0 0 0 0 0 0 0 ben 0 4 0 0 0 0 5 0 0 0 0 0 ben 2 0 2 5 3 0 2 2 0 0 1 1 0 big black 0 0 0 0 0 0 0 0 0 0 0 0 big lank brown 5 0 1 4 3 3 7 0 0 1 0 3 flamebutt 0 0 0 0 8 0 0 0 0 0 0 2 godzilla 0 0 0 0 0 0 0 0 0 0 0 0 golden jaw 0 1 1 0 0 0 0 0 0 0 0 0 half ass 0 0 0 0 0 17 0 0 0 2 40 20 half ass #2 0 0 0 0 0 0 0 0 0 0 0 0 hammerhead 0 0 0 0 2 0 1 0 0 0 0 0 jesus 0 0 0 1 0 1 0 0 0 0 0 0 lank black ass 1 0 21 2 0 0 11 49 1 0 0 1 lucifer 0 0 0 0 0 0 0 0 0 0 0 0 mean red 0 0 0 1 0 0 0 0 0 0 0 0 med land brown 0 0 0 0 0 0 2 0 0 0 0 0 narrow blac 3 3 0 0 0 0 0 0 0 0 0 1 panch 0 0 0 0 0 0 0 0 0 0 0 0 pee-wee 0 0 0 0 0 0 0 0 0 1 0 0 pseudo argentine 0 0 16 0 3 1 0 0 1 1 0 0 sausage 0 0 1 0 0 0 0 0 0 0 0 0 scott 0 0 0 0 0 0 1 0 0 0 0 0 sharp ass 28 3 0 0 0 7 1 0 4 4 2 6 small red 0 1 2 0 5 0 11 0 11 1 10 5 john major 0 0 0 0 0 0 0 75 0 0 0 0 weiner 0 0 0 0 0 0 0 0 0 0 0 0 dark sharp ass 0 0 0 0 0 0 0 0 0 0 0 0 the boxer 0 0 0 0 0 0 0 0 0 0 0 0

95 Appendix 2. Species NC1 NC2 NC3 NC4 NC5 NC6 NC7 NC8 NC9 NC10 NC11 NC12 baby gold 0 3 0 0 0 0 1 0 0 0 0 0 baby gold 2 0 0 0 0 0 0 0 0 0 0 0 0 ben 0 0 0 0 0 0 0 0 0 1 0 6 ben 2 0 0 0 0 0 0 0 0 0 0 0 0 big black 0 1 0 1 0 0 0 0 1 0 0 0 big lank brown 0 0 0 0 0 0 0 0 0 0 0 2 flamebutt 0 0 0 0 0 0 0 0 0 0 0 0 godzilla 0 1 0 0 0 1 0 0 2 2 1 0 golden jaw 0 0 0 8 0 0 0 0 0 0 0 0 half ass 44 0 4 9 1 6 13 2 24 5 16 8 half ass #2 0 0 0 2 1 2 0 0 0 3 2 0 hammerhead 0 1 0 2 0 2 0 2 2 1 0 0 jesus 0 0 0 0 0 0 0 0 0 4 0 0 lank black ass 0 1 0 0 0 0 0 0 0 0 4 89 lucifer 0 0 0 0 0 0 0 0 0 0 0 0 mean red 0 0 0 2 0 0 0 0 0 0 0 0 med land brown 0 0 0 0 4 0 0 0 0 0 0 0 narrow blac 0 0 0 0 0 0 1 0 0 0 1 0 panch 0 0 0 0 0 0 0 0 0 0 0 0 pee-wee 0 0 0 0 0 0 0 0 0 0 0 0 pseudo argentine 0 1 0 2 1 0 0 0 3 0 0 0 sausage 0 0 0 0 0 0 0 0 0 0 0 0 scott 0 0 0 0 0 0 0 0 3 0 0 0 sharp ass 1 1 0 4 1 0 9 0 3 5 0 0 small red 0 0 0 0 0 0 0 0 0 0 0 0 john major 0 0 0 0 0 0 0 0 0 0 0 0 weiner 0 0 0 0 0 0 0 0 0 0 1 0 dark sharp ass 0 0 0 0 0 0 0 0 0 0 0 0 the boxer 0 0 0 0 0 0 0 0 0 0 0 0

96 Appendix 3. Species SA1 SA2 SA3 SA4 SA5 SA6 SA7 SA8 SA9 SA10 SA11 SA12 baby gold 0 0 0 0 0 0 0 0 9 0 0 2 baby gold 2 0 0 0 0 0 0 0 0 0 0 0 0 ben 0 0 0 0 0 0 0 0 0 0 3 0 ben 2 2 0 0 0 4 0 2 0 0 1 0 0 big black 0 1 0 0 0 0 0 0 0 0 0 0 big lank brown 0 5 0 2 0 0 1 0 7 0 1 0 flamebutt 0 0 0 0 0 0 0 0 0 0 0 0 godzilla 0 0 0 0 0 0 0 0 0 0 0 0 golden jaw 0 0 0 0 0 0 0 0 0 0 0 0 half ass 0 0 3 0 5 2 1 3 0 2 0 4 half ass #2 0 0 1 0 0 0 0 0 0 0 0 0 hammerhead 0 0 0 0 0 0 0 0 0 0 0 0 jesus 0 0 0 0 0 8 0 0 0 0 0 0 lank black ass 64 90 0 122 0 23 97 11 100 23 61 289 lucifer 0 0 0 0 0 0 0 0 0 0 0 0 mean red 0 0 0 0 1 0 0 0 0 0 0 0 med land brown 0 0 0 2 1 1 0 0 0 0 0 0 narrow blac 0 0 1 3 0 0 3 0 1 2 1 1 panch 0 0 0 0 0 0 0 5 0 0 0 0 pee-wee 0 0 0 0 0 0 0 0 0 0 0 0 pseudo argentine 0 0 0 0 0 0 0 0 0 0 0 0 sausage 0 0 0 0 0 0 0 0 0 0 0 0 scott 0 0 0 0 2 0 0 0 0 0 0 0 sharp ass 0 0 1 1 2 4 0 2 2 4 0 0 small red 0 0 0 0 0 0 0 0 0 0 0 0 john major 0 0 0 0 0 0 0 0 0 0 0 0 weiner 0 0 0 0 0 0 0 0 0 0 0 0 dark sharp ass 0 0 0 0 0 0 0 0 0 0 0 0 the boxer 0 0 0 0 0 0 0 0 0 0 0 0

97 Appendix 4. SC1 SC2 SC3 SC4 SC5 SC6 SC7 SC8 SC9 SC10 SC11 SC12 baby gold 0 0 0 0 0 0 0 0 0 0 0 0 baby gold 2 0 0 0 0 0 0 0 0 0 0 0 0 ben 0 0 0 0 3 0 0 2 0 0 0 0 ben 2 0 0 0 0 0 0 0 0 0 0 0 0 big black 0 0 1 0 0 0 0 0 0 0 0 0 big lank brown 0 0 0 0 0 0 0 1 0 0 0 0 flamebutt 0 0 0 0 0 0 0 0 0 0 0 0 godzilla 0 0 0 0 0 0 0 0 0 0 0 0 golden jaw 0 0 0 0 0 0 0 0 0 0 0 0 half ass 3 3 26 15 9 20 13 7 13 7 8 6 half ass #2 1 0 1 0 0 1 1 0 0 0 0 0 hammerhead 0 0 0 0 0 0 0 1 0 1 0 0 jesus 0 0 0 0 1 0 0 1 0 0 2 0 lank black ass 0 0 1 0 0 0 0 0 0 0 0 0 lucifer 0 0 0 0 0 0 1 0 0 0 0 0 mean red 0 0 0 0 0 0 0 0 0 0 0 0 med land brown 0 0 0 0 146 2 0 0 0 0 0 0 narrow blac 1 0 0 0 0 0 0 0 0 0 0 0 panch 0 0 0 0 0 0 0 0 0 0 0 0 pee-wee 0 0 0 0 0 1 0 0 0 0 0 0 pseudo argentine 0 0 0 4 1 0 0 1 0 0 0 5 sausage 0 0 0 0 0 0 0 0 0 0 0 0 scott 0 0 0 0 0 0 0 0 0 0 0 0 sharp ass 0 0 1 0 0 5 1 3 0 0 0 3 small red 0 0 0 0 0 0 0 0 0 0 0 0 john major 0 0 0 0 0 0 0 0 0 0 0 0 weiner 0 0 0 0 0 0 0 0 0 0 0 0 dark sharp ass 0 0 0 0 0 0 0 2 0 0 0 0 the boxer 0 0 0 0 0 0 0 1 0 0 0 0

98 Termites and Fire: The burning question. The effect of different fire regimes on termite activity.

Category: Independent Project Participants: Fahiema Daniels Site: Kambeni and Shabeni Burn Plots. Kruger National Park, Mpumalanga Province. South Africa.

Key words: fire treatments, termite activity, Shabeni and Kambeni experimental burn plots

Abstract: This experiment was carried out in the Shabeni and Kambeni experimental burn plots in the Pretoriuskop region of the Kruger National Park, Mpumalanga Province, South Africa. Sampling was conducted in the annually burned and control plot of each of the sites. Four 50m transects and a 200mx 200m quadrat was set up in each site to get an estimate of termite density and activity. The results show that there was a greater density of termite activity on the annually burned plot than the control plots in both sites ( Shabeni-df=1,P<.01, Kambeni- df=1,P<.01). The conclusion drawn from the results is that termites do better in fire prone areas because fires promote habitat heterogeneity.

Samevatting: Hierdie ekperiment was uitgedra in die Shabeni en Kambeni experimentale verbrande komplotte in die Pretoriuskop gebied van die Kruger Nasionalle Park, Mpumalanga Provinsie, Suid Afrika. Toetsing was uitgedra in die jaarliks verbrand en kontrol komplot van die Shabeni en Kambeni eksperimentalle komplotte. In elke kompolt was vier 50m lyne en ŉ 200m by 200m kwadrant gestuig om ŉ idea van witmier digtheid te kry. Die uitslae duig aan dat die digtheid van witmiere in die jaarlikse verbrande komplotte hoёr is dan die digtheid in die controle komplotte vir die Shabeni en Kambeni komplotte ( Shabeni-df=1,P<.01, Kambeni- df=1,P<.01. Die gevolgtrekking wat van die utslae geskep was, is dat witmiere beter vorder in die jaarliks verbrande komplotte omdat vier woonplaas heterogeneteit bevorder.

Introduction Fire is one of the important drivers in South African savannas as it is a fundamental top down system controller in many savanna systems. Fire consumes an enormous amount of plant biomass in the tropics al one it removes about 2700-6500 tonnes of plant carbon annually. Terrestrial vegetation has been influenced by fire since at least the Mesozoic times and grass has been creating open plains since the mid tertiary. Today the savanna in the tropics continues to erode in to the tropical forests, under the influence of fire ( Bond and Van Wilgen 1996). In South Africa the effect of fire is equally important as fire fire prone communities have taken a different course from the mainstream plant commutnites. Adaptations in plants like resprouting, fire simulated seed set, fire simulated seed release and fire simulated seed germination and fire simulated flowering in the Fynbos are examples of characteristics tha have evolved in response to fire. Grasses in the savanna biome are also fireadapted as different fire frequencies alow various species to dominate ( Bond and Van Wilgen 1996). The role of fire in vegetation structure in the Kruger Park is not as well defined as it is else where and this is mainly because vegetation structure seems to be largely driven by soil characteristcis ( Du Toit, Rogers and Biggs, 2003). Active fire management in the Kruger Park started in the 1950’s with the installation of firebreak networks controlled rotational burning and perimeter burns. In 1992 the management was changed to allow for natural fires only and no fires were lit by management. From 2002 on, fires are managed by a system of integrated fire management (Govender pers.comm) (Figure 1). Experimental burn plots were designed to evaluate vegetation changes in response to different fire treatments. The main aim of the experimental burn plots was to understand fire dynamics better with respect to the effect of seasonality and frequency on vegetation structure. Each burn site has 12 replicated treatments ( Govender pers.comm). Different burn treatments have an effect on the fauna that inhabit these areas as they affect the vegetation these species use as resources. Fire frequency, intensity and seasonality are the most important in determining the effect the fire has on the system (Scholes and Walker 1993). Fire serves as a nutrient recycling agent as it burns wood and leaf litter and returns nutrients back into the system (Davies 1997). The role of termites in a savanna is also to remove dead wood and litter and thus the role of both fire and termites are considered to be complementary.

99 Termites are widely distributed in Northern and South America, Africa, Asia and Australia (Jones 1990) and are the dominant arthropod decomposers in savanna systems (Davies 1997). Termites are important for humification, soil conditioning, fragmentation of organic detritus and nitrogen fixation (Eggelton et al. 1996). Termites can be grouped according to whether they are Epigeal (tree dwelling) or hypogeal (below ground), forage for wood, litter or humus, construct distinct nest or cultivate fungi (Eggelton et al. 1996). Various groups would be expected to respond differently to environmental disturbances as they feed on dissimilar material. The effect of disturbance on termites has been studied in the past. This paper will focus on the effects of fire on termite denisities under two fire regimes. The hypothesis set forth in this paper is: H1: There is more termite activity (both below ground and above ground) on less frequently burnt plots than on frequently burned plots. H0: There is no difference in the termite nests densities on the plots.

Methods The experiment was carried out in the Shabeni and Kambeni burn plots in the Kruger National Park, Mpumalanga Province, South Africa (Figure 1). The Kambeni site was sampled on the 22nd - 23rd February 2004 and Shabeni site was sampled on the 24th of February 2004. The sampling times were between 0700 hrs and 1300 hrs each day. A frequently burned plot (burned annually in August) and a less frequently burned plot ( where fire is excluded as much as possible). In each of the plots a direct and indirect measure was used to determine termite activity . As an indirect method, a series of four transects of 50m each was sampled in each plot and below ground activity was recorded at every meter along the transect. The termite activity for a 1m x 1m plot from the transect was assessed by estimating the percentage of termite activity. A 200m x 200m plot was set out in each plot and the amount of termite mounds in the plot was counted as a direct measure of above ground termite activity. The number of trees with termite activity was also counted in the 200m x 200m plot. A chi squared contingency test was used to determine if there was a difference in the proportion of activity in the annually burned plots relative to the control plots.

Results In the Shabeni burn plots the mean number of below ground termite activity is much higher in the annually burned plot than the control plot (Figure 2). The number of trees with termites and the number of termitaria are also higher in the annually burnt plot than in the control plot (Figure 2). The Kambeni plots show the same pattern. The number of below ground activity, the number of termitaria and trees with termite activity is lower in the Kambeni control plot than in the annually burnt plot (Figure 2). It was also found that the annually burned plots had a higher concentration of termite activity than the control sites (Table 1).

Discussion The results suggest that my initial hypothesis was flawed, as previous work done on termite responses to fire show, like my results, that termites have higher species numbers and biomass in fire prone sites. An example of this is a study from Thailand that showed that fire maintains habitat heterogeneity, thus more species can occupy the same area as there are more niches available (Davies 1997). The higher termite densities in the annually burnt plots are probably due to the increased availability of dead and burnt wood in the annually burned plots. In the Shabeni annual burn plot the vegetation was more open, in terms of canopy cover and it also had a lot more dead wood material than all of the other plots sampled and this may account for the difference in termite activity between the sites ( personal observation). The higher proportion of dead wood is because of the fire intensity and frequency, i.e. the fire regime is effective in burning canopies and keeping trees in the fire trap. The fire trap allows termites to increase in density because they have more dead wood as a food source The open space that is created after a fire increases the probabilty of finding a mate and thus termite densities increase in frequent fire environments ( Davies 1997). Termites are resilient to environmental perturbations (Eggleton et al. 1996 )and thus the effect of fireon termites is not severe and there population numbers are not significantly affected by fire frequency. There were also differences in the degree of termite activity between the two sites. The Kambeni plots ( both annual and control) had a lower termite density than the Shabeni plots. The difference may be due to tree and grass species found in the sites and thus fire intensity differs because

100 of these factors. As fire intensity depends on the grass community species comosition, variations in grass communities could account for differences in the intesity of the burns in these sites. Because the fire intensity is dissimilar in these plots the amount of dead wood varies and so too does the relative abundance of termite acitivty. In conclusion, termites are important in nutrient cycling and many trees grow on top of termite mounds because they provide nutrient rich hotspots. Thus the effect of fire on termite diversity and abundance is important as it plays a role in vegetation structure. I recommend future studies should be done to see if this pattern exists in all of the Kruger Burn Plots. An inventory of species richness for each of the burn plots would also be a good indicator of termite responses to differing fire treatments.

Acknowledgments I would like to thank Laurence Kruger for his help with analysing the results and Bruce Anderson for his help in data collection.

Literature cited Bond, W.J. and Van Wilgen, B.W. 1996. Fire and plants. Chapman and Hall. London. Davies, R.G. 1997. Termite species richness in fire prone and fire protected dry nodipterocarp forest in Du Toit, J.T., Rogers, K.H., and Biggs, H.C. 2003. The Kruger Experience. Island Press. London. Chapter 7. Eggleton, P.E., Bignell, D.E., Sands, W.A., Mawdsley, N.A.., Lawton, J.H., Wood, T.G., and Bignell, N. C. 1996. The diversity, abundance and biomass of termites under differing levels of disturbance in the Mbalmayo Forest reserve, southern Cameroon. Philosophical Transactions of the Royal Society of London. 351, 51-68 Jones, J.A. 1990. Termites, soil fertility and carbon cycling in dry tropical Africa: a hypothesis. Journal of Tropical Ecology. 6, 291-30 Scholes, R.J. and Walker, B.H. 1993. An African Savanna: Synthesis of the Nylsvley study. Cambridge University Press. Cambridge.

101 Table 1. Chi squared test results for the differences between the annually burned plot and the control plots of the different experimental burn plot sites. Site degrees of freedom P value Kambeni 1 P<.05 Shabeni 1 P<.05

Figure 1. A map of the Pretoriuskop area. The Kambeni and Shabeni sites are the sampling sites (Source: Navashni Govender)

102

20 mean density termitaria 15 trees with termite activity

10 Termite presence Termite

0 shabeni annual shabeni control -5 Fire treatment

Figure 2. Differences in termite activity in different fire treatments. The mean number of activity is higher in the Shabeni annually burnt plot than in the Shabeni control plot that is only bunt when a fire goes through the park.(df=203 , P< .001)

8 7

5 mean density 4 termitaria 3 trees with termite 2 activity Termite presence Termite 1

0 kambeni annual kambeni control -1

-2 Fire treatement

Figure 3. Differences in termite activity in different fire treatments. The mean number of termite activity is higher in the Kambeni annually burnt plot than in Kambeni control plot (df=203, P<.05)

103 Appendix 1. Raw data for termite densities in the Shabeni and Kambeni sites shabeni kambeni 51 N 0 N 0 N 0 A 5 shabeni annual annual kambeni 52 N 0 N 0 N 0 N 0 control burn bur control 53 N 0 A 5 A 5 N 0 1 N 0 N 0 N 0 N 0 54 N 0 A 20 N 0 N 0 2 N 0 A 20 N 0 N 0 55 N 0 A 15 A 10 N 0 3 N 0 A 10 A 15 N 0 56 N 0 A 10 N 0 N 0 4 N 0 A 10 N 0 N 0 57 N 0 N 0 N 0 N 0 5 N 0 A 30 A 25 N 0 58 N 0 A 5 N 0 A 5 6 N 0 A 15 A 25 N 0 59 N 0 A 20 N 0 A 10 7 N 0 N 0 A 5 N 0 60 A 12 N 0 A 5 N 0 8 A 5 A 5 N 0 A 10 61 A 5 N 0 N 0 N 0 9 A 5 A 5 A 5 N 0 62 A 5 A 20 N 0 N 0 10 A 0 A 5 N 0 N 0 63 A 5 N 0 A 5 N 0 11 N 0 A 10 N 0 N 0 64 A 5 N 0 N 0 N 0 12 N 0 A 15 A 5 A 10 65 A 5 A 90 N 0 A 5 13 A 5 A 20 A 5 N 0 66 A 5 N 0 A 85 N 0 14 A 30 A 15 A 5 N 0 67 N 0 A 0 N 0 A 15 15 A 0 A 5 N 0 N 0 68 N 0 A 85 N 0 N 0 16 N 0 A 45 N 0 N 0 69 N 0 N 0 N 0 N 0 17 N 0 N 0 A 10 A 5 70 N 0 N 0 A 15 N 0 18 N 0 N 0 N 0 N 0 71 N 0 A 20 N 0 N 0 19 N 0 N 0 N 0 N 0 72 A 15 A 25 N 0 N 0 20 N 0 A 10 A 10 N 0 73 N 0 A 5 A 15 N 0 21 A 10 A 10 N 0 N 0 74 N 0 A 10 N 0 A 50 22 N 0 A 10 A 15 N 0 75 N 0 A 10 N 0 N 0 23 N 0 A 10 N 0 N 0 76 A 5 A 15 N 0 N 0 24 N 0 A 10 N 0 N 0 77 N 0 A 40 N 0 N 0 25 N 0 N 0 A 5 N 0 78 N 0 A 50 A 0 N 0 26 N 0 A 15 N 0 N 0 79 N 0 N 0 N 0 N 0 27 N 0 A 5 N 0 N 0 80 N 0 A 15 N 0 N 0 28 N 0 A 5 A 5 N 0 81 N 0 A 15 A 0 N 0 29 N 0 A 10 N 0 A 25 82 N 0 A 40 A 0 N 0 30 N 0 A 15 N 0 N 0 83 N 0 A 25 N 0 N 0 31 N 0 A 5 N 0 A 25 84 N 0 N 0 N 0 N 0 32 A 5 A 5 N 0 N 0 85 A 5 A 25 N 0 N 0 33 N 0 A 10 N 0 N 0 86 N 0 A 10 N 0 N 0 34 A 5 A 5 N 0 N 0 87 N 0 A 5 A 10 N 0 35 N 0 A 10 N 0 N 0 88 A 5 N 0 A 10 N 0 36 A 5 A 50 A 15 A 15 89 N 0 N 0 N 0 N 0 37 A 5 A 5 A 5 A 15 90 N 0 A 5 N 0 N 0 38 N 0 A 10 A 5 N 0 91 A 5 A 5 N 0 N 0 39 N 0 A 20 A 5 N 0 92 N 0 A 50 N 0 N 0 40 N 0 A 5 N 0 N 0 93 N 0 A 5 N 0 N 0 41 N 0 A 5 N 0 N 0 94 N 0 A 5 A 10 A 20 42 A 5 A 15 A 50 N 0 95 N 0 A 5 N 0 N 0 43 N 0 A 40 A 5 A 5 96 A 5 A 15 N 0 N 0 44 N 0 A 25 A 5 N 0 97 N 0 N 0 N 0 N 0 45 N 0 A 5 A 10 N 0 98 N 0 A 10 N 0 N 0 46 N 0 A 70 A 5 N 0 99 N 0 A 20 N 0 A 10 47 N 0 N 5 A 15 N 0 100 N 0 A 25 N 0 N 0 48 N 0 A 15 N 0 N 0 101 N 0 A 15 N 0 N 0 49 N 0 A 5 N 0 N 0 102 N 0 A 10 A 25 N 0 50 N 0 A 5 N 0 N 0 103 N 0 A 5 A 20 N 0

104 104 N 0 N 0 N 0 N 0 158 N 0 A 5 A 10 N 0 105 A 5 A 5 N 0 N 0 159 A 5 A 5 N 0 N 0 106 N 0 A 15 N 0 N 0 160 N 0 A 15 N 0 N 0 107 N 0 A 50 N 0 N 0 161 N 0 A 25 N 0 N 0 108 A 5 N 0 N 0 N 0 162 A 10 A 5 N 0 N 0 109 N 0 A 25 A 5 A 5 163 N 0 A 5 A 15 N 0 110 N 0 N 0 N 0 N 0 164 N 0 A 45 N 0 A 10 111 N 0 N 0 N 0 N 0 165 N 0 A 85 N 0 N 25 112 N 0 A 15 N 0 N 0 166 N 0 A 25 N 0 A 0 113 A 10 A 15 N 0 N 0 167 A 5 A 30 A 5 N 0 114 A 5 N 0 A 10 A 5 168 N 0 A 40 A 5 N 0 115 A 5 A 30 A 5 N 0 169 N 0 A 85 N 0 N 0 116 N 0 A 5 A 15 N 0 170 N 0 A 50 N 0 N 0 117 A 5 A 5 N 0 N 0 171 N 0 A 50 A 5 N 0 118 N 0 A 25 N 0 N 0 172 N 0 A 25 A 70 N 0 119 A 75 A 5 A 5 N 0 173 A 5 A 25 N 0 N 0 120 N 0 A 10 N 0 N 0 174 N 0 A 15 N 0 N 0 121 N 0 A 10 N 0 N 0 175 N 0 A 5 N 0 N 0 122 N 0 N 0 N 0 N 0 176 N 0 N 0 A 5 N 0 123 N 0 A 5 A 25 N 0 177 N 0 A 50 N 0 A 5 124 N 0 N 0 N 0 N 0 178 N 0 A 5 N 0 N 0 125 N 0 N 0 A 15 N 0 179 A 20 A 5 N 0 N 0 126 N 0 A 10 N 0 N 0 180 N 0 A 5 N 0 N 0 127 N 0 A 10 N 0 N 0 181 N 0 A 5 N 0 N 0 128 N 0 A 5 N 0 N 0 182 N 0 A 10 N 0 N 0 129 N 0 A 5 A 10 N 0 183 N 0 A 15 N 0 N 0 130 N 0 N 0 N 0 N 0 184 N 0 A 50 A 5 N 0 131 N 0 A 5 N 0 A 10 185 N 0 A 40 N 0 N 0 132 A 0 A 5 A 5 N 0 186 N 0 A 30 A 5 N 0 133 A 0 A 10 A 5 N 0 187 A 5 A 20 N 0 N 0 134 N 0 A 5 N 0 N 0 188 N 0 A 5 N 0 N 0 135 N 0 A 5 N 0 N 0 189 N 0 A 35 N 0 N 0 136 N 0 A 25 N 0 A 5 190 A 5 N 25 A 5 N 0 137 N 0 N 0 N 0 N 0 191 N 0 A 5 N 0 N 0 138 N 0 N 0 A 15 N 0 192 N 0 A 15 N 0 N 0 139 N 0 A 25 A 10 N 0 193 N 0 A 5 N 0 N 0 140 N 0 A 40 N 0 A 85 194 N 0 A 10 N 0 A 50 141 A 5 A 40 N 0 N 0 195 N 0 A 10 N 0 N 0 142 N 0 A 45 N 0 A 15 196 N 0 A 25 N 0 N 0 143 N 0 A 40 N 0 N 0 197 N 0 N 0 A 15 N 0 144 N 0 A 25 N 0 N 0 198 N 0 N 0 N 0 N 0 145 N 0 A 10 A 10 N 0 199 N 0 A 25 A 65 N 0 146 A 20 A 50 N 0 N 0 200 N 0 A 50 N 0 N 0 147 N 0 A 50 N 0 N 0 201 N 0 A 90 A 5 N 0 148 N 0 A 10 N 0 N 0 202 N 0 A 50 N 0 N 0 149 A 5 A 10 N 0 N 0 203 A 5 N 0 N 0 N 0 150 N 0 A 50 N 0 N 0 204 N 0 N 0 N 0 N 0 151 N 0 A 5 N 0 N 0 152 N 0 A 5 N 0 N 0 153 N 0 A 5 A 5 N 0 154 N 0 A 10 N 0 N 0 155 N 0 A 10 N 0 N 0 156 N 0 A 0 N 0 N 0 157 N 0 N 0 N 0 N 0

105

Appendix 2. Raw data for number of termite mounds and trees with an indication of termite activity. Kambeni Shabeni Kambeni annual annual Shabeni control burn burn control 200x 200m plots termite mounds 1 6 7 4 trees with termite activity 1 32 3 0

106 Fire and bush nucleation in broadleaf savannas

Category: Independent Project Participants: Carla Staver Site: Skukuza, Kruger National Park, Mpumalanga Province, South Africa

Key words: bush clumping, fire, nucleation

Abstract: The effects of fire on the spatial distributions of trees and shrubs have been largely ignored in investigations on the role of fire in savanna ecosystems. This study examined bush clumping around large trees in experimental burn plots near Pretoriaskop, Kruger National Park in South Africa. Nucleation of established individuals (ht>1m) occurs in the absence of fire. These bush clumps are characterized by a switch to forest and mixed forest species. Annual burn plots and open areas of no burn plots are dominated by bushveld species. This suggests that fire plays a vital role in preventing forest encroachment into savannas and in maintaining the savanna state in the Pretoriaskop area. This area is characterized by higher rainfall (>650mm/annum) which may make the area appropriate for forest encroachment.

Introduction Kruger National Park has a long and diverse history of fire management policy (Van Wilgen et al. 2003). Although fire policy in the park is no longer as rigid and structured as it has been in the past, fire is considered by many of Kruger’s staff to be one of the most powerful management tools available (Van Wilgen 2003). Nonetheless, the ecology of fire in savannas and in Kruger specifically is not well understood; debate continues to rage about the effects of fire on savanna ecosystems and the extent to which fire drives the savanna biome. Two theories dominate the discussion on how and why savannas exist in the form they do. The fire-interval hypothesis suggests that differences in fire frequency are the major drivers of ecosystem level heterogeneity in savannas (Bond and van Wilgen 1996). The density dependence hypothesis suggests that populations are more or less self regulating (ibid.); this implies that savannas are relatively stable and resilient state rather than an intermediate transition phase between grassland and forest. Most work thus far has focused on the effects of fire on population structure and demographics (e.g. recruitment, size class distribution) and species richness. Little work has been done on the spatial distribution of trees and shrubs in savannas. Community level spatial distribution is as much a part of community dynamics; as such, its relationship to fire merits attention. This study will investigate the effects of fire on spatial distribution of trees and shrubs in the savanna landscape and to examine the extent to which fire is essential to the existence of savanna ecosystems. The project will examine the occurrence and dynamics of bush nucleation around large trees. Preliminary observations suggest that trees and shrubs clump around large trees (e.g. Marula (Sclerocarya birrea) or fig (Ficus spp.)) in the absence of fire near Pretoriuskop. The project explores whether bush nucleation occurs around large trees in areas where fire is excluded (control plots) and whether nucleation occurs where fire burns frequently (August annual burn plots). Possible drivers of nucleation fall into two broad categories. Bush clumps may be a naturally occurring component of savannas or they may be indicative of a transition to a forest state. Several possibilities fall under the first category. Large trees are associated with the uplift of water from deep soil to shallow soil, which can sometimes result in greater nutrient availability. In areas of water and nutrient stress, this could greatly alter the micro-environment over a large tree's root system and under its canopy and modify potential for tree growth. Ludwig et al. (2003) documented hydraulic lift for Acacia tortilis in East African savannas. Alternatively, large trees may be associated with elevated concentrations of essential nutrients from leaf litter decomposition or animal droppings (Belsky 1994). Either of these mechanisms could drive bush clump formation in an unaltered natural savanna state. Then again, bush clumps may be indicative of forest encroachment. For instance, birds may be the main drivers of tree and shrub encroachment; if birds use large trees for roosting and feeding, bird- dispersed plants, which frequently have fleshy seeds characteristic of forest species, will occur in higher concentrations around a large tree. The same may be true of mammal or insect dispersed plants. Finally, shade provided by the large tree may also play a role in the formation of clumps; some forest tree species may require protection from the sun for germination and establishment (Belsky 1994). Thus, patterns in the species composition of bush clumps may give insights into the role of fire in determining community structure in savannas and in excluding forest encroachment.

107

Methods Kruger National Park established a series of experimental burn plots in 1954 with a variety of fire treatments, including no burn, annual burn, and two and three year burns in various seasons (Van Wilgen et al. 2003) with a view to facilitating controlled study of the effects of fire on various ecosystem characteristics in savannas. Four burn plot areas cover different soil types (granite and basalt derived) and climates (high and low rainfall). The Pretoriuskop burn plots are characterized by granitic soils and high rainfall (700-750 mm/yr). Within the group are four burn strings, which replicate each of twelve burn treatments. Burn frequency is assumed to be the only variable. Fire intensity and spread rates can also affect vegetation dynamics in burned areas (Scholes & Walker 1993), but they are closely linked to vegetation type, which is constant within a burn string, and burn frequency. I sampled areas within two burn treatments, annual burn and no burn, in two experimental burn plot strings, Shambeni and Kambeni in the Pretoriuskop area (see Table 3). Within each of these plots, I defined two distinct areas: areas under the canopy of trees of DBH ≥30 cm and areas not under the canopy of trees of DBH ≥30 cm. In non-canopy areas I randomly sampled 25-30 plots of radius 2 m. I also randomly sampled 10-15 trees of DBH ≥30 cm and sampled within a 2 m radius. For each tree or plot, I collected the following data: species, height, diameter (DBH), and number of stems for all woody individuals. These data compare tree and shrub distribution in annual burn and no burn plots. A comparison of woody vegetation densities in the sub-canopy and non-canopy of large trees determines the occurrence of bush clumping around the nucleus of a large tree in savannas. Such directed sampling limits the other types of analyses I can carry out on the data collected, but allows me to evaluate this question more easily.

Results The density of established individuals (height ≥ 1m) is greater under the canopy of a large tree than in the open in no burn plots at both Shabeni and Kambeni (Tables 1 and 2, Figures 1 and 3). The density of established individuals does not change under the canopy of a large tree in annual burn plots at both Shabeni and Kambeni (Tables 1 and 2, Figures 1 and 3). Densities of established individuals in the open are similar in burn and no burn plots at both Shabeni and Kambeni, while densities of established individuals under the canopy of a large tree are significantly different in burn and no burn areas (Tables 1 and 2, Figures 1 and 3). Patterns of species richness of established individuals are similar to those of densities of established individuals, possibly due to the close relationship between the two (Spearman’s Rank=0.964668, P<.05, see Figure 7). Species richness of established individuals increases under the canopy in no burn plots but not in annual burn plots (Tables 1 and 2, Figures 2 and 4). Species richness in open areas is similar in no burn and burn plots, but in canopy areas no burn plots have higher species richness than annual burn plots (Tables 1 and 2, Figures 2 and 4). Data for individuals with height less than 1m, categorized as seedlings or gullivers, were collected at only one string. Density and species richness did not change from open to canopy areas of the no burn plot. Areas of annual burn showed a marked increase in density of individuals in the sub- canopy (Tables 1 and 2, Figures 5 and 6). Here again, there is a close relationship between species richness and density (Spearman’s Rank=0.944792, P<.05) (Figure 8). Density and species richness were found to have no relationship to the size of the focus tree (DBH≥30cm) for established individuals in no burn sub-canopy areas. Cross-sectional area of the focus is not related to density (Spearman’s Rank= 0.108232, p>0.05) or species richness (Spearman’s Rank= -0.148732, P>.05) for established individuals. I did not examine seedlings/gullivers because they do not clump around a focus tree (see above). Community analyses based on tree species found in each of the burn plots show that a switch in community structure and species composition does occur from open areas of no burn plots to sub- canopy areas. The MDS (multi-dimensional scaling, Figure 9) for established individuals, i.e. height of more than 1 m, shows that canopy communities in burn sites are closely related and open area communities are closely related. Open areas in no burn sites closely resemble annual burn areas. Annual burn sites showed no significant community composition differences between open and canopy areas. MDS analysis on seedlings/gullivers, i.e. height of less than 1 m, shows that no burn sub-canopy communities are closely related. Annual burn sites in the sub-canopy and in the open are also closely related. However, in this case, annual burn sites are not as distinctly separated from no burn sites. Preliminary analysis suggests that sub-canopy areas of no burn plots have high proportions of forest and mixed forest (forest margin and thicket) species (Figure 11). Annual burn plots are

108 dominated almost exclusively by bushveld type species, with some mixed ecosystem species (bushveld and thicket). Again, open areas of no burn plots represent the intermediate; they are dominated by bushveld species but have higher proportions of forest, mixed forest, and mixed type species than annual burn plots do. Seedlings/gullivers show similar patterns, although forest, mixed forest, and mixed species seem more prevalent in most areas. Sub-canopies of no burn sites are dominated by non-bushveld species.

Discussion Large trees form the core for nucleation of established trees and shrubs in areas of no burn in the Pretoriuskop area, but not in frequently burned areas. This suggests that fire is an important driver of variation in the spatial distribution of trees and shrubs. Moreover, bush nucleation around a large tree does not arise from the prevalence or dominance of a savanna species that thrives when fire is excluded; species richness increases as density increases. The additional species that colonize in the absence of fire represent a switch from an overwhelming dominance of savanna and bushveld species to a significant proportion of forest, forest margin, and forest precursor species. The composition of seedlings/gullivers undergoes a similar switch from bushveld to forest species from open areas to sub-canopy in areas of no burn. This suggests that seeds are dispersed preferentially into the sub-canopy of a large tree or that they germinate and establish more successfully there. However, in the case of seedlings/gullivers, mixed and mixed forest species also start to appear under the canopy of large trees in annual burn plots. However, the frequency of fire prevents forest, mixed forest, and mixed species from recruiting into larger size classes. Bushveld species are nevertheless much more successful even under the canopy in annual burn plots. These species are presumably better equipped to deal with and survive fire than forest species. The fact that small individuals nucleate around large trees in annual burn areas supports the idea that frequent fire creates a demographic bottleneck (Bond and van Wilgen 1996) in which individuals become trapped. Gullivers are knocked back to ground level yearly until they are able to escape (Bond and van Wilgen 1996). In this way, bushveld species and ultimately grasses are given a competitive advantage by the occurrence of fire, to which they are adapted. Thus, at least in the Pretoriuskop area, forest encroachment occurs in areas in which fire is excluded. Perhaps, as the fire-interval hypothesis suggests, fires are the determining factor in savannas, maintaining the balance between grasses and trees in the savanna landscape. However, it is important to keep in mind that the Pretoriuskop is by far the wettest area of the park, receiving between 700-750 mm/yr. Cowling et al. (1997) suggest that 625 mm of rainfall in general or 725 mm of rainfall per year in areas of summer rain is the minimum requirement for forest development. The Pretoriuskop area has the minimum amount of rainfall required for the prevalence of the forest biome. However, the remainder of the park receives much lower annual rainfall 400 mm in the driest northern regions to 650 mm in wetter areas around Skukuza and the wetter south of the park. Nucleation and bush clumping may still occur and probably does, but species switches and community dynamics may be different. The topic provides exciting possibilities for research on the role and importance of fire in savannas.

Literature Cited Belsky, JB. 1994. Influences of Trees on Savanna Productivity: Tests of Shade, Nutrients, and Tree- Grass Competition. Ecology 75: 922-932. Bond, WJ and BW van Wilgen. 1996. Fire and Plants. Population and Community Biology Series 14. London: Chapman and Hall. Cowling, RM, DM Richardson, and SM Pierce. 1997. Vegetation of Southern Africa. Cambridge: Cambridge UP. Pages 258-266 and 278-299. Ludwig, F, TE Dawson, H Kroon, F Berendse, and HHT Prins. 2003. Hydraulic lift in Acacia tortilis trees on an East African savanna. Oecologia 134: 293-300. Schmidt, E, M Lotter, and W McCleland. 2002. Trees and Shrubs of Mpumalanga and Kruger National Park. Johannesburg: Jacana Earth. Scholes, RJ and BH Walker. 1993. An African Savanna: Synthesis of the Nylsvley Study. Cambridge; Cambridge UP. Pages 111-125. Van Wilgen, BW, W Trollope, HC Biggs, A Potgieter, and BH Brockett. 2003. Fire as a Driver of Ecosystem Variability. Pgs 149-170 in Du Toit, JT, KH Rogers, and HC Biggs, (eds.), The Kruger Experience:Ecology and Management of Savanna Heterogeneity. Washington: Island Press.

109

Table 1. Summary statistics. String Burn Treatment Canopy Variable Valid N Mean Std. Dev. Kambeni No Burn Open Density ht<1m 29 0.178 0.192 Species richness ht<1m 29 1.448 1.213 Density ht>1m 29 0.145 0.187 Species richness ht>1m 29 1.345 1.396 Canopy Density ht<1m 11 0.383 0.425 Species richness ht<1m 11 3.455 2.505 Density ht>1m 11 0.875 0.368 Species richness ht>1m 11 6.727 1.794 Annual Burn Open Density ht<1m 25 0.048 0.061 Species richness ht<1m 25 0.480 0.586 Density ht>1m 25 0.146 0.148 Species richness ht>1m 25 0.920 0.702 Canopy Density ht<1m 10 0.382 0.324 Species richness ht<1m 10 2.400 1.506 Density ht>1m 10 0.119 0.101 Species richness ht>1m 10 1.400 1.265 Shabeni No Burn Open Density ht>1m 43 0.074 0.070 Species richness ht>1m 42 0.833 0.730 Canopy Density ht<1m 10 0.135 0.146 Species richness ht<1m 10 1.200 0.919 Density ht>1m 10 0.517 0.291 Species richness ht>1m 10 4.400 2.171 Annual Burn Open Density ht<1m 31 0.144 0.156 Species richness ht<1m 31 1.129 0.922 Density ht>1m 31 0.036 0.064 Species richness ht>1m 31 0.323 0.541 Canopy Density ht<1m 15 0.456 0.240 Species richness ht<1m 15 3.267 1.033 Density ht>1m 15 0.064 0.069 Species richness ht>1m 15 0.733 0.799

110

Table 2. Test statistics showing significant differences between treatments. String Treatment Variable U P Kambeni No Burn v. Annual Burn Open Density ht<1m 194 0.003466 Species richness ht<1m 189 0.002614 Density ht>1m 338 0.670825 Species richness ht>1m 326 0.526611 Canopy Density ht<1m 52 0.832689 Species richness ht<1m 40.5 0.307227 Density ht>1m 0 0.000108 Species richness ht>1m 0 0.000108 Open v. Canopy No Burn Density ht<1m 108 0.118773 Species richness ht<1m 82 0.018901 Density ht>1m 6 0.000003 Species richness ht>1m 2 0.000002 Annual Burn Density ht<1m 29.5 0.000488 Species richness ht<1m 31 0.000598 Density ht>1m 120 0.855132 Species richness ht>1m 101.5 0.390839 Shabeni No Burn v. Annual Burn Open Density ht>1m 432 0.010196 Species richness ht>1m 392.5 0.003916 Canopy Density ht<1m 16 0.001065 Species richness ht<1m 11.5 0.000428 Density ht>1m 8 0.000202 Species richness ht>1m 7 0.000162 Open v. Canopy No Burn Density ht>1m 25 0.000016 Species richness ht>1m 21.5 0.000012 Annual Burn Density ht<1m 55 0.000032 Species richness ht<1m 31 0.000002 Density ht>1m 176 0.185529 Species richness ht>1m 166.5 0.121977

Table 3. Site codes. Site Site Code Shabeni No Burn S7C Kambeni No Burn K1C Shabeni Annual Burn S3AB1 Kambeni Annual Burn K7AB1

111 0.7

0.6

0.5

0.4

0.3

0.2

0.1 Density of Individuals ht>1m (ind per m^2)

0.0

-0.1 Open Canopy

No Burn August Annual Burn

Figure 1. Density of Established Individuals v. Canopy Status, Shabeni . 5.0

4.5

4.0

3.5

3.0

2.5

2.0

1.5

1.0

Species Richness of ind ht>1m (species per plot) 0.5

0.0

-0.5 Open Canopy

No Burn August Annual Burn

Figure 2. Species Richness of Established Individuals v. Canopy Status, Shabeni

112 1.2

1.0

0.8

0.6

0.4

0.2 Density of Individuals ht>1m (ind per m^2) (ind ht>1m of Individuals Density 0.0

-0.2 Open Canopy

No Burn August Annual Burn

Figure 3. Density of Established Individuals v. Canopy Status, Kambeni.

>1m (species per plot) 4

1 Species richness of ind ht

-1 Open Canopy

No Burn August Annual Burn

Figure 4. Species Richness of Established Individuals v. Canopy Status, Kambeni.

113 0.6

0.5

0.4

0.3

0.2

0.1 Density of ind ht<1m (ind per m^2)

0.0

-0.1 Open Canopy No Burn August Annual Burn

Figure 5. Density of Seedlings/Gullivers v. Canopy Status, Kambeni.

5.0

4.5

4.0

3.5

3.0

2.5

2.0

1.5

1.0

Species richness of ind ht<1m (species per plot) 0.5

0.0

-0.5 Open Canopy No Burn August Annual Burn

Figure 6. Species Richness of Seedlings/Gullivers v. Canopy Status, Kambeni.

114 y= 0.2796+6.9608*x 14

ht>1m (species per plot) 6

2 Species Richness of Ind. Species Richness of Ind. 0

-2 -0.2 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0 Density of Individuals ht>1m (ind per m^2)

Figure 7. Species Richness v. Density for Established Individuals, Spearman’s Rank=0.964668, P<.05 y = 0.536+5.2405*x 9

1 Species Richness for Ind. ht<1m (species per plot)

-1 -0.2 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 Density of Ind. ht<1m (individuals per m^2) Figure 8. Species Richness v. Density for Seedlings/Gullivers, Spearman’s Rank=0.944792, P<.05

115

Figure 9. Multi-dimensional scaling community analysis based on tree species found in each plot (individuals ht≥1m).

Figure 10. MDS community analysis based on tree species in each plot (individuals ht≤1m).

116 1.2

0.8 bushveld mixed 0.6 forest mixed forest 0.4 Proportion of Individuals of Proportion

0.2

0 No Burn No Burn No Burn No Burn Annual Annual Annual Annual Open (Sh) Open (Ka) Canopy Canopy Open (Sh) Open (Ka) Canopy Canopy (Sh) (Ka) (Sh) (Ka) Plot

Figure 11. Proportion of Individuals (height >1m) v. Plot. The graph depicts the proportion of individuals in each plot of bushveld, mixed bushveld, mixed forest, and forest type species.

1.2

0.8 bushveld mixed 0.6 forest mixed forest 0.4 Proportion of Individuals of Proportion

0.2

0 No burn No burn No burn Annual burn Annual burn Annual burn Annual burn Open (Ka) Canopy (Sh) Canopy Open (Sh) Open (Ka) Canopy (Sh) Canopy (Ka) (Ka) Plot

Figure 12. Proportion of Individuals (height <1m) v. Plot. The graph depicts the proportion of individuals in each plot of bushveld, mixed bushveld, mixed forest, and forest type species.

117 Biological nitrogen fixation nodules in legumes in three different sites in Kruger National Park

Category: Independent Project Participants: Sally Koerner, Taryn Morris, and Justine Norman (secretaries), Laura Buckley and Kinesh Chetty (field assistants) Site: Skukuza, Kruger National Park, Mpumalanga Province, South Africa

Key words: fine-leaf, broad-leaf, fire, legumes, nitrogen fixation, nodules

Abstract: Broad-leaf savannas are more nutrient limited than fine-leaf savannas. Plant productivity is constrained by Nitrogen availability thus some legume plants have developed adaptations to obtain atmospheric nitrogen in a usable form by forming a symbioses with nitrogen fixing bacteria that form gall like nodules on the roots. Our study aims to determine if herbaceous leguminous plants in the nutrient limited broad leaf site have a higher occurrence of nitrogen fixing nodules in comparison to the fine leaf site. Additionally as fire causes a loss of Nitrogen to occur in the system we aimed to see if the burnt broad leaf site has a higher occurrence of nodules in the burn plot than the unburnt plot. We found the broad-leaf and burnt broad leaf sites have a similar proportion of nodulated plants (P<.834188) with fine-leaf site comparatively lower (P<.001877) in comparison to the broad-leaf site.

Second Language Abstract: Breed-blaar grasvlakte is meer voedend beperk as fyn-blaar grasvlakte. Stikstof biskikbaarheid beperk die produktiwiteit van plante en toe n’ paar peuldraend plante het aanwendings ontwikkel om lugstoornis stikstof te kry wat kan gebruik wees. Bakteria wat stikstof vasmaak vorm knoesies wat soos galnote lyk op die wortels van plante. Die doel van ons studie is om te sien as die voedend beperk breed-blaar plek het n’ hoer voorval van die knossies wat stikstof vasmaak as die fyn-blaar grasflakte lokasie. As vuur kan n’ verlooring van stikstof in n’ sisteem veroorsaak ons will ook probeer sien as daar a hoer voorval van knossies in die gebrande breed-blaar lokaasie sal wees. Ons uitslae het gewys dat die breed-blaar loakasie wat nie gebrand is nie het n’ gelyksoortig eweredigheid van knossies as die gebrand lokasie (P<.834188). Ons het ook gevind dat in vergelyking die fyn-blaar lokasie n’ laaer eweredigheid van knossies gehad het.

Introduction In a Southern African savanna there are two predominant woody vegetation communities: broad leaf and fine leaf. Broad leaf areas are situated on sandy/granitic soils that are nutrient poor while fine leaf sites are situated on clayey / basaltic soils that are comparatively nutrient rich. There are several factors such as soil moisture, fertility, and fire that help to maintain this dynamic interaction between woody plants and the continuous annual / perennial herbaceous layer (Scholes et al. 1999 in Woghiren 2002). Future changes in these three factors, such as climate and management practices, could have far-reaching effects on the ecosystem (Woghiren 2002). These same changes could affect the long-term sustainability and biogeochemistry of carbon and nutrients such as nitrogen (Woghiren 2002). There are certain micronutrients and macronutrients that are essential for plant growth and functioning. Nitrogen (N) and phosphorous (P) are both considered as macronutrients. These nutrients are required in large amounts by plants in order to photosynthesize, grow, and reproduce. The amount of nitrogen available in the soil frequently limits the ability of the plant to carry out these functions (Gurevitch et al. 2002) and hence a potential constraint on productivity in plants (Scholes and Walker 1993). Our study focuses on nitrogen limitation and on biological nitrogen fixation by certain legume plants in savannas. The largest nitrogen pool in savannas is organic nitrogen in the soil; this is spatially variable in both width and depth (Scholes and Walker 1993). Nitrogen fixation is the process by which atmospheric nitrogen is converted to ammonium, and this is performed mainly by nitrogen fixing bacteria of Rhyzobium type located in gall like nodules on the roots of legume plants (Scholes and Walker 1993). As mentioned broad leaf sites are more nutrient limited than fine leaf; therefore, one would expect plants found here to have adapted to this limitation by forming this relationship more readily than plants in the fine leaf sites. Also nitrogen fixation uses a great deal of energy; it requires

118 more energy for a plant to support nitrogen fixing bacteria compared to obtaining nitrogen from the soil, when it is readily available(Gurevitch et al. 2002). Plants that are capable of supporting nitrogen fixing bacteria usually cease to do so when soil nitrogen is abundant (Gurevitch et al. 2002). We therefore hypothesise that that there will be a higher occurrence of nitrogen nodules in the broad leaf site compared to the fine leaf site. Pyrodenitrification is a loss of nitrogen in a system due to burning, thus in regularly burnt systems, it would be expected that less nitrogen will be available. Most of the nitrogen released into the atmosphere comes directly from the vegetation, but substantial amounts can come from the burning of accumulated litter, and very hot fires can burn the soil organic mater in the uppermost part of the profile, thus taking away nitrogen from the soil (Gurevitch et al. 2002). Trapnell et al. (1976) (as cited in Scholes and Walker 1993) showed a 0-8% decrease in nitrogen in the top 15 cm of Miombo woodland in Zambia when annually burned. Our second hypothesis, therefore, is that due to less nitrogen being available in frequently burnt plots, there will be a higher occurrence of nodules on plants in order to facilitate uptake of nitrogen in a depleted system.

Methods Site selection and sampling Our study was conducted in the Kruger National Park, Mpumalanga Province, South Africa during summer. Nitrogen fixing activity has been found to be associated exclusively with nodulated legumes. It has however, been shown that the subfamily Caesalpinoidae do not contain members that fix nitrogen, Mimosoidae have few members that fix nitrogen and the majority of Papillionacae members do (Scholes pers. comm.). In a study conducted by Grobbler and Rosch (1981) in Nylsvley Nature Reserve (a typical South African savanna) it was found that N fixation was definitely predominant in certain genera of this subfamily namely: Elephantorrhiza, Tephrosia and Indigofera sp. (Scholes and Walker 1993). Thus we decided to concentrate on plants from these genera. Two transects were sampled in a fine-leaf and broad-leaf site in the Skukuza region as well the Shabeni February biannual burn site near Pretoriaskop. The biannual burn site was chosen as it approximately resembles the burning regime of uncontrolled areas. Each transect was 100m in length and 5m wide. Herbaceous Papillionaceae plants were excavated within the transect and presence and activity of nodules were noted. Species were identified into morpho-species due to lack of inflorescences making identification difficult. Analysis Data were analysed in Statistica using Chi-squared tests to determine if there was significant differences in proportional nodulation between the fine-leaf and broad-leaf sites as well as between the burnt broad-leaf site and non-burnt broad-leaf site. Chi-squared tests were also performed on two morpho-species from the broad-leaf and burn sites. Comparing morpho-species in different sites allows us to eliminate fluctuations in proportional nodulation that may be linked to the ability of certain species to form nodules more readily than others.

Results The broad-leaf site had the highest density of herbaceous legume plants, followed by burnt broad leaf with the fine leaf site having the lowest total density (Figure 1). Broad leaf and burnt broad leaf sites had a similar proportion of nodulated plants with fine-leaf site comparatively lower (Figure 1). Chi-squared tests indicate that the broad-leaf has a higher proportion of nodulated plants than the fine-leaf site (P<.001877). Proportions of nodulated plants in the broad-leaf and burnt sites are similar (P<.838488) (Table 1). There is no significant difference between the proportion of nodulated “flat pod” morphospecies in the burnt broad-leaf and non-burnt broad-leaf sites (P < 0.903807). Similarly, the proportions of nodulated “fine-leaf” morpho-species plants in the burnt broad-leaf and non-burnt broad leaf sites do not differ greatly (P<.834188) (Table 2).

Discussion Our results supported our first hypothesis which stated that there would be a higher proportion of nitrogen nodules in the broad-leaf site compared to the fine leaf-site. The proportions of nodulation in these two sites are significantly different (P<.05). Both morpho-species showed no difference in proportional nodulation between the burnt and broad leaf site, supporting that there is minimal difference in nodulation between the burnt and broad-leaf site. The flux of nitrogen through a fine-leaved savanna system is substantially higher than through a broad-leaved savanna (Scholes and Walker 1993) due to the higher mean of N and P content in the

119 fine-leaved litter fall (Woghiren, 2002). Plants in a fine-leaved savanna have less of a need to fix nitrogen, as it is more easily available to them; thus, they have a lower occurrence of nodules. Our results don’t however support our second hypothesis that there will be a higher occurrence of nodules in the burnt broad-leaf site compared to the broad-leaf site as we found that there is no significant difference (P <.84) between the two sights and therefore many factors may have affected these findings. According to Woghiren (2002), savanna structure and productivity is strongly correlated with plant water availability. Thus one may conclude from this that as the burnt site was situated in a higher rainfall region (719mm per annum) than that of the broad-leaf site (537 mm per annum) it would be wrong to make a direct comparison between nodule proportions. When considering the status of nodules, we found that the percentage of nodules that were inactive was extremely small, and so the presence of nodules was analyzed disregarding whether they were inactive or active. Sampling was restricted due to time constraints, but the study would be more conclusive if other factors that affect nitrogen content or soil fertility could be investigated in future studies. Phosphorous for example, if limited in soils will directly affect and further limit the uptake of nitrogen (Gurevitch 2002). If we take this into consideration when examining our results between the burnt and broad leaf sites, it may give insight into why the burnt site did not actually have a higher proportion of nodulation. Other interesting avenues could be explored such as size classes in plants and whether size / age of a plant will affect the amount of nodules found on an individual. The degree of nodulation would also be important to determine the extent of fixation taking place and, therefore, may also be useful in future studies. In the broader context, our findings contribute to current information pertaining to biological nitrogen fixation in savannas especially when taking into account that most of the previous studies conducted were done on Acacia species. Understanding of nitrogen fixing in any context also helps further our understanding of global change on an international level. As Woghiren (2002) points out, 40% of the African continent consists of arid and semi-arid savannas, and vegetation changes within this biome may become more significant to regional carbon and nitrogen cycles, particularly with regard to its role as a sink. Adding to this in formation in this context information such as this will enable prediction of how these ecosystems will respond to the complex combinations of future climate and land use (Woghiren, 2002).

Acknowledgements: We would like to thank all those who helped support team nodule. Laura Buckley for her superb recording, and Kinesh Chetty for his motivation and digging. We would also like to thank Julie Coetzee and Laurence Kruger for their help with the fieldwork and Statistica.

Literature Cited Grobbler, N., and M. W. Rosch. 1981. Biological nitrogen fixation in Northern Transvaal savanna. South African Journal of Botany. Gurevitch, Jessica, S. M. Scheiner, and G. A. Fox. 2002. The Ecology of Plants. Sinauer Associates, Inc. Scholes, R.J. and B. H. Walker. 1993. An African Savanna Synthesis of the Nylsvley Study. Cambridge University Press, Cambrdge. Woghiren, A. J. 2002. Nitrogen characterization of the savanna flux site at Skukuza, Kruger National Park. MSc Thesis. University of Witwatersrand, Johannesburg.

Table 1: Chi-squared tests of proportion nodulating plants between sites Plots Chi-squared value P-value FL, BL 9.666083 P<.001877 BBL, BL 0.0415459 P<.838488

120

Table 2: Chi-squared tests of morpho-species between the burnt and non-burnt sites Plots Chi-squared value P-value BBL, BL 0.0146056 P<.001877 BBL, BL 0.0438208 P<.838488

26 0.30 24 22 0.25 )

2 20 0.20 18 16 0.15 14 12 0.10 10 8 0.05 Proportion nodulated plants nodulated Proportion

Total Density of plants(/100m of Density Total 6

4 0.00 2 0 -0.05 density(L) Fine Leaf Burnt Broad Leaf prop(R) Broad Leaf

Figure 1. Total density of herbaceous leguminous plants and proportion of nodulated plants in each site.

121

Appendix 1. Fine leaf raw data - transect one and two

Transect 1 Transect 2 Distance Species Yes / No Active? Distance Species Yes / No Active? 27.2 short 0 1 flat 0 27.3 curly 0 2 curly 1 1 28.4 curly 0 15 short 0 34.3 curly 0 17.3 flat 0 90 curly 0 17.3 flat 0 97.5 bunched 0 25.3 curly 0 97.5 bunched 0 38.5 curly 0 Total 0 60 short 0 60.8 bunched 0 60.9 bunched 0 61.3 short 0 63.7 bunched 0 63.8 bunched 0 65.2 curly 0 67.4 short 1 1 77 curly 0 81.6 curly 1 1 85 curly Total 3

122

Appendix 2. Broad leaf burnt site raw data - transects one and two

Transect 1 Transect 2 Distance Species Yes / No Active? Distance Species Yes / No Active? 5.8 short 0 15.2 short 0 6.7 flat 1 1 16 short 0 7.4 short 0 20 short 1 0 8.4 flat 0 20 short 0 9.7 fine 0 20 short 0 13.3 thin 0 20 short 0 14.8 flat 1 1 20 flat 1 1 14.8 fine 1 1 21 short 1 1 17.4 fine 0 21 short 0 18.1 fine 0 21 short 0 18.4 short 1 1 21.6 short 1 1 18.5 fine 0 23 thin 0 18.5 fine 0 23.8 short 1 0 18.8 fine 0 23.8 short 0 19.2 fine 0 25 short 0 Transect 1 - continued Transect 2 - continued Distance Species Yes / No Active? Distance Species Yes / No Active? 19.5 fine 0 43.3 short 0 21.2 short 0 45.9 thin 0 23.3 flat 0 45.9 short 0 25.2 flat 1 1 46 short 0 28.4 short 1 1 46 short 1 1 29.2 thin 0 47.6 short 1 1 29.6 flat 0 47.6 flat 1 0 30.4 thin 0 47.6 flat 0 31.2 thin 0 47.6 flat 0 32.1 flat 1 1 47.6 flat 0 32.3 thin 0 50 short 0 33.2 short 0 50 short 0 33.3 fine 0 53.3 flat 0 33.4 short 1 1 53.7 flat 0 33.6 flat 1 1 56 flat 0 33.6 flat 0 56 flat 0 33.8 flat 0 57.5 flat 0 34 short 0 60 flat 1 1 36 short 0 94.3 flat 1 1 36.5 flat 0 94.3 flat 1 1 37 flat 1 1 94.3 flat 0 38 thin 0 95.4 flat 0 39 flat 1 1 Total 11 8 42 fine 0 42.5 fine 1 1 45 thin 1 1 45.6 fine 0 47.2 flat 1 1

123 48 short 0 49 thin 0 56.5 short 1 1 57.4 flat 1 1 58.2 thin 0 58.4 thin 0 58.4 thin 0 59.6 flat 1 1 59.8 short 0 62.5 thin 0 62.5 short 1 1 62.5 short 0 63.2 short 0 64.6 short 0 66.3 flat 0 66.3 flat 0 77 flat 0 77.6 short 0 79 thin 0 83.9 flat 1 0 84.3 flat 0 84.5 flat 0 84.5 flat 1 0 84.5 flat 0 84.5 flat 0 85.5 flat 1 1 87.3 flat 0 88.3 flat 0 90.2 fine 0 90.2 flat 0 90.2 flat 0 92.8 flat 1 0 92.9 flat 0 92.9 flat 1 0 93.4 flat 0 93.4 flat 0 93.4 flat 0 93.4 flat 0 96 flat 0 96 flat 0 99 flat 0 Total 24 20

124

Appendix 3. Broad leaf site raw data - transects one and two Transect 1 Transect 2 Distance Species Yes / No Active? Distance Species Yes / No Active? 1.4 flat 0 1 flat 3.3 flat 0 1.1 flat 3.3 flat 1 1 1.6 flat 3.7 flat 1 1 1.6 fine 5.5 flat 1 1 2.3 flat 5.5 fine 0 4.4 fine 5.7 flat 1 1 8.6 fine 5.7 fine 0 9.7 fine 1 1 6.2 flat 0 9.7 fine 6.3 fine 1 1 10 fine 1 1 6.8 fine 0 10 fine 7.2 fine 0 10.1 fine 8.6 flat 1 1 11.4 fine 12.5 fine 0 12.3 fine Transect 1- continued Transect 2 – continued Distance Species Yes / No Active? Distance Species Yes / No Active? 12.5 fine 0 15.4 flat 1 1 12.5 fine 0 16 fine 12.5 fine 0 16.3 fine 13.2 flat 0 20.7 flat 13.5 flat 0 21 fine 14.8 fine 1 1 21.3 flat 15.3 fine 0 21.4 flat 15.8 fine 0 33.4 flat 15.8 fine 0 33.6 long 16 fine 0 34.9 flat 16 fine 0 34.9 fine 16 fine 0 35.8 fine 16 fine 0 36 fine 16.8 fine 0 36.4 fine 18 fine 1 1 36.7 flat 18 fine 1 1 37 flat 1 1 18 fine 1 1 39 fine 18 fine 1 1 40 round 1 1 18.5 fine 0 40 flat 19 fine 1 1 40 flat 20 fine 1 1 41 fine 30.2 fine 1 1 42.3 fine 30.7 fine 1 1 42.3 flat 33 fine 0 42.8 flat 33 fine 1 0 43.4 flat 33.5 fine 1 1 43.4 flat 33.5 fine 0 45 fine 34 fine 0 45.6 flat 35.6 fine 0 48.9 fine 41.7 flat 0 48.9 fine

125 43.3 flat 0 50.4 fine 44.4 flat 0 50.6 fine 44.5 flat 0 50.7 fine 44.9 flat 0 50.7 fine 47 fine 0 51.3 fine 47 fine 0 51.3 fine 47.7 flat 0 52.6 fine 47.7 flat 0 55 fine 47.8 fine 0 55.5 fine 1 1 54.4 flat 0 56 flat 1 1 58.5 fine 0 57 flat 1 1 58.5 flat 0 57.7 fine 58.5 round 0 59 fine 1 1 58.5 round 1 1 62.3 fine 1 1 Transect 1- continued Transect 2 – continued Distance Species Yes / No Active? Distance Species Yes / No Active? 61.2 flat 0 63 fine 1 1 61.3 round 0 63.9 fine 1 1 61.3 round 0 64.4 fine 1 1 61.5 fine 0 64.6 flat 62.1 flat 0 65.5 fine 62.5 round 0 67.8 fine 63.3 round 67.8 fine 1 64 fine 68 fine 64.5 fine 68 fine 1 1 67.9 fine 68 fine 70 fine 68.6 fine 71.2 fine 69.6 fine 71.9 fine 69.6 flat 1 1 72.2 fine 69.6 flat 1 1 72.3 fine 70 fine 1 1 73 fine 1 1 70 flat 73 fine 72.2 fine 78 flat 73.5 fine 1 1 80.5 flat 75.8 fine 80.7 flat 1 1 76 fine 1 1 82.6 fine 1 1 76.1 fine 84 flat 1 1 76.7 fine 1 1 87.6 flat 77.4 fine 87.9 flat 1 1 79 fine 88.3 fine 80 flat 88.3 fine 81 fine 89.4 flat 81 fine 91.1 fine 81 fine 91.7 flat 82.5 fine 99 fine 82.9 fine Total 23 22 83.4 fine 83.7 fine 1 1 83.7 short 1 1 84 fine

126 85 flat 1 1 85 fine 1 1 85.4 fine 85.8 fine 86.5 fine 86.5 fine 87.7 fine 1 1 87.7 fine 1 1 87.7 fine 88.3 fine Transect 2 - continued Distance Species Yes / NoActive? 89.3 flat 1 89.5 fine 89.5 fine 90 flat 91.4 flat 1 1 92.3 flat 93.4 fine 93.8 flat 99 flat 99 flat Total 29 28

127 Ligno-tubers, obligate or facultative?

Category: Independent Project Participants: Benjamin Wigley Site: Skukuza, Kruger National Park, Mpumalanga Province, South Africa

Key words: broad-leafed savanna species, facultative, lignotubers, obligate, resprouting

Abstract: To overcome the challenge of frequently occurring fires in savannas, some woody species have evolved the ability to resprout from underground lignotubers. This study set out to determine if the formation of these tubers is an obligate or facultative adaptation to disturbance. This was done by comparing root to shoot ratios of three savanna broad-leafed species, Sclerocarya birrea, Terminalia sericea and Euclea natalensis growing in annually burnt, no burn and no disturbance (control) treatments. Significant differences were found between the root/shoot ratios of E. natalensis annual burn and no burn treatments (P=.028) and between S. birrea annual burn and control treatments (P=.028). No significant difference was found between root/shoot ratios for T. sericea from annual and no burn treatments. The outcomes of this study suggest that S. birrea and E. natalensis are facultative tuber forming species and T. sericea is an obligate tuber forming species. This suggests that some savanna species are influenced by disturbance in terms of tuber formation and invest in carbohydrate storage accordingly.

Second language abstract: Houtagtige spesies in savannas het deur evolusie die vermoë ontwikkel om vanuit ondergrondse storingsorgane te herstel na vuur. Die doel van hierdie studie was om vas te stel of die vorming van storingsorgane in reaksie op versteuring gedoen word en of dit voorkom in alle individue van ‘n spesie onafhanklik van die versteurings waaraan dit onderwerp is. Dit is gedoen deur die wortel tot groeipunt verhoudings van drie savanna spesies, Sclerocarya birrea, Terminalia sericea en Euclea natalensis te vergelyk. Individue wat in areas van drie verskillende vuur frekwensies groei is vergelyk. Die drie verskillende vuurfrekwensies wat ondersoek is, is jaarlikse vuur, geen vuur en kontrole (geen versteuring). Betekenisvolle verskille is gevind in die wortel tot groeipunt verhoudings van E. natalensis wat tussen jaarlikse brand en geen vuur areas vergelyk is (P=.028). Sclerocarya birrea in jaarlikse brand areas het betekenisvol verskil van dieselfde spesie in die kontrole studie area (P=.028). Geen betekinsvolle verskille is gevind in T. sericea nie. Die studie verskaf bewyse vir die moontlikheid dat S. birrea en E. natalensis slegs ondergrondse storinsorgane ontwikkel wanneer benodig bv. wanner dit in areas wat gereeld versteur word deur vuur groei. In teenstelling hiermee word dit gestel word dat T. sericea storingsorgane ontwikkel onder enige groei versteurings omstandighede. Sekere savanna species reageer dus op versterings op verskillende maniere deur bv. ondergrondse storingsorgane vir voedingstowwe te ontwikkel.

Introduction Fire is a common occurrence in savanna ecosystems and plays an important role in ensuring that trees and grasses co-exist, with the exclusion of fire the grass/tree balance can be altered in favour of trees (Bond and van Wilgen 1996). Most savanna woody species have overcome the frequent disturbance generated by fire by evolving the ability to resprout which allows them to survive and escape the fire trap (Bond and van Wilgen 1996). When a fire passes through a savanna most of the small trees (gullivers) will lose their aboveground parts unless they have grown tall enough to escape the fire. Thus with frequent fires these gullivers are continually losing their above ground biomass and rely on energy reserves stored in underground tubers to send out more stems (Bond and van Wilgen 1996).

128 Hansen et al. (1991) as cited in Bond and Midgley (2003) found sprouters to allocate more reserves to underground structures and to have reduced aboveground growth rates as a result. After a fire sprouters, send out a number of stems (coppices) which grow rapidly and are very leafy to maximise photosynthesis. It is thought that these coppices are used primarily to replenish the starch reserves in the tuber (Bond and van Wilgen 1996). The plant reaches a point at which it sends out one pole-like stem which it uses to try escape the fire trap (Bond and van Wilgen 1996). The point at which it does this is not well understood as the plant faces a trade-off because if it doesn’t make it out of the fire trap before the next fire it needs to have some reserves left to send out new coppices and try again (Bond personal communication 2004). In savannas most woody angiosperms are known to resprout as seedlings and saplings and all should face similar trade-offs between allocation to above ground growth versus belowground storage at the seedling stage (Bond and Midgley 2003). There is an extensive literature on the ecology of sprouting in woody plants (see Bellingham 2000, Bond and van Wilgen 1996, Bond and Midgley 2001, Del Tredici 2001, Bond and Midgley 2003). There is however very little literature on the underground biology and physiology of gullivers growing in savannas. A number of studies have been done on the lignotubers of Australian Eucalypt species (Blake 1972, Whitcock et al. 2003), these studies focused on the types of resprouting and physiology of the ligno-tubers. Jahnke et al. (1983) as cited in Bellingham (2000) found that genetically isolated populations of Eucalylyptus camaldulensis may lose their capacity to resprout in low frequency disturbance regimes; however this is probably as a result of thousands of years of isolation. Another study by Schwilk (2002) as cited in Bond and Midgley (2003) found no growth differences in a comparison of intraspecific variants of sprouting and non-sprouting forms of Ceanothus tomentosus. Bond and Midgley (2003) have suggested that sprouting carries a considerable establishment cost for sprouter seedlings since they have to accumulate belowground reserves at the expense of shoot and root growth. Thus it would be more efficient for a plant not to invest large quantities of starch in tubers if the plant is growing in conditions where they are unlikely to need these reserves. It is quite clear that the formation of tubers by woody species in savannas is an important adaptation which ensures the survival of woody plants through frequent disturbance. However previous studies of savanna ecosystems have failed to address whether these tubers are obligative of facultative structures. The fire plot experiments that exist in the Kruger National Park provide the ideal opportunity to test the effects of frequent versus no disturbance on tuber development. A number of experimental burn plots were set up in the Kruger National Park in 1954. They consist of a series of plots (approx. 6 ha each) all exposed to different burning regimes since they were set up. I plan to test whether all small trees or gullivers automatically form tubers regardless of their extant disturbance regimes, suggesting an obligatory need for tubers. However if plants growing in low disturbance areas have small or no tubers this would suggest that the tubers are facultative structures. I aim to test the hypothesis that small trees (gullivers) growing in frequently burnt areas will have large underground tubers as they depend on them for survival and need large quantities of carbohydrate reserves in order to escape the fire trap, suggesting that lignotubers are facultative structures. Alternatively small trees growing in areas where fire has been excluded since germination should have very small or no tubers as they have no need for reserves to survive fire, this would also suggest that lignotubers are facultative structures.

Methods Study site The data were collected at the Kambeni fire plots in the Pretoriaskop region of the Kruger National Park. The fire plots are part of an ongoing experiment in the park so plants were not allowed to be removed from the plots. However the fire breaks surrounding the annual burn plots have also been burnt annually since 1954 and could therefore be treated as an annual burn area. The Kambeni site also contains a quarry area which has effectively excluded fire since its creation and could therefore be treated as a low disturbance area for this study. Sampling Three common broad-leafed savanna trees species (Sclerocarya birrea, Terminalia sericea and Euclea natalensis) occurring in and around the Kambeni experimental fire plots in KNP were chosen for this study. In each of these areas five plants of similar size for each of the above species were located and dug up ensuring that all root material was recovered. Five S. birrea and E. natalensis seedlings were bought from the nursery at Skukuza, no T. sericea seedlings were available from the nursery. The nursery plants were used as the control treatment as they had never been exposed to disturbance in the form of fire or browsing. The plants were then separated into root and shoot material

129 which was dried in an oven for five days at 70 ˚C after which the dry weight was measured. The root/shoot ratios were then calculated by dividing the dry weight of the root material by the dry weight of the shoot material. Analysis Nonparametric statistical analyses were performed to test for significant differences between the root/shoot ratios for each treatment. The Man-Whitney U-test was used to test for differences between treatments using Statistica 6.0.

Results The mean root/shoot ratios were calculated and plotted for each species and each treatment. Figure 1 shows that E. natalensis had the highest root/shoot ratio in the control treatment (1.6) with a lower ratio in the annual treatment (1.1) and the lowest ratio in the no burn treatment (0.6). The standard error was also the highest in the control (0.4) with similar values in the other two treatments (0.2). Sclerocarya birrea was found to have the highest root/shoot ratio in the annual burn treatment (2) with a lower value in the no burn treatment (1.4) and much lower ratio in the control treatment (0.4). The standard error values were similar in the no burn and annual burn treatments (0.5) with a much lower value in the control (<0.1). Terminalia sericea had the highest root/shoot ratios with the annual treatment being higher than the no burn treatment (3.3 and 2.5 respectively). The annual burn also showed the highest standard error (1.2) while the no burn had a much smaller standard error (0.5). The results of the Man-Whitney U-tests are shown in Table 1. Significant differences in root/shoot ratios were found for E. natalensis annual treatment vs. no burn treatment and no burn vs. control both with following statistics (U=2, P=.028). There was also a significant difference in root/shoot ratios for S. birrea annual vs. control (U=0, P=.009). All other comparisons failed to yield any significant differences in root/shoot ratios.

Discussion The results show a gradient of different responses to the three treatments by the three species. For E. natalensis I found significant differences in the root/shoot ratios between the annual burn and no burn treatments and between no burn and control treatments (Table1). The significant difference between the no burn and annual burn treatments suggest that this species is phenotypically plastic and has the ability to control the formation of tubers. Thus the formation of tubers in E. natalensis can be thought of as a facultative adaptation. The significant difference between tuber size of plants in the no burn and control treatment is somewhat surprising. However this difference is most likely as a result of the control plants from the nursery being root bound. All the plants from the nursery appear to have remained in their bags for an extensive period and therefore had extremely extensive and matted root systems. Sclerocarya birrea showed a highly significant difference in root/shoot ratios between the annual burn and control treatments. From Figure 1, a distinct gradient can be seen in the root/shoot ratios across the treatments with a decrease in these ratios with decreasing disturbance. Note should be taken of the low standard error exhibited by the nursery plants showing all plants to have very similar root to shoot ratios when disturbance is excluded. This result strongly suggests that the formation of tubers in S. birrea is a facultative adaptation used to overcome disturbance. Thus it would appear that both E. natalensis and S. birrea are influenced by disturbance in terms of tuber formation and invest in carbohydrate storage accordingly. The results would suggest that T. sericea is an obligate tuber forming species; the root/shoot ratios shown in Figure 1 are both relatively high compared to the other two species and the ratios of the two treatments are similar in magnitude, both with high standard errors. The result of the Man- Whitney U-test shows a highly insignificant difference (P >.9) between the no burn and annual treatments. Unfortunately I was unable to compare these to a control treatment as there were no T.sericea plants available in the nursery. In conclusion this study suggests that some woody species growing in frequently disturbed areas such as savannas do possess the ability to control the extent of tuber formation to a certain degree. This would potentially give them a major competitive advantage over obligate tuber formers as they are able to invest in energy storage according to the extant disturbance regime. Obligate tuber formers on the other hand have to invest in energy storage regardless of the disturbance regime, which could cost their competitive ability when growing in low disturbance environments. In contrast if facultative tuber formers growing in previously undisturbed areas were suddenly exposed to disturbance they may not have sufficient reserves to regenerate and are at risk of dying or out-competed by obligate tuber formers. Thus both strategies have both advantages and risks involved. This is

130 especially true for plants growing in an areas exposed to highly variable disturbance regimes such as savannas. This study was not extensive enough to determine whether the age of the plants affects the size of the tuber. A study by Malanson and Trabaud (1998) as cited in Bellingham (2000) found that after fire destroyed the above ground biomass of Quercus coccifera, nine year old shrubs sprouted more vigorously than three year old shrubs, which presumably had less developed below-ground storage reservoirs. While Del Tredici (2001) claims that the lignotuber continues to expand throughout the life of the tree and eventually forms a basal burl with high carbohydrate content. The sample size of this study was very small, thus in order to come to a definite conclusion with regards to whether tuber formation in savanna woody species is obligate or facultative a much more extensive study would need to be undertaken. The outcomes of this study do suggest that S. birrea and E. natalensis are facultative tuber forming species and T. sericea is an obligate tuber-forming species.

Literature Cited Bellingham, P. J. 2000. Resprouting as a life strategy in woody plant communities. Oikos 89(0):1-7. Blake, T. J. 1972. Studies on the lignotubers of Eucalyptus oblique L’Herit. New Phytol. 71:327- 334. Bond, W. J. and B.W. van Wilgen. 1996. Fire and plants. Chapman and Hall. London. Bond, W. J. and J. M. Midgley. 2001. Ecology of sprouting in woody plants: the persistence niche. Trends in Ecology and Evolution. 16 (No.1): 45-51. Bond, W. J. and J. M. Midgley. 2003. The evolutionary ecology of sprouting in woody plants. International Journal of Plant Science. 164 (3 Suppl.):103-114. Del Tredici, P. 2001. Sprouting in Temperate Trees: A Morphological and Ecological Review. The Botanical Review 67(2): 121-140. Ligavha, M., W. D. Stock, and W. J. Bond. 2001. The physiology of sprouting and its ecological implications in savanna and woodland dynamics. www.dwaf.gov.za/Forestry/IFM/Docs/ CD1/Doc%2029%20-%20Ligavha.DOC Whitcock, S. P., L. A. Apiolaza., C. M. Kelly, and B. M. Pots. 2003. Genetic control of coppice and lignotuber development in Eucalyptus globules. Australian Journal of Botany 51:57-67.

131

Table 1. The results of the Man-Whitney U-tests used to test for significant differences in root/shoot ratios for the three species and three treatments. U-values and p-levels are both given for each comparison. Species Treatments being tested U-value P-level Euclea natalensis Annual vs. No burn 2 0.028 Euclea natalensis Annual vs. Control 8 0.347 Euclea natalensis No burn vs. Control 2 0.028 Sclerocarya birrea Annual vs. No burn 8 0.347 Sclerocarya birrea Annual vs. Control 0 0.009 Sclerocarya birrea No burn vs. Control 5 0.117 Terminalia sericea Annual vs. No burn 12 0.917

4.5

3.5

2.5

Root/Shoot Ratio Root/Shoot 1.5

0.5

l) ) ) ) a rn) ol rn) ol r u r nu bu nt b nt n o o o o (a n (c t t(c r (annual a ir n na bir (n b se c nat( u c cl cl Scl bir (annual) S E Eu Euc S Ter Ter ser (no burn) Treatment

Figure 1. Mean root:shoot ratios of the three species from the Kambeni fire plots(error bars represent standard error of the mean). Euc nat = Euclea natalensis, Scl bir = Sclerocarya birrea and Ter ser = Terminalia sericea.

132 Root suckering dynamics of Dichrostachys cinerea

Category: Independent Project Participants: Gareth Hempson (secretary), Laurence Kruger (resource person) Site: Pretoriuskop, Kruger National Park, Mpumalanga, South Africa

Keywords: Demographic Bottleneck Model, Dichrostachys cinerea, disturbance, root sucker

Abstract: Root suckering is the phenomenon whereby a plant is able to produce new stems from buds on its roots. Dichrostachys cinerea is a widely occurring species well known for its ability to encroach in disturbed areas. Reports of its ability to spread vegetatively by means of root suckering prompted this investigation into its rooting morphology, as an attempt to begin to understand the regeneration dynamics of D. cinerea in savanna systems. It does not appear to conform to the savanna plant life history strategy proposed by the Demographic Bottleneck Model, in that it thrives within the fire trap. Twelve individuals were excavated in the mesic savanna region surrounding Pretoriuskop in the Kruger National Park, South Africa. Six of these individuals were found to be linked root suckers, while no root connection was found in the other six individuals. None of the root suckering individuals had a tap root, suggesting that a lateral spreading network of roots without a tap root may be a feature of D. cinerea where it is root suckering. Of the individuals where no root connection was observed, three individuals had a tap root and three did not. The stem basal area was found to be positively correlated to bole area (r2 = 0.817), and also to root area (r2 = 0.5301), suggesting a consistent pattern of resource allocation throughout the lifespan of an individual. The relationship between above ground and below ground biomass as well as observations of the nature of root suckering is hoped to provide a useful starting point for future studies. The observation of root suckering in D. cinerea is important for future management strategies, and also requires the definition of a new life history functional group for savanna systems.

Introduction Dichrostachys cinerea is a widely occurring member of the Mimosaceae, its natural distribution extending from northern Australia, through Burma and India into all parts of Africa (Coates Palgrave 2002). It has also been introduced into various regions including Cuba and Florida, U.S.A., where in parts it has become highly invasive (Weir 1927). The ability of D. cinerea to form impenetrable thickets has been widely reported, generally accredited to being in response to disturbances such as fire, overgrazing and attempts at mechanical removal (e.g. Coates Palgrave 2002, Van Wyk and Gericke 2000, Weir 1927). The capacity for this rapid regeneration and colonization response has been linked to its ability to spread rapidly via reseeding, but also to its potential to form root suckers (Weir 1927). Root suckering is the phenomenon whereby a plant is able to produce new stems from buds on its roots. Two types of root buds are known: additional buds which are formed from deep tissues in uninjured roots (Del Tredici 2001), and reparative buds which are formed on the surface of roots in response to injury or senescence (Bosela and Ewers 1997 as cited in Del Tredici 2001). Root suckering in trees has been found to largely be in response to injury to the primary trunk, while shrubs tend to produce new stems spontaneously as a normal part of their development (Del Tredici 2001). The amount of light received by a tree has also been shown to affect root suckering, with uninjured primary trunks being more likely to produce root suckers in full sunlight than in shady conditions (Del Tredici 1995 as cited in Del Tredici 2001). Root suckering in temperate trees native to mesic habitats has generally been observed as being a reparative response, with clonal growth being a secondary consequence (Del Tredici 2001). Frequent fire and heavy logging are disturbances that have been shown to promote the spread of root suckering species over non-root suckering species (Burns and Honkala 1990 as cited in Del Tredici 2001). The recently proposed Demographic Bottleneck Model for savanna plant life histories (Bond personal communication) does not take into account the implications of a root suckering response to fire disturbance in savanna plants. This model proposes two bottlenecks in savanna plant life histories: during the establishment of seedlings, and in the process of escaping from the fire trap. Seedling establishment appears to be limited by a combination of competition for water and light while within the grass layer, fire and browsing (Scholes, et al. 2003). The escape from the fire trap requires the plant going through the ‘Gulliver Phase’, where individuals are burnt down to the ground by regular fires

133 (Bond and Van Wilgen 1996). Release from this phase occurs when the period between fires is long enough to allow plants to resprout and grow above the fire zone. Gullivers are able to reduce the time necessary to escape the fire trap by building up large underground reserves and then growing a tall straight pole to boost leaves above the danger zone. It is only after escaping from the fire trap that individuals mature into fully reproductive adults. Dichrostachys cinerea seldom grows tall enough to escape the fire trap (personal observation), and rather reaches reproductive maturity and thrives within this zone. If this species were to be found to show a root suckering response to fire disturbance, with fire not impeding its spread through savannas, it would necessitate the definition of a new strategy axis or plant functional group for inclusion in the Demographic Bottleneck Model of savanna plant life history strategies. This study is an exploration into the root morphology of D. cinerea subsp. africana in the mesic savannas of the south western Kruger National Park. Particular attention is focused on searching for evidence of root suckering in this species within the region. The investigation of root morphology will include attempts to determine if rooting strategies are uniform in the species, or whether a range of root growth forms can be identified.

Methods Study site This study was carried out in the experimental burn plots laid out near to Pretoriuskop in the south western corner of the Kruger National Park, South Africa. These burn plots form part of an experiment that has been running for 50 years, aimed at assessing the impacts of different burning seasons and fire frequencies on savanna vegetation. Fieldwork took place in two sessions, 15 – 20 January 2004 and 20 – 24 February 2004. Six individuals were excavated in the Shabeni February 2 year burn plot, and six in the Kambeni August 3 year burn plot. The choice of these sites was arbitrary, the criterion rather being to fit in logistically with other fieldwork programs in progress at the same time. Fieldwork Twelve individuals of D. cinerea were excavated. The extent of digging varied from one individual to the next, but two general procedures were adhered to. Firstly, the bole at the base of the stem was excavated, and where present, the tap root was excavated to a depth of at least 30cm, or until it was felt that no substantial lateral roots would be found with deeper excavation. The second focus while digging was to follow all substantial lateral roots which were viewed as being potential sources of root suckers. No root system was excavated in entirety and it must thus be recognised that some root suckering connections may have been overlooked. The following measurements were made on each individual: height (dead/alive), canopy width (dead/alive), number of stems (dead/alive), stem base diameters (dead/alive), maximum diameter of bole, number and diameter of lateral roots, digging depth, distance to nearest neighbour and the length of the connecting root where root suckering was observed. Digital photos were taken of all excavations. Analysis Correlations pertaining to the biology of D. cinerea were performed using graphing functions in Microsoft Excel. Statistica was used to compare ratios of above ground to below ground material between individuals grouped by root growth form.

Results Root growth form Six (Dc 7 - 12) of the 12 individuals excavated in this study were found to be linked root suckers. None of these root suckering individuals had a classic tap root i.e. there was no major root extended vertically from the bole into the ground. The rooting morphology was rather that of an extensively spreading lateral network of roots. One individual (Dc 11) had its largest root projecting downwards at an angle of 45 degrees, but this was not followed to any great depth. No root suckering connection was found in any of the six other individuals that were excavated. Three of these individuals (Dc 1 – 3) had large tap roots that extended vertically into the soil and tapered with depth. No deviation in this vertical path was observed. The remaining three individuals (Dc 4 -6) in the study did not have tap roots, their general root morphology being quite similar to that of Dc 7 – 12, although as stated, no root suckering connection was observed in these three individuals. A significant difference in the ratio between stem base area and tap root area of root suckers and individuals where no connection was observed (P =.0455), and is illustrated in Figure 1. The mean stem base area to tap root area ratio was 0 for root suckering individuals and 0.12 for individuals where no root connection was observed. Biology

134 Figure 2 shows how live stem basal area was directly related to bole area (r2 = 0.817), despite root suckering and non-root suckering individuals being analysed together. Root area also showed a positive relationship with live stem basal area (Figure 2), although this was not as closely correlated (r2 = 0.5301). Root suckering observations A plan diagram of the root connections between the root suckering individuals (Dc 7 – 12) is shown in Figure 3. The nature of the connections varied quite considerably. The connections between Dc 9 and Dc 10 and between Dc 10 and Dc 11 are both large roots that travel directly to the next stem, just below the surface. The other three connections are all via branching roots. The connection between Dc 7 and Dc 10 was the most circuitous and travelled at least 40cm below the soil surface before emerging as a root sucker. The roots connecting Dc 8 to Dc 12 and connecting Dc 10 to Dc 12 were small, having diameters less than 5mm. A further observation on visiting the excavation site a month later was that of small green root suckering stems that had appeared in places on the exposed roots.

Discussion This study, although limited to a large degree by the amount of time available for fieldwork, does raise a number of issues with scope beyond that of this report. The most important result is that of root suckering being found to occur in D. cinerea in southern African savanna systems. The fact that D. cinerea is root suckering, coupled with the observation of it flourishing within the fire trap would strongly suggest that its life history strategy differs from that of Gullivers as modelled in the Demographic Bottleneck Model. Dichrostachys cinerea would conceivably be able to spread through a frequently burnt region by tolerating fire damage, resprouting basally, reseeding and root suckering. At the same time, the Demographic Bottleneck Model suggests that other species may be trapped in the Gulliver phase, relying on reseeding from individuals that have previously managed to escape the fire trap to extend their distribution (Bond and Van Wilgen 1996). The question is thus raised as to what in fact does limit the expansion of the range of D. cinerea. Interspecific competition may prove to be one of these limiting factors, with D. cinerea being shaded out by other species that have managed to escape the fire trap. The discovery of root suckering also has important implications for managers in savanna systems where there is a need to control encroachment by D. cinerea. Burning and mechanical bush clearing could perhaps assist D. cinerea invasions; its root suckering response may prove to be more effective in colonising the disturbed areas which are formed than that of its non-root suckering competitors. This promotion of the spread of root suckering species over non-root suckering species has previously been observed in North American systems as a response to frequent fire and heavy logging (Burns and Honkala 1990 as cited in Del Tredici 2001). This study provides potential evidence for different root growth forms being associated with root suckering and non-root suckering individuals. None of the root suckering individuals had a tap root. Should this observation be found to hold true by future studies, it would provide a useful tool for rapidly assessing whether an individual is likely to be a root sucker or not. Research in this field is considerably slowed by the need for extensive excavation. It would also have important implications for managers if, for example, it were to be found that controlling the spread of a root suckering form of D. cinerea was substantially different from that of a reseeding form. Much work remains to be done on the biology of D. cinerea. The high linear correlation between stem area and bole size (Figure 2) may suggest a consistent pattern of resource allocation as an individual increases in size. This would be in contrast to the strategy employed by Gullivers, where a large resource base is first built up in the bole and is then later used to rapidly grow a tall straight pole to above the level of the fire trap (Bond and Van Wilgen 1996). The factors determining the extent of basal resprouting and also possibly root suckering after a fire still require extensive investigation, with this study hopefully providing a useful starting point. The variability in the nature of root suckering connections between individuals is interesting. It would initially appear that the production of root suckering stems is fairly opportunistic, with roots producing stems simply wherever they are exposed to sunlight. Dichotomies may well, however, exist in responses to high and low disturbance. More directed root suckering (e.g. Dc 9-10 and Dc 10-11) may be associated with periods of high disturbance, being a response designed to spread the risk of an individual being eliminated. The more circuitous connections between root suckering individuals (e.g. Dc 7 – Dc 10) could occur during periods of high or low disturbance, and may simply reflect the result of a root’s chance exposure to sunlight. The processes controlling root suckering require much further study. One useful observation may however be that all root suckering observed in this study was associated with the bole at the base of the stem. No root suckering was observed from deep lateral roots off a tap root.

135 It is hoped that this study will provide a useful basis for future research into the ecology of D. cinerea. The simple suggestions of relationships between above and below ground material and the observations of root suckering connections are hoped to provide insight into the design of future studies. The documentation of root suckering in a species that thrives within the fire trap begs the description of an additional life history strategy functional group to that already described by the Demographic Bottleneck Model. It also poses questions regarding the best strategy for future management of encroaching D. cinerea.

Literature cited Bond, W. J. and B. W. van Wilgen. 1996. Fire and plants. Chapman and Hall, London. 263pp Bosela, M. J. and F. W. Ewers. 1997. The mode of origin of root buds and root sprouts in the clonal tree Sassafras albidum (Lauraceae). American Journal of Botany 84: 1466-1481 Burns, R. M. and B. H. Honkala (eds.) 1990. Silvics of North America. 2 volumes. U. S. Forestory Service Handbook. 654 Coates Palgrave, M. C. 2002. Trees of Southern Africa 3rd Edition. Struik Publishers, Cape Town. 1212pp Del Tredici, P. 1995. Shoots from roots: A horticultural review. Arnoldia 55(3): 11-19 Del Tredici, P. 2001. Sprouting in Temperate Trees: A Morphological and Ecological Review. The Botanical Review 67(2): 121-140 Scholes, R. J., W. J. Bond and H. C. Eckhardt. 2003. Chapter 11: Vegetation Dynamics in the Kruger Ecosystem in The Kruger Experience. Ed. J. H. Du Toit, K. H. Rogers and H. C. Biggs. Island Press, Washington. 519pp Van Wyk, B-E and N. Gericke. 2000. People’s plants. Briza Publications, Pretoria. 351pp Weir, J. R. 1927. The problem of Dichrostachys nutans, a weed tree, in Cuba with remarks on its pathology. Phytopathology 17: 137-146 http://www.hear.org/pier/species/dichrostachys_cinerea.htm

0.35

0.30

0.25

0.20

0.15

0.10

0.05

0.00 Stem area:Tap base root area

-0.05

-0.10

-0.15 Mean No connection Root sucker Mean±SE Mean±SD

Figure 1. Ratio of stem base area to tap root area of individuals showing no root connection with individuals found to be root suckering. Root suckers have a significantly lower stem base area to tap root area than individuals showing no connection (P=.0455).

136

Stem base area vs Bole area & Root area

16000

14000

) 2 2 R = 0.817 12000

10000 Stem vs Bole 8000 Stem vs Root

6000

4000 Total of area bole/roots (mm 2000 R2 = 0.5301 0 0 500 1000 1500 2000 2500 3000 3500 4000 Total stem base area (mm2)

Figure 2. Relationship between stem base are and bole area and stem base area and root area. (Stem vs. bole r2 = 0.817; Stem vs. roots r2 = 0.5301).

Figure 3. Plan diagram showing root suckering connections of individuals Dc 7 – Dc 12.

137 Appendix 1. Raw data for individuals Dc 1 to Dc 6 for which no root suckering connection was found.

138 Appendix 1. (Continued) Raw data for individuals Dc 7 to Dc 12 which were all connected root suckers.

139 Waterhole preferences among large mammalian herbivores in the Skukuza region of the Kruger National Park, South Africa, based on track evidence

Category: Independent Project Participants: Tammy Baudains, Megan Eastwood & Govan Pahad Site: Skukuza area, Kruger National Park, Mpumalanga Province, South Africa

Key words: large herbivores, predator avoidance, preferences, vegetation cover, water availability, waterhole

Abstract: Preference of large herbivores for waterholes of varying size and vegetation cover was investigated. Nine waterholes were sampled by identifying and counting tracks in 1m2 quadrates every fifth meter around the waterhole’s perimeter. Waterholes were compared in terms of percentage individual species composition, as well as size class and predator-avoidance strategies of visiting species. Mega-herbivores (elephants, giraffes, and hippos) preferred large waterholes. Preference for degree of cover at a waterhole was related to the species’ predator-avoidance strategies: species that flee from predators preferred open waterholes; species that hide from predators preferred abundant cover close to the water’s edge; and impala preferred intermediate cover some distance from the water’s edge as they scatter and hide once predators have been detected.

Samevatting: Ons het die voorkeur van groot plantetend diere vir drinkplekke van verskillende grootte en planteryk ondersoek. Opname was uitgedra op nege drinkplekke en die spore wat in een vierkantige meter viekant gevind was, was getel en geidentifiseer. Die viekants was elke vyf meter op die omtrek van die watergat geplaas. Drinkplekke was gevergelyk met betrekking tot persent individueel spesie komposisie, so wel as grootte orde en roofdier ontduiking strategies van spesies wat die watergat gebruik het. Groot plantetend diere het die groter drinkplekkke voorkeur omdat hulle individueel benodigdhede vir afstand en water het. Voorkeur vir die bedrag van planteryk oordekking by ‘n drinkplek was verwant na die spesie se roofdier ontduiking strategies: spesies wat van roofdiere vlug, het drinkplekke in die ooplug voorkeur; spesies wat van roofdiere wegkruip, het drinkplekke met baie planteryk ontdekking voorkeur; en rooibok het ver maar baie planteryk ontdekking van die drinkplek se kant voorkeur omdat hulle hardloop en vegkruip wanneer hulle roofdiere sien.

Introduction Water availability is an important landscape-scale constraint on dry-season herbivore distributions and affects all herbivore species aggregations, impacts, and range overlaps (Redfern et al. 2003). The provision or exclusion of waterholes is one of the few tools available to ecosystem managers that can aid them in bringing about changes to achieve complex long term goals such as ensuring vegetation heterogeneity and thus faunal diversity in the park. Water provision therefore ranks with fire and culling as one of the main intervention tools (Gaylard et al. 2003). The Kruger National Park has a long history of manipulation of water sources, much of which was unsuccessful due to a lack of understanding of how surface water availability affects large mammals and other aspects of the ecosystem. When the Kruger National Park was established in 1898, few natural sources of perennial water existed between its five perennial rivers, and large sections of the Kruger supported very few herbivores, especially in the dry season (Gaylard et al. 2003). Also, the completion of the western boundary fence in 1961 prevented any animal movements that may have taken place westward towards the wetter foothills of the escarpment (Gaylard et al. 2003). A water provision program was established in the 1930s with the aims of supporting herbivore populations during droughts and minimizing the influences of temporal variability of rainfall (Gaylard et al. 2003). It was feared that without such a policy, rare antelope species such as roan, tsessebe and sable could be lost during severe droughts. However, despite large scale water provision, populations of rare antelope fell dangerously during the extended drought period of 1982-1987. Research suggests that the provision of artificial water-points may have reduced herbivore diversity by expanding the range of common water-dependant species such as zebra and wildebeest and concomitantly predator species such as lion into the range of the less common antelope (Gaylard et al. 2003). The zebra and wildebeest

140 not only compete with the rarer antelope, but their heavy grazing also degrades the antelopes’ preferred habitat of long, lightly-used grasses into a short grass mat (Harrington et al. 1999). Recognition of the importance of understanding such processes has led to a reassessment of the practice of artificial water provision and stimulated various recent studies on waterhole interactions with herbivorous mammals. Scientists and managers in Kruger are removing artificial waterholes to help the rare antelope populations recover but also realize they need to study waterhole use further to better predict the consequences of their actions (Redfern et al. 2003). Redfern et al. (2003) studied whether species distributions relative to water sources corresponded to their degree of water dependence. Other studies have covered a range of topics including how herbivore activity affects vegetation around a waterhole, differences in waterhole use between the wet and dry seasons and even the times of day at which species prefer to drink. Relatively few studies, however, have looked at species preferences for different waterhole types, which would be useful information to have when deciding which waterholes to remove. Elephants prefer drinking from cement reservoirs, while buffalo, zebra and wildebeest prefer earth dams; and roan antelope reportedly drink most often from natural pans or springs (Gaylard et al. 2003). To supplement such observations, we tested two hypotheses concerning animal preferences for waterholes. Hypothesis 1 There will be a difference in the species composition of large mammalian herbivores visiting waterholes of varying sizes. 1A: Herding species will prefer larger waterholes due to demand for drinking space. 1B: Large species will prefer larger waterholes due to high individual water demands. Hypothesis 2 There will be a difference in the species composition of large mammalian herbivores visiting waterholes with varying degrees of cover. 2A: Species whose predator avoidance strategy is to hide will prefer waterholes with abundant cover. 2B: Species whose predator avoidance strategy is to flee will prefer open waterholes.

Methods All waterholes surveyed in this study were in the region surrounding the Skukuza rest camp in the Kruger National Park, South Africa. The sampling period extended over several days in late February 2004, during the wet season. We sampled tracks at nine waterholes with as great a variation in size and surrounding cover as possible. Quadrates of 1m2 were laid down every 5th meter around the perimeter of the waterholes, so that 20% of the perimeter of each waterhole was sampled regardless of its size. This means that approximately 20% of the tracks around each waterhole should have been sampled, representing the same proportion of animals that had visited each waterhole. As the area directly around the waterhole has the most mud and the least vegetation (both of which are favourable conditions for clear, identifiable tracks) we placed each quadrate as close to the water’s edge as possible. Within each quadrate, we identified and counted the tracks present but only recorded tracks that could be clearly identified as belonging to a particular species. Track field guides and the assistance of guides enabled us to identify more obscure tracks. Cover at each waterhole was assessed in terms of percentage woody cover and distance in meters from the water’s edge to the nearest cover. To assess species’ preference for waterhole types, we calculated the number of tracks found for each species visiting a particular waterhole as a percentage of the total number of tracks of that species seen at all waterholes. This should nullify any bias caused by some species’ tracks lasting longer or being easier to identify than others. We assumed that finding a great deal of a species’ tracks around a particular waterhole type would indicate a preference for waterholes of similar size and cover. Due to the small sample sizes for individual species, we grouped the herbivores together according to: a) predator avoidance strategies and b) size. The groups according to predator avoidance strategies were: those that hide from predators and escape by dodging amongst vegetation (common duiker, steenbok and kudu); those that flee from predators (Burchell’s zebra, blue wildebeest, buffalo and warthog); those that stand up to predators (giraffe, elephant and hippopotamus); and those that scatter into nearby vegetation to escape from predators (impala). The species were divided into these groups according to information from a behavioural guide by Estes (1999). The groups according to mass were: less than 100kg (steenbok, common duiker, warthog and impala); between 100kg and 1000kg (kudu, buffalo, Burchell’s zebra and blue wildebeest); and greater than 1000kg (elephant, hippopotamus and giraffe). The masses were taken from a field guide by Frandsen (2002). We assessed animal communities by using Primer to create a multi-dimensional scaling (MDS) plot (Primer 5).

141 Results The MDS plot for individual species composition at each waterhole revealed similarities between the two biggest waterholes (3 and 4) and among all the smaller waterholes (1, 2, 5 and 6). Waterholes sampled after rainfall (7, 8 and 9) also grouped together. The groupings are fairly strong as the average stress value for the plot was 0.1. (Figure 1) Herbivores of over 1000kg were most commonly detected at the larger waterholes, while those between 100kg and 1000kg showed no clear pattern in relation to waterhole size. The less than 100kg size class was found most commonly at the smaller waterholes. Impala were excluded from this analysis as impala tracks occurred in such large numbers, especially at large waterholes, that they obscured the results from other species in the less than 100kg size class. The impala were not excluded from the predator-avoidance group analyses, as they were in their own group (scatter group) and therefore did not hide results from less common species. (Figure 2) The hide/dodge group was found most commonly at the smaller waterholes (perimeter<55m), while the fight group was found most commonly at larger waterholes (perimeter>100m). Of the groups containing species that display herding behaviour, the scatter group was more commonly detected at larger waterholes whereas the flight group was randomly distributed relative to waterhole size. (Figure 3) Tracks from the hide/dodge group were most common around waterholes with a high percentage (>40%) of cover less than 3m from the water’s edge. Tracks from the scatter group were most common around waterholes which had an intermediate percentage of cover (15-39%) approximately 3-5m from the water’s edge. The only clear pattern related to cover shown by the flight group was that their tracks were uncommon around waterholes with a high percentage of cover close to the waters edge. Visitation by the fight group showed no pattern in relation to waterhole cover. In Figures 4 and 5, waterholes 8 and 9 are not included as we did not measure vegetation cover there due to heavy rain during sampling. (Figures 4 and 5).

Discussion The waterholes grouped by size in the MDS analysis (Figure 1), as the two biggest waterholes, 3 and 4, had very similar numbers of elephant, giraffe and impala tracks. The smaller waterholes grouped loosely in figure 1. We expect that these waterholes simply grouped by exclusion as the waterholes sampled in the rain (7, 8 and 9) grouped due to low track counts and the large waterholes (3 and 4) grouped due to similar species numbers and we could not find any distinct patterns between the species compositions for the small waterholes (1, 2, 5 and 6). Our results did not support hypothesis 1A, although impala (the scatter group) showed a preference for large waterholes. The large number of impala tracks at the large waterholes was expected because impala forage and drink in large herds and a waterhole with a large perimeter would be advantageous as it allows more individuals to drink at one time. However, the flight predator- avoidance group also includes herding animals, but it did not follow the expected pattern of higher numbers at larger waterholes, thus nullifying hypothesis 1A (Figure 3). The prevalence of tracks of species over 1000kg at the larger waterholes supports hypothesis 1B that big species would prefer bigger waterholes due to a large individual demand for water (Figure 2). Elephants, for example, require large amounts of water in which to bathe and large quantities of clean drinking water (Gaylard et al. 2003). Our results support hypothesis 2A, which predicted that species that hide from predators would prefer waterholes with lots of close vegetation cover to hide amongst and dodge through. As seen in figures 4 and 5, the hide/dodge group tracks were indeed more common at waterholes with a high percentage of close cover. Hypothesis 2B was also supported. The tracks of the flight group were less common at waterholes with close dense cover, probably because dense vegetation would obstruct their view of predators approaching and delay their escape (figures 4 and 5). Although trends in preference of the scatter group were not part of our main hypotheses, we noticed interesting patterns in the graphed data (figures 3, 4 and 5). The scatter group (the impala) seemed to prefer larger waterholes, which could be due to either to the large size or the cover available at those particular waterholes (intermediate cover some distance from the water’s edge). As discussed for hypothesis 1A, other herding species, i.e. the flight group, did not show a preference for large waterholes, suggesting that cover and thus predator-avoidance strategy was the primary cause of the impala’s preference. Large herbivores at low risk from predation appear to prefer larger waterholes, while other species seem to prefer waterholes where the vegetation cover facilitates their predator-avoidance strategy. If this information is accurate, park managers may be able to favour certain species or species

142 groups through manipulation of the type of water sources available, not just by waterhole number and distribution. The information might also enable the park to produce a waterhole guide for tourists interested in viewing particular species. Further studies performed throughout the year would be valuable to determine how strong these preferences are and whether the patterns seen are simply by- products of the abundance of water sources in the wet season. However, even if animals are willing to drink at unfavourable water sources during the dry season, it may still influence their distribution by increasing their risk of predation.

Acknowledgements We wish to thank Mike Smith for all of his assistance throughout the project; game guards Simione Mudhlovu, Lucas Masinga, Alione Phalene, and Godfrey Sekhula for their assistance in identifying tracks; Deedra McClearn, Julie Coetzee, and Laurence Kruger for their help with project design, statistics, and write-up, and Kinesh Chetty for providing transport.

Literature Cited Estes, R. D. 1999. The safari companion: a field guide to watching African mammals. Chelsea Green Publishing Company. White River Junction. Vermont. USA. Frandsen,, R. 1998. Southern Africa’s mammals: a field guide. Honeyguide Publications. Sandton. South Africa. Gaylard, A., N. Owen-Smith, and J. Redfern. 2003. Surface water availability: implications for heterogeneity and ecosystem processes. Pages 171-188 in Du Toit, J. T., K. H. Rogers, and H. C. Biggs, eds. The Kruger experience: ecology and management of savanna heterogeneity. Island Press. Washington DC. USA. Harrington, R., N. Owen-Smith, P. C. Viljoen, H. C. Biggs, D. R. Mason, and P. Funston. 1999. Establishing the causes of the roan antelope decline in the Kruger National Park, South Africa. Biological Conservation 90: 79-78. Primer 5 for Windows, Version 5.1.2. 2000. PRIMER-E Ltd. Redfern, J. V., R. Grant, H. Biggs, and W. M. Getz. 2003. Surface-water constraints on herbivore foraging in the Kruger National Park, South Africa. Ecology 84: 2092-2107.

Figure 1. MDS plot of the waterholes (1 to 9) according to individual species composition. (Stress = 0.1).

143 90

p 70

waterhole visiting 30

Percentage of size-class grou size-class of Percentage 20

under 55m 56-100m over 100m Waterholes grouped by perimeter size

% x<100kg % 100kg1000kg

Figure 2. Incidence function of waterhole perimeter size versus visitation by size classes (impala excluded from analysis). Twenty percent of each waterhole’s perimeter was sampled.

c 70

waterhole visitng group 30

predator-avoidan of Percentage 20

0 under 55m 55-100m over 100m Waterholes grouped by perimeter size

SCATTER HIDE/DODGE FLIGHT FIGHT

Figure 3. Incidence function of waterhole perimeter size versus visitation by predator avoidance groups. Twenty percent of each waterhole’s perimeter was sampled.

144

80 c 70 60

waterhole visiting group 20 predator-avoidan of Percentage 10

0 less than 15% 15-39% over 40%

Waterholes grouped by percentage of cover

SCATTER HIDE/DODGE FLIGHT FIGHT

Figure 4. Incidence function of percentage of vegetation cover at a waterhole versus visitation by predator avoidance groups. Twenty percent of each waterhole’s perimeter was sampled.

100 90

c 80 70

50 40

waterhole visiting group 30 20 predator-avoidan of Percentage 10

0 less than 3m 3-5m over 5m Waterholes grouped by distance to cover

SCATTER HIDE/DODGE FLIGHT FIGHT

Figure 5. Incidence function of distance to closest vegetation cover at a waterhole versus visitation by predator avoidance groups. Twenty percent of each waterhole’s perimeter was sampled.

145

Appendix 1. Size and vegetation cover of each waterhole sampled. We did not measure vegetation cover at waterholes 8 and 9 due to heavy rain during sampling.

Waterhole Perimeter size(m) % Woody vegetation cover Distance to closest vegetation cover(m)

1 50 66 <1 2 24 5 4 3 144 20 2-3 4 256 35 7 5 50 10 2-3 6 67 40 0 7 86 2 10-20 8 190 - - 9 70 - -

Appendix 2. Track counts for individual species per waterhole.

Species 1 2 3 4 5 6 7 (rain) 8 (rain) 9 (rain)

Loxodonta africana (Elephant) 0 0 24 20 0 3 0 0 0 Hippopotamus amphibious (Hippo) 0 0 0 0 0 0 0 9 2 Giraffa camelopardalis (Giraffe) 0 9 2 1 0 0 0 2 0 Syncerus caffer (Buffalo) 5 0 0 0 0 0 0 0 0 Connochaetes taurinus (Blue Wildebeest) 0 0 0 0 0 0 6 7 15 Tragelaphus strepsiceros (Kudu) 2 2 0 0 0 0 0 0 0 Equus burchelli (Burchell’s Zebra) 0 12 0 50 31 0 0 2 0 Phacochoerus aethiopicus (Warthog) 0 0 17 0 3 4 0 0 0 Raphicerus campestris (Steenbok) 0 0 1 0 0 0 0 0 0 Sylvicapra grimmia (Common Duiker) 0 0 0 0 0 14 0 0 0 Aepyceros melampus (Impala) 2 29 162 220 27 11 29 0 0

146 Spatial distribution of Ploceus velatus (Greater Masked Weaver) nests within the canopy

Category: Independent Project Participant: Laura L. Buckley Site: Skukuza, Kruger National Park, Mpumalanga Province, South Africa

Key words: colonial nesting, nesting site selection, Ploceus velatus, spatial distribution

Abstract: Greater Masked Weavers are common colonial nesting birds found throughout South Africa. There are several advantages to nesting communally and in certain spatial arrangements. This study focuses on the spatial distribution of Masked Weaver nests within the canopy and along a vertical stratum in the canopy. I performed spatial distribution analyses using photographs of four different nesting sites. The nests are generally arranged in a clumped fashion across the canopy, with a bias towards selecting sites in the upper-middle region of the canopy. More sampling will give insight into the potential relationship between tree size, nest abundance, and distribution in the canopy.

Introduction The Greater Masked weaver, Ploceus velatus, is the most widely distributed weaver in Southern Africa, found commonly throughout South Africa except most coastal and southern areas of Kwa-Zulu Natal and parts of the southeastern Western Cape. They utilize a wide variety of habitats, including semi-arid brush, savannas, riverine thickets, woodland edges, and farmland with tree clumps (Allan et al. 1997). They are often found near water and around areas with human activity. As gregarious birds, they forage and form breeding colonies in large groups. The males weave neat, round, oval-shaped nests from strips of grass, reeds, and leaves. These nests are most often attached to drooping tree branches or between upright reed stems (Gordon et al. 1993). Colonial nesting provides several benefits to the members of the colony. In colonies that forage together, information on food sources can be distributed throughout the group. Less successful foragers can follow more adept ones to sites as they leave the colony. Colonial nesting also provides a degree of protection from predators. Large numbers increase the chance of spotting approaching predators, as well as decrease the chance of any single nest in particular being preyed upon (Dobkin et. al. 1988). While research has been conducted on the advantages of nest distribution on a periphery- interior gradient (Dobkin et al. 1988), little has been done on the vertical distribution of weaver nests and the potential advantages of nest selection at different heights. My objective is to study the spatial distribution of Greater Masked weaver nests along a vertical gradient and distributional homogeny across the entire canopy. Based on assumptions of predator avoidance and avoidance of damage from browsers, I hypothesize that the nests will exist in a clumped distribution, localized in the upper regions of the canopy.

Methods I selected four trees that served as nesting sites for Masked Weaver communities in different locations near Skukuza: at the buffalo enclosure, at the golf course, at Kruger Gate, and on the road towards Pretoriuskop. Using triangulation, I measured the height of each tree, as well as the height of the canopy above the ground. I then took a photograph of each entire tree from an angle that would maximize the number of nests in the image. For analysis, I laid a grid over each image and performed a spatial distribution analysis. I then laid a second grid over the images, dividing each tree into four relative height classes (lower, lower- middle, upper-middle, and upper), and performed a Chi-square test for homogeny. I also used linear regression to determine the relationship between tree height and number of nests, and canopy height and number of nests.

Results Each tree was analyzed separately to assess individual nesting community structure and to determine whether patterns exist between nesting sites. The spatial analysis on the trees at Kruger Gate, the buffalo enclosure, and the golf course yielded results indicating a clumped distribution of nests (Table 1). The analysis on the tree on the road to Pretoriuskop indicated that the nests here are distributed randomly.

147 I also performed a Chi-square test on each tree to determine the vertical stratification of nesting sites. The nests at the Kruger Gate and the golf course sites show a non-homogenous distribution (Table 2). The nests at the buffalo enclosure site show a more homogenous distribution across height classes, while the nests on the Pretoriuskop road site were highly homogenous with relation to distribution across height levels. When looking at the distribution of nests within specified height classes, I observed that the greatest percentages of nests in each tree were found in one of the two middle zones, with a slight bias towards the upper-middle zone (Figure 1 and Table 3). The regression analysis on the relationship between canopy height and number of nests indicated that there was no relationship (R2=0.02, P=.12). Since there is a strong linear correlation between canopy height and tree height, a regression between tree height and number of nests yielded similar results (R2=0.96, P=.12) (Figures 2 and 3).

Discussion The clumped nature of the nests in the three sites suggests that there may be some advantage for weavers to place their nests in a certain area on the tree and in a dense fashion. Exactly where this clumping occurs within the canopy will reveal more about what types of advantages there may be. One possible reason for this arrangement is to increase predator detection and avoidance. The random distribution on the fourth tree contradicts this explanation of advantage by clumping; however, the tree size and number of nests found there were much smaller, which could be a compounding variable. When the relationship between tree and canopy size and number of trees per tree is examined, the regression shows no relationship. The single outlier, the Kruger Gate site, with a vastly different height-to-number of nests ratio, draws the regression away from what appears to be an otherwise straightforward linear relationship. More sampling must be done to determine if this is an isolated phenomenon and a relationship does exist. Afterwards, I will be able to make a better assessment of the relationship between spatial distribution of nests, the number of nests, and the size of the tree. The Chi-square tests indicate that there is a random distribution of nests between height classes on the trees on the Pretoriuskop road and at the buffalo enclosure. With regard to the buffalo enclosure site, these results indicate that although distribution within the canopy is clumped, it is not clumped in a particular height class. The chi-square tests for the Kruger Gate and golf course sites indicate that the clumped nest distribution was indeed within certain height classes. Analysis of the distribution of nests between height classes shows that in each site, the largest percentage of nests is found in either of the middle two classes. All sites except the buffalo enclosure show the majority of nests located in the upper-middle zone. This suggests that the extreme top and bottom of trees are not desirable sites for weavers to nest. Avoidance of the lower regions may be a means to avoid potential damage from browsers. Avoidance of the upper region may be a form of protection from potential predators from above. Other factors must be taken into consideration when examining nest site choices with individual trees, such as the availability of suitable branches and the requirements for such suitable branches. Greater Masked Weavers tend to build their nest colonies in a clumped distribution across the canopy. This clumping tends to occur in the upper-middle region of the canopy along a vertical gradient, although the data suggest that other dimensions, such as horizontal and depth measurements, must be taken into consideration to explain the clumping patterns. Tree and canopy size may play a role in the number of nests, and this number may play a role in their distribution. Initially, no correlation has been found, but the low sample size suggests that more sampling will give a better indication of the correlation. After the distribution patterns have been well examined, behavioral observations will give valuable insight into any relationship between nest site selection and external influences.

Acknowledgements I would like to thank Deedra McClearn, Laurence Kruger, Julie Coetzee, and Kinesh Chetty for their valuable suggestions and guidance in project design, statistical analysis, and last-minute crisis- management.

Literature Cited Allan, D.G., Brown, C. J., Harrison, J.A., Herremans, M., Parker, V., Tree, A.J., Underhill, L.G. The Atlas of Southern African Birds (Volume 1). 1997. Bird Life South Africa. Dobkin, D.S., Ehrlich, P.R., Wheye, D. Colonial nesting. 1988. http://www.stanfordalumniservices/birdsite.htm Gordon, Lindsay, Maclean. 1993. Robert’s Birds of Southern Africa (Sixth Edition). Trustees of the John Voelcker Bird Book Fund: Cape Town, South Africa.

148

Table 1. Test statistics for spatial distribution analysis Location χ2 Degrees of P-value freedom (ν) Kruger Gate 270.28 18 <0.05

Pretoriuskop 37.2 25 >0.05 Road

Buffalo 70.0 24 <0.05 enclosure

Golf course 187.26 28 <0.05

Table 2. Test statistics for homogenous distribution along a vertical gradient Location χ2 Degrees of p- freedom (ν) value Kruger Gate 30.46 3 <0.05

Pretoriuskop 3.60 3 >0.05 Road

Buffalo 9.56 3 <0.01 enclosure Golf course 20.50 3 <0.05

Table 3. Percentages of total nests per tree located in each height class Location Height Class* 1 2 3 4 Kruger Gate 27.7% 41.9% 22.6% 7.8%

Golf Course 25.6% 57.8% 16.7% 0%

Buffalo 0% 17.5% 60.0% 22.5% Enclosure

Pretoriuskop 10.0% 50.0% 20.0% 20.0% Road *Class 1 is highest, class 4 is lowest to ground

149 100 90 80 70 60 Kruger Gate 50 Golf Course 40 Buffalo Boma 30 Pretoriuskop 20

Percentagenests per tree 10 0 1234 He ight cla sse s

Figure 1. Nest abundance along a vertical gradient.

350 300 250 200

150 2

# nests R = 0.0239 100 50 0 0 10203040 Canopy height

Figure 2. Relationship between canopy height and number of nests per tree.

350 300 250 200 2 150 R = 0.1104 # nests 100 50 0 0 5 10 15 20 25 30 Tree height

Figure 3. Relationship between tree height and canopy height.

150

Punda Maria Independent Projects

151 The revenue generation of ecotourism and ecosystem services: two primary practices by South African rural communities

Category: Independent Project Participants: Scott Briscoe Site: Punda Maria, Kruger National Park, Mpumalanga Province, South Africa

Key Words: ecosystem services, ecotourism, natural resource management, revenue generation

Abstract: Ecosystem services and ecotourism are two primary sources of generating revenue for communities in South Africa today. Ecotourism has been an effective method of resource management in generating viable economic development in rural South African Communities. Ecotourism concentrates on conserving the environment while at the same time assisting in sustaining the best interests of local people. Ecosystem services include, but are not limited to, fuelwood, medicinal plants, river sand and industrial timber, provide rural employment and cash income through sales. In investigating the total projected revenue generation of both practices, I have concluded that by combining the two methods South African rural communities can ensure economic diversity and economic, social, political and psychological empowerment.

Introduction Advocates of ecotourism argue that it is an effective method for providing revenue from tourists and the private sector, while at the same time providing economic opportunities for rural communities living near parks and protected areas (World Resource Institute 2000). The Ecotourism Society defines ecotourism as “responsible travel to natural areas that conserves the environment and sustains the well-being of local people” (The International Ecotourism Society 2004). The Makuleke Region within the KNP is one example of a South African rural community which has embodied the ideas of conservation and economic rural development through ecotourism. The Makuleke community, who is descendant of the Tsonga-speaking Maluleke, is known to have inhabited the north eastern region of South Africa since the early 1800's; the areas between the Levuvhu and Limpopo rivers. In 1906 the region was officially designated to the Makuleke by the British. In 1969, as a result of interests by the National Parks Board (NPB) to expand the KNP, the Makuleke community was forcibly removed from their land and relocated outside the Kruger gates. In the light of 1994 and the end of legal apartheid, the Makuleke community had filed to regain their land back through the National Land Claims Commission (NLCC) and the Department of Land Affairs (DLA), whose responsibility was the units Directorate Restitution (Steenkamp 2000). Through the support of other non- governmental organizations (NGO’s) and approximately four years of arduous negotiations, the signing ceremony for the transfer of the land ownership of 22734ha took place in April of 1998 (Steenkamp 2000). Rather than relocating their existing village outside of the park to their reclaimed original land, the Makuleke have used the region to develop responsibly and use the natural resources in a sustainable manner in order to assist in their own community development. The primary source of revenue for the community, which now exists outside of the KNP, is ecotourism. As a first step in insuring equitable development for the Makuleke community, trophy hunting was implemented as a form of ecotourism. Although trophy hunting was suspended in 2001, it has been the single source of revenue since 1998, generating more than R2 million (Maluleke 2004). Those funds have been directly redistributed back into the Makuleke community, funding a community project for electricity, a school project, creating 4 new classrooms, the building of 2 new primary and secondary schools and the maintenance of the Makuleke Communal Property Association (CPA) (Maluleke 2004). The primary function of the CPA includes the conservation status of the land within KNP and working to maximize sustainable community benefit through commercial development of the region (Steenkamp, Jana Uhr 2000). Since the suspension of trophy hunting, the Makuleke have invested in ecotourism concessions with premier resorts. The Outpost Lodge has been the first of these concessions and is situated at the confluence of the Luvuvhu and Mutale rivers. In addition to the creation of jobs (the majority of employees, under contract, must come from the Makuleke community), 14% of the total annual income generated from the Outpost goes into a trust account which is used for development projects (Maluleke, L. 2004). The current contractual agreement between the Outpost Lodge and the Makuleke community is for 30 years at which time the concession ends and the community takes over the ownership and all operations of the lodge (Maluleke, L. 2004). Critics of ecotourism claim that some effects of ecotourism include habitat fragmentation, air 152 pollution from vehicles traveling in the conserved area and litter (World Resource Institute 2000). However, the Makuleke have proven that with proper management practices the impacts of ecotourism on can be minimized. One alternative practice to ecotourism as a revenue generating source is ecosystem services. In addition to providing a vital income, bio-resources enable rural communities to generate employment and fulfill domestic consumption needs (Sharma 2000). The quality and quantity of the species is critical in determining the value of resources within a specified region. It is important to differentiate here the applied values to resources; there is a direct-use value and an indirect-use value. In Robin Grimble’s paper on biodiversity management, she defines direct- use values as those physical goods used by people and indirect-values as those ecological functions which help to maintain the stability and productivity of the environment. Although both use values are significant to economic and sustainable development, in this paper I will only be focusing on direct-use values.

Methods In researching ecotourism as a source for revenue generation, the Makuleke model for ecotourism was used. A personal interview with Tinyeko Maluleke, from the Makuleke community was held at the Outpost Lodge in the Makuleke region in KNP. Tinyeko Maluleke is a current employee at the Outpost and holds a managerial position. A personal interview was also conducted with Livingston Maluleke, the spokesperson for the CPA and the a chairperson who sits on the Joint Management Board (JMB). The JMB handles decisions regarding land and conservation and is comprised of three members from South African National Parks (SANP) and three members from the Makuleke community. The interviews consisted of a list of questions (fig 1) regarding employment generated income, the allocation of funds generated from ecotourism ventures, methods of reassuring conservation principles, average mean incomes, sources of household incomes. Because of the sensitivity surrounding the Makuleke land claim with SANP and the South African government, collecting specific income numbers was a timely process and not all the data needed for an effective comparison was gathered. In investigating the revenue generating possibilities of ecosystem services, several previously written papers were used as sources. I then compared the two income projections, i.e. ecotourism and ecosystem services, to help in determining the best possible practice for a South African rural community.

Results Since 1998 when the land claim settlement was settled, the Makuleke Region has generated approximately R2 million (Maluleke, L. 2004). In addition, 14% of the annual income generated from the Outpost Lodge is allocated to the Makuleke community. There is also a contractual agreement between the two parties for the majority of employment to come directly from the Makuleke community. In total the projected income generated by the lodge is approximately $35,000/R210,000 a year (at current exchange rate) (Unesco 2001). In 2003 direct-use values were calculated by multiplying the mean annual consumption per village by the current local price. This projected total places the economic value of resources for each household at R3959 (fig. 2). This total underscores the economic viability of natural resources within South African rural communities.

Discussion In investigating ecosystem resources and ecotourism as effective methods of revenue generation for South African rural communities, the limited data collected suggests that there are particular situations which allow for the use of one method over the other. In other situations, the integration of both practices would be most effective. It is also, necessary to look at the what the community has access to. Many villages are in agriculturally marginalized areas, decreasing the potential for them to utilize any resources as a viable income (Twine 2003). In circumstances like these, ecotourism seems to be the most viable option for generating an income. The development of ecotourism in South African villages could be in any form, from larger commercial ventures such as the Makuleke and the Outpost Lodge to smaller scale enterprises such as tours through heritage sites or small scale markets for locally made jewelry. These are only a few examples of how ecotourism can empower rural communities. In situations where communities have access to agriculturally sustainable land, as Wayne Twine’s projections show, ecosystem services are an important consideration for development that is both equitable and sustainable. However, looking at the projected numbers for both methods and the potential circumstances which would make the land unusable, the integration of a well managed practice using both, ecosystem resources and ecotourism could be the best option for the empowerment and economic development of a South African community.

Acknowledgments I would like to gratefully acknowledge Deedra McClearn, Laurence Kruger and Julie Coetzee for their 153 patience, understanding and support and their seemingly tireless commitment to the achievement of all of the OTS students, 2004. Sincere thanks and appreciation to Tinyeko Maluleke and Livingston Maluleke, for their time and effort in conducted interviews with me. I would also like to thank Eric Caldera for his assistance in documentng interviews and Kyle Harris for his support and friendship. In addition, I would like to sincerely thank and acknowledge the Mellon Foundation for allowing me the opportunity to enhance my academic experience in South Africa.

Literature Cited Bosch, Dawie (compiled by). 1999. The Makuleke Land Claim Settlement Agreement. Grimble, Robin, Martyn Laidlaw. 2002. Biodiversity Management and Local Livelihoods: Rio Plus 10. Natural Resource Perspectives 73. Maluleke, Livingston. 2004. A personal interview conducted by Scott Briscoe. Spokesperson for CPA and chairperson for Joint Management Board. Sharma, Narendra P., Simon Rietbergen, Claude, R. Heimo and Jyoti Patel. The Strategy for the Forest Sector in Sub-Saharan Africa. World Bank Technical Paper Number 251: Africa Technnical Department Series. 13-14. Steenkamp, Conrad, Jana Uhr. 2000. The Makuleke land claim: power relations and community-based natural resource management. Evaluating Eden series. Discussion paper no. 18 The International Ecotourism Society (TIES). 2004. http://www.ecotourism.org/index2.php?what-is-ecotourism. Twine, Wayne, D. Moshe, T. Netsiluvhi, V. Siphungu. 2003. Consumption and direct-use values of savanna bio- resource used by rural households in Mametja, a semi-arid area of Limpopo province, South Africa. South African Journal of Science 99:467-473 Unesco. 2001. http://www.unesco.org/courier/2001_07/uk/planet2.htm. World Resources Institute. 2000 - 2001. People and ecosystems the fraying web of life. Ecotourism and conservation: are they compatible? Box 1.15

Table 1. Series of questions that were asked at key informant interviews. 1. How much income, per annum, is generated directly from the Outpost Lodge? 2. Is there a reserve account which monies go into? 3. What is the average mean income of residents in the Makuleke community? 4. What is the primary source of income generated from within KNP? 5. How much has the Makuleke accrued from the Makuleke Region since the land claim settlement in 1998? a. What projects have those funds been allocated towards/where will they go? 6. What reassurances of the conservation principles outlined in the Land Settlement Agreement has the Makuleke offered? 7. What is the population of the Makuleke community? 8. Has the CPA and the Makuleke community considered the options of the commercialization of resources within the Makuleke Region?

Table 2. Mean direct-use values of natural resources per household for three separate villages sampled. Numbers in brackets are per capita. Resources Finale Mabins Willows Mean % of total wild herbs 1328 (184.5) 2067 (287) 1238 (188) 1544 (219) 39 wild fruit 1047 (145.4) 1381 (192) 649 (98.30) 1026 (145) 25.9 fuelwood 472 (65.50) 580 (80.50) 709 (107) 587 (84.5) 14.8 edible insects 548 (76) 735 (102.) 412 (62.40) 565 (80.2) 14.3 reed mats 60.3 (8.4) 82.70 (11.50) 67.20 (10.20) 70.10 (10.) 1.8 fencing poles 38.5 (5.4) 50.20 (7.00) 53.60 (8.10) 47.40 (6.8) 1.2 bushmeat 10.4 (1.4) 34.60 (4.80) 60.70 (9.20) 35.20 (5.10) 0.9 wooden utensils 24.6 (3.4) 35.70 (5.00) 27.50 (4.20) 29.30 (4.20) 0.7 grass brooms 24.5 (3.4) 15.20 (2.10) 34.10 (5.20) 24.60 (3.60) 0.6 twig brooms 13.5 (1.9) 18.90 (2.60) 21.00 (3.20) 17.80 (2.60) 0.5 thatching grass 3.6 (0.5) 16.20 (2.30) 6.40 (1.00) 8.70 (1.20) 0.2 housing poles 5.8 (0.8) 2.90 (0.40) 2.90 (0.40) 3.90 (0.50) 0.1 Total 3576 (497) 5019 (697) 3280 (497) 3959 (564) 100

154 Structural defense of Acacias: hooks, spines and architecture

Category: Independent Project Participants: Gareth Hempson (secretary), Laurence Kruger (resource person) Site: Punda Maria, Kruger National Park, Mpumalanga Province, South Africa

Key words: Acacia, defense, feeding simulation, index, thorns

Abstract: African Acacias form attractive browse for many large mammals, necessitating their development of various structural defenses. Hooked thorns, straight spines and various branching architectures are used to defend against herbivory. The defenses of four Acacia species are quantified using a stripping index, biting index and branching index, which are combined in an overall defensive index. Three feeding simulations were also performed in an effort to quantify the effect of Acacia defenses on browsers in a more holistic manner. Results showed the biting index and clamp tests to be a good measure for the effectiveness of spines. The sock test, stripping index and branching index appear to be useful tools for assessing the defensive functioning of hooks, and the associated importance of architecture. This study provides a useful starting point for the development of a measure of the overall defense of Acacias, which would greatly enhance the understanding of the life history and distributional patterns observed in this genus.

Introduction African Acacias are typically fine-leaved savanna trees, generally more prevalent in regions with higher nutrient levels. This pattern is observed in their occurring towards the bottom of savanna catenal sequences, but also on a broader scale in their being more dominant in the dry, fertile savannas versus the less fertile mesic savanna systems (Scholes and Walker 1993). The often high nutrient status of Acacia species makes them attractive browse for many savanna herbivores, necessitating their development of various defenses to limit this herbivory pressure. Plants defenses can be grouped into three main forms: structural defenses, chemical defenses and phenological defenses (Scholes and Walker 1993). Acacias typically have weak chemical defenses (e.g. distasteful or toxic compounds), and rather rely on structural defenses in the form of thorns to reduce browsing pressure to tolerable levels. Phenological defense mechanisms will vary within and between species, and refer to the mechanisms plants use to reduce the time that their more vulnerable parts are exposed to the risk of herbivory (Scholes and Walker 1993). Acacias use thorns as a mechanical defense against large mammal herbivory. These thorns can be divided into two types, hooks and spines, which appear to be designed for different protective strategies (Midgley et al. 2001). Midgley et al. (2001) suggest that hooks may primarily be designed to limit leaf stripping, while spines act to prevent stems being bitten through. They also note that while hooked thorns are generally dark and inconspicuous, spines are usually white and very obvious, and suggest that this may make them strong visual deterrents to browsers. Previous attempts to quantify how well defended an Acacia individual is have included the use of a spinescence index (Midgley et al. 2001), which was calculated as thorn length (x100)/leaf length x inter-thorn distance. This index is designed to assess how well thorns protect leaves, but as argued by Midgley et al. (2001), this may not always be the primary function of thorns (e.g. spines), and also may not be an adequate quantification of the means by which hooks, for example, act in protecting leaves. A further technique whereby Acacias are able to increase the effectiveness of their mechanical defense is by increasing the degree of branch ramification. This can act to limit herbivory in three ways: bite size can be reduced if long shoots are not accessible (e.g. Cooper and Owen-Smith 1986 as cited in Archibald and Bond 2003), it can create a cage effect whereby younger shoots are protected by older shoots (e.g. Archibald and Bond 2003), and on the scale of a whole tree, a wide spreading architecture can limit herbivore access to the inside of the tree (e.g. Brown 1960 as cited in Archibald and Bond 2003). The degree of defense of Acacias varies considerably within and between species. Different selective pressures (e.g. fire, browsing and light availability) on A. karroo results in markedly different architectures and degrees of spinescence in different regions in its range (Archibald and Bond 2003). Individuals occurring in the more frequently burning mesic savannas were found to have the smallest spines and low degrees of branching, a response seemingly determined by the need for rapid growth to escape the fire trap (Bond and Van Wilgen 1996). Archibald and Bond (2003) however also pointed out that this rapid growth would help elevate branches out of the reach of browsers. 155 Variability in Acacia defenses has also been shown to be caused by nutrient availability (Gowda et al. 2002), herbivory (Young et al. 2003) and size (Brooks and Owen-Smith 1994). Gowda et al. (2002) observed a significant relative increase in the mass of long spines of A. tortilis in greenhouse experiments, as a result of an increase in nutrient availability. Natural and simulated herbivory resulted in A. drepanolobium having longer spines than when unaffected by herbivory (Young et al. 2003). This response was extremely localised, with only affected branches showing the induced response to herbivory. Ant pruning of leaves stimulated similar responses to that of large mammal herbivory. Brooks and Owen-Smith (1994) observed size class differences in spinescence in A. nilotica and A. tortilis. They concluded that juvenile A. nilotica individuals were better defended than adults, but could not explain the effect of longer hooks and a higher hook to spine ratio in juveniles compared to adults of A. tortilis. Acacia defense is thus governed by a range of factors. It is however possible to quantify how well defended a species is within a region, although the differing functions of hooks and spines, and the effect of the degree of branching have yet to be individually assessed. This study aims to quantify the effectiveness of mechanical defenses i.e. hooks, spines and branch ramification, as anti-large mammal herbivory defenses separately, but also to combine them in an overall defense index for different species. This should assist in understanding the positioning of Acacias in the landscape and also offer insight into their key life history traits (e.g. Archibald and Bond 2003).

Methods Study sites This study was carried out in the northern regions of the Kruger National Park, South Africa (figure 1), from the 24 - 31 March 2004. The area lies in the savanna biome, in the Lowveld regions of the Limpopo Province, South Africa. The average annual rainfall for the Punda Maria rest camp is 540mm, but this decreases further north, with approximately 430mm being recorded annually in the Pafuri region (Venter et al. 2003). This area would thus be referred to as a xeric savanna system (Bond 1997). Sampling of A. erubescens and A. tortilis was done in the Pafuri area, on the slopes and the valley bottom near the Thulamela heritage site. The A. erubescens individuals were all found on the fairly rocky slopes, while the A. tortilis specimens occurred on the valley bottom. A. robusta was sampled just outside the entrance gate to the Punda Maria rest camp, in a slight depression subject to seasonal flooding. Two individuals of A. nigrescens were sampled on the Flycatcher walking trail in the Punda Maria rest camp, and the remaining eight individuals 3km south on the southern section of the Mahonie loop road. Sampling methods Ten individuals of each of A. erubescens, A. nigrescens, A. robusta and A. tortilis were sampled. These four species were selected to represent different mechanical defense strategies. A. erubescens and A. nigrescens both have only paired hooked thorns. A. robusta has paired spines, and A. tortilis has both paired hooked thorns and paired spines. Measurements Three branches were cut from each individual at a branch diameter of 12mm. Thorn densities were calculated by counting the number of thorns on the last 50cm of the longest branch. Thorn measurements were made on the left hand thorn of the first five pairs of thorns moving outwards along the end 50cm of the longest branch. For A. tortilis, which has hooks and spines, five thorns of each type were measured. The following measurements were made on each thorn: length (thorn base to tip), gape width (perpendicular length from branch to thorn tip), angle of orientation (measured with 0˚ pointing back along branch to tree interior) and size of base (thorn attachment to branch). The length of ten mature leaves was measured on each branch sampled. The length of the longest branch was measured, and also the total length of branches in the sample. Feeding simulations Three feeding simulations were performed. The sock test involved throwing a tennis ball in a sock into each of the study individuals. This was then pulled out of the tree using a 10kg persola scale, and the effort to dislodge the sock recorded. This test is aimed at quantifying defense against a pulling motion out of the tree. This would simulate an animal pulling its head and ears or even whole body out from the low branches it was feeding on. On a smaller scale it also simulates leaf stripping. This test was performed ten times on each individual. The muzzle test involved pushing a wooden block (6cm x 6cm x 8.5cm) covered in plasticine in and out through the outer branches of each individual five times. The total number of scratches in the plasticine was recorded, as well as the number of scratches deemed to potentially be inflicting serious damage to the intruder. This test is designed to try and imitate a browser moving its muzzle in and out of the tree as it feeds. Not all contact with thorns is going to effectively deter a browser, but it is argued that more severe stabbing, gouging and shredding will impede its feeding. Scoring this test as a ratio of total number of scratches: serious scratches, relies on the thought that its not only about having lots of thorns, but that their design and presentation are also very important in determining their 156 effectiveness as physical defenses. This test was performed three times on each individual. The third feeding simulation, the clamp test, made use of a large pair of wooden scissor apparatus with 8.5cm x 7.5cm plates as jaws. These were covered with plasticine. The clamps were then used to simulate an animal biting five branches of the study individual. The number of thorn impressions and branch impressions in the plasticine was recorded. A high score can be attained from this simulation by two means. Firstly, a low number of branch impressions can be caused by the thorns preventing the jaws actually closing onto the branch, and secondly, a high thorn density increases the number of thorn impressions. This test was performed three times on each study individual. Analysis All analyses were carried out using Microsoft Excel and JMP IN 5.1 Statistical package. Measurements Thorn parameters (angle, base size, length and gape), thorn density and leaf length were tested for significant differences (p = 0.05) between species (and also against hooks and spines of A. tortilis separately where appropriate) using a Tukey-Kramer pairwise ANOVA comparison. Feeding simulations The mean score of each species for each of the feeding simulations was compared to the others for statistical significance (p = 0.05) using a Tukey-Kramer pairwise ANOVA comparison. For the sock test this took the form of a direct comparison between species of the mean effort required to dislodge the sock from each individual tree included in the study. The mean of the three muzzle test scores (ratio of the total number of scratches: serious scratches) for each individual was calculated and compared between species. Similarly, the mean of the clamp test scores (ratio of thorn impressions: branch impressions) was calculated and compared between species. Indices Three indices were used to attempt to quantify three anti-herbivory defenses of Acacias: to protect against leaf stripping (stripping index), branch biting (biting index), and also branch ramification to restrict access to the plant interior (branching index). A fourth index, the defense index, combined these three measures in an attempt to score the total defensive capacity of each species. The stripping index was calculated as: Stripping index = Density * (180 – Angle of orientation)/10 * Base This index is designed to assess defense against a pulling motion along a branch, directed out of the tree. A higher density of thorns is thought to increase defense effectiveness. The angle of thorn orientation is said to be most effective against leaf stripping when a thorn points directly back along a branch, and lowest when it points towards the tree exterior. Dividing the angle by ten was used weight the contribution of the index components more evenly. Base size is used to quantify the strength of attachment of thorns, with a larger base providing greater attachment strength. The biting index was defined as: Biting index = Density * Thorn length * (90 – Absolute (90 – Angle of orientation)) A higher thorn density is again thought to provide greater protection against herbivory. An increase in thorn length increases the range over which the thorn acts i.e. the biting action is resisted while the jaws are further from the branch. The contribution of the angle of thorn orientation to the index is maximised at 90˚, at which the greatest protection against a biting motion from above and below is offered. The method used to calculate the branching index is the similar to that used by Archibald and Bond (2003). The total length of branches in a sample is divided by the length of the longest branch. An alternative method for assessing the degree of branching was suggested during the study, whereby the length of the longest branch is divided by the direct length from the tip to the base of the longest branch. These two methods are likely to be differentially suited to capturing this property of Acacias under different circumstances. The index used by Archibald and Bond (2003) was adopted in this study. The stripping index, biting index and branching index were combined to form the defensive index in the following manner: Defense index = Branching index * (Stripping index/Mean stripping index + Biting index/Mean biting index) This formula aims to combine the effect of thorns against leaf stripping and branch biting, which are not necessarily mutually exclusive, and then to expand this by multiplying by the degree of branching, which would conceivably act to intensify the effect of thorns. The scores calculated for each of these four indices were compared across the four species using a Tukey- Kramer pairwise ANOVA comparison. The stripping index, biting index and defense index were also calculated 157 separately for the hooks and spines of A. tortilis, and were included in the statistical analysis.

Results Measurements In addition to the mean values for each species, the separate values obtained for the hooks and spines of A. tortilis are also included. All results the presented in this section are shown in table 1, where mean values that do not differ significantly at the 95% confidence level are grouped in boxes or ellipses. Significance difference was found in thorn densities between species (F5, 174 = 122.6222, p = .05). The thorn densities of A. erubescens (39), A. nigrescens (34.5) and A. robusta (38.2) were not significantly different (p =.05). A. tortilis had a significantly higher (p = .05) thorn density (56.4), despite its spine density being very low (4.3). There was a significant difference in thorn angles between species (F5, 970 = 322.6246, p = .05). The angles of the three hooked thorn groups did not differ significantly at the 95% confidence level (A. erubescens (32), A. nigrescens (29.7) and A. tortilis (hooks) (36.6)). The mean angles of the spines of A. robusta (107.4) and A. tortilis (spines) (91.8) differed significantly (p = .05), and were also both significantly higher (p = .05) than the hooked thorns. The size of the basal attachment of thorns differed significantly between all species (F5, 970 = 264.4610, p = .05). The two largest based groups were the hooked thorns of A. nigrescens (5.3) and A. erubescens (4.2). A. robusta (3.7) had the third largest mean thorn base. The mean base sizes of the hooks (1.9) and the spines (2.8) of A. tortilis were both smaller than those measured for the other species. Gape and thorn length measurements showed significant difference between species (F5, 970 = 142.5644, p = .05 and F5, 970 = 138.5427, p = .05), and showed the same trends. The hooked groups showed no significant difference (p =.05) in mean length or gape measurements, but were significantly lower (p = .05) than the spine measurements. Of the hooked groups A. nigrescens (5.3 and 5.3) had the longest thorns with the largest gapes, followed by A. erubescens (3.9 and 2.9) and then A. tortilis (hooks) (3.5 and 1.9). The spine length and gape measurements of A. tortilis (45.6 and 2.8) were significantly greater (p = .05) than for A. robusta (21.6 and 2.3). Leaf lengths differed significantly between species (F5, 1376 = 765.3441, p = .05). The mean leaf lengths of A. robusta (66.7) and A. nigrescens (66.9) did not differ significantly (p = 0.05), but were significantly larger than those of both A. erubescens and A. tortilis. The leaves of A. tortilis were significantly smaller (26.3) than those of A. erubescens (46.0). Feeding simulations Significant differences were observed between sock test (figure 2) scores of each species (F3, 396 = 69.1912, p = .05). A. nigrescens (6.6) and A. tortilis (6.4) scored significantly higher (p = .05) in the sock test than either A. erubescens or A. robusta, but did not differ significantly (p = .05) from each other. A. robusta had the lowest score for the sock test (2.5), which differed significantly (p = .05) from that of A. erubescens (4.5). Muzzle test (figure 3) scores differed significantly between species (F3, 116 = 10.9700, p = .05). A. nigrescens (1.8), A. robusta (2.1) and A. tortilis (2.6) did not differ significantly (p = .05) in their scores obtained for the muzzle test. These three scores were all however significantly lower (p = .05) than the score calculated for A. erubescens (3.6). There was significant difference in the clamp test (figure 4) scores between species (F3, 116 = 14.7500, p = .05). The clamp test produced two significantly different (p = .05) groups of scores. A. robusta (7.1) and A. tortilis (6.8) scored significantly higher (p = .05) than A. nigrescens (2.8) and A. erubescens (2.7).

Indices The results of the stripping index comparison (figure 5) show all four species to have significantly different (F5, 965 = 217.3387, p = .05) means. The comparison between A. tortilis (1476.7) and A. tortilis (hooks) (1408.5) was however not statistically significant (p = .05). A. nigrescens had the highest mean stripping index score (2799.4), followed by A. erubescens (2494.7), A. tortilis (1476.7) and finally A. robusta (1030.2). The spines of A. tortilis, when scored separately, only had a stripping index score of 148.9. The results of the biting index (figure 6) comparison were significantly different between species (F5, 965 = 75.9343, p = .05). Two significantly different (p = .05) sets were separated out by the test. A. tortilis (92928.2) and A. robusta (85013.8) both had high mean biting index scores. The four other groups in the comparison all had significantly lower (p = .05) mean scores: A. tortilis (spines) (22941.6), A. tortilis (hooks) (6435.9), A. nigrescens (5433.2) and A. erubescens (4477.3). The branching index (figure 7) results differed significantly between species (F3, 116 = 35.6404, p = .05). There was no significant difference (p = .05) between the mean index scores of A. nigrescens (7.0), A. erubescens 158 (6.9) and A. tortilis (6.6). A. robusta had a mean branching index score of 2.4, which was significantly lower (p = .05) than any of the three other mean scores. The significance relations between species (F5, 965 = 65.7113, p = .05) in the defense index comparison was rather complex. The spread of values for each group is shown in figure 8, and the significance interactions in table 2. A. tortilis (20.6) scored significantly higher than any other group in this comparison. A. nigrescens (11.0) had the second highest mean score, followed by A. erubescens (9.5) and A. robusta (6.3). The two lowest mean defense index scores were of the hooks (1.0) and spines (3.5) of A. tortilis when assessed separately.

Discussion Despite being subject to variability arising from a range of factors (e.g. Archibald and Bond 2003, Gowda et al. 2002), patterns are never the less evident in Acacia defenses. Similar life history strategies are employed by different groups in order to overcome the same obstacles posed by an environment. Mechanical defenses are designed as anti-herbivory devices, and this study shows how different patterns in the design of these defenses afford Acacias protection against different aspects of browsing. The results suggest that there is a strong distinction between the functioning of a hooked defense and a spiny defense. Spines The most apparent pattern in the data is perhaps the separation of the two hooked species and the two species with spines into significantly different groups by the biting index and the clamp feeding simulation. In both situations the hooked species scored much lower than the species with spines. This strongly suggests that the one of the primary functions of spines is to act against a biting motion, while hooks offer little protection against biting. Midgley et al. (2001) argue that spines also act as a visual defense against herbivory. This suggestion makes good sense in that once a certain spine length is exceeded, they become too long to fit into a herbivore’s gape, and would thus no longer seem to function as anti-biting defenses. A further suggestion to that of spines acting as a visual defense is that long spines may function defensively in a similar way to a branch ramification defense. The long spines effectively form a sharp pointed cage, which limits mouth access and threatens eyes and lips. These spines would conceivably offer very little nutritional value to browsers, making browsing a way into a spine-based cage defense less rewarding than browsing into a branch-based cage defense. Hooks The two hooks only species scored much higher in the stripping index than did the spines only and hooks and spines species. This suggests that the role of hooks, as suggested by Midgley et al. (2001) may well primarily be anti-leaf stripping. A. tortilis has a higher hook density than either of the two hooks only species, but still scores lower than them in the stripping index. This apparent incongruity can possibly be explained by looking at the thorn design measurements. The forces of a stripping motion combine to pull a thorn off a branch. The strength of attachment is thus an important design parameter in this regard. Two factors are likely to be largely responsible for increasing the strength of attachment of thorns, the size of the basal attachment area and the size of the gape. A larger gape offers greater leverage for breaking the thorn off. It could thus be argued that the ratio of basal area to gape is a measure of the strength of attachment. The base area of the two hook only species is significantly larger (almost double) than that of A. tortilis hooks. Secondly, the gapes of the hook only species are both smaller than their bases, while the hook gape of A. tortilis is greater than the hook base. The fact that A. tortilis scored much lower in the stripping index than either of the two hook only species, despite the fact of its significantly higher hook density, is as a result of its poorly designed hooks. Overall defense The defense index suggests that A. tortilis is the most well defended of the four species included in this study. This result, on face value, may not appear unexpected, in that it has both paired hooks and paired spines. What is interesting though is that when these two components of the total thorn defense are analysed separately, they both score very low. The whole of the defense is much greater than the sum of its parts. The key to the defense of A. tortilis, whose hooks do not score particularly well on the stripping index and whose spines don’t perform exceptionally in the biting index, is most likely its very high thorn density. Midgley et al. (2001) found A. tortilis to be the most well defended of their 17 study species based on their spinescence index. The two paired hook species, A. nigrescens and A. erubescens, scored similarly in the defense index. Both of these species scored well in the stripping index, but poorly in the biting index. However, their degree of branch ramification was high. Based on personal observation of these two species in the field, it did come as some surprise that A. nigrescens did not outperform A. erubescens to a greater degree in the defense index. This could be as a result of a lack of sensitivity in the index designs, particularly the stripping index. It must however be noted that interactions with the study species during data collection did not necessarily mimic natural animal feeding activities around these 159 trees, and that the indices may in fact be sound. What can be taken from this study though is that these two species seem to rely quite heavily on their branching structure for protection. They both do have seemingly well designed anti-stripping hooked thorns, but the presentation of these i.e. architecture, is also very important in deterring herbivores. A. robusta scored the lowest in the defense index. This is due to its very low branching index score, and also its low stripping index score. While working with this species in the field was less harrowing than the other three species, a parameter of its defense may have been overlooked. Despite the mean length of A. robusta spines being lower than those of A. tortilis, the longest individual thorns measured in this study belonged to A. robusta. As mentioned above, these long spines may act to form a spine-based cage defense. This was not quantified in this study and would improve the defensive rating of A. robusta, mainly by increasing the contribution of its ‘branching’ index. Predictions Using the defense index as scored in this study, the following predictions could be made about the distribution of the study species in a landscape. A. tortilis is well defended against herbivory, and would thus be expected to be found in nutrient rich regions such as the lower end of the catenal profile, or on a larger scale to be more prevalent in xeric, more nutrient rich savanna systems. The least well defended A. robusta would be expected to occur in areas with low herbivory levels, or alternatively in regions where herbivory is not a main ecosystem driver. Phenological defenses, e.g. growing tall rapidly, could offer A. robusta an alternative to structural defenses. The differences in defenses of the two hooked thorn species may explain the apparently less well defended A. erubescens occurring on less accessible rocky slopes, while A. nigrescens occurs widely in the savanna landscape. However, differences in physiological adaptations could form an alternative to this suggestion, particularly in the light of the fact that the difference in their defense index scores was only slight. Future work Much work remains to be added to this study. An important aspect requiring attention is the refinement of the methodology, in particular in terms of index design. It has been suggested that additive indices may prove a more fair comparison than the current multiplicative indices (L. Kruger personal communication). Increasing the suite of study species is in important future addition. The selection of species is intended to allow more insight into differences between the roles of hooks and spines, but also to begin to attempt to understand the effect of various gradients in landscapes. The difference in Acacia defenses between mesic and xeric savannas warrants investigation, and also between more and less accessible growing sites. Increasing the number of individuals per species should show intra-specific variation in defenses (e.g. Archibald and Bond 2003). This would be expected to occur on scales from between biomes (Archibald and Bond 2003) to between size classes in a locality (Brooks and Owen-Smith 1994). Field observation of mammal herbivory and thorn clipping selection experiments would further enhance the study. The examination of hooked versus spiny Acacias for general characteristics is important. The existence of such characters would prove very useful in understanding the evolutionary history of Acacias. Conclusions This study has given good insight into the differing defensive functioning of hooks and spines. It has also emphasised the role of branching architectures in structural defenses, seemingly most importantly in a hooked thorn defense. The clamp test biting simulation gave a good measure of the effectiveness of a thorn defense against biting, with the results correlating well with those of the biting index. The sock test gave good insight into the way in which architecture and thorns combine to form a herbivory defense. It is hoped that this study provides a useful starting point for further investigation in the field of Acacia defenses. The indices are aimed at trying to facilitate the quantification of aspects of this defense, assisting in comparisons between studies. It is also hoped that innovations such as the feeding simulations will stimulate fresh, holistic thought on how Acacias protect themselves from herbivores.

Acknowledgements Many thanks to Laurence Kruger for his help throughout the course of this project - from the initial concept, assistance in the field and for the discussions and comments on the data during and after analysis. Also a big thank you to Scott Briscoe, Jasper Slingsby, Julie Coetzee, Eric Caldera, Godfrey Sekhula and Wachi Masinga for their assistance during data collection.

Literature cited Archibald, S. and W. J. Bond. 2003. Growing tall vs growing wide: tree architecture and allometry of Acacia karroo in forest, savanna, and arid environments. Oikos 101: 1-12 Bond, W. J. and B. W. van Wilgen. 1996. Fire and plants. Chapman and Hall, London. 263pp. 160 Bond, W. J. 1997. Chapter 18: Fire in Vegetation of Southern Africa. Eds: R. M. Cowling, D. M. Richardson and S. M. Pierce. Cambridge University Press, Cambridge. 615pp. Brooks, R. and N. Owen-Smith. 1994. Plant defenses against mammalian herbivores: are juvenile Acacia more heavily defended than mature trees? Bothalia 24: 211-215 Brown, W. 1960. Ants, acacias and browsing animals. Ecology 41: 587-592 Cooper, S. M. and N. Owen-Smith. 1986. Effects of plant spinescence on large mammalian herbivores. Oecologia 64: 446-455 Gowda, J. H., B. R. Albrectson, J. P. Ball, M. Sjoeberg and R. T. Palo. 2002. Spines as a mechanical defense: the effects of fertilizer treatment on juvenile Acacia tortilis plants. Acta Oecologica 24: 1-4 Midgley, J. J., M. A. Botha and D. Balfour. 2001. Patterns of thorn length, density, type and colour in African Acacias. African Journal of Range and Forage Science 18: 59-61 Scholes, R. J. and B. H. Walker. 1993. An African Savanna synthesis of the Nylsvlei study. Cambridge University Press, Cambridge. 306pp. Venter, F. J., R. J. Scholes and H. C. Eckhardt. 2003. Chapter 5: The Abiotic Template and Its Associated Vegetation Pattern in The Kruger Experience. Eds: J. T. du Toit, K. H. Rogers and H. C. Biggs. Island Press, Washington. 519pp. Young, T. P., M. L. Stanton and C. E. Christian. 2003. Effects of natural and simulated herbivory on spine lengths of Acacia drepanolobium in Kenya. Oikos 101: 171-179

161 Table 1: Comparison of the mean values of a range of measurements taken on the study species’. Values joined by a box or ellipse do not differ significantly (p = 0.05) from one another when compared using a Tukey-Kramer pairwise ANOVA comparison.

Table 2: Mean defense index scores for the six groups analysed in this study. Rows joined by the same letter do not differ significantly from one another at the 95% confidence level when compare using a Tukey-Kramer pairwise ANOVA comparison.

Level Mean A. tortilis A 20.6 A. nigrescens B 11.0 A. erubescens BC 9.5 A. robusta CD 6.3 A. tortilis (spines) DE 3.5 A. tortilis (hooks) E1.0

Figure 1: Location of study sites in the northern Kruger National Park, South Africa. 162

Figure 2: Comparison of the mean (horizontal bar) and spread of values obtained with the sock test. Circles that intersect at an angle greater than 90˚ indicate data groups that do no differ significantly at the 95% confidence level in a Tukey-Kramer pairwise ANOVA comparison.

Figure 3: Comparison of the mean (horizontal bar) and spread of values obtained with the muzzle test. Circles that intersect at an angle greater than 90˚ indicate data groups that do no differ significantly at the 95% confidence level in a Tukey-Kramer pairwise ANOVA comparison.

Figure 4: Comparison of the mean (horizontal bar) and spread of values obtained with the clamp test. Circles that intersect at an angle greater than 90˚ indicate data groups that do no differ significantly at the 95% confidence level in a Tukey-Kramer pairwise ANOVA comparison. 163

Figure 5: Comparison of the mean (horizontal bar) and spread of values obtained with the stripping index. Circles that intersect at an angle greater than 90˚ indicate data groups that do no differ significantly at the 95% confidence level in a Tukey-Kramer pairwise ANOVA comparison.

Figure 6: Comparison of the mean (horizontal bar) and spread of values obtained with the biting index. Circles that intersect at an angle greater than 90˚ indicate data groups that do no differ significantly at the 95% confidence level in a Tukey-Kramer pairwise ANOVA comparison.

Figure 7: Comparison of the mean (horizontal bar) and spread of values obtained with the branching index. Circles that intersect at an angle greater than 90˚ indicate data groups that do no differ significantly at the 95% confidence level in a Tukey-Kramer pairwise ANOVA comparison. 164

Figure 8: Comparison of the mean (horizontal bar) and spread of values obtained with the defense index. Circles that intersect at an angle greater than 90˚ indicate data groups that do no differ significantly at the 95% confidence level in a Tukey-Kramer pairwise ANOVA comparison.

165 An investigation into the fine-scaled variation in tree diversity and the varied architecture of Colophospermum mopane in the Mopaniveld

Category: Independent Project Participants: Benjamin Wigley and Jasper Slingsby (secretaries), Laurence Kruger (editor) Site: Punda Maria, Kruger National Park, Mpumalanga Province, South Africa

Key words: abiotic factors, competition, Colophospermum mopane, disturbance, Mopaniveld, species richness, diversity

Abstract: In this study we investigate the drivers behind variation in species richness and diversity at a fine scale in the Mopaniveld (i.e. what causes monospecific vs. multispecies stands?) and the drivers behind variations in stand and individual C. mopane architectures (height, basal area, multistemmedness). We hypothesize that the basic compositional structure of the plant communities is determined by physical and edaphic soil characteristics. In the absence of disturbance, competition should determine densities and stand architectures of C. mopane given the abiotic template. In highly disturbed savanna environments the effects of competition are most likely less obvious, and densities and stand architectures of C. mopane would be determined by its response to the disturbance regime. Our data indicated the presence of catenal sequences. Plots with sandy soils were located at higher elevations and had lower conductivity, calcium and carbon concentrations than plots with clay soils (U=57, P=.022; U=45, P=.01; U = 56, P = .02; U=35, P<.005). Variation in species richness and diversity at a fine scale in this part of the Mopaniveld appears to be determined by catenal sequences as species richness and diversity was higher in plots with sandier soils (U=52.5, P<.05 and U=62, P=.05 respectively). C. mopane appears to grow in higher densities on clay soils (U=66, P<.055), but grows larger on sandy soils (Beta=0.652, P<.05 and Beta =0.553, P<.01 respectively) with high calcium concentrations (Beta=0.487, P<.05 and Beta =0.463, P<.05 respectively). Variations in disturbance regimes and competition at the catenal scale could be important in determining stand and individual C. mopane architecture.

Introduction The Mopaniveld is an African savanna vegetation type so named because of the dominance of Colophospermum mopane (J. Kirk ex J. Léonard), a tree species of the Caesalpinaceae in the Fabaceae (Palgrave 1983). Large scale studies have been performed relating climate and rainfall to the distribution and variation within community structure and composition of the Mopaniveld (Sebego et al. 1999, Ringrose et al. 2003). The dominant vegetation patterns in Kruger closely follow variation in climate, geology and soils (Venter et al. 2003). C. mopane dominates in areas with a mean annual rainfall of less than 500mm, and soils with greater than 15% clay content (Venter et al. 2003). These patterns may be true at the landscape scale, but there seems to be far more variation in Mopaniveld at a finer community scale (pers. obs.). The three main drivers of this variation in vegetation types and community compositional structure are thought to be the abiotic template (Venter et al. 2003), disturbance (Scholes et al. 2003) and competition (Tilman 1997). For plants the most important aspects of the abiotic template are determined by soil characteristics. Soils influence plants through influence on water availability, provision of nutrients needed for growth and metabolic processes, and the physical or chemical conditions of soils may inhibit the penetration of plant roots and the volume of soils they use. These characteristics greatly influence the species composition and structure of plant communities (Venter et al. 2003). In the Kruger National Park the two main disturbance factors are fire (van Wilgen et al. 2003, Scholes et al. 2003), and elephant damage of trees (Whyte et al. 2003, Scholes et al. 2003). Fires in African savannas damage small trees that are too short to escape the “fire trap”, causing them to either resprout (Bond and van Wilgen 1996, Higgins et al. 2000) or perish. Kennedy and Potgieter (2003) studied the effects of fire on the size and architecture of C. mopane in long term experimental burn plots in the KNP. They found trees in burnt plots to have aggregated distributions and significant differences in a range of morphological parameters including lower tree height, canopy diameter, mean stem circumference and a higher number of stems per individual than no burn plots. They also found burnt plots had a higher proportion of coppiced stems than no burn plots. Thus the effects of fire have been seen to have a major 166 impact on the size and architecture of C. mopane, which in turn may have impacts on the community structure. We hope to investigate this further. It is commonly observed that elephant feeding is highly concentrated in the savanna landscape which results in a patchwork landscape with stands in different stages of recovery from intensive use by elephants (Ben-Shahar and Macdonald 2002, Scholes et al. 2003). C. mopane is utilized by elephants in two main ways. Trees are either pushed over (Scholes et al. 2003) or have branches stripped of the main stem (Halzka Hraber pers comm.). When pushed over, C. mopane resprouts basally which reduces the height profile of the canopy and in extreme cases may even cause bushlands to change into shrublands (Scholes et al. 2003). When branches are stripped by elephants the tree responds through epicormic resprouting (canopy resprouting) which doesn’t appear to significantly alter the stand structure. The third major ecological driver of community structure is competition. In areas with a favorable abiotic template, and little disturbance, competition becomes the most important factor driving community processes (Tilman 1997). The dominance of C. mopane in vast areas of Southern Africa suggests it to be a superior inter-specific competitor, or a better tolerator of poor conditions. In these areas where C. mopane dominates and disturbance is low, intra-specific competition is likely to result in very dense stands of single stemmed trees with low basal area as a result of self thinning (Smit 2001). In this study we investigate the drivers behind variation in species richness and diversity at a fine scale in the Mopaniveld (i.e. what causes monospecific vs. multispecies stands?) and the drivers behind variations in stand and individual C. mopane architectures (height, basal area, multistemmedness). We hypothesize that the basic compositional structure of the plant communities is determined by physical and edaphic soil characteristics. In the absence of disturbance, competition should determine densities and stand architectures of C. mopane given the abiotic template. In highly disturbed savanna environments the effects of competition are most likely less obvious, and densities and stand architectures of C. mopane would be determined by its response to the disturbance regime.

Methods Study site The study site we selected occurs in the Bulweni land system near Punda Maria rest camp in the northern Kruger National Park, South Africa. It occurs on soils derived from the Karoo sediments, Ecca shale and mudstone. The area receives a mean annual rainfall of 450-500mm, making it one of the driest areas in the park (Venter et al. 2003). Data collection We sampled thirty 400m2 plots, one every c. 250m along a 7.25km transect (latitudes and longitudes given in Appendix 1). All trees were identified, their height and diameter at 50cm measured (D50cm), and the total number of stems counted. The number of stems with branches damaged by mechanical damage were counted (this was assumed to be as a result of elephants), and the presence of dead, charred stems was noted, as an indication of fire disturbance. In each plot a 50cm deep hole was dug in order to determine if a calcrete layer was present and at what depth. A soil sample was taken from the centre of each plot. This was used to determine pH, conductivity and the concentrations of available phosphorous, calcium, nitrogen and carbon for each site. Soil texture was established using the field assessment method described in Tongway and Hindley (1995). Soil pH was determined using calcium chloride and a 1:2.5 ratio of soil to salt solution according to the method outlined by Rowell (1994). Conductivity was determined by using a 1:5 ratio of soil to distilled water. The solution was shaken for 15 minutes using a soil shaker before conductivity was measured using a conductivity meter. Data analysis The plots were divided into two categories for soil texture and Shannon Weiner diversity. Soil texture was divided into sandy and clay soils. Clay to sandy clay loam (Tongway and Hindley 1995) were classified as clay soils. Sandy loam, loamy sand and sand (Tongway and Hindley 1995) were classified as sandy soils. Shannon Weiner diversity categories were set as those plots with indices less than one versus those plots with indices greater than one. These categories were used to compare various biotic and abiotic properties using Mann-Whitney U tests. We performed a number of correlation and multiple regression analyses testing for significant relationships between the main vegetation components and physical and chemical components.

167 Results Soil texture The difference in soil texture between the clay and sand soil type categories was significant (U=0, P<.000, Figure 1). The mean particle size on clay soil types was 4.5 (Figure 1) corresponding to a mix of clay loams and sandy clays (Tongway and Hindley 1995). The mean particle size of the sandy soil types was 11.5 (Figure 1) corresponding to a mix of sands and loamy sands (Tongway and Hindley 1995). There appeared to be more C. mopane individuals in the plots with clay soils, but this difference was not significant (U=66, P<.055, Figure 2). The differences in mean C. mopane numbers were seen to be substantially different on the two soil types with a mean of 45 on sandy soils and a mean of 100 on clay soils (Figure 2). Multiple regression analysis showed a negative relationship between soil particle size and number of C. mopane individuals (Table 2). Species richness and Shannon Weiner diversity indices were significantly higher in plots with sandy soils (U=52.5, P<.05, Figure 3 and U=62, P=.05, Figure 4 respectively). Species richness was significantly positively correlated with increased soil particle size (R2=0.18, F=6.24, P<.05). Conductivity readings, carbon content and calcium concentrations were significantly higher in the plots with clay soils (U=45, P=.01, Figure 5 and, U=56, P=.02, Figure 6, U=35, P<.005, Figure 7 respectively). Nitrogen and phosphorus showed similar trends, but the differences between the two soil types were not significant. The plots that had sandy soils were found to be at significantly higher elevations (U=57, P=.05, Figure 8). Shannon Weiner diversity There was a significant difference in diversity between the two diversity categories (U=0.00, P<.000, Figure 9), validating the relevance of the division. That the means for the two categories were 0.5 and 1.5 respectively further supported the relevance of the division. The number of C. mopane individuals was significantly higher in the plots with diversity indices of less than one (U=41, P<.005, Figure 10). The number of individuals of all species other than C. mopane was higher in plots with higher diversity indices (U=16.5, P<.000, Figure 11). Table 1 shows there to be a significant negative correlation between the basal area of all species excluding C. mopane and calcium concentration (R =.16, F=5.32 and P=.029). A significant negative relationship was found between total number of plants and soil texture (R2 =0.17, F=5.81 and P=.023). This shows that there were more plants found in plots with smaller particle size i.e. clay soils. This was most likely as a result of the high numbers of C. mopane found on the clays as shown in Figure 2 and Table 2. A significant positive relationship was found between species richness and soil texture (R2=.18, F=6.24 and P=.019). Thus species richness was higher on sandy soils. Despite the higher number of C. mopane individuals in plots with clay soils, total basal area, mean basal area per individual, and mean height of individuals increased with increased soil particle size (Table 2). They were also found to increase with increased calcium concentrations, and total basal area increased with increased phosphorus concentrations too (Table 2). The number of C. mopane individuals resprouting from a burnt stem increased with increased phosphorus concentrations (Table 2). Shannon Weiner diversity and Pielou’s evenness both increased with increased soil particle size (Table 2). The summary statistics for the multiple regression models are shown in appendix 2.

Discussion On an apparently level landscape subtle changes in elevation appeared to be driving major changes in the soil characteristics. Plots with sandy soils were located at higher elevations and had lower conductivity, calcium, carbon, nitrogen and phosphorus concentrations than plots with clay soils indicating the presence of catenal sequences. Variation in species richness and diversity was determined by catenal sequences. Species richness and diversity was positively correlated with increased soil particle size. There were more C. mopane individuals in the plots that had clay soils, in accordance with Scholes (1997) and Venter et al. (2003). The dominant number of C. mopane individuals may have resulted in the exclusion of some species, and a consequent reduction in species richness and diversity. The decline in the number of C. mopane individuals with increased soil particle size is most likely related to the effects particle size has on the physical and chemical properties of the soil. Conductivity, carbon and calcium concentrations were found to be significantly higher in plots with clay soils, while not significant phosphorus and nitrogen concentrations showed a similar trend. 168

The mean basal area, total basal area and mean height of C. mopane individuals was greater in areas with higher calcium concentrations and larger soil particle sizes. Calcium concentrations, however, were found to be significantly higher in plots with clay soils. This would imply that C. mopane prefers areas with high calcium concentrations, but has difficulty growing large in soils with high clay content. The reasons for this are unclear and require further investigation. It is possible that the calcrete layer limits rooting depth in the clay areas and thus limits plant height. It is also possible that fewer C. mopane individuals establish on sandier soils, but those that do establish experience less intra-specific competition and, should sufficient calcium be present, are able to grow larger. This would represent a trade off between favourable abiotic conditions and competition. A possible factor limiting recruitment in these areas could be a lack of available water for seedlings in the upper soil layer due to the higher porosity of sandier soils. Unfortunately the effects of fire and elephants on the plots was difficult to quantify meaningfully from evidence in the field, and the records of these disturbance factors were not detailed enough to be useful at a scale as fine as this study. Previous studies have suggested that the variation in C. mopane individual architecture could also be related to elephant and fire disturbance (Scholes 1997, Ben-Shahar and Macdonald 2002, Kennedy and Potgieter 2003). Variation in these disturbance factors at a catenal scale could have important implications. Differences in fire frequency and/or intensity along catenal sequences could result in variations in the number of trees that escape the fire trap and individual architectures such as height, basal area, and multistemmedness. Fire intensity is known to be determined by grass biomass which is limited by water availability, nutrient availability (Higgins et al. 2000) and grazing pressure (Glynn Alard unpublished data). These factors are known to vary along catenal gradients (Venter et al. 2003). It is thus likely that there is variation in fire intensity along catenal sequences. For example our data showed that the number of C. mopane individuals resprouting from a burnt pole was higher in plots with higher phosphorus concentrations. The higher phosphorus concentrations could have resulted in greater grass biomass (Landon 1991) which in turn would have resulted in more intense fires. Elephant feeding has been observed to be highly concentrated in the savanna landscape resulting in a patchwork landscape with stands in different stages of recovery from intensive use by elephants (Ben-Shahar and Macdonald 2002, Scholes et al. 2003). Elephant disturbance tends to result in stands of short multistemmed C. mopane (Scholes et al. 2003) such as those seen on the clay soils. The lack of elephant disturbance would allow C. mopane trees to get taller, such as those seen on sandy soils with higher calcium concentrations. Once such a pattern developed it is possible that elephants would maintain the pattern through a possible preference to forage in the shorter stands because leaves are more abundant and easier to access. The higher densities of C. mopane individuals on the clay soils could have initiated the development of the pattern as forage in these areas would have been more abundant. The third major factor that could be affecting the density and stand architecture of C. mopane is intraspecific competition. Currently very little is known about the competitive interactions occurring in C. mopane dominated savannas (Smit 2001). Due to the naturally high occurrence of disturbance in KNP in the form of fire and elephant damage, the effects of competition are most likely to be fairly low. Furthermore it would be extremely difficult to quantify the effects of competition in a short-term study of this nature. Smit (2001) performed a study in order to determine the influence of tree thinning on the growth of C. mopane. He found that thinning reduced inter- tree competition which resulted in significant increases in the growth of the remaining trees and that with increased intensity of thinning the magnitude of growth in the remaining plants increased. This could explain why the C. mopane individuals that occurred on clay soils in higher densities were significantly shorter and had smaller basal areas than the individuals growing on sandy soils at lower densities. Future investigations into the variation in dominance and stand architecture of C. mopane should study each of the major influences in isolation by controlling for the other possible drivers. Experimental burn plots and elephant exclusion plots could be useful in this respect. It is also advised that longer term studies be undertaken such that the effects of competition, fire and elephant disturbance can be properly quantified. This could be achieved be the installation of long term monitoring plots. Soil properties such as water availability, variations in soil horizons and depth of the calcrete layer should also be taken into account. Variation in species richness and diversity at a fine scale in this part of the Mopaniveld is determined by catenal sequences. C. mopane grows in higher densities on clay soils, but grows larger, and comprises larger stand basal area on sandy soils with high calcium concentrations. Variations in disturbance regimes and competition at the catenal scale could be important in determining stand and individual C. mopane architecture.

169 Acknowledgements Special thanks to Laurence Kruger for correctly predicting that we would struggle to perform this study, but sticking with us anyway. Thanks to Godfrey Sekhula and Alione Ndlopfu for defending us from the forces of nature while we were performing our fieldwork. Thanks to the Organization for Tropical Studies for making this study possible.

Literature cited Ben-Shahar, R. and D. W. Macdonald. 2002. The role of soil factors and leaf protein in the utilization of mopane plants by elephants in northern Botswana. BMC ecology 2(3). Bond, W. J. and B. W. van Wilgen. 1996. Fire and plants: population and community biology series 14. Chapman and Hall, London. Crawley, M.J. 1997. Life History and Environment. Pages 73-131 in M.J. Crawley, editor. Plant Ecology. Oxford: Blackwell Science. Higgins, S.I., Bond, W.J. and Trollope, W.S. 2000. Fire, resprouting and variability: a recipe for grass-tree coexistence in savanna. Journal of Ecology 88:213-229. Kennedy, A. D. and A. L. F. Potgieter. 2003. Fire season affects size and architecture of Colophospermum mopane in Southern African savannas. Plant Ecology 167: 179-192. Landon, J. R. 1991. Booker tropical soil manual. Longman scientific and technical, London. Palgrave, K. C. 1983. Trees of Southern Africa. C. Struik Publishers, Cape Town. Ringrose, S., W. Matheson, P. Wolski and P. Huntsman-Mapila. 2003. Vegetation cover trends along the Botswana Kalahari transect. Journal of Arid Environments 54: 297-317 Rowell, D. L. 1994. Soil acidity and alkalinity. Pages 159-161 in Soil Science: Methods and Applications. Scholes, R. J. 1997. Savanna. Pages 258-277 in R. M. Cowling, D. M. Richardson and S. M. Pierce, editors. Vegetation of South Africa. Cambridge University Press, Cambridge. Scholes, R. J., Bond, W. J. and Eckhardt, H. C. 2003. Vegetation dynamics in the Kruger ecosystem. Pages 242-262 in J. T. Du Toit, K. H. Rogers and H. C. Biggs, editors. The Kruger Experience. Washington: Island Press. Sebego, R. J. G., W. Arnberg and S. Ringrose. 1999. Relation between cold cloud data, NDVI and mopane in eastern Botswana. The International Archives of the Photogrammetry, Remote Sensing and Spatial Information Sciences. 34(6): 170-182. Smit, G. N. 2001. The influence of tree thinning on the vegetative growth and browse production of Colophospermum mopane. South African Journal of Wildlife Research 31(3&4): 99-114. Tongway, D. and N. Hindley. 1995. Manual for soil condition assessment of tropical grasslands. Division of Wildlife and Ecology. CSIRO Australia, Canberra. Tilman, D. Mechanisms of plant competition. Pages 239-261 72 in M.J. Crawley, editor. Plant Ecology. Oxford: Blackwell Science. Venter, F.J., Scholes, R.J. and Eckhardt, H.C. 2003. The abiotic template and its associated vegetation pattern. Pages 83-129 in J. T. Du Toit, K. H. Rogers and H. C. Biggs, editors. The Kruger Experience. Washington: Island Press. Whyte, I. J., R. van Aarde and S. L. Pimm. 2003. Krugers elephant population: Its size consequences for ecosystem heterogeneity. Pages 332-348 in T. Du Toit, K. H. Rogers and H. C. Biggs, editors.The Kruger Experience. Washington: Island Press.

Table 1: Regression relationships of vegetation components against calcium and soil texture Vegetation component Soil property r2 F Ratio P Relationship Basal area excluding C. mopane Calcium concentration 0.16 5.32 .029 negative Total number of plants Soil texture 0.17 5.81 .023 negative Species richness Soil texture 0.18 6.24 .019 positive

170

Table 2: The importance of each soil factor in each multiple regression model Conductivity (uS/cm) Soil texture Ca (mg/kg) P (mg/kg) Beta P Beta P Beta P Beta P - # C. mopane individuals NS 0.456 * NS NS Total C. mopane basal area NS 0.405 * 0.733 *** 0.363 * Mean C. mopane basal area per individual NS 0.652 ** 0.487 * NS Mean C. mopane height NS 0.553 * 0.463 * NS # C. mopane individuals resprouting from a burnt stem NS NS NS 0.549 ** Shannon Weiner Diversity (H') NS 0.447 * NS NS Evenness (J’) NS 0.457 * NS NS *=<.05, **=<.01, ***=<.001

Soil texture rating Soil texture 6

3 Clays Sands Soil texture

Figure 1: The difference in soil texture ratings between the plots considered having clay soils and the plots considered to have sandy soils. A Mann-Whitney U test found the difference to be significant (U=0, P<.0001).

171 130

120

110

100

90 individuals 80

70 C. mopane C.

# of # of 60

30 Clays Sands Soil texture

Figure 2: The difference in the number of C. mopane individuals between plots with clay versus sandy soil types. A Mann-Whitney U test did not find the difference to be significant (U=66, P=.055).

6 # Species

0 Clays Sands Soil texture

Figure 3: The difference species richness between plots with clay versus sandy soil types. A Mann-Whitney U test found the difference to be significant (U=52.5, P=.013).

172 1.4

1.2

1.0

0.8

0.6

0.4 Shannon Weiner diversity (H')

0.2

0.0 Clays Sands Soil texture Figure 4: The difference in Shannon-Weiner diversity between plots with clay versus sandy soil types. A Mann-Whitney U test found the difference to be significant (U=62, P=.038).

24 Conductivity (uS/cm) Conductivity 22

16 Clays Sands Soil texture

Figure 5: The difference in conductivity between plots with clay versus sandy soil types. A Mann-Whitney U test found these differences to be significant (U=45, P=.006). 173 1.4

1.2

1.0

0.8 C % 0.6

0.4

0.2

0.0 Clay Sand Soil texture

Figure 6. The difference in carbon content between plots with clay versus sandy soils. A Mann-Whitney U-tests found the differences to be significant (U = 56, P = .02).

1500

1400

1300

1200

1100

1000 Ca (mg/kg) 900

800

700

600

500 Clay Sand Soil texture

Figure 7: The difference in calcium concentrations between plots with clay versus sandy soil types. A Mann-Whitney U test found this difference to be significant (U=35, P=.001).

174 406

404

402

400 Elevation (m)

398

396

394 Clays Sands Soil texture

Figure 8: The difference in elevation between plots with clay versus sandy soil types. A Mann-Whitney U test found these differences to be significant (U=57, P=.022).

2.0

1.8

1.6

1.4

1.2

1.0

0.8

0.6 Shannon Weiner diversity (H')

0.4

0.2

0.0 <1 >1 Shannon Wiener diversity (H')

Figure 9: The difference in Shannon Weiner diversity between plots with diversity greater than one and plots with diversity less than one. A Mann-Whitney U test found these differences to be significant (U=0, P=.000). 175 160

140

120

100 individuals

80 C. mopane mopane C. # 60

20 <1 >1 Shannon Weiner diversity (H')

Figure 10: The difference in the number of C. mopane individuals between plots with diversity greater than one and plots with diversity less than one. A Mann-Whitney U test found these differences to be significant (U=62, P=.05).

C. mopane C. 35

20 # Individuals# excluding

10 <1 >1 Shannon Weiner diversity (H')

Figure 11: The difference in the total number of individuals, excluding C. mopane, between plots with diversity greater than one and plots with diversity less than one. A Mann-Whitney U test found these differences to be significant (U=16.5, P=.000).

176

Appendix 1: Latitudes and longitudes of the study plots Plot # Latitude (S) Longitude (E) Elevation Degrees Minutes Seconds Degrees Minutes Seconds (m above sea level) 1 22 43 11 31 3 44 416 2 22 43 15 31 3 51 415 3 22 43 18 31 3 59 415 4 22 43 21 31 4 7 412 5 22 43 22 31 4 16 409 6 22 43 25 31 4 25 402 7 22 43 29 31 4 32 405 8 22 43 37 31 4 43 400 9 22 43 42 31 4 50 404 10 22 43 45 31 4 58 405 11 22 43 49 31 5 6 388 12 22 43 54 31 5 15 394 13 22 43 58 31 5 22 394 14 22 44 1 31 5 31 396 15 22 44 2 31 5 35 395 16 22 44 6 31 5 42 395 17 22 44 10 31 5 50 393 18 22 44 12 31 5 55 392 19 22 44 16 31 6 7 391 20 22 44 19 31 6 19 402 21 22 44 20 31 6 28 396 22 22 44 22 31 6 37 402 23 22 44 26 31 6 45 399 24 25 22 44 29 31 6 52 399 26 22 44 34 31 6 59 403 27 22 44 36 31 7 6 404 28 22 44 37 31 7 16 402 29 22 44 39 31 7 24 397 30 22 44 40 31 7 27 395

177

Appendix 2: Summary statistics for each multiple regression model r2 df F-Ratio P # C. mopane individuals 0.389 4, 25 3.974 0.012 Total C. mopane basal area 0.428 4, 25 4.678 0.006 Mean C. mopane basal area 0.405 4, 25 4.255 0.009 Mean C. mopane height 0.31 4, 25 2.803 0.047 # C. mopane individuals resprouting from a burnt stem 0.389 4, 25 3.978 0.012 Shannon Diversity (H') 0.322 4, 25 2.971 0.039 Evenness (J’) 0.294 4, 25 2.608 0.059

178 Biotic determinants of soil characteristics in the Northern Plains of Kruger National Park

Category: Independent Project Participants: Carla Staver Site: Punda Maria, Kruger National Park, Mpumalanga Province, South Africa

Key words: herbivory, Northern Plains, soil nutrients, vegetation

Abstract: The study investigates the effects that changing herbivory patterns in the Northern Plains of Kruger National Park have had on vegetation structure and composition and on soil characteristics, both chemical and physical. Type I herbivores such as zebra and buffalo have encroached into the area, which has been historically inhabited by type II herbivores such as roan and sable. The study found that the more intense and less selective grazing associated with type I herbivores has had a marked effect on vegetation and soil chemical characteristics. Overall grass and tree cover has decreased, and the size of the remaining trees is reduced. Colophospermum mopane seems to thrive especially under disturbance conditions. Calcium, available phosphorus, salinity, and pH have homogenized under the influence of type I grazers, while organic carbon and total nitrogen have diversified. Differential land use patterns and grazing habits of type I and type II herbivores may drive these changes. However, until these patterns are better understood, their role in determining vegetation and soil properties will remain unclear.

Introduction At its most basic level, soil formation is defined by the interaction of abiotic factors, including parent material and climate, and biotic factors. The effects of parent material and climate on soils and thus on vegetation structure have been well studied and are fairly well understood in the savanna biome. In South Africa as a whole, savanna can be divided into two broad types, moist, infertile savannas and dry, infertile savannas . Kruger National Park is generally characterized as dry and fertile, although parts around Pretoriuskop and at the far north of the park are more mesic. At a smaller scale within the Kruger, the west is dominated by granitic parent material, yielding infertile soils with a mixture of sand and clay particles, while the east is dominated by basalt-derived soils, which are more fertile and clayey than granitic soils (Venter et al. 2003). In general annual rainfall decreases northward in the park, yielding shallower soil profiles (Venter et al. 2003). Meanwhile, the biotic factors involved in shaping soils have been largely ignored. Most work on the effects of vegetation on soils in savannas has concentrated on agriculturally useful plants, specifically nitrogen-fixing legumes, which have been shown to rehabilitate degraded soils by building up soil nitrogen content from atmospheric nitrogen (Wortman 2000, Vanlauwe 2000) and phosphorous content from rock phosphorus, which is unusable to most plants (Vanlauwe 2000). Studies conducted by Witkowski (1991) on the influence of Australian alien Acacia species on nutrient cycling patterns found significant differences in nutrient turnover rates and soil nutrient and organic content in areas invaded by Acacia spp. Moreover, large trees are known to drive elevated nutrient concentrations in their immediate vicinity through a variety of mechanisms, including nutrient, ion, and hydraulic uplift (Belsky et al. 1989; Belsky et al. 1993; Ludwig 2001). They absorb water and nutrients, leaving salts and some ions behind, resulting in higher concentration of ions in the area of large trees. The potential for vegetation to influence soil chemical and physical properties is significant and diverse, but studying the relationship between vegetation structure and type and soils in the natural savanna context is complicated for practical reasons. Many variables are involved in defining soils, including parent material, climate, vegetation, herbivory, and fire, making them locally extremely heterogeneous. Determining the relationship between vegetation and soils is difficult. Parent material and climate affect vegetation structure and composition through their influences on soils, while herbivory and fire feed back as top-down controls on vegetation. They also have direct links to soil properties, making the role of vegetation ambiguous and at times indistinguishable. In natural savanna systems, the direct relationship between herbivory and soils can be difficult to determine. Mega-herbivores concentrate in nutrient-rich, fine-leaved savannas, especially near valley bottoms and on sodic sites (Blackmore et al. 1990, Rina Grant pers. com. 2004). While this might suggest that herbivores play a role in concentrating nutrients in the areas they frequent, a more likely and well supported conclusion is that herbivores feed preferentially in areas with more nutritious grasses and trees. However, even within areas of similar base soil type, mega-herbivores have been shown to accelerate and enhance nutrient cycling in savannas (McNaughton et al. 1997). Naiman et al. (2003) outline a mechanism whereby herbivory by both mega- and smaller herbivores influences soils, 179 which in turn feed back on herbivore populations, creating a temporally heterogeneous environment of herbivore populations, vegetation structure, and soil characteristics. Determining the influences of herbivory and changing vegetation structure is of particular interest on the northern plains of Kruger National Park, where the addition of permanent waterpoints to the landscape has resulted in a change in natural herbivory patterns, resulting in differences in vegetation structure (Gaylard et al. 2003). The area is a traditional and natural habitat for Type II herbivores, such as roan and sable, which are characterized by having little to no effect on vegetation structure where they live. These herbivores generally require tall areas of underutilized grasses together with shorter stands of grasses for survival, and are generally considered to be water- independent species (Collinson and Goodman 1982). The addition of waterpoints has made the area suitable for water dependent species, like elephant, white rhino, buffalo, and zebra, which are classified as Type I herbivores. These have a marked effect on vegetation structure, due to both grazing habits and population density (Collinson and Goodman 1982). However, the effects of herbivory and vegetation compositional shifts on salient soil characteristics have not been extensively investigated. Levick (2001) documented a decrease in woody vegetation density and cover in the absence of browsing, but discussed changes in soil characteristics only with reference to the riparian/upland boundary. Thus, opportunities for determining the role of herbivory and vegetation structure in determining soil properties are ample. The effects of fire has well documented effects on nutrient concentrations in soils is well documented. Frequent fire leads to decreases in organic carbon, total nitrogen, and soil moisture (Jones et al. 1990; Shackleton and Scholes 2000) and increases in calcium, magnesium, potassium, and sodium (Snyman 2002). For this reason, this study will control for fire frequency, season, and intensity.

Methods The roan enclosure directly north of Shingwedzi provides an ideal location in which to examine the effects of herbivory and vegetation structure on soil characteristics in the Northern Plains. The enclosure lies on olivine-rich basalt in the Karoo geologic sequence and receives between 350 and 450 mm of rainfall per year. The enclosure was established in 1968 in an area upslope from a river; riparian and fluvial zones were added to the enclosure in 1985. This study examined upslope areas, where vegetation structure and soils have had more time to adjust to the exclusion of herbivory. The roan enclosure is divided into four quarters. One lies on black soil, which is dominated by marula (Sclerocarya birrea) and knobthorn (Acacia nigrescens), and was burnt 4-5 years ago. Fifteen sites were sampled in this quarter of the roan enclosure. The remainder of the enclosure lies on mopane-veld (dominated by Colophospermum mopane), characterized by red soils, which extends beyond the fence of the enclosure. Fifteen plots were sampled in the south-west corner of the enclosure, an area dominated by C. mopane, which was burned 4-5 years ago. Corresponding plots were sampled outside the enclosure on each soil type (see Figure 1 for a map of the enclosure with sampling layout). The fifteen plots in each site lie every 30 meters along three transects. These transects run parallel to the slope of the hill in order to minimize bias from effects of down-slope water, nutrient, and small particle flow on soil properties. The following parameters were sampled in each plot: • Trees within 400m2 circle (r=11.3): species, height, DBH • Grasses within circle of 2m radius: species, abundance, total cover • Soil samples at the center of these circles Soil nutrients (total N, available P, Ca) were analyzed at a soil lab according to the standard method. Soil pH was determined using the method outlined by Rowell (1994), using a 1:2.5 ratio of soil to 0.01 M CaCl2 solution. Conductivity was determined with a 1:5 ratio of soil to water and converted to standard-scale conductivity (ECe) using the method outlined by Landon (1991). Volumetric soil moisture content was determined by weighing soil moisture samples and then drying them at 105oC for 24 hours. Porosity was determined in the field using a surrogate measure; I measured the absorption time of 100 mL of water over an area of ~20 cm2. I used a field method for determining soil texture outlined by Tongway and Hindley (1995).

Results Vegetation Grass cover decreased on both soil types from inside to outside the enclosure (see Figure 2). On C. mopane soil type, grass cover decreased from 74.5% inside the enclosure to 59.0% outside the enclosure (U=53, p=.014). On S. birrea soil type, grass cover decreased from 78.0% inside to 50.7% outside (U=28, p<.000). There were no

180 significant differences in grass cover between soil types inside (U=97, p=.520) or outside the enclosure (U=84.5, p=.245). Total tree density of individuals with height greater than two meters decreased on both soil types from inside to outside the enclosure (see Figure 3). On C. mopane soils, tree density decreased from 8.7 individuals per plot inside the enclosure to 3.2 individuals per plot outside (U=16.5, p<.000). On S. birrea soils, tree density decreased from 5.3 individuals per plot inside to 1.4 individuals per plot outside (U=16, p<.000). Variance for tree density inside was significantly higher than variance for tree density outside the enclosure for both soil types together (F=4.10, p<0.05, see Table 1 for F-statistics). Mean tree height decreased on both soil types from inside to outside the enclosure (see Figure 4). On C. mopane soils, mean tree height decreased from 4.0 m inside the enclosure to 2.9 m outside the enclosure (U=45, p=.009). On S. birrea/A. nigrescens soils, mean tree height decreased from 4.8 m inside to 2.9 m outside the enclosure (U=30.5, p=.013). There were no differences in mean tree height between soil types either inside (U=94, p=.443) or outside the enclosure (U=68.5, p=.930). Variance in mean tree height inside was significantly greater than variance outside (F=9.20, p<.05). Multi-dimensional scaling analysis on tree species and abundance reveals a clear community switch from inside to outside the enclosure (see Figure 5). Communities on the different soil types inside the enclosure appear to be more closely related that communities outside, but this largely ignores the prevalence of dominant species and may give undue importance to species that exist only inside because they cannot survive outside. Analysis of prevalence of dominant species was necessary for this reason. The density of C. mopane individuals of all heights remained constant on C. mopane dominated soils inside and outside the enclosure at ~5.5 individuals per plot (U=107, p=.820). On S. birrea type soils, the density of C. mopane increased significantly from 1.5 individuals per plot inside the enclosure to 4.7 individuals per plot outside the enclosure (U=66.5, p=.046). This suggests that vegetation undergoes the same homogenization outside the enclosure as soil characteristics do (see Figure 6). Indeed, C. mopane densities inside the enclosure are very different (U=45.5, p=.005), while outside the enclosure C. mopane densities are similar (U=90, p=.351). Size class distributions for C. mopane in all four study sites show similar patterns. Inside on C. mopane type soils, mopane individuals are taller and more abundant than they are on S. birrea type soils inside the enclosure. However, outside the enclosure, mopane individuals are shorter and of similar abundance on both soil types (see Figure 7). However, variances for the density of C. mopane individuals inside and outside the enclosure were not significantly different (F=1.22, p>.05). Changes in density of large individuals, defined as having DBH ≥ 20 cm, show similar homogenization patterns (see Figure 8). S. birrea type soils inside the enclosure are dominated by large trees (1.1 individuals per plot), while the same soil type outside has no large trees (U=45, p=.005). C. mopane type soils have a higher density of large trees (0.5 individuals per plot) than the same soil type outside the enclosure (0.0 individuals per plot), but this difference is not as marked as on S. birrea type soil nor is it significant (U=75, p=.120). Further sampling might reveal significant differences in density of large trees between soil types inside the enclosure, but sampling area was insufficient to yield results (U=79.5, p=.171). There is no difference between density of large trees on the different soil types outside the enclosure (U=112.5, p=1.000). Figure 9 shows the change in distribution of S. birrea and A. nigrescens abundance for the whole site from inside to outside the roan enclosure and between soil types. S. birrea and A. nigrescens were more abundant on marula/knobthorn soil type, as expectd; S. birrea abundance decreased to zero outside the enclosure, while A. nigrescens increased outside the enclosure. S. birrea decreased on C. mopane type soils, while A. nigrescens remained constant across both treatments at a low level. Soil Characteristics Neither absorption rate (C. mopane type, inside v. outside: U=62, p=.102; S. birrea type: U=94, p=.631) nor soil texture (C. mopane type: U=105, p=.756; S. birrea type: U=105, p=.756) exhibited any differences between from inside to outside the enclosure. Moreover, both absorption rate (inside, C. mopane v. S. birrea type: U=88, p=.458; outside: U=89, p=.695) and soil texture (inside: U=111.5, p=.967; outside: U=111, p=.950) were similar between soil types, variously dominated by C. mopane and S. birrea. On C. mopane-dominated soil type, pH increased from 6.0 inside the enclosure to 6.3 outside the enclosure (U=58, p=.023). On S. birrea-dominated soil type, pH decreased from 6.6 inside to 6.2 outside (U=50, p=.010). Inside the enclosure, pH was higher on S. birrea-dominated soil (U=24, p<.001); outside the enclosure, pH values were similar on the two soil types (U=102, p=.663). pH appears to homogenize outside the enclosure (see Figure 10). pH increases significantly from C. mopane dominated area to S. birrea dominated area inside the enclosure along the slope of the transect; no such increase is evident outside the enclosure (see Figure 11). pH determined in 0.01M 181 CaCl2, a salt solution, are 0.5 to 0.9 pH values lower than pH values determined in water (Landon 1991). For this reason, actual pH values on C. mopane dominated soil probably increase from 6.5 – 6.9 inside to 6.8 – 7.2 outside. On S. birrea dominated soil actual pH values may decrease from 7.1 – 7.5 inside to 6.7 – 7.1. While most soils are still neutral (pH 5.5 – 7.0) after the conversion, soils dominated by S. birrea inside the enclosure have a high pH (pH 7.0 – 8.5) (Landon 1991). pH values outside the enclosure exhibited a lower variance than values inside (F=3.25, p<.05). Similarly, conductivity increases from 0.22 mS cm-1 inside the enclosure to 0.38 mS cm-1 on C. mopane type soils (U=18, p<.001). On S. birrea type soils, conductivity decreases from 0.39 mS cm-1 inside to 0.28 mS cm-1 outside (U=65, p=.049). Conductivity differed significantly inside the enclosure between soil types (U=25, p<.001); conductivities on the various soil types were similar outside the enclosure (U=72.5 , p=.097). Conductivity also seems to homogenize outside the enclosure (see Figure 12). Along the transect slope, conductivity inside the enclosure jumps at the boundary between C. mopane soil type and S. birrea type soil; aside from a peak at C. mopane plot 5, which has a high error value, conductivity appears more uniform outside the enclosure (see Figure 13). However, although conductivity values homogenized outside the enclosure, there was no difference in variance outside the enclosure (F=1.09, p>.05). Calcium and phosphorous analyses showed similar patterns, but these results were not significant. On C. mopane type soil, calcium content increased from 4373 mg/kg inside the enclosure to 4911 mg/kg outside the enclosure (U=16, p=.475). On S. birrea type soil, calcium content decreased from 6334 mg/kg inside to 4338 mg/kg outside (U=11, p=.262). Although it is not significant, the trend toward homogenization (see Figure 14) outside the enclosure may be important for further investigation. Calcium content values inside on both soil types together are characterized by a significantly higher variance than calcium values outside the enclosure (F=13.19, p<.05) (see Figure 15). Phosphorus shows the opposite change, but similar homogenization, although differences are likewise not significant. On C. mopane type soil, phosphorus content decreased from 16.7 mg/kg inside the roan enclosure to 10.6 mg/kg outside the enclosure (U=17, p=.568). On S. birrea type soil, phosphorus content increased from 6.6 mg/kg inside to 9.4 mg/kg outside (U=13, p=.423). Here again, this trend toward homogenization (see Figure 16) may be of interest for further investigation. Phosphorus values inside on both soil types together have a higher standard variance than values outside, but this difference was not quite significant (F=2.53, p<.05) (see Figure 17). Volumetric moisture content was soil type specific. Inside the enclosure, moisture content was different between C. mopane and S. birrea/A. nigrescens dominated soils (U=54, p=.015). The same was true outside the enclosure (U=38, p=.002). Moisture content remained constant from inside to outside the enclosure on both C. mopane dominated soils (U=95, p=.950) and S. birrea/A. nigrescens dominated soils (U=89, p=.330). The various soil types also responded differently in terms of carbon and nitrogen content. Percentage organic carbon content by weight was similar on both soil types inside the roan enclosure (U=19, p=.775). Organic carbon decreased from inside to outside the enclosure on C. mopane type soil, but this decrease was not quite signficant (U=7.5, p=.054). Organic carbon did not change significantly on S. birrea/A. nigrescens soil from inside to outside the enclosure (U=12, p=.336) Organic carbon was significantly lower on C. mopane soil type than on S. birrea/A. nigrescens outside the enclosure (U=3, p=.016) (see Figure 18). As average height increased, organic carbon content of soil increased on all soil types together (r2=0.434, p=.002) (see Figure 19). Organic carbon content was negatively correlated with density of C. mopane (r2=0.251, p=.011). Total nitrogen was closely correlated to organic carbon content (r2=0.567, p<.001, see Figure 20) and followed similar patterns. Nitrogen levels were similar on the two soil types inside the enclosure (U=19, p=.775). On C. mopane type soils, total nitrogen decreased from inside to outside (U=6.5, p=.038). Again, no change occurred on S. birrea/A. nigrescens soil type (U=11, p=.262). Nitrogen levels were significantly lower on C. mopane type soil outside the enclosure between soil types (U=1, p=.006). Thus, nitrogen levels were similar inside the enclosure and divered outside the enclosure (see Figure 21). Total nitrogen was also correlated with both average tree height (r2=0.296, p=.013) and density of C. mopane individuals (R2=.284, p=.006).

Discussion The influx of type I herbivores, including zebra and buffalo, has had a marked effect on vegetation and, consequently or directly, on soil chemical properties. Grass cover and tree density and general tree height have decreased under the influence of intense grazing. While vegetation types are still distinguishable by the occurrence as such key species as Acacia nigrescens outside the enclosure, the two types resemble each other much more closely outside the enclosure than inside. Colophospermum mopane seems to thrive under disturbance conditions of type I grazers, resulting in a landscape dominated by C. mopane outside the enclosure. 182 A range of soil chemical properties follow the same patterns, including calcium, phosphorus, pH and conductivity. In general, the S. birrea/A. nigrescens soil type inside the enclosure is characterized by higher nutrient and ion concentrations. Available phosphorus is a notable exception, but this may be due to the interaction of high pH and calcium content values, which acts to convert available phosphorus to insoluble and unusable calcium phosphate. Under the influence of type I herbivores, these marked differences disappear, giving way to a more homogeneous set of chemical characteristics. However, this is not true for soil physical properties or for all chemical properties. Available water content did not change according to herbivory pattern and seems to be soil type-specific. Organic carbon content and total nitrogen were homogeneous inside the enclosure across soil types, but divered outside, creating a more heterogeneous environment. However, these differences are to be expected. To begin with, C. mopane and S. birrea/A. nigrescens soil types are distinct types, and may well react differently to the same herbivory pattern. Alternatively, herbivory patterns within a treatment, inside or outside, may be different according to the nutritional and protection value of the vegetation. At this point, the grazing patterns of type I and type II herbivores is not well understood. It is thought that type II herbivores prefer to graze in the vlei area, in the riparian/upland boundary, and in the S. birrea/A. nigrescens dominated area, possibly due to increased nutrient content (Levick 2001). Type II herbivores are thought to graze preferentially in the S. birrea/A. nigrescens area (Rina Grant, pers. comm.), but their land use patterns have not been studied in detail. A simple preference for marula/knobthorn bush would not explain the decrease in total nitrogen and organic carbon content on C. mopane dominated soils, although it might explain the homogenization of some of the other soil chemical characteristics and slightly elevated levels of nitrogen and carbon on the S. birrea/A. nigrescens soil type itself. Intensive grazing and land use by grazers and browsers increases resident soil nutrients through an acceleration in nutrient cycling (McNaughton et al. 1997). Land use and grazing patterns of both type I and type II herbivores will have to be studied in much more detail before we can gain an in depth understanding of their possible direct effects on vegetation structure and on soil chemical properties. Large trees may also play a significant role in changing chemical dynamics of soil on S. birrea/A. nigrescens dominated soils inside the enclosure, where large trees are common. Large Acacia trees (dbh ≥ 50cm) can create hydraulic lift and to act as ion pumps in the micro-environment under their canopies (Belsky et al. 1989, Belsky et al. 1993). This may in large part contribute to elevated salinity, calcium, and pH values found on S. birrea/A. nigrescens type soils inside the enclosure. However, large trees were not found with any abundance in the other three sites, and thus their potential to explain overall patterns is limited. A third, and largely unexplored, possibility is differential herbivory patterns of insects in each of the four sites. While mega-herbivores certainly drive changes in large-scale vegetation shifts, insect communities may be vastly different in the different sites. Insect herbivory can even have significant effects on vegetation structure, in the shape of seedling mortality (Shaw et al 2002). Additionally, the mopane worm is unique to areas dominated by C. mopane, where it can have huge effects. Mopane worms are the single biggest consumer of C. mopane and constitute more browsing on C. mopane than all other browsing put together (Halzka Hraber, pers. comm.). Moreover, termites are known to greatly accelerate nutrient cycling in savannas (Holt et al. 1990; Naiman et al. 2003). Insect communities could have huge effects on nutrient cycling and soil chemical characterisitics, which this study has ignored. Potential drivers of these unique patterns in soil chemical characteristics are many. While this study can by no means solve the mechanism behind them, it identifies the need and possibility for future research. A greater understanding of soil patterns surrounding shifts in herbivory patterns associated with the switch from type II to type I grazers could have significant implications for a more complete understanding of savanna dynamics on the Northern Plains and the extent of conservation concerns in the area.

Literature Cited Belsky, A.J., R. G. Amundson, J. M. Duxbury, S. J. Riha, A. R. Ali, and S. M. Mwonga. 1989. The effects of trees on the physical, chemical, and biological environments in a semi-arid savanna in Kenya. Journal of Applied Ecology 26: 1005-1024. Belsky, A. J., S. M. Mwonga, R. G. Amundson, J. M. Duxbury, and A. R. Ali. 1993. Comparative effects of isolated trees on the undercanopy environments in high- and low-rainfall savannas. Journal of Applied Ecology 30: 143- 155. Blackmore, A. C., M. T. Mentis, and R. J. Scholes. 1990. The origin and extent of nutrient-enriched patches within a nutrient-poor savanna in South Africa. Journal of Biogeography 17: 463-470. Bond, W. J. 1997. Fire. Pages 421-442 in Cowling, R. M., D. M. Richardson, and S. M. Pierce. The Vegetation of Southern Africa. Cambridge: Cambridge UP. 183 Collinson, R. F. H. and P. S. Goodman. 1982. An assessment of range condition and large herbivore carrying capacity of the Pilansberg Game Reserve, with guidelines and recommendations for management. Inkwe 1: 1-54. Fitter, A. 1997. Nutrient acquisition. Pages 51-72 in Crawley, M. J. Plant Ecology. Oxford: Blackwell Science. Gaylard, A., N. Owen-Smith, and J. Redfern. 2003. Surface water availability: implications for heterogeneity and ecosystem processes. Pages 171-188 in Du Toit, J. T., K. H. Rogers, and H. C. Biggs. The Kruger Experience: ecology and management of savanna heterogeneity. Washington: Island Press. Holt, J. A. and R. J. Coventry. 1990. Nutrient cycling in Australian savannas. Journal of Biogeography 17: 427-432. Jones, C. L., N. L. Smithers, M. C. Scholes, and R. J. Scholes. 1990. The effect of fire frequency on the organic components of a basaltic soil in the Kruger National Park. South African Journal of Plant and Soil 7: 236-238. Joubert, S.C.J 1970. A study of the social behavior of the roan antelope, Hippotragus equinus equinus in the Kruger National Park. M.Sc thesis, University of Pretoria, Pretoria, South Africa. Landon, J. R. 1991. Booker Tropical Soil Manual: A handbook for soil survey and agricultural land evaluation in the tropics and subtropics. London: Booker Agriculture International Ltd; New York: Longman. Levick, S. 2001. Effects of large mammalian herbivore exclusion on the physiognomy, species composition and boundary dynamics of woody vegetation, across a vlei/upland boundary. Honors thesis, University of the Witwatersrand, Johannesburn, South Africa. Ludwig, F. 2001. Tree-grass interactions on an East African savanna: the effects of competition, facilitation and hydraulic lift. Tropical Resource Management Papers 39. Wageningen University. Marshall, T. J. and J. W. Holmes. 1979. Soil Physics. Cambridge: Cambridge UP. McCarron, J. K., A. K. Knapp, and J. M. Blair. 2003. Soil C and N responses to woody plant expansion in a mesic grassland. Plant and Soil 257: 183-192. McNaughton, S. J., F. F. Banyikwa, and M. M. McNaughton. 1997. Promotion of the Cycling of Nutrients by African Grazers. Science 278: 1798-1800. Naiman, R. J., L. Braack, R. Grant, A. C. Kemp, J. T. du Toit, and F. J. Venter. 2003. Interactions between species and ecosystem characteristics. Pages 221-241 in Du Toit, J. T., K. H. Rogers, and H. C. Biggs. The Kruger Experience: ecology and management of savanna heterogeneity. Washington: Island Press. Rowell, D. L. 1994. Soil science: methods and applications. Scholes, R. J. 1990. The influence of soil fertility on the ecology of southern African dry savannas. Journal of Biogeography 17: 415-419. Shackleton, C. M. and R. J. Scholes. 2000. Impact of fire frequency on woody community structure and soil nutrients in Kruger National Park. Koedoe 43: 75-81. Shaw, M. T., F. Keesing, and R. S. Ostfeld. 2002. 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184 Table 1. F-statistics for differences in variance in selected vegetation and soil characteristics. Inside v. Outside

2 2 s inside s outside F P Ninside; Noutside

Tree density 12.76 3.11 4.10 <0.05 30; 30

Average tree height 3.65 0.4 9.20 <0.05 30; 30

Density of C. mopane 25.47 31.2 1.22 >0.05 30; 30

pH 0.26 0.08 3.25 <0.05 30; 30

Conductivity 0.02 0.03 1.09 >0.05 30; 30

Ca (mg/kg) 4530000 344000 13.19 <0.05 13; 12

P (mg/kg) 138.12 54.67 2.53 >0.05 13; 12

N vlei

Transects Transects Transects on C. on S. on C. mopane birrea/A. mopane inside nigrescens outside inside

Transects on S. birrea/A. nigrescens outside

Figure 1. Aerial photograph of the roan enclosure, Northern Plains, KNP, with locations of transects and major geographic landmarks.

185 1.0

Soil type - C. mopane dominated 0.9 Soil type - S. birrea/A nigrescens dominated

0.8

0.7

0.6 Percentage grass cover grass Percentage

0.5

0.4

0.3 inside outside Treatment

Figure 2. Grass cover v. treatment and soil type.

10 Soil type - C. mopane dominated Soil type - S. birrea/A. nigrescens dom 9

5 Total tree density 4

0 inside outside Treatment

Figure 3. Tree density (individuals per 400 m2) v. treatment and soil type.

186

6.5

6.0

5.5 Soil type - C. mopane dominated Soil type - S. birrea/A. nigrescens dom 5.0

4.5

4.0

3.5

3.0 Average height (m) Average height 2.5

2.0

1.5

1.0

0.5 inside outside Treatment

Figure 4. Average height (m) v. treatment and soil type.

Figure 5. Multi-dimensional scaling of communities based on tree species and abundance. 187

9 Soil type - C. mopane dominated Soil type - S. birrea/A. nigrescens dominated 8

5 individuals (per sq. 400 m)

C. mopane 3

2 Density of 1

0 inside outside Treatment

Figure 6. Density of C. mopane individuals (per 400 m2) v. treatment and soil type.

C. mopane soil type S. birrea soil type 40 40 35 35 30 30 25 25 20 20 15 Inside 15 ubrof Number individuals

Number of 10 10 individuals 5 5 0 0 012345678 012345678 Height (m) Height (m)

40 40 35 35 30 30 25 25 20 20 Outside 15 15 10 10 5 5 Number of individuals of Number 0 Individuals of Number 0 012345678 012345678 Height (m) Height (m) Figure 7. Size class distributions of C. mopane on the various treatments and soil types.

188 1.8

1.6 Soil type - C. mopane dominated Soil type - S. birrea/A. nigrescens dominated 1.4

1.2

1.0

0.8

0.6

0.4 Density of individuals with dbh>20cm (per 400 sq. 400 m) dbh>20cm (per with individuals of Density 0.2

0.0 inside outside Treatment

Figure 8. Density of individuals with dbh ≥ 20 cm (per 400 m2) v. treatment and soil type.

Sclerocarya birrea 10 Acacia nigrescens

Number of Individuals (whole site) (whole Number Individuals of 4

0 Col mop inside Col mop outside Scl bir/Aca nig inside Scl bir/Aca nig outside

Figure 9. Number of S. birrea and A. nigrescens individuals for each site (6,000 m2).

189 7.0

Soil type - C. mopane dominated 6.8 Soil type - S. birrea/A. nigrescens dominated

6.6

6.4 pH

6.2

6.0

5.8

5.6 inside outside Treatment

Figure 10. pH v. treatment and soil type.

7.5 pH inside pH outside 7

6.5 pH

5.5

5 Mop 1 Mop 2 Mop 3 Mop 4 Mop 5 Marula 1 Marula 2 Marula 3 Marula 4 Marula 5 Position along slope

Figure 11. pH v. position along transect. Mopane 1 is the farthest upslope, and marula 5 is the farthest downslope. Transition from C. mopane dominated soils to S. birrea/A. nigrescens dominated soils occurs between mopane 5 and marula 1.

190 0.50

0.45

0.40

0.35

0.30

0.25 Conductivity (mS/cm)

0.20

Soil type - C. mopane dominated 0.15 Soil type - S. birrea/A. nigrescens dominated

0.10 inside outside Treatment

Figure 12. Conductivity (mS/cm) v. treatment and soil type.

1.2 conductivity inside conductivity outside 1

0.8

0.6

0.4 Conductivity (mS/cm) Conductivity

0.2

0 Mop 1 Mop 2 Mop 3 Mop 4 Mop 5 Marula 1 Marula 2 Marula 3 Marula 4 Marula 5 Position on slope

Figure 13. Conductivity v. position along slope. Mopane 1 is the farthest upslope, and marula 5 is the farthest downslope. Transition from C. mopane dominated soils to S. birrea/A. nigrescens dominated soils occurs between mopane 5 and marula 1.

191 8500

8000 Soil type - C. mopane dominated Soil type - S. birrea/A. nigrescens dominated 7500

7000

6500

6000

5500

Ca (mg/kg) 5000

4500

4000

3500

3000

2500 inside outside Treatment

Figure 14. Calcium content v. treatment and soil type.

8000

7500

7000

6500

6000

5500

5000 Ca (mg/kg)

4500

4000

3500

3000

2500 inside outside Treatment Figure 15. Calcium content v. treatment. Soil types within a treatment are pooled to give variation within a treatment. Error bars denote standard deviation.

192 30

25 Soil type - C. mopane dominated Soil type - S. birrea/A. nigrescens dominated

P (mg/kg) 10

-5 inside outside Treatment

Figure 16. Phosphorus content v. treatment and soil type.

12 P (mg/kg)

0 inside outside Treatment Figure 17. Phosphorus content v. treatment. Soil types within a treatment are pooled to give variation within a treatment. Error bars denote standard deviation.

193 3.2

Soil type - C. mopane dominated 3.0 Soil type - S. birrea/A. nigrescens dominated

2.8

2.6

2.4

% Organic Carbon 2.2

2.0

1.8

1.6 inside outside Treatment

Figure 18. Organic carbon content v. soil type and treatment.

3.0

2.5

2.0

1.5 % Organic Carbon

1.0

0.5

0.0 012345678 Average height (m) Figure 19. Organic carbon content for plot v. average height for plot (R2=.434 p=.002).

194 0.16

0.14

0.12

0.10

0.08

% Total Nitrogen 0.06

0.04

0.02

0.00 0.00.51.01.52.02.53.0 % Organic Carbon Figure 20. Total nitrogen content v. organic carbon content (R2=.567 p<.001).

0.15

0.14 Soil type - C. mopane dominated Soil type - S. birrea/A. nigrescens dominated

0.13

0.12

0.11

0.10 % Total Nitrogen

0.09

0.08

0.07

0.06 inside outside Treatment

Figure 21. Total nitrogen content v. treatment and soil type.

195 Assessment of the proportion and extent of elephant damage on Acacia nigrescens and Sclerocarya birrea in the Punda Maria area in the Kruger National Park

Category: Independent Project Participants: Laura Buckley, Shannon Hatmaker, Justine Norman, Simon Thomson Site: Punda Maria, Kruger National Park, Mpumalanga Province, South Africa

Key words: density, elephant impact, population structure, recover, tree damage

Abstract: The Kruger National Park has responded to growing concern over the increasing elephant population by first implementing a population control management strategy through culling and in recent years, suspending culling and implementing a policy of promoting heterogeneity through maintaining differential elephant numbers in different areas in the park. The concern stems from the potential of elephants to change the population structure of certain keystone tree species in the park drastically, such as Acacia nigrescens and Sclerocarya birrea. This study surveys four sites in the Punda Maria area and assesses the population structure, damage intensity, and degree of recovery of over eight hundred A. nigrescens individuals and over one hundred S. birrea individuals. Almost twenty-five percent of all living A. nigrescens and over eighty-two percent of S. birrea sampled showed some sign of elephant damage; however, only a small proportion of these trees showed high levels of intense damage, and the intensity of damage was distributed differentially across height classes. Also, these species showed a great deal of recovery in the forms of resprouting after branch damage and bark regrowth after stripping. Large proportions of damage were recorded in both populations but a very low proportion of this damage was severe. A. nigrescens and S. birrea both show immense capability to recover. Survival strategies differ between the two species. Elephants prefer to remove bark from marula trees and in response, these trees have adapted an ability to repair removed bark efficiently. Complete recovery of bark damaged areas was commonly noted. Elephants readily break branches of A. nigrescens and in response, this species resprouts quite readily.

Introduction Over recent years, elephant damage has been a major concern for Kruger Park management as the density of elephants has been increasing since the discontinuation of culling in 1995 (Whyte 2001 as cited in Jacobs and Biggs 2002). Between 1976 and 1994, the elephant population was managed through culling, maintaining the population at 7,000 individuals (Whyte 2003). At a public debate in May 1995, SANParks undertook to review its policy on elephant management and as a result culling was discontinued (Whyte 2003). As stated in the new elephant management proposal (Whyte 2003) the new policy is based biodiversity and has three fundamental principles for the management of elephants in the KNP: 1. Flux in ecosystems is natural and desirable as it contributes to biodiversity, 2. Elephants are important agents of disturbance and thus create heterogeneity, which will contribute to biodiversity, 3. Elephant populations that are confined but not managed will increase to a number that will have negative impacts on biodiversity. The basic principles of elephant management have shifted from only managing the elephant population to now managing the elephant impact and other ecosystem processes, such as fire, to promote biodiversity in general (Whyte 2003). The reduction of diversity of habitats, elimination of plant species, and structural changes to vegetation caused by elephants at high densities may greatly affect biodiversity (Whyte 2003). This is why Kruger’s elephant policy uses Thresholds of Potential Concern (TPCs) that focus on the extent and intensity of damage on biodiversity, and not just elephant numbers. The new elephant management policy as explained by Whyte, Aarde and Pimm (2003) divides the park into six zones, each to be managed differently in terms of their elephant populations. There will be two botanical reserves, where it is felt that the vegetation types are placed at risk due to excessive elephant impacts. There will be two high elephant impact zones, where elephant numbers will be allowed to increase until one or more TPCs are crossed. Lastly, there will be two low elephant impact zones, where the elephant numbers will be decreased by 7% each year until one or more TPC’s are crossed. Should a TPC be crossed, the zones will be swapped around, with the exception of the botanical reserves. This study focused on Acacia nigrescens and Sclerocarya birrea because these are two of the preferred tree 196 species selected by elephants (Jacobs and Biggs 2002). Bark removed by elephants can kill trees directly or else make them more susceptible to fire or wood-boring insects (Barnes 1980 as cited in Jacobs and Biggs 2002). In the Kruger National Park, serious concern over the damaged condition and reduction of mature trees has been raised by studies indicating the decrease in range, shift in population structure, and decrease in density of these trees (Whyte 2003, Jacobs and Biggs 2002). Elephants are completely removing many adult trees from the population (Croze 1974, Dublin 1995, Tchambe 1995 as cited in Gadd 2002) as well as negatively affecting recruitment by eating the seedlings (Gadd 2002). A. nigrescens is a medium to large deciduous tree that can grow up to 30m high in arid bushveld on a variety of soils. It is an important browse tree for game (Schmidt, Lötter, McCleland and Burrows 2002). S. birrea is also a medium to large deciduous tree that can reach heights of up to 18m and grows found in bushveld and woodland. It has the widest distribution of any tree in the Kruger National Park. It is a keystone species for many insects breed and feed on the tree, many types of game browse its leaves, and elephants often strip its bark (Schmidt et al. 2002). According to Gadd (1997) elephants broke 18% of marula trees in patches of woodland where they fed, thus making it the most favoured tree relative to abundance. The marula tree is one of the most highly valued indigenous trees (Coates Palgrave, 1993 as cited in Gadd 2002) for the variety of uses it serves. For example, its bark is used for medicinal purposes and to obtain dye (Schmidt et al. 2002) and its fruits are widely used locally and recently are being marketed internationally (Gadd, 2002). The marula tree is an officially protected tree in South Africa (Coates Palgrave, 1993 as cited in Gadd, 2002). A study done on A. nigrescens was conducted in 1979 in the Punda Maria area (Engelbrecht 1979). Its results indicated that 65.3% of the 951 trees examined were damaged and 27.7% were dead or dying due to toppling or ring barking. The extent of damage was not included in this study, therefore making it difficult to compare the extent of past damage in the area to what is seen now. A recent study on S. birrea conducted by Jacobs and Biggs 2002 found that almost half of the surveyed population over four different landscapes in the Kruger National Park suffered damage from elephant activity. This study took into consideration damage at different intensities and height classes. Many factors are involved in evaluating elephant impacts on trees, including the type of damage, the ability of the tree to recover, and the role that the tree plays in the ecosystem (Engelbrecht 1979). The aim of this study is to determine the population structure of A. nigrescens and S. birrea in the Punda Maria area and the effects that elephants are having on the demography of these two species. We will also be assessing the proportion of trees damaged and to what extent this damage is taking place.

Methods Four sites were selected in the Punda Maria area along main roads (Fig 1). Roads may influence the vegetation through increased runoff and increased herbivore activity, especially when roads are used as pathways (Coetzee, et al. 1979 as cited in Jacobs and Biggs, 2002). By restricting the location of transects along these four roads, the possible influence of the roads will be kept constant. The first site, Pafuri Road, was 300 m in length and 40 m off each side of the road. The second transect was conducted at Loop 1 and was 600 m in length and 50 m off one side of the road for A. nigrescens and was 2.8 km and 30 m off both sides of the road for S. birrea. The third transect was conducted 40 m off each side of the Main Road and was 400 m in length. The last transect was performed on one side of Loop 2 and was 400 m in length and 50 m off the road. At each of these sites, all A. nigrescens and S. birrea trees, both alive and dead, were examined for population structure, damage due to elephants, and ability to recover. This was done by placing the trees into height and damage classes. For A. nigrescens, the height classes are as follows: <2 m, 2-5 m, 5-10 m, 10-15 m, and >15 m. Height classes for S. birrea include: <2 m, 2-8 m, 8-15 m, 15-20 m, and >20 m. Five damage classes were adapted from Jacobs and Biggs 2002 study. Nil represents trees that had zero elephant damage. Light damage was marked by having less than 50 percent bark removed and/or secondary branch damage. Trees with moderate damage had less than 50 percent bark removed and less than 50 percent primary branch damage or less than 50 percent primary branch damage. Heavy damage was noted by having greater than 50 percent bark removed and less than 50 percent primary branch damage or greater than 50 percent bark removed or greater than 50 percent primary branch damage. Trees considered to have extra heavy damage had 100 percent bark removed or 100 percent primary branch damage. The percentage of bark regrowth, presence of resprouting, and the presence of wood-boring insects were also noted. Dead trees were also recorded and their cause of death was attributed to either elephant or other causes. A tree’s death was classified as caused by elephants only if it had been toppled (roots showing), if there was 100 percent main stem breakage, or if it had been ring barked. Dead trees were not classified into height classes but instead into 197 DBH (diameter at breast height) classes because some dead trees were stumps. A G-test was used to detect differences between the distribution of trees across height classes and damage classes for A. nigrescens and S. birrea. A Post hoc analysis (Tukey-Kramer Honestly Significant Difference Test) was used to analyze relationships between height and damage classes for both tree species. A linear regression analysis was performed to assess the relationship between the extent of bark damage and recovery and between recovery of bark and tree height for both tree species. Statistical analysis was done using JMP IN software (Version 5.1).

Results Acacia nigrescens Population Structure Density of Acacia nigrescens in the total area sampled was 7.70 trees/km2 (Table 1). Of the 919 trees sampled, the largest proportion (61.9%) of trees is less than two meters tall. The population structure depicts an inverse j-curve (Fig. 1). There was significant difference between the distribution of trees across height classes [G = 795.0015, df = 4, P < .05]. Sampling of dead trees was difficult and without certainty, it can be said that 44.66% of the 103 dead trees were due to elephant damage. Over sixty-seven percent of these were in the >20cm DBH size class. Only 10.86% of the dead trees attributed to elephants were killed by toppling and the remaining trees were due to 100% main stem breakage (Fig.2). All of the toppled trees had a DBH of more than 20cm. Damage In the sampled population, 24.4% of all trees show some sign of damage attributed to elephants. The least amount of damage is seen on the smallest trees (<2m), where less than 2% of this size class display any damage. The proportion of damaged trees increases as the size of the tree increases. The 10-15m class has the most combined moderate, heavy, and extra heavy damage (Fig. 3). There is a significant difference in the distribution of all damage classes across height classes[G = 1251.398, df = 4, P < .05]. However, a Post hoc analysis (Tukey-Kramer Honestly Significant Difference Test) showed that there is no significant difference between the distribution of moderate and extra heavy damage classes (Table 2). Figure 4 shows that trees in the 10-15m class had the highest proportion of trees showing primary branch damage as compared to other classes. The larger height classes (10-15m and >15m) suffered much more from bark removal than the smaller classes. Trees less than 2m tall showed no bark removal. Recovery Figure 5 shows that 26.40% of trees with stem damage displayed resprouting. There is a greater proportion (36.36%) of resprouting individuals in the 5-10m size class than in any other class. Fifty-five percent of trees that incurred bark damage showed no signs of bark regrowth. The smallest proportion (23.52%) of trees within a size class displaying any regrowth occurs in the 10 to 15m class, and the largest proportion (70%) of trees within a size class showing regrowth is found in the 2 to 5m height class (Fig. 6). There is no significant linear relationship between the amount of bark removed on a tree and the extent of its recovery [R2 =.00809, P>.05] (Fig. 7). There is also no significant linear relationship between the percentage of bark regrowth and the height of a tree [R2 = 0.0058414, P > .05] (Fig. 8). Sclerocarya birrea Population Structure Density of marulas in the total area sampled was 5.12 trees/km2 (Table 1). Of the 128 trees sampled, over half (54%) are in the 8 to 15m size class. Trees in the 2-8m class are nearly absent (Fig. 8). There was a significant difference between the distribution of trees across height classes [G = 87.72, df =3, P < .05]. Only three dead trees were found and they had DBHs of greater than half a meter. None of their deaths could be attributed to elephant damage with certainty. Damage In the sampled population, 82.4% of trees show some sign of damage attributed to elephants. Very little damage is seen on the smallest trees (<2m). The trees in the 2 to 8m class show the greatest extent of damage; all of the trees sampled show either moderate or extra-heavy damage (Fig. 9). All trees in the 2 to 8m height class suffered from primary branch damage while less than 20% of trees in each of the remaining classes displayed this type of damage (Fig. 10). The two larger size classes were most heavily impacted by bark removal; nearly every individual showed some signs of removal. Trees less than 2m in height did not suffer bark removal. There is significant difference in the distribution of all damage classes across height classes [G4 = 103.23, df = 4, P<.05]. However, a Post hoc analysis (Tukey-Kramer Honestly Significant Difference Test) revealed many cases in which the distribution of damage classes show an insignificant difference: light and heavy, light and extra heavy, heavy and moderate, 198 heavy and extra heavy, moderate and extra heavy, and nil and extra heavy (Table 3). Recovery Few individuals (5.32%) with stem damage displayed resprouting (Fig. 11). Resprouting occurred only in the 2-8m and 8-15m classes. All of the trees that had bark damage displayed some degree of bark regrowth (Fig. 12). A greater extent of bark regrowth is seen as the tree size increases. Trees that showed the greatest amount of bark removal (95% bark removed) displayed bark regrowth between 55% and 95%. There is no significant linear relationship between the amount of bark damage a tree suffers and the extent of its recovery [R2 =.016241, P >.05] (Fig. 14). There is also no significant linear relationship between the percentage of bark regrowth and height of a tree [R2 = .03269, P > .05] (Fig. 15).

Discussion The purpose of this study was to examine the proportion and extent of elephant damage on Acacia nigrescens and Sclerocarya birrea in the Punda Maria area in the Kruger National Park. This enables a broader understanding of the utilization of these two species by elephants in this immediate area and a comparison with previous studies done, namely the study by Engelbrecht (1979). Acacia nigrescens The population structure of A. nigrescens in the area sampled shows that the most predominant size class is less than two meters in height, and the amount of individuals decreases readily as the height class increases (Fig. 1). There is an abundance of trees less than two meters that appear to be either saplings or gullivers; further investigation would provide insight into the exact proportion of each. Gullivers are plants, typically stunted multi-stemmed shrubs, which dominate communities as adults but struggle emerge from the herbaceous layer as juveniles due to frequent fires which kill or stunt these plants so that they cannot escape the danger zone (Bond and Wilgen 1996). Trees that are indeed gullivers have been retained in that size class because of frequent burning and herbivory pressure. The presence of a large proportion of these individuals would indicate that fire plays a large role in this area. A high proportion of saplings in the <2m category would imply that A. nigrescens is very capable of recruiting new individuals into the population. Most of the dead trees have a DBH of greater than 20cm (Fig. 2). The cause of death was difficult to ascertain and our figures carry a large degree of uncertainty. We attributed the deaths of 44.66% of the sampled dead population to elephant damage. Van Wyk and Fairall (1969) as cited in Engelbrecht (1979) stated that A. nigrescens is easily toppled by elephants because they have a shallow root system, yet only 0.56% (5 individuals) of our entire sample population were toppled. There is some form of elephant damage on trees across all the size classes but it must be noted that almost 100% of trees less than two meters were untouched by elephants (Fig. 3). There are significant differences between the amount of damage between all the height classes and it can be seen from Figure 3 that as height increases, there is a larger proportion of trees showing any sort of damage. Trees in the five to ten meters height range show the most amount of ‘extra heavy’ damage, but in general most of the damage done to the trees is light and therefore should have no significant negative impact on the survival of the tree. Trees that are greater than 15 meters high have the most damage overall when including light, moderate, heavy, and extra heavy damage by elephants. It is difficult to tell whether damage in the larger height classes is because of accumulation over the years or if the larger trees are indeed being targeted. When examining the predominant type of damage suffered across height classes, it was apparent that the 2- 5m size class had considerable primary branch damage (Fig. 4.). The extent of bark damage increases with increasing height (Fig. 4). According to Engelbrecht (1979), elephants are discouraged from debarking younger trees because the side branches protect the stem and prevent the removal of strips of bark. From personal observations in the field, this did indeed seem the case as young trees seemed to form cage-like structures with their branches. Since these trees are so small they are also less likely to be targeted by larger herbivores because of the low return on effort. A similar impact study was conducted by Engelbrecht (1979) in the Punda Maria area on A. nigrescens trees greater than 6 meters. Over sixty-five percent of the trees were damaged by elephants. We found 69.2% of trees (greater than 6 meters) to be impacted by elephants. Thus there has been very little change over 25 years. However, Engelbrecht (1979) reported only 4.2% of the sampled trees were dead due to toppling or ring barking while we recorded 16.8% dead individuals. Yet these results cannot be directly compared, as Engelbrecht (1979) did not attribute death incurred by 100% main stem breakage as due to elephants and this form of fatal damage was recorded 83.3% of our dead individuals. Engelbrecht (1979) stated that regrowth of bark by A. nigrescens is minimal and not important because even small markings don’t grow closed or recover (Engelbrecht 1979). However, forty-five percent of the bark damaged 199 trees we sampled displayed signs of recovery. The 10 to 15m height class had the highest number of individuals with zero recovery (Fig. 6). This same size class also had the lowest proportion of resprouting individuals (Fig. 5). This perhaps implies that a threshold may exist after which recovery becomes more difficult. Resprouting occurred in 26% of the trees that displayed stem damaged. Resprouting is an efficient strategy by which woody plants regain biomass lost during a disturbance event (Bellingham 200), such as herbivory. Sprouting was more predominant in juveniles (up to 10m), in frequently burnt savannas, sprouting is an essential prerequisite for juvenile survival (Bond and Midgley 2001). Resprouting in A. nigrescens took the form of basal, stem and branch epicormic (Bond and Midgley 2001). Sclerocarya birrea The most largely represented height class was the 8-15m class. Only three trees in the 2-8m height class were found, a size class that has been noted to be notoriously ‘missing’ in other studies (Fig. 9) (pers. comm. Michelle Hofmeyer). Severe browsing by Aepyceros melampus (Impala) on marula seedlings has been noted and may play a large role in seedling mortality. This may contribute to the explanation of this gap in the population structure (Lewis 1987 and Haig 1999 in Jacobs and Biggs 2002). Jacobs and Biggs (2002) also suggest that the fixed triannual burns in Kruger Park between 1954 and 1992 have hampered the establishment and development of marula seedlings into the upper canopy as their suggested fire escape height is between 2.5 and 3m (Jacobs and Biggs 2002). Hofmeyer (pers. comm.) suggests that February burns be implemented in areas where marula populations are of concern. These fires are cooler and do not seem to have as negative of an effect on the trees and allow them to escape the fire trap. Jacobs and Biggs (2002) found that there was a decrease of individuals with increasing height and suggest that the marula population is changing towards a shrub category. However, our study contradicts this finding as 80% of the living population that we sampled was larger than 8 meters. We suggest that concern should center on the cause of the ‘missing size class’ rather than the adult population. Previous studies on marula population characteristics in other nature reserves in Southern Africa have found markedly similar results to that of our study and report unstable population structures with no immature trees and little or no evidence of successful regeneration and recruitment (Walker et al. 1986, Lewis 1987 and Gadd 1997 in Jacobs and Biggs 2002). Out of the 128 S. birrea sampled only three dead individuals were found and they were all approximately ten meters high. Mortality of these trees could not be attributed elephants as none of them were ring barked, toppled, or had 100% main stem damage. Jacobs and Biggs (2000) found that elephant damage is the main cause of mortality among marulas greater than two meters. We found very few dead trees, this suggests that either our sample size and area were too small to assess the population structure, or that the damage the elephants cause is not having a fatal effect on the S. birrea in this area. A study conducted by Coetzee et al. (1979) found that 6.5% of S. birrea trees sampled across the whole park were felled or ring barked. This finding is far greater than the 0.8% of dead trees that we found in the Punda Maria area. We found no felled or ring barked individuals, only trees with 100% primary stem breakage. Jacobs and Biggs’ (2002) findings support our observations and state that ring barking and uprooting is very rare. Every size class that was sampled had some form of damage with the less than two meters class being the least affected and the 2-8m size class being the most affected (Fig 10). However, it must be kept in mind that the latter size class only consists of three individuals. A very small proportion (0.8%) of living trees displayed extra heavy damage and 14.4% of living trees displayed heavy damage. This contradicts Jacobs and Biggs (2000) study that found a ‘significant’ proportion of trees had suffered extreme damage. Most of the damage that was recorded was light damage, so that although 83.6% of S. birrea that were sampled had elephant damage, the degree is of no serious concern. Beyond the 2-8m height class bark removal is more common and occurs on more than 80% of all the trees from 8m to above 20 meters (Fig 11). Bark removal has been noted to be the main type of damage inflicted by elephants on marulas (Jacobs and Biggs 2002). Primary branch damage was not as common as bark removal. It must be noted that most of the damage that was recorded can be considered relatively old damage and new or fresh damage was hardly ever seen other than the occasional tusk mark (personal observation). It would be interesting to do a further study that incorporates the age of scarring. In contrast to A. nigrescens, bark recovery in S. birrea was very common and was noted in all trees that had incurred damage above two meters (Fig 13). Trees greater than 15 meters displayed the most recovery. This could be because they are older and therefore have more resources to put into recovery instead of growth, which would be a priority in the smaller size classes. The presence of wood-boring insects was noted in 21.68% of all trees that had bark removed from them. This may in turn reduce the further regrowth of bark as suggested by Engelbrecht (1979), who stated that cessation of regrowth was found to be strongly associated with the deteriorating condition of exposed 200 wood. Resprouting ability varies with the age or size of a plant and with the type and severity of injury (Bond and Midgley 2003). Juvenile sprouting ability is considered part of the recruitment strategy of a species, whereas adult sprouting behaviour indicates potential persistence (Bond and Midgley 2001). Our study found resprouting by trees between 2 and 15 meters (Fig 12). Resprouting took the form of branch epicormic and stem epicormic sprouting (Bellingham 2000). Proportions of resprouting individuals, however, were relatively low: 33.33% in the 2 to 8m class and 6.45% in the 8 to 15m class. No trees with branch damage above 15 meters were found to be resprouting. This may suggest that trees could reach an age when they lose their ability to resprout (Bond and Midgely 2003). Jacobs and Biggs (2002) found that there was a high percentage of coppicing trees in the 2 to 5m category but we found no incidence of resprouting in this size class. This is most likely because only approximately 3% of trees in this class had any primary branch damage necessitating resprouting for recovery.

Conclusions The aim of our study was to assess the proportion and extent of elephant damage on A. nigrescens and S. birrea and the effects of this impact on the demography of these two species. Large proportions of damage were recorded in both populations but a very low proportion of this damage was severe. A. nigrescens and S. birrea both show immense capability to recover. Survival strategies differ between the two species. Elephants prefer to remove bark from marula trees and in response, these trees have adapted an ability to repair removed bark efficiently. Complete recovery of bark damaged areas was commonly noted. Elephants readily break branches of A. nigrescens and in response, this species resprouts quite readily. It is difficult to comment on densities that we recorded, as similar studies have not been conducted in the Punda Maria area. The studies conducted by Jacobs and Biggs (2002) and Engelbrecht (1979) have been used as a guidelines, however it has been difficult to directly compare our results with theirs. Jacobs and Biggs surveyed the marula population across the entire park while our study focused only on the Punda Maria area, and our methodology differed greatly from Engelbrecht’s. Further studies to increase the understanding of the exact extent of impact that elephants are having on woody species must incorporate fire and other herbivory so that an accurate assessment can be made.

Acknowledgements We would like to thank Angela Gaylard, Michelle Hoffmeyer, Laurence Kruger, Deedra McClearn, Alione Ndlopfu and Godfrey Sekhula.

Literature Cited Bellingham, P. J. 2000. Resprouting as a life history strategy in woody plant communities. OIKOS 00:0. Bond, W.J. and J.J. Midgley. 2003. The evolution ecology of sprouting in woody plants. Int. I. Plant Sci. 164(3 Suppl.):S103-S114. Bond, W.J. and J.J. Midgley. 2001. Ecology of sprouting in woody plants: the persistence niche. Trends in Ecology and Evolution. 16(1): 45-50. Bond, W.J. and B.W. Wilgen. 1996. Fire and Plants. Chapman and Hall. London. Coetzee, B.J., A.H. Engelbrecht and S.C. Joubert. 1979. Elephant impact on Sclerocarya caffra trees in Acacia nigrescens tropical plains thornveld of the Kruger National Park. Koedoe. 22: 39-60. Croze, H. 1974 The Seronera bull problem. 11. The trees. East African Wildlife journal. 12: 1 – 27. Du Toit, J.T., K.H. Rogers and H.C. Biggs. 2003. The Kruger Experience: ecology and management of savanna heterogeneity. Island Press, Washington. Dublin, H.T. 1995 Vegetation dynamics in the Serengeti-Mara ecosystem: the role of elephants, fire and other factors. In: Serengeti II: Dynamics, Management and Conservation of an Ecosystem (Eds A. R. E. Sinclair and M. Norton-Griffith). University of Chicago Press, Chicago, IL. Engelbrecht, A.H. 1979. Olifantinvloed op Acacia nigrescens bome in ‘n gedeelte van die Nasionale Kruger wildtuin. Koedoe. 22: 29-38. Gadd, M.E. 2002. The impacts of elephants on the marula tree Sclerocarya birrea. African Journal of Ecology. 40: 328-336 Jacobs, O. S. and R. Biggs. 2002. The impact of the African elephant on marula trees in the Kruger National Park. South African Journal of Wildlife Research. 32(1): 13-32. Jacobs, O. S. and R. Biggs. 2002. The status and population structure of the marula in 201 the Kruger National Park. South African Journal of Wildlife Research. 32(1): 1-12. Schmidt, E. Lötter, M. McCleland, W. and Burrows, J. 2002 Trees and shrubs of Mpumalanga and Kruger National Park. NBI. Whyte, I. J., H.C. Biggs, A. Gaylord, and L.E.O. Braack. 2003. Policy for the Management of the Elephant Population of the Kruger National Park.

Table 1. Tree densities of Acacia nigrescens and Sclerocarya birrea for each transect and total area sampled Density (trees/km2) Density (trees/km2) Transect Acacia nigrescens Sclerocarya birrea Pafuri rd 16.5 0.9583 Loop 1 2.1 0.4524 Main rd 7.125 0.3438 Loop 2 6.5 0.9 Density of total area 7.698 0.512

Table 2. Comparison between damage classes of Acacia nigrescens using a post hoc analysis (Tukey-Kramer Honestly Significant Difference) * = significant difference Damage Heavy Light Moderate Extra heavy Nil Classes Heavy -3.3168 0.9278 * 2.5526 * 3.7026 * 8.2020 * Light 0.9278 * -1.1841 0.2680 * 1.4180 * 6.2499 * Moderate 2.5526 * 0.2680 * -1.8453 -0.6953 3.9869 * Extra heavy 3.7026 * 1.4180 * -0.6953 -1.8453 2.8369 * Nil 8.2020 * 6.2499 * 3.9869 * 2.8369 * -0.4814

Table 3. Comparison between damage classes of Sclerocarya birrea using a post hoc analysis (Tukey-Kramer Honestly Significant Difference) * = significant difference Damage Classes Light Heavy Moderate Extra heavy Nil Light -2.062 -1.716 1.064 * -3.752 8.786 * Heavy -1.716 -4.067 -1.238 -5.508 6.394 * Moderate 1.064 * -1.238 -4.455 -8.601 3.161 * Extra heavy -3.752 -5.508 -8.601 -17.255 -9.202 Nil 8.786 * 6.394 * 3.161 * -9.202 -3.765

600 500 400 300 200 100

Number of individuals of Number 0 <2m 2-5m 5-10m 10-15m >15m Height classes

Figure 1. Population structure of living Acacia nigrescens

202

80 70 60 other 50 40 100% 30 mainstem 20 toppled 10

Number of dead individuals dead of Number 0 >5cm 5-10cm 10-20cm >20cm Siz e cla sse s (DBH)

Figure 2. Attributed cause of death in Acacia nigrescens

100 Xtra Heavy 80 Heavy 60 Moderate 40 Light population 20 Nil

Percentage of sample 0 <2m 2-5m 5-10m 10-15m >15m Height classes

Figure 3. Intensity of damage across height classes in sample population of Acacia nigrescens

6 5 Bark 4 damage

3 Primary

population 2 branch damage

Percentage of sample 1 0 <2m 2-5m 5-10m 10-15m >15m Height classes

Figure 4. Extent of primary branch damage and bark damage across height classes in Acacia nigrescens

203

100% 90% 80% 70% 60% 50% 40%

individuals 30% 20%

Percentage resprouting resprouting Percentage 10% 0% <2m 2-5m 5-10m 10-15m >15m Height classes

Figure 5. Resprouting in Acacia nigrescens displayed by trees with branch damage

100%

80% 100% 60% 50-99% 40% 1-49% individuals 20% 0%

0% Percentage of bark damaged <2m 2-5m 5-10m 10-15m >15m Height classes

Figure 6. Bark regrowth of damaged Acacia nigrescens. Note: No trees in the <2m class had any bark damage.

100 90 80 70 60 50 40 bark regrowth 30 20 10 0 0 10 20 30 40 50 60 70 80 bark damage

Figure 7. Percentage bark removal vs. extent of bark regrowth in Acacia nigrescens

204

100 90 80 70 60 50 40 bark regrowth 30 20 10 0 0 10 20 height

Figure 8. Percentage bark regrowth vs. height in Acacia nigrescens

80 70

50 40 30

20 Number of individuals of Number 10

0 <2m 2-8m 8-15m >15m Height classes

Figure 9. Population structure of living Sclerocarya birrea

100% R2R2 = = 0.0080904 0.058414 80% P > 0.05 Xtra heavy 60% Heavy

40% Moderate Light 20% Nil 0% Percentage of sample population sample of Percentage <2m 2-8m 8-15m 15-20m Height classes

Figure 10. Intensity of damage across height classes in sample population of Sclerocarya birrea

205

50 45 40 35 bark 30 damage 25 20 primary 15 branch 10 damage 5 0 Percentage of sample population <2m 2-8m 8-15m >15m Height classes

Figure 11. Extent of primary branch damage and bark removal across height classes in Sclerocarya birrea

35% 30% 25% 20% 15% 10% 5% 0%

% resprouting individuals% resprouting <2m 2-8m 8-15m >15m Height classes

Figure 12. Resprouting in Sclerocarya birrea displayed by trees with branch damage

100% 90% 80% 70% RECOVERY 100% 60% 50% RECOVERY 50-99% trees 40% 30% RECOVERY 1-49% 20% 10%

Percentage of bark damaged damaged of bark Percentage 0% <2m 2-8m 8-15m >15m Height classes

Figure 13. Bark regrowth in damaged Sclerocarya birrea. Note: No individuals in the <2m height class had any bark damage.

206

110 100 90 80 70 60 50 40 bark recovered 30 20 10 0 0 10 20 30 40 50 60 70 80 90 100 bark removed

Figure 14. Percentage bark removal vs. extent of bark regrowth in Sclerocarya birrea

110 100 90 80 70 60 50 40 bark recovered 30 20 10 0 0 5 10 15 20 25 height

Figure 15. Percentage bark regrowth vs. height in Sclerocarya birrea

← Site 3:Main Rd. ⁭ Site 4: ⁭ Site 2: Site 1: Pafuri Rd. Loop 2 Loop 1 ⁭

Figure 16. Map of site locations in the Punda Maria area 207 The effects of elephant damage on Adansonia digitata distributions along a slope in the Kruger National Park, South Africa

Category: Second Independent Project Participates: Michael Chazan, Kyle Harris Site: Punda Maria, Kruger National Park, Limpopo Province, South Africa

Key Words: accessibility, Baobab, elephant damage, management policy

Abstract: Seventy-eight baobabs were sampled in the Pafuri area in Kruger National Park (KNP). The percentage damage, percentage rockiness and relative position on the slope for each individual was recorded. Results suggest that baobabs found on rocky, steep, inaccessible slopes had little or no damage, while baobabs found in open accessible areas had significantly more damage. Further data suggests that refugia are comprised of physical characteristics including a percentage of 60% or higher, a slope value of 0.6 or higher and general landscape features such as rocky outcrops and steep valleys. Baobabs sampled showed no signs of mortality and had no excessive damage that they could not recover from. The overall condition of the baobab population in the Pafuri area seems stable and the current KNP management policy regarding elephants should ensure that the baobab population remains stable.

Introduction Elephants (along with fire) have been singled out as one of the main contributors to the decrease in large tree densities within Kruger National Park (KNP) (Trollope et al. 1998). Analyses of fixed-point photographs in KNP has shown that large trees are generally decreasing in numbers and only fewer of the smaller trees were able to escape the high fire frequencies of the past and develop into larger trees (Eckhardt pers comm). Elephants and fire have been held responsible for these trends, as they are the main architects affecting vegetation. Devastation caused by elephants is well documented throughout conservation areas in Africa (Barnes et al. 1994; Swanepoel 1993; Weyerhaeuser 1985) and may be a result of the high density of elephants found within a closed area. Since culling operations were brought to a halt in 1995, elephant numbers have passed the 10 000 mark (Whyte et al. 2003). The increase in elephant densities coupled with vegetation destruction in other parks around Africa, has lead park managers in KNP to implement a new policy. The policy focuses on the extent and intensity of elephant impacts, rather than the number of elephants found in the park (Whyte et al. 2003). The new policy is based on the premise that ecosystems are not static, conditions may fluctuate and population responses are an inherent and desirable attribute which may contribute to biodiversity. Therefore a range of elephant impacts achieved through different densities of elephants in different areas at different times may also be seen as natural and desirable (Whyte et al. 2003). Studies in Tanzania (Barnes et al. 1994) and Zimbabwe (Swanepoel 1993) have shown that elephants have a severe impact on baobab trees (Adansonia digitata) and were directly responsible for a 29% mortality rate in adult trees (Swanepoel 1993). Bark removal by elephants can kill woody plants directly, or may increase their susceptibility to fire or wood boring insects (Jacobs and Biggs 2002). Elephants are particularly fond of the soft, pulpy wood of baobabs and will gouge out large amounts of the wood with their tusks, often leading to the collapse and death of the tree (Weyerhaeuser 1985). In Mana Pools National Park, Zimbabwe, elephant foraging on baobabs was shown to be distinctly seasonal and was related to the relative position of the trees to water. Elephants in the study concentrated their feeding on individual trees, and no significant prevalence of damage was found in the size classes (Swanepoel 1993). Research in Tanzania’s Lake Manyara Park has shown that mortality and damage to baobabs are much lower in parts of the park that are inaccessible to elephants (Weyerhaeuser 1985). Several baobabs found in the Pafuri area are found on koppies that are inaccessible to elephants; these inaccessible koppies may act as refugia for populations of baobabs. The aims of the study would be to investigate the distribution of baobabs across the slope, the extent of damage inflicted on the baobabs, what physical parameters constitute refugia and to investigate the implications of our findings for managers.

208 Methods The study site was conducted around the Pafuri area near the Punda Maria camp of KNP. The fieldwork was carried out from the16th to the 23rd March 2004. Areas with a high concentration of baobabs were located and as many trees as possible were surveyed in each area. Surveys were completed by walking from one tree to the next closest in no specific direction. The data collected for each baobab tree included diameter at breast height (DBH), height, damage, percent rockiness, slope, and accessibility. Elephant damage was determined by observing the first three meters of the trunk and a percentage damage of that area was determined. Damage was only considering when the fibrous tissue of the bark was exposed. Ancient damage was noted but not considered in the analysis. DBH was measured with a tape measure and the height of each tree was estimated. Percent rockiness was estimated within a five meter radius. Only boulders that would cause instability for elephants were considered in the estimation. Plant accessibility to elephants was rated using four rating categories one being the most accessible to elephants and four being inaccessible to elephants. Trees in accessibility class one grew on no or very low slope, no rockiness, and were usually on valley bottoms. Trees in accessibility class two were denoted by Low to medium slope, low to medium rockiness, and usually were on the lower slope of the koppies. Accessibility class three trees were on medium to steep slope, medium to high rockiness, mid or high slope, and many were found in the river valleys. Accessibility class four was completely inaccessible to elephants and is described as Steep slope, high rockiness. Other factors such as rocky outcrops, cliffs, and deep ravines were also taken into account when considering accessibility. A tree described as an accessibility class one or two may have been bumped up an accessibility class if we felt it was necessary. The analysis involved a series of statistical methods. Two correlations were performed between damage and slope, and damage and rockiness. A Kruskal-Wallis ANOVA was completed in order to analyze the difference between the accessibility classes. Size class distribution graphs were created in order to look at variation of damage throughout the different size classes.

Results There is a significant relationship between damage on baobabs and degree of slope (R²=.3051, P <.01). The four baobab individuals found above a slope of 0.6 showed no damage at all, while individuals situated below a slope of 0.6 showed varying degrees of damage (Fig. 1). There is a significant relationship between percentage damage and percentage rockiness (R²=.2144; P<.01). The three baobabs that are found in a rocky area (>60% rockiness) do exhibit less damage than individuals found at a less rocky site. However these three individuals represent a small portion of the total population sampled (Fig. 2). The size-class distribution graph (Fig. 3) shows that there are few individuals (twelve) in the first size-class. All twelve individuals were above a DBH of 0.4 m and no smaller individuals or saplings were found at all. Size- classes two, three, four and five had thirty three, nineteen, seven and seven individuals respectively. The percentage damage and size class graph (Fig. 4) shows that the smallest size class has the least damage (18%). Generally the larger the size class the higher the damage. Baobabs grouped into the various accessibility classes, displayed varying degrees of damage (Fig. 5). Individuals grouped in Class 1 fared the worst, with 58.57% damage inflicted upon them. Individuals in Class 2 had an average of 39.91% damage and individuals in Class 3 had an average of 8.75% damage. Individuals within Class 4 displayed no damage. Statistical analysis showed a significant difference (X²=19.2395; P=.002) in damage between the accessibility classes.

Discussion Steepness of slope acts as an effective barrier for elephants that attempt to feed on baobabs. Baobabs found on a steep incline were less likely to be damaged by elephants compared to baobabs found in the low-lying areas. Rockiness does seem to be an effective deterrent to elephants. Inaccessibility may be increased when baobabs are situated on a rocky substrate that is made up of loose, large boulders. Our data suggests that refugia may be defined by specific physical characteristics. These characteristics include; a slope of above 0.6, a percentage rockiness of above 60% and landscape features including steep valleys and rocky outcrops. However, these areas may never be totally “out of bounds” for elephants, but may be less preferable. Damage to baobabs is also less or completely reduced when baobabs are found in inaccessible regions. Populations of baobabs that are found on a steep slope, a rocky substrate and are generally inaccessible may act as source populations, while baobabs that are found in low-lying areas may act as a sink population (Michelle Hofmeyer pers. comm.). Baobabs in the accessible areas are more likely to be damaged by elephants and seedlings in these areas may 209 be more susceptible to a variety of browsers or non-selective feeders such as impala (Michelle Hofmeyer pers. comm.). However, baobabs sampled in inaccessible areas seem to recover well and seem to cope well with elephant impacts. KNP’s elephant management policy should be sufficient in protecting current baobab populations. Under the new policy the KNP will be divided into six zones which will receive different treatments according to their respective elephant populations. The Pafuri area falls within the Northern Botanical Reserve, which encompasses vegetation types (including A. digitata populations) that should not be placed at risk of excessive elephant populations (Whyte. et al. 2003). Elephant populations will be maintained at a density of one elephant for every 2.86km² and this will amount to a total of 500 elephants in the area (Whyte et al. 2003). A reduction in elephant densities could allow for the regeneration of adult trees and the recruitment of seedlings. However, previous studies (Lewis 1987; Prins and van der Jeugd 1993) points to impala as being the most likely candidate in the suppression of seedling recruitment in the Manyara system, Tanzania. If baobab numbers are decreasing through herbivore pressures, KNP should aim at protecting sink populations found in the low-lying accessible areas. However, baobabs that were sampled and had 100% damage showed no signs of collapsing and no dead baobabs were found. The presence of refugia and the fact that the baobabs seem to recover well from elephant damage means that it is unlikely that the baobab population in the Pafuri area is under threat. Several baobabs are found in inaccessible areas and elephant density in the area is lower than the rest of the park (pers. obs.).

Acknowledgements We would like to thank Laurence Kruger for his invaluable assistance in the field and in analyzing the data. Our appreciation also goes to Lucas Masinga for protecting us in the field and Scott Briscoe and Eric Caldera for their assistance in the field. Big up to you mofo’s.

Literature Cited Barnes, R.F.W., Barnes, K.L. and Kapela, E.B. 1994. The long-term impact of elephant browsing on baobab trees at Msembe, Ruaha National Park, Tanzania. African Journal of Ecology. 4: 177-184. Jacobs, O.S.and Biggs, R. 2002. The impact of the African elephant on marula trees in the Kruger National Park. South African Journal of Wildlife Research. 32(1): 13-22. Lewis. D.M.1987. Fruiting patterns, seed germination, and distribution of Sclerocarya caffra in elephant-inhabited woodland. Biotropica. 19(1): 50-56. Prins, H.H.T,. and van den Jeugd, H.P. 1993. Herbivore population crashes and woodland structure in East Africa. Journal of Ecology. 81: 305-314. Swanepoel, C.M. 1993. Baobab damage in Mana Pools National Park, Zimbabwe. African Journal of Ecology. 31: 220-225. Trollope, W.S.W., Trollope, L.A., Biggs, H.C., Pienaar,D. and Potgieter, A.L.F. 1998. Long term changes in the woody vegetation of the Kruger National Park, with special reference to the effects to elephants and fire. Koedoe. 41(2): 103-112. Weyerhaeuser, F.J. 1985. Survey of elephant damage to baobabs in Tanzania’s Lake Manyara National Park. African Journal of Ecology. 23: 235-243. Whyte, I.J., Biggs, H.C., Gaylard, A. and Braack, L.E.O. 2003. Policy for the management of the elephant population of the Kruger National Park.

Table 1. Accessibility classes Class Description 1 No or very low slope, no rockiness, easily accessible to elephants, usually valley bottoms 2 Low to medium slope, low to medium rockiness, lower slope 3 Medium to steep slope, medium to high rockiness, mid or high slope, river valleys 4 Steep slope, high rockiness, completely inaccessible to elephants

210 100

40 Percentage Damage

0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4

Slope Figure 1. Relationship between slope of the hill and percentage elephant damage on A. digitata individuals (R²=.3051; P < .01).

100

% Damage 40

0 20 40 60 80 100 120 % Ro ckiness

Figure 2. Relationship between percentage elephant damage on A. digitata individuals and percentage rockiness within a five meter radius of each A. digitata individual (R²=.2144; P < .01).

211 35

15 Frequency

0 0-1m 1m-2m 2m-3m 3m-4m 4m+

Size Class (DBH)

Figure 3. Size class frequency (by DBH) for the total number of A. digitata individuals sampled in the Pafuri District, KNP.

% Damage 30

0 0-1m 1m-2m 2m-3m 3m-4m 4m+ Size Class (DBH)

Figure 4. Histogram showing the average percentage damage for the five size classes of A. digitata individuals.

212 90

60 e

40 % Damag 30

0 1234 Accessibility Class

Figure 5. Histogram showing the percentage elephant damage for each of the four accessibility classes of A. digitata individuals (X²=19.2395; P=.002).

213 Herbivore density, impala group size, and vigilance of herbivores while feeding

Category: Independent Project Participants: Blanchie Asberry, editor, Deedra McClearn (resource person) Site: Punda Maria, Kruger National Park, Mpumalanga Province, South Africa

Key Words: group size, herbivore, impala, mixed feeders, vigilance

Abstract: There are three types of herbivorous feeders: browsers, grazers, and mixed feeders. The habitats they feed in determine whether they feed with their heads lowered or raised. The mixed feeder, Aepyceros melampus (impala) must be vigilant when feeding. My hypothesis that individuals in smaller herds would be more vigilant than larger herds was supported by my results of a significant negative correlation between group size and the number of times an individual looked up. My other findings are that there is more game in the south of Kruger National Park versus the north and that there is no significant difference in group size of herds based on location (north versus south), and vigilance of herds based on location.

Introduction Browsers, grazers, and mixed feeders differ in feeding habitats which in turn determines whether they feed with their heads up or down and the amount of time they spend being vigilant while feeding. Browsers tend to feed in dense, bushy areas and usually feed with their heads raised. Grazers tend to feed in open areas with their heads down. Because mixed feeders eat in the same types of habitats that browsers and grazers eat in, their feeding behavior would depend on what they are eating. The mixed feeder, Aepyceros melampus became the focus of my study because they are quite common in the park and are easy to observe. These herbivores must be vigilant when feeding to prevent predation and they are a major food source to several of the park’s carnivores. Vigilance is important for many reasons. Mainly and most importantly it plays a large role in prey avoiding the risk of predation. It is clear that predator detection is a major function in many species (Roberts 1996). According to the ‘many eyes’ hypothesis, by taking advantage of the vigilance of other group members, individuals can reduce their own vigilance. Therefore, this allows individuals an increased time to feed. The group size of feeders should play an important role in their feeding behavior. The more animals in a group, the lower the predation risk for each individual. Many birds and mammals respond to a heightened risk of predation, especially that associated with smaller group sizes, with an increase in vigilance (Lima et al. 1999). A reduction in individual vigilance with an increase in group size is one of the most frequently reported relationships in the study of animal behavior. There is evidence for a direct relationship between group size and vigilance where other variables have been controlled (Roberts 1996). As an additional part of my project, I was interested in game density in the northern and southern parts of the park. It has been said that more game is found in the South of the park. If there is more game in the south, this could also affect impala group size and the vigilance of individuals in the north and the south. The aim of my study was to answer four related questions. Is there really more game in the south? In regards to impala, are smaller herds more vigilant than larger herds? Is there are significant difference of vigilance of herds due to location? Is there a significant difference of group size of herds due to location?

Methods This study was carried out in Punda Maria (north) and Skukuza (south) of Kruger National Park, Mpumalanga Province, South Africa. The study was conducted on 21-24 March 2004 in Punda Maria and 1-2 April 2004 in Skukuza. On 23-24 March there was light and heavy rain. In Punda Maria ten 25 km transects were driven and in Skukuza five, 25 km transects were driven. For each 25 kilometer driven transect, different routes were taken to ensure that the same animals were not counted twice. Along each transect the presence of all large mammals was noted. Once a group of impalas was spotted, I counted the number of individuals within the herd and noted the vegetation type they were feeding in. A focal animal that was feeding was then chosen to be observed. This observation was done in increments of two minutes. With a tally counter, I counted the number of times that an animal looked up while feeding. If for any reason the focal animal did not feed consistently for two minutes, data collection from the animal was abandoned and another animal was chosen. A Mann-Whitney U test was performed to analyze the number of sighting of groups of animals based on location, group size and location, and vigilance of herds based on location. A linear regression was made to analyze 214 vigilance of impala in smaller groups versus larger groups.

Results There are more groups of animals in south of the park versus the north (p = .01), although there was no significant difference between the group size and location, (p = .10, Figure 1). Also, there are more groups of impala in the south versus the north, although there was no significant difference between group size and location, p = .006 (Figure 2). When only observing impalas, smaller groups spend more time being vigilant than larger groups. There was a significant correlation between group size and the amount of times they looked up, although the R2 = .11 was not very high, p=.004. Vigilance between herds in the south and the north was not significantly different, p=.25. When analyzing all groups of herbivores, vigilance between herds in the south compared to the north was not significant, p=.29.

Discussion As of now, it is still unclear as to why more game is found in the south of the park versus the north. This has however influenced tourism within the park so that tourist could choose to stay in the southern part of the park. There is no significant difference in impala group size between sites and there was such a high variance to the number of individual groups at each site. The number of individuals within each herds ranged from three to 75. My hypothesis about smaller groups of impala being more vigilant than larger groups was supported by the results of the linear regression performed between group size and the number of times an animal looked up (vigilance) (Figure 3). Although the r2 was not high, this could have been due to other contributing factors such as vegetation type, etc. From the data collected many herds of impala were found in open areas; however, even more herds of impala in which data was unable to be collected were observed in dense brush. Data was unable to be collected from some animals because they would always feed consistently and sometimes tended to quite nervous or jumpy. As for vigilance of herds based on location, it is also unclear why there is no significant difference; however, again, this could be due to other contributing factors not observed (Figure 4).

Acknowledgments I would like to thank Deedra McClearn for taking out the time to drive me around, assisting with data collection and data analysis, and revision of several things throughout this study. Also, I would like to thank Kinesh Chetty for driving me around and assisting with data collection. Next, I would like to thank Godfrey Sekhula for assistance with data collection and taking pictures. Then I would like to thank Laurence Kruger for his assistance with the idea for this study.

Literature Cited Lima, S.L. and Bednekoff, P.A. 1999. Back to the basics of antipredatory vigilance: can nonvigilant animals detect attack?. Animal Behavioral. 58, 537-543. Roberts, G. 1996. Why vigilance declines as group size increases. Animal Behavior. 51:1077- 1086.

215 80

30 GROUP SIZE 20

0 PM SKU

LOCATION

Figure 1. All groups: The size of the herds are not significant, p = .10.

10 number of sightings 5

0 PM SKU

location

Figure 2. Impala: More groups in the South versus the North, p = .006.

216 11

LOOKING UP 3

-1 0 10 20 30 40 50 60 70 80 NUMBER IN GROUP

Figure 3. Impala: Smaller groups are more vigilant than larger groups, R2 = .11. p = .004.

LOOKING UP 3

-1 PM SKU

LOCATION

Figure 4. Impala: No significant difference of vigilance between herds in the North versus the South, p = .25.

217 The effect of overlapping piospheres on landscape heterogeneity

Category: Independent Project Participans: Sally Koerner and Zoe Layton (Secretaries), Abri de Buys Site: Punda Maria, Kruger National Park, Mpumalanga Province, South Africa

Key words: decreaser, grass species, heterogeneity, increaser II, piosphere, water hole

Abstract: In 1933, Kruger National Park implemented artificial sources of surface water. Many studies have been conducted on the effect these waterholes have on herbivore distribution and the related impacts. One such finding is that piospheres, patches created by herbivores through their grazing, browsing and trampling activities focusing around a water source (Owen-Smith 1996 cited in Gaylard et al 2002) occur around waterholes and contribute to the patchiness in the landscape. The aim of this study was to determine if the proportion of increaser II grass species would drop below 50% and be replaced by decreaser grass species at a shorter distance away from the closer spaced water holes than from the isolated water hole and to determine if the vegetation between two closely spaced water holes will be dominated by increaser II grass species. Two sites were selected, one with an isolated water hole and one with two closely spaced water holes. Transects were preformed at increasing distances from each water hole, and grass species were recorded along these transects at 2m intervals. Our data supported our hypotheses. We found that transects going away from an isolated water hole and transects going away from two closely spaced water holes changed from a system dominated by increaser II grasses to a system dominated by decreaser grass species. The transects running between the two closely spaced water holes, however, never became dominated by decreaser grass species. The area between the two closely spaced water holes was an example of increased homogeneity on a small scale; however, when looked at on a large scale the space in between the water holes creates a distinct patch in the ecosystem that is supportive of a certain type of grazer.

Introduction In 1933, Kruger National Park implemented artificial sources of surface water for three main reasons, namely, to stabilize existing natural supplies of water, to provide additional artificial water supplies in areas where natural supplies had previously existed, and to construct dams for periods of intense drought (Pienaar 1970 cited in Gaylard et al 2002). In recent years, many studies have been conducted on the effect these waterholes have on herbivore distribution and the related impacts. One such finding is that piospheres, patches created by herbivores through their grazing, browsing and trampling activities focusing around a water source (Owen-Smith 1996 cited in Gaylard et al 2002) occur around waterholes and contribute to the patchiness in the landscape. The sizes of these piospheres differ depending on the distribution of water points relative to each other (Gaylard et al 2002). Davidson showed in her 1996 study that herbaceous vegetation composition changes with increasing distance from water points. The grass layer also responds in certain ways to different grazing intensities, the most obvious of which is the characteristic sacrifice zone in the immediate vicinity of the water point (Thrash 1998 cited in Gaylard et al 2002). Sacrifice zones are normally heavily overgrazed and can sometimes be completely devoid of grass. The impact slowly becomes less intense as distance from water increases. Grasses that are able to tolerate high intensity herbivore impact (increaser II) occur in greater abundance closer to water while species that are tolerant of more moderate impacts (decreaser) occur further away from the water point. At a certain distance from the water the proportion of increaser IIs and decreaser species in the grass layer are equal (50%) (Davidson 1996). The distance at which this 50:50 ratio occurs, has been used to determine the recovery of the grass layer after the closure of waterholes (Davidson 1996). Closely spaced water holes allow herbivores to move over the entire landscape because they will never be far from a water source, leading to a homogenous impact pattern (Gaylard et al 2002). This study aims to investigate whether this pattern is found in the grass layer between two closely spaced water points on the Northern Plains of the Kruger National Park and compares the size of piospheres around closely spaced versus isolated water points. Our hypotheses were that the proportion of increaser II grass species will drop below 50% and be replaced by decreaser grass species at a shorter distance away from the closer spaced water holes than from the isolated water hole and that the vegetation between the two closely spaced water holes will be dominated by increaser II grass species (i.e. homogenous). 218 Methods Using Arcview two sites were selected. One of the selected sites consisted of two water holes spaced 1.67km apart. These two water holes were Mandadzidzi and Elandskuil. The other site comprised a single, isolated water hole named N'warihlangari (Figure 1). Grass species composition was recorded along transects perpendicular to the water holes. Transects were placed 0m, 100m, 200m, 400m, 600m and 800m away from the water holes. At each water hole two series of transects were performed in opposite directions from the water hole. At the N'warihlangari water hole, one of the series of transects required an additional transect 950m away from the water hole because the 50:50 ratio of increasers to decreases had still not been reached at the 800m transect. At the Mandadzidzi water hole, the series of transects going away from the two water holes, as opposed to towards the Elandskuil water hole, was stopped early. The last transect was performed 400m away from the water hole because the 50:50 ratio was reached around 200m away from the water hole. The series of transects starting at the Elandskuil water hole and going away from the two water holes, as opposed to towards the Mandadzidzi water hole, was also completed early. The last transect was performed at 600m because the 50:50 ratio had been reached. At the site with these two closely spaced water holes, two of the transect series met at the middle, thus, ultimately creating one 1600m transect running between the two water holes. Along all transects, grass species were recorded at 2m intervals for a distance of 100m. After the grass species were identified they were classified as either increaser IIs or decreaser species according to Trollope (1990). The grass species were also identified as either low palatability, average high palatability or high palatability. The data were then manipulated in Excel and Jump. In Jump, we preformed a Mann-Whitney U test to compare the palatability of the grass species around Mandadzidzi to the palatability of the grass species around Elandskuil.

Results Around the isolated water hole, N'warihlangari, we found that at both transects 0m from the water hole and intersecting the sacrifice zone there were very few decreaser grass species. The 0m transect on one side of the water hole comprised seven spots of bare ground, one decreaser grass species (Cenchrus ciliaris), and the rest Urochloa mosambicensis, a dominant increaser II grass species. The 0m transect on the other side of the water hole comprised nine spots of bare ground, no decreaser grass species and the rest Urocholoa mosambicensis. The subsequent transects in the series showed a gradual increase in the percentage of decreaser grass species (Table 1, Figure 2). The first transect series reached the 50:50 increaser to decreaser ratio at approximately 800m where it was comprised of 98.0% decreaser grass species. The second transect series reached the 50:50 increaser to decreaser ratio at a slightly further distance of approximately 950m where is what composed of 86.3% decreaser grass species (Table 3, Figure 2). At the second site with two closely spaced water holes, the transect series moving away from the water holes reached the point of fifty percent decreaser grass species at a more rapid rate (Table 3, Figure 3). The transect series moving away from the Elandskuil water hole had 3.9% decreaser species on the 0m transect, 15.7% decreaser grass species on the 100m transect, and then quickly erupted to 41.2% decreaser grass species on the 200m transect. The 50:50 increaser to decreaser ratio was achieved at approximately 600m and this transect was composed of 74.5% decreaser grass species. The transect series moving away from the Mandadzidzi water hole had 5.9% decreaser grass species on the 0m transect, and then rapidly increased to 37.3% decreaser grass species on the 100m transect. The percentage of fifty percent decreaser grass species was achieved at approximately 200m and this transect was composed of 62.7% decreaser grass species. The transects in between the two closely spaced water holes never achieved the 50:50 increaser to decreaser ratio (Figure 3). Of the twelve transects performed between Elandskuil and Mandadzidzi, the average percentage of decreaser grass species was 19.6%. The six transect series were divided into three types of transect series; transects going away from an isolated water hole, transects going away from two closely spaced water holes, and transects running between two closely spaced water holes (Figure 4). The slope of the average of the two transects running between the two closely spaced water holes is .0208 was much smaller than the slopes of the transect groups running away from water points. This shows that the rate change from increaser II grass species to decreaser grass species was much smaller. In fact, the slope is so small that the rate of change is almost not existent. The slope of the average of the two transects going away from the isolated water hole and the slope of the average of the two transects moving away from the closely spaced water holes are very similar; .0894 and .8628 respectively. The y-intercepts of the two sets of transect series are very different, however. The y-intercept of the average of the transects moving away from the isolated water hole is –3.8771 while the y-intercept of the average of 219 the transects moving away from the closely spaced water holes is 15.095 (Figure 4). This shows that the rates of change from increaser II grass species to decreaser grass species is not much different but that the change from increaser II grass species to decreaser grass species begins earlier moving away from the closely spaced water holes than moving away from the isolated water holes The percentage of highly palatable grass species around water holes was also taken into account. Around the isolated water hole we found that there was generally high palatability at all distances around the water hole however the highly palatable grass species close to the water hole were mostly increaser II grass species while the highly palatable grass species further away from the water hole were mostly decreaser grass species (Figure 5). Around the closely spaced water holes we found only that there was a significantly higher amount of highly palatable grass species around Elandskuil than around Mandadzidzi (Figure 6). We found this significance by running a Mann- Whitney U-test (p=.0117).

Discussion: Our data support the hypotheses that the proportion of increaser II grass species would drop below 50% and be replaced by decreaser grass species at a shorter distance away from the closer spaced water holes than from the isolated water hole, as the 50:50 increaser to decreaser ratio was reached much earlier on transects moving away from the closely spaced water hole. At the transect series going away from the Elandskuil water hole the decreaser grass species dominated the transect at 600m with 74.5% decreaser grass species. It is likely that the decreaser grass species would have dominated at a closer distance, but the presence of a seasonally flooded pan between the 200m and 400m transects may have resulted in the persistence of more increaser II grass species. At the transect series going away from the Mandadzidzi water hole the decreaser grass species dominated at a point even closer to the water hole. At the isolated water hole the 50:50 increaser to decreaser ratio was reached at a transect that was much further from the water hole. This shows that, as expected, the overutilized grass species have a shorter radius when going away from the closely spaced water holes. Thus, the piospheres around the closely spaced water holes are smaller in area. Our data also supported the hypothesis that the vegetation between two closely spaced water holes would be dominated by increaser II grass species thus making it more homogeneous. Unlike the transects going away from Elandskuil and Mandadzidzi, the transects in between the two closely spaced water holes never reached the 50:50 increaser to decreaser ratio. The percentage of decreaser grass species only reached 45.1%. This transect is an outlier, however, because the next highest percentage of decreaser grass species was 29.4%. This outlier, at 400m away from Elandskuil, most likely occurred because of a thicket where shade-loving decreaser grass species such Panicum maximum and Sporobolus fimbriatus often thrive. It has been noted that both P. maximum and S. fimbriatus prefer shade under trees and shrubs (Oudtshoorn 1999). Since five of the decreaser grass species on this transect were either P. maximum or S. fimbriatus, it is clear that the percentage of decreaser grasses on this transect were affected by the shade of trees and shrubs. If we were to discount these five grasses, the percentage of decreaser grass species on this transect would be 35.3% decreaser grass species rather than 45.1%, making the percentages of decreaser grass species in between the closely spaced water holes more uniform. When analyzing the palatability data we found that the percentage of highly palatable grass species was concentrated around the isolated N'warihlangari water hole. The percentage of highly palatable grass species was high in transects around the water hole due to high concentrations of U. mosambicensis. Similar trends emerged when looking at the palatability of grass species between and around the two closely spaced water holes. What was of more striking interest however, was that the grass species around Elandskuil had much higher palatability than the grass species around Mandadzidzi. Further testing is necessary in order to determine possible factors, such as the presence of the road dividing the two water holes possibly serving as a fire break, that may influence higher palatability around the Elandskuil water hole. Several studies in the past have shown that herbaceous composition changes with increasing distance from water points (Davidson 1996). Our study supported this with the grass layer. We found that transects going away from an isolated water hole and transects going away from two closely spaced water holes changed from a system dominated by increaser II grasses to a system dominated by decreaser grass species. Transects running between the two closely spaced water holes, however, never became dominated by decreaser grass species. The area between the two closely spaced water holes was an example of increased homogeneity on a small scale as the area is dominated by increaser II grasses and never reaches a percentage of decreaser grasses greater than 50%. On a larger scale however, the space in between the water holes creates a distinct patch in the ecosystem that is supportive of a certain type of grazer. This is a very complex phenomenon that we plan to follow up in the next couple of months. We plan to assess whether such patches are useful to the ecosystem or detrimental to selective grazers such as the rare roan antelope. 220 This fits into the biodiversity management objective in the Kruger management plan regarding terrestrial research on herbivory with hopes of monitoring spatial patchiness and herbaceous vegetation (Biggs and Rogers 2002).

Acknowledgements We would like to thank Abri du Buys for all his invaluable help, for protecting us in the field, for driving us around, for getting us out of the mud, and for not panicking when we lost our data. We would also like to thank Julie Coetzee for showing us how to love grass and for all her wonderful knowledge that she shares so freely. Thanks to Laurence Kruger and Deedra McClearn for all their help and time as well. Ons doen wat ons wil.

Literature Cited Biggs, H. C., K. H. Rogers. 2002. An adaptive system to link science, monitoring, and management and practice. Page 64 in J.T. Du Toit, K. H. Rogers, H.C. Biggs (ed.), The Kruger Experience: Ecology and Management of Savanna Heterogeneity. Washington: Island Press. Gaylard, A., Owen-Smith, N. and Redfern, J. 2002. Surface water availability: implications for heterogeneity and ecosystem processes. Pages171-188 in J.T. Du Toit, K. H. Rogers, H.C. Biggs (ed.), The Kruger Experience: Ecology and Management of Savanna Heterogeneity. Washington: Island Press. Davidson, T.M. 1996. Recovery of the herbaceous vegetation around three closed water holes on the northern plains of the KNP. Unpublished B.Sc. honors thesis, University of the Witwatersrand, Johannesburg, South Africa. Oudtshoorn, Frits. 1999. Guide to Grasses of Southern Africa. Pretoria: Briza Publications.

Table 1. Percentage of decreaser grasses around an isolated water hole, N'warihlangari. Distance from Waterhole (m) # Bare Ground # Increasers #Decreasers %Decreasers 800 0 1 50 98.0 600 0 36 15 29.4 400 0 34 17 33.3 200 0 37 14 27.5 100 0 48 3 5.9 N'warihlangari 7 43 1 2.0 N'warihlangari 9 42 0 0.0 100 0 48 3 5.9 200 0 47 4 7.8 400 0 43 8 15.7 600 0 29 22 43.1 800 0 25 26 51.0 950 0 7 44 86.3

221 Table 2. Percentage of decreaser grasses around two closely spaced water holes, Elandskuil and Mandadzidzi. Distance from waterhole (m) # Bare Ground # Increasers # Decreasers % Decreasers 600 0 13 38 74.5 400 0 28 23 45.1 200 0 30 21 41.2 100 0 43 8 15.7 Elandskuil 5 44 2 3.9 Elandskuil 3 46 2 3.9 100 0 47 4 7.8 200 0 41 10 19.6 400 0 28 23 45.1 600 0 41 10 19.6 800 0 36 15 29.4 800 0 39 12 23.5 600 0 40 11 21.6 400 0 44 7 13.7 200 0 40 11 21.6 100 0 38 13 25.5 Mandadzidzi 3 46 2 3.9 Mandadzidzi 3 45 3 5.9 100 0 32 19 37.3 200 0 19 32 62.7 400 0 21 30 58.8

Table 3. Approximate distance that fifty percent decreaser grass species was achieved. Water hole Direction from water hole Distance (m) Percentage decrease grass species N'warihlangari away 800 98.0% N'warihlangari away 950 86.3% Elandskuil away 600 74.5% Elandskuil towards Mandadzidzi not Mandadzidzi away 200 62.7% Mandadzidzi towards Elandskuil not

222 Map of Northern Plains with piospheres at study sites.

# ðð # ð ð ## ð ð # ð ð ð b b Mandadzidzi ## # ð ð ð # ð ð ## Elandskuil ð ð ð ð# b ð ð ð ## ð ðð ðð ð #

# ## b ð ð ð ð # # # #

## ## #

# # b b # ## # # bb b # b # # # # #b b #

## ð b # ##

N'warihlangari # ## b # # b b ð # # #

0 102030Kilometers

b Springs.shp ð Pans.shp # Boreholes.shp N Roads.shp Disused Jeep / 4x4 Road Management / Firebreak Road W E Need GPS Other Private / Concession Road S Rest Camp Roads Tourist Gravel Road Tourist Tar Road Boundary.shp

Figure 1. Map of the study site.

223

Figure 2. Percentage of decreaser grasses around an isolated water hole.

Figure 3. Percentage decreasers around two closely spaced water holes.

224

Figure 4. Average rate of recovery in several transect series.

Figure 5. Percentage of highly palatable grass species with increasing distance from an isolated water hole.

225 Appendix 1. Species composition for isolated water hole. N'war 1 N'war 2 Species Category 0m 100m 200m 400m 600m 800m 0m 100m 200m 400m 600m 800m 950m Ari con bar I2 1 Ari con con I2 1 3 Bot rad I2 5 1 Bra def I2 1 Cen cil D 1 2 5 6 2 Dig eri D 4 1 Ech col I2 Enn cen I2 8 6 4 2 2 Era rac I2 Era sup I2 2 1 1 4 3 9 3 Era tri I2 Fin afr D 5 5 Het con I2 4 2 2 Isc afr D 3 Pan col D 16 1 1 2 Pan max D 1 9 6 6 3 27 2 5 5 13 24 Sch pap I2 12 16 13 1 6 1 Set inc D 30 1 1 1 16 8 13 Set sph D 4 Spo fim D The tri D 1 2 Tri mon I2 2 Uro mos I2 43 48 12 2 12 42 23 40 29 19 22 7 Uro oli D Uro pan I2 Bare ground 7 9 # of Species 2 3 8 9 10 5 1 3 7 9 6 7 5

226 Appendix 2. Species composition for Mandadzidzi. Mand 1 Mand 2 Species Category 0m 100m200m 400m 0m 100m200m 400m 600m 800m Ari con bar I2 3 1 2 1 Ari con con I2 Bot rad I2 2 1 2 1 1 Bra def I2 Cen cil D 1 2 2 1 2 4 1 Dig eri D 6 11 4 1 1 3 Ech col I2 1 Enn cen I2 Era rac I2 Era sup I2 20 16 8 11 28 28 34 23 18 9 Era tri I2 7 Fin afr D Het con I2 1 2 2 4 3 4 1 Isc afr D Pan col D 5 9 16 1 3 1 5 4 Pan max D 4 3 9 4 7 2 3 Sch pap I2 1 1 1 2 6 4 2 Set inc D 1 1 2 Set sph D Spo fim D The tri D 6 5 3 1 3 Tri mon I2 Uro mos I2 22 15 7 4 17 3 1 9 12 21 Uro oli D 2 Uro pan I2 2 Bare ground 3 3 # of Species 6 7 11 10 5 10 7 7 10 9

227 Appendix 3. Species composition for Elandskuil. Eland 1 Eland 2 Species Category 0m 100m200m 400m 600m 0m 100m200m 400m 600m 800m Ari con bar I2 4 10 8 Ari con con I2 Bot rad I2 1 2 3 1 4 4 Bra def I2 Cen cil D 1 3 3 5 8 1 2 2 Dig eri D 8 15 14 2 1 1 Ech col I2 Enn cen I2 Era rac I2 1 Era sup I2 8 6 11 1 1 9 12 1 7 9 10 Era tri I2 1 3 3 9 3 2 4 1 5 4 Fin afr D Het con I2 5 1 2 1 Isc afr D Pan col D 1 4 2 1 8 4 1 14 6 13 Pan max D 1 3 1 6 3 2 1 Sch pap I2 9 5 11 2 11 1 7 Set inc D Set sph D Spo fim D 2 2 The tri D 3 5 1 Tri mon I2 Uro mos I2 34 19 3 6 6 35 34 25 16 10 6 Uro oli D 2 2 2 Uro pan I2 Bare ground 5 3 # of Species 6 9 12 9 11 4 4 8 10 10 9

228 The impact of vegetation disturbance around water holes on rodent abundance and diversity in the north of Kruger National Park, South Africa.

Category: Independent Project Participants: Tammy Baudains, Taryn Morris and Govan Pahad Site: Punda Maria, Kruger National Park, Mpumalanga Province, South Africa

Key words: Disturbance, diversity, Kruger Park, monitoring, piospheres, rodents, Sherman traps

Abstract: Many studies have demonstrated a causal relationship between disturbance and species diversity and furthermore between losses in diversity and declines in ecosystem function, resilience and resistance. Our study supports this and additionally the notion that monitoring of small mammal diversity can be used as a quick and inexpensive indicator of ecosystem integrity. The movement of herbivores to and from waterholes results in an increasing gradient in their grazing, browsing and trampling impacts towards waterholes, creating circular zones of impact known as piospheres. An increase in artificial waterholes in the past has resulted in an overlap in these impacted areas, creating a homogenised, disturbed habitat. The aim of our study was to determine if the vegetation structure and composition is indeed impacted in the piosphere and whether this affects rodent diversity and abundance. We tested an open waterhole and one which had been closed for about two years. We found similar number of rodents at both waterholes. However the open waterhole had a lower diversity of rodents than the closed waterhole and was dominated by the multimammate mouse, Mastomys natalensis, which in other studies has been associated with disturbed areas.

Samevatting: Baie studies het die verhouding tussen steuring en species verskeidenheid verklaar. Verhoudings tussen verlies van veskeidenheid in ekosiesteem funksie, elastisiteit en teëstand was ook gevind. Ons studie ondersteun dit en ook die begrip dat afluistering van klein soogdier verskeidenheid kan as ‘n vinnig en goedkoop wysen van systeem volledigheid gebruik word. Die beweging van plantetend diere heen en weer van drinkplekke volk in ‘n toename helling in hul skraming, grasing en vertrapping versteuring teen drinkplekke. Hierdie versteuring skep sirkulêre sone van steuring wat “piospheres” genoem is. ‘n Vermeerdering in kunsmatige drinkplekke in die gelede het in ‘n gedeeltelike dek van hierdie versteurde sone en ‘n homogenisere, gesteurde habitat was geskep. Die doel van ons studie was om te besluit of die plante gebou en komposisie regtig geslag binne die “piospheres” was en of dit die knaagdier afluistering en oorvloed beinvloed. Ons het twee drinkplekke getoets. Een was twee jaar gelede toegemaak en die ander is oop. Ons het ‘n ooreenkomstige nommer van knaagdiere by albei drinkplekke gevind. Egter het die oop drinkplekke ‘n laer verskeiheid knaagdiere as die toe drinkplek. Die oop drinkplek was by die muis Mastomys natalensis gedomineer. Hierdie spesies is met versteuring bybehorende.

Introduction The availability and distribution of water systems can influence ecosystem structure and function at a range of scales and organizational levels by influencing processes and feedbacks affecting both plants and animals. However, our understanding of these relationships is limited and has not been considered within a hierarchy of scales (Gaylard et al. 2003). The movements of herbivores to and from waterholes results in an increasing gradient in their grazing, browsing, and trampling impacts towards waterholes, creating circular zones of impact known as piospheres (Gaylard et al. 2003). The most severely impacted zone occurs within 200–300m of a water source (Gaylard et al. 2003). Piospheres were traditionally viewed as undesirable due to the changes brought about in vegetation structure and composition. However with respect to mulitscale approaches, piospheres represent patches of contrasting vegetation and therefore contribute to heterogeneity. Increasing the number of water points may have affected this heterogeneity by allowing piospheres to overlap creating a more homogeneous landscape (Gaylard et al. 2003). The policy of establishing artificial water points was adopted in the Kruger National Park to support herbivore populations during droughts. Specifically, it aimed to provide an adequate network of reliable water points through construction of borehole and earth dams (Gaylard et al. 2003). Although this policy may have seemed reasonable, the outcomes were not as desired. Populations of rare antelope such as roan, tsessebe, sable, and reedbuck fell to dangerously low levels during the extended drought period of 1982–1987 (Grant 1999, Western 229 1975). Kruger has subsequently altered its water policies resulting in the closure of many artificial water points (Gaylard et al. 2003). Many studies have been carried out to test the effects of artificial water sources on rare or charismatic large herbivore species. However, little work has been done on the indirect impacts of grazing on small mammal communities (Salvatori et al. 2001) Small mammal community structure and species richness have been related to variables such as habitat structure, area, predation, trampling and grazing, among others (Avenant and Kuyler 2002). Many studies have shown that fewer species are found in more disturbed areas, indicating that a causal relationship exists between disturbance and species diversity, and between losses in diversity and declines in ecosystem function, resilience and resistance (Avenant and Kuyler 2002). Disturbance in rodent communities may have consequences for the whole ecosystem and can affect populations of medium sized predators (both mammals and birds) (Salvatori et al. 2001). Thus monitoring of small mammals may provide an important link to general ecosystem functioning. Thus, not only is this type of study essential for conservation and for the understanding of small mammal biodiversity in terrestrial ecosystems, but direct monitoring of small mammals may also be used as a relatively quick and inexpensive method of indication of ecological disturbance (Avenant and Kuyler 2002). Our study examined the effects of disturbance on vegetation and hence rodent populations. Our first hypothesis is that the vegetation at the open waterhole is in fact more disturbed than the closed waterhole. Our second hypothesis that there will be a higher rodent diversity around a waterhole that has subsequently been closed than around a waterhole that has remained open since construction. Our third hypothesis is that there will be a higher number of rodent individuals caught around a closed waterhole than around an open waterhole.

Methods Sampling was undertaken on four consecutive nights from 21 to 24 March 2004. March is considered near the end of the breeding season of small mammals (Monadjem and Perrin 1997). Two waterholes were sampled in the Punda Maria region of the Kruger National Park (Figure 1). One waterhole, which we named Elephant waterhole, had been closed approximately two years prior to our study. The other, Elandskuil waterhole, had remained open since construction. The vegetation was sampled along two transects of 50m in each of the two grids at each waterhole to investigate correlations with rodent diversity. Percentage grass cover was assessed every 10m along the transect by randomly placing a 0.5x0.5m quadrate on the ground and estimating the percentage grass cover within. Vegetation composition was determined by recording the species present at points every 2m along each transect. Structure was sampled by determining foliage profiles using boards 1.5m high and 10cm wide, divided into height classes of 25cm. At five points (every 10m) along each transect, we measured the distance from the point at which 50% of each height class on the board was obscured. The reciprocal of this distance is directly proportional to foliage density and is thus an approximation of the vertical distribution of biomass. From these values, a foliage profile of each site was constructed. Sherman traps were set in two 5 by 6 grids of 30 traps spaced ten metres apart on opposite sides of each waterhole. At the open waterhole, grids were laid approximately 20m away from the waterhole so that the absence of vegetative cover due to excessive over grazing and trampling would not affect our results. Traps were left open day and night and were checked and rebaited in the early mornings and late afternoons. Traps were baited using a mixture of peanut butter, rolled oats, and whole seeds including corn and sorghum. Rodents trapped were identified to species level, sexed, weighed and were marked on the third day by cutting a patch of fur off the belly), before being released in the same area. Reproductive status was assessed by males either being scrotal or not scrotal and females being nulliperous or sexually active. This was checked by feeling whether or not the pubic symphasis was open or closed (many females were pregnant or suckling and so were easily identified as sexually active). Data were analyzed using Microsoft Excel 2000, Primer 5, JMP IN 5.1 and Statistica 6. The mean percentage vegetation cover was derived for the open and closed waterhole and graphed. Additionally a student’s t- test for independent samples was performed in Statistica 6 to determine whether the difference in percentage cover was significant or not. The vegetation species composition at the four open waterhole transects and four closed waterhole transects were compared by producing a cluster analysis in Primer 5. A foliage density graph was created to illustrate the vertical distribution of biomass at each waterhole. Additionally student t-tests for independent samples were performed on 0.25m, 0.5m and 0.75m height classes to determine if there was significant difference in the amount of foliage at those heights. The rodent captures were summed and divided by the number of trap days to determine percentage trap success at 230 each waterhole. All calculations took into account the loss of two traps at the open waterhole due to damage by an elephant, and the disappearance of one trap. Species data were graphed using Microsoft Excel so that graphical comparisons could be made i.e. what were the most common species at each of the two sites A Shannon-Wiener diversity test was performed in Primer 5 on the species data to see whether either site was more diverse in rodent species than the other. It is a measure of both the number of species and the equality of representation of the individuals of all the species i.e. evenness. The Peterson method was used to obtain an estimation of the population size using the unbiased estimator equation: N = (((M+1)(C+1))/(R+1)) -1, where M= number of individuals marked in the first sample, C = Total number of individuals captured in the second sample and R = Number of individuals in second sample that are marked (Krebs 1999). G-tests were performed to compare male versus female ratios, perimeter versus interior trap success and trap success in long model Sherman traps versus short model ones, in JMP IN 5.1. Lastly mean weights per species were determined.

Results Vegetation. Percentage cover was found to be significantly higher (t38 = -12.0132, P<.05) at the closed waterhole in comparison to the open waterhole (Figure 2). Vegetation composition differed between the open and closed waterholes, with the two sites being approximately 25% related (Figure 3). The open waterhole was comprised of 5 plant species with only two prominent species. In contrast the closed waterhole had 8 plant species that were more evenly spread. Urochloa mosambicensis dominated the open waterhole and although it was present at the closed waterhole it was not as prominent (Table 1). From the foliage profiles, the open waterhole appears to have a relatively lower foliage density at all heights (figure 4). When statistically tested at certain heights, it was found that the open waterhole had significantly denser vegetation than the open waterhole (Table 2). Rodents . Four rodent species were caught; the Natal multimammate mouse, Mastomys natalensis, the red veld rat, Aethomys chrysophilus, the bushveld gerbil, Tatera leucogaster and the pouched mouse, Saccostomys campestris. M natalensis and Mastomys coucha are cryptic and cannot be differentiated on appearance alone (Apps 2000), thus all Mastomys were treated as Mastomys natalensis. Additionally Aethomys chrysophilus is split into a second chromosomal species. However identification on morphological grounds is not possible and so all were considered to be A. chyrsophilus (Apps 2000). There was a total of 125 captures in 472 trap days, i.e. 26.5% capture rate. Of these, 56 were caught at the open waterhole (24.1% capture rate) and 69 were caught at the closed waterhole (28.8% capture rate), showing that there was little difference in the number of rodents caught at each waterhole (χ2 =1.3544, P>.05) (Table 4). The Multimammate mouse, M. natalensis was the most common species at the open (50 individuals) and closed (33 individuals) waterhole, although the closed waterhole also had a high number of A. chrysophilus (29 individuals) (Figure 5). A population estimate was derived using the Peterson method (Krebs 1999). The estimated population for the open waterhole was 55 rodents and 66 rodents at the closed waterhole (Table 5). Mark and recapture was only performed on one day’s capture and thus it seems as though over the four days almost the entire population in that area was captured. The closed waterhole grids had a higher diversity index than the open waterhole (Table 3). A G-test performed on total male to female ratio showed males were significantly more abundant than females (χ2=4.6807, P<.05) (Table 6). However, when the two most abundant species were tested separately, it was found that A. chrysophilus had similar numbers of males and females (χ2=.1251, P>.05) with a ratio of almost 1:1, and M. natalensis had almost twice the number of males than females (χ2=5.65, P<.05) (Table 6). The number of total captures in long versus short traps was tested using a G-test. The result showed that overall there was no difference in the number of animals caught in either the long or the short traps (χ2=2.5923, P>.05) (Table 7). With regard to individual species, Aethomys chrysophilus showed no preference for long traps (χ2=.5619, P>.05), however M. natalensis showed a slight preference for long traps rather than short traps (χ2=4.9303, P<.05) (Table 7). Additional G-tests were performed to determine whether traps in the interior of the grid caught significantly more rodents than the interior traps. The results revealed that there was no difference in the captures on the perimeter or the interior (χ2=.132, P>.05) (Table 8).

Discussion We found significant differences in vegetation parameters between the open and closed waterholes. The closed waterhole had a significantly higher percentage ground cover in comparison to the open waterhole (Figure 2). Many studies have reported correlations between distributions of small mammal species and ground cover. For example, Monadjem (1997b) reported that areas with a higher ground cover support a higher diversity of small 231 mammal species. The results of our study support this because we found that the closed water hole supported a higher rodent species diversity (Figure 5). The grass communities comprised different species at each site and were only approximately 25% related (Figure 3). The open waterhole site was dominated by Urochloa mosambicensis (Table 1), a species that typically grows in, and is therefore a good indicator of, disturbed places such as roadsides and overgrazed or trampled veld (Van Wyk and Van Oudtshoorn 1999). The vegetation at the closed waterhole was more diverse and more evenly spread (Table 1). Additionally it was also characterised by a higher diversity of rodents with a higher evenness (Figure 5). Thus the diversity in vegetation may allow a higher diversity of rodents to be supported as they occupy different niches, food sources and habitats. With regard to foliage density, we found that the closed waterhole site had much denser vegetation, which was considerably taller than that of the open waterhole site. The open site’s vegetation was noticeably less dense with very short grass (Figure 4). As mentioned before, the open waterhole was dominated by Urochloa mosambicensis which although is an indicator of disturbed sites is actually highly palatable, which would suggest that the low density of vegetation and short heights are as a result of grazing pressure by larger herbivores.. Bond et al. (1980) stated that sites with a lower vegetation density and cover exposed rodents to a higher degree of predation and other mortality factors. Some species may be able to tolerate these circumstances better than others which would explain why one species dominated the site. The above three components of the vegetation structure and composition analysis show obvious differences between the open and closed waterhole sites. The differences indicate that the open waterhole has been impacted by grazing and trampling and thus can be considered a disturbed area. Additionally, rodent communities vary between the sites. The open water-hole site was dominated by Mastomys natalensis (Figure 5), which is a species with generalized ecological requirements that tolerates disturbances well (Fuller and Perrin 2001). It is a prolific breeder under favourable conditions and breeds throughout the year (Apps 2000). This enables M. natalensis to out-compete other species and dominate an area, and can thus be considered a good indicator of disturbed areas. Although M. natalensis was present at the closed waterhole, it did not dominate. Other species were present in high numbers with Aethomys chrysophilus present in almost equal numbers to M. natalensis (Figure 5). According to Avenant (2000a), a lack of dominance by Mastomys, high species richness and high diversity are all small mammal community characteristics that indicate ecosystem integrity. As the closed waterhole was not dominated by M. natalensis and had a higher diversity index that indicates that the system is in fairly good condition. Our findings support the hypothesis that less impacted waterholes will have a higher rodent diversity (Table 3); however we found that less impacted waterholes did not have a significantly higher abundance of rodents as expected (Table 4) but was rather characterized by the dominance of M. natalensis (Figure 5), because of its tolerance of disturbed areas. These findings are in agreement with other studies that have demonstrated a causal relationship between disturbance and rodent diversity (Avenant and Kuyler 2002), and between losses in diversity and declines in ecosystem functioning. Together they support the proposal by Avenant (2000a and 2000b) that rodent monitoring can be used as an inexpensive and quick indication of ecosystem disturbance. Additional descriptive information from the study may prove useful for future studies or other scientists by providing a comparison. The G tests performed on the male to female ratios of all the species together showed that there were significantly more males trapped than females (P=.03). This probability is primarily influenced by the dominant rodent species M. natalensis, which showed a significant dominance of males (P=.01) (Table 6). There was no significant difference between the numbers of male and female A. chrysophilus (P=.72) (Table 6). A possible reason for this may be that A. chrysophilus breeds much less frequently than M. natalensis and thus fewer females would be sacrificing foraging trips to care for young (Apps, 2000). Further G tests revealed that there was no significant difference in rodents caught in long traps versus short traps (P>.05) (Table 7). However, individually, M. natalensis showed a significant preference for long traps (P<.05). In a study by Slade et al. (1993), it was found that species with lower adult body mass were more commonly caught in long traps. Our study supports this as M. natalensis had the mean lowest body mass (Figure 6). This may be important for future studies as long traps may have a higher trap success. Traps on the perimeter of the grid may catch more rodents than the interior due to edge effects. This trend has been observed in past studies (McClearn Pers. obs.) A G-test indicated that there was no significant difference between catch success of traps in the two parts of the grids (P=.72) (Table 8). One concern of our study was that the closed waterhole was situated in a vlei area which may support different rodent species due to differences in vegetation type rather than the lack of disturbance. The vegetation had vlei elements (sedges and wet area grasses) however on the whole, vegetation was not dominated by wetland species. Additionally the four rodent species found at this site are not typical vlei or wetland species (Apps, 2000), and this 232 suggests that differences in rodent diversity are indeed due to lack of disturbance rather than the presence of vlei vegetation. Other factors that could have influenced our captures during the study would be a heavy rainfall on the third day of sampling. However as the sites are situated in relatively close proximity, both sites were subjected to the same weather conditions and so comparisons between the two would not have been affected. Additionally, results displayed no evidence of being affected. Due to time and equipment limitations, only two waterholes were sampled. This is not an ideal sample as each type of waterhole is not necessarily representative of others. Further sampling at several sites would provide a better understanding of vegetation-rodent interactions. It would be of interest to conduct another study in mid-autumn to mid-winter as these months are considered to be the best time to conduct small mammal sampling, while spring and summer are considered less favourable as significantly fewer species are found in the traps during this time of year (Avenant and Kuyler 2002). This may be due to the availability of food resources in an otherwise limited environment. The possibility of a higher diversity of rodent species being trapped may further support our results.

Acknowledgements We thank Deedra Mclearn, Robert Timm and Mike Smith for their continuous, invaluable assistance, knowledge and support. Additional thanks to Kinesh Chetty, Fahiema Daniels, Megan Eastwood and Stephanie Johnson for their assistance in the field.

Literature cited Apps, P. 2000. Smithers’ mammals of Southern Africa. Struik Publishers. South Africa. Avenant, N. L., and P. Kuyler. 2002. Small mammal diversity in the Maguga Dam inundation area, Swaziland. South African Journal of Wildlife Research 32:101–108. Avenant, N.L. 2000a. Small mammal community charachtersitics as indicators of ecological disturbance in the Willem Pretorius Nature Reserve, Free State, South Africa. South African Journal of Wildlife Research. 30: 26- 33 in Avenant, N. L., and P. Kuyler. 2002. Small Mammal diversity in the Maguga Dam inundation area, Swaziland. South African Journal of Wildlife Research 32: 101–108. Bond, W., Ferguson, M., Forsyth, G. 1980. Small mammals and habitat structure along altitudinal gradients in the Southern Cape mountains. South African Journal of Zoology 15: 34-43. Fuller, J. A. and Perrin, M. R. 2001. Habitat assessment of small mammals in the Umvoti Vlei Conservancy, KwaZulu-Natal, South Africa. South African Journal of Wildlife Research 31: 1-12. Gaylard, A., N. Owen-Smith, and J. Redfern. 2003. Surface water availability implications for heterogeneity and ecosystem processes. J. T. Du Toit, K. F. Rogers, and H.C. Biggs, (editors). The Kruger Experience. Island Press. USA. Grant, R. 1999. Status report on the Northern Plains Projects. South African National Parks Scientific Report. 01/99. South African National Parks, Skukuza, South Africa. Krebs, C.J. 1999. Ecological methodology 2nd ed. Addison Wesley Longman. United States of America. Monadjem, A. R. A. 1999. Geographic distribution of small mammals in Swaziland in relation to abiotic factors and humans land-use activity. Biodiversity and conservation 8: 223-237 Monadjem, A. and Perrin, M.R. 1997. Population dynamics of Lemniscomys rosalia (Muridae:Rodentia) in a Swaziland grassland: effects of food and fire. South African Journal of Zoology 32:129-135 in Avenant, N. L., and P. Kuyler. 2002. Small mammal diversity in the Maguga Dam inundation area, Swaziland. South African Journal of Wildlife Research 32:101–108. Salvatori, V., Egunyu, E., Skidmore, A. K., de Leeuw, J. and van Gils H. A. M. 2001. The effects of fire and grazing pressure on vegetation cover and small mammal populations in the Masai Mara National Reserve. East African Wild Life Society, African Journal of Ecology 39: 200-204. Slade, N.A., Eifler, M.A., Gruenhagen, N. M. and Davelos, A. L. 1993. Differential effectiveness of standard and long Sherman livetraps in capturing small mammals. Journal of mammalogy 74: 156-161. Van Wyk, E. and Van Oudsthoorn, F. 1999. Guide to grasses of southern Africa 1st ed. Briza Publications. Pretoria, South Africa. Western, D. 1995. Water availability and its influence on the structure and dynamics of a savanna large mammal community. East African Wildlife Journal 13: 265-286.

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Table 1. Percentage species composition of vegetation at the open and closed waterhole. Species Open (%) Closed (%) Aristida congesta 1.9 0.0 Cenchrus cilliaris 1.0 0.0 Digitaria sp. 0.0 10.6 Eragrostis sp. 5.8 0.0 Eragrostis superba 31.7 0.0 Ischmum sp. 0.0 6.7 Lala palm 0.0 1.9 Lawn grass 0.0 16.3 Panicum maximum 0.0 13.5 Sedge sp. 0.0 18.3 Tall purple sp. 0.0 9.6 Urochloa mosambicensis 59.6 23.1

Table 2. Results of students t-tests comparing vegetation of open and closed waterhole at 0.25m, 0.50m and 0.75m height classes. Height t value d.f. P value 0.25m -2.03479 38 0.048892 0.50m -2.80539 38 0.007879 0.75m -3.21644 38 0.002652

Table 3. Shannon-Wiener diversity indices for rodent species at open and closed study sites. Sample S N d J' H'(loge) 1-Lambda' A (Open) 3 45 0.5254 0.3856 0.4236 0.2081 B (Open) 2 11 0.417 0.4395 0.3046 0.1818 C (Closed) 3 38 0.5498 0.902 0.9909 0.6216 D (Closed) 4 31 0.8736 0.6689 0.9273 0.5742

Table 4. G-test comparing captures at the open and closed waterhole. Level Estim Prob Hypoth Prob ChiSquare DF P value Closed 0.552 0.5 1.3544 1 0.2445 Open 0.448 0.5

Table 5. Unbiased estimator of population size of each site. A B C D A+B (Open) C+D (Closed) Total 35 14 63 17 55 66 131

234 Table 6: Test probabilities of G-tests for male to female ratios of captured rodents Level Estim Prob Hypoth Prob ChiSquare DF P value M:F ratio of all rodent species F 0.38947 0.5 4.6807 1 0.0305 M 0.61053 0.5 M:F ratio of Aethomys chrysophilus F 0.46875 0.5 0.1251 1 0.7236 M 0.53125 0.5 M:F ratio of Mastomys natalensis F 0.36709 0.5 5.65 1 0.0175 M 0.63291 0.5

Table 7: Test probabilities for Chi-square test for long versus short traps Level Estim Prob Hypoth Prob ChiSquare DF P value Long vs short trap ratio of all captures Long traps 0.224 0.168 2.5923 1 0.1074 Short traps 0.776 0.832 Long vs short trap ratio of Aethomys chrysophilus Long traps 0.12121 0.168 0.5619 1 0.4535 Short traps 0.87879 0.832 Long vs short trap ratio of Mastomys natalensis Long traps 0.26506 0.168 4.9303 1 0.0264 Short traps 0.73494 0.832

Table 8: Test probabilities for G-test comparing interior and perimeter trap successes Level Estim Prob Hypoth Prob ChiSquare DF P value Interior 0.424 0.408 0.132 1 0.7164 Perimeter 0.576 0.592

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Figure 1. Site map of a section of the Vlakteplaas management unit showing the position of Elandskuil waterhole (22° 42’ 18” S; 31° 09’ 43” E ) and Elephant waterhole (22° 49’ 40” S; 31° 14’ 21” E).

100

60 Percentage cover (%) cover Percentage

Mean 30 Mean±SE OPEN CLOSED Figure 2. Mean percentage vegetation cover at the open waterhole (39.30%) and closed waterhole (87.15%).

236 Open waterhole Closed waterhole

Figure 3. Cluster analysis illustrating similarity between species composition of the eight vegetation transects (A and B are transects in Open waterhole site, and C and D are transects in the closed waterhole site.

1.5 Open Closed

0.75 Height (m)

0.5

0.25

0.00 1.00 2.00 3.00 4.00 5.00 6.00 1/distance (1/m)

Figure 4. Foliage profile of vegetation at the open and closed waterhole. 237

70 1

60 2 50 33 T. leucogaster 40 S. campestris

M. natalensis

30 A.chrysophilus 50

captures of Number 20 29 10

4 0 Open (A+B) Closed (C+D)

Figure 5. Absolute number of each species found at the open and closed waterhole.

45 Weight (g) Weight

Tatera leucogaster Mean

Mastomys natalensis Mastomys Mean±SE Aethomys chrysophilus Aethomys

Saccostomys campestris Saccostomys Figure 6. Mean weights of the four species of rodents. 238 Reduced kin recognition in the success of a widespread ant: Lepisiota capensis

Category: Independent Project Participants: Eric Caldera Site: Punda Maria, Kruger National Park, Mpumalanga Province, South Africa

Key words: Argentine ant, kin recognition, Lepisiota capensis, intraspecific aggression

Abstract: Intraspecific competition for resources is often the factor limiting worker population size in ants. Energy expended on competition would otherwise go toward colony growth and reproduction. A loss of intraspecific aggression has been key to the success of many invasive ants, including the Argentine ant. The Black Sugar Ant (Lepisiota capensis) may be capable of displacing the Argentine ant. Here I tests whether L. capensis has also experienced a similar reduction in intraspecific aggression by conducting aggression assays within and among four sites in Kruger National Park, South Africa. My results show that the L. capensis does have reduced intraspecific aggression. This reduction in intraspcific aggression may help L. capensis escape energetic costs associated with intraspecific competition, making this species a better competitor. Because of this reduced intraspecific aggression, L. capensis may pose a threat as a potentially invasive species and consequently a threat to biodiversity in South Africa.

Introduction Kin recognition is an important factor in determining the ecological and evolutionary success of ants (Holldobler and Wilson 1990), particularly within the context of kin selection theory. The ability for ants to distinguish nestmates from non-nestmates is important given their haplodiploid method of reproduction. In most ant species, worker and soldier casts do not reproduce and thus cannot directly pass on their genes. Because of the haplodiploid system of reproduction in ants, if a worker or soldier were to have offspring, that individual would be less related to its own son or daughter than a reproductive sister. It is important that worker and soldier casts only “help” colonies with queens from which they were born. If an ant were not able to recognize kin then it would have the potential to accidentally help non-kin individuals from different nests, which would violate kin selection theory. There are a variety of kin recognition systems in the social insects. In ants, genetic and environmental cues control nestmate recognition (Tsutsui et al. 2000; Holldobler and Wilson 1990; Carlin 1989); however, they are not necessarily mutually exclusive. Some kin recognition systems rely on queen-emitted pheromones to mark nestmates, while other systems rely on the specific odor of the nest to mark individuals for nestmate recognition (Holldobler and Wilson, 1990; Vander Meer and Alonso 1998). The consequence of kin recognition is that a large amount of energy is spent on intraspecific competition that would otherwise be directed toward colony growth, maintenance, and reproduction (Holldobler and Lumsden 1980). This behavior of defending territories is common in most ants and leads to multicolonial behavior, where several single queen nests exist independently without sharing resources (Holldobler and Wilson 1990). Multicoloniality is also thought to limit population density in ants (Holldobler and Wilson 1990). The success of many ants is dependent on the loss of this energetic cost associated with multicoloniality. In contrast, unicolonial species have multiple queens and nests that share resources and behave functionally as one colony. Worker populations of unicolonial species often attain high densities (Porter and Savignano 1990; Macom and Porter 1996; Holway 1998). Achieving high worker densities is key to the success of many native and introduced ant species (Holldobler and Wilson 1983; Lynch et al. 1980; Fellers 1987; Adams 1990). For example, two worldwide invasive ants, the Red Imported Fire Ant, Solenopsis wagneri (formely Solenopsis invicta), and the Argentine ant: Linepithema humile, both have ant colonies primarily containing one queen in their native ranges, while colonies in introduced ranges often have multiple queens (Kaspari 2000). The Argentine ant is also known to exchange workers and even queens between different nests in introduced areas (Holway 1999). Because these invasive ants do not compete for resources intraspecifically, they are able to achieve abnormally high worker densities which allow the colony to out-compete almost all other ground dwelling arthropods (Holway 1999). In single queen colonies of S. wagneri, queen emitted pheromones control nestmate recognition; however this system breaks down in multiple queen colonies (Vander Meer and Alonso 2000). Nestmate recognition in introduced colonies of L. humile is not controlled by queen pheromones (Caldera and Holway 2004), but may have a strong genetic component. Some studies imply that introduced colonies of L. humile often pass through a genetic bottleneck that causes colonies to loose kin recognition capabilities, allowing them to form large “supercolonies” (Holway et al. 1998; Tuitsani et al. 2000). At the supercolony scale 239 nestmate recognition can occur across hundreds of kilometers rather than meters between nests (Holway 1995; Holway et al.1998). The Argentine ant is a worldwide invasive pest that displaces mutualisms, and can disassemble community structure (Bond and Slingsby 1984; Sanders et al. 2003; Holway 1999). The Argentine ant is invasive in Africa, the Americas, and Europe. One ant species that may be capable of competing with and displacing the Argentine ant is the African Small Black Sugar Ant, Lepisiota capensis (Picker et al. 2002). Here I test whether the ability for L. capensis to displace the Argentine ant is due to reduced nestmate recognition.

Methods Collection sites. I collected ants from four sites in the Kruger National Park (KNP), South Africa: Punda Maria camp (PM), Letaba camp (LB), Skukuza camp (SK), and the Luvuvhu River west of the Pafuri camp (LR). At each site, I collected from two nests that were a minimum of 100m apart to reduce the risk of sampling from the same nest more than once. To further ensure that individuals from each sample belonged to different nests, I only collected ants where queens and brood were present. I maintained ant colonies in plastic containers and fed them 20% sugar water once a week.

Behavioral assays and analysis. I used aggression assays that have been previously used to investigate levels of intraspecific aggression in Argentine ants to quantify kin recognition in L. capensis (Holway et al. 1998; Tsutsui et al. 2000; Caldera and Holway, 2004). Pairwise aggression assays were performed between ants from the two nests collected at each site and among ants collected from nests from the four sites (Figure 1). For each aggression assay two workers, one from each of two nests, were placed into a vial, after which a smaller plastic container was inserted into the vial so that the space for the ants to move was minimized, thus increasing the probability for the ants to interact. Interactions between the ants were observed for 5min and scored using a 4 point scale of escalating aggression: touch = 1 (contacts that included prolonged antenation), avoid = 2 (contacts that resulted in one or both ants quickly retreating in opposite directions), aggression = 3 (lunging, biting, and pulling legs or antennae), or fight = 4 (prolonged aggression between individuals including spraying of formic acid). Each aggression assay was repeated 10 times, using different ants each time, and the average was compared to the average level of aggression between nestmates using a Kruskal-Wallis non parametric test. The average level of aggression between nestmates (1.1±0.447) was determined by conducting 10 aggression assays between workers from the same nest, repeating the assays using a different nest, and averaging them together. For the sites where aggression between the two nests sampled was not significantly different from the average level of aggression between nestmates only one of the nests was used to conduct aggression assays between the different sites. However, where aggression between two nests collected was high, both nests were used in aggression assays between sites. To ensure that aggression seen in L. capensis was not due to aspects of being housed in lab colonies interfering with recognition sensory mechanisms, I separated one nest into two lab colonies and conducted aggression assays after 2 weeks.

Results There was no significant aggression between nests at three sites (Letaba, Luvuvhu River, and Punda Maria) compared to the average level of aggression between nestmates (1.10+0.447), while aggression at the Skukuza camp was significantly high between nests (Figure 2). Of the nine aggression assays conducted between sites, six had significantly high aggression while the three did not have significantly high aggression (Figure 3). In all three aggression assays containing Skukuza site A, there was no significant aggression. For aggression assays between the two control nests that were isolated from the same original nest, no score aggression score above 1 was seen. I document here that L. Capensis maintains multiple queen nests.

Discussion My results show that some colonies of L. capensis have experienced a loss of intraspecific aggression, as demonstrated by the 6 of 13 aggression assays that did not differ significantly from the average level of aggression between nestmates. Of the four aggression assays conducted within the camp sites one had significantly higher aggression than the average aggression of ants taken from the same nest, and within the nine aggression assays among different camps, six had significantly high aggression. This shows that L. capensis is in fact capable of intraspecific aggression. 240 In the aggression assays conducted within camps, it is possible that the ants sampled at the different sites within a camp belonged to the same large nest. This could possibly account for the lack of aggression between three of the four tests. In order to account for this, I only sampled at locations where queens and brood were present. However, some ant species do have multiple queen colonies that are capable of spreading long distances and maintaining kin recognition (Holldobler and Wilson 1990). This also may account for the low intraspecific aggression within camps. I have documented here that L. capensis does contain polygyne nests. While these cases of reduced intraspecific aggression within camps, and across relatively short distances, may be explained by the polygynous biology of L. capensis, further explanation is required to explain the reduced aggression between the camps. It may be that L. capensis is capable of forming large scale super colonies due to a loss of genetic variation and ultimately intraspecific aggression. This is the case with the invasive Argentine ant and the Red Imported fire ant (Holway et al. 1998; Tuitsani et al. 2000; Ross and Keller 1995; Ross et al. 1996). In Argentine ants, it has been shown that as genetic variation and variation in cuticular hydrocarbons of ants between nests increases, so does intraspecific aggression (Holway. et al. 1998; Tuitsani et al. 2000) A great deal of research will be required to investigate whether reduced nestmate recognition in L. capensis is due to a loss of genetic variation, but this study implies that this may be the case. Because L. capensis is polygynous, along with the fact that there was low aggression between nests at different camps, it is not likely that kin recognition is currently governed by queen emitted pheromones. It may be that kin recognition in L. capensis is determined by cuticular hydrocarbons and that aggression assays that yielded low intraspecific aggression were between nests with similar hydrocarbons. Further research should use biochemical and molecular methods to see if genetic and cuticular hydrocarbon patterns correlate with levels of intraspecific aggression. Reduced intraspecific aggression in L. capensis may make this species a potential threat to biodiversity. The loss of intraspecific aggression in the invasive Argentine ant has allowed this species to maintain extremely high worker densities, which in turn allows large colonies to out compete ants and other ground dwelling arthropods that are larger and more aggressive (Holway et al. 1998; Holway 1999; Holway and Case 2001). In the Cape region of South Africa, however, L. capensis may be capable of displacing the Argentine ant (Picker et al. 2002). If L. capensis’ ability to compete with L. humile is due to a similar loss of intraspecific aggression, then it is possible that L. capensis is capable of becoming just as successful as an invasive species as the Argentine ant. L. capensis is already widely distributed in South Africa. From general observations while collecting in camps around Kruger National Park, I noticed that L. capensis is very abundant in the camp sites. It is likely that these colonies were established through long distance jump dispersal events through transport by automobiles and landscaping plans and equipment. Because many of the camps are regularly watered, these areas also provide good habitat for the ants to thrive. Long distance jump dispersal events are typically the dominant dispersal pattern for invasive ants (Suarez et al. 2001). A great deal of research is required to fully understand kin recognition in L. capensis; however, this study shows that this species does have reduced intraspecific aggression. This reduced intraspecific aggression may give this widespread species an ecological advantage that allows it to out-compete ground dwelling arthropods and ultimately threaten biodiversity.

Acknowledgements I thank Kyle Harris, Scott Briscoe, Gareth Hempson, and our game guard Watchie (Lucas) Masinga for their assistance in field collection. For their dedication to teaching I thank Laurence Kruger, Deedra McClearn, and Julie Coetzee.

Literature Cited Adams, E.S. 1990. Boundary disputes in the territorial ant Azteca trigona: Effects of asymmetries in colony size. Animal Behaviour 39:321-328 Bond, W. and P. Slingsby. 1984. Collapse of an ant plant mutualism: The Argentine ant and mymrmecochorous proteaceae. Ecology 65:1031-1037 Caldera, E.J. and D.A. Holway. 2004. Evidence that queens do not influence nestmate recognition in the invasive Argentine Ant. Insectes Sociaux in press Carlin, N.F. 1989. Discrimination between and within colonies of social insects: two null hypo-theses. Netherlands Journal of Zoology 39: 86 –100 Fellers, J.H. 1987. Interference and exploitation in a guild of woodland ants. Ecology 68: 1466-1478

241 Holldobler, B. and C. J. Lumsden. 1980. Territorial strategies in ants. Science 210:732-739 Holldobler, B. and E.O. Wilson. 1983. Queen control in colonies of weaver ants (Hymenoptera:Formicidae). Annual review of Entomological Society America 76:235-238 Holldobler, B. and E.O. Wilson. 1983. The evolution of communal nest-weaving in ants. American Science. 71:490- 499 Holldobler, B. and E.O. Wilson.1990. The ants. Belknap Press of Harvard University Press, Cambridge, Massachusetts, USA Holway, D.A. 1995. Distribution of the Argentine ant (Linepithema humile) in Northern California. Conservation Biology 9:1634-1637 Holway, D.A., 1998. Factors controlling rates of invasion: a natural experiment using Argentine ants. Oecologia 115:206-212 Holway, D.A., 1999. Competitive mechanisms underlying the displacement of native ants by the invasive Argentine ant. Ecology. 80: 238-251 Holway, D.A., A. V. Suarez, and T.J. Case. 1998. Loss of intraspecific aggression in the success of a widespread invasive social insect. Science 282:949-952 Kaspari, M. 2000. A primer on ant ecology. Pp 12. In Agosti, D., J. D. Mayer, L. E. Alonso, and T. R. Schultz (ed.), Ants: standard methods for measuring and monitoring biodiversity. Smithsonian Institution, Washington DC. Lynch, J. F., E. C. Balinsky, and S. G. Vail. 1980. Foraging patterns of three sympatric forest ant species, Prenolepis imparis, Paratrechina melanderi, and Aphaenogaster rudis. Ecological Entomology 5:353-371 Malcom, T.E. and S.D Porter, 1996. Annual review of Entomological Society of America 98:535. Picker, M., C. Griffiths, and A. Weaving. 2002. Pp 426. Field guide to insects of South Africa. Struik Publishers, Cape Town Porter, S. D., and D. A. Savignano. 1990. Invasion of polygyne fire ants decimates native ants and disrupts arthropod community. Ecology 71:2095-2106 Ross, K. G., and L. Keller. 1995. Evolution of social organization: insights from fire ants and other highly eusocial insects. Annual Review of Ecology and Systematics, 26:631-656 Ross, K. G., E. L. Vargo, and L. Keller. 1996. Social evolution in a new environment: the case of introduced fire ants. Proceedings of the National Academy of Sciences 93: 3021-3025 Sanders, N.J., N.J. Gotelli, N.E. Heller, and D.M. Gordon. 2003. Community disassembly by an invasive species. Proceedings of the National Academy of Sciences (100)5:2474-2477 Tsutsui, N.D, A.V. Suarez, D.A. Holway, and T.J. Case. 2000. Reduced genetic variation and the success of an invasive species. Proceedings of the National Academy of Sciences (11) 97:5948-5953 Vander Meer, R.K. and L.E. Alonso. 2000. Pheromone Directed Behavior in Ants, pp.159-192. In R.K. Vander Meer, M. Breed, M. Winston, and K.E. Espelie [eds.], Pheromone Communication in Social Insects. Westview Press, Boulder, CO. 368 p.1998 Vander Meer, RK, and LE Alonso. 2001. Queen primer pheromone affects conspecific fire ant (Solenopsis invicta) aggression: Behavioral Ecology and Sociobiology 51:122-130

242

Figure 1. Diagram of 13 aggression assays conducted in four sites in Kruger National Park, South Africa. Each line represents 10 aggression assays conducted between the two sites. Four assays were performed between different nest at the same site, and nine were performed between sites. The Skukuza site was the only site to show intraspecific aggression within one site and therefore both the A and B nests within Skukuza were tested against all other sites.

2 Level of aggression of Level

0 Punda Maria Luvuvhu River Letaba *Skukuza avg. nesmate aggression

Figure 2. Mean aggression and standard deviation between Lepisiota capensis nests separated by >100m at four different camp sites in Kruger National Park, South Africa. A Kruskal Wallace non-parametric test was used to determine if levels of aggression were significantly higher than the average level of aggression between ants from the same nest 1.10±0.46. * Aggression was significantly higher than the average level of aggression between nestmates, (P<.01). 243

2 Level of aggression of Level

0 *Punda *Punda *Luvuvhu *Skukuza *Skukuza *Skukuza Skukuza B Skukuza B Skukuza B Avg. Maria & Maria & River. & A & A & A & & Letaba & Punda & nesmate Luvuvhu Letaba Letaba Letaba Punda Luvuvhu Maria Luvuvhu aggression River Maria River River

Figure 3. Mean aggression and standard deviation between Lepisiota capensis nests at four different sites in Kruger Nation Park, South Africa. A Kruskal Wallace non-parametric test was used to determine if levels of aggression were significantly higher than the average level of aggression between ants from the same nest 1.10±0.46. *Aggression was significantly higher than the average level of aggression between nestmates, (P<.05).

244 Differences in dung beetle diversity on granite and basalt-derived soils and between periods of diel activity in Kruger National Park, South Africa

Category: Independent Project Participants: Stephanie Johnson, Fahiema Daniels, and Megan Eastwood Site: Punda Maria, Kruger National Park, Mpumalanga Province, South Africa

Key Words: biodiversity, diel activity, dung, Scarabaeidae, soil type

Abstract: Dung beetles are important biodiversity indicators and also play integral roles as detritivores in many ecosystems by removing large amounts of dung from the landscape. However, relatively little is known about dung beetle diversity, especially in South Africa. As part of an ongoing effort to assess dung beetle diversity in Kruger National Park, we used pitfall traps to examine dung beetle diversity near Punda Maria rest camp. Our study focused on differences between dung beetle assemblages in areas with basalt and granite-derived soils and between nocturnal and diurnal sampling periods. As expected, the nutrient-rich basalts yielded a higher biomass of beetles than the granites, and nocturnal sampling also caught a greater biomass of beetles than diurnal sampling.

Samevatting: Muskruiers is belangrikke aanwysers van biodiversiteit. Hulle speel ook ŉ integraal rol in baie ecosisteem deur die verwydering van groot bedraë mis vanaf die landskap. Egter, is daar min studies wat aandag skep ann hul verskeidenheid, vir al in Suid Afrika. As deel van ŉ projek in die Kruger Nasionale Park het ons valstik betrapers verbruik om misruikers se verskuidenheid te skat naby die Punda Maria rus-kamp. Ons studie is gefokus op die verskil tussen miskruier versamelings op die ystermarmer en die graniet grond. Ons kyk ook na die verskil tussen die nagtelik en daaglikse miskruiers in die ystermarmer en granite grond. Soos ons vermoet het, het die ystermarmer terrein ŉ hoër massa van miskruiers gehet as die graniet terrein. Nagtelik toetsing het ook ŉ hoër massa gekry as daglikse toetsing.

Introduction Dung beetles are good indicators of biodiversity on both a global and regional scale: they are found on every continent and their distribution has been linked to climate and biogeography, yet because of their small size they also discriminate on a very fine scale between habitats (Davis 2002). They are responsive to a variety of environmental factors including moisture levels, vegetation, and soil, and they are generally sensitive to alteration in habitat structure (Davis 1997, Davis 2002, Van Rensburg et.al. 1999). Therefore, dung beetles are a useful indicator species for monitoring habitat changes, especially as it is easy to assess dung beetle abundance with pitfall traps baited with dung (Davis 2002). Southern Africa has a very rich dung beetle fauna that includes over 780 species (Doube 1991). South Africa in particular has many endemic species of dung beetles, especially in the highveld, the winter rainfall areas, and the eastern coastal region. Knowledge about dung beetle diversity in this country is still limited, although attempts have been made to accurately assess their abundance (Davis 2002). Recently, Pretoria University has begun a project to assess dung beetle diversity in the Kruger National Park, but while the species in the south of the park have been well surveyed, very little sampling has occurred in northern Kruger. Many factors contribute to dung beetle species diversity. In South Africa, dung beetle diversity at the local level is mainly influenced by the types of soil, vegetation, and dung available (Davis 2002). In KNP, most soil is derived from either granite or basalt bedrock (Venter et al. 2003). Granites form sandy, well-draining soils that are generally low in nutrients, while basalts form relatively nutrient-rich, slow-draining clay soils that support a higher grass biomass and therefore more grazers than granite-derived soils (Venter et al. 2003). A higher number of mammalian grazers would mean a greater variety of types of dung and more resources for dung beetles, perhaps leading to a greater diversity of dung beetles as many South African dung beetles are highly specialized on specific dung types (Doube 1991). Dung beetle diversity is also increased by the spatial, seasonal, and temporal variability in dung beetle activity (Davis 2002). For example, species are most active either during the day (diurnal) at sunrise or sunset (crepuscular), or at night (nocturnal), and this temporal separation should reduce competition and allow diverse communities of beetles to exist during different diel periods (Davis 2002, Hanski and Cambefort 1991). A study conducted in the semi-arid Karoo/Kalahari area margins of South Africa found that there is a north-east to south-west 245 gradient in dung beetle diversity that corresponds to a climatic gradient in rainfall seasonality and its concurrence with appropriate temperatures for activity (Davis and Scholtz 2004). To supplement the Pretoria project on Kruger dung beetle diversity and to further explore how this diversity varies spatially with soil type and temporally by times of peak diel activity, we examined two hypotheses: Hypothesis 1: Dung beetle diversity will be higher on basalt-derived soils than on granite-derived soils, and a greater biomass of beetles will be collected in the traps on the basalt soils. Hypothesis 2: There will be a difference in beetle species composition among the catches collected from pitfall traps in the evening and the morning.

Methods The experiment was conducted in the Punda Maria area of the Kruger National Park, Mpumalanga Province, South Africa between the 21st and 25th of March. We sampled six sites, three on granite-derived soils and three on basalt-derived soils (Figure 1). Pitfall traps baited with elephant dung were set in three by three grids of nine traps spaced ten meters apart. Although vegetation cover varied somewhat between sites, all traps were placed in unshaded areas. Pitfall traps were filled with 350ml of soap solution in order to prevent any beetles from escaping, as the soap breaks the surface tension on the water and the beetles consequently drown. We checked and re-baited the traps with fresh dung every day between seven and nine AM and again between four and five PM. One sample of nocturnal beetles, henceforth referred to as a night sample, and one of diurnal beetles, a day sample, was collected from each plot. To collect samples, we poured the water from the traps through a strainer to remove beetles. The samples were cleaned to remove dirt, debris, soap, and all non-scarab insects. We sun-dried the beetle samples for four hours, then weighed and counted the beetles to determine total biomass and total number of beetles collected in each sample. Beetles were lined up in a row and five beetles were randomly selected from each sample using a random number generator. The five beetles in this sub-sample were then identified to morphological species. Davis (2002) recommended this sub-sampling technique for studies where time was limited and large numbers of beetles were collected. Data were analyzed using Kruskal Wallace tests in Statistica 6.0 and multidimensional scaling (MDS) plots, cluster analysis, and Shannon-Weiner diversity indices in Primer 5.

Results Thirty-four dung beetle morphospecies were observed in total; samples from the basalt day sites contained ten different species, while the basalt night sites contained seventeen species, the granite day sites had fifteen, and the granite night sites had fourteen. A multidimensional scaling plot (MDS) revealed highly significant groupings between sites with a stress value of 0.04, and a cluster analysis also revealed many similarities between sites (Figures 2 and 3). Both analyses group the three basalt day samples, and the cluster analysis revealed that these samples are more related to each other than any other sample (~75% similarity) (Figure 3). The granite and basalt night samples group closely on the MDS plot, and the cluster plot also shows them all to be at least 58% similar with some plots being even more closely related (Figures 2 and 3). The granite day samples were different from each other and all of the other samples in both of the plots (Figures 2 and 3). When we compared the specific species seen at each of the sites, we found that the basalt day and basalt night sites had no species in common but that other sites showed some species overlap. The granite day and night sites shared five common species, the granite and basalt day sites shared five species, and the granite and basalt night sites shared twelve species. The sites on basalt-derived soils yielded a greater total number of beetles than the sites on the granite-derived soils (H=16.460, N=108, P<.001) (Figure 4). The scarab biomass was also found to be greater on the basalt sites than on the granite sites (H=8.067, N=108, P<.001) (Figure 5). Using the Shannon-Weiner diversity index, we found no difference in species richness or diversity between granite and basalt sites (Table 1). Differences between day and night samples were compared within sites on the same soil type. In the samples from basalt-derived soils, the night samples contained a greater number of beetles and a larger scarab biomass than the day samples (H=15.520, N=54, P<.001; H=19.325, N=54, P<.001) (Figures 6 and 7). A very similar trend was seen on the granite-derived soils: the total beetle number and the total biomass were higher for the night samples than the day samples (H=24.070, N=54, P,.001; H=25.379, N=54, P<.001) (Figures 8 and 9). The average diversity of morphospecies was higher in the night samples than those collected during the day, but the difference is not significant (Table 1).

246 Discussion Samples from basalt-derived soils showed a greater total number of beetles and biomass of beetles than samples from granite-derived soils. These results are consistent with our predictions. Basalt-derived soils support more grasses and thus more herbivores than granite-derived soils (Venter et al. 2003). Consequently, the basalts should be able to support more dung beetles than the granites. We also hypothesized that the basalts would support a greater diversity of beetle species, but this was not seen. Although the basalt and granite sites supported different species of dung beetles, species richness was very similar between the two soil types. The observed differences between samples from basalt and granite-derived soils may be due to the soil differences mentioned above, but it may also have been influenced by a rainshower in the Punda Maria area during the granite day sampling period. Insects are ectotherms and their environment affects their body temperatures and level of activity. The cooler temperatures during the rain may have prevented many beetles from flying and actively foraging for dung. This would contribute to the small number of beetles collected in the granite day samples and would also explain why the granite day samples were not closely related in the MDS or cluster plots. However, little or no rain appeared to have fallen on the actual sampling sites and the pitfall traps did not overflow and were consequently still functional. Therefore, the samples collected during the rain can still provide useful data. The most dramatic differences were seen in the comparisons of day and night samples. As predicted, we found a difference in species composition between night and day samples: there was very little species overlap between day and night samples, only five species, and the night samples were generally more diverse. Additionally, we found that night samples contained a significantly greater number of beetles and beetle biomass than day samples. Our findings support earlier research demonstrating differences in dung beetle distribution on different soil types. For example, soil hardness has been shown to influence the species of dung beetles found: sandy soils, such as granite-derived soils, usually support different species than clay or loam soils, such as basalt-derived soils (Davis 1996, Doube 1991). In fact, Davis (1996a) found that in the Gauteng Provence bushveld, soil type had more a significant effect on dung beetle distribution than vegetation type or air temperature. We found no differences in species richness between basalt and granite-derived soils, which supports Davis’ study in the Pretoria bushveld that found no significant variation in species richness between soil types (Davis 1996b as cited in Davis 2002). Previous studies on diel diversity also found nocturnal and diurnal communities of dung beetles to be very distinct, possibly because flying during the day requires different thermoregulatory adaptations than flying at night (Cambefort 1991, Hanski and Cambefort 1991). Dung beetles appear to be adapted to fly at certain light intensities (Van Rensburg et al. 1999). However, few other studies seem to have compared beetle biomass or number of beetles between nocturnal and diurnal species. Hanski and Cambefort (1991) did hypothesize that nocturnal or crepuscular species would be more numerous than diurnal species, as their flying times correspond with the periods when herbivores produce the most dung, allowing them to reach the dung before many other possible competitors. Dung beetles play an important role as detritivores in many ecosystems: they consume large amounts of dung and recycle the nutrients from this dung back into the environment (Doube 1991). This role is especially vital in habitats with high mammalian herbivore populations that produce large amounts of dung, such as African savanna and pasturelands. When setting aside areas for conservation, it will be important to consider that dung beetles appear to specialize on specific soil types and therefore a heterogeneous area must be conserved to maximize dung beetle species abundance. Furthermore, dung beetles’ specializations to particular environmental factors such as soil type make them an ideal indicator species. The abundance of dung beetle species showing especially high habitat-fidelity can be used as a measure of changes in vegetation structure and herbivore densities (Van Rensburg et al. 1999). It is important to know what diversity of dung beetles exists and what factors influence this diversity to better conserve dung beetles and to identify habitats in which they can function as indicator species. Future studies should be conducted at different times throughout the year to encompass the seasonal variation of dung beetles. Studies using dung from various herbivore species as bait and assessing if different sized beetles select dung with different particle sizes would also be insightful.

Acknowledgments We wish to thank Mike Smith for all of his help throughout the entire project, Julie Coetzee for her help in planning the project and analyzing the data, Simione Ndlovhu for being our game guard and helping us relocate sampling sites, Deedra McClearn for her help with sample collection and statistics, Kinesh Chetty for taking us to collect dung and helping with sample collection, and Bob Timm, Tarryn Morris, Govan Pahad, and Tammy Baudains for help with sample collection.

247 Literature Cited Cambefort, Y. 1991. Dung beetles in tropical savannas. Pages 156-178 in Hanski, I., and Y. Cambefort, eds. Dung Beetle Ecology. Princeton University Press. Princeton, New Jersey, USA. Davis, A. 1996a. Seasonal dung beetle activity and dung dispersal in selected South African habitats: implications for pasture improvement in Australia. Agriculture, Ecosystems and Environment 58:157-169. Davis, A. 1997. Climatic and biogeographical associations of southern African dung beetles (Ceoptera: Scarabaeidae s. str.). African Journal of Ecology 35: 10-38. Davis, A. 2002. Dung beetle diversity in South Africa: influential factors, conservation status, data inadequacies, and survey design. African Entomology 10: 53-65. Davis, A. and C. Scholtz. 2004. Local and regional species ranges of a dung beetle assemblage from the semi-arid Karoo/Kalahari margins, South Africa. Journal of Arid Environments 57: 61-85. Doube, B. 1991. Dung Beetles of Southern Africa. Pages 133-155 in Hanski, I., and Y. Cambefort. Dung Beetle Ecology. Princeton University Press. Princeton, New Jersey, USA. Hanski, I., and Y. Cambefort. 1991. Species Richness. Pages 350-365 in Hanski, I., and Y. Cambefort, eds. Dung Beetle Ecology. Princeton University Press. Princeton, New Jersey, USA. PRIMER-E Ltd. 2000. Primer 5 for windows installation. http://www.primer-e.com/. StatSoft, Inc. (2001). STATISTICA (data analysis software system), version 6. www.statsoft.com. Van Rensburg, B. J. , M. A. McGeoch, S. L. Chown and A. S. Van Jaarsveld. 1999. Conservation of heterogeneity among dung beetles in the Maputaland Centre of Endemism, South Africa. Biological conservation. 18 : 145- 153. Venter, F., B. Scholes, L. Otter, and A. Woghiren. 2003. The abiotic template and its associated vegetation pattern. Pages 83-129 in Du Toit, J. T., K. H. Rogers, and H. C. Biggs, eds. The Kruger experience: ecology and management of savanna heterogeneity. Island Press. Washington DC. USA.

Table 1. Shannon-Weiner H’ values for each sample. There were no differences in H’ values between the granites and basalts samples (H=0.231, N=12, p =0.6310). H’ was generally higher in the night samples than in the day samples, although this was not a significant difference (H=3.102, N= 12, p =.0782).

Sample H'(loge) Basalt Site 1 Day 1.64 Basalt Site 2 Day 1.701 Basalt Site 3 Day 1.352 Basalt Site 1 Night 2.285 Basalt Site 2 Night 1.768 Basalt Site 3 Night 1.853 Granite Site 4 Day 2.008 Granite Site 5 Day 1.129 Granite Site 6 Day 1.154 Granite Site 4 Night 2.022 Granite Site 5 Night 1.933 Granite Site 6 Night 1.196

248

Site 3 Site 1 Site 2

Site 4

Site 5

Site 6

Figure 1. Map of sampling sites near Punda Maria, Kruger National Park, Mpumalanga Province, South Africa. Sites 1-3 are situated on basalt-derived soils and sites 4-6 are situated on granite-derived soils. There was at least 1.5km between each site, and 3 km between the closest basalt and granite sites. Each grid square on the map represents 1km x 1km.

249

Figure 2. MDS plot based on the sub-sample of morphospecies present in each sample. (Stress=0.04).

Figure 3. Cluster analysis based upon the sub-sample of morphospecies present in each sample.

250 30

26 24

22 20

16 14

12 Total Number of Beetles 10 8

6 4 Mean 2 ±SE Basalt Granite ±1.96*SE Soil Type

Figure 4. Comparison of total number of beetles per sample between sites on basalt-derived soils and granite-derived soils (H= 16.460, N=108, P<.001).

1.4

1.2

1.0

0.8

(g) Biomass Scarab 0.6

0.4

Mean 0.2 ±SE Basalt Granite ±1.96*SE Soil Type Figure 5. Comparison of scarab biomass per sample between sites on granite-derived soils and granite-derived soils (H=8.067, N= 108, P<.001).

251 50

Total Number of Beetles 15

Mean 0 ±SE Day Night ±1.96*SE Sampling Period Figure 6. Sampling period versus total number of beetles per sample on basalt-derived soils (H=15.520, N=54, P<.001).

2.4

2.2

2.0

1.8

1.6 1.4

1.2

1.0

Biomass Scarab 0.8

0.6

0.4

0.2 Mean 0.0 ±SE Day Night ±1.96*SE Sampling Period

Figure 7. Sampling period versus scarab biomass per sample on basalt-derived soils (H=19.325, N=54, P=.001).

252 20

Beetles of Number Total 6

2 Mean 0 ±SE Day Night ±1.96*SE Sampling Period

Figure 8. Sampling period versus total number of beetles per sample on granite-derived soils (H= 24.070, N=54, P=<.0001).

1.8

1.6

1.4

1.2

1.0

0.8

BiomassScarab 0.6

0.4

0.2

0.0 Mean -0.2 Mean±SE Day Night Mean±1.96*SE Sampling Period

Figure 9. Sampling period versus scarab biomass per sample on granite-derived soils (H= 25.379, N=54, P<.0001).

253 Appendix 1. Total number of beetles for each site. Sites 1-3 were on granite-derived soils and sites 4-6 were on basalt-derived soils Granites Basalts Site 1 AM Site 4 AM Sample Total No of Beetles Sample Total No. Beetles 1 7 1 5 2 1 2 2 3 15 3 3 4 4 4 1 5 5 5 2 6 2 6 0 7 6 7 1 8 3 8 0 9 0 9 0

Site 2 AM Site 5 AM Sample Total No.of Beetles Sample Total No. Beetles 1 1 1 1 2 10 2 1 3 9 3 2 4 11 4 3 5 22 5 4 6 20 6 1 7 16 7 1 8 9 8 2 9 9 9 3

Site 3 AM Total Number of Site 6 AM Sample Beetles Sample Total No. Beetles 1 7 1 0 2 2 2 0 3 4 3 0 4 3 4 1 5 2 5 0 6 6 6 2 7 2 7 2 8 8 8 1 9 8 9 1

Site 1 PM Total Number of Site 4 PM Sample Beetles Sample Total No. Beetles 1 35 1 23 2 0 2 28 3 4 3 14 4 21 4 22 5 7 5 35 6 30 6 13 7 5 7 18 8 83 8 5 9 6 9 13 254 Granites Basalts Site 2 PM Site 5 PM Sample Total No. Beetles Sample Total No. Beetles 1 27 1 11 2 104 2 14 3 97 3 17 4 6 4 44 5 143 5 1 6 26 6 12 7 14 7 8 8 21 8 7 9 38 9 52

Site 3 PM Site 6 PM Sample Total No. Beetles Sample Total No. Beetles 1 27 1 3 2 4 2 6 3 25 3 4 4 34 4 1 5 48 5 1 6 43 6 5 7 11 7 1 8 22 8 1 9 8 9 5

255 Appendix 2. Scarab biomass collected in each pitfall trap. Site 1-3 were on basalt-derived soils, and sites 4-6 were on granite-derived soils.

Basalts Granites Site 1 AM Site 4 AM Sample Scarab Biomass (g) Sample Scarab Biomass(g) 1 0.5 1 0.01 2 0.005 2 0.005 3 0.43 3 0.14 4 0.03 4 0.09 5 0.06 5 0.005 6 0.005 6 0 7 0.05 7 0.18 8 0.05 8 0 9 0 9 0

Site 2 AM Site 5 AM Sample Scarab Biomass (g) Sample Scarab Biomass(g) 1 0.01 1 0.002 2 0.58 2 0.002 3 0.08 3 0.004 4 1.59 4 0.006 5 0.31 5 0.04 6 0.43 6 0.002 7 0.24 7 0.002 8 0.16 8 0.004 9 0.22 9 0.008

Site 3 AM Site 6 AM Sample Scarab Biomass (g) Sample Scarab Biomass(g) 1 0.1 1 0 2 0.21 2 0 3 0.02 3 0 4 0.04 4 2.18 5 0.48 5 0 6 0.07 6 0.005 7 0.02 7 0.005 8 0.41 8 0.002 9 0.46 9 0.06

Site 1 PM Site 4 PM Sample Scarab Biomass(g) Sample Scarab Biomass(g) 1 1.68 1 5.03 2 0 2 0.99 3 0.12 3 0.72 4 1.87 4 1.84 5 0.15 5 1.56 6 1.33 6 1.47 7 0.17 7 0.5 8 3.52 8 0.99 256 Basalts Granites 9 0.15 9 0.52 Site 2 PM Site 5 PM Sample Scarab Biomass (g) Sample Scarab Biomass (g) 1 0.97 1 2.04 2 4.05 2 1.86 3 3.64 3 0.48 4 0.58 4 4.3 5 7.36 5 0.57 6 2.01 6 0.54 7 0.46 7 1.14 8 1.49 8 0.22 9 1.67 9 4.16

Site 3 PM Site 6 PM Sample Scarab Biomass (g) Sample Scarab Biomass (g) 1 1.36 1 0.27 2 0.19 2 0.11 3 0.45 3 0.09 4 3.16 4 0.01 5 1.71 5 0.002 6 2.78 6 0.13 7 0.47 7 0.05 8 0.62 8 0.002 9 0.94 9 0.16

257 Appendix 3. Appendix 3: Sub-sample of five morphological species identified from each catch (if less than five beetles were found in a trap, the sub-sample consisted of all beetles in the trap). Morphological species were numbered 1-33. Sites 1-3 were on basalt-derived soils, and sites 4-6 were on granite-derived soils.

Site Sample Sub-Sample of 9 None 3 PM 1 15, 15, 23, 25 Morphospecies 5 AM 1 15 2 24, 15, 15, 30 1 AM 1 2,2,3,4,5 2 15 3 15,15,13, 15, 15 2 1 3 15, 5 4 23,15,15, 15, 22 3 4, 2, 5, 6, 5 4 13, 13, 13 5 26,18,20, 27, 15 4 2, 2, 3, 3 5 15, 15, 15, 15 6 23,15,20, 15, 15 5 2, 6, 2, 2, 5 6 31 7 12,12,20, 15, 20 6 2,3 7 31 8 15, 13,15 13, 18 7 6, 7, 5, 2, 3 8 15, 15 9 26, 5, 20, 20, 15 8 6, 2, 2, 9 15, 15, 3 4 PM 1 23,23,23, 26, 24 9 None 6 AM 1 None 2 18,13,15, 27, 24 2 AM 1 9, 3, 2 None 3 15,15,14, 18, 15 2 5, 5, 2, 2, 2 3 None 4 20,13,22, 15, 15 3 9, 5, 2, 2, 9 4 33 5 15,23,14, 12, 15 4 1, 2, 2, 4 5 None 6 15,15,23, 18, 15 5 5, 5, 2, 9, 3 6 34, 15 7 15,15,15, 18, 15 6 2, 2, 5, 6, 1 7 15, 15 8 15,27,23, 15, 23 7 2, 6, 10, 5, 2 8 15 9 20, 5, 18, 15, 16 8 8, 5, 2, 3, 2 9 6 5 PM 1 15,27,12, 20, 15 9 2, 9, 6, 2, 2 1 PM 1 11,12,13, 14, 15 2 20,20,20, 14, 15 3 AM 1 2, 2, 9, 2, 2, 2 None 3 15,14,26, 23, 27 2 4, 5, 2 3 15,12,12, 15, 18 4 13, 13, 26, 27 3 2, 3, 9, 3 4 17,15,14, 14, 13 5 26 4 2, 2, 2, 5 12,15,18, 15, 15 6 27,13,26, 27, 20 5 4, 4, 6 12,17,17, 15, 15 7 15,20,20, 27, 27 6 3, 3, 2, 2, 2 7 18, 12,12,15 8 18, 15, 15, 2, 13 7 2 8 12,15,15, 13, 19 9 20,15,15, 15, 15 8 4, 6, 2, 2, 6 9 15, 15, 18, 20 6 PM 1 15, 20, 15 9 3, 2, 2, 3, 2 2 PM 1 11, 13, 15, 17 2 15,15,15, 15, 12 4 AM 1 29, 29, 29, 8, 30 2 15,15,18, 18, 20 3 13, 15, 18, 15 2 15, 32 3 18,15,15, 15, 15 4 None 3 3,2,8 3 19,20,15,15, 21 5 None 4 8, 8, 8, 5 22,15,15,15, 12 6 2, 15, 15, 15, 15 5 29, 18 6 15,18,15,13, 12 7 18 6 None 7 15,15,24, 13, 13 8 15, 8 7 20, 7 8 24, 24, 23, 15 9 18,24,15, 15, 15 8 8 9 15, 15, 20, 8

258

Appendix 4. Descriptions of Morphological Species Observed

Morpho spp #1: Approximately 4.5-5mm in length, coloring is metallic green. Body is highly punctuate, especially the pronotum. The elytra are large with longitudinal striations.

Morpho spp #2: Approximately 4.5mm in length, coloring is metallic brown or brownish red and metallic. Elytra are very hairy with distinct longitudinal striations and taper to a blunt curve. Pronotum has lateral fringe of curved hair. Hind tibia is elongate with distinct ridges.

Morpho spp #3: Approximately 3.5 mm in length, coloring is brown. Pronotum has wrinkly pattern. Hind and middle legs are elongate and slender, while the middle coxae is flat and elongate.

Morpho spp #4: Approximately 17mm in length, The beetle’s underside is metallic gold with brown markings, the antennae are yellow at the base with a dark brown club, the pronotum and head are bright blue or blue green with pale yellow on the margins, and the elytra are black and with pale yellow margins.

Morpho spp #5: Approximately 3mm in length, coloring is brown and underside is metallic brown. Pronotum is large, fairly punctuate, and extends backwards for ½ the length of the thorax and abdomen. Elytra have longitudinal striations.

Morpho spp #6: Approximately 6.5mm in length, coloring is light brown with metallic green patterns, while the underside is pale yellow w/ brown markings on thorax, the coxae is pale yellow distally and dark brown basally, and the antennae are totally brown.Generally flattened dorsally. There are many punctations on the pronotum, often with brown highlighting. Elytra are weakly striated with metallic green stripes.

Morpho spp#7: Approximately 5.5mm in length. Beetle is elongate with a black pronotum and head that are broadly joined. Elytra are light brown and with longitudinal black markings and striations. The coxae are light brown with lighter colored spots.

Morpho spp # 8: Approximately 7.5mm in length. Coloring is metallic brown. Elytra are hairy and have faint longitudinal striations. The pronotum is quadrate with short robust hairs. The hind legs are very long with short thin spines on the tibia and two robust spines on the femur. The femur, forelegs, tibia and coxae have long hairs, and there are also lateral fringes on the femur and the tibia.

Morpho spp#9: Approximately 4mm in length. The pronotum is punctuate with a fringe of hairs at the base of the head and can be metallic green to dark brown. The elytra are mottled light brown and black with a black outline. The tibia of the hind legs are metallic green and the underside is bronze. Male has two horns on head.

Morpho spp #10: Approximately 8mm in length. Pronotum is metallic green and tapers to a rounded angle. The elytra are light brown with black spots, the head is green and iridescent, while the underside is pale yellow and brown with green accents at the coxae and the fore tarsi.

Morph spp #11: Approximately 13mm in length, whole beetle is black. Pronotum is highly punctuate, quadrate has rounded frontal edges. Elytra is striated. Head is very shovel like and with two rounded bump things on tip of head. Tibia and tarsae are very spinose.

Morpho spp #12: Approximately 7mm in length. Coloring is green metallic with greenish-bronze underside, elytra have black striations, and tips of tibia have distinct row of spines that increase gradually in size.

Morpho spp #13: Approximately 9mm in length, coloring is all black, highly punctuate, like spp11 but smaller, underside brown. Base of head is slightly lighter than underside of body.

259 Morpho spp #14: Approximately 8mm in length, body coloring is brown with brown underside, pronotum is rounded and highly punctuate with longitudinal striations. Hind femurs have very few spines, and few spines arise btw base of mid coxae.

Morpho spp # 15: Approximately 5.5mm in length, coloring is metallic bronze to brown, underside also bronze. Beetle generally very punctate and hairy (underside too); base of tarsae has row of spines increasing gradually in size.

Morpho spp#16: Approximately 6mm in length, pronotum dark brown with many punctures but no hair. Elytra are light brown with longitudinal black markings, also punctate with no hair. Bottom side is metallic brown w/ strong green sheen. Coxae are brown and metallic but femur is light brown. The last segment of abdomen is also light brown.

Morpho spp #17: Approximately 5.5mm in length, coloring is metallic greenish-brown, pronotum is punctate without hair, elytra hairy with longitudinal striations, underside is dark metallic brown. Femurs are metallic brown or green at base, distally they become light brown.

Morpho spp #18: Approximately 8.5mm in length, completely light brown with two dark brown markings on pronotum at base of head and two other dark brown markings at back of pronotum. Head is dark brown. Thorax is darker brown than rest of underside, which is light brown

Morpho spp #19: Approximately 5mm in length. Beetle is elongate with smooth brown pronotum and hairy, striated, light brown elytra. The abdomen and legs are light brown on underside, and the underside of the thorax is brown.

Morpho spp #20: Approximately 16mm in length. Coloring is black. Male has one very large horn on top of head, four smaller horns on a quadrate pronotum. Elytra are weakly striated and smooth; in fact, whole body is smooth. Bottom of pronotum has fringe of small hairs on backside, few hairs beneath middle coxae on outside of thorax.

Morpho spp # 21: Approximately 5mm in length. Coloring generally light brown with greenish tint and spotted pattern of dark brown spots. Longitudinal striations on elytra are dark brown. The femur and coxae of each leg are pale yellow. Underside of thorax is dark brown with green metallic sheen. Abdomen is light brown on underside.

Morpho spp #22: Approximately 28mm in length. Beetle is reddish metallic with a rounded pronotum that has a broad lateral fringe. Underside is black with metallic red sheen.

Morpho spp #23: Approximately 19mm in length. Elytra are light brown with dark brown outlines and faint striations; pronotum is rounded and light brown with distinct metallic green or brown markings rather like butterfly. More elongate than #18. Coxae are dark metallic brown, rest of legs are pale yellow. Underside of body is light yellow with dark brown spots between coxae.

Morpho spp #24: Approximately 10mm in length. Coloring is metallic greenish brown, pronotum has pale yellow margins. Elytra are light brown. Male has head with two spines at back. Underside is mostly light brown with dark brown markings betwen mid and hind coxae. Coxae are light brown with dark brown ovals on them.

Morpho spp# 25: Approximately 6mm in length. Beetle is black and elongate with a smooth pronotum. Elytra have distinct longitudinal striations. Underside is dark brown.

Morpho spp #26: Approximately 6.5mm in length. Beetles is elongate and highly punctuate with a metallic reddish sheen. Underside is also highly punctate and metallic red; antennae are also bright red.

Morpho spp #27: Approximately 14mm in length. Coloring is metallic red, elytra are flattened with distinct striations, and the pronotum is punctate. Antennaeare light brown with a dark brown base. Underside is metallic reddish-brown. Hind femur have distinct sharp bump to them.

Morpho spp #28: Approximately 5mm in length. Pronotum is bronzy-red and punctate with small hairs. Elytra are light brown with darker brown margins and semi-transparent with small hairs. Underside is metallic bronzy-red. 260

Morpho spp #29: Approximately 5mm in length. Beetle is elongate and flattened. Pronotum has one plane and then two side slanting down, elytra also have one plane and then two slanting down on sides. Beetle is generally black, and the underside also black with greenish tint.

Morpho spp #30: Approximately 9mm in length. Coloring is metallic green and the pronotum has pale yellow margins. Elytra are light brown with black dashes and metallic green margins. Femurs are light brown with dark brown ovals. Area between mid and hind coxae is dark brown.

Morpho spp #31: Approximately 4mm in length. Beetle is elongate with light brown and a laterally wrinkly pronotum.

Morpho spp #32: Approximately 4.5mm in length. The entire body is generally smooth though slightly punctate. Beetle is generally black but elytra are semi-transparent. The bottom of the thorax is dark brown, and legs are lighter brown.

Morpho spp #33: Approximately 30mm in length. Coloring is red metallic; legs are very hairy.

Morpho spp #34: Approximately 4mm in length. Beetle has black, punctate pronotum. Elytra have longitudinal ridges with big hairs at the end. Underside has a middle ridge between the mid coxae and is punctate.

261

Back of the Book

262 Terminology that every American should know before coming to South Africa

South Africa American Bakkie Truck Torch Flashlight Stiffy Floppy Disk Jello Jelly Jelly Jam Tap Faucet Fish fingers Fish sticks Cold Drink Soda Just Now In a few minutes Hey huh, what, yeah, etc How’s It What’s up, How are you Ag Shame Sad, pity Robot (Robo) Traffic light Stumpy Cigarette butt Tick Check mark Pleasure You’re welcome Jug Pitcher Braai Similar to a barbeque Napkins (Nappy) Diaper Sub B 2nd Grade Sub A 1st Grade Slip slop Flip flop (Hot) Chips French fries Biscuit Cookie Nought and Crosses Tic Tac Toe Nought Zero Hash Pound (#) Sweet Melon Cantaloupe Potjie Small pot of food (ex. Stew) Lecturer Professor Takkies Tennis shoes Chemist Pharmacist Square Brackets Brackets Round Brackets Parenthesis Full Stop Period (.) Tomato Sauce Ketchup Tin Can Rubbish Garbage, Trash Boot Trunk of a car Hooter Horn of a car Ear bud Q-tips Bill Check (when you’re out eating) Note Bill ($)

Compiled by Blanchie w/ the help of the OTS Class of 2004

263 OTS Goldsworthy Images

Gareth and Jasper

264

Carla and Ben

265

Tammy

266

Eric

267

Taryn

268

Kyle

269

Gareth

270

Fahiema and Megan

271

Megan and Fahiema

272

Scott and Justine

273

Simon

274

Laurence

275

Tammy

276

Zoe and Sally

277

Deedra

278

Gareth

279

Laurence

280

Blanchie

281

Jasper and Gareth

282

Michael

283

Stephanie

284

Laura and Govan

285

Gareth

286

Deedra

287

Zoe and Sally

288

Julie

289 Nylsvley

290 Nylsvley

291 Nylsvley

292 Nylsvley

293 Wits Rural Facility

294 Wits Rural Facility

295 Kruger National Park

296 Kruger National Park

297 Kruger National Park

298 Kruger National Park

299 Skukuza

300 Skukuza

301 Skukuza

302 Skukuza

303 Skukuza

304 Skukuza

305 Skukuza

306 Skukuza

307 Skukuza

308 Skukuza

309 Skukuza

310 Ladysmith

311 Ladysmith

312 Ladysmith

313 Shingwedzi

314 Shingwedzi

315 Shingwedzi

316 Shingwedzi

317 Punda Maria

318 Punda Maria

319 Punda Maria

320 Punda Maria

321 Punda Maria

322 Punda Maria

323 Punda Maria

324 Train Ride

325 Train Ride

326 Cape Town

327 Robben Island

328 De Hoop

329 De Hoop

330 De Hoop

331 Skukuza 2

332 Skukuza 2

333 Skukuza 2

334 Skukuza 2

335 Skukuza 2

336