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smithsonian contributions to zoology • number 632

Conserving Wildlife 

in African Landscapes

Kenya’s Ewaso Ecosystem

Edited by

Nicholas J. Georgiadis
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smithsonian contributions to zoology • number 632

Conserving Wildlife

in African Landscapes

Kenya’s Ewaso Ecosystem

Edited by

Nicholas J. Georgiadis

washington d.c.

2011
ABSTRACT

Georgiadis, Nicholas J., editor. Conserving Wildlife in African Landscapes: Kenya’s Ewaso Ecosystem. 

Smithsonian Contributions to Zoology, number 632, vi + 123 pages, 37 figures, 10 tables, 2011.—During 

the last two decades, conservation strategies in Africa have changed from an almost exclusive focus on large 

mammals in protected areas to an emphasis on conserving ecological processes at the level of entire landscapes 

and on the role of human communities. The papers assembled in this volume address diverse aspects of 

conserving the Ewaso landscape in northern Kenya, where concerted and prodigious efforts to conserve 

wildlife and natural resources have achieved substantial progress. Topics range from interpreting evidence 

for continuity and change in patterns of human settlement in the region to describing ecological interactions 

between wildlife, people, and livestock that are harmful or helpful; from the challenges of adapting livestock 

management in the presence of predators to legal mechanisms for conserving wildlife habitat on private land. 

In the final chapter, results of a strategic planning exercise are described for conserving essential elements 

in the entire landscape—the first of its kind in Kenya. Today, national policy and political will are still 

insufficiently aligned with this landscape conservation imperative to effect the changes that are necessary to 

conserve Kenya’s biodiversity. We hope this volume will help propagate awareness about the importance and 

threatened status of Kenya’s ecosystems and promote confidence that a policy can be crafted that will reverse 

their decline.

Cover image: Giraffes in the Ewaso landscape of Kenya. Photo by Nicholas J. Georgiadis.

Published by SMITHSONIAN INSTITUTION SCHOLARLY PRESS

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Library of Congress Cataloging-in-Publication Data

Conserving wildlife in African landscapes : Kenya’s Ewaso ecosystem / edited by Nicholas Georgiadis.

 p. cm. — (Smithsonian contributions to zoology ; no. 632)

Includes index.

1. Wildlife conservation—Kenya—Ngiro River Watershed. I. Georgiadis, Nicholas. 

QL84.6.K4C664 2011

333.95’416096762—dc22

2010042306

ISSN: 0081-0282 (print) / 1943-6696 (online)

The paper used in this publication meets the minimum requirements of the American National Standard 

for Permanence of Paper for Printed Library Materials Z39.48–1992. 
Contents

Acknowledgments v

Introduction: Conserving Wildlife in Kenya’s Ewaso Landscape 1

Nicholas J. Georgiadis

An Outline of the Later Holocene Archaeology and 

Precolonial History of the Ewaso Basin, Kenya 11

Paul Lane

Reassessing Aerial Sample Surveys for Wildlife 

Monitoring, Conservation, and Management 31

Nicholas J. Georgiadis, Nasser Olwero, Gordon Ojwang’, 

and George Aike

Ecological Importance of Large Herbivores in the Ewaso Ecosystem 43

Robert M. Pringle, Todd M. Palmer, Jacob R. Goheen, 

Douglas J. McCauley, and Felicia Keesing 

Pathways for Positive Cattle–Wildlife Interactions in 

Semiarid Rangelands 55

David J. Augustine, Kari E. Veblen, Jacob R. Goheen, 

Corinna Riginos, and Truman P. Young

Living with Lions: Lessons from Laikipia 73

Laurence Frank

Attitudes Toward Predators and Options for Their Conservation 

in the Ewaso Ecosystem 85

Stephanie S. Romañach, Peter A. Lindsey, and Rosie Woodroffe

Conserving Wildlife on Private Land: The Legal Framework 

for Landownership and New Tools for Land Conservation 

in Kenya 95

Nyokabi Gitahi and Kathleen H. Fitzgerald
i v • smithsonian contributions to zoology

Landscape-Scale Conservation Planning of the Ewaso Nyiro: 

A Model for Land Use Planning in Kenya? 105

Karl A. Didier, Alayne Cotterill, Iain Douglas-Hamilton, 

Laurence Frank, Nicholas J. Georgiadis, Max Graham, 

Festus Ihwagi, Juliet King, Delphine Malleret-King, 

Dan Rubenstein, David Wilkie, and Rosie Woodroffe
Acknowledgments

This book was made possible by a Lone Mountain Fellowship to me at 

the Property and Environment Research Center (PERC) in Bozeman, 

Montana, where I was a resident fellow during its development. I 

was also supported by a Senior Fellowship from the Mpala Wildlife 

Foundation.

Nicholas J. Georgiadis
FRONTISPIECE. Typical assemblage of herbivores (giraffes and zebras) in the Ewaso landscape of Kenya. Mount Kenya is in the background. 

Photo by Nicholas J. Georgiadis.
Introduction: Conserving Wildlife in Kenya’s 

Ewaso Landscape

Nicholas J. Georgiadis

The papers assembled in this volume are disparate in subject, discipline, 

treatment, and outlook, but they have two unifying threads. The first 

is that directly or indirectly, they all address aspects of what in recent 

decades has become a global environmental imperative, generally referred to as “landscape conservation.” The second is that all chapters focus 

geographically on the Ewaso landscape in northern Kenya (defined in Box 1; 

outlined by a rectangle in the inset of Figure 1), where concerted and prodigious efforts to conserve wildlife and natural resources over the last 20 years 

have achieved substantial progress. This work is not a “how-to” handbook, nor 

does it aim to cover all aspects of this increasingly complex and multifaceted 

endeavor. Rather, it provides an opportunity to document progress in ways intended to facilitate further advancement toward landscape conservation in this 

and other regions like it. 

In this introduction, a brief history recounts how conservation strategies at 

global, national, and local levels have changed from an almost exclusive focus 

on charismatic species and protected areas to emphasizing landscape-level processes, human communities, and economic incentives. The structure of the book 

is described, with sketches of the rationale and content of each paper. The final 

section sketches the outlook for wildlife in Kenya’s landscapes.

CHANGING PARADIGMS IN CONSERVATION

The late 1980s marked a turning point for conservation in a district of 

northern Kenya, across much of Africa, and throughout the world in distinct 

but related ways. In the Laikipia rangelands of northern Kenya, ranchers who 

had once regarded wildlife as a liability began to collaborate with each other 

in the conservation and management of wildlife on their land. At the time, the 

Nicholas J. Georgiadis, Property and Environment Research Center, 2048 Analysis Drive, Suite 

A, Bozeman, Montana 59718, USA. Correspondence: georgiadis.nick@gmail.com.

Manuscript received 7 January 2009; accepted 18 

May 2010.
2 • smithsonian contributions to zoology

dominant conservation paradigm in Africa, which had 

fixated on large mammals in protected areas for over 60 

years, began to be replaced by a more inclusive approach, 

which recognized humans as an integral component of 

biodiverse landscapes and the need for conservation to account human needs and effects. The persistence of biodiversity was seen to depend on active conservation not only 

in parks and reserves but also in areas that lacked formal 

protection, and were occupied by humans and livestock 

(see discussion in Hutton et al., 2005). At the same time, a 

conceptual revolution among public and private agencies 

in the United States sparked what is today a global conservation agenda, originally known as “ecosystem management” (Grumbine, 1994, 1997; Keough and Blahna, 

2006), that aims to sustain the integrity of ecological processes across vast landscapes. Twenty years on, the papers 

in this volume document some of the approaches and 

progress toward landscape conservation that have been 

achieved in the Ewaso landscape of northern Kenya.

Ranchers in Laikipia District, the southwestern sector 

of the Ewaso region, were motivated to promote wildlife 

as a source of income by a weakening beef market (Heath, 

2000) and by growing recognition of the potential value 

of wildlife on their land. After a 14-year hiatus, Kenya 

was once again experimenting with sanctioned harvesting 

of wildlife on private land (Kock, 1995). Those wishing to 

participate were required by the national conservation authority, Kenya Wildlife Service (KWS), to create and join 

management associations. To meet this need, the Laikipia 

Wildlife Forum (LWF) was launched in 1991 as a company of private and communal landholders with a shared 

interest in wildlife management and conservation. With 

investors foreseeing greater returns from nonconsumptive 

uses of wildlife, ecotourism enterprises proliferated on 

private and communal properties in the region. Cooperation among landholders favoring wildlife was spurred by 

growing awareness of the need to maintain sufficient space 

for species like elephants, lions, and wild dogs, whose survival depended on continuing freedom to range across far 

greater areas than even the largest individual property.

That most of Africa’s protected areas were never designed to maintain, on their own, ecologically viable populations of these wide-ranging mammals has long been 

recognized and often repeated, at least in the conservation 

BOX 1. Defining the Ewaso ecosystem.

For the purposes of this book, the Ewaso ecosystem is arbitrarily defined by the geographical extents of two large-scale processes that have been selected as long-term targets for conservation: dry-season river flow and elephant migration. Two “water 

towers” that collect much of the region’s rainfall, the Aberdare Range (3,999 m) in the southwest and Mount Kenya (5,199 m) 

in the southeast, are drained to the north by many streams that ultimately combine to form two perennial rivers (Figure 1). These 

are the Ewaso Nyiro and Ewaso Narok, which are confluent in the central “Laikipia Plateau,” continuing to flow northward and 

then eastward through the Samburu, Buffalo Springs, and Shaba National Reserves and ultimately to the Lorian Swamp. This 

drainage system is not exceptional in that water abstraction for irrigation and household use has escalated in recent decades, 

particularly in the upper reaches of the catchment, leaving progressively less water available for people, livestock, and wildlife 

in the drylands below. Accordingly, the functional ecosystem “boundary” encompasses the upper drainage basin of the Ewaso 

Nyiro and Ewaso Narok rivers as far as the eastern boundary of Samburu District.

Elephants have been shown to migrate immense distances up and down elevation and rainfall gradients in this region, 

seeking food and water on a seasonal basis. Reasoning that the viability of elephants in the region depends on their seasonal 

migrations, and that if elephant migration can be conserved, then many other species and ecosystem processes will also benefit, 

elephants have assumed strategic prominence in conservation planning for this landscape. Accordingly, the known extents of 

their migrations beyond the Ewaso Basin are also encompassed within the rectangle that delineates the Ewaso ecosystem.

These criteria define a vast (~40,000 km2) and geographically diverse region, including the entirety of Laikipia District 

(9,666 km2) and parts of adjacent Samburu, Meru, Isiolo, and Marsabit Districts to the north and east (Figure 2). A wide diversity 

of habitats, land cover types, and land uses (Figure 2) are associated with the elevation and climatic gradients that characterize 

this region, from cool, wet highlands in the south to hot, dry lowlands in the north. Most of the wildlife occur in the wetter, southwestern sector of the region, on private ranches in Laikipia District (Figure 3). These range from alpine moorlands above 3,000 m 

on Mount Kenya and the Aberdare Range, through protected montane rain forests and an intensively cultivated moist zone to dry 

savanna grasslands and bushlands at lower elevations. Mean annual rainfall varies from around 300 mm in the northeast to more 

than 1000 mm in the south, increasing with elevation to around 1300 mm in the forested zones on Mount Kenya. Rainfall is highly 

variable but trimodal, falling mainly in April–May, November, and August, with a pronounced dry season in January–March.
number 632 • 3

literature (Soulé et al., 1979; Western and Ssemakula, 

1981; Western, 1982, 1989; Western and Gichohi, 1993; 

Newmark, 1996, 2008). In most cases, the wild animals 

that generate tourism revenue for national or local governments within unfenced parks and reserves would soon 

perish if denied access to surrounding lands that are privately or communally held. But to most who own or use 

this land, wildlife is a liability. They may incur losses from 

crop raiding or be threatened or even killed by elephants, 

and their livestock are vulnerable to predators and disease (Woodroffe et al., 2005). Almost invariably, wildlife 

ultimately lose these conflicts with humans, becoming displaced by development and incompatible land use changes. 

As a result, wildlife is increasingly confined to protected 

FIGURE 1. Relief and drainage in the Ewaso ecosystem. The region is bounded in the west by the Rift Valley 

and in the south by Mount Kenya and the Aberdare Highlands. Protected areas are outlined in white as follows: 1, Samburu National Reserve (NR); 2, Shaba NR; 3, Buffalo Springs NR; 4, Mount Kenya National 

Park (NP); 5, Aberdare NP; 6, Karisia Hills Forest Reserve (FR); 7, Maralal FR; 8, Matthews Range FR; 

9, Loisai NR; 10, Ndoto Mountains FR. Major rivers are as follows: 11, Ewaso Nyiro; 12. Ewaso Narok.
4 • smithsonian contributions to zoology

islands that are too small to ensure their long-term viability. With diminishing options to expand existing parks and 

reserves, attention turned toward habitat and biodiversity 

conservation in areas that were not formally protected, 

most of them communally owned or used. A variety of 

conservation initiatives, labeled integrated conservation 

and development projects, community-based conservation 

projects, and numerous variations on this theme spread 

across the continent from southern Africa, aiming to create incentives for landholders and land users to conserve 

wildlife and natural resources (e.g., Western and Wright, 

1994; Barrow and Murphree, 2001; Fabricius et al., 2004; 

Child, 2004).

Similar concerns in North America about the inadequacy of national parks to sustain viable populations of 

large mammals like grizzly bears and progressive habitat 

fragmentation in natural areas adjacent to parks prompted 

a search for ways to conserve large-scale ecological processes across diverse landscapes, only parts of which were 

formally protected. Today, an ecosystem or landscape approach dominates most conservation strategies, aspiring 

to integrate the use and management of natural resources 

FIGURE 2. Land use in the Ewaso ecosystem.
number 632 • 5

in sustainable ways across spatial scales that better match 

the historical extents of natural ecosystem processes (e.g., 

Pirot et al., 2000).

Economic rationale provided the impetus for these 

changing conservation strategies. In many African cases, 

landholders and land users were not expected to conserve 

biodiversity unless they benefited, largely by realizing 

market values of wildlife via ecotourism, trophy hunting, or sale of live animals (e.g. Child et al., 1997). The 

ecosystem management approach went still further, ultimately calling for the full value of “ecosystem services” 

to be assessed and captured (e.g., Daily, 1997), for the full 

costs of economic development to be accounted, and for 

development and exploitation of natural resources to be 

appropriately priced and thus truly sustainable.

About to come of age, these community- and marketbased approaches to landscape conservation have had 

mixed results. Some community-based projects in Africa have made remarkable progress (e.g., Weaver and 

Skyer, 2003), but most have struggled to alleviate poverty, 

conserve biodiversity, or both, for several reasons (Newmark and Hough, 2000). The conception and operation of 

community-owned enterprises have inherent uncertainties 

as to individual rights and communal benefits, hampering solutions that are both workable and equitable (Logan 

and Moseley, 2000; Mburu and Birner, 2002; Walpole and 

Thouless, 2005). National or local governments have been 

reluctant to devolve to local communities the responsibility for and ownership of—or at least rights to—land and 

wildlife (Child and Jones, 2006). The management and 

governance capacity of communities often fall short of the 

minimum needed to sustain small businesses (Spiteri and 

Nepal, 2006). Wildlife and natural habitats continue to 

lose ground wherever competing land uses provide greater 

benefits to landholders (Norton-Griffiths and Southey, 

1995). Conserving ecosystem services will not necessarily 

conserve biodiversity (Ghazoul, 2007). Basing a strategy 

on biodiversity does not necessarily conserve large-bodied 

wildlife. And it seems unlikely that the inclusive costs of 

development, or the full value of ecosystem services, will 

be captured by markets soon enough to make a substantial 

difference (Egoh et al., 2007; Kroeger and Casey, 2007). 

For many large African mammals, their habitats, and the 

large-scale ecological processes that they define, overall 

trends continue downward.

These discouraging trends have led some to reconsider 

the validity and potential of incentive-based approaches 

to conservation in human-occupied landscapes and to 

recommend a return to providing sanctuary for wildlife 

in additional protected areas (see discussion in Hutton et 

al., 2005), or Transfrontier Conservation Areas (Hanks, 

2003; Wolmer, 2003; Munthali, 2007). Although this will 

help, opportunities to create or expand protected areas 

continue to diminish. Moreover, sustaining natural processes in areas that are not formally protected is hardly 

optional, not only in the survival interests of biodiversity 

but also of humanity. A more hopeful outlook sees progress to date as an exercise in defining the complexity and 

scope of the challenge (Adams and Hulme, 2001), which 

is colossal and has been persistently underestimated.

TOWARD LANDSCAPE CONSERVATION 

IN NORTHERN KENYA

Despite media-generated impressions of teeming wildlife, Kenya is not exceptional in that most of its wildlife 

populations have been declining, and their habitats fragmenting, for decades (Box 2). These trends and their 

causes have been documented at both local (e.g., Lamprey 

Marsabit

District

Samburu

District

Isiolo

District

Laikipia

District

↑

↑

FIGURE 3. Relative abundance and distribution of wildlife (black 

dots) in the Ewaso ecosystem, with the majority on pro-wildlife 

ranches in Laikipia District (source: Department of Resource Surveys and Remote Sensing survey, February 2001).
6 • smithsonian contributions to zoology

and Reid, 2004) and national levels. Habitat loss and fragmentation due to expanding settlement, land subdivision, 

and spread of incompatible land uses such as cultivation, 

clearing of woodland for charcoal production, consumption of bushmeat, displacement by livestock, proliferation 

of firearms, and local outbreaks of the rinderpest virus 

(today eradicated globally) are widely recognized to be the 

proximate causes. There is broad agreement that transformative policy measures are needed to arrest, preferably to 

reverse, these trends, but consensus is lacking as to what 

those measures should be. Projecting these seemingly irresistible trends forward, it is hard to conclude other than 

that Kenya’s wildlife, its wilderness image, and its tourism 

sector have a bleak outlook.

The primary challenge is to justify enough space for 

wildlife so that large-scale ecological processes that ensure 

their survival, such as seasonal migration by elephants, 

can continue as before. Toward these ends, much effort 

and donor funding have been spent in Kenya, attempting 

to engage landholders in wildlife conservation, to some 

effect. Stable or increasing wildlife trends on some private 

and group-owned properties provide rare exceptions to 

nationally declining trends (Box 2).

The most substantial progress toward landscape conservation has been made in the Ewaso region of northern 

Kenya (Figure 1, Box 1). In its southwestern sector (Laikipia District), wildlife was once suppressed to promote beef 

production on large-scale ranches. Following the switch 

from ranching to conservation and ecotourism, which 

began in the 1980s, wildlife abundance increased dramatically and today is exceeded only in Kenya’s premier 

reserve, the Masai Mara (which is much larger and wetter, 

and therefore more productive). The region holds more 

endangered mammals than anywhere else in the nation, 

including the world’s largest remaining concentrations of 

Grevy’s zebra and reticulated giraffe. More than 30 ecotourism ventures and community enterprises based on 

natural products have sprouted. Seven rhino sanctuaries 

BOX 2. Trends and status of wildlife in Kenya since 1977.

The dynamics of large-bodied (>20 kg) herbivore species have been monitored throughout Kenya’s rangelands since 1977 

by the Department of Resource Surveys and Remote Sensing (DRSRS, formerly the Kenya Rangeland Ecological Monitoring Unit, 

KREMU). Using mostly aerial sampling methods (Norton-Griffiths, 1978) and repeatedly surveying entire districts at a time, results 

are sufficient to measure coarse changes over vast areas. There have been several summaries of overall trends, all showing net 

declines in most areas (Figure 4). Grunblatt et al. (1996) concluded that wildlife in Kenya’s rangelands declined by 33% between 

1977 and 1994. Further analyzing results of de Leeuw (1998), Western et al. (2009) concluded that wildlife abundance in 17 

districts declined by 38% between 1977 and 1997. A separate analysis by Ojwang’ et al. (2006) showed declining wildlife 

trends in Kajiado District to continue into the early 2000s (excepting plains zebra and elephants). At these rates, and assuming 

a linear trend, the overall decline by 2011 is projected to be about 60%.

Supplementing data from additional sources, Western et al. (2009) showed that trends inside protected areas were also 

mainly downward. Between 1977 and 1997 wildlife declined in most protected areas, precipitously in Tsavo and Meru National Parks. In the 1990s steep declines were recorded in Nairobi, Nakuru, and Amboseli National Parks. In areas adjacent 

to protected areas, wildlife declined sharply around Tsavo National Park and the Masai Mara National Reserve (Western et al., 

2009; see also Ottichilo et al., 2000; Homewood et al., 2001; Thompson and Homewood, 2002).

There have been some notable exceptions. Steeply declining elephant and rhino numbers in the 1970s and 1980s were 

reversed in the 1990s. For elephants, this was achieved by an international ban on ivory trading and stringent enforcement 

within Kenya. For black rhinos, a well-organized collaboration between national wildlife authorities and private landowners has 

allowed numbers in small but heavily guarded sanctuaries to increase.

Perhaps most promising, wildlife numbers in private conservation areas have remained steady or even tended to increase 

in recent decades (Western et al., 2009). In Laikipia District (the southwestern sector of the Ewaso ecosystem), wildlife increases 

between 1977 and the early 1990s were largely due to a resurgence of plains zebra and elephants. Since 1990, declines in 

some prey species, especially on properties investing in conservation, were ultimately found to signal a further conservation 

success: the restoration of predators in that region (Georgiadis et al., 2007a).

Today, the largest proportion of Kenya’s wildlife, about 40%, is found on private and communally owned land (Western 

et al., 2009). An additional 25% is scattered at low density across vast stretches of dry rangeland, much of it held in trust by 

local governments. Of the 35% remaining within formally protected areas, fully two-thirds are in the Masai Mara Reserve, with 

Amboseli and Nakuru National Parks holding most of the rest. The future of Kenya’s wildlife depends on the success of conservation on private and communally owned lands that lack formally protected status.
number 632 • 7

have been established. Furthermore, an initial strategic 

conservation plan for the landscape has been drafted.

The importance of this region is that wildlife share 

most of the largely unfenced landscape with varying densities of people and livestock. Only about 5% of the total 

land area of around 40,000 km2 has formally protected 

status, with an additional 1% set aside exclusively for 

wildlife in (private) fenced reserves. Land uses and management practices have varied widely as landownership 

and attitudes toward wildlife among landholders have 

changed over the last century. The result is a mosaic of 

properties with contrasting histories, land uses, management practices, ownership attitudes, and densities of 

livestock and wildlife (Georgiadis et al., 2007b). Such diversity over time and space provides a microcosm of the 

variability in land use existing across much of dryland Africa, and an instructive model system for large-mammal 

conservation in human-occupied landscapes.

Conservation success in the Ewaso region has spawned 

new problems, including aggravated conflict between humans and wildlife in some areas. However, necessity, in 

turn, is breeding creative solutions to these problems. Solutions have emerged from original research, based largely 

at Mpala Research Centre, from engagement with communities, through the Laikipia Wildlife Forum, Lewa Wildlife 

Conservancy, Northern Rangeland Trust, and others, and 

FIGURE 4. Declines and increases in wildlife densities in each 10 km grid square across Kenya’s rangelands between 1977 and 1996. The Ewaso 

ecosystem is outlined by the rectangle. Declines predominate nationally, but the southwestern sector (Laikipia District) of the Ewaso region was 

one of few areas in which wildlife abundance increased (source: World Resources Institute, 2007).
8 • smithsonian contributions to zoology

from a strong entrepreneurial spirit among many of the 

landholders themselves. Although the collaboration that 

produced these advances is likely to remain unique, many 

of the emerging solutions are applicable, with appropriate 

adaptation, elsewhere.

THE PAPERS

The papers in this volume document approaches, 

tools, and practices that have led to conservation advances 

in the Ewaso region. They add to the record of progress in 

ways intended to promote conservation in this and other 

unprotected landscapes like it. The targeted readership is 

landholders and land users, wildlife authorities, the donor 

and investment sectors, and conservation communities in 

similar landscapes elsewhere. The rationale and focus of 

each of the papers in this collection are sketched below.

In the paper following this introduction, Lane begins 

by outlining the pre-Colonial history of human occupation in the Ewaso region, describing interactions between 

pastoralist and foraging communities, and interpreting evidence for continuity and change in patterns of settlement.

In the third paper, Georgiadis et al. address the challenge of monitoring wildlife and the human attributes that 

compete with wildlife for space and resources in unprotected landscapes. Aerial sample surveys, a method widely 

applied to census wildlife and livestock in Africa, are reassessed in the context of meeting needs for finer-grained 

information at a time of escalating costs.

In the fourth paper, Pringle et al. review results of 

long-term studies that examine the indirect ecological effects of large herbivorous mammals, including cattle, on 

less charismatic species that share the landscape, including 

small mammals, reptiles, invertebrates, and soil microbes. 

These studies paint a clearer portrait of how humans and 

livestock modify natural systems and what conservationists mean when they refer to maintaining or restoring 

“ecological integrity.”

In the fifth paper, Augustine et al. synthesize research 

on rangelands in central Laikipia, highlighting positive

interactions between cattle and wild herbivores. They 

describe how livestock management practices can benefit 

wildlife through soil nutrient redistribution and how native browsers can enhance forage production for grazers.

The sixth and seventh papers focus on how humans 

and predators live with each other in this landscape. 

Frank’s studies reveal how management of both livestock 

and lions can be adapted to enhance coexistence between 

them. Romañach et al. assess attitudes toward predators 

by landholders and land users in the Ewaso region and 

their tolerance of livestock losses to predators, relative to 

socioeconomic factors.

In the eighth paper, Gitahi and Fitzgerald describe the 

legal framework for landownership in Kenya, highlighting 

existing and proposed mechanisms for conserving wildlife habitat on private land, such as easements and leases. 

These tools are not yet widely used, but they present some 

of the most promising mechanisms for providing the space 

wildlife will require to survive outside of protected areas.

Finally, Didier et al. describe the results of a strategic 

planning exercise for conserving the Ewaso landscape, the 

first of its kind in Kenya. More than 40 researchers and 

conservationists collaborated to define and prioritize conservation and investment needs for specific biodiversity 

and development objectives in a spatially explicit manner.

OUTLOOK

Kenya formally recognized the importance of conserving ecosystem processes in 1995 and was one of the first 

countries in the world to adopt an objective, systematic 

approach to mapping and prioritizing landscapes and seascapes for conservation at the national level. The policy 

was abandoned after only a year, and institutional turmoil prevented readoption of an ecosystem conservation 

agenda until 2005. In the interim, with no clearly defined 

strategy to guide conservation action, the nation’s biodiversity continued to dwindle. Today, national policy and 

political will are still insufficiently aligned with this ecosystem imperative to effect the changes that are necessary 

to conserve Kenya’s biodiversity. The principal challenge 

is to justify enough space for wide-ranging species like elephants and large predators to persist in ecologically viable 

populations. Given human population demands for land, 

that nonhuman land users do not vote politicians into 

office, and that economic returns from wildlife are typically lower than from competing interests, the likelihood 

of meeting this challenge for many areas is low. We hope 

this volume will help to propagate awareness about the 

importance and threatened status of Kenya’s ecosystems 

and promote confidence that a policy can be crafted that 

will reverse their decline.

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An Outline of the Later Holocene 

Archaeology and Precolonial History 

of the Ewaso Basin, Kenya

Paul Lane

ABSTRACT. Drawing on a range of archaeological, paleoenvironmental, linguistic, ethnographic, and historical data, this chapter outlines what is currently known about the 

trajectories of pastoralism in the Ewaso Basin from its initial appearance some 4000 

years ago until the early twentieth century, by which time new systems of land use and 

tenure had been established within the context of British colonial rule. Overall, the evidence indicates that many different groups have occupied parts of this area at different 

times in the past, and that these have encompassed both hunter-gatherer and pastoralist 

communities and speakers of various Nilotic, Cushitic, and Bantu languages, among 

others. There is also good evidence to suggest that the boundaries between different “ethnic,” “subsistence,” and even linguistic groups have been fairly fluid and that cultural 

intermixing as well as interaction through exchange and other social mechanisms was 

common. Additionally, at least four broad phases of pastoralist practices can be identified for the period prior to the twentieth century. These phases entailed different forms 

and levels of mobility, including population migration and seasonal movements, and had 

diverse environmental consequences, some of which appear to have enhanced biodiversity and ecological resilience whereas others initiated significant changes to the structure 

of the vegetation mosaic and may have even resulted in localized rangeland degradation.

INTRODUCTION

The primary aim of this article is to outline what is currently known about 

the trajectories of pastoralism in the Ewaso Basin from its initial appearance some 

4,500 years ago until the early twentieth century, to develop a better understanding of the changing nature of different pastoralist societies over this time period. 

Before considering this evidence in detail, it is worth noting that in contrast to 

some other areas of East Africa (such as the Central Rift Valley), the Ewaso 

Basin has not been subject to intensive archaeological research until recently. 

Consequently, the archaeological record concerning the trajectory of pastoralism 

is both patchy and partial. Surviving archaeological traces nevertheless include a 

combination of rock shelters and open sites with surface scatters of flaked stone, 

Paul Lane, Department of Archaeology, University of York, The King’s Manor, York YO1 7EP, 

UK; also, School of Geography, Archaeology and 

Environmental Studies, University of the Witwatersrand, Johannesburg, South Africa. Correspondence: paul.lane@york.ac.uk.

Manuscript received 7 January 2009; accepted 18 

May 2010.
1 2 • smithsonian contributions to zoology

pottery, iron slag, and iron smelting furnaces. Several of the 

rock shelters also contain rock paintings. The ceramics can 

be linked to regional archaeological typologies, with occurrences of Pastoral Neolithic (PN) Nderit, Elmenteitan, 

Marangishu, and Akira ceramic types all being reported, as 

well as a ceramic tradition known as “Kisima Ware” (Siiriäinen, 1984), which is believed to be associated with post 

AD 800 Pastoral Iron Age (PIA) communities (Table 1). 

At least five different types of stone cairn are also known, 

some of which were definitely used for human burial (Lane 

et al., 2007). On the basis of oral traditions among different ethnic groups that have a history of utilizing parts of 

Laikipia and Leroghi plateaus, however, others may have 

been used as markers for stock routes, as cenotaphs for 

fallen warriors, as markers associated with captured (or 

recaptured) cattle, and even simply as meat stores.

The available historical sources indicate that from ca. 

AD 1750 to the early twentieth century, various pastoral, Maa-language-speaking Iloikop sections (including 

Laikipiak, Samburu (Burkeneji), Mumonyot, LeUaso, 

and Purko Maasai communities) coexisted alongside 

scattered groups of other peoples that included huntergatherers (notably, Dikirri Dorobo and Mukogodo); the 

area probably also included pockets of Oromo-speaking 

Warra Daaya and possibly Turkana and Rendille along 

the northern fringes and various Bantu language speakers (including Meru and Kikuyu) along the more southern boundaries (Kenya Land Commission, 1934; Cronk, 

1989, 2004; Lamphear, 1993; Sobania, 1993).

In 1904, a large portion of the Laikipia and Leroghi 

plateaus was designated part of the Northern Maasai Reserve, and in 1911, following pressure from various quarters, a new agreement was signed between Maasai elders 

and the British East Africa (BEA) protectorate administration, under the terms of which Maasai occupying the 

Northern Reserve were to be relocated to the Southern 

Reserve and some 4,500 square miles (approx. 11,655 sq. 

km) of land handed over to the BEA protectorate for European settlement (Hughes, 2005, 2006). At the time of 

this move, the Maa-speaking population of Laikipia was 

estimated to have numbered around 10,000 and was 

dominated by members of the Purko, Keekonyukie, and Il 

Dalalekutuk sections, following their defeat of the Laikipiak Maasai (who belonged to a different moiety) during 

the intersectional Iloikop wars in the 1870s (Weatherby, 

1967; Waller, 1979; Sobania, 1993). The bulk of the Maaspeaking population was moved from the designated area 

between 1911 and 1913, along with up to approximately 

one million sheep and goats and 175,000 cattle (Hughes, 

2006:50), and the area became part of the “White Highlands” set aside for European settlement.

After the First World War, soldier settlement schemes 

started to bring European settlers to Laikipia, gradually 

creating an export-oriented economy, with land use strategies being transformed so as to maximize levels of production. Settlement took off in the early 1920s, with the 

number of settlers in occupation of their farms increasing from 18 in March 1920 to 42 in March 1921. Not 

TABLE 1. Approximate dates of Pastoral Neolithic (PN) and Pastoral Iron Age (PIA) pottery traditions for historically documented 

ethnic groups (based on Gifford-Gonzalez, 1998, 2000; Karega-Mu ̃ nene, 2002). Abbreviations not defined below: LSA, Later Stone 

Age; MIA, Middle Iron Age.

Age 

BP Archaeological Selected LSA stone LSA, PN, and PIA Iron age

(kyr) periodization tool industries ceramic traditions ceramic traditions

1 Later Iron Age (LIA), Kisima, ca. Cal AD 1200–1750; MIA and LIA regional variants,

 Pastoral IA (PIA) Lanet/Sirikwa, ca. Cal AD 750–1600 from ca. Cal AD 1000

2 Early Iron Age, overlaps Elmenteitan Akira, ca. Cal AD 100–800; Kwale, ca. Cal AD 150–1000;

 with PN and LSA Maringishu, ? to ca. Cal AD 350; Urewe, ca. 650 Cal BC to

 Elmenteitan, ca. 1600 Cal BC to Cal AD 700; Cal AD 1000

 Narosura, ca. 1000 Cal BC to Cal AD 620 

4.5 PN, overlaps with Eburran 5, ca. 3000 Cal BC 

 later LSA to Cal AD 750; Nderit and Ilert, ca. 3800–1800 Cal BC (PN)

 LSA “bone harpoon” sites, 

 ca. 8100–2500 Cal BC

6 Late LSA Kansyore; Eburran 4 Kansyore, ca. 6300 Cal BC to Cal AD 700
number 632 • 1 3

all the land was immediately taken into productive use, 

however. Some prospective settlers simply visited but left 

soon afterward, finding the land either waterless or too 

stony or bushy for their liking, such that even in 1922, 

200 alienated farms still had to be taken up. Even as late 

as 1937, there were still substantial areas of unoccupied 

surveyed land in the east of the district (see Waller, 2004, 

and Vaughan, 2005, for further detail concerning colonial 

land use history of Laikipia District). With independence 

in 1963 and a policy of Africanization, several, but not 

all, Europeans sold their ranches, a number of which were 

subdivided and settled by small-scale farmers originating 

in the densely populated Kikuyu areas located south of 

the district (Carey Jones, 1965; Kohler, 1987). The sale of 

large-scale commercial ranches to cooperatives and landbuying companies continued at least into the 1990s, and 

as a consequence, approximately one-third of the original large-scale ranches have now been subdivided. Today, 

Laikipia District comprises a mosaic of different land use 

types, including large-scale ranching, small-scale arable 

farming, traditional pastoralism, and forestry, with ~8.4% 

under cultivation, mostly concentrated in West Laikipia 

and around Nanyuki, even though only 1.7% of the 

district is classified as having high agricultural potential 

(Huber and Opondo, 1995; Kiteme et al., 1998). These 

and related events transformed earlier systems of land use 

and management, replacing them with a largely marketoriented system of production and land tenure, albeit 

in a piecemeal fashion drawn out over several decades. 

Pastoralists have nevertheless remained in this landscape 

and are still a significant presence in some areas of the 

Ewaso Basin, especially on the Leroghi Plateau. Moreover, 

when taken together and despite its limitations, the available archaeological evidence provides a good indication of 

the antiquity of pastoralism in the area, confirmation of a 

long history of interaction between pastoralist and foraging communities, and a record of continuity and changes 

in settlement practices, local environments, and material 

culture traditions, and it is these aspects that are reviewed 

in detail in the following sections. 

LATER STONE AGE HUNTER-GATHERERS 

AND THE BEGINNINGS OF PASTORALISM

The final phase of the Later Stone Age (LSA) in central 

Kenya dates from ca. 4400 Cal1 BC to Cal AD 300 and is 

associated with “Eburran Phase 5” stone tool flaking technologies (Ambrose, 1985) and, in its initial phases at least, 

with a purely hunting and gathering subsistence economy. 

Evidence from several excavated rock shelter and cave 

sites in the Central Rift Valley, most notably the site of Enkapune ya Muto, indicates that these LSA hunter-gatherer 

groups were fairly mobile populations who exploited a 

range of medium- to large-sized animals and participated 

in wide-ranging exchange networks that gave them access 

to different lithic raw materials, such as obsidian and chert 

(Ambrose, 1998; Kusimba and Kusimba, 2005). The evidence also suggests that there were different adaptations 

to highland and lowland environments, with the groups 

utilizing the former tending to select caves and rock shelters situated close to the boundary between montane forest and grassland habitats, whereas the latter generally 

occupied open-air sites on the savannas.

Four rock shelter sites on the Laikipia Plateau have 

thus far been investigated in some detail, and exploratory 

investigations have been carried out at a sample of other 

similar locations. Of the four key sites, those at Shulumai and Kakwa Lelash close to Dol Dol in the Mukogodo 

Hills (see Figure 1 for map) provide long, dated sequences 

that span terminal phases of the Middle Stone Age (MSA), 

the subsequent development of the LSA, and the appearance of PN ceramics and stone tool industries. Additionally, the upper levels at Shulumai contained extensive 

faunal and other evidence concerning more recent transitions from foraging subsistence to pastoralism among 

Mukogodo (Mutundu, 1999). The earliest horizon at 

Shulumai is dated to around 45,000 Cal BC and so represents the final phases of the MSA. The MSA horizon 

is overlain by successive deposits containing microblades 

and other lithic materials considered typical of the LSA 

in East Africa, spanning a date range from ca. 40,000 to 

ca. 3,000 Cal BC toward the upper levels (Dickson and 

Gang, 2002). There is a gradual increase in the percentage of nonlocal raw materials through the LSA levels. An 

even more marked increase in the use of chert and obsidian is evident from the excavated LSA levels at the nearby 

rock shelter of Kakwa Lelash (Kuehn and Dickson, 1999; 

Gang, 2001; Dickson and Gang, 2002). This increased use 

of nonlocally available raw materials is generally regarded 

as evidence for increased patterns of mobility or, at least, 

the exploitation of larger territories. Similar trends toward 

greater mobility and more widely ranging exchange networks have been noted for the LSA at sites around Lukenya Hill in Kenya (Kusimba, 2001) and elsewhere in the 

region, although the date of this transition is quite variable 

geographically (Kusimba and Kusimba, 2005). Although 

there remain some uncertainties with regard to the dating 

of the sequences at Kakwa Lelash, it is evident that by 

20,000 Cal BC, and probably considerably earlier, LSA 
1 4 

• smithsonian contributions to zoology

FIGURE 1. Map of the Ewaso Basin showing location of main Pastoral Neolithic and Pastoral Iron Age sites mentioned in the text. Prepared by Tim Evans.
number 632 • 1 5

hunter-gatherers occupying Laikipia possessed a complex and sophisticated lithic technology, were relatively 

residentially mobile with extended territories, and were 

linked with other hunter-gathering groups in the Central Rift by long-distance exchange that provided access 

to high-quality lithic raw materials (Dickson and Gang, 

2002:19–20; Dickson et al., 2004).

Excavations at a limited number of other rock shelters on the Laikipia Plateau by Jacobs (1972a) and more 

extensive investigations at Porcupine Cave and two other 

rock shelters (KFR-A4 and A12) on Kisima Ranch by Siiriäinen (1977, 1984) suggest the first appearance of domestic stock in the Ewaso Basin was between 2000 and 

1000 Cal BC. These dates are broadly consistent with the 

current evidence from other parts of the region, which 

tends to suggest that there was a gradual southward 

spread characterized by a series of stops and starts, with 

the transition to generalized, mixed cattle (Bos taurus L.) 

and caprine (Ovis sp.) pastoralism occurring in some parts 

of the northern lowlands bordering Lake Turkana by ca. 

3800 Cal BC (Barthelme, 1985), by ca. 2000–1400 Cal BC 

in different parts of highland central, southern, and western Kenya (Robertshaw, 1989; Karega-Mu ̃ nene, 2002; 

Marean, 1992; Marshall, 2000), and by ca. 1800 Cal BC 

in the Tsavo region of southeastern Kenya (Wright, 2005, 

2007). By ca. 1000 Cal BC, PN traditions were present 

across much of eastern Africa.

Archaeological and genetic evidence indicates that 

the initial introduction of livestock into East Africa was 

due in part to the southward expansion of small groups 

of herders from Sudan, Ethiopia, and possibly Somalia (Bower, 1991; Marshall, 2000; Hoelzmann et al., 

2001; Hanotte et al., 2002). The diffusion of livestock, 

knowledge of herding practices, material culture traditions, and technological styles is also likely to have contributed to the southward spread of pastoralist systems 

of food production and their archaeological signatures. 

The historical linguistic evidence suggests that these early 

pastoralists may have been speakers of proto-Southern 

Cushitic languages (Ehret, 1998). Their southward expansion from the Sahara was probably stimulated (and 

also facilitated2

) by the onset of a mid-Holocene dry 

phase; however, this phase of hyperaridity may have constrained early PN groups to the Lake Turkana Basin until 

climatic amelioration sometime after ca. Cal AD 0. The 

uptake farther south was certainly by no means rapid, 

and many of the earliest dated occurrences of domestic 

stock come from sites and contexts more closely associated with LSA hunter-gatherers (e.g., Barthelme, 1985; 

Bower and Chadderdon, 1986; Marean, 1992; Ambrose, 

1998). The reasons for this initially slow uptake of herding may have been the relative abundance of wild animals 

and plants as alternative food sources in many areas and, 

probably, the spatial distribution of livestock diseases, 

including trypanosomiasis, foot and mouth, and malignant catarrhal fever, that were endemic to forested areas 

(Gifford-Gonzalez, 1998, 2000).

It is also important to stress that although the southward spread of pastoralism almost certainly involved some 

population migration, once herding economies had become 

established in a particular area, it is highly likely that some 

of the autochthonous hunter-gatherer communities gained 

access to livestock and may have ultimately changed their 

subsistence strategies. This change certainly seems to have 

been the case in at least some areas of the Ewaso Basin, 

given that Siiriäinen (1984) found considerable typological continuity in the stone tool assemblages at the sites he 

excavated spanning levels dated ca. 1000 Cal BC to the 

twelfth century AD and perhaps even up to the fifteenth 

century AD. He also found a wide range of different PN 

pottery types represented at the sites he investigated, with 

no single type dominating until the emergence of Kisima 

Ware during the second millennium AD. Both observations 

would be consistent with a view that the rock shelters were 

occupied by hunter-gatherer communities that gradually 

made the transition to food production, although it is possible that PN pastoralists were also utilizing rock shelters, 

as has been documented in the Central Rift. 

The limited information on the faunal assemblages of 

these sites, which contain a mix of wild and domestic taxa 

(Siiriäinen, 1984:88), nevertheless indicates that hunting 

remained an important aspect of the subsistence strategies 

of these early herding communities. Similarly, at Kakwa 

Lelash and Shulumai, the upper levels, dated to ca. Cal 

AD 900 (Gang, 2001:14, 16), contained assemblages of 

PN and historic pottery associated with worked stone and 

wild and domestic fauna. Detailed analysis of the faunal 

remains from this horizon at Shulumai indicates that the 

assemblage was dominated by remains of wild animals, 

with a preference for small species such as hyrax, dik-dik, 

and bushpig, with only small numbers of domestic cattle 

and sheep or goats (Mutundu, 1999:50–56). Although 

this pattern of resource use may have been influenced by 

the ecological conditions in the immediate vicinity of these 

rock shelters, Mutundu argued that the characteristics 

of the assemblage are fairly typical of specialized huntergatherers with relatively restricted settlement mobility in 

the process of adopting food production.
1 6 • smithsonian contributions to zoology

The results from recent excavations at Ol Ngoroi rock 

shelter on Lolldaiga Hills Ranch, undertaken as part of 

a British Institute in Eastern Africa (BIEA) research project during 2002–2006, lend support to these arguments. 

Specifically, despite relatively shallow deposits, excavations revealed the presence of domestic livestock remains 

associated with wild fauna, a microlithic stone tool assemblage on obsidian, and in situ hearth deposits. A charcoal 

sample from the lowest of these hearths, of which there 

were seven in all, produced an AMS radiocarbon date 

of around 2700 Cal BC. Although sheep or goat bones 

were present, the bulk of the identifiable elements in the 

faunal assemblage is of wild fauna, and the range of the 

taxa reflects the structure of the habitats proximal to the 

site today, with hyraxes and small bovids predominating 

(Mutundu, 2005). In broad terms, the composition of the 

assemblage resembles that recovered from the upper levels 

at Shulumai rock shelter in the Mukogodo Hills, which 

is known to have been occupied by Mukogodo groups 

during a period of transition from hunting and gathering 

to a herding subsistence strategy (Mutundu, 1999). However, until open-air sites associated with diagnostic early 

PN ceramic and/or lithic assemblages are investigated, 

the processes whereby the transition to pastoralism occurred across the Ewaso Basin will only be partially understood (as discussed below, locating such sites has been 

challenging despite systematic survey). It is worth noting 

here that the Ol Ngoroi shelter, which looks out across 

the flat plateau land to the west of the Lolldaiga Hills, 

contains several well-preserved panels of white, geometric 

rock art (Figure 2). This art is similar in design to that 

recorded at Lukenya Hill south of Nairobi and is generally attributable to Maasai meat-feasting or initiation rituals (Gramly, 1975; Smith et al., 2004). Several other rock 

shelters with similar art are known elsewhere in the Ewaso 

Basin (Chamberlain, 2006), and many appear to have remained important places in the ritual landscapes of Laikipia and Leroghi. The artifactual and faunal remains from 

FIGURE 2. Example of white geometric pastoralist art, Ol Ngoroi rock shelter, Lolldaiga Ranch, Laikipia Plateau. 

Photo ©Paul Lane, BIEA.
number 632 • 1 7

the upper levels at Ol Ngoroi are consistent with the later 

use of the site by pastoralist communities and have yielded 

a date of ca. Cal AD 1260.

CONSOLIDATION OF PASTORALISM

Recent judgmental and systematic transect surveys 

undertaken as part of the BIEA research project between 

2002 and 2006 on Lolldaiga Hills, Mugie, and Borana 

Ranches, supplemented by information collected during 

rock art surveys on Mpala, Jessels, and Chololo Ranches, 

have located more than 250 previously unrecorded sites 

(Lane, In press). More than 50% of these sites are stone 

cairns or complexes of stone cairns (Figure 3), and another 

5% are rock art sites. Of the remaining sites, roughly six 

times as many (29% compared to 5%) probably date to 

the PIA compared to those associated with the earlier PN. 

Although this probably reflects a steady increase in human 

activity and presence on Laikipia during the second millennium AD, it is also possible that a proportion of the older 

sites have yet to be detected because they lie buried beneath 

colluvial and alluvial sediment that has accumulated along 

valley floors over the millennia. Scatters of PN Elmenteitan 

and SPN (Savanna Pastoral Neolithic) pottery associated 

with flaked obsidian and other worked lithic material have 

recently been reported from Lolldaiga (Causey, 2008:316, 

2010) and are also known to occur on Kisima and Mpala 

Ranches and in the vicinity of Baawa on Leroghi Plateau. 

Since these have yet to be dated or intensively investigated, 

little can be said at present about the later phases of the 

PN. Far more evidence is available concerning pastoralist 

activities in the Ewaso Basin during the PIA.

Current evidence suggests that the adoption of iron 

manufacturing technologies among pastoralist communities of the Central Rift occurred around Cal AD 800–900 

(Collett and Robertshaw, 1983). Archeologists term this 

phase the Pastoral Iron Age. One of the earliest dated 

occurrences of this transition comes from the Deloraine 

“main site,” near Rongai to the northwest of Lake Nakuru (Ambrose, 1984). This site contains abundant remains 

of cattle as well as evidence of cereal processing and the 

manufacture and use of iron implements. The ceramics 

may well represent a development of earlier PN Elmenteitan traditions (Sutton, 1993a:123). The ensuing centuries 

witnessed the consolidation of PIA economies along the 

Central Rift and adjacent highlands, out of which many of 

the ethnic identities and linguistic clusters that characterize these areas today were formed.

The BIEA surveys were supplemented by excavations 

and test excavations at 12 different sites, selected to provide 

evidence from a range of sites of different ages and dates. 

Of particular importance to the discussion here of the archaeology of pastoralism in the Ewaso Basin are the results 

from excavations at the Mili Sita open settlement site in the 

Lolldaiga Hills and the Maasai Plains site on Mugie Ranch, 

aspects of which are discussed below. More detailed excavation reports on both sites will be published elsewhere.

FIGURE 3. Stone cairn on Mugie Ranch, Laikipia 

Plateau, prior to excavation. Photo ©Paul Lane, 

BIEA.
1 8 • smithsonian contributions to zoology

PASTORAL IRON AGE SETTLEMENTS

Both the Mili Sita and Maasai Plains sites are large, 

open-air pastoralist settlements that contain a pottery 

type known as Kisima Ware, which has been tentatively 

associated with the Laikipiak (Siiriäinen, 1984), although 

whether the Laikipiak made the pottery themselves or they 

obtained it through exchange with neighboring communities (possibly hunter-gatherer groups) for other products is 

not clear (see below), and it should not be assumed (rather 

than demonstrated) that a particular pottery style can be 

equated in a straightforward manner with ethnicity. Of 

the two open-air sites thus far excavated, the Maasai 

Plains site on Mugie is the older one, and on the basis of 

available radiocarbon dates it was occupied around Cal 

AD 1400–1480. The site is situated in an area of open 

grassland surrounded by woody vegetation, about 1.5 km 

to the west of Loitigon Vlei, which has been sampled for 

pollen and other environmental remains (Taylor et al., 

2005). The Maasai Plains site consists of three concentric 

and roughly circular arrangements of low ash mounds between ~0.35 and 1.0 m high, covering an area ~750 m in 

diameter, with a pair of mud wallows, or shallow waterholes, in the center (Figure 4). The mounds, although not 

FIGURE 4. Maasai Plains site plan, showing concentric arrangement of low ash mounds, Mugie Ranch, Laikipia. 

Prepared by Guy Hopkinson and Dave Pinnock from BIEA data.
number 632 • 1 9

especially distinctive on the ground, show up as darker 

patches on vertical aerial photographs and even on moderately low resolution satellite images (see below). Two of 

the mounds have been examined by excavation, as have 

some of the level areas between these mounds. The results 

of these excavations indicate that the mounds comprise a 

series of interleaved layers of ashy soil, containing pottery 

and bone fragments and occasional flaked stone artifacts. 

In terms of the composition of the lithic assemblage, modified blades and outils écaillés dominate, pointing to links 

with the earlier Elmenteitan, with the main raw materials 

being obsidian, chert, and quartz. Traces of burnt dung, 

charcoal, and other organic materials also occur. Preliminary analysis of the faunal remains indicates that the assemblage is dominated by domestic stock (almost equally 

cattle and small stock), although a few wild fauna are also 

represented (Kennedy Mutundu, Kenyatta University, personal communication 2006). Sampling for paleobotanical 

remains was also undertaken at the site, but the results 

are as yet unavailable. Phytolith evidence recovered from 

soil samples and charred seed material may help confirm 

whether crops were cultivated in the vicinity. Isotopic analysis of human remains recovered from burials at Baawa, 

~50 km to the north, and also from the single burial thus 

far recovered at Mili Sita might also indicate whether C4

-

based plants, such as millet and sorghum, formed part 

of their diet (for details of these burials, see Lane et al., 

2007; unfortunately, although three cairns at Mugie have 

been excavated, none have produced any human remains). 

Moreover, although current rainfall regimes make cultivation in the general vicinity a risky undertaking, slightly 

higher rainfall, as is documented regionally for the main 

period of the site’s occupation (see below), could have 

made farming a viable subsistence strategy.

In terms of their form, composition, stratigraphic 

structure, and contents, these mounds closely resemble 

the dung-and-trash heaps commonly observed outside 

modern Samburu homesteads (nkan’g; in Maa, enkang), 

formed by the regular cleaning of dung from animal kraals 

and the discarding of household waste (Figure 5). No 

house remains have yet been located in the areas between 

the mounds to confirm this theory, although the presence 

of low-density mixed artifact and faunal scatters in these 

areas would be consistent with the kind of debris generated around a living area (Causey, 2008:185–267). The 

regular patterning of the mounds and the obvious clear 

area at the center of the site suggest that at a minimum, 

the mounds making up at least one “ring” were created at 

the same time, and given the layering in the mounds, each 

mound was probably built up progressively. The extent of 

the site is much larger than any other recorded archaeological pastoralist site in the region, and it covers an area 

even larger than most ethnographically documented Maasai enkang. It is entirely possible, however, that other rings 

were created at different times and that not all mounds relate to separate houses. It is also possible that the site was 

only occupied for short periods at a time and was perhaps 

reoccupied on several occasions; it is even possible that 

it was used for ceremonial purposes similar to those of 

ethnographically documented Maasai manayatta, which it 

resembles in terms of its formal spatial organization.

The Mili Sita site is located on a low col toward the end 

of a northward-trending ridge in the Lolldaiga Hills, and 

like the Maasai Plains site, it appears to have been used as 

FIGURE 5. Example of a dung midden outside 

a contemporary Samburu homestead, Laikipia, 

2005, which is possibly analogous to the ash 

mounds at Maasai Plains. Photo ©Paul Lane, 

BIEA.
2 0 • smithsonian contributions to zoology

an area of pastoralist settlement (Figure 6). On either side 

of the ridge, gentle to moderate slopes run roughly east 

and west to alluvial valley floors. At the center of the col is 

a large grass-covered area (Figure 7) with several discrete 

concentrations of archaeological material, including later 

variants of Kisima Ware that are similar but not identical 

to those from the Maasai Plains site. Some scatters appear 

to represent the remains of rubbish dumps, whereas others 

have been shown by excavation to mark the sites of former dwellings or stock enclosures. As at the Maasai Plains 

site, provisional assessment of the faunal remains indicates 

a subsistence economy largely orientated to herding cattle 

and small stock, which is also confirmed by provisional 

fungal spore data from the site. Interestingly, preliminary 

analysis of phytoliths recovered from different contexts 

also indicates that domestic crops were present throughout the occupation of the site, although these never exceed ~20% of the entire phytolith assemblage (Veronica 

Muiruri, National Museums of Kenya, personal communication, 2009). Additional research on these samples and 

equivalent material from the Maasai Plains site is ongoing.

About 300 m northeast of this area, on the upper 

slopes of the ridge, there is a heavily eroded area nearly 

bare of grass cover and with only a sparse covering of low 

FIGURE 6. Mili Sita site plan, Lolldaiga Hills, Laikipia. Prepared by Michael Causey and Dave Pinnock 

from BIEA data.
number 632 • 2 1

acacia thorn trees. Scattered across this area are numerous 

distinct scatters of iron slag mixed with tuyere fragments 

and the remains of several smelting furnaces (Iles and 

Martinón-Torres, 2009). To the south of the main settlement area there are at least 55 stone cairns, one of which 

has been excavated and has been shown to have been 

used for human burial (see Lane et al., 2007:43). Close by 

are two flat stone slabs, each with parallel rows of small 

ground hollows, similar to known variants of mankala or 

bao gaming boards. There is also a line of cairns running 

at right angles to the ridge at its northern end. On the basis 

of the available radiocarbon dates, the site appears to have 

been occupied around Cal AD 1640–1730, which on the 

basis of available oral histories of the region, could suggest 

an association with the Laikipiak.

LAIKIPIAK ORIGINS

One current major uncertainty is precisely when Laikipiak entered the Ewaso ecosystem. Most scholars now 

consider Laikipiak to have been speakers of an Eastern Nilotic language similar to Maa, although Jacobs (1965:66) 

initially suggested that they may have been Galla originally, a term generally used to refer to speakers of Eastern 

Cushitic languages, such as Oromo and Borana. If Siiriäinen’s (1984) association of Kisima Ware pottery with Laikipiak is correct, then on the basis of the available dates for 

its occurrence on the Maasai Plains site, Laikipiak were 

already using the plateau by the mid-fifteenth century AD. 

However, there are various reasons to be cautious about 

making such a direct correlation between the ceramics and 

past identities. First, over the last three decades archaeological research across the globe has shown that variations in material culture forms and styles are influenced 

by a wide range of different factors, and such variations 

can signal a multitude of different meanings and identities, not just those related to ethnicity (see, e.g., Jones, 

1997; Conkey, 2006). Second, it is possible that pots were 

obtained via exchange with neighboring groups (including Bantu language speakers to the south and the various 

pockets of hunter-gatherers known to have coexisted with 

FIGURE 7. View looking south across Mili Sita, 2005. Photo ©Paul Lane, BIEA.
2 2 • smithsonian contributions to zoology

Maa-speakers on the Laikipia and Leroghi plateaus) for 

much the same reasons as have been proposed with reference to PIA Sirikwa pottery and the Okiek groups of 

central Kenya (Blackburn, 1973). Additionally, differences 

in scholarly opinion over when the first Maa speakers in 

general arrived in the rift and adjacent highlands further 

complicate the picture. Arguing from the position of historical linguistics, Sommer and Vossen (1993:25) suggest 

that the ancestors of modern Maa speakers reached the 

Rift Valley “by the end of the ninth century” AD, whereas 

Galaty (1993) has argued that these represent “early Maa 

speakers” who can be distinguished from “later Maasai,” 

who represent another expansionary phase of pastoral 

genesis in the area stretching from Lake Baringo to Lake 

Natron from which they spread in a spiral fashion (see 

also Sutton, 1993b). In this model the Laikipiak only attained dominance over the Ewaso Basin around the nineteenth century.

Maasai Purko oral traditions, on the other hand, refer 

to the Laikipiak arriving from the northeast sometime prior 

to AD 1600, encountering a group of people (the Il Tatua, 

or Tatoga) already inhabiting Laikipia. According to these 

traditions, the latter were driven away by the newcomers 

(Jacobs, 1972b:82). Some support for these Purko traditions is provided by the fact that although the language 

(Yaaku) originally spoken by Mukogodo hunter-gatherers 

(who might well be descended from the autochthonous 

LSA hunter-gatherers on Laikipia) is considered to belong 

to Eastern Cushitic (Heine, 1974–1975), it also contains 

Southern Cushitic loanwords (Ehret, 1974). These loanwords could have entered Yaaku as a result of prolonged 

interaction with a Southern Cushitic population, such as 

the Il Tatua. The evidence from rock shelter excavations 

at Ol Ngoroi, Shulumai, and possibly Porcupine Cave also 

supports a model of forager-herder interaction. However, 

if the Laikipiak did indeed enter the area around AD 1600, 

one would expect to see a noticeable change in pottery and 

other artifact styles dating to this period. No such marked 

changes are evident in the available archaeological data, 

however, and there is no evidence to support a model of 

population replacement over this time period. Moreover, 

although the Il Tatua are believed to have been speakers of 

a proto-Southern Cushitic language (which would be consistent with them being descended from the initial phase of 

pastoralists entering eastern Africa), it is now generally believed that they lived farther south in the Crater Highlands 

rather than in the Ewaso Basin. Clearly, further research is 

required before these issues might be resolved.

Regardless of the precise origins of the Laikipiak, the 

available historical sources do indicate that from ca. AD 

1750 to the early twentieth century pastoral Maa-speaking 

communities, including Laikipia and Samburu, coexisted 

on Laikipia and Leroghi with scattered groups of other 

peoples that included hunter-gatherers (notably, Dikirri 

Dorobo and Mukogodo). Furthermore, their communities probably also encompassed pockets of Maa-speaking 

Mumonyot and LeUaso, Oromo-speaking Warra Daaya, 

and possibly Rendille along the northern fringes together 

with various Bantu language speakers, including Meru 

and Kikuyu, along the southern boundaries (Kenya Land 

Commission, 1934; Herren, 1987; Cronk, 1989, 2004; 

Sobania, 1993). These ethnicities were not necessarily permanently fixed entities, and in fact, there seems to have 

been widespread ethnic shifting among different groups 

within the region that often corresponded to (or resulted 

in) changes in subsistence strategy (see, e.g., Galaty, 1982, 

1986; Schlee, 1989; Waller and Sobania, 1994:55–63; 

Cronk, 2002). Individuals, rather than entire communities, may also have shifted identity as a result of marriage 

or some other reason, as has been well documented elsewhere, including more recently in Isiolo, where individuals 

of Turkana descent subsequently formed the Ilgira section 

of the Samburu (Hjort, 1981). It is also known that certain times in the past were politically unstable and were 

characterized by increased levels of intercommunity violence and warfare, especially during the Iloikop wars of 

the 1840s–1880s, which resulted in the defeat of the Laikipiak Maasai and the penetration of the area by Purko 

Maasai (Weatherby, 1967; Waller, 1985; Galaty, 1993; 

Sobania, 1993). Maasai oral traditions place the defeat of 

the Laikipiak as having taken place during the last Iloikop 

war during the time of the Laimer age set (1866–1886) 

(Jennings, 2005:1130). However, this defeat did not result in the complete disappearance of Laikipiak, as is often 

commonly believed and was seemingly implied by Joseph 

Thomson, one of the earliest European explorers in the 

area. Thomson (1968:243) claimed that following the 

Purko defeat of Laikipiak, “not a man in the entire land” 

was left. In fact, many Laikipiak families were absorbed 

by other neighboring groups, including Turkana (Lamphear, 1993), Il Chamus around Baringo (Little, 1998), 

Samburu on Leroghi (Bilinda Straight, Western Michigan 

University, personal communication, 2006), and Meru 

and Kikuyu to the south (Waller, 1985); others temporarily became “Dorobo” or Mukogodo hunter-gatherers, 

and others still were assimilated into the Purko (Hughes, 

2005). The oral traditions of various groups to the north 

of Laikipia, including those of Rendille, Turkana, Borana, 

and Samburu, also attest to a continuing presence of Laikipiak as a distinct entity—as exemplified by numerous 
number 632 • 2 3

cattle raids—until at least the last decade of the nineteenth 

century (Sobania, 1993).

Various external events and historical processes are 

also known to have had repercussions within northern 

“Maasailand,” such as the southward expansion of the 

Boran into areas around Marsabit, the extension of Somali territory to include Wajir and sections of the Tana 

River emanating from the rise of Ethiopian imperialism 

during the nineteenth century (Fratkin, 2001), and Turkana incursions into the Samburu grasslands after ca. 

AD 1900 (Lamphear, 1993). Moreover, from the midnineteenth century there was a significant expansion of the 

caravan trade in the region, fed partly by an increase in demand in Europe and North America for ivory, encouraging greater penetration of the interior by traders, hunters, 

and porters from the Swahili coast and the Bantu-speaking 

heartlands. The large-scale removal of elephants from the 

Ewaso Basin as a consequence of this expansion could 

have also had major ecological consequences, at least over 

the short term, and potentially much longer (Håkansson, 

2004). The East African region was also heavily impacted 

by a combination of epizootics, famine, and epidemics in 

the latter part of the nineteenth century, which included 

widespread famine between 1890 and 1891, smallpox 

outbreaks during 1883–1890, and rinderpest and bovine 

pleuropneumonia in the 1880s. These events are likely 

to have had significant repercussions for local demography and economic productivity (Kjekshus, 1977), as did 

the enclosure of commons, the restructuring of the tenets 

of landownership and property, and changes in the approaches to disease control that accompanied the establishment of British colonial rule (Waller, 2004).

EVIDENCE FOR HUMAN-INDUCED 

ENVIRONMENTAL CHANGE

In common with other parts of the region, the vegetation of the Ewaso Basin has undergone considerable 

change during the last 5,000–6,000 years. The main shift 

has been from dry evergreen upland forest dominated by 

cedar (Juniperus procera) and African olive (Olea africanus), which still survives in some areas, to savanna vegetation communities of various types. Evidence for these 

changes is provided by the results from recent paleoenvironmental investigations of swamp soils and sediments 

at Loitigon vlei on Mugie Ranch in the northern part of 

Laikipia and research on sediment cores collected from 

Ewaso Narok swamp just north of Rumuruti and Marura 

Swamp along the Mutara River adjacent to Ol’Pejeta 

Ranch (Muiruri, 2008). The data from Loitigon vlei, for 

example, suggest that there was a marked decline in the 

extent of Afromontane forest in the catchment at around 

100 Cal BC and that this decline coincided with a pulse 

of biomass fires, as evidenced by an abundance of largesized fragments of charcoal in the sediments (Taylor et al., 

2005). At Marura, there is limited evidence for vegetation 

disturbance at the base of the record dated to around Cal 

200 BC. Fossil pollen at this level is overwhelmingly from 

montane rainforest taxa (such as Podocarpus, Olea, and 

Juniperus, with scattered secondary species that included 

Croton, Rapanea, and Cyathea), suggesting a densely 

wooded environment. This conclusion is further supported by the presence of fungal spores associated with 

forest conditions (such as Trichoglossum cf. hirsutum and 

Glomus sp.) at these levels. Microscopic charcoal is also 

relatively scarce, suggesting a low incidence of fires. From 

ca. Cal AD 300, there is a marked decline in montane 

forest taxa and a corresponding increased abundance of 

Poaceae. An increase in forest fires is also indicated by the 

dramatic rise in the overall amount of charcoal and by the 

size of fragments, suggesting that fires were locally abundant (Muiruri, 2008).

Paleoclimatic factors such as drought cannot be fully 

ruled out as several of the larger lakes in the region experienced low-stands around this time, e.g., Lake Tanganyika, 

ca. 200 Cal BC (Alin and Cohen, 2003); Lake Edward, 

ca. Cal AD 0 (Russell et al., 2003); and Lake Victoria, ca. 

700–500 Cal BC (Stager et al., 2003). Nevertheless, these 

changes on Laikipia seem more likely to have been related to an expansion of herding and human-induced bush 

clearance, especially as evidence from nearby Mount Kenya indicates that the period from ca. 900 Cal BC to Cal 

AD 100 was one of heavy convective rainfall, enhanced 

soil erosion, neoglacial ice advances, and forest expansion, rather than contraction, as documented on Laikipia 

(Barker et al., 2001). Further increases in burning are implied by an increased abundance of the largest-sized fraction of charcoal in the Loitigon sediments from ca. Cal 

AD 300 to 1300, and according to the pollen and d13Cbulk 

data, the Acacia bushland was replaced by fire-adapted C4 

grassland. Rather similar trends are evident in the Marura 

and Ewaso Narok cores, which document a steady replacement of Afromontane vegetation by woodland and 

bushland taxa (including Capparis, Acacia, and Grewia), 

indicative of open disturbed savannah becoming widely 

established in these catchments between ca. Cal AD 300 

and 850. There is also a marked rise in the proportions of 

the spores of the dung-colonizing fungi Cercophora after 

Cal AD 1300 in these cores. This rise is probably indicative 
2 4 • smithsonian contributions to zoology

of an increase in the overall numbers of herbivores (which 

could include domestic stock), as might be expected following a shift to more open habitats (Muiruri, 2008).

When taken together, the changes in vegetation structure, increased evidence for local as well as regional fires, 

and the fungal spore evidence suggest that the period between ca. 200 Cal BC and Cal AD 300 (coinciding with 

the later stages of the PN) witnessed a marked expansion 

of human activity on Laikipia that included widespread 

forest clearance through the use of fires aimed at improving and extending grazing. This interpretation is only 

partially consistent with the available archaeological data 

since only limited traces of PN activity have been documented thus far. This apparent gap in the archaeological 

record may be due more to the limited extent and specific 

geographical distribution of archaeological research and/

or various taphonomic processes that have either buried 

or eroded PN sites, than to a genuine absence of pastoralist activity in the Ewaso Basin at this time. However, 

it must be noted that the first few centuries AD were a 

period of considerable climatic variability across the region and especially increased unpredictability of rainfall 

(Alin and Cohen, 2003). Thus, settlements may have been 

more transient as pastoralist groups became more mobile 

and populations dispersed during the dry-season months. 

It is interesting to note, nonetheless, that the apparent date 

of occupation of the Maasai Plains site at around Cal AD 

1400 coincides with evidence from Lake Naivasha for 

a brief high-stand, possibly indicative of a period of increased precipitation (Verschuren et al., 2000; Lamb et al., 

2003), which could have facilitated a phase of settlement 

and population aggregation across the Ewaso Basin.

High-resolution environmental records from various locations throughout the region also indicate that the 

period from the late AD 1500s through the late 1700s appears to have been characterized by long intervals when 

levels of effective precipitation in eastern Africa were much 

reduced relative to the present (Taylor et al., 2000; Verschuren et al., 2000; Alin and Cohen, 2003; Robertshaw et 

al., 2004). Yet as the records from Lake Naivasha and elsewhere indicate, against this general drying trend there were 

periodic wet episodes interspersed with periods of very low 

rainfall and ensuing drought, some of which are remembered in various local oral traditions. It is also clear from 

the most recent work that there have been several massive, 

supraregional droughts within the last millennium (Verschuren et al., 2000; Verschuren, 2004; Holmgren and Öberg, 

2006). Of these, an interdecadal drought dated to ca. Cal 

AD 1760–1840 (Bessems et al., 2008), which Verschuren 

et al. (2000) describe as much more serious than any of 

the twentieth century droughts, may have been particularly 

instrumental in triggering a variety of economic and settlement changes. Among other things, this severe climate 

event led to the complete desiccation of Lakes Baringo and 

Nakuru (Verschuren et al., 2000; Kiage and Liu, 2009) and 

had profound consequences for vegetation (Lamb et al., 

2003; Bessems et al., 2008).

Nevertheless, paleoecological evidence from the same 

area also suggests that swamp formation around Lakes 

Baringo and Bogoria was both extensive and rapid during wetter climatic phases, with relatively abrupt shifts 

between C3

 dry scrubland (mixed warm-season grasses 

and Acacia) and C3 wetland (dominated by Typha) vegetation mosaics (Driese et al., 2004). Moreover, it has been 

shown that droughts and other climate-induced stresses 

were not experienced uniformly across the region (Russell et al., 2007). There is also plentiful geomorphological 

evidence to suggest that rainfall regimes have fluctuated 

widely across the Ewaso Basin over the millennia. In commenting on the results of recent surveys in the Kipsing 

and Tol River watersheds in the Mukogodo Hills area immediately north of Lolldaiga, Pearl and Dickson (2004) 

noted that archaeological site distributions in these areas 

have been heavily affected by erosion and often reburied 

beneath recent alluvia. In all, Pearl and Dickson identified five stratigraphic units in this area, which related to 

different depositional events dated to between ca. 35,000 

Cal BC and Cal AD 1550. Of particular interest is the evidence for a major episode of floodplain aggradation after 

ca. 6500 Cal BC and the subsequent formation of a paleosol near the top of the sequence under cooler and moister 

conditions than prevail today around Cal AD 600. After a 

period of relative stability, erosion resumed during the last 

400–500 years (Pearl and Dickson, 2004:571–576). 

Whether this erosion was induced by grazing pressure 

generally across Laikipia and Leroghi has yet to be established. Nevertheless, preliminary evidence from around 

the Mili Sita site in the Lolldaiga Hills suggests that there 

was a phase of severe soil erosion here before and/or concurrent with the use the site around Cal AD 1640, which 

would imply that overgrazing and/or the creation of deep 

drove ways by frequent cattle movements may have contributed to severe degradation in at least some catchments. 

Elsewhere, in contrast, pastoralist activities also had major 

beneficial effects on the transfer and local concentration of 

soil nutrients, with long-term consequences for vegetation 

patterns. An example is at the Maasai Plains site on Mugie, where concentrated pastoralist activity has resulted 

in the formation of stable, less-erodible, and fertile soils, 

with a number of possible positive consequences for local 
number 632 • 2 5

wildlife populations resulting from the formation of longlived grassland glades within the wider mosaics of Acaciadominated low-tree and shrub savanna and bushland.

More specifically, glades are a common feature of African savanna environments, and researchers from several 

disciplines have sought to understand their origin and dynamics, both from an ecological perspective and in relation to the implications for environmental management 

(e.g., Dublin et al., 1990; Reid and Ellis, 1995; Young 

et al., 1995; Augustine, 2003; Augustine and McNaughton, 2004a). This ecological research has confirmed that 

some glades represent nutrient-enriched patches related to 

abandoned pastoralist enclosure settlements (“bomas”) 

that become dominated by nutrient-rich grass species that 

are particularly palatable to wild grazing ungulates and 

also enhance local biodiversity more generally. On parts of 

Laikipia glades have been shown to be dominated by Cynodon plectostachyus and the annual forb Tribbalus terrestris (Augustine and McNaughton, 2004b:831), whereas 

on the Athi Plains in southern Kenya, Cynodon nlemfuensis is the dominant grass (Stelfox, 1986). In the Amboseli 

Basin, Kenya, longitudinal study of abandoned bomas has 

indicated that recolonization of these sites by different 

plant species tends to follow a distinct sequence, culminating in the overall dominance of Acacia tortilis (Muchiru et 

al., 2008, 2009). Glades are perpetuated in the landscape 

by browsing and grazing activities that act to suppress the 

invasion of tree and shrub seedlings, and different studies 

have shown that glades may survive for periods up to at 

least 100 years. Numerous factors, including changes in 

management regime, climate change, fire history, periodic 

fluctuations in herbivore populations, and the effects of 

large herbivores such as elephants and giraffes, are known 

to influence the ecology of glades, and it is likely that different glades have had quite different histories.

Comparison between the Mili Sita and Maasai Plains 

archaeological sites lends support to such an argument 

but also indicates that some glades may be much older 

than currently estimated. Specifically, both sites survive in 

the landscape today as distinct areas of open grassland, 

against a broader context of expanding bushland, and 

both exhibit signs of intensive use for human settlement by 

pastoralist communities. The Maasai Plains site on Mugie Ranch dates to around Cal AD 1450 and is ~200–250 

years older than the Mili Sita site. This site is surrounded 

by Acacia-dominated low-tree and shrub savanna on relatively fertile Chromic Luvisols that exhibit little evidence 

of soil erosion. Analysis of time series aerial photography 

and satellite imagery of several glades on Mugie Ranch 

indicates that changes in their spatial extent since 1950 

(the date of the earliest available aerial photograph coverage) have been variable and that bush encroachment has 

been influenced by a range of anthropogenic and natural 

processes (Causey, 2008:157–184). Surveys of a sample of 

these indicate that some but not all of the existing glades 

contain traces of archaeological material indicative of former pastoralist settlement activity, although the cultural 

and temporal attributions of this material were variable 

(Causey, 2008). Of all the glades studied, the greatest decrease in open area was at the Maasai Plains site (31.7% 

since 1950). Nevertheless, the glade has continued to persist within this landscape as a nutrient “hot spot” that 

originated directly as a consequence of pastoralist activity 

some 750 years ago.

The soils around Mili Sita on Lolldaiga Hills Ranch 

are also Chromic Luvisols, but unlike those in the vicinity of the Maasai Plains site, they are rarely more than 

40 cm in depth, usually lack a well-developed A horizon, 

and frequently have been stripped, leaving only quartz lag 

deposits. All soils across the col and down the flanks show 

signs of having been severely eroded in the past, despite 

the lack of steep rocky slopes that could have generated 

significant runoff during heavy rainstorms. Both sides of 

the ridge to the south of Mili Sita are cut by substantial 

erosion gullies that are in places well over 6 m deep. In 

light of this evidence for erosion, the persistence of a large 

expanse of grassland across the col is probably due to the 

effects of the presence of pastoralist settlement here over 

a significant period of time. Soil test pits excavated along 

transects across the col seem to confirm this, as the results 

showed that the soil here had higher organic carbon levels 

than the truncated Chromic Luvisols around the Mili Sita 

ridge (see Payton, 2005). The soil pH was also always neutral and supported by elevated base saturation measurements of 100% and strong enrichment of exchangeable 

calcium and exchangeable potassium. Both factors lend 

support to the hypothesis that the soils have been enriched 

even though their morphology and physical characteristics indicate that the col had been severely eroded prior to

settlement and the use of the area for penning livestock in 

the seventeenth century AD. Whether there were earlier 

phases of settlement on Mili Sita that might have initiated this erosion or whether it arose from climate change 

(the period from ca. Cal AD 1250 to the late 1500s, for 

instance, is believed to have been characterized across the 

region by relatively humid conditions; see Alin and Cohen, 

2003) or a combination of factors remains to be established. Nevertheless, it would still appear that although 

overgrazing might have contributed to severe degradation in the surrounding catchment, pastoralist activities 
2 6 • smithsonian contributions to zoology

also had a major effect on the transfer and local concentration of soil nutrients in the degraded soils on the Mili 

Sita ridge, with long-term consequences for vegetation 

patterns. In contrast to the situation at Mili Sita, at the 

Maasai Plains site on Mugie the contrasting environment 

of the phonolite plateau has resulted in the formation of 

more-stable, less-erodible, and more-fertile soils located in 

a landscape of low relief. Even with similar land use and 

grazing pressures to those at Mili Sita, these soils would 

probably resist soil erosion. 

Elsewhere on Lolldaiga, Causey’s research has shown 

that the majority of PIA sites located during the BIEA transect surveys fell within open grassy glades and typically on 

sloping ridge tops (Causey and Lane, 2005), indicating a 

settlement location preference similar to what has been recorded among contemporary Maa speakers (e.g., Western 

and Dunne, 1979), although some settlement activity in 

the valley floors and open plains also took place. Causey’s 

transect surveys of the lowlands toward the northern end 

of the ranch, however, found much more evidence for PN 

activity than encountered during the BIEA surveys. These 

sites are all smaller in extent than the later PIA examples 

and occur on slightly sloping ridge tops, but their distribution does not correlate closely with the distribution of 

open glades (Causey, 2008:294–311, 2010).

In summary, these provisional results suggest that the 

patterns of soil erosion and deposition and bush encroachment are highly variable in space and time. Further, and 

of greater significance to contemporary environmental narratives, these results also suggest that pastoralist activity, 

although a cause of land degradation under some circumstances, can have the potential to encourage local stabilization and a reversal of degradation processes even in those 

areas where it had acted previously as a catalyst for erosion. 

In other words, the links between grazing patterns, herd 

sizes, range quality, and soil erosion appear more complicated than is implied by many current environmental narratives, whether these see pastoralists as the root cause of 

land degradation and biodiversity loss or as the traditional 

guardians of an Edenic wilderness. Far more integrated 

cross-disciplinary research on this topic is, nevertheless, 

needed to clarify certain human–environment relationships.

SUMMARY

Archaeological, historical, and ethnographic information concerning the Ewaso Basin indicate that many different groups have occupied parts of this area at different 

times in the past and that these groups have encompassed 

both hunter-gatherer and pastoralist communities and 

speakers of various Nilotic, Cushitic, and Bantu languages, 

and possibly others as well. There is also good evidence 

to suggest that at least in the recent past, the boundaries 

between different “ethnic,” “subsistence,” and perhaps 

even linguistic groups were fairly fluid and that cultural 

intermixing as well as interaction through exchange and 

other social mechanisms was common. All of these factors 

make it likely that the archaeological traces that survive in 

the landscape were the product of multiple social groups 

at different times in the past, with multiple meanings and 

social significance. Similarly, the available data caution 

against making simplistic correlations between, for example, a particular cairn or artifact type and either the 

linguistic or ethnic affiliation of their makers. Problems 

with dating and uneven preservation further complicate 

the interpretation of this evidence.

However, on the basis of the combined results of recent work on Laikipia taken together with the broader regional evidence and with the proviso that elements of the 

sequence and specific details may change as more research 

is completed, four broad phases of pastoralist practices in 

the Ewaso Basin can be proposed for the period prior to 

the establishment of colonial rule, spanning the last ~4000 

years. The first three of these are still rather speculative 

and need to be tested by further research; the evidence for 

the last phase, however, is more comprehensive. The proposed phases are as follows:

1. An initial “moving frontier” of pastoralism ca. 2700–

1000 Cal BC.

2. The formation of a “static frontier” and emergence 

of specialized pastoralism ca. 1000 Cal BC to Cal AD 

100.

3. A shift to more mixed herding-hunting economies and 

fluid ethnic boundaries ca. Cal AD 100–800.

4. The reappearance of specialized pastoralism and the 

creation of Maa identities ca. Cal AD 800–1900.

These phases entailed different forms and levels of mobility that at times may have incorporated both population 

migration and seasonal movements and perhaps even genuine nomadism. At others times, such as between ca. 2000 

BC and AD 100, the patterns of mobility may have been 

characterized more by a system of seasonal transhumance 

between relatively clearly defined and stable territories. It 

is also evident that the boundaries between “pastoralists” 

and nonpastoralists, including both hunter-gatherers and 

agriculturalists, have at times been more fluid and negotiable than at others, and hence “identity” mobility and not 
number 632 • 2 7

just spatial mobility was at times constrained and at others 

much less restricted. Other processes and events also acted 

to stimulate or reduce mobility, all of which would have 

had rather different effects on local environments and vegetation mosaics.

Acknowledgments

The research on which this chapter is based was 

conducted as part of a wider British Institute in Eastern 

Africa project (Landscape and Environmental Change in 

Semi-arid Regions of East & Southern Africa: Developing Interdisciplinary Approaches; Lane, 2005) generously 

funded by the British Academy under the auspices of its 

BASIS New Initiatives Scheme, launched in 2002. The research was undertaken under the auspices of Kenya Research Permit MOEST 13/014 issued by the Ministry of 

Education. The following people and institutions also provided additional funding, support in kind, use of laboratory and computing facilities, and/or logistical assistance, 

all of whom I thank: Michael and Nicky Dyer and staff 

at Borana Ranch, Laikipia; Michael Causey (formerly of 

Department of Archaeology, University of Oxford); Benjamin Smith and Carolyn Thorp (Department of Archaeology and the Rock Art Research Institute, University of 

the Witwatersrand); Kennedy Mutundu (Department of 

Environmental Studies, Kenyatta University); David Taylor and his students (Department of Geography, Trinity 

College Dublin); Bill Adams and Max Graham (Department of Geography, University of Cambridge); Veronica 

Muiruri and other staff at the Department of Palynology, 

National Museums of Kenya; Robert Wells, Fumi Mizutani, and staff at Lolldaiga Hills Ltd., Laikipia; Nick Pearson (OnSite Archaeology, York); Claus Mortenson and 

staff at Mugie Ranch, Laikipia; Malilangwe Trust; Mpala 

Ranch, Laikipia; Robert Payton (School of Agriculture, 

Food and Rural Development, University of Newcastle 

upon Tyne). 

I also thank the following people for their assistance 

and input: John Sutton (St. Antony’s College, Oxford), 

Fiona Marshall (Department of Anthropology, Washington University), John Giblin and Louise Iles (Institute of 

Archaeology, University College London), Bilinda Straight 

(Department of Anthropology, Western Michigan University), Charles Hilton (Department of Anthropology, 

Grinell College), Ann Powys (Kisima Ranch, Laikipia), 

Richard Waller (Department of History, Bucknell University), Rob Marchant (Environment Department, University of York), Kathleen Ryan (University of Pennsylvania 

Museum of Archaeology and Anthropology), Philip Owiti 

and Stephen Manoa (formerly of BIEA), and all the students who have assisted on the field projects.

I thank Nick Georgiadis for his encouragement and 

patience during the writing of this paper. I also thank Benson Kimeu (BIEA), Tim Evans (Archaeology Data Service, 

University of York), Michael Causey and Dave Pinnock 

(OnSite Archaeology, York), and Guy Hopkinson (formerly OnSite Archaeology, York) for help with preparing 

several of the figures.

NOTES

1. Owing to fluctuations in the ratio of 14C to 12C in the atmosphere 

over the millennia, largely as a result of changes in the solar magnetic 

field, “raw” radiocarbon dates do not correspond directly with calendar 

years. These raw dates can be calibrated by using calibration curves produced by systematic dating of materials of known absolute date (such as 

tree rings), and comparing the differences between the actual calendar 

age of the material and its radiocarbon age. Whereas raw radiocarbon 

dates are conventionally reported in “radiocarbon years BP,” calibrated 

dates are reported in calendar years in the form Cal BC or AD.

2. Although increased aridity in parts of Ethiopia and Sudan may 

have made these areas more marginal for pastoralism, it would have had 

the reverse effect in, for example, the Central Rift Valley, where reduction in rainfall and a fall in lake levels would have encouraged the retreat 

of forest margins to higher altitudes, allowing the expansion of grasslands more suited to grazing (Ambrose and Sikes, 1991; Marean, 1992).

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Reassessing Aerial Sample Surveys 

for Wildlife Monitoring, Conservation, 

and Management

Nicholas J. Georgiadis, Nasser Olwero, 

Gordon Ojwang’, and George Aike

ABSTRACT. Concerns about the cost-effectiveness of conservation monitoring prompt 

a reassessment of systematic aerial sample surveys, which have been widely applied to 

census wildlife and livestock in African savannas for more than 30 years. First, we use 

results from high-resolution sample surveys in Laikipia District (northern Kenya) to directly compare results from aerial total and sample surveys, showing few systematic differences in their estimates of wildlife abundance but great differences in cost and scope. 

Second, we quantify how the precision of population estimates is affected by survey 

resolution and species density. Results suggest that lower survey resolutions widely used 

in the past to census wildlife and livestock resources have been insufficient to reliably 

estimate all but the most abundant species. Third, we describe how sample survey data 

can be used to map the potential for human–wildlife conflict across large landscapes. 

High-resolution sample surveys in Laikipia have revealed causes and consequences of 

ecosystem change, advanced our understanding of ungulate population dynamics, guided 

wildlife management and conservation action, and increased confidence in sample survey 

methodology. However, further refinements in sample survey methods are needed to improve cost-effectiveness.

INTRODUCTION

There is a growing need to inventory and monitor wildlife resources as well 

as threats to those resources, for several reasons. One is to track the status of 

threatened and endangered species: Are they dwindling or recovering? Another 

is to ensure that exploited species are harvested sustainably. There is also a need 

to map the changing human “footprint” across landscapes since this is typically 

the principal threat to the extent and integrity of wildlife habitat. Finally, there is 

a need to evaluate the success of measures designed to achieve specific conservation targets (Stem et al., 2005). Judging how best to allocate limited funds either 

to direct conservation actions or to monitoring the success of those actions is not 

Nicholas J. Georgiadis, Property and Environment Research Center, 2048 Analysis Drive Suite 

A, Bozeman, Montana 59718, USA. Nasser Olwero, Mpala Research Centre, Nanyuki, Kenya; 

present address: World Wildlife Fund–US, 1250 

24th Street, NW, Washington, D.C. 20037-1193, 

USA. Gordon Ojwang’, Department of Resource 

Surveys and Remote Sensing, Post Office Box 

47146, Nairobi, Kenya. George Aike, Mpala Research Centre, Nanyuki, Kenya. Correspondence:

N. Georgiadis, georgiadis.nick@gmail.com. 

Manuscript received 7 January 2009; accepted 18 

May 2010.
3 2 • smithsonian contributions to zoology

straightforward (Tear et al., 2005). Within this context of 

enhancing conservation efficiency, we reassess aspects of 

the utility and role of systematic aerial sample surveys as a 

tool for conservation monitoring.

Aerial sample surveys originated in Serengeti National 

Park in the early 1970s as an affordable way to monitor 

the rapidly growing wildebeest population (Sinclair and 

Norton-Griffiths, 1982). The method was later refined to 

estimate the abundance of all large herbivores, both wild 

and domestic, and to map their distributions across vast 

regions (Norton-Griffiths, 1978). For more than 30 years 

this monitoring method has been widely applied in African savannas to track wildlife and livestock dynamics 

(see Box 1 in the Introduction to this volume: Georgiadis, 

2011:2).

A reassessment of systematic sample surveys was 

prompted by several concerns. Increasing costs have sharpened questions about affordability and cost-effectiveness. 

Declining numbers and distributions of many wildlife 

populations call for data at ever increasing levels of spatial resolution, prompting questions about the precision of 

sample survey estimates and when precision is sufficient 

for specific conservation and management applications. 

Cutting costs by reducing survey resolution seems counterproductive, given its attendant penalties of diminished 

precision and spatial resolution. Another option is to increase resolution but limit the spatial extent of surveys to 

areas of particular concern. A compromise is to stratify 

surveys, covering different areas at different resolutions. 

A final option is to abandon aerial surveys entirely. Other 

than some initial attention to these issues (International 

Livestock Centre for Africa, 1981; Redfern et al., 2002), 

little quantitative information has been published that 

helps us choose among available options.

We address these issues empirically, using results from 

high-resolution sample surveys that have guided the conservation and management of large herbivores in Laikipia 

District (northern Kenya) for more than two decades. We 

begin with a direct comparison of results from aerial total 

and sample counts of wildlife in Laikipia District, discussing their strengths and weaknesses. We then show how 

the precision of population estimates from sample surveys 

is affected by survey resolution and species density, inferring thresholds at which density may be too low to yield 

reliable results. Finally, we describe a simple but instructive application of sample survey data, in which the spatial 

proximity of wildlife and human attributes is used to map 

the potential for human–wildlife conflict. Further refinements in sample survey methods are suggested to improve 

cost-effectiveness and conservation applications.

Comparing Results from Total

and Sample Counts, and Seasonal

Effects on Visibility Bias

Sample surveys have provided an unparalleled record 

of wildlife and livestock dynamics in Kenya since 1977 

(Georgiadis, 2011, this volume:2 [Box 1]). However, reservations about the reliability of sample surveys have persisted, resulting from several surveys that appeared to yield 

wildly inaccurate results. Such reservations prompted a 

census of wild herbivores in Laikipia District in September 

1996 by total aerial counting, a more expensive method, 

but with greater intuitive expectation among nonscientists 

to yield accurate results. A high-resolution sample survey 

covering the same area followed only five months later, 

in February 1997. The two censuses by different methods 

were sufficiently close in time for actual changes in wildlife numbers to be minor and for differences in population 

estimates to be largely due to counting method (elephants 

were omitted from the comparison as the only species 

capable of migrating in or out of the study area in the 

interim). This provided a rare opportunity to directly compare population estimates obtained from total and sample 

surveys over a relatively large area.

Our “null” expectation was that results from these 

surveys should not be identical. We expected sample surveys to yield slightly higher estimates than total counts 

for two reasons. First, searching efficiency by observers 

should be higher within the narrow (150 m) strip transects 

of a sample survey than in a total count, in which the effective strip width may be up to 1000 m wide. Second, in 

a total aerial count there is always a strip underneath the 

aircraft that is obscured, contributing further to an undercount. Consequently, more individuals should be missed 

in a total count than a sample survey.

Given that the total count took place during a wet 

phase and the sample survey took place during a dry phase, 

seasonal factors such as background foliage density and 

color may also have biased the results because of changes 

in observer visibility (Redfern et al., 2002). An opportunity to test the effect of season on visibility, without the 

confounding influence of different survey methods, was 

afforded by an identical sample survey conducted in June 

1997, just five months after the dry season sample survey in February 1997. This coincided with the wettest El 

Niño event on record. Since Laikipia District is effectively 

a closed system for wild herbivores (excepting elephants), 

differences in population estimates between February and 

June 1997 would more likely reflect a seasonal visibility 

bias than actual population changes.
number 632 • 3 3

Total Counting Method: Over three 

days in September 1996, numbers of wild ungulate species were counted over an area of about 7,000 km2

 within 

Laikipia District, using 10 aircraft, all but one provided 

and crewed by local landowners (Kenya Wildlife Service 

provided one aircraft, and the U.S. Agency for International Development (USAID) provided fuel). To cut costs, 

areas with little or no wildlife (partially cultivated land 

in southwestern Laikipia) were omitted from the survey 

zone. Counting was organized by dividing the survey zone 

into three sections, with one section counted per day. 

Within each section, daily counting blocks of 200–500 

km2 were allocated to each aircraft. Counting began at 

around 0700 local time and ended before 1030 in the 

morning, and it took place again between 1530 and 1830 

in the evening. Blocks were searched systematically, using 

transects separated by 1–2 km from a height of 75–122 

m above ground. Flight paths were tracked using Trimble

GPS, and the locations of all herds were recorded as numbered waypoints (Figure 1, top left, bottom left). Flight 

paths and waypoint maps were printed out immediately 

upon landing and handed back to crews for checking and 

annotation so that overlaps and double counts between 

adjacent blocks could be identified and corrected.

Sample Survey Method: Aerial sample 

surveys of Laikipia District were carried out using one or 

two high-wing, twin-engine Partenavia P68s, each with a 

crew of four, consisting of the pilot, a front seat observer 

(FSO), and two rear seat observers (RSO). Topographic 

sheets of scale 1:250,000 were used in preflight planning 

to define the survey area boundary and the location of parallel transects, oriented north–south, each separated by a 

fixed distance (5 km for a low-resolution survey or 2.5 km 

for a high-resolution survey). Flying at about 190 km/h 

and at a height above the ground of 122 m (using a radar 

altimeter), the pilot navigated transects oriented north–

south using GPS. Observers counted animals that fell 

within narrow strips of known width (150 m) on either 

side of the aircraft, defined by rods attached to the wing 

struts. Both wild and domestic herbivore species were 

counted, including cattle, donkeys, camels, and sheep and 

goats, the latter two treated as a single “species” because 

they cannot be distinguished from the air. Herds of 10 or 

fewer animals were counted directly. Herds exceeding 10 

animals were estimated and obliquely photographed using 

35 mm digital cameras for subsequent counting. Since the 

sampling fraction is known (about 6.25% of the area for 

transects separated by 5 km or 12.5% for 2.5 km transect 

spacing), population estimates within the entire survey 

zone can be extrapolated from densities estimated within 

strip transects. Transects were subdivided into 5 km sections using GPS, defining survey subunits of 2.5 × 5 km or 

5 × 5 km. Human dwellings were also counted, and the 

presence of cultivation was recorded, the spread of both 

indicative of human population trends and loss of wildlife 

habitat. The spatial location of all observations within a 

given subunit was assigned to the center point of the subunit (Figure 1, top right, bottom right). 

Although the sample survey covered the entirety of 

Laikipia District, the total count omitted parts with low or 

zero wildlife densities. To allow a direct comparison with 

total count results, sample survey results were adjusted 

such that the same zone was covered in both counts. 

Results: Population estimates from sample 

and total counts differed by <23% for all but two species (Figure 2a; dotted line is x = y; solid line is the least 

squares regression: ln(y) = 0.93.ln(x)+0.69; r2 = 0.94, n = 

11). Estimates for oryx and Grant’s gazelle differed by 

89% and 67%, respectively. As expected, estimates for all 

but one species (Thompson’s gazelle) were greater from 

the sample survey than from the total count, an outcome 

unlikely to be caused by chance (P = 0.01). However, this 

bias was not great, yielding a regression with a slope not 

significantly different from unity and an intercept not significantly greater than zero (P > 0.33, n = 11). 

Comparing results from dry and wet season sample 

surveys, population estimates were again in fair agreement (Figure 2b; bars denote standard errors; solid line is 

least squares regression: ln(y) = 0.96.ln(x)+0.32; r2

 = 0.97, 

n = 11), differing by <23% for all species except eland 

(31.2%) and buffalo (37.1%). Once again, there was no 

overall bias, in that neither the regression slope nor intercept differed significantly from 0 or 1, respectively (P > 

0.63, n = 11). However, this agreement could have resulted from opposing biases for individual species canceling each other out.

Discussion: Neither total nor sample surveys necessarily yield accurate population estimates. Both 

methods tend to yield undercounts, with bias varying 

among species, observers, and habitats (Caughley, 1974; 

East, 1999; Redfern et al., 2002). For example, underestimates of more than 50% are not uncommon for impala 

in woodland (their preferred habitat). In this comparison, 

undercounting bias appeared to be slightly greater in the 

total count than the sample survey, as expected, but on 

average, differences were not large. However, comparison of additional factors, such as cost and the scope of 

information provided, revealed huge differences in favor 

of sample surveys. Compared to total counts, for example, costs of sample surveys diminish by a factor related 
3 4 • smithsonian contributions to zoology

FIGURE 1. Comparison of results from a total count and a sample survey of wild herbivores in Laikipia District. Top left: Combined flight 

paths of 10 aircraft involved in the total count of herbivores in September 1996, transects spaced 1–2 km apart (low-density areas omitted). Top 

right: Flight path of the single aircraft used to conduct a high-resolution sample survey in February 2007, transects spaced 2.5 km apart (only the 

restricted zone in vicinity of Nanyuki Air Force Base in southeast was excluded). Bottom left: Relative size and distribution of herbivore herds 

resulting from the total count. Bottom right: Impression of herbivore relative abundance and distribution from the sample survey. In the latter, 

animals are featured only if they were observed within strip transects having a combined width of 300 m. The sample survey yielded a realistic 

impression of wildlife distribution at less than half the cost of the total count.
number 632 • 3 5

100

1,000

10,000

100,000

100 1,000 10,000 100,000

Jun 1997 Sample Survey Estimates

Feb 1997 

Sample Survey Estimates

100

1,000

10,000

100,000

100 1,000 10,000 100,000

Sep 1996 Total Count Estimates

Feb 1997 Sample Survey Estimates

a

b

Symbols

Plains zebra

Eland

Impala

Buffalo

Grant’s gazelle

Oryx

Giraffe

Hartebeest

Thomson’s gazelle

Waterbuck

Grevy’s zebra

FIGURE 2. Comparison of population estimates for 11 herbivore species in Laikipia District. (a) Total count 

in September 1996 (x axis) versus a high-resolution sample count in February 1997 (y axis; axes are logtransformed). (b) High-resolution sample surveys in June 1997 (x axis; wet season) versus February 1997 

(y axis; dry season). Symbols in legend apply to both graphs.
3 6 • smithsonian contributions to zoology

to transect spacing (e.g., by about 60% for a survey with 

2.5 km transect spacing, as in this case). Perhaps most importantly, whereas only wildlife species can typically be 

censused in a total count, livestock, human settlement of 

various kinds, cultivation, and many other variables can 

also be censused in a sample survey. For wildlife conservation in nonprotected landscapes, tracking human and livestock attributes is vital because they present the principal 

conservation threats (see below). 

At least for abundant herbivore species, these comparisons should enhance confidence that total counts and 

sample surveys yield population estimates that do not differ greatly or in unexpected ways, at least for abundant 

species. In many ways, aerial total counts can be viewed 

as a special kind of sample survey, with narrowly spaced 

transects and wide counting strips, but lacking any estimate of precision. Focusing on sample surveys for the remainder of this chapter, we highlight the importance of 

this omission.

Optimizing Sample Survey Resolution

Population estimates from sample surveys are often 

used to guide conservation and management actions. 

Rarely, however, is the precision of those estimates assessed as to its adequacy for a given application. Precision 

is especially important when census results are used to 

set harvesting rates, as they were in Laikipia from 1997 

to 2003, because periodic overestimates due to random 

sampling errors can lead to overharvesting, even if population estimates are unbiased (Milner-Gulland et al., 

2001). However, the tendency for sample surveys to yield 

underestimates reduces the likelihood that harvest rates 

based on sample survey results will lead to overharvesting. Precision is also important when seeking trends in 

low-density species, for which variances associated with 

population estimates are inevitably high. In this section 

we describe an empirical approach used to assess the precision of population estimates for ungulates in Laikipia 

District. Specifically, we show how precision varies with 

survey resolution (transect spacing) and animal abundance (density).

Methods: Data were used from seven highresolution sample surveys (2.5 km transect spacing) across 

the entirety of Laikipia District. The first survey was in 

1991; there were two in 1997 (discussed in the section 

above) and others in 1999, 2001, 2003, and 2004. In each 

case, a second set of population estimates was generated 

as if they were from low-resolution surveys (5 km transect 

spacing) simply by omitting alternating transects from the 

high-resolution data.

Results: A way to visualize how population 

estimates vary with survey resolution is to plot, for each 

species in each survey, the ratio of estimates made at low 

and at high resolution against estimated density at high 

resolution (Figure 3a). The overall pattern shows this ratio 

varying widely at low densities but converging on unity 

at higher densities. This is the expected pattern, given the 

inflation in sampling error for low-density species. At least 

in Laikipia, estimates made at high and low resolution can 

diverge markedly (by >40%) for species with densities less 

than about 1 individual per km2. This threshold density 

was exceeded only by the three most abundant species 

(sheep and goats, cattle, and plains zebra).

At low resolution, a density of 1 individual per km2

corresponds to a proportional standard error of, on average, about 20%, a subjective threshold commonly considered to be a working limit for management purposes 

(Figure 3b). The density associated with this level of precision at high resolution was, on average, 0.17 km−2 (Figure 

3b). Doubling the survey resolution from 5 to 2.5 km between transects therefore improved by more than fivefold 

the threshold density at which relatively stable population 

estimates could be expected. 

Discussion: This assessment of precision 

served to validate the Laikipia Wildlife Forum’s choice of 

higher-resolution sample surveys to guide the conservation 

and management of wild herbivores in Laikipia District. 

In addition to the one wild species abundant enough to 

be reliably censused at low resolution (plains zebra), eight 

additional species had mean densities exceeding the lower 

threshold afforded by high-resolution surveys. Estimates 

for impala, Grant’s and Thomson’s gazelles, buffalo, and 

eland were used to set harvesting quotas. Estimates for giraffe, hartebeest, and waterbuck revealed declining trends 

(Georgiadis et al., 2007b). Estimates for species with mean 

densities <0.17 km2

 (Grevy’s zebra, warthog, kudu, oryx, 

and gerenuk) were considered to vary too widely to guide 

conservation action reliably.

The choice of high-resolution surveys was further validated by modeling analyses based on sample survey data, 

which provided mechanistic insights into the causes of observed herbivore population dynamics and guidelines for 

harvesting. For example, a simulation model established 

that plains zebra dynamics in Laikipia District were driven 

by an interaction between annual rainfall and zebra density (Figure 4; details in Georgiadis et al., 2003). Best fit 

solutions were fit to data from 1985 to 1999 for two alternative models, one purely rainfall dependent (RD; thin 

line) and the other rainfall mediated, density dependent 

(RMDD; thick line). Both were able to reconstruct the 

population history from 1985 to 1999 reasonably well. 
number 632 • 3 7

However, when both models were projected beyond 1999 

(gray dots and lines), with census estimates and model predictions independent of each other, the RD model deviated 

progressively from the census series (thin gray line), but 

the RMDD model continued to perform quite well (thick 

gray line). The RMDD model suggests how “equilibrial” 

(density-related) and “nonequilibrial” (rainfall-related) 

factors interact in the bottom-up control of a savanna ungulate population.

Linear modeling also implicated rainfall or population 

density in the dynamics of other abundant species, and 

predators were implicated in the decline of less-abundant 

species (Georgiadis et al., 2007a, 2007b). To our knowledge, this was the first time that results of sample surveys 

have been used to suggest causality in the dynamics of an 

ungulate community.

Mapping the Potential for Human–Wildlife Conflict

Human–wildlife conflict (HWC) is not only a liability 

to people, most of whom cannot afford crop or livestock 

losses to wildlife, it is also a severe threat to conservation. 

Where they are opposed, human interests almost invariably prevail over wildlife interests over the long term. Understanding and reducing HWC is therefore a conservation 

priority in human-occupied landscapes. In this section 

we show how a snapshot of the potential for HWC can 

be mapped across a landscape using sample survey data 

and a few simple assumptions. Our first assumption was 

that conflict (or at least the potential for conflict) increases 

with geographical proximity between wildlife and one or 

0.0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1.0

1.E-03 1.E-02 1.E-01 1.E+00 1.E+01 1.E+02

Estimated Density (km-2)

SE as Proportion of Estimate 

0.0

0.5

1.0

1.5

2.0

2.5

1.E-03 1.E-02 1.E-01 1.E+00 1.E+01 1.E+02

Est. Density at High Resolution (km-2)

Ratio

a

b

Symbols

Cattle

Sheep and goats

Plains zebra

Grevy’s zebra

All other species

FIGURE 3. Variation of precision with sample survey resolution. (a) 

Ratio of population density estimates at low survey resolution (5 km 

transect spacing) to high resolution (2.5 km transect spacing), plotted against density estimates at high resolution for 21 species in Laikipia District. Data are from 7 surveys conducted between 1997 and 

2004. Dotted line is y = 1. The x axis is log-transformed. (b) Decline 

of the standard error of population estimates with increasing density 

for 21 species from 7 sample surveys at low resolution (white circles) 

and high resolution (black circles). Standard errors are expressed as 

a proportion of the density estimate. The x axis is log-transformed.

0

5,000

10,000

15,000

20,000

25,000

30,000

35,000

40,000

1984 1988 1992 1996 2000 2004 2008

Years

Number of Zebras

0

200

400

600

800

1,000

1,200

1,400

1,600

Rainfall (mm) .

FIGURE 4. Time series of Laikipia zebra population estimates (large 

dots; vertical bars are standard errors), with best fit solutions for 

two alternative models (solid lines; see details in text). Dotted line is 

mean annual rainfall as measured from five gauges scattered around 

Laikipia.
3 8 • smithsonian contributions to zoology

more of the human-related elements of conflict, namely, 

livestock, cultivation, and habitation. For example, cooccurrence of wild herbivores with cultivation in the same 

subunit raises the potential for crop raiding or signals loss 

of wildlife habitat. Likewise, co-occurrence of wildlife with 

human habitation implies conflicting land use and loss of 

habitat. Finally, co-occurrence of wild herbivores with 

livestock implies competition for forage or the presence of 

predators that could prey on livestock. This approach was 

used to map the potential for different types of HWC on 

different types of land use in Laikipia District and to record 

how patterns of potential conflict changed over time.

Methods: Presence or absence of wildlife, 

livestock, habitation, and cultivation in each survey subunit (2.5 × 5 km) from six high-resolution sample surveys conducted between 1991 and 2003 were used for 

this analysis. Within each survey, three principal land 

use types were distinguished. Pro-wildlife properties are 

large-scale, private holdings on which wildlife is actively 

favored, often as the basis for ecotourism. Group ranches

are large-scale holdings communally owned by a registered group of families, mostly practicing pastoralism. 

Transitional properties were formerly large-scale farms or 

ranches that have been subdivided into small plots (1–10 

ha), with titles sold to thousands of smallholders, only 

some of which have been occupied and cultivated when 

rainfall permits. Where not occupied by the owner, many 

transitional properties are grazed by pastoralists. Larger 

properties that either only tolerate or actively discourage 

wildlife were included in this category. For each survey, 

co-occurrences within each subunit of wildlife with other 

elements of HWC (livestock, cultivation, and settlement) 

were tallied using geographic information systems (GIS).

Results: The potential for different types of 

conflict varied markedly across the landscape and among 

different land uses in Laikipia (Figure 5a, Table 1). By definition, there was little or no potential for HWC wherever 

wildlife was rare or absent, as in southwest Laikipia, or 

where wildlife was abundant and conflicting human attributes relatively uncommon, as on the pro-wildlife properties. The potential for crop raiding (co-occurrence of 

wildlife with cultivation) was prevalent in sectors of southern Laikipia where wildlife is still present and cultivation 

is possible in wet seasons. In contrast, the potential for 

predation on or competition with livestock (co-occurrence 

of wildlife with livestock) was more prevalent in drier 

areas, especially on group ranches and transitional areas 

adjacent to pro-wildlife properties.

The potential for HWC tended to decline between 

1991 and 2003, particularly between wildlife and livestock 

on all property types (Table 1; also compare results for the 

first and last survey in Figure 5). At first impression, this 

decline may appear to signal a conservation success, but 

closer scrutiny of changing distribution patterns for each of 

the components of conflict showed the cause to be a steady 

shrinking of wildlife distribution, significantly so on transitional and group ranches (see wildlife-only column in Table 

1). Agriculture, settlement, and especially livestock tended 

to increase over time on transitional properties, but wildlife declined markedly. On pro-wildlife and group ranch 

properties a different pattern of change was evident over 

the same period, in that not only wildlife but also livestock 

tended to decline, although not significantly (Table 1).

Discussion: Data routinely collected in aerial 

sample surveys were used to derive a plausible impression 

of the potential for HWC in Laikipia District and how it 

has changed over space and time. Validation of this impression was provided by an independent study of the actual incidence of HWC in Laikipia, in which the location and type 

of incident (livestock predation, crop raiding, infrastructure 

damage, etc.) were systematically recorded in a dedicated 

database, known as the Ewaso Incident Reporting System. 

The potential for HWC was in fair agreement with the observed distribution of incidents, with livestock predation 

occurring largely on group ranches adjacent to pro-wildlife 

properties and crop-raiding incidents scattered mostly 

across transitional properties in southern and western Laikipia (compare Figure 5b and 5c). The only exception was 

a cluster of crop-raiding incidents around Rumuruti forest in southwestern Laikipia. These incidents were caused 

by elephants that inhabited the forest during the day and 

raided crops in adjacent fields at night. This pattern did not 

feature in the potential HWC map because elephants were 

not detected under the forest canopy during sample surveys.

Opposing trends in the spatial prevalence of livestock 

on transitional and pro-wildlife ranches suggest different 

factors contributed to declines in wildlife abundance on 

different land use types. Increasing livestock abundance, 

particularly of sheep and goats, probably displaced wildlife on transitional properties over the last decade (Georgiadis et al., 2007b). In contrast, the most likely cause of 

wildlife decline on pro-wildlife properties and, to a lesser 

extent, on adjacent group ranches as well turned out to 

be the restoration of predators (Georgiadis et al., 2007a).

This validated example enhances confidence that sample survey data, especially at high resolution, can provide 

a reliable impression of the potential for and distribution 

of human–wildlife conflict. A more exacting spatial analysis using actual figures rather than presence/absence data 

may provide further insights into spatial and temporal 

dynamics of potential conflict wherever wildlife share the 

landscape with humans and livestock. 
number 632 • 3 9

CONCLUSIONS AND RECOMMENDATIONS

We used data collected in Laikipia District over two 

decades to assess the utility of aerial sample surveys for 

monitoring wildlife and human attributes across large 

landscapes, understanding wildlife dynamics, and guiding 

conservation action. A direct comparison between results 

from sample surveys and total counts showed relatively 

minor differences in population estimates but great reductions in cost and a substantial increase in the scope and 

FIGURE 5. Potential for human–wildlife conflict is mapped across three principal land use types in Laikipia District (group ranches in light 

gray, private ranches in white, and transitional properties in dark gray) by the spatial co-occurrence of wildlife with agriculture, livestock and 

settlement in survey subunits (described in the text). Visual comparison of the distribution of potential HWC in (a) February 1991 and (b) February 2003 shows the apparent decline in potential for human–wildlife conflict. (c) Recorded distribution of human–wildlife conflict incidents 

between June 2003 and February 2004, showed fair correspondence with potential for conflict in (b). Source: Ewaso Incident Reporting System, 

unpublished report, 2005.
4 0 • smithsonian contributions to zoology

utility of variables that can be monitored with sample 

surveys. An assessment of precision suggested that survey resolutions widely used in the past to census wildlife 

and livestock resources in Kenya have been high enough 

to estimate reliably only the most abundant species or 

combinations of species, such as “all wildlife.” This, in 

turn, suggests that persistent reservations about the reliability of sample surveys are largely due to a tendency 

to accept population estimates at face value, without accounting for degrees of precision. Finally, we described 

how sample survey data can provide a useful impression 

of the spatial incidence of different types of human–wildlife conflict across a landscape. Overall, data from higherresolution sample surveys in Laikipia have advanced our 

understanding of ungulate population dynamics, revealed 

causes and consequences of ecosystem change, provided 

a quantitative basis for consumptive management of wild 

ungulates, and guided conservation action (Georgiadis et 

al., 2003, 2007a, 2007b). This assessment should improve 

confidence in the validity and utility of sample surveys. 

Conservation challenges in Laikipia District are 

echoed across much of Kenya’s rangelands, where, in theory at least, regular high-resolution sample surveys would 

likewise yield sufficient information for conservation and 

management purposes. However, costs of monitoring all 

of Kenya’s rangeland districts at higher resolution would 

be prohibitive, particularly with escalating fuel prices. Survey costs in Laikipia were largely born by donors and will 

continue to be for the foreseeable future. It is unlikely that 

survey costs somehow could be met from income generated by wildlife enterprises, landholders, or conservation 

authorities in this or any other region.

TABLE 1. Proportion of the total number of survey subunits in which the elements of human–wildlife conflict were observed in highresolution sample surveys of Laikipia District between 1991 and 2003, partitioned by three principal land use types.a

 Trends over time 

in the incidence of these elements of human–wildlife conflict are indicated by correlation coefficients (r) between these percentages and 

survey date (proportional data were root/arcsine-transformed for analysis; n = 6 in all cases; an asterisk (*) denotes significance at α = 

0.05; a dash denotes statistic is not applicable).

Land use Survey date, Wildlife Agriculture Livestock Settlement

type statistic (W) (A) (L) (S) W+A W+L W+S W+A+L+S

Transitional Nov 1991 42.0% 36.3% 70.2% 53.4% 7.1% 26.4% 15.0% 6.4%

Feb 1997 43.0% 39.4% 73.5% 67.6% 6.9% 27.1% 22.9% 5.5%

Jun 1997 39.1% 47.3% 73.1% 55.6% 8.3% 23.7% 11.1% 6.9%

Feb 1999 34.2% 43.7% 79.0% 65.9% 4.5% 21.3% 15.1% 4.3%

Feb 2001 34.3% 40.0% 79.9% 54.1% 7.4% 20.9% 10.4% 5.2%

Feb 2003 28.1% 41.6% 78.7% 64.4% 4.0% 18.1% 12.4% 3.1%

Correlation, r −0.856* 0.396 0.897* 0.373 −0.509 −0.863* −0.351 −0.751

Pro-wildlife Nov 1991 53.4% 3.8% 24.2% 11.8% 2.4% 18.9% 8.9% 1.6%

Feb 1997 54.8% 3.4% 21.5% 16.1% 2.6% 16.3% 13.3% 1.1%

Jun 1997 50.7% 4.8% 20.2% 10.6% 2.7% 14.8% 6.7% 2.1%

Feb 1999 45.4% 4.0% 24.1% 12.0% 1.6% 16.0% 7.2% 1.6%

Feb 2001 47.9% 4.8% 21.1% 12.4% 3.9% 13.8% 6.9% 2.3%

Feb 2003 43.8% 4.1% 20.9% 12.0% 2.0% 11.4% 8.4% 1.7%

Correlation, r −0.793 0.345 −0.547 −0.031 0.076 −0.943* −0.267 0.343

Group ranch Nov 1991 7.8% 0.0% 15.4% 5.9% 0.0% 5.5% 2.4% 0.0%

Feb 1997 7.9% 0.0% 15.3% 11.5% 0.0% 6.0% 4.2% 0.0%

Jun 1997 6.3% 0.0% 11.2% 6.6% 0.0% 3.6% 1.3% 0.0%

Feb 1999 5.0% 0.3% 14.9% 10.1% 0.0% 3.2% 2.2% 0.0%

Feb 2001 5.6% 0.5% 11.7% 4.6% 0.0% 2.0% 0.8% 0.0%

Feb 2003 3.4% 0.0% 12.4% 8.3% 0.0% 1.7% 1.2% 0.0%

Correlation, r −0.859* 0.397 −0.571 0.086 – −0.844* −0.483 –

a

 For example, on transitional properties in November 1991, wildlife was recorded in 42.0% of survey subunits, agriculture in 36.3% of subunits, livestock in 70.2%, and 

settlement in 53.4%. Both wildlife and agriculture were observed in 7.1% of subunits, wildlife and livestock in 26.4%, etc.
number 632 • 4 1

The analysis suggests several changes in monitoring 

methodology would enhance the utility of sample surveys. 

First, a more precise record of change would be generated 

by a shift from lower-resolution surveys done relatively frequently to higher-resolution surveys done less frequently. 

Second, sample survey methods should be tailored to conservation needs in nonprotected areas, where awareness 

about the distribution, activities, and impact of humanity 

is vital to conservation. In particular, tasks assigned to the 

FSO in sample surveys should be reviewed and revised. 

For example, we have found that subjective estimates of 

woody vegetation cover by the FSO are not reliable (comparing woody cover estimates for individual subunits over 

time, there was no correspondence from one survey to the 

next). Instead, the FSO could collect more refined data on 

human activities, especially the type and spatial extent of 

habitation and cultivation. Third, a further improvement 

at no added cost would be to double survey resolution by 

dividing 2.5 × 5 km subunits in half to yield a 2.5 × 2.5 

km grid, a logistical problem involving doubling the frequency of GPS readings. Fourth, electronic advances are 

available that could greatly increase the amount of imaging information collected automatically along transects, 

for example, with vertically and obliquely mounted digital 

cameras taking georeferenced images. 

Finally, an ongoing collaboration between the Department of Resource Surveys and Remote Sensing and Mpala 

Research Centre aims to add distance sampling methodology (Buckland et al., 2001) to aerial sample surveys (M. 

Kinnaird, Mpala Research Centre, personal communication, 2007). Application of this relatively recent method 

could improve both precision and accuracy of population 

estimates from sample surveys while reducing costs. Distance sampling differs from the method currently used 

primarily by quantifying, and correcting for, the decline 

in visibility of counted objects with distance from the observer (that is, the ability of the observer to detect and 

count objects). In distance sampling, therefore, accuracy 

of the population estimate is improved by applying an 

empirically derived “detection function,” which allows a 

correction for individuals missed by the observer. Ideally, 

different detection functions are applied for different habitat types (e.g., open grassland versus woodland) and even 

different species. Precision may be improved with this approach because sampling strip width is increased, allowing 

a greater fraction of the survey area to be sampled. Costs 

are reduced because fewer transects can be flown for the 

same degree of precision. To map areas of human–wildlife conflict, survey stratification may be necessary, such 

that rapidly changing areas would be covered at higher 

resolution. Initial capital investment for refining these enhancements would be high, not least because both types 

of survey would have to be run simultaneously in order 

to compare them directly. However, this cost would be 

justified by the improved quality of information gathered 

thereafter. 

Acknowledgments

Principal support for data collection and analysis was 

from USAID and the Mpala Wildlife Foundation. Additional support was provided by African Wildlife Foundation, Laikipia Wildlife Forum, and Kenya Wildlife Service. 

This paper was written while NJG was a Lone Mountain 

Fellow at the Property and Environment Research Center, 

Bozeman, Montana.

REFERENCES

Buckland, S. T., D. R. Anderson, K. P. Burnham, J. L. Laake, D. L. Borchers, and L. Thomas. 2001. Introduction to Distance Sampling: Estimating Abundance of Biological Populations. Oxford, UK: Oxford 

University Press.

Caughley, G. 1974. Bias in Aerial Survey. Journal of Wildlife Management, 38:921–933.

East, R. 1999. African antelope database 1998. IUCN/SSC Antelope 

Specialist Group, pp. 91–92. Gland, Switzerland and Cambridge, 

UK: IUCN.

Georgiadis, N. J. 2011 (this volume). “Introduction: Conserving Wildlife in Kenya’s Ewaso Landscape.” In Conserving Wildlife in African Landscapes: Kenya’s Ewaso Ecosystem. ed. N. J. Georgiadis. 

Smithsonian Contributions to Zoology, No. 632:1–10. Washington, D.C.: Smithsonian Institution Scholarly Press.

Georgiadis, N. J., M. Hack, and K. Turpin. 2003. The Influence of Rainfall on Zebra Population Dynamics: Implications for Management. 

Journal of Applied Ecology, 40:125–136.

Georgiadis, N. J., F. Ihwagi, N. Olwero, and S. S. Romañach, 2007a. Savanna Herbivore Dynamics in a Livestock-Dominated Landscape: 

II. Ecological, Conservation and Management Implications of Predator Restoration. Biological Conservation, 137:473–483.

Georgiadis, N. J., N. Olwero, G. Ojwang’, and S. S. Romañach. 2007b. 

Savanna herbivore dynamics in a livestock-dominated landscape: 

I. Dependence on land use, rainfall, density, and time. Biological 

Conservation, 137:461–472.

International Livestock Centre for Africa. 1981. Low-Level Aerial Survey Techniques. ILCA Monograph 4. Addis Ababa: ILCA. 

Milner-Gulland, E. J., K. Shea, H. Possingham, T. Coulson, and C. Wilcox. 2001. Competing Harvesting Strategies in a Simulated Population under Uncertainty. Animal Conservation, 4:157–167.

Norton-Griffiths, M. 1978. Counting Animals. Nairobi: African Wildlife 

Foundation.

Redfern, J. V., P. C. Viljoen, J. M. Kruger, and W. M. Getz. 2002. Biases 

in Estimating Population Size from an Aerial Census: A Case Study 

in the Kruger National Park, South Africa. South African Journal 

of Science, 98:455–461.
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Sinclair, A. R. E., and M. Norton-Griffiths, eds. 1979. Serengeti: Dynamics of an Ecosystem. Chicago: University of Chicago Press.

Stem, C., R. Margoluis, N. Salafsky, and M. Brown. 2005. Monitoring and Evaluation in Conservation: A Review of Trends and Approaches. Conservation Biology, 19:295–309.

Tear, T. H., P. Kareiva, P. L. Angermeier, P. Comer, B. Czech, R. Kautz, 

L. Landon, D. Mehlman, K. Murphy, M. Ruckelshaus, J. M. Scott, 

and G. Wilhere. 2005. How Much Is Enough? The Recurrent Problem of Setting Measurable Objectives in Conservation. BioScience, 

55:835–849.
Ecological Importance of Large Herbivores 

in the Ewaso Ecosystem

Robert M. Pringle, Todd M. Palmer, 

Jacob R. Goheen, Douglas J. McCauley, 

and Felicia Keesing

Abstract. Large ungulate herbivores are widely considered to be functionally important components of African savannas, but this belief is based largely on what is known 

about their direct interactions (as consumers of plants and food for predators). Their 

indirect interactions, both as key consumers within food webs and as “engineers” of 

habitat structure, are less well understood. Here we review the results of nearly 15 years 

of experimental research at Mpala Research Centre (Laikipia, Kenya), focusing primarily 

on the indirect effects of ungulates on the abundance, distribution, and diversity of other 

species. We highlight several broad conclusions: (1) ungulate exclusion increases densities 

of plants and many small consumers; (2) the strength of these indirect effects is context 

dependent and may vary systematically along environmental gradients; (3) impacts of 

different ungulate species vary both quantitatively and qualitatively. Although some of 

the observed effects of ungulate exclusion were predictable, others were surprising. These 

insights are obliquely relevant to management, but there remains a mismatch between the 

kinds of information most interesting to academics on the one hand and most useful to 

land mangers on the other. We conclude by identifying some areas of likely convergence 

between questions of basic and applied importance.

INTRODUCTION

Much early research on African savanna ecosystems focused on the natural 

history and population dynamics of particular wildlife species. Pioneering work 

on conspicuous species such as elephants (Laws et al., 1975), buffalo (Sinclair, 

1977), lions (Schaller, 1972), and hyenas (Kruuk, 1972) helped lay the foundation for studies that illuminated relationships between herbivores and plants and 

between predators and prey (e.g., McNaughton, 1976, 1978; Hilborn and Sinclair, 1979). Recently, synthetic works have drawn on this knowledge to paint 

more complete pictures of entire ecosystems (Scholes and Walker, 1993; Sinclair 

and Arcese, 1995; du Toit et al., 2003). 

Robert M. Pringle, Department of Ecology and 

Evolutionary Biology, Princeton, New Jersey 

08544, USA, and Society of Fellows, Harvard 

University, Cambridge, Massachusetts 02138, 

USA. Todd M. Palmer, Department of Biology, 

University of Florida, Gainesville, Florida 32611, 

USA. Jacob R. Goheen, Departments of Zoology 

& Physiology and Botany, University of Wyoming, Laramie, Wyoming 82071, USA. Douglas 

J. McCauley, Department of Biology, Stanford 

University, Stanford, California 94305, USA. Felicia Keesing, Department of Biology, Bard College, Annandale-On-Hudson, New York 12504, 

USA. Correspondence: R. Pringle, rpringle@fas

.harvard.edu.

Manuscript received 7 January 2009; accepted 18 

May 2010.
4 4 • smithsonian contributions to zoology

As a result, we now know a great deal about the biology of large mammals in African savannas: what they 

eat, how they eat it, what eats and what parasitizes them, 

and, increasingly, how their activities affect vegetation 

communities and other key landscape attributes. Most of 

these processes fall into the category of what ecologists 

call direct interactions—immediate physical contact and/

or exchanges of energy between species. But such direct 

exchanges represent only a fraction of the total number 

of interspecific interactions in an ecosystem since each direct interaction creates the potential for numerous indirect 

interactions. Indirect effects occur when a direct interaction between two species influences a third species in some 

meaningful way (Abrams et al., 1996). One well-known 

example from African ecosystems involves the Serengeti 

ungulate migrations, where grazing by wildebeest (Connochaetes taurinus) prevents grass senescence, enhances 

productivity, and therefore determines the subsequent pattern of grazing by Thomson’s gazelles (Gazella thomsonii) 

(McNaughton, 1976).

Typically, however, indirect interactions are difficult 

or impossible to observe in real time. To take just one 

example, detecting system-wide indirect effects of lions 

on grasses via their suppressive effects on grazers might 

require years of observation and experimentation. Partly 

for this reason, indirect effects remain poorly understood 

relative to direct ones. Nevertheless, ecological studies 

over the past 40 years have shown repeatedly that indirect 

interactions are powerful determinants of the abundance, 

distribution, and diversity of organisms in many types of 

ecosystems and that they are frequently even more important in these respects than direct interactions (Paine, 1966, 

1980; Wootton, 1994; Menge, 1995; Abrams et al., 1996; 

Schmitz et al., 2000; Ripple and Beschta, 2004).

Thus, we cannot understand any ecosystem or predict 

the likely consequences of perturbing it without characterizing the most important interaction chains within it. 

Indirect interactions help define the emergent patterns and 

properties of all ecosystems, including how many species 

it supports, how many individuals of each species, where 

these individuals occur, and how they behave. Indirect 

interactions can also define characteristics of savannas at 

very broad scales, such as the relative balance of trees and 

grasses and the amount of carbon stored in the ecosystem (Holdo et al., 2009). Even seemingly straightforward 

direct interactions, such as the impact of ungulates on 

plants, are often mediated by cryptic indirect effects (as 

we will describe below) (Goheen et al., 2010). 

Studying the indirect effects of large mammals in African rangelands is appealing for several reasons, some 

esoteric and some more pragmatic: it allows us to test 

general hypotheses about ecological organization; it may 

enable us to predict the consequences of wildlife declines 

and extinctions (or, conversely, wildlife translocations and 

reintroductions); and at least in theory, it can provide insights about how best to manage landscapes simultaneously for wildlife conservation and livestock production.

Many advances in our understanding of the functional 

importance of large mammals in savanna landscapes have 

emerged from the past 15 years of research at the Mpala 

Research Centre and Conservancy in Laikipia, Kenya (an 

area under constant but relatively low-intensity human 

use). Much of this work has utilized a series of large-scale 

(0.5–4 ha) experimental plots that employ electric fences 

of different designs to selectively exclude different kinds 

of large mammals, allowing investigators to assess the responses of plant and small-animal communities. There are 

two ways to view this research. One is as an effort in experimental “community disassembly” for the purpose of basic 

understanding. The other is as a simulation of extinction: 

if the declining trends in many wildlife populations continue, what are the likely consequences for biodiversity and 

ecosystem functioning? In the remainder of this chapter, 

we summarize salient findings from this body of work and 

discuss ways in which future research might speak more 

directly to conservationists and land managers. 

METHODS

Mpala Research Centre supports several large-mammal exclusion experiments (“exclosures”). The oldest 

is the Kenya Long-term Exclosure Experiment (KLEE), 

located in the southwestern corner of Mpala Ranch on 

black-cotton vertisol (“clay”) soils (Young et al., 1998). 

This experiment was established in 1995 and uses electric wires at different heights along with visual barriers 

to selectively exclude cattle, “megaherbivores” (elephants 

and giraffes), and all other large ungulates (>15 kg) from 

three replicated sets of 4 ha plots, allowing researchers to 

isolate the impacts of each of these groups. The high clay 

content of these recent volcanic soils impedes infiltration 

and drainage and causes pronounced shrinking and swelling with changes in water content. At least partly for these 

reasons, the clay soils of Mpala are highly productive but 

low in diversity and evenness of almost every group of 

organisms except large mammals: a single tree (the whistling thorn, Acacia drepanolobium) makes up ~97% of 

the overstory (Young et al., 1997), five dominant grasses 

constitute ~90% of the understory (Riginos and Young, 
number 632 • 4 5

2007), the pouched mouse (Saccostomus mearnsi) represents 75% of all rodent captures (Keesing, 1998), etc. 

Another experiment, comprising only two treatments 

(total exclusion plots surrounded by 3 m tall electric fences 

along with unfenced “control” plots), was established by 

David Augustine in 1999 (Augustine and McNaughton, 

2004). This experiment is located in red sand soil and includes three replicated sets of exclusion (70 × 70 m) and 

control plots in ordinary bushland and another three in 

nutrient-rich treeless glades (abandoned cattle enclosures 

called “bomas”; Young et al., 1995). The sand soils are 

older and coarser and contain fewer nutrients than the 

clay soils; they are less productive but support more heterogeneous and diverse biotic communities than the clay 

soils. Further details on the design of these experiments 

and the sampling strategies of studies conducted within 

them can be found in the original sources cited throughout 

this chapter.

RESULTS

Influence of Ungulates on Population

Densities of Other Species

Perhaps the most straightforward prediction from 

these experiments was that standing plant biomass would 

increase in the absence of ungulates. In general, this prediction has been supported, although the net effects of 

ungulates on plants are not straightforward (Figure 1). 

In sand soil bushland, woody plants responded rapidly 

and dramatically to ungulate exclusion: after three years, 

Acacia leaf density and biomass were lower by 60% and 

80%, respectively, in unfenced plots relative to ungulate 

exclosures (Augustine and McNaughton, 2004). Likewise, 

live aboveground grass and herb biomass was greater in 

ungulate exclusion than control plots in both bushland 

and glade sites, although these effects were contingent on 

rainfall: in a low-rainfall year grazers reduced aboveground productivity regardless of soil nutrient availability, 

whereas in a wetter year grazers reduced productivity in 

bushland but increased productivity in nutrient-rich glades 

(Augustine, 2002). 

Plant responses to ungulate exclusion in black-cotton 

clay soils have been more muted. After a decade, plots 

without wildlife exhibited no substantial increases in the 

density, cover, or biomass of the whistling thorn tree (Acacia drepanolobium) (Okello, 2007; Pringle et al., 2007; 

Ogada et al., 2008; Augustine et al., 2011 [this volume]). 

Changes in grass and forb cover materialized slowly, 

with no appreciable difference between the experimental 

treatments during the first two years of the experiment 

(Keesing, 1998). By the sixth year of the experiment, grass 

density was 28% greater in cattle exclusion plots than in 

plots with cattle, while forb density was 8% greater in 

plots from which only megaherbivores had been excluded 

(this latter change is attributable to elephants since giraffes 

do not eat forbs) (Young et al., 2005). 

There are several potential explanations for this difference in plant community response between the two soil 

types, and these factors might be operating singly or in 

conjunction with one another. The greater productivity in 

clay soils relative to sands might contribute to this result 

because, with all else equal, plants in high-productivity environments can compensate more rapidly for each unit of 

production lost to herbivores (Chase et al., 2000; Pringle 

et al., 2007). Differential ungulate activity levels across 

soil types would also influence the relative magnitude of 

effects in different sites. Moreover, some of the dominant 

plants in clay soils are exceptionally well defended, most 

notably the whistling thorn, which is protected by both 

large thorns and symbiotic ants. Recent research by Goheen and Palmer (2010) shows that ants effectively defend 

trees against elephants, which reduces the impact of elephants on tree density at the landscape scale. Moreover, 

the removal of large herbivores precipitates a shift in the 

structure of this ant community; specifically, an ant species 

that weakly defends its host trees gains dominance over a 

more aggressively defending species (Palmer et al., 2008). 

Because trees occupied by the poorly defending species 

are more likely to die, the shift toward this species in the 

absence of ungulates may, counterintuitively, increase tree 

mortality. Finally, the removal of large mammals causes 

a dramatic increase in the density of rodents (see below). 

Rodents are efficient predators upon whistling thorn seeds 

and seedlings (Figure 1), which means that reduction in 

browsing of Acacia by large mammals (which largely affects resilient mature trees) is offset by increased mortality 

of juvenile tree stages due to mice (Goheen et al., 2010). 

Thus, although more seeds are produced by trees in the 

absence of large browsing mammals (Goheen et al., 2007; 

Young and Augustine, 2007), fewer are able to establish.

A second set of predictions involved the responses of 

small consumers. Since many small mammals consume 

vegetation, Keesing (1998, 2000) tested the hypothesis 

that rodent populations would be released from competition and increase in plots without large herbivores. Years 

of trapping within the clay soil KLEE plots have shown a 

sustained doubling of the dominant rodent species in the 

black-cotton community, the pouched mouse, S. mearnsi

(Keesing, 1998, 2000; Goheen et al., 2004; McCauley et 
4 6 • smithsonian contributions to zoology

al., 2006). This increase occurred before any detectable 

difference in the amount of grass cover (Keesing, 1998), 

suggesting two things. First, it is unlikely that mice became 

more abundant simply because they were better concealed 

from predators, a conclusion further supported by the fact 

that mouse survivorship rates were similar in areas with 

and without large herbivores (Keesing, 2000). Second, as 

discussed above, mice might be consuming plant biomass 

in ungulate removal plots before it can accumulate, providing one explanation for the relatively mild shifts in vegetation cover documented in KLEE. Consistent with this 

interpretation, Keesing (2000) installed small rodent exclosures within the larger ungulate exclosures and found 

that plant biomass in these smaller ungulate- and rodentfree plots increased by 50%.

Predators and parasites of mice have also responded to 

ungulate removal. Nineteen months of sampling in KLEE’s 

total-exclusion and control plots revealed that rodent-eating sand snakes (Psammophis mossambicus) were roughly 

twice as abundant within the exclosures (McCauley et al., 

2006). The number of fleas (Siphonaptera) that parasitize 

mice increased as a linear function of mouse density, also 

FIGURE 1. Direct and indirect effects of wild and domestic ungulates on Acacia establishment. Browsers such as giraffes and gazelles (pictured) 

directly suppress seed production. However, this negative effect is cancelled by the positive indirect effects of browsers’ suppressing rodents, 

which depredate seedlings. Cattle, in contrast, suppress rodents and grasses, leading to positive net effects on tree establishment. Reprinted, with 

permission, from Goheen et al., 2010.
number 632 • 4 7

leading to a near doubling, on average, of flea density in 

ungulate exclusion plots (McCauley et al., 2008). 

In contrast to the consistent results for mice and 

their predators and parasites, studies of free-living insect 

populations have yielded variable results that defy easy 

explanation. Ogada et al. (2008) sampled arthropods 

within the clay soil KLEE plots and found that cattle suppressed grass-dwelling arthropods (captured in sweep nets) 

whereas megaherbivores (elephants and giraffes only) suppressed ground-dwelling arthropods (captured in pitfall 

traps). Goheen et al. (2004) used both pitfall traps and 

sweep nets and found that true bugs (Hemiptera) increased 

more than twofold in total-exclusion plots but that beetles 

(Coleoptera) and grasshoppers (Orthoptera) did not respond detectably. Several years later, Pringle et al. (2007) 

used the same sampling methods in both KLEE and the 

sand soil exclusion plots and found that only beetles increased significantly following ungulate removal, again 

by nearly twofold on average, but with much stronger responses in sand soils than clay soils (Figure 2). Thus, insect 

responses, when detected, tend to be in the same direction 

(greater in the absence of ungulates), but the responses are 

not always detected and vary spatially, temporally, and 

taxonomically. The lack of clarity regarding the impacts 

of large herbivores on insect populations may stem in part 

from inadequate taxonomic resolution in these studies (in 

which insects were identified only to order) and/or from insufficient sampling duration or repetition. It is also possible 

that arthropod populations respond much more strongly 

to abiotic climatic variables such as precipitation than to 

shifts in the biotic variables influenced by large mammals. 

Among the arboreal fauna, Pringle et al. (2007) found 

that densities of the Kenya dwarf gecko (Lygodactylus keniensis) increased in the absence of ungulates, by about 60% 

on average, because of increases in the availability of both 

prey (insects) and habitat (trees). As was the case with the 

beetles and trees studied by Pringle et al. (2007), the response of geckos to ungulate exclusion was more dramatic 

in sand soil than clay soil habitats (Figure 2). One potential 

explanation, as mentioned above, is that this variability in 

effect size was related to productivity, with stronger effects 

in lower-productivity areas (i.e., where rainfall is low or soils 

are poor). Although this hypothesis is consistent with evidence that grassland plants are more sensitive to herbivory 

in low-productivity areas (Chase et al., 2000), the existence 

of multiple confounding factors across the two soil types 

means that further testing is required. A novel ungulate 

exclusion experiment (UHURU, see below), which is replicated across a rainfall gradient, is now underway and will 

shed light on the context dependence of ungulate effects.

Intriguingly, the negative indirect effects of ungulates 

on geckos in general concealed a positive effect of elephants 

in particular. Elephant browsing alters tree structure, often 

creating splintered crevices, which geckos use for shelter 

(Pringle, 2008). As a result, geckos were most abundant 

where there were no ungulates whatsoever, but where ungulates were present, geckos were more numerous in areas with elephant-damaged trees. This finding underscores 

the importance of physical, “ecosystem-engineering” effects on habitat structure and illustrates two additional 

points: not all ungulates are equivalent in the effects that 

they exert on other species, and some species may simultaneously exert positive and negative indirect effects on 

other species via distinct pathways. Thus, in making difficult management decisions about individual wildlife species—for example, whether to reduce or augment elephant 

FIGURE 2. Spatial variation in the strength of ungulate effects across 

six pairs ungulate exclusion and control plots. The strength of the 

experimental effect of ungulate exclusion on the densities of trees 

(diamonds), beetles (squares), and lizards (circles) is negatively related to the normalized difference vegetation index (NDVI), a commonly used proxy for primary productivity. Effect size is calculated 

as the natural logarithm of the ratio of each variable in the exclosure 

versus the control plot at each site. White background indicates three 

relatively low-productivity sites on red sand soils, and gray indicates 

three relatively high-productivity sites on black clay soils. Redrawn 

from data in Pringle et al. (2007).
4 8 • smithsonian contributions to zoology

populations—comparisons of treatments that exclude all 

ungulates versus open-access controls may not be particularly informative, since they do not allow experimenters to 

discern the impacts of individual species.

Effects on Biodiversity

Many fewer studies have addressed how ungulate removal influences community composition and biodiversity. Theoretical arguments suggest that forces promoting 

disturbance and habitat heterogeneity, such as large herbivores, should enhance diversity (Ritchie and Olff, 1999), 

but there are currently insufficient data to allow a compelling test of this proposition in the Ewaso ecosystem.

The few studies that have compared community composition across different ungulate exclusion treatments 

have revealed complex and context-dependent effects. 

Huntzinger (2005) studied grasshoppers. In sand soil 

glade habitats, grasshopper biomass increased dramatically following the exclusion of ungulates from red soil 

glade habitat. In the clay soil KLEE plots, total grasshopper biomass did not differ between plots with and without cattle, but the number of grasshopper individuals was 

actually slightly higher where cattle were present. This 

seemingly counterintuitive result arose from differential 

responses within the grasshopper community: two smallbodied species increased in abundance while a mediumsized species decreased. Hence, cattle facilitate some 

grasshopper species while suppressing others, which may 

result from structural impacts of cattle on the architecture 

of the herbaceous layer. 

Invertebrate predators also exhibited subtle responses 

to changes in the large-mammal community. Warui et al. 

(2005) sampled spiders in the KLEE plots and found that 

native ungulates did not influence total spider diversity. 

Cattle, however, decreased the diversity of the spider community as well as the abundance of the three most common species. One species of jumping spider (Salticidae), 

in contrast, increased in abundance where cattle were 

present. Because cattle decreased herbaceous cover and 

because spider diversity was positively correlated with 

relative vegetation cover, Warui et al. (2005) proposed 

that cattle indirectly affect spiders via their effects on vegetation structure. It is also possible that spiders suffered 

from reductions in their insect prey (Goheen et al., 2004; 

Pringle et al., 2007; Ogada et al., 2008). However, these 

studies of grasshoppers and spiders seem to suggest that 

for many arthropod communities the structural attributes 

of the grass layer may be at least as important as food 

availability.

Finally, Ogada et al. (2008) showed that elephants 

and giraffes significantly reduced total bird activity and 

the diversity of insectivorous birds in the 4 ha KLEE 

plots, whereas native ungulates in general depressed the 

diversity of granivorous birds. Again, multiple causality is 

likely: removal of elephants and giraffes increased the canopy area of subdominant tree species (i.e., species other 

than A. drepanolobium), providing a greater number of 

perches and less exposure to predators, whereas enhanced 

seed production in ungulate-free plots may have attracted 

granivores.

Implications for Managed Landscapes

Several studies have revealed effects of ungulate exclusion that may be relevant for the maintenance of productive rangelands. For example, Sankaran and Augustine 

(2004) showed that ungulates indirectly suppressed the 

abundance of microbial decomposers in red soil glade and 

bushland habitats by decreasing the input of plant matter 

to the soils. Because these decomposers play a pivotal role 

in the breakdown of organic matter such as dead plants, 

dung, and animal carcasses, reductions in ungulate biomass could reduce rates of nutrient cycling, a key ecosystem function with important economic ramifications.

Another example involves behavioral interactions 

among native and domestic ungulates. Grazing wildlife 

such as zebra may sometimes compete with cattle for food 

(Odadi et al., 2007). As described above, the KLEE plots 

independently manipulate the presence of cattle, megaherbivores, and all other wildlife. Young et al. (2005) reported 

data suggesting that the strength of competition between 

cattle and zebra depended upon whether elephants were 

present. The presence of cattle caused zebra to forage elsewhere (but not vice versa). However, this effect was significantly weaker in plots accessible to elephants, perhaps 

because of changes in the composition of the herbaceous 

layer. Cattle feed readily on certain forbs, and elephants reduced forb cover by 33% in this experiment. Young et al. 

(2005) hypothesized that cattle may feed less in plots accessible to elephants because of lower forb availability, which 

may then make these plots more attractive to grazing zebra.

Evidence from other African savanna systems suggests that ungulates are an important link between the 

biosphere and the climate. In the Serengeti ecosystem of 

northern Tanzania, vaccination of cattle against the rinderpest virus in the late 1950s led to the eradication of 

the disease among the native grazers of Serengeti National 

Park (Dobson, 1995). The resulting four- to fivefold expansion of the wildebeest population between 1960 and 
number 632 • 4 9

1980 reduced fuel loads and fires, which, in turn, increased woody cover. This shift from an open grassland 

to a more heterogeneous savanna doubled the amount of 

carbon stored in trees, switching the Serengeti from a net 

source of carbon to the atmosphere to a net sink (Holdo 

et al., 2009). Thus, this chain of indirect effects involving 

large herbivores determines a key ecosystem function with 

major implications for climatic change.

DISCUSSION

Conclusions and Gaps in Basic Knowledge

The work summarized above helps to illuminate the 

ecological structure of the Ewaso ecosystem and suggests 

several broad conclusions. In general, the removal of large 

herbivores has a net positive effect on the densities and/or 

activity levels of other populations, although not all species respond. Taxa increasing in density, biomass, or local 

habitat usage following removal of large ungulates include 

woody and herbaceous plants, small mammals, lizards, 

snakes, and at least some species of birds and arthropods. The ecological interpretation of these results seems 

straightforward: the large-mammal guild represents a very 

large warm-blooded biomass, which is maintained by the 

consumption of vast quantities of plant material. The reduction or elimination of large mammals from the system 

frees that primary production for use by other consumers 

(and, by extension, the consumers of those consumers), 

either as energy or as habitat. 

However, the mechanisms underlying these net effects 

may be complex and counterintuitive, necessitating some 

caution in extrapolating the results. For example, although 

tree density is higher on average in ungulate exclusion plots 

than in control plots, ungulates actually facilitate tree seedling establishment and survival by suppressing herbivorous 

rodents and competing grasses (Goheen et al., 2004, 2010; 

Riginos, 2009). In the special case of the whistling thorn 

tree, aggressive ant defenders also limit the impacts of 

mammalian herbivory. Likewise, arboreal geckos are suppressed by the ungulate community at large but facilitated 

by structural impacts of elephant on trees (Pringle et al., 

2007; Pringle, 2008). Indeed, most of the experimental effects documented to date are likely to have multiple causes, 

although only for trees have these been systematically and 

experimentally untangled (Goheen et al., 2004, 2010; Riginos and Young, 2007; Riginos, 2009). 

An obvious conclusion, but one whose implications 

may often be overlooked, is that different large-herbivore species are not ecologically equivalent. A thousand 

kilograms of zebra does not equal a thousand kilograms 

of elephant or a thousand kilograms of cattle. Each species eats different sets of plants in different ways, and each 

therefore exerts a unique impact on the broader system. 

The distinction between predominantly grazing and predominantly browsing species is one example, but there are 

many ecologically important differences between species 

within these loosely defined guilds: body size, water requirements, and, in the case of elephants, the ability to 

uproot mature trees. The effects of altering livestock density are likely to be particularly pronounced, in part because humans shield domestic animals from many factors 

that control wildlife populations: drought, food shortage, 

and predation (Hairston et al., 1960). Thus, livestock may 

continue to exert strong ecological influence even at times 

when other large-herbivore populations are declining and 

in circumstances (e.g., drought) when the rest of the community may be least resilient to their impacts. 

Finally, the relative ecological importance of large herbivores is variable in space. Pringle et al. (2007) showed 

that the indirect effects of ungulates on the densities of 

trees, beetles, and arboreal geckos varied in strength across 

a gradient in soil type and rainfall. Because the magnitude of these effects were negatively related to productivity (Figure 2), the authors hypothesized that productivity 

may modulate the strength of herbivore impacts. Studies 

elsewhere (Larson et al., 2008; Asner et al., 2009) have 

yielded at least partial support for this notion. However, 

multiple factors confounded the comparison of effects 

across soil types. 

To better understand both the modulating influence 

of environmental context and the differential effects of 

different native ungulate species, J. R. Goheen, R. M. 

Pringle, and T. M. Palmer (unpublished) initiated a novel 

ungulate exclusion experiment in September 2008. Ungulate Herbivory Under Rainfall Uncertainty (UHURU) was 

designed to focus on the effects of three particularly abundant and influential wildlife species: dik-dik (Madoqua 

kirkii), impala (Aepyceros melampus), and elephant (Loxodonta africana). Thus, the experiment includes a treatment that excludes all mammals larger than hares (Lepus

spp.), another that allows dik-dik but excludes impala and 

all larger species, and a third that excludes elephant and 

giraffes only. A fourth treatment is unfenced and accessible to all wildlife (but not to livestock, which are kept 

out by herders). Each plot is 1 ha (100 × 100 m), and 

each type of plot is replicated three times in the northern, 

central, and southern parts of the Mpala Conservancy (all 

on red sand soils), with precipitation increasing ~80% 

from north to south. This experiment will reveal potential 
5 0 • smithsonian contributions to zoology

interactions between herbivory and precipitation regimes, 

and planned warming treatments may also reveal how important ecosystem variables like soil carbon storage will 

vary as global temperature increases.

What is most striking in reviewing these studies is how 

much we still do not know. It is odd how few studies have 

addressed the effects of ungulate exclusion on community 

diversity, especially with respect to grass and forb communities. There is a similar paucity of data about the influence 

of large herbivores on basic ecosystem functions such as 

decomposition and nutrient-cycling rates. Even the influence of ungulates on primary productivity has received 

little study, and no attempt has been made to understand 

how the impacts of herbivores on plants scale up to determine levels of carbon sequestration and storage (cf. Holdo 

et al., 2009). The role of predators is likewise poorly understood. Studies in the Yellowstone ecosystem of North 

America have shown that the reintroduction of wolves in 

the mid-1990s has profoundly affected the landscape by 

altering the behavior and physiology of elk (Ripple and Beschta, 2004; Creel et al., 2007). Similar processes may be at 

work in the Ewaso ecosystem, with the apparent recent recovery of many predator species on properties in Laikipia 

since the 1990s (Georgiadis et al., 2007a, 2007b), but no 

data exist with which to evaluate this possibility. Finally, 

there has thus far been a bias in this research toward questions that are academically interesting but dubiously useful 

from the perspective of day-to-day landscape management, 

an issue that we address in the following section.

Reconciling Basic and Applied Interests

At the most ethereal level, it can be argued that preserving any complex system in working order—from automobiles to ecosystems—requires a functional understanding 

of how that system works and what its different parts actually do. The studies that we synthesize above, which paint 

an increasingly complex picture of how ungulates shape 

the abundance, distribution, and diversity of other species 

in savannas, are certainly an effort in this direction. On a 

slightly more pragmatic level, the information generated by 

these studies should be considered when managing largeherbivore populations. Harvesting, culling, contraception, 

translocation, and reintroduction of ungulates are all common practices (van Aarde et al., 1999; Georgiadis et al., 

2003; Tambling and du Toit, 2005; Druce et al., 2006; 

Hayward et al., 2007a, 2007b; Whitman et al., 2007). 

These actions are always expensive and often contentious 

(e.g., Landeman, 1978; Pimm and van Aarde, 2001), and 

it therefore seems prudent to know how such deliberate 

manipulations are likely to affect biodiversity and ecosystem processes. Understanding the direct and indirect effects 

of elephants in particular (i.e., how ecosystem properties 

change as a function of elephant abundance) may inform 

the emotionally charged issue of elephant management. 

The indirect effects of elephants have received little formal 

study. Given the high stakes of elephant management and 

control programs, this is an area where rigorous experimental data are surprisingly few.

These issues aside, very different imperatives drive 

academic science, on the one hand, and land management, 

on the other. The reward structures in most parts of academia (including the large grants necessary to finance major research efforts) demand advances at the “forefront” 

of scientific theory (wherever that might be at the time). 

Many rangeland management issues are widely considered 

(rightly or wrongly, and with definite exceptions) to fall 

outside the narrow boundaries of modern ecology’s leading edge. This situation can lead to frustration on behalf 

of both local stakeholders (“Why are researchers spending 

so much time and money on projects with no immediate 

practical benefit?”) and researchers (“Why won’t managers and community members look at the ‘big picture’ of 

interconnected parts?”). 

From the perspective of the research community, 

there are at least two ways to help resolve this dilemma. 

One is to identify research questions that simultaneously 

meet both needs: to extend the frontiers of ecology and 

to address issues that genuinely matter to managers and 

pastoralists. (This suggestion is not to diminish the importance of research questions that are more purely basic 

or more purely applied but merely to point out that all of 

these things are compatible.) Each of the concerns most 

commonly cited by those who raise livestock or manage 

multiuse landscapes in the Ewaso ecosystem dovetails in 

some way with issues currently considered exciting by academic ecologists and ecological economists: the relationship between climate and plant (i.e., forage) production, 

the potential for competition between native and domestic herbivores, the behavior and population dynamics of 

large carnivores, the transmission and management of 

infectious diseases, the determinants of range quality, the 

market price of livestock, and strategies for generating alternative revenue streams, such as from tourism. 

We provide just a few examples relevant to the topic 

of this chapter. As discussed above, early research in 

KLEE showed that small mammals and their potentially 

disease-bearing ectoparasites doubled in the absence of 

ungulates. The logical next steps, some of which are currently under investigation by McCauley, Keesing, and 
number 632 • 5 1

colleagues, are to understand whether this shift actually 

alters the prevalence of disease in humans and domestic animals and to what extent this is true in landscapes 

where livestock have completely replaced native ungulates. The effort to address these questions links recent 

intensive efforts in disease ecology with very real problems encountered in managed landscapes. Another example, being addressed by Goheen’s research group, is 

to understand how both wild and domestic ungulates 

might be better managed to limit predation on the latter 

(or similarly, how common wild ungulates might be managed to minimize depredation of rare ones). This research 

fuses long-standing academic interest in issues such as 

apparent competition (Holt, 1977) with the goal of enabling coexistence of people and predators and/or the 

stabilization of rare and declining wildlife species such as 

hartebeest (Georgiadis et al., 2007a, 2007b). Rangeland 

quality, from the perspective of livestock production, is 

largely a product of factors that have long intrigued basic 

biologists and that are determined in part by the abundance and community composition of ungulates, such as 

the relative abundance of trees and grasses. Along similar lines, an ecological-economic quantification of how 

the relative abundance of native wildlife influences per 

hectare profitability of single-use and multiuse landscapes 

would be warmly received both by managers and by academics in the thriving hybrid discipline of ecosystem services. Finally, there is enormous interest within academia 

in projecting the biological impacts of ongoing climatic 

change. Such changes, including shifts in rainfall regime, 

will determine the quantity, quality, and spatial distribution of forage over the coming century, but they remain 

poorly understood.

None of these research efforts, however conclusive 

their results might be, will yield the desired objective of 

linking basic science to conservation without another element, which is to aggressively disseminate results and 

their implications to nonscientists while simultaneously 

being forthright about the limits of current knowledge in 

informing range management. Experimental results are 

unlikely to be applied if they remain confined to scientific journals; of equal importance, but perhaps easier to 

forget, is that nothing diminishes respect for science more 

efficiently than hypotheses passed off as certainties in an 

effort to be relevant.

Acknowledgments

We thank Truman Young (University of California) 

and David Augustine (U.S. Department of Agriculture), 

who built exclosure experiments and welcomed us to work 

in them. The KLEE exclosures were built and maintained 

by grants from the Smithsonian Institution (to A. Smith), 

the National Geographic Society (4691-91), the National 

Science Foundation (BSR-97-07477 and 03-16402), and 

the African Elephant Program of the U.S. Fish and Wildlife Service (98210-0-G563) (to T. Young). Funding for 

the research and writing of this chapter was provided 

by the following sources: NSF DEB-0709880 and OISE0852961, the William R. and Sara Hart Kimball Stanford 

Graduate Fellowship, the Heinz Environmental Scholarships, and grants from the Sherwood Family Foundation 

and the Phi Beta Kappa Northern California Association 

to R. M. Pringle; NSF DEB-9726663 and DEB-0089706 

to T. M. Palmer; the American Society of Mammalogists, 

the Natural Sciences and Engineering Research Council of 

Canada, the Smithsonian Tropical Research Institute, and 

the U.S. Environmental Protection Agency to J. R. Goheen; and NSF CAREER 0196177 to F. Keesing. Thoughtful comments from Nick Georgiadis and an anonymous 

reviewer helped to improve the manuscript considerably.

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Pathways for Positive Cattle–Wildlife 

Interactions in Semiarid Rangelands

David J. Augustine, Kari E. Veblen, 

Jacob R. Goheen, Corinna Riginos, 

and Truman P. Young

ABSTRACT. Livestock-wildlife interactions in rangelands are often viewed in terms of 

competition, but livestock and native ungulates can also benefit each other through longterm modifications of rangeland habitats. Here we synthesize research on rangelands in 

central Laikipia focusing on two types of cattle-wildlife interactions that have implications for their long-term coexistence. The first interaction occurs via redistribution of soil 

nutrients within the ecosystem, which is a consequence of the use of bomas (temporary 

corrals) to manage livestock. Our studies on two different soil types show that rotational 

boma management creates hectare-scale patches in the landscape that are enriched in soil 

and plant nutrients and persist for decades to centuries. In both of the predominant soil 

types in Laikipia, forage phosphorus content is low relative to ungulate demands during 

peak lactation. Nutrient-rich boma sites (hereafter referred to as glades) provide a key 

wet-season forage resource of nutritional sufficiency for lactation. Our studies further 

show that a wide range of native ungulates selectively use glades relative to surrounding 

nutrient-poor habitats. Impala (Aepyceros melampus) in particular show intensive use of 

glades on sandy soils and are rare in portions of the landscape lacking glades. A second 

important pathway for cattle-wildlife interaction occurs through the influence of native 

browsing ungulates on woody vegetation. Shrub and tree cover has been increasing in 

Laikipia over the past century, followed by increases in native browsers in recent decades 

on ranches where wildlife are allowed to coexist with cattle. Our exclosure experiments 

in central Laikipia indicate that native browsers suppress shrub encroachment on both 

dominant soil types. However, the strength of browser effects are three to seven times 

greater on sandy soils, where two browsers, dik-dik and elephants, are both abundant, 

compared to heavy clay soils, where elephants are the only dominant browser. In the 

clay soils, native browsers still exert a significant influence on dynamics of the dominant tree, Acacia drepanolobium, and suppress encroachment by subdominant shrub 

species. Browser effects on woody vegetation likely enhance forage production for cattle 

and maintain open habitats favored by native grazers for predator avoidance. Taken 

together, our studies indicate that boma rotation and browser control of shrub encroachment are key interaction pathways that promote cattle-wildlife coexistence in the Ewaso 

ecosystem.

David J. Augustine, USDA–Agricultural Research 

Service, Rangeland Resources Research Unit, Fort 

Collins, Colorado 80526, USA. Kari E. Veblen, 

Department of Plant Sciences, University of California, Davis, California 95616, USA. Jacob R. 

Goheen, Departments of Zoology & Physiology 

and Botany, Laramie, Wyoming, USA 82071. 

Corinna Riginos, Department of Ecology, Evolution and Behavior, Princeton University, Princeton, New Jersey 08544, USA. Truman P. Young, 

Department of Plant Sciences, University of 

California, Davis, California 95616, USA. Correspondence: D. Augustine, David.Augustine@

ars.usdalgov.

Manuscript received 7 January 2009; accepted 18 

May 2010.
5 6 • smithsonian contributions to zoology

INTRODUCTION

In East Africa, most native ungulate populations still 

occur on rangelands outside of formally protected areas, 

and long-term coexistence of livestock and native species is critical if wildlife is to be conserved in the region 

(Western, 1989; Prins et al., 2000). Negative relationships 

between wildlife and livestock on some pastoral rangelands in Kenya have been attributed to poaching and 

harassment by humans associated with livestock, rather 

than direct competition for forage (de Leeuw et al., 2001). 

However, where such direct human effects are removed 

and livestock are maintained at moderate stocking rates, 

the magnitude of competition between livestock and native herbivores is a concern for the production of livestock 

and the conservation of the native fauna. Livestockwildlife interactions in rangelands are often viewed in 

terms of competition because herbivores with overlapping 

habitat and diets have obvious potential for competition. 

However, multiple pathways of interaction are possible 

in diverse herbivore communities. Prins (2000) provided 

a comprehensive review of negative pathways, including 

direct resource competition, apparent competition, diffuse 

competition, predation, and competition through habitat 

modification. He suggested that despite considerable resource overlap between wildlife and cattle, competition 

is largely asymmetrical and diffuse, with cattle having a 

competitive effect on some wildlife species, but wildlife 

often having little or no net competitive effect on cattle 

(Prins, 2000). Several recent large-scale experiments have 

demonstrated competition between cattle and native grazers in grasslands under controlled conditions, but the magnitude of competition is often less than expected on the 

basis of diet overlap (Hobbs et al., 1996; Young et al., 

2005; Derner et al., 2006; Odadi et al., 2009). 

The relationship between livestock and cattle can 

also be positive (Arsenault and Owen-Smith, 2002). Prins 

(2000) discussed two modes of positive cattle-wildlife interactions: facilitation of forage quality or accessibility and facilitation through habitat modification. In diverse herbivore 

communities, the existence and strength of facilitation can 

determine the net effect of cattle and wildlife on one another and hence are important to long-term cattle-wildlife 

coexistence. Facilitation of forage quality or accessibility 

is usually defined as a short-term effect that bulk-feeding 

herbivores can have on forage for smaller herbivores (Bell, 

1971; Gordon, 1988; Verweij et al., 2006). Facilitation 

through habitat modification involves long-term effects of 

different herbivore species on soil nutrient distribution and 

plant community composition (Arsenault and Owen-Smith, 

2002). Our studies in the Ewaso ecosystem have addressed 

trophic interactions among soils, plants, and herbivores and 

identified two types of facilitation through habitat modification that have implications for the coexistence of livestock 

and native ungulates. The first interaction occurs through 

redistribution of soil nutrients within the landscape, which 

is a consequence of the use of bomas (temporary livestock 

corrals) to manage livestock. Boma-derived nutrient redistribution, in turn, has substantial long-term effects on the 

heterogeneity of forage nutrient content within the landscape, with potential benefits to both cattle and native 

herbivores. The second interaction is the ecosystem service 

that native browsers provide by suppressing and reversing 

shrub encroachment into grasslands, with implications for 

wildlife and livestock use of the landscape. In addition, the 

old boma sites are themselves treeless and attract increased 

wildlife use for this reason as well as for their nutrient-rich 

grasses. Here we synthesize studies examining these interaction pathways on two widely distributed soil types in the 

Ewaso ecosystem. 

In central Laikipia, the Ewaso ecosystem encompasses 

a transition in soils, elevation, and vegetation, where gently undulating plains dominated by grassland and Acacia 

drepanolobium savanna switch to a more dissected lower 

plateau on Precambrian metamorphic rocks dominated by 

Acacia mellifera, A. etbaica, and A. brevispica bushland 

with a discontinuous grass understory (Ahn and Geiger, 

1987; Taiti, 1992; Augustine, 2003b). Black cotton soils 

(Pellic Vertisols) occur on the plains in the south and west, 

whereas red sandy loams (Ferric and Chromic Luvisols) 

occur at lower elevations to the north and east (Ahn and 

Geiger, 1987). These soils differ dramatically in texture, 

with black cotton containing, on average, 50% clay and 

24% sand (Young et al., 1998) and red sands containing 

15% clay and 74% sand (Augustine, 2003a). The Mpala 

Research Centre (MRC) spans this transition zone, which 

occurs at elevations of 1740–1800 m above sea level. Although these two soils support distinctly different assemblages of plants, native herbivores, and other biota (Table 

1), we show that on both of these soil types, herbivoremediated nutrient heterogeneity and shrub control are 

important pathways through which cattle and wildlife 

benefit each other.

LIVESTOCK MANAGEMENT AND 

NUTRIENT HETEROGENEITY

Tropical rangelands are notoriously poor in mineral nutrients, such that mineral supplementation is recommended 
number 632 • 5 7

as a standard practice for tropical livestock (McDowell, 

1985). Wild herbivore biomass is also strongly affected by 

large-scale variation in geology and soil nutrient availability 

(Bell, 1982; Fritz and Duncan, 1994). Furthermore, spatial 

heterogeneity in soil and plant nutrients in African rangelands can strongly influence the distribution of wild herbivores at multiple spatial scales (McNaughton, 1988, 1990; 

Blackmore et al., 1990; Ben-Shahar and Coe, 1992; Murray, 1995; Muchiru et al., 2008). Natural sources of soil 

nutrient heterogeneity include termites (Brody et al., 2010; 

Fox-Dobbs et al., 2010) and tree canopies (Riginos et al., 

2009). Here we discuss a major anthropogenic source of 

landscape heterogeneity: livestock corrals.

Abandoned livestock corrals, or bomas, are a widespread feature of African rangelands. Bomas are temporary 

structures consisting of a dense ring of thorn scrub branches 

that contain and protect livestock overnight from theft and 

predation for periods ranging from weeks to years. A major 

consequence of boma use is the concentration of large quantities of dung and urine within a small area, where livestock 

excrete and redistribute nutrients from the surrounding savanna. Following abandonment, boma sites often support a 

nutrient-enriched plant community and potentially alter the 

spatial pattern of nutrient cycling within the ecosystem. Recently abandoned boma sites in East Africa support unique 

herbaceous plant communities with nutrient enrichment 

both in soils and grasses (Stelfox, 1986; Augustine, 2003a; 

Treydte et al., 2006; Muchiru et al., 2009). Studies in South 

Africa even suggest that plant communities on abandoned 

sites of human occupation can persist in a nutrient-enriched 

state for centuries (Blackmore et al., 1990). 

Soil and Grass Nutrients on Red Sands

Versus Black Cotton Soil

For increased nutrient heterogeneity to benefit wild 

herbivores, forage nutrients must be limiting to herbivores 

in all or a portion of the landscape during part of the year. 

Previous studies on nutrient limitations to savanna ecosystem processes and herbivore abundances have primarily 

relied upon geology or cation exchange capacity as general 

indices of soil fertility at broad spatial scales (Bell, 1982; 

Fritz and Duncan, 1994; Olff et al., 2002). We used direct 

measures of soils and grasses to examine forage nutrient 

availability on black cotton soils, red sands, and a transition zone between them at MRC in central Laikipia. On 

the basis of the previous discussion of nutrient limitations 

to ungulates in Africa, we focus here on N, P, and Ca. 

We sampled soil and grass leaf blades at 100 locations within MRC during the wet season in August 2001. 

We randomly selected 60 locations within the red sands 

soil type that underlies the majority of MRC, 20 locations within a region of transitional soils between 1740 

and 1800 m elevation, and 20 locations within the black 

cotton soil type above 1800 m elevation in the southwestern portion of MRC. At each location, we sampled 30–40 

of the youngest, fully expanded grass leaves from a 1 m2

plot dominated by any of the dominant grasses (red sands: 

Digitaria milanjiana, Cynodon dactylon, Pennisetum mezianum, Pennisetum stramineum, Enteropogon macrostachyus, Cymbopogon pospischilii, Themeda triandra, 

Sporobolus ioclados, Harpachne schimperi, or Chloris 

roxburghiana; transitional soils: Digitaria milanjiana, 

Pennisetum stramineum, Themeda triandra, Cymbopogon 

pospischilii, Pennisetum mezianum, and Cenchrus ciliaris; 

black cotton soils: Pennisetum stramineum, Themeda triandra, Pennisetum mezianum, Brachiaria lachnantha, Setaria sphacelata, and Lintonia nutans. 

Leaves were analyzed for N, P, and Ca content following McNaughton (1988). At each location, we also collected a soil core (15 cm depth). Cores were stored at 5°C 

at the MRC laboratory, and within 48 hours of collection, 

a 15 g subsample was extracted with 1 M KCl to measure NH4

+

 and NO3

–

 content, and a 2.5 g subsample was 

TABLE 1. Comparison of plant and herbivore communities on two widespread soil types in Laikipia, Kenya.

Variable Red Sands Habitat Black Cotton Habitat

Soil texture 74% sand, 15% clay 24% sand, 50% clay

Dominant grasses Digitaria milanjiana, Cynodon dactylon, Pennisetum Pennisetum stramineum, Bracharia

stramineum lachnantha, P. mezianum, Themeda triandra

Dominant shrubs and trees Acacia mellifera, A. etbaica, A. brevispica Acacia drepanolobium

Dominant native ungulates Impala, dik-dik, elephant Plains zebra, Grant’s gazelle, hartebeest, giraffe

Less abundant native ungulates Plains zebra, Grevy’s zebra, eland, waterbuck, giraffe Elephant, Grevy’s zebra, eland, buffalo, oryx
5 8 • smithsonian contributions to zoology

extracted with Mehlich-3 solution to measure extractable 

P and Ca (Mehlich, 1984). Another 15 g subsample from 

each core was dried and analyzed for total soil N, P, Ca, 

and Na following McNaughton (1988). For extractable 

soil nutrient pools, we report geometric means because of 

high variability within soil types. Post hoc, pairwise comparisons among soil types were tested using Tukey’s honestly significant difference (HSD), and all letters in figures 

indicate significant differences at the P < 0.05 level.

Black cotton soils contain greater concentrations of N 

and Ca compared to red sands (Figure 1A,C; F2,97 > 9.41, 

P < 0.0002 for one-way analysis of variance (ANOVA)), 

but both soils have similar total P concentration (Figure 

1B; F2,97 = 2.37, P = 0.10). Mean soil C concentration is 

more than twice as high in black cotton soils compared 

with red sands (mean ± 95% confidence interval = 2.44 

± 0.35% versus 1.12 ± 0.13%), and soil C:N ratios are 

also higher in black cotton compared to red sands (17.3 

± 2.4 versus 10.1 ± 0.2). Extractable soil nutrient pools 

mirrored patterns for total soil nutrients, with inorganic 

Ca significantly lower on the red sands (F2,97 = 43.9, P

< 0.0001; Figure 1F) and inorganic N only marginally 

lower on red sands (F2,97 = 2.50, P = 0.087; P > 0.05 for 

all pairwise comparisons; Figure 1D). Extractable soil P 

was similar on red sands and black cotton and lower for 

the transitional soils (F2,97 = 12.8, P < 0.0001; Figure 1E). 

Grass P content was relatively consistent (F2,97 = 1.1, P

= 0.34; Figure 1H) and low across the different soil types, 

reflecting a similar pattern for total and extractable soil P. 

In contrast to soil nutrient patterns, grass N content was 

significantly greater on red sands relative to black cotton 

(F2,97 = 5.2, P = 0.007; Figure 1G). Grass Ca content did 

not vary across soil types (F2,97 = 0.5, P = 0.59; Figure 1J), 

even though black cotton soils contain significantly more 

total and extractable Ca. 

Variation in grass nutrient content across soil types 

has clear implications for large herbivores. Most notably, wet-season forage P content is lower than estimated 

demands for lactating ruminants on both red sands and 

black cotton soils (Figure 1H). For example, Murray 

(1995) estimated minimum forage P content of 3,900 mg/

kg dry weight necessary for lactating wildebeest (Connochaetes taurinus), and recommendations for lactating 

cattle are in the range of 3,100–4,000 mg/kg (McDowell, 

1985). Despite considerable differences in parent material and properties of red sands versus black cotton soil, 

our results indicate that forage P content is likely to affect 

growth rates and distribution of wild grazers in both habitats. Calcium can also be a critical nutrient for large herbivores in tropical grasslands (McNaughton, 1988, 1990; 

Murray, 1995), but in the Ewaso ecosystem, grass calcium 

content is relatively constant across soil types, insensitive 

to variation in soil Ca availability, and sufficient to meet 

demands of grazing ungulates (Figure 1F,J). 

Nitrogen content of grasses during the wet season on 

both the red sands and black cotton soils also appears sufficient to meet the protein requirements of ungulate grazers. 

For example, grasslands in the Serengeti ecosystem supporting high wet-season ungulate densities contain leaves with 

2.2%–2.6% N dry weight (Murray, 1995), similar to grass 

leaf N content on black cotton soil (mean = 2.3%, 1 standard error (SE) = 0.1%) and lower than grass leaf N content 

on red soil (mean = 2.9%, 1 SE = 0.1%). For high-yielding 

dairy cattle and actively growing beef cattle, optimum N 

content in the diet is typically in the range of 2.1%–3.2% 

dry weight (McDowell, 1985; Whitehead, 1995). We note 

that our analysis does not address dry-season N limitations, 

when forage N content below 1% N can limit rumen microbes (Van Soest, 1994). Although our results emphasize P 

as a critical limiting nutrient for ungulates in Laikipia, the 

significantly lower N content in black cotton grasses compared to red sands suggests that future studies in the black 

cotton habitat should consider factors influencing seasonal 

and spatial variation in grass protein content. 

Influence of Abandoned Bomas on Landscape

Nutrient Heterogeneity

In the Ewaso ecosystem, shortgrass and shrub-free 

lawns on the order of 0.2–1 ha in size are a widespread 

and conspicuous feature of the landscape. These shortgrass lawns are derived from abandoned livestock bomas, 

contain nutrient-enriched soil, and support a unique 

plant community dominated primarily by the stoloniferous grass Cynodon plectostachyus (Young et al., 1995; 

Augustine, 2003a). In the red sands habitat, abandoned 

boma sites can persist in this shrub-free, shortgrass state 

for at least four decades and likely for centuries (Augustine, 2003a). In the black cotton soils, Cynodon lawns 

usually develop to a Pennisetum stramineum–dominated 

but still shrub-free stage over one to three decades and 

likely remain in the latter state many decades more (Veblen and Young, 2010). Hereafter, we refer to all of these 

nutrient-enriched, shrub-free communities that have developed on abandoned boma sites as “glades.” In some of 

our studies, we used aerial photographs from the 1960s 

to determine ages of abandoned boma sites, and we examined soils, plant nutrients, and herbivore distribution 

in relation to glades created by boma abandonment during the past four decades (young glades) versus those 
number 632 • 5 9

FIGURE 1. Variation in concentrations of N, P, and Ca in (A–C) the total soil pool, (D–F) inorganic soil pool, and (G, H, J) wet-season 

grass leaves across soil types (n = 60 red soil sampling locations, 20 transition soil locations, and 20 black cotton soil locations) at the 

Mpala Research Centre in central Laikipia. Dashed lines in (H) and (J) show forage phosphorus and calcium requirements for lactating 

wildebeest (Murray, 1995).

0

0.02

0.04

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a a b

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Total Soil P (mg/kg dry soil)

NS

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Extractable NH4

+ + NO3

- (mg N/g dry soil)

 NS

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Extractable P (mg/g dry soil)

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 a ab b

G
6 0 • smithsonian contributions to zoology

derived from bomas abandoned more than 40 years ago 

(old glades). 

In the red sands habitat, soil nutrient content is substantially enriched on abandoned boma sites relative to 

the surrounding landscape and also varies in relation to 

the age of abandoned boma sites. Over the first four decades following boma abandonment, soil carbon and nitrogen decline relatively rapidly, while soil phosphorus is 

more tightly retained; old glades contain 1.9 times more 

N and 7.4 times more P than surrounding soils (Augustine, 2003a). Analysis of soil phosphorus availability using the Mechich-3 extractant (Mehlich, 1984) from soil 

cores (0–15 cm depth) collected at four glades and four 

paired bushland sites revealed 27 times more extractable 

phosphorus in glade versus bushland soils (Figure 2, left; 

paired t3

 = 4.82, P = 0.02). In the black cotton habitat, 

soil cores (0–30 cm depth) collected from nine glades and 

nine black cotton sites showed 1.84 times more total soil 

N in glades, whereas phosphorus availability measured 

using the Olsen-P extractant (Olsen and Sommers, 1982) 

revealed 20 times more extractable phosphorus in black 

cotton glades compared to nonglade soil (nitrogen: t8 = 

4.47, P = 0.0002; phosphorous: t8 = 3.48, P = 0.008; Figure 2, right). Although the Olsen extraction appears to 

provide lower estimates of soil P availability compared to 

the Mehlich extraction, both methods revealed the >20-

fold larger soil P pools in glade compared to nonglade 

soils. 

A study of grass nutrient content in the red sands habitat during the wet season evaluated variation associated 

with glades, topography, and plant species (Augustine, 

2004). Results from this study, combined with previous 

analyses of spatial variation in grass species distributions 

(Augustine, 2003b), illustrate the variation in forage P 

content that a selectively foraging impala could experience 

in the red sands habitat (Figure 3). Mineral requirements 

specific to impala have not been evaluated; hence, we rely 

upon previously discussed requirements on the order of 

3100–4000 mg P/kg dry matter for wildebeest and cattle 

(McDowell, 1985; Murray, 1995). Although a selectively 

foraging ungulate moving through the red sands habitat 

experiences substantial small-scale variation in forage P 

content among grass species and topographic positions, 

access to P-rich forage in glades is clearly important for 

meeting dietary demands (Figure 3). Within the nutrientpoor red sands habitat, P content of grasses is similar to 

grasslands in the Serengeti ecosystem of Tanzania, which 

Red Sands

Glade Non-Glade Non-Glade

Mehlich-3 Extractable Soil P (mg/kg dry soil)

0

100

200

300

400

500

Glade

Olsen Extractable Soil P (mg/kg dry soil)

0

10

20

30

40

50

Black Cotton

FIGURE 2. Differences in extractable soil phosphorus content of 

glade and surrounding nonglade soils for the red sands (four glade 

versus four paired nonglade sites; paired t

3

 = 4.82, P = 0.02) and 

black cotton soil (nine glade versus nine paired nonglade sites; t

8

 = 

3.48, P = 0.008) at Mpala Research Centre in central Laikipia.

0

1000

2000

3000

4000

5000

6000

0 200 400 600 800 1000

Distance (m)

Grass leaf P content (mg/kg dry mass)

Topographic Variation

Variation among 

grass species

Glade

FIGURE 3. Spatial representation of variation in forage phosphorus 

content that an herbivore in the red sands landscape of Laikipia, 

Kenya, might encounter while traveling along a 1 km transect beginning at a ridge, passing through a glade, and ending in a drainage, based on forage P content analyses (Augustine, 2004) and grass 

species distributions in the red sands habitat (Augustine, 2003b). 

Variation at scales of 5–15 m results from the patchy distribution of 

dominant grasses. The dashed line shows the forage P requirement 

measured for lactating wildebeest (Murray, 1995).
number 632 • 6 1

migratory ungulates are only able to use during the dry 

season, outside of calving and peak lactation periods. In 

contrast, P content in glades is similar to the shortgrass 

Serengeti plains to which ungulate herds migrate in the 

wet season for calving (McNaughton, 1990). Phosphorus 

content in glade grasses at MRC is also similar to grassland “hot spots” in the Serengeti ecosystem that support 

high concentrations of resident herbivores, including impala (McNaughton, 1988). If a selectively foraging impala 

were able to acquire a dietary intake of 2800 mg P/kg forage from the red sands bushland habitat, for example, by 

foraging in lowlands and feeding on red sands grasses such 

as Cynodon dactylon and Pennisetum stramineum, the 

impala would still need to acquire an estimated 42% of its 

forage from glades in order to achieve an overall dietary P 

content of 3900 mg P/kg. 

Similar measures of grass nutrient content in black cotton glades have not yet been conducted. However, the soil 

P and N enrichment documented in black cotton glades 

combined with low grass P content found in grasses on 

the background black cotton soil suggests that (1) grasses 

on black cotton glades are likely to be enriched in N and 

P and (2) the level of P enrichment may be important for 

meeting nutritional requirements of wild grazers. Quantitative measures of how grass nutrient content varies in the 

black cotton habitat in relation to glades versus other factors such as seasonality, termite mounds, topography, and 

plant community composition are clearly needed.

Influence of Abandoned Bomas on Ungulate

Distribution

On the basis of the previous findings for forage nutrient content, it is not surprising that in the landscape 

encompassing both black cotton and red sands habitats, 

some native ungulate species exhibit preferential use of 

abandoned bomas. In both habitats, we used dung group 

surveys of transects in abandoned bomas compared to the 

surrounding background vegetation to quantify these patterns. In the red sands habitat, we measured dung group 

densities in December 1999 at (1) all known glades abandoned between 10 and 40 years ago (n = 38), (2) glades 

abandoned >40 years ago (n = 16), and (3) a grid of transects distributed systematically across all bushland habitat at MRC (n = 46). Dung groups were counted within 

three randomly located 20 m transects using a 2 m width 

for impala dung and a 4 m width for all other species at 

each glade site and within a 50 × 4 m area for each bushland transect. Cattle dung group counts could possibly include cape buffalo dung, but direct counts of both species 

indicated that buffalo were extremely rare (65 times less 

abundant than cattle) in the study area (D. J. Augustine, 

unpublished). Data were analyzed by standard one-way 

ANOVA, with Tukey’s HSD for pairwise comparisons. 

In the black cotton habitat, we measured dung group 

densities in August 2003 at five glades abandoned <42 

years ago and five glades abandoned >42 years prior to 

sampling. Inside each of the 10 glade sites we identified 

and counted all herbivore dung within two 400 m2

 blocks 

(except at one site where we had only one block). We also 

sampled dung group densities outside of each of the 10 

glade sites along two 4 × 100 m transects, one located 

100 m away from the boma site and one located 200 m 

away. Data were analyzed as a split-plot ANOVA, treating glade age as the main plot effect and location (inside, 

100 m or 200 m) as the subplot effect. Pairwise comparisons were made with planned orthogonal independent 

contrasts testing (1) inside versus outside and (2) 100 m 

versus 200 m. 

In the red sands habitat, impala dung group densities 

were 21–35 times greater in old and young glades compared with surrounding bushland habitat (Figure 4; F2,95 

= 28.2, P < 0.0001). For zebra (Equus burchelli and E. 

grevyi combined), dung group densities were four to five 

times greater in glades compared with surrounding bushland (Figure 4; F2,95 = 11.98, P < 0.0001). For eland (Taurotragus oryx), densities were four times greater in young 

glades versus bushland habitat, with intermediate densities in old glades (Figure 4; F2,95 = 11.82, P < 0.0001). For 

cattle, densities were three times greater in glades compared to surrounding bushland habitat (Figure 4; F2,95 = 

29.6, P < 0.0001). Waterbuck (Kobus ellipsiprymnus) did 

not show significant selection for glades (F2,95 = 2.19, P = 

0.12). The two browsers (elephant, Loxodonta africana,

and giraffe, Giraffa camelopardalis) also did not select 

glades (F2,95 < 0.65, P > 0.53). 

In the black cotton habitat, dung group densities of 

hartebeest (Alcelaphus buselaphus) and Grant’s gazelles 

(Gazella granti) were three to seven times higher inside 

boma sites of both age classes relative to surrounding 

bushland (Figure 5; inside versus outside, hartebeest: F1,16 

= 5.56, P = 0.03; Grant’s gazelles: F1,16 = 8.75, P = 0.01). 

Cattle densities were three times higher in young glades 

(new: in versus out, F1,16 = 3.92, P = 0.07), whereas oryx 

(Oryx beisa) densities were six times higher in old glades 

(old: in versus out, F1,16 = 8.10, P = .01). Eland showed 

an inconsistent response to black cotton glades (location, 

F2,16 = 0.54, P = 0.59). In contrast to the red sands habitat, 

elephant dung density was six times higher in newer (<42 

years) boma sites in black cotton habitat (new: in versus 
6 2 • smithsonian contributions to zoology

FIGURE 4. Red Soils: Selection of old glades (>40 

years since boma abandonment, n = 16 sites) and 

young glades (10–40 years since boma abandonment, 

n = 38 sites) relative to surrounding bushland habitat 

(n = 44 sites) for ungulates in the red sands landscape 

in central Laikipia, Kenya, as indexed by dung group 

densities. “Zebra” includes both plains and Grevy’s 

zebra. Error bars show 1 SE above the mean. For each 

species, bars with different letters indicate significant 

(P < 0.05) differences between the three habitats, and 

NS indicates species with no significant differences 

among the three habitats.

Zebra

Cattle

Dung groups per hectare

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Elephant

Eland

Hartebeest

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Giraffe

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<42 yr Glades

Bushland

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e

a b c

>40 yr Glades

10 - 40 yr Glades

Bushland

FIGURE 5. Black Soils: Selection of old glades (>42 years since boma abandonment, n = 5 sites) and young glades (<42 years since 

boma abandonment, n = 5 sites) relative to surrounding savanna habitats (100–200 m distant from glades, n = 10 sites) on black 

cotton soil in central Laikipia, Kenya, as indexed by dung group densities. “Zebra” includes both plains and Grevy’s zebra. Error 

bars show 1 SE. For each species, bars with different letters indicate significant (P < 0.05) differences between the three habitats, and 

NS indicates species with no significant differences among the three habitats.
number 632 • 6 3

out, F1,16 = 5.62, P = 0.03). Giraffe avoided boma sites (in 

versus out, F1,16 = 7.67, P = 0.06). Likewise, zebra densities 

were not elevated inside black cotton glades and instead 

showed a nonsignificant trend toward avoidance of glades 

relative to woodland (location, F2,16 = 1.52, P = 0.25). 

We have also used infrared camera monitors to conduct more detailed analyses of how glades influence seasonal and spatial distribution of impala in the red sands 

habitat (Augustine, 2004). Camera monitors showed that 

impala preferentially used glades compared to surrounding bushland in both wet and dry seasons but increased 

glade use ninefold during wet compared to dry seasons 

(Figure 6A). Furthermore, impala density was highly correlated with the distance to the nearest glade or abandoned boma (Figure 6B), with impala density declining to 

extremely low levels in areas where glades were more than 

600 m distant. 

Overall, our studies demonstrate that across diverse 

soil types, the management of livestock via bomas creates 

nutrient hot spots within the landscape. These hot spots 

have a strong influence on wild ungulate distribution and 

potentially influence overall carrying capacity. Our work on 

impala showed that (1) glades are critical for meeting wetseason phosphorus requirements, (2) impala show strong 

preferential use of glades, especially during the wet season, 

and (3) portions of the landscape distant from glades support few impala. Fine-scale distribution analyses have not 

been conducted for other wild herbivores, but many species show clear preferential use of glades (Figures 4 and 

5). Our surveys have also documented some notable differences between red sands and black cotton habitats, especially with zebra and elephant use of glades. Factors such 

as differences in grazing pressure or in the structure and 

composition of the plant communities in the two habitat 

types may be involved and merit future research attention.

Our studies also suggest that anthropogenic glades 

can contribute to the viability of ungulate species of conservation concern in the Ewaso ecosystem. Although 

species such as plains zebra and Grant’s gazelles may be 

adaptable to livestock-dominated landscapes, others such 

as impala, hartebeest, eland, waterbuck, and buffalo may 

require deliberate management efforts (Georgiadis et al., 

2007a). The hartebeest population in Laikipia is the largest of only three distinct populations of A. buselaphus

remaining in Kenya that are hybrid between two morphotypes (A. b. cokei and A. b. lelwel), the remainder having 

been extirpated by human activities. Hartebeest have been 

declining in Laikipia over the past decade and serve as one 

of several focal species for conservation efforts in the region (Georgiadis et al., 2007b). In the black cotton habitat, hartebeest are strongly attracted to glades, especially 

FIGURE 6. (A) Seasonal variation in the abundance of impala in 

glades versus adjacent red sands bushland habitat and (B) spatial 

variation in wet-season impala abundance within bushland habitat 

as a function of the distance to the nearest glade or abandoned boma 

in Laikipia, Kenya, 2001–2002. Impala herds were present in glades 

at a significantly higher rate during wet versus dry seasons (season × 

habitat interaction: F1,27 = 23.96, P < 0.001). Error bars show 1 SE 

above the mean, and the number above each bar in (B) shows the 

number of locations sampled within each distance class. Reprinted 

from Augustine (2004) with permission; © The Wildlife Society.
6 4 • smithsonian contributions to zoology

younger glades (Figure 5). Hartebeest also preferentially 

forage on abandoned bomas in southern Kenyan rangelands (Stelfox, 1986). Hartebeest are primarily associated 

with open, treeless grasslands; provision of nutrient-rich 

glades in combination with such open areas may be one 

important means to enhance habitat for the species. 

The Grevy’s zebra is another rare and declining species, with much of the remaining population found in the 

Ewaso ecosystem (Williams, 2002). Grevy’s zebra at MRC 

are primarily found in the red sands and adjacent transitional habitats dominated by Acacia bushland. Grevy’s 

zebra habitat use at MRC differs for lactating versus nonlactating mares, with lactating mares selecting the shortest, greenest grass swards in the landscape (Sundaresan 

et al., 2008). Abandoned bomas are the primary source 

of these shortgrass swards. Use of glades by Grevy’s zebra may be constrained by availability of drinking water 

(which lactating females require frequently) and the locations of active bomas (which Grevy’s zebra avoid; Sundaresan et al., 2008). Although Grevy’s zebra avoid areas 

with herders and cattle in the short term, they also appear 

to benefit from long-term consequences of boma rotation. 

Consideration of how water sources and shrub density interact with glade distribution could assist in boma placement and rotation planning to provide critical habitat for 

such declining wildlife species.

Long-­Term Glade Persistence

Glades in the Ewaso ecosystem persist for at least four 

decades and could potentially persist for centuries, as reported for nutrient-enriched patches in a southern African 

savanna (Blackmore et al., 1990). During 1961–2000 at 

MRC, bomas in the red sands habitat were abandoned 

at a mean rate of 1.15 per year, while long-term glades 

(i.e., those present in 1961) were invaded by Pennisetum

at a rate of only 0.08 per year (Augustine, 2003a). In 

other words, glade density in the landscape was increasing during this four-decade period primarily because 

existing glades persisted in a nutrient-enriched, Cynodondominated state for well beyond 40 years. Soil nitrogen 

can be lost from glades over time via several pathways, 

including leaching, denitrification, and volatilization, such 

that long-term glade persistence likely depends upon nitrogen inputs to replenish these losses. Detailed analysis 

of glade nitrogen budgets at MRC showed that during dry 

seasons, impala bed within glades (Figure 6) while foraging in adjacent bushland, causing a substantial net input of 

nitrogen via dung and urine deposition (Augustine et al., 

2003). This N input may facilitate long-term glade persistence in the Cynodon-dominated state and represents a 

pathway by which impala benefit all ungulate grazers that 

make use of glades, including livestock. 

In the black cotton soils, there is more rapid transition 

from Cynodon-dominated glades to glades dominated by 

Pennisetum stramineum (in as few as 20–30 years), but the 

latter also persist for a minimum of 40 years, and probably 

much longer. In these black cotton soils, the long-term persistence of glades may be associated with increased termite 

activity (not seen in the red soil glades), which not only enriches soil fertility but also changes soil texture and reinforces the treelessness of glades. Another factor may be the 

length of time a boma is occupied by livestock, which in 

black cotton soils can be limited by rainy-season accessibility. In addition, a recent study of multiple proximate glades 

in black cotton soils in Laikipia (Porensky, in press) has 

shown that the areas between glades within 150 m of each 

other have elevated tree densities and concomitant low use 

by herbivores, perhaps because of low predator visibility 

(see below). This result suggests that although bomas can 

be useful in enhancing wildlife habitat, details of boma density, configuration, and persistence can produce emergent 

patterns that will also need to be considered.

Taken together, our studies have several implications 

for boma and nutrient management in this ecosystem. First, 

boma rotation is key to glade formation and hence ungulate hot spots, indicating boma placement can be used as 

a tool to attract wild ungulates into particular locations 

in the future. Fertilization at the hectare scale, particularly 

with phosphorus, may also be a means to induce herbivore 

hot spots in desired locations. Sedentarization of bomas or 

elimination of boma use may have negative long-term consequences for wildlife species such as impala, hartebeest, 

Grevy’s zebra, eland, and Grant’s gazelle. Long-term persistence of glades may also be contingent on use by wild ungulates. In bushlands on red soils, where shrub-free glades 

attract herbivores year-round, dung and urine inputs from 

native ungulates such as impala can prevent long-term nitrogen loss from glades and hence can maintain them in 

a shortgrass, Cynodon-dominated state (Augustine et al., 

2003). In purely grassland habitats, soil and plant nitrogen 

in glades may decline, and Pennisetum cover may increase 

over time. However, enriched P content in glade grasses is 

likely to persist for decades in all soil types and to be an 

important wet-season forage resource for native ungulates. 

NATIVE BROWSERS AND RANGELAND 

SHRUB ENCROACHMENT

Worldwide, rangeland ecosystems are experiencing increases in woody plant abundance and associated declines 
number 632 • 6 5

in their suitability for cattle production (Scholes and 

Archer, 1997; Van Auken, 2000). Shifts from grassland to 

shrubland may be related to changes in climate (Polley et 

al., 1997; Morgan et al., 2007), fire frequency and grazing 

intensity (Dublin et al., 1990; Roques et al., 2001; Tobler 

et al., 2003), and browsing pressure (Dublin et al., 1990). 

In some areas of Africa, woodlands have declined in protected areas coincident with a concentration of elephants 

into these areas (e.g., Laws, 1970; Dublin et al., 1990; Van 

de Vijver et al., 1999). Conversely, woodland expansion is 

often observed in African rangelands managed for cattle 

production (Scholes and Archer, 1997; Oba et al., 2000; 

Roques et al., 2001; Tobler et al., 2003). A consequence of 

lowered tree or shrub densities may be increased availability of understory nutrients (Treydte et al., 2007), whereas 

increased densities of trees and shrubs can negatively affect grass productivity (Scholes and Archer, 1997; Smit 

and Rethman, 2000). High densities of spinescent Acacia

shrubs can also physically impede cattle access to the understory, effectively reducing forage availability. 

Our studies in Laikipia have addressed the effects 

of native browsers on woody plant dynamics on rangelands where native fauna coexist with cattle. Here we first 

present baseline information on the abundance of native 

browsers in the red sands and black cotton landscapes 

and the history of woody plant dynamics in the district. 

We then synthesize results of exclosure studies in both red 

sands and black cotton soils, addressing the effect of native 

browsers on woody plant dynamics. Finally, we discuss 

the long-term implications for both cattle and wildlife. 

Historic Trends in Shrubs and Browsers

in the Ewaso Ecosystem

The distribution of woody vegetation and elephants 

across the Ewaso ecosystem has changed substantially 

since settlement by European ranchers in Laikipia began 

in 1912. As reviewed by Thouless (1995), none of the 

early European explorers, hunters, and ivory traders who 

traveled through the Ewaso ecosystem during 1880–1910 

found elephants in central Laikipia, and this area consisted 

mostly of open grass plains. Laikipia is now dominated by 

Acacia bushland on red soils and contains extensive areas of A. drepanolobium woodland on black cotton soils 

(Taiti, 1992; Young et al., 1998; Augustine, 2003b; Table 

1). As recently as a few decades ago, much of Laikipia was 

open grassland on both the red and black soils (Heady, 

1960; T. P. Young, personal communications with living 

ranch owners and herders).

Early in the twentieth century, elephants in this region were found only in Samburu district to the north of 

Laikipia, but by the 1950s elephants were increasing and 

expanding their range toward Laikipia (Thouless, 1995). 

Heavy poaching during the 1970s and 1980s strongly affected elephant numbers and distribution in Samburu. By 

the 1990s, telemetry studies identified a significant population of elephants (estimated at ~800 individuals) that 

regularly migrated 80–100 km between pastoral rangelands in Samburu and the commercial ranches in Laikipia 

(Thouless, 1995). Our studies have examined savanna dynamics on the commercial rangelands in Laikipia utilized 

by this migratory elephant population. Our own surveys 

of ungulate densities at MRC during 2000–2002 (distance 

sampling of 93 km of ground transects in the red sands 

habitat at 6 month intervals; densities calculated following Thomas et al., 1998) show a continued migration of 

substantial numbers of elephants into this region during 

June–October each year (Table 2). In addition, the Acacia 

bushland habitat on the red soils supports high densities 

of dik-dik, a small antelope that forages predominantly 

on woody vegetation (Table 2). Dik-dik were unlikely to 

have been abundant in grasslands that dominated Laikipia 

early in the twentieth century. Rather, the high abundance 

of dik-dik and migratory elephants are likely a response, 

in part, to increases in woody vegetation. Other browsing species also likely to have increased as the density of 

woody species has increased include giraffe, eland, steinbuck, and kudu. 

Current landscape-scale estimates of shrub and tree 

cover in central Laikipia are approximately 28% for the 

red sands (Augustine, 2003b) and 31% for the black cotton soils (Riginos et al., 2009). A continental-scale analysis of woody cover in African savannas predicts that in 

the long-term absence of disturbances (i.e., no browsers or 

TABLE 2. Densities of dik-dik and elephant (number/km2) in the 

red sands habitat at Mpala Research Centre during dry seasons 

(March) and wet seasons (August) of 2000–2002. CI, confidence 

interval.

Dik-dik Elephant

Mean Mean

Season density 95% CI density 95% CI

2000 dry season 130 (99, 170) 0.0 (0.0, 0.0)

2000 wet season 140 (102, 194) 5.2 (1.4, 19.8)

2001 dry season 139 (106, 182) 0.04 (0.01, 0.39)

2001 wet season 139 (119, 164) 2.9 (1.3, 6.4)

2002 dry season 145 (121, 173) 0.3 (0.1, 1.7)

Five-survey mean 139 (124, 156) 1.7 (0.8, 3.4)
6 6 • smithsonian contributions to zoology

fire), the rainfall-limited bound for woody cover on both 

soil types in central Laikipia based on 500 mm mean annual precipitation is approximately 56% (Sankaran et al., 

2005). Continued increases in shrub and tree abundance 

toward this maximum could have at least two negative 

effects on herbivores. First, high densities of shrubs and 

trees can substantially suppress grass production (Smit 

and Rethman, 2000; Smit, 2004) and hence reduce forage for cattle. In the black cotton habitats, sites with a 

low density of trees were found to have double the grass 

production as sites with a high density of trees (Riginos et 

al., 2009). Further research is needed to quantify the effects of tree density on grass productivity in the red sands 

habitat. Second, woody encroachment is likely to directly 

affect some native wild herbivores by reducing the quality of the habitat available to them. In the black cotton 

habitat, a recent study examined patterns of herbivore use 

in A. drepanolobium woodland habitats that varied from 

240 to 2,784 trees per hectare, or 6 to 45% canopy cover. 

All but the largest herbivore species (elephants) spent significantly more time in areas with low tree density and 

avoided areas with high tree density (Riginos and Grace, 

2008). This preference was independent of grass cover or 

species composition. Rather, herbivore habitat preference 

was strongly related to visibility; in more open habitat 

with low tree density and cover, predators are more visible, making those areas favorable for vulnerable herbivore 

species. Thus, areas with high tree density represent lowerquality habitat for small- and medium-sized herbivores, 

including the declining hartebeest. Areas with high tree 

density, in contrast, actually attracted elephants and experienced more damage per tree (Riginos and Grace, 2008). 

These relationships suggest that elephants and other native browsers can have important feedbacks on landscapescale woody vegetation dynamics, with potential benefits 

to native grazers and cattle alike.

Effects of Native Browsers on Shrub

and Tree Dynamics in Laikipia

Two large-scale herbivore exclusion experiments at 

MRC have demonstrated the effects of native browsers on 

shrubs and trees. In 1999, a set of exclosures was established in the red sands habitat to examine effects of browsers on bushlands dominated by A. mellifera, A. etbaica,

and A. brevispica (Augustine and McNaughton, 2004). 

Exclusion of browsers for just three years led to rapid 

shrub encroachment, measured in terms of shrub cover, 

density, and biomass (see Table 4). This response was 

due to the effects of two vastly different native species. 

First, high densities of dik-dik (Madoqua kirkii) strongly 

suppressed growth of shrubs less than 0.5 m high; exclusion of dik-dik led to rapid recruitment of shrubs into 

taller height classes, which was the primary determinant 

of increases in shrub density. Second, browsing elephants 

focused mainly on large (>2.5 m high) shrubs, significantly 

suppressing rates of change in shrub cover and aboveground biomass (Augustine and McNaughton, 2004). 

On the black cotton soils, the Kenya Long-term Exclusion Experiment (KLEE), initiated in 1995, has examined browser effects on woodland dynamics (Young et al., 

1998). Studies at this site have confirmed that cattle do 

not eat the dominant woody plant, Acacia drepanolobium

(Odadi et al., 2007), whereas native browsers regularly 

feed on this species. We used three measurements from 

the KLEE plots to examine how the exclusion of native 

browsers has affected woodland encroachment in this ecosystem. First, we monitored annual changes in size of A. 

drepanolobium in KLEE plots from which all large herbivores have been excluded and in KLEE plots to which all 

large herbivores have access. Briefly, we measured crown 

breadth, diameter at breast height, and height of all A. 

drepanolobium trees in two 150 × 10 m strip transects 

at each plot each year during 2004–2007 (see Goheen et 

al., 2007). We use this data set to quantify differences in 

canopy cover of adult A. drepanolobium (>1 m) between 

the ninth and twelfth year of large-herbivore exclusion 

and also changes in height of large trees (>2.5 m) during 

this time. Second, we measured the height, diameter at 30 

cm above ground, and location for every tree >1 m tall 

within the central hectare of plots exposed to large herbivores and plots from which large herbivores had been 

excluded. We use these data to compare differences in density and biomass of A. drepanolobium stemming from the 

10 year exclusion of large herbivores. Third, we measured 

the cover of all shrub species other than A. drepanolobium

in the KLEE plots in 2001, which we use to quantify effects of herbivores on the rare shrub species after six years 

of herbivore exclusion. We used mixed-model ANOVA to 

analyze these measurements, treating large-herbivore exclusion as a fixed factor and replicate as a random factor. 

Large herbivores reduced the density of tall (>1.0 m) 

A. drepanolobium by 32% (F1,2 = 16.14, P = 0.06), reduced mean canopy cover by 28% (F1,2 = 2.54, P = 0.25), 

and reduced woody biomass of Acacia drepanolobium by 

29% (F1,2 = 5.88, P = 0.14; Figure 7). Even more dramatic 

was the increase in canopy cover of the several species 

of subordinate trees and shrubs, which accounted for a 

relatively small fraction of total woody abundance (Table 

3). Density of subordinate species was not affected by 

the exclosure treatment, but established plants that had 

been suppressed by constant browsing were released when 
number 632 • 6 7

protected from herbivory. The mean canopy sizes of some 

species increased by as much as twentyfold, and the percent contribution to woody cover across all these species 

was 12.8% in exclosures compared to 3.8% in the presence of browsers (Table 3). 

Our studies have shown that browsers strongly influence woody vegetation dynamics in both soil types, and 

they also highlight differences in the strength of browser 

effects (Table 4). Following the experimental removal of 

native browsers, the cover, density, and biomass of shrubs 

have been increasing on both soils, but at three to seven 

times greater rates on red sands compared to black cotton soils (Table 4). One potential factor underlying these 

rate differences may be high dik-dik densities combined 

with substantial elephant browsing pressure in the red 

sands, compared to elephants being the only high-density 

browser in the black cotton habitat. 

A second important factor influencing shrub encroachment rates may be differences in herbivore resistance traits of the dominant A. drepanolobium on black 

cotton soils compared to the dominant Acacia species in 

the red sands. Results for the subdominant shrub species 

on the black cotton soil, which showed a fourfold increase in cover over just six years (Table 3), also indicate 

that browsers exert less control over A. drepanolobium

compared to other co-occurring woody species. Ant symbionts that occur on A. drepanolobium effectively protect 

host trees and greatly reduce browse intake by elephants 

(Goheen and Palmer, 2010). In addition, subordinate species in the black cotton habitat may experience low rates 

of growth and reproduction on black cotton soils relative to red soils and therefore may not be able to compensate after being browsed by large herbivores (Harper, 

1969; Holt and Lawton, 1994). Fire is another management tool that has substantial potential to influence shrub 

dynamics and interact with the effects of browsing ungulates (Dublin et al., 1990). There is widespread belief 

among ranchers, pastoralists, herders, and ecologists that 

fire, and probably intentional burning, was once far more 

frequent and widespread in Laikipia in previous centuries 

and that it suppressed woody plant densities. Fire effects 

may be stronger in the black cotton habitat because of 

the continuous understory grass cover, but grass cover 

in the red sands would likely have been great enough in 

at least some years to carry a fire. In addition, there can 

be synergistic effects of fires and herbivores. In a recent 

study in the black cotton habitat at MRC, burnt A. drepanolobium trees were found to survive the fires themselves 

but later to succumb to increased browsing pressure in 

the burned areas (Okello et al., 2008). Using ground and 

satellite-derived measurements of canopy cover, we estimate that although shrub cover on unburned KLEE plots 

accessible to native wild herbivores was 18% (Figure 7), 

the combination of fire and native herbivore presence 

reduced average canopy cover to 13%. These findings 

emphasize the need to understand herbivore responses to 

FIGURE 7. Effects of removing large herbivores on the density of 

tall (>1.0 m) individuals, percent canopy cover, and woody biomass 

of Acacia drepanolobium in the Kenya Long-term Exclusion Experiment. Error bars show 1 SE above the mean; n = 3 exclosures and 3 

paired control sites for each comparison.
6 8 • smithsonian contributions to zoology

fires in semiarid rangelands in order to take full advantage of burning as a management tool.

CONCLUSIONS

Two themes in African wildlife conservation have 

emerged over the past decade. The first is the underlying 

importance of habitat heterogeneity across multiple spatial scales in sustaining biodiversity (du Toit et al., 2003; 

Cromsigt and Olff, 2006). The second is the critical role 

of nonprotected lands, especially managed rangelands, in 

sustaining Africa’s large mammalian fauna (Prins et al., 

2000). These themes are related, in that protected areas 

of Africa often do not encompass sufficient heterogeneity 

across scales to maintain the range of habitats that sustain diverse and abundant mammalian communities. Our 

studies show that facilitation via habitat modification is 

an important pathway for cattle-wildlife interactions in 

semiarid rangelands. First, nutrient-enriched patches created by abandoned cattle bomas can persist for decades 

to centuries and provide a key source of nutrient-rich forage for both native and domestic grazers. Nutrient-rich 

glades that develop from abandoned bomas in the red 

sands habitat not only sustain local impala abundance but 

are also maintained as glades by nitrogen inputs from impala (Augustine et al., 2003). Such feedbacks and interactions among cattle bomas, soil and plant nutrients, and 

wild ungulates indicate that ranch managers can influence 

the long-term distribution and abundance of wild ungulates though the placement and rotation of current bomas. 

Second, although cattle can directly compete with native 

grazers for forage (Young et al., 2005), native browsing 

ungulates can provide an important compensatory benefit 

to cattle by controlling shrub encroachment. There is even 

evidence that the ability of elephants to reduce forb cover 

may interact with effects of cattle grazing in a manner that 

reduces competition between cattle and zebra (Young et 

al., 2005). Soils have a strong influence on the structure 

of woody plant communities and rates of shrub encroachment, but the influence of native browsers appears to be 

robust across such variation. 

To directly measure the importance of such facilitation through habitat modification, we would ideally compare ranches with dynamic boma management and the 

presence of native fauna to ranches where bomas are not 

rotated and native fauna are excluded. An example of 

the latter is provided by a large ranching enterprise in the 

coastal savannas of Tanzania. There the 462 km2

 Mkwaja 

ranch was managed for cattle production using a system of 

permanent paddocks where cattle were kept at night; native ungulates were rare throughout the property (Tobler 

et al., 2003; Treydte et al., 2006). Over a 48 year period, 

this intensive, cattle-only ranching operation proved to be 

economically and ecologically unsustainable, primarily 

because of severe bush encroachment that developed in 

TABLE 3. Differences in canopy cover of nine subdominant shrub species in the black cotton 

habitat after six years of large-herbivore exclusion. Although these species are typically rare 

within the plant community, following browser exclusion, they exhibited more rapid rates of increase than the dominant woody species, Acacia drepanolobium. For three species, there was insufficient sample size for reliable statistical analysis. Here n.a. denotes statistic is not applicable.

Mean Percent Canopy Cover

Herbivores Herbivores

Species excluded present Ratio F P

Cadaba farinosa 3.36 1.42 2.4 15.5 0.0002

Balanites aegyptiaca 4.23 1.06 4.0 36.4 <0.0001

Acacia mellifera 2.95 0.89 3.3 4.59 0.038

Lycium europaeum 0.86 0.36 2.4 n.a. n.a.

Rhus natalensis 0.92 0.04 22.0 2.72 0.14

Grewia spp. 0.17 0.02 6.9 7.80 0.019

Boscia angustifolia 0.15 0.03 4.4 3.67 0.07

Acacia brevispica 0.14 0.01 10.6 n.a. n.a.

Lippia javanica 0.05 0.004 12.5 n.a. n.a

Total 12.82 3.84 3.3 n.a. n.a.
number 632 • 6 9

wide bands around the permanent paddocks (Tobler et al., 

2003). Similarly in Laikipia, ranches intolerant of browsing wildlife are often characterized by greater tree densities. Although numerous ecological factors differ between 

coastal savanna and the Ewaso ecosystem, the long-term 

patterns observed in these rangelands illustrate the potential importance of boma rotation and browsing ungulates 

in sustaining the Ewaso ecosystem.

Acknowledgments

We thank Dana Augustine, Jackson Ekadeli, James 

Ekiru, Frederick Erii, Abdikadir Ali Hassan, Dan Kelly, 

Simon Lima, John Lochukuya, Francis Lomojo, and Margaret McNaughton for invaluable assistance in the field 

and lab. We also thank the Mpala Research Centre staff 

for logistical support. We thank Justin Derner and Nick 

Georgiadis for helpful reviews of the manuscript. Financial support was provided the James Smithson Fund of the 

Smithsonian Institution, the National Geographic Society, 

the National Science Foundation (DEB-9813050; DGE0114432; LTREB BSR-97-07477 and 03-16402; National 

Science Foundation Graduate Research Fellowships), the 

African Elephant Program of the U.S. Fish and Wildlife 

Service (98210-0-G563), the Jastro-Shields Award, the 

American Society of Mammalogists, the U.S. Environmental Protection Agency, and the U.S. Department of Agriculture’s Agricultural Research Service. 

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Ahn, P. M., and L. C. Geiger. 1987. Kenya Soil Survey: Soils of Laikipia 

District. Kabete, Kenya: Ministry of Agriculture, National Agricultural Laboratories. 

Arsenault, R., and N. Owen-Smith. 2002. Facilitation versus Competition in Grazing Herbivore Assemblages. Oikos, 97:313–318.

Augustine, D. J. 2003a. Long-Term, Livestock-Mediated Redistribution 

of Nitrogen and Phosphorous in an East African Savanna. Journal 

of Applied Ecology, 40:137–149.

TABLE 4. Summary of the effects of native browsing ungulates on woody plant dynamics at the Mpala Research Centre in central Laikipia, Kenya. Browser effects in the red sands habitat are from Augustine and McNaughton (2004) and Young and Augustine (2007). 

Browser effects in the black cotton habitat are derived from Gadd et al. (2001), Goheen et al. (2007), Riginos and Young (2007), and 

this chapter (see Figure 7).

Dynamic Red sands (A. etbaica and A. mellifera) Black cotton (A. drepanolobium)

Seed production No effect Reduced by half in the presence of browsers

Twig growth rate Annual twig growth rates negative or zero in the presence Simulated giraffe browsing resulted in compensatory

of dik-diks; positive and rapid in the absence of twig regrowth (i.e., no negative short-term effect 

dik-diks; no detectable browser effects above dik-dik on twig biomass), but this may occur at the

foraging height expense of long-term growth and/or seed 

production

Sapling recruitment Recruitment into 0.5–1.5 m height class reduced Enhanced in the presence of large herbivores, despite 

sixfold by browsers, primarily due to effects of dik-dik suffering greater browse damage; enhanced 

recruitment likely due to reduced density of 

conspecific adult trees

Shrub growth in large size class Browser presence reduced canopy volume and height for Browser presence reduced vertical growth by 49% 

shrubs >2.5 m tall, primarily due to elephant damage over three years for trees >2.5 m tall, primarily 

due to elephant damage

Net browser effect on woody plant abundance

Change in shrub density No change in the presence of browsers; rapidly increasing Adult trees >1.0 m increasing (+37 individuals

(+137 individuals ha-1 yr-1) with browser removal ha-1 yr-1) with browser removal

Change in shrub cover Shrub cover increasing rapidly with browser removal A. drepanolobium cover increasing slowly with 

(+1.92% per year) and declining (−0.75% per year) browser removal (+0.7% per year); cover of other 

in the presence of browsers shrub species three times greater after six years of 

browser exclusion.

Woody biomass accumulation Near zero (−26 kg ha-1 yr-1) in the presence of browsers; Increasing with browser removal (+137 kg ha-1 yr-1)

rapidly increasing (+962 kg ha-1 yr-1) with browser 

removal.
7 0 • smithsonian contributions to zoology

———. 2003b. Spatial Heterogeneity in the Herbaceous Layer of a Semiarid Savanna Ecosystem. Plant Ecology, 167:319–332.

———. 2004. Influence of Cattle Management on Habitat Selection by 

Impala on Central Kenyan Rangeland. Journal of Wildlife Management, 68:916–923.

Augustine, D. J., and S. J. McNaughton. 2004. Regulation of Shrub Dynamics by Native Browsing Ungulates on East African Rangeland. 

Journal of Applied Ecology, 41:45–58.

Augustine, D. J., S. J. McNaughton, and D. A. Frank. 2003. Feedbacks 

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Living with Lions: Lessons from Laikipia

Laurence Frank

ABSTRACT. Lion numbers are in steep decline in Africa. Lions have been eliminated 

from agricultural areas, and pastoralists have lost tolerance for predators; ready availability of cheap and effective poison poses a critical threat to their survival in Kenya 

outside of the largest national parks. However, lions still thrive on the commercial ranchlands of Laikipia District, where landowners use traditional low-cost African methods of 

livestock husbandry to effectively protect domestic stock from predators. Thus, Laikipia 

has been an ideal laboratory in which to investigate the ecology of predators living in human-dominated landscapes and factors in management of both livestock and lions which 

allow coexistence between them. Although socioeconomically unrepresentative of most 

African rangelands, the Laikipia experience shows that it is easy to raise livestock in the 

presence of viable populations of large carnivores if people have incentive and motivation to take basic, traditional measures to keep them apart. The current lack of financial 

incentives for rural Kenyans to tolerate wildlife, however, may limit the usefulness of the 

Laikipia model to privately owned lands.

INTRODUCTION

Laikipia District is unusual in many ways, but unique in one: it is the 

only place in the world where commercial ranchers actively conserve large 

predators and go to significant lengths to coexist with them. In Europe, North 

America, and Asia, humans have eliminated large carnivores to protect livestock, but at a time when Africa’s great predators are also rapidly disappearing, the Laikipia example shows that carnivore extinction is not a necessary 

consequence of livestock production and rural development. Nor does carnivore conservation require resources beyond traditional methods available 

and familiar to pastoralists throughout East Africa. Laikipia ranchers have 

found that the ancient techniques which traditionally allowed humans to rear 

livestock in natural ecosystems are still highly effective at protecting domestic 

animals from predators.

Laurence Frank, Living with Lions, Panthera, 

8 West 40th Street, 18th Floor, New York, New 

York 10018, USA; and Museum of Vertebrate 

Zoology, University of California, Berkeley, California 94720, USA. Correspondence: lgfrank@

berkeley.edu.

Manuscript received 7 January 2009; accepted 18 

May 2010.
7 4 • smithsonian contributions to zoology

In 1995 wolves (Canis lupus) were reintroduced into 

the northern Rocky Mountains despite the vehement 

and nearly universal opposition of the livestock ranchers 

of Wyoming, Montana, and Idaho. In 2005, the entire 

71,600 km2 Yellowstone ecosystem lost 13 cattle and 71 

sheep to wolves (Defenders of Wildlife, 2006); in spite of 

full monetary compensation and prompt removal of the 

offending wolves, the great majority of U.S. ranchers are 

still intensely opposed to wolf recovery. By comparison, 

in 1995–1996, the average (132 km2) ranch in Laikipia 

lost 10.6 cattle and 52.3 sheep to predators (Frank, 1998). 

Ranchers there are not compensated for their losses, and 

many make little money from tourism, yet few advocate 

the elimination of predators. Unlike American ranchers, 

most accept costs of livestock protection measures and occasional livestock losses as a part of doing business.

STATUS OF LIONS IN AFRICA

Ten thousand years ago, the lion (Panthera leo) was 

the most widely distributed mammal on earth, ranging 

across most of Europe, Asia, Africa, and North and South 

America (Turner and Anton, 1997; Yamaguchi et al., 

2004). It was eliminated from the New World at the end of 

the Pleistocene and from Europe and nearly all of Asia in 

historical times. In Africa, it has been eliminated from 83% 

of its historic range since European colonization, but only 

in the last few years have conservationists come to realize 

how serious the situation has become: the best current estimate is that fewer than 30,000 lions exist in the wild, and 

nearly half of those are in Tanzania (Chardonnet, 2002; 

Bauer and van der Merwe, 2004; IUCN, 2006). Most of 

the rest are in small and isolated national parks, vulnerable 

to stochastic events such as disease and political upheaval. 

Only six protected or managed areas are large enough 

to ensure long-term survival of lions and other wide-ranging mammals (IUCN, 2006); three of those are in Tanzania 

and one, the Tsavo complex, is in Kenya. Given Africa’s 

history of political instability and rapid human population 

growth, lion populations may soon come to resemble tiger 

populations: scattered among small parks, few of which 

are sufficiently large and interconnected to maintain a viable metapopulation. It is a truism of population biology 

that small, isolated populations cannot survive indefinitely 

without gene flow among them (Harcourt et al., 2001). 

Because human population growth and land pressures 

make creation of new large protected areas unlikely, maintaining viable populations of large mammals outside and 

between national parks is critical to their conservation. 

STATUS OF LIONS IN KENYA

European settlement of Kenya at the turn of the twentieth century had a major impact on wildlife generally 

and predators in particular. Because they readily prey on 

livestock, large carnivores were considered vermin, and 

settlers exterminated them from farming and ranching 

areas. As an example of the zeal with which lions were 

killed, safaris to the Serengeti area in the early 1900s 

sometimes shot over 100 lions (Turner, 1987), and clients of just one Nairobi safari company killed 700–800 

lions in 1911 alone (Herne, 1999:78–89). In 1908, over 

150 lions were killed “on license” in Laikipia District 

(Playne, 1909). This scale of slaughter was not exclusive 

to the early days of settlement: between 1946 and 1952, 

one Laikipia game warden shot 434 lions “on control” 

(Herne, 1999), and several individuals killed over 300 

lions apiece in the course of ranching in other parts of 

Kenya in the 1970s and 1980s (Anonymous individuals, 

personal communications). 

Much of this killing took the form of “sport” but 

was motivated primarily by the perceived need to protect 

domestic animals. Although ranchers in East Africa used 

traditional African cattle husbandry methods that effectively minimized losses (see below), western practice was 

to eliminate predators rather than try to live with them. 

Poison (strychnine and organochlorine cattle dips) was 

widely used on East African ranches, continuing well into 

the latter half of the twentieth century (Denney, 1972). At 

least until recently, both the Kenya Wildlife Service and the 

Kenya Veterinary Department poisoned and shot spotted 

hyenas (Crocuta crocuta) on a wide scale, affecting lions 

and other scavengers as well (Mwaniki, 1997; L. Frank, 

personal observation, 1991). 

In Kenya, lions were eliminated from settled agricultural areas in the last century, and today they persist only 

in national parks and reserves and on some rangelands. 

Extrapolating from known numbers in a few areas, we 

estimate a current Kenya population of less than 2,000 

individuals (Frank et al., 2006); there are no data upon 

which to base estimates for earlier periods. Until quite recently, they were still widespread in Masailand of southern 

Kenya and throughout much of northern Kenya (covering 

roughly half of Kenya’s land area), but that has changed 

dramatically in the past 20 years. Although there are no 

data available for the north, it seems clear that overgrazing and ubiquitous firearms have largely eliminated most 

wildlife, including lions. Again, there are no historical data 

for lion numbers in southern Kenya, but they were still 

abundant in much of Masailand until the current century, 
number 632 • 7 5

which has seen a dramatic decline due to spearing and poisoning (Ogutu et al., 2005; Frank et al., 2006). 

LIVING WITH LIONS: THE LAIKIPIA 

PREDATOR PROJECT

In spite of its overwhelming importance in modern 

lion conservation, there has been remarkably little research on lion-human conflict. Laikipia ranchers’ commitment to conservation, including tolerance of predators, 

has made this an ideal laboratory in which to study all 

aspects of carnivore-livestock interactions and to improve 

humans’ ability to live with lions. Working in central and 

northern Laikipia, we have identified over 180 lions and 

radio collared 145 since 1998. 

STATUS OF LIONS IN LAIKIPIA

Communal Lands

Lions are rare on most communal lands of Laikipia. 

Overgrazing by domestic livestock has reduced wild ungulate numbers on group ranches, leaving superabundant 

goats and cattle as the most available and vulnerable prey 

for carnivores. Incoming lions kill livestock and are then 

poisoned. Because lions are so rare on communal lands, our 

community conservation efforts among Laikipia pastoralists have concentrated on reducing losses to spotted hyenas.

Lions avoid community areas, remaining largely within 

the bounds of commercial ranches by day (Figure 1). Note, 

however, that most of our data are obtained from radio 

tracking flights in the early morning; current studies with 

GPS collars will reveal the extent to which some individuals 

may use communal areas at night. The cues used by lions to 

avoid communal areas are not known but may involve low 

density of wild prey, high density of livestock and humans, 

or greater experience of persecution in those areas. 

Use of poison is widespread in Laikipia’s communities. We have records of at least 52 lions being poisoned 

on communal lands since 2003; 22 of these are known to 

have originated on commercial ranches but were poisoned 

when they moved onto community lands. Spotted hyenas 

are in steep decline as well: at least nine were known to 

have been poisoned on one Laikipia group ranch in seven 

months in 2007 (S. Dolrenry, Living With Lions/University of Wisconsin, personal communication, 2007). 

Much poisoning is meant for hyenas, but lions often die 

instead; an entire northern Laikipia pride died this way 

in 2004, and we have had similar reports from eastern 

Laikipia. Poisoning is not restricted to Laikipia: a minimum of 68 lions have been poisoned in the comparably 

sized Amboseli-Tsavo ecosystem since 2001 (Frank et al., 

2006; L. Hazzah, Living with Lions and University of 

Wisconsin-Madison, unpublished data). Although shooting can target specific problem animals, poison is indiscriminate and often removes whole prides at once, as well 

as large numbers of other predators and scavengers. 

Although strychnine was identified in one case, a more 

common agent is carbofuran, sold as Furadan for use in agriculture. Although banned in Europe and North America 

because of its toxicity to birds and mammals, Furadan is 

cheap and was widely available in Kenya until 2009, when 

adverse publicity in the United States caused the American 

manufacturer to temporarily withdraw it from the Kenya 

market (Gavshon and Magratten, 2009). Vultures have become noticeably uncommon in most of Kenya, including 

Laikipia, the victims of poisoned carcasses put out to kill 

lions, hyenas, and leopards. Bushmeat snaring also affects 

predators (Hofer et al., 1993), and in some parts of Laikipia 

a substantial proportion of hyenas and lions bear snare scars. 

Commercial Ranches

The commercial ranches of Laikipia support one of 

the few stable population of lions in Kenya and the only 

one outside a protected area; most other populations are 

known or thought to be in decline. Most Laikipia lions 

are wary of humans, staying in dense bush by day, and 

we have found that they do not respond reliably to callin census techniques that are useful in protected areas 

(Mills, 1996; Mills et al., 2001). Thus, it is difficult to 

obtain precise numbers, but extrapolating from known 

pride sizes and home ranges, we estimate a stable population of 200–250 of all ages, representing a density of 

6–7 adult and subadult lions/100 km2

 on the commercial 

ranches. In comparison to ecologically comparable bush 

ecosystems lacking livestock, densities in Kruger National 

Park varied between 3.3 and 9.6 adult and subadult lions/100 km2

 (Mills, 1995). Creel and Creel (1997) estimated the lion density of the hunted population in the 

Selous Game Reserve to be between 8 and 13 adults/100 

km2, and Yamazaki (1996) estimated densities of between 

5 and 6 adult and subadult pride females/100 km2, or 

12–13 lions/100 km2 if calculated for all pride members, 

in a hunted population in Zambia. In Kenya, Patterson et 

al (2004) estimated about 4 adult females/100 km2 in the 

bush country surrounding Tsavo National Park. 

Although wild ungulates in Laikipia are outnumbered 10 to 1 by livestock (N. Georgiadis, N. Olwero, and 
7 6 

• smithsonian contributions to zoology

FIGURE 1. Land use map of Laikipia showing that lions remain on commercial ranches by day, avoiding communal pastoralist and settlement areas. Each dot represents an aerial radio tracking fix obtained for an individual lion (n = 3,658 fixes from 136 lions, 1998–2007). The fixes in the northeast are of two female lions collared on Il Ngwesi, one of the few group ranches which had resident lions; they have since been extirpated by poisoning (R. Moller, Lewa Wildlife Conservancy, personal communication).
number 632 • 7 7

G. Ojwang’, unpublished report, “Numbers and Distributions of Large Herbivores in Laikipia District, Leroghi 

and Lewa Conservancy,” presented at Laikipia Wildlife 

Forum, Nanyuki, Kenya, 2003), wild prey form the bulk 

of the lions’ diet. Burchell’s zebra (Equus burchelli) and 

eland (Taurotragus oryx) have been the primary prey of 

Laikipia lions since 1998 (Living With Lions, unpublished 

data), but buffalo (Syncerus caffer), which were seriously 

reduced in the 2000–2001 drought, may become more important as numbers recover. Because smaller prey such as 

impala are eaten very quickly, their contribution to the 

lions’ diet may be underrepresented in observational data.

COSTS OF LIVING WITH LIONS

In all studies of livestock losses, disease and drought 

account for a much higher proportion of livestock mortality than do predators (Frank, 1998). However, losses to 

predators or human raiders elicit a much stronger emotional response from owners. 

Both commercial ranchers and Mukogodo Masai pastoralists use traditional African livestock husbandry techniques: cattle, sheep, goats, and camels are closely herded 

as they graze by day, and at dusk they are brought back 

into thornbush bomas (corrals or kraals), with people 

living in huts around them. On the commercial ranches, 

Frank (1998) found that lions took 0.51% of cattle and 

0.27% of sheep annually. In 1996, it cost on average 

$360 in lost livestock to support a lion on the commercial ranches of Laikipia, or $1.54/head of cattle. However, 

there was great variation in loss rates among ranches, and 

losses on most ranches have declined with improved husbandry in subsequent years. 

Data from one Laikipia group ranch and one settlement scheme (both communally owned by Mukogodo 

Masai pastoralists) showed losses of 0.69% of their cattle and 1.40% of sheep and goats annually to predators, 

largely spotted hyenas. This result may be compared to 

commercial ranches adjoining Tsavo, which lost 2.2% 

of cattle to lions annually (Patterson et al., 2004). In one 

communal area of Zimbabwe, 1.2% of cattle and 3.4% 

of shoats were taken by predators (calculated from Butler, 

2000). On the Mbirikani Group Ranch in Kenyan Masailand, predators took 2.28% of the livestock herd annually, 

but lions accounted for only 4.4% of those losses, whereas 

spotted hyenas accounted for 62% (Maclennan et al., 

2009). However, lions have been reduced to only about 

1.5/100 km2

 on Mbirikani (Frank et al., 2006; Maclennan 

et al., 2009), a quarter of the density in Laikipia, and the 

relatively high livestock losses are due to poor herding, as 

most incidents involve stock left out in the bush at night.

DEPREDATION CIRCUMSTANCES

The great majority of lion depredation occurs at night 

(Frank, 1998; Hemson, 2003; Ogada et al., 2003). In Kenya, lions most frequently take cattle simply by approaching a boma, causing the cattle inside to panic. If the boma 

is not sufficiently strong or if it has weak points (most 

often the “gate,” which may be just a bush pulled into the 

opening), the cattle stampede out of the boma and may 

then be taken by the lions or by hyenas; rounding them up 

often takes several men and vehicles most of the next day. 

Aside from the actual loss of cattle killed, the stress may 

cause loss of weight and, hence, profit. Depending on the 

structure of the boma (see below), some lions may learn 

to leap over the wall, particularly when taking small stock 

from stone or wicker bomas.

Less commonly, lions take stock by day, which seems 

to be more opportunistic than taking them from bomas at 

night and usually occurs when a herd wanders into lions 

sleeping in the bush. Most ranchers consider this loss to 

be simply bad luck and do not hunt down the responsible 

lions. On one ranch that stopped all lion shooting, however, lions learned that they could take stock by day with 

impunity, and losses rose to 79 cattle in one year (M. Dyer, 

Borana Ranch, personal communication, 2005).

Data from Laikipia (Frank et al., 2005; Woodroffe 

and Frank, 2005) and from the Tsavo region (Patterson et 

al., 2004) support ranchers’ and pastoralists’ reports that 

livestock losses are higher during rainy periods. We saw 

few losses to predators during a severe multiyear drought, 

but losses skyrocketed when the rains finally came, and 

many lions were shot in response. We speculate that listless 

wild prey and ready availability of carcasses during dry 

periods provide easy meals and that lions are likely to turn 

to livestock when abundant grass makes wildlife harder to 

catch because they are well nourished, alert, and energetic. 

In the Makgadikgadi Pan, Botswana, and Ngorongoro 

Conservation Area, Tanzania (Hemson, 2003; Ikanda and 

Packer, 2008), clear seasonal trends in livestock predation 

were recorded. These were related to wild prey availability: stock raiding decreased when migratory wild prey was 

present in large numbers and increased when migrant zebra and wildebeest moved to other areas. 

In livestock areas, conservative grazing practices and 

effective control of poaching not only promote wildlife but also probably help minimize depredation losses. 
7 8 • smithsonian contributions to zoology

Although more data are needed, in Laikipia it seems clear 

that depredation is strongly affected by the availability of 

wild prey (Figure 2); lions are more likely to take livestock 

when other ungulates are in short supply. Thus, when livestock production is coupled with wildlife for ecotourism 

or sport hunting, losses to predators are reduced. Such 

“mixed systems” also maximize profit from land that is 

unsuitable for agriculture (Mizutani et al., 2005; NortonGriffiths et al., 2006). 

LETHAL CONTROL

Although Laikipia ranchers are remarkably tolerant of 

predators and willing to absorb a certain amount of loss, 

they do shoot persistent stock raiders, usually by tracking 

lions from a kill or by “sitting up,” waiting for them to return to the carcass of a cow killed the night before. This is 

highly selective; “innocent” lions are rarely shot. Between 

1998 and 2002, an average of 19.4% of the adult population was shot annually, amounting to 30–40 lions per year, 

divided equally among males and females (Frank, 1998; 

Woodroffe and Frank, 2005). Although this seems a high 

mortality rate, lion numbers are capable of rapid recovery; 

cub survival as high as 78.5% (Hunter et al., 2007) and annual population growth rates of 30%–80% have been documented (Maddock et al., 1996; Kissui and Packer, 2004). 

In Laikipia, cub survival is high, and the only emaciated 

lions we have seen have been very old, solitary individuals. 

Importantly, lions originally collared in association 

with livestock kills were four times more likely to be shot 

in response to subsequent livestock damage than were lions 

collared on wildlife kills (12.9% versus 49.0%), strongly 

supporting ranchers’ contention that certain individuals 

or prides are chronic livestock killers and others are not. 

More generally, ranches with good livestock husbandry 

rarely lose stock and rarely shoot lions, whereas both livestock and lions are killed at higher rates on ranches with 

poor practices. 

Given that most Laikipia lions move over several 

ranches (which average 132 km2

 in size), a single ranch 

that kills many lions serves as a local sink, draining lions 

from a much larger area (Woodroffe and Frank, 2005). In 

a recent example, a pride known since 1998 had a home 

range centered on Ranches C, D, and E, which have excellent cattle husbandry and minimal depredation problems 

(Figure 3). Only 2.7% of all aerial fixes (n = 300) of the 

four collared females from this pride were on Ranch F, 

which does not use bomas for its cattle or tolerate wildlife, yet 9 of the 10 pride members, including five breeding 

females, were shot there. As we only learn of deaths in 

collared prides, these represent an unknown fraction of 

the total killed on this single ranch. A small number of the 

25 commercial ranches in Laikipia account for the great 

majority of lion deaths in the region. Thus, even if a community of landowners wants to support predators, a few 

with poor husbandry and no interest in conservation can 

jeopardize the population of an entire region.

Because of the high mortality rate of stock-killing females, those not known to take livestock had four times 

higher cub production (0.98 cub/female/year versus 0.23 

cub/female/year) and 2.7 times higher cub survival than 

did stock killers. Moreover, this population is producing 

a skewed cub sex ratio, 69:31 favoring males (Woodroffe 

and Frank, 2005). It is not known whether this is an effect 

of excess male mortality or other ecological factors.

SOLUTIONS

Ogada et al. (2003) assessed the efficacy of traditional 

African methods of livestock husbandry in protecting 

0

1

2

3

4

5

0 0.5 1 1.5 2

Ratio of Wildlife:Livestock Biomass

Percent Lost

to Predators

CATTLE

SHOATS

FIGURE 2. Relationship between livestock loss 

rate in 1995–1996 and wildlife density on individual ranches in Laikipia. (Wildlife data are from 

N. Georgiadis, N. Olwero, and G. Ojwang’, unpublished report, “Numbers and Distributions of 

Large Herbivores in Laikipia District, Leroghi and 

Lewa Conservancy,” presented at Laikipia Wildlife Forum, Nanyuki, Kenya, 2003.)
number 632 • 7 9

livestock from predators on commercial ranches. These 

practices evolved in response to the twin threats of both 

predators and livestock-stealing humans and are probably 

similar to those used when pastoralism was first introduced in Africa over 9,000 years ago (Marshall, 2000). 

Not surprisingly, Ogada et al. (2003) found that ranchers 

kill significantly more predators on ranches where predators kill more livestock. Thus, implementation of any practice that reduces the vulnerability of livestock is critically 

important for reducing retaliatory killing of predators. 

Seventy-five percent of depredation of cattle, sheep, 

and goats took place at night, and lions were responsible 

for over 75% of the total; depredation on Laikipia ranches 

occurs largely at the boma. Well-built bomas effectively 

constrain cattle and keep predators out. Bomas in Laikipia are made from native thornbush (Figure 4), stone walls, 

wooden posts, or wire mesh; of these, thick strong thornbush was most effective at keeping lions out and panicked 

cattle in. However, thornbush deteriorates over time, and 

many ranchers are reluctant to keep cutting mature trees to 

rebuild them. Stone is an excellent building material if there 

is a fence on top to prevent lions from leaping onto the wall 

and into the boma. Although most expensive to build, stone 

bomas last essentially forever and need little maintenance. 

Thornbush bomas are most effective if divided into 

inner “rooms” that make it harder for cattle to reach the 

main gate, and the gate must be very strong, preferably 

made from lumber. The normal practice of using a tree 

or bush as a gate is ineffective, as it does not contain panicked cattle and allows hyenas to enter.

Wire mesh is a poor barrier if not well supported, 

but John Harris on Suyian Ranch and Giles Prettejohn of 

FIGURE 3. Central Laikipia, showing the home range of female lions LF5, LF69, LF85, LF110, and their 

seven pride mates. These animals were not known to take livestock on Ranches A–E, but nine were shot 

on Ranch F, which does not use bomas for its cattle.
8 0 • smithsonian contributions to zoology

Ol Pejeta Conservancy have developed modular, moveable bomas made of mesh or old fencing wire welded into 

panels of interconnecting steel frames that are highly resistant to predators and easily transported (Ol Pejeta Conservancy, 2011). Not only are they predator proof, but 

they also allow ranchers to move bomas every few days to 

improve pastures. Although expensive (cattle bomas cost 

on the order of US$2,000 each), these bomas have proven 

100% effective in preventing lion depredation by night 

and are being rapidly adopted by other Laikipia ranches. 

We found that lions are reluctant to approach bomas 

that are located in close proximity to large numbers of 

people. However, for security and environmental reasons, some ranches do not allow herders to have their 

families at the bomas. Of course, in traditional societies 

bomas usually have large numbers of people and dogs. 

Dogs are also highly effective deterrents by both day and 

night (Woodroffe et al., 2007); they do not chase predators but warn herders, who then chase the lions. Again, 

however, some ranches do not allow dogs, as herders may 

use them for hunting wild game. Dogs can carry lethal 

carnivore diseases, but they are such an effective deterrent that vaccinated dogs are an essential component of 

livestock husbandry. Bright lights or noise-making devices 

like a shotgun or thunder flash (firecracker) also effectively 

repel loitering lions.

PROBLEM ANIMAL CONTROL

In the absence of totally reliable methods for protecting 

livestock from lions, some depredation is inevitable, and 

some lions will become habitual livestock killers. There is 

as yet no alternative to lethal removal of chronic offenders 

through Problem Animal Control (PAC). “Translocating” 

problem predators to parks is not a humane alternative, 

as translocated predators often sustain severe tooth and 

claw damage in the trap, are severely persecuted by residents of the same species where released, and usually end 

up being killed, often after causing further problems as 

they try to find their way home (P. Jenkins, Kenya Wildlife 

Service, 1997; Frank et al., 2003). Because of trauma to 

the animals and the low likelihood of survival, we strongly 

recommend against translocation. 

Most commercial ranchers are able to deal with problem lions, but small-scale rural farmers and pastoralists 

usually do not have the means. In Kenya, rural people consistently complain that wildlife authorities do not respond 

FIGURE 4. Thornbush boma used on commercial ranches in Laikipia. Photo by Laurence Frank.
number 632 • 8 1

effectively when people report chronic stock raiders, leading to resentment not only against wildlife, but also government, conservation, and tourism. A well-trained and 

reliable PAC team, able to respond efficiently, effectively, 

and rapidly, is an essential element of large-carnivore management in livestock areas. In their absence, rural people 

have little alternative besides such indiscriminate methods as poisoning, which probably poses the most serious 

threat to predator populations. Problem Animal Control 

teams should be trained not only in the reliable identification of problem animals and humane removal but also, 

most importantly, in educating rural people in livestock 

husbandry techniques that better protect stock. 

CONCLUSIONS

Large carnivores are among the most difficult animals 

to conserve because their feeding habits inevitably bring 

them into conflict with humans. At the same time, their 

wide ranging movements and need for substantial prey 

populations require very large areas, and thus, only the 

biggest protected or well-managed landscapes currently 

provide long-term security for viable populations; only six 

such areas currently exist in Africa.

Conflict with humans over livestock depredation is 

the single most important factor causing the decline in African lion populations. With growing numbers of people 

and livestock throughout the continent, conflict mitigation 

must be implemented on a wide scale if lions are not to become restricted to a few very large protected areas or wellmanaged hunting blocks. Traditional methods of livestock 

husbandry are remarkably effective at minimizing conflict, 

but these are being lost to modernization. Building good 

bomas and conscientiously tending livestock require time 

and effort at a time when poison is readily available and 

spearing lions is the only traditional test of manhood left 

for young warriors. As pastoralists increasingly engage in 

a cash economy, they have lost their tolerance of predators 

and are likely to continue eliminating lions unless lions 

bring in financial benefits that outweigh costs. 

In the 95% of Kenya’s rangeland that does not support tourism (Norton-Griffiths et al., 2006), wildlife are 

only an expensive nuisance to the people who share their 

habitat: wild animals kill livestock, destroy crops, and occasionally kill and injure humans. Few rural people benefit 

from tourism revenues, but in combination with poor law 

enforcement, national policy, which prohibits legal and 

sustainable consumptive utilization, ensures that the only 

value for wildlife is as bushmeat, snared or hunted illegally 

and unsustainably. Thus, current policy in Kenya, essentially determined by foreign animal rights groups with 

neither interest nor expertise in conservation, effectively 

makes rural Kenyans resent and eliminate wildlife.

In 2007, tourist hunters in Botswana paid US$140,000 

to shoot a lion, $100,000 of which went to the community in which it was taken (J. Rann, Rann Safaris, personal 

communication, 2006). Given Kenya’s romantic history 

of hunting (Herne, 1999), today’s wealthy trophy hunters would probably pay substantially more than that for 

a Kenyan lion. Laikipia is currently the only nonprotected 

area in Kenya with a lion population large and stable 

enough to permit a carefully regulated offtake of several 

old males per year (Whitman et al., 2004). If equitably distributed and combined with help in preventing depredation 

losses, this potential income could significantly improve 

community perceptions of predators and the potential role 

of wildlife in raising rural living standards. However, many 

argue that Kenya’s pervasive corruption would make it 

impossible to sustainably manage and regulate a lucrative 

hunting industry except on privately owned lands.

People will either learn to live with lions or we will 

lose them. We have shown that ancient livestock husbandry methods effectively protect livestock from lions, 

but spears, bullets, and poison are always cheaper and 

easier solutions than managing livestock, lions, or growing rural human populations. Thus, rural people must perceive lions and other wildlife as valuable commodities if 

they are to accept the burden of living with them: the benefits of wildlife must outweigh the costs. Successful lion 

conservation must combine effective management of risks 

with development of viable wildlife-based economies that 

improve the lives of rural Africans. Traditional peoples 

and wildlife managers already have most of the techniques 

necessary to manage depredation, but the greater challenge of managing ecologically sustainable rural development lies in the realm of policy, social science, and politics.

Acknowledgments

A great many people have contributed to our work 

on Laikipia’s predators, including Ian Ross and Jonathan 

Burchell, who gave their lives. None of this would have 

been possible without the enthusiastic cooperation of Laikipia ranchers, who deserve full credit for creating a remarkable conservation success. Stephen Ekwanga, Alayne 

Cotterill, and Rosie Woodroffe participated in field work 

for years. Bruce Patterson contributed many helpful comments that improved the manuscript. Many individuals 

and organizations have contributed financially, including 
8 2 • smithsonian contributions to zoology

the Wildlife Conservation Society, Panthera, the Wildlife 

Conservation Network, the Banovich Wildscapes Foundation, the Philadelphia Zoo, and the Potreronuevo Fund.

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Attitudes Toward Predators and Options for 

Their Conservation in the Ewaso Ecosystem

Stephanie S. Romañach, Peter A. Lindsey, 

and Rosie Woodroffe

ABSTRACT. The act of killing predators over livestock predation has been the principal 

cause of declining predator populations throughout Africa. Finding solutions for the 

coexistence of people with predators in the Ewaso region is of great importance for the 

local tourism industry and for the persistence of Kenya’s wildlife outside of protected 

areas. We report results of an assessment of attitudes toward predators by landholders 

and land users in the Ewaso region and their tolerance of livestock losses, relative to 

socioeconomic factors. Although prospects are good for predator conservation on largescale private ranches, the future of predators on communally owned properties remains 

uncertain. Prospects would be improved not only by finding solutions to reduce livestock 

depredation but also by ensuring that individuals tolerating losses from predators receive 

benefits from having predators on their land. Additionally, policy changes are needed to 

allow landholders to capitalize on benefits from wildlife, which would encourage their 

participation in wildlife conservation.

PREDATORS AND PEOPLE

Predators are among the hardest animals to conserve because they range 

widely, often outside of protected areas and onto land not used for wildlife 

conservation. Predators can become a liability to landholders, who incur financial losses when livestock are killed. As a result, people kill predators, often 

preemptively. For example, lions (Panthera leo) are killed in Kenya because of 

conflict over livestock (Woodroffe and Frank, 2005). Predators, including wild 

dogs (Lycaon pictus) and cheetahs (Acinonyx jubatus), are suppressed on some 

wildlife ranches in southern Africa where they kill antelopes, which are valuable 

for ecotourism and trophy hunting (Sillero-Zubiri and Laurenson, 2001; Ogada 

et al., 2003; Marker et al., 2003; Lindsey et al., 2005). Persecution of predators 

has resulted in serious declines in the population sizes and distributions of many 

predator species (Woodroffe and Ginsberg, 1998), and it remains the most important source of their mortality. 

Stephanie S. Romañach, African Wildlife Conservation Fund, 15064 NW 57 Street, Doral, Florida 

33178, USA. Peter A. Lindsey, Mammal Research Institute, University of Pretoria, 0002, Pretoria, South Africa. Rosie Woodroffe, Institute of 

Zoology, Zoological Society of London, Regent’s 

Park, London NW1 4RY, UK. Correspondence: 

S. Romañach, stephanie.romanach@gmail.com.

Manuscript received 7 January 2009; accepted 

18 May 2010.
8 6 • smithsonian contributions to zoology

Finding ways to increase tolerance of predators is 

critical for their survival (Hackel, 1999; Woodroffe, 

2000). Strenuous attempts have been made to improve 

local citizens’ attitudes toward predators in rural Africa 

through community-based conservation projects, education campaigns, and outreach programs (e.g., Painted 

Dog Research Project, Zimbabwe; Cheetah Conservation 

Fund, Namibia). Some successful programs have focused 

on generating income from predators through ecotourism 

or sport hunting (Adams and Hulme, 2001). For such programs to work, financial returns from wildlife must reach 

individuals who bear the costs of living with wildlife and 

must be adequate to offset not only direct costs but also 

opportunity costs. This prerequisite is difficult to achieve 

because the benefits any individual can receive from predators are often diluted by the number of community members among whom benefits must be shared (Walpole and 

Thouless, 2005), whereas costs are accrued individually. 

Accordingly, predator population sinks can form outside 

of protected areas, threatening predators both outside and 

inside of protected areas.

CONSERVATION CHALLENGES IN KENYA

Kenya’s national parks and reserves house 35% of the 

country’s wildlife; the remaining populations are found 

on land that is privately or communally owned or held in 

trust by local government (Grunblatt et al., 1996; Western 

et al., 2009). The persistence of wildlife is important for 

the persistence of Kenya’s primary industry, tourism (Ottichilo et al., 2000). Human populations continue to grow, 

resulting in the movement of people from densely populated areas to Kenya’s rangelands to practice subsistence 

livestock farming, where they overlap with the majority of 

the country’s wildlife (Ottichilo et al., 2000). Over the last 

decade Kenya’s human population has grown nearly 4% 

per year (Shikwati, 2004). Rising human populations have 

driven habitat conversion to accommodate subsistence 

pastoralism and agriculture, resulting in increased conflict 

between people and wildlife and population declines of 

many wildlife species (Awere, 1996). 

EWASO ECOSYSTEM

Kenya’s Ewaso ecosystem is an area dominated by 

livestock, but it retains the potential to host viable wildlife 

populations. With six large carnivore species indigenous 

to the area (cheetahs, lions, leopards [Panthera pardus], 

spotted and striped hyenas [Crocuta crocuta and Hyaena 

hyaena], and the endangered African wild dog), reducing 

conflict between people and predators is essential for the 

persistence of wildlife in this area. This need is particularly 

true for predators, which range across a mosaic of land 

uses, including subsistence pastoralism, patchy small-scale 

agriculture, large-scale commercial livestock ranches, and 

wildlife conservation and tourism areas. In this chapter we 

report results of an assessment of attitudes toward predators by local landholders and land users, and their tolerance of livestock losses, relative to socioeconomic factors. 

Our goals were to examine the challenges facing people 

and predators and to identify ways to improve their potential for coexistence in the region.

Interviews

We interviewed 416 community members and commercial livestock ranchers in the Ewaso ecosystem in 

2004–2005 to assess their attitudes toward predators, 

tolerance for livestock kills, and options for predator 

conservation (Figure 1). We conducted interviews with 

23 commercial ranch managers or owners. Nine local assistants interviewed 393 rural community members who 

owned livestock, including people from 10 tribes (Maasai and Samburu are the predominant tribes in the area, 

although people from other tribes, including Borana, 

Kalenjin, Kikuyu, Merian, Pokot, Rendile, Somali, and 

Turkana, were also interviewed).

Before each interview, we assessed the respondents’ 

familiarity with predators (e.g., hunting behavior) using 

an illustrated card with color drawings of each predator 

(for interview details, see Romañach et al., 2007). The 

questionnaires were designed to assess respondents’ attitudes toward and tolerance for livestock losses from cheetahs, jackals (Canis mesomelas), leopards, lions, spotted 

hyenas, striped hyenas, and wild dogs. We quantified respondents’ tolerance for predator attacks on livestock by 

recording how many livestock (measured as the number of 

cattle or the combined number of sheep and goats) respondents were willing to lose before trying to kill the predator 

responsible. Older interviewees were also asked what their 

tolerance for livestock losses to predators was 20 years 

ago. Interviewers asked respondents whether they would 

want predators on their properties if they were to start 

tourism operations. We asked whether people thought trophy hunting should be legalized again in Kenya, an industry banned in Kenya in 1977 because of poor control and 

coordination and large-scale poaching (Outoma, 2004). 

Interviewees were also asked, in an open-ended question, 
number 632 • 8 7

to provide suggestions for methods to promote coexistence between people and predators.

During each interview, we recorded information on 

characteristics of the property where the interviewee was 

living at the time of the interview: land use type (e.g., 

group ranch, commercial ranch); if the property had wildlife tourism as a form of income at the time of the interview; primary source of property income; benefits received 

from wildlife (e.g., dams built, communications radios purchased); whether the property had a wildlife conservation 

area set aside (where livestock did not graze); whether there 

was an immediate (within two years) plan to start a wildlife 

conservation area if there was not one at the time of the interview; presence of each predator on the property; and for 

commercial ranchers, whether ranch activities served as the 

primary source of income for the ranch owner, compared 

to having a subsidized income, for example. We recorded 

personal information about each interviewee: age, gender, 

number of years of formal education, primary source of 

household income, and the number of their livestock killed 

by each predator in the last year. 

FINDINGS AND IMPLICATIONS FOR 

PREDATOR CONSERVATION

Attitudes and Tolerance

More than 90% of commercial ranchers were in favor 

of having predators on their land (Figure 2). We found 

that commercial ranchers want to have all predators in 

question on their land, regardless of the rancher’s personal 

characteristics (e.g., age, education) or the ranch characteristics (e.g., primary source of property income, presence 

of tourism). Commercial ranchers were also more tolerant 

of livestock losses from all predators compared to the tolerance of community members. 

FIGURE 1. Map of locations for commercial rancher interviews (×) and community member interviews (•). (Source: Romañach et al., 2007.)
8 8 • smithsonian contributions to zoology

Community members were less enthused about the 

presence of predators. For example, only about 20% 

wanted to have spotted hyenas on their land, and about 

40% wanted to have other predator species. Community 

members were more likely to want predators on their land 

if benefits from having predators reached the individual, 

not if benefits were only given to the community as a whole 

(e.g., dams built). We found that people with more education were more likely not to want predators on their communal land; however, the higher the level of education, the 

higher interviewees’ stated level of tolerance for predators 

that were on their land. These findings suggest the value of 

education in conservation of biodiversity in that although 

people might not like predators, higher levels of education 

lead to greater tolerance. Another factor contributing to 

interviewees’ increased tolerance was if their community 

was planning to start a wildlife conservancy in the next 

two years, as opposed to if they already had a conservancy. 

Our results also show the danger of false hope brought by 

the promise of tourism when planning to start a wildlife 

conservancy if the plan does not come to fruition and also 

for individuals who are part of a conservancy but who 

do not personally receive benefits. Distribution of benefits 

so that they reach individuals, particularly reaching those 

that tolerate livestock losses, will be important to the future of predators on communal land.

Over half (53%) of all interviewees reported having lost 

livestock to at least one predator in the year prior to the interview. Lions, leopards, and spotted hyenas were reported 

to be the most troublesome killers of livestock (Table 1). 

Lions are thought to be the most problematic of predators 

and therefore may have the most uncertain future. 

Some respondents said they would not kill predators in 

response to livestock attacks, no matter how many head of 

stock were lost; this “no-kill” policy was largely for stated 

tolerance of cheetahs and wild dogs. Conversely, other respondents stated they would kill predators on sight, even 

if no attack had occurred. Of commercial ranchers, only 

one rancher held a “kill-on-sight” policy, and it was for 

spotted hyenas. Of community members, depending on 

predator species, 77%–88% of interviewees stated having a kill-on-sight policy. Community members were least 

likely to want to kill wild dogs, which we attributed partially to the belief of Samburu respondents that wild dogs 

are sacred and partially to the perception that a successful 

wild dog research and conservation project in the region 

has provided jobs. This sentiment was particularly notable in areas where full-time field scouts are employed to 

FIGURE 2. Percentages (95% confidence intervals) of commercial ranchers and community members who wanted each predator on their property. (Source: Romañach et al., 2007.)

0

10

20

30

40

50

60

70

80

90

100

Cheetah Jackal Leopard Lion Spotted

hyena

Striped

hyena

Wild dog

Interviewees with positive attitudes (%)

Commercial ranchers Community members
number 632 • 8 9

monitor movement of wild dogs on their community land 

(e.g., Kijabe Group Ranch). The Samburu-Laikipia Wild 

Dog Project (SLWDP; headed by Rosie Woodroffe) had 

been running in the region for three years (since 2001) at 

the time of the interviews.

Despite the high number of community members 

reporting a kill-on-sight policy, all native large predator 

species are at least present over large areas of communal 

land in the region. Lions are not common on communal 

lands, however, and are heavily persecuted (L. G. Frank, 

personal communication, University of California, Berkeley, 2004). Seventy-seven percent of respondents stated a 

kill-on-sight policy for wild dogs, but none reported having killed wild dogs in the previous year, and only 15% 

reported having killed any predators in the previous year. 

Additional data from the study area show that at the time 

of our interviews, only 2 of 33 (6%) radio-collared wild 

dogs had been killed by humans (R. Woodroffe, unpublished data, 2005), and the population size of wild dogs 

continues to increase. This increase is particularly striking given that wild dogs primarily occur on community 

land in the region (Woodroffe et al., 2004), where they are 

generally unwelcome. Two of the goals of the SLWDP are 

to help reduce livestock depredation by wild dogs and to 

encourage wild dog ecotourism for income. SLWDP has 

been successful in meeting its goals and, as a result, has 

been increasingly gaining the support of locals for the protection of wild dogs.

Thirty-one percent of all interviewees (commercial 

and community) who owned or managed livestock for the 

last 20 years reported increased tolerance for predators 

today, and 6% reported being less tolerant of predators 

today. Commercial ranchers are prepared to lose between 

four and eight stock (cattle or sheep/goats) before eliminating the responsible predator, which is about twice the 

tolerance they said they had 20 years ago. Community 

members were less tolerant and were unwilling to lose 

more than one head of stock at any time before attempting to eliminate the predator.

Lethal Control

We found that people (commercial ranchers and 

community members) who lost livestock to predators in 

the last year were much more likely to have killed them 

compared to people who had not experienced livestock 

losses. This finding emphasizes the importance of improving methods of livestock protection (e.g., stronger corrals) 

to decrease retribution killings and thus to increase coexistence of people with predators through greater tolerance. Commercial ranchers were more likely to have killed 

predators in the last year if the income generated by their 

ranch was the ranch owner’s primary source of income 

(compared to ranches that have subsidized incomes, for 

example, through foreign investment). This finding highlights the financial difficulty of tolerating livestock losses 

to predators and the difficultly of coexistence between 

people and predators even on large-scale ranching areas.

Interviewees were asked how they would kill predators responsible for livestock attacks. Commercial 

TABLE 1. Survey results of commercial ranchers and community members asked to rank predators as worst attacker of livestock and 

owners who lost livestock to predators in the preceding year. Some respondents did not give a ranking for all predators, and some had 

livestock killed by multiple predators; thus rows will not necessarily total 100%.

 Spotted Striped Wild

Measurement Cheetah Jackal Leopard Lion hyena hyena dog

Commercial ranchers (%)

Ranking predator worst for cattle attacks 0 0 0 70 30 0 0

Owning cattle killed by predator 0 0 17 78 22 0 0

Ranking predator worst for sheep/goat attacks 9 0 35 9 48 0 0

Owning sheep/goat killed by predator 17 4 48 26 43 0 4

Community members (%)

Ranking predator worst for cattle attacks 2 5 3 71 9 1 1

Owning cattle killed by predator 1 0.3 24 21 11 1 0.3

Ranking predator worst for sheep/goat attacks 7 6 27 2 35 6 9

Owning sheep/goat killed by predator 14 21 29 5 31 8 4
9 0 • smithsonian contributions to zoology

ranchers and community members differed in their preferred means of lethal control. Ranchers who would kill 

predators would largely shoot them; however, when considering lethal control of spotted hyenas, one respondent 

would gas them inside of their dens, and another would 

use general carcass poisoning. No commercial ranchers 

would set wire snares to kill predators. 

Of community members, almost half (47%) would 

use carcass poisoning to kill predators, and 9% would set 

wire snares. To kill cheetahs, 60% of respondents would 

use spears or knives, 13% would use poisoned arrows, 

8% would trap the cheetah with no mention what they 

would do with the animal once trapped, 6% said they 

would call the government authority (Kenya Wildlife Service) to take care of the problem, 4% would use clubs, 4% 

would shoot the animal, and 1% would hunt predators 

with domestic dogs. 

Communal Land

Communal land makes up 60% of Laikipia District 

and nearly all of the surrounding districts. It is therefore 

the land use type of greatest concern—and potential—for 

conservation. We found that community members living 

on group ranches (i.e., those with title deeds) were most 

positive toward predators compared to those on government land and squatters. The Laikipia Wildlife Forum 

(LWF), African Wildlife Foundation, other nongovernmental organizations, and commercial landowners have 

been helping communities to gain title deeds to their land 

in the region. Promoting group landownership could have 

positive implications for predator conservation over the 

large areas in Africa. Property rights provide residents 

with incentives for the sustainable use of natural resources, 

given that they will have access to those resources in future. Group ranches have the benefit of being a size large 

enough for holding suitable populations of wildlife species 

compared to smaller-sized, individually owned land. Such 

a model is preferable to the subdivision of formerly communally owned land into small privately owned parcels. 

Such land tenure shifts in parts of Kenya have been demonstrated to be associated with declining diversities and 

densities of wildlife (Norton-Griffiths, 2007). 

Community members who lost livestock to predators 

in the last year killed more predators than those whose 

livestock were not attacked. Of community members interviewed, Masais and Samburus were the most positive 

toward predators. However, Samburus had a relatively 

high incidence of killing predators in response to livestock 

attacks. We found no association between livestock lost 

to a particular predator and likelihood for that specific 

predator species to be killed. This finding suggests that 

if people lost livestock to any predator, they were more 

likely to kill all predators in retaliation.

Commercial Ranch Land

Attitudes were most positive and tolerance was greatest among commercial ranchers. These positive attitudes 

are noteworthy given that 30% of commercial ranchers 

receive no benefits from tourism and no income subsidies 

and yet are interested in conserving predators that can 

have negative effects on their livestock ranching activities. 

Tolerance for predators among commercial ranchers has 

improved compared to their tolerance 20 years ago. Most 

respondents attributed their increased tolerance to realizing the importance of wildlife conservation (citing education from LWF) and, for some, experiencing economic 

return from wildlife. Commercial ranchers in the Ewaso 

region hold more positive attitudes than southern African 

ranchers (Marker et al., 2003; Lindsey et al., 2005), indicating great promise for the Ewaso ecosystem. Although 

prospects for predator conservation are greatest on commercial ranches, ranches cover only 40% of the Laikipia 

District and very little to none of the neighboring districts; 

therefore, great efforts are needed to promote wildlife conservation participation on communal land for wildlife to 

persist across the mixed land uses.

OPTIONS TO PROMOTE COEXISTENCE 

BETWEEN PEOPLE AND PREDATORS

When asked as an open-ended question about how attitudes toward predators could be improved, the most common suggestion from all respondents (commercial ranchers 

and community members) was to give value to predators, 

with most citing using ecotourism and trophy hunting. 

Tourism is the world’s fastest growing industry, with 

ecotourism its fastest growing sector (Gössling, 2000), 

and developing nations are increasingly popular destinations (Goodwin, 1996; Gössling, 1999). With the number of successful tourism operations in the Ewaso region 

growing and with visitors increasingly interested in both 

wildlife and human cultures, tourism should continue to 

improve attitudes toward predators. There are a number of successful tourism operations in the region, several of which are on community land (e.g., Il Ngwesi 

Lodge, Sarara Camp, Koija Starbeds), with more lodges 

in the development stages. The Ewaso ecosystem houses 
number 632 • 9 1

globally important populations of endangered wild dogs 

and Grevy’s zebra (Equus grevyi), as well as cheetahs and 

lions, all of which are of interest to tourists. 

The development of tourism operations on communal land in the region would be facilitated through the 

development of public-private-community partnerships, 

whereby commercial ranchers form partnerships with adjacent communities in the development of an ecotourism 

enterprise. A community neighboring Savé Valley Conservancy in Zimbabwe, for example, has voluntarily removed 

all livestock and incorporated 40 km2 of their land into the 

conservancy (Lindsey et al., 2008). This community will 

receive full financial gains from wildlife as well as become 

part of wildlife management decision making.

When interviewees were asked directly about their attitudes toward legalizing trophy hunting in Kenya, older 

community members were generally in favor of trophy 

hunting, mentioning benefits brought by employment. 

Younger community members had mixed views, with negativity stemming from the fear that trophy hunters would 

kill all wildlife and leave nothing to show visitors. Hunting was banned in Kenya in 1977, so most Kenyans lack 

firsthand experience of tourist hunting, and many are concerned that it would pose a threat to the country’s wildlife 

populations (see Box 1). However, most wildlife species 

in Kenya have declined significantly since hunting was 

banned, suggesting that the ban has not been an effective 

conservation measure (Western et al., 2009). In fact, the 

hunting ban may have exacerbated the decline by removing a key option for landholders to derive financial benefits 

from conserving wildlife (Norton-Griffiths, 2007). One 

of the major challenges with the management of trophy 

hunting (and ecotourism) is ensuring that a sufficiently 

large proportion of revenues accrue to communities. 

Current government policy in Kenya is such that its citizens would not be able to adequately benefit from trophy 

hunting as a wildlife land use because wildlife belongs to 

the state. If trophy hunting were legalized, changes would 

also be required to grant full user rights over wildlife to 

private landowners and communities. In Zimbabwe, for 

example, positive effects for wildlife conservation were 

seen as a result of the Communal Areas Management Programme For Indigenous Resources (CAMPFIRE), which resulted in devolution of user rights over wildlife from central 

government to local government (Child, 2000). However, 

most practitioners believe that CAMPFIRE would be more 

successful still if user rights over wildlife were devolved further to the community level. Without the devolution of user 

rights to wildlife, locals cannot benefit adequately from 

wildlife and therefore lack incentive to protect it.

Predators can pose a cost to people living with or near 

them, so to conserve these species, conservationists must 

BOX 1. Potential for trophy hunting in Kenya.

Reinstating trophy hunting in Kenya would provide incentives for wildlife conservation over a wide area. Hunting and ecotourism are largely complementary land uses; trophy hunting is often conducted in areas not desired by ecotourists, such as areas 

with low wildlife densities and where people and livestock predominate. Trophy hunting has been successful in creating incentives 

for conservation on communal lands in other parts of Africa, particularly in Namibia, Botswana, and Zimbabwe. In Namibia, following legislative changes that made provision for communities to obtain user rights over wildlife, income from trophy hunting has 

been the primary stimulant for the development of a network of communal conservancies now covering more than 126,000 km2

in which wildlife populations are recovering rapidly (Jones and Weaver, 2008; M. Lamprechts, Trophy Hunting in Namibia from 

the 1960s to the Present Day, http://www.huntersnamibia.com/Lamprecht_Paper.doc). The Namibian conservancy example 

clearly illustrates the importance of allocating user rights over wildlife to communities and allowing the option of utilizing wildlife 

consumptively. Properly managed trophy hunting has potential to offset losses of livestock due to depredation and to create incentives for predator conservation. Visiting hunters pay particularly high prices for opportunities to hunt predators. In southern Africa, 

hunters pay US$6,000–32,000 in trophy fees per safari to hunt lions in addition to US$1,300–4,000 per day for a minimum 

of 15–28 days (Booth, 2009). Potential earnings from trophy hunting are particularly high in Kenya because of the presence of 

a number of species and subspecies that do not occur or are rare elsewhere and because the country is viewed with nostalgia 

by many hunters as being the original hunting safari destination (Lindsey et al., 2006). There are problems associated with trophy hunting that prevent the industry from benefiting conservation and local development to its potential. For example, there is 

inadequate regulation of the hunting industry in some areas, and in some cases local communities do not receive an adequate 

proportion of hunting revenues. However, the net impact of trophy hunting on conservation is almost certainly positive through the 

creation of financial incentives for the retention and development of wildlife as a land use, especially in areas where alternative 

income sources such as ecotourism are not viable (Lindsey et al., 2007).
9 2 • smithsonian contributions to zoology

find ways to reduce costs, increase the value of predators, 

or, preferably, both. If benefits from wildlife are increased, 

communities might be encouraged to establish wildlife areas on their land, as has occurred elsewhere in southern 

and eastern Africa. In Zambia, for example, some communities chose to stop agricultural production in favor of 

forming wildlife areas for trophy hunting for the greater 

financial benefit (Lewis and Alpert, 1997). The conversion 

from agriculture- to wildlife-based land use has provided a 

greater amount of habitat, which is critical for the persistence of wide-ranging species. A key means of reducing the 

costs of living with predators is reducing livestock losses, 

for example, through improved livestock husbandry (e.g., 

Treves and Karanth, 2003; Jackson and Wangchuk, 2004; 

Woodroffe et al., 2007) and seasonal stocking strategies 

(Patterson et al., 2004). In another pastoralist area in 

Kenya, Kruuk (1981) suggested that constructing stronger, 

less-penetrable corrals should reduce livestock losses, but 

Ogada et al. (2003) found that corral height and thickness 

had no effect on livestock losses in Laikipia (see Frank, 

2011, this volume, for further discussion on livestock predation). Setting aside conservation areas to conserve wild 

prey may also reduce livestock losses by increasing the 

abundance of available wild prey (Woodroffe et al., 2005). 

Positive attitudes of community members regarding 

predators were generally based on the hope of financial 

gain from wildlife. Our interview results show that community members in the Ewaso region were generally more 

tolerant of predators if they had tourism on their properties and thus were receiving a benefit. Thirty-five percent 

of community respondents had tourism operations on 

their land at the time of our interviews, and an additional 

21% had plans to start a wildlife conservancy in the next 

two years. As a result, a maximum of 56% of respondents 

should be able to earn income from wildlife tourism in 

the near future. If the planned tourism schemes come to 

fruition, community members’ attitudes toward predators should improve. However, previous work has shown 

that when people do not receive expected benefits, their 

attitudes may ultimately become worse (Western, 1994; 

Walpole and Goodwin, 2001; Walpole and Thouless, 

2005). Although we found that community members’ tolerance for wild dogs was greater if some of their household income came from tourism, we found no increased 

tolerance for predators if benefits were shared among the 

community as a whole (e.g., schools built). This finding is 

in keeping with the increasing realization that if benefitsharing schemes are to succeed, benefits must reach individuals (Walpole and Goodwin, 2001; Walpole and 

Thouless, 2005). 

The best prospects for enhancing predator conservation lie in promoting wildlife-based land uses. Options 

for communities to benefit financially from predators and 

other wildlife species would be significantly enhanced if 

government policies were changed such that user rights to 

wildlife were granted and consumptive forms of wildlife 

utilization were permitted. 

Among commercial ranchers, increased tolerance for 

predators compared to 20 years ago is encouraging. Commercial ranchers are considerably wealthier than their 

communal counterparts, and a significant proportion of 

them favor wildlife, including predators. On the commercial ranches in the region that favor wildlife, prospects for 

predator conservation in the Ewaso ecosystem are therefore excellent.

CONCLUSION

Attitudes toward predators have generally improved 

in Laikipia over the last 20 years, but further improvement is still possible and is necessary to ensure the persistence of predators. Kenya has recently been going 

through a process to try to improve its wildlife policies 

for the benefit wildlife and the people living with wildlife. 

Implementation of policies that allow people to maximize 

benefits from the wildlife they live with should advance 

local citizens’ interest in expansion of habitat for wildlife 

conservation.

Acknowledgements

We thank the residents of the Ewaso region who 

participated in our interview; to our assistants who conducted interviews, N. Gitonga, M. Kahindi, D. Lomoe, 

P. Lekitasharan, B. Lengalen, S. Lesowapir, D. Masere, 

J. Musiany, and S. ole Ranah; to LWF; and to the Mpala 

Research Centre and its staff.

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Conservation, 8:91–98. 

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of Populations inside Protected Areas. Science, 280:2126–2128.

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ole Ranah. 2005. Livestock Predation by Endangered African Wild 

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Conserving Wildlife on Private Land: The 

Legal Framework for Landownership and 

New Tools for Land Conservation in Kenya

Nyokabi Gitahi and Kathleen H. Fitzgerald

ABSTRACT. The threat to wildlife conservation in Kenya continues to escalate, largely 

because of an increase in habitat fragmentation, land use change, and human population pressure. To maintain viable populations of native wildlife, lands outside protected 

areas must be conserved through innovative measures. This paper describes the legal 

framework for landownership in Kenya and existing legal mechanisms used to conserve 

wildlife habitat and recommends new tools for land conservation on private land. We 

stress that none of the existing land tenure systems provide an absolute right of use and 

that most existing legal mechanisms for conservation cannot be relied upon to effectively 

conserve wildlife habitat outside protected areas. Although environmental statutes such 

as the Forests Act (Laws of Kenya, 2005) prescribe conservation guidelines, most environmental legislation in Kenya relies upon enforcement and compliance. This approach 

is not effective because of the lack of institutional capacity to enforce. Because existing 

measures are not sufficient, other mechanisms such as environmental easements, land 

purchase, and conservation leases are urgently needed. Even if Kenya’s environmental 

legislation was fully enforced and implemented, environmental easements, land leases, 

and land purchase would still be vital tools for conservation. Although these tools have 

great promise, they have not been widely used in Kenya because of legal limitations and 

lack of precedence. Legislative change and the testing of these tools through existing 

statutes are required to enable their broader application in Kenya. The Kenya Land Conservation Trust is introduced as an institution providing the framework and structure to 

utilize conservation mechanisms, such as environmental easements.

INTRODUCTION

The majority of Kenya’s large mammals inhabit land outside the network of 

formally protected national parks, reserves, and forest reserves, which comprise 

approximately 7.5% of Kenya’s land area (Western and Wright, 1994). It is also 

agreed that the survival of wide-ranging wildlife species found within protected 

areas depends on seasonal access to surrounding non-protected lands. Large mammals and their access to dispersal areas outside protected areas have been declining in Kenya for decades, creating a severe threat to their viability. Existing 

Nyokabi Gitahi, African Wildlife Foundation, 

Post Office Box 310, 00502, Nairobi, Kenya. 

Kathleen H. Fitzgerald, African Wildlife Foundation, Post Office Box 310, 00502, Nairobi, 

Kenya. Correspondence: K. Fitzgerald, kfitzgerald

@awfke.org. 

Manuscript received 7 January 2009; accepted 18 

May 2010.
9 6 • smithsonian contributions to zoology

conservation mechanisms are not adequate to address the 

threat to wildlife and habitat. New mechanisms are needed. 

This paper outlines existing and proposed legal provisions 

and mechanisms for conserving wildlife habitat outside 

protected areas. In particular, we outline the potential for 

landholders to conserve wildlife habitat through new legal 

mechanisms, based on property rights.

LAND RIGHTS IN KENYA

The Constitution of Kenya1 recognizes and protects 

the right of individuals to hold and enjoy private property, 

either individually or in association with others, as one of the 

Rights and Fundamental Freedoms of the individual. It protects the individual from being arbitrarily deprived of their 

property without compensation (Laws of Kenya, 2010: section 40). The right to property as used in the Constitution 

includes property rights in land and is further set out in various statutes that provide for the manner in which property 

rights in land may be held and transferred. The statutes include Government Lands Act (Laws of Kenya, 1915), Registration of Titles Act (Laws of Kenya, 1920), Land Titles Act 

(Laws of Kenya, 1908), Land (Group Representatives) Act 

(Laws of Kenya, 1968c), Trust Land Act (Laws of Kenya, 

1970), Registered Land Act (Laws of Kenya, 1963b), Land 

Acquisition Act (Laws of Kenya, 1968a), and the Land Control Act (Laws of Kenya, 1967). The statutes relating to land 

will be revised to ensure that they are consistent with the 

new Constitution of Kenya of 2010 adopted and enacted in 

a national referendum on 4 August 2010.

The Constitution classifies landownership into three 

categories: public land; private land, in which individuals 

or corporate entities hold lease or freehold interests; and 

community land, in which freehold interests are vested in 

communities, including registered groups whose lands are 

referred to as “group ranches,” lands registered to a specific community, and lands vested in county governments 

to hold in trust on behalf of the communities residing 

within them (Laws of Kenya, 2010: sections 62–64).

Public lands comprise what were formerly Crown 

Lands under the Crown Land Ordinance of 1915 (Laws 

of Kenya, 1915) and consist now of unalienated land; that 

is, land that had not been leased to any other person as 

of the effective date of the Constitution of Kenya of 2010 

(by which date the Commissioner of Lands had not issued any letter of allotment). Public land also includes 

government forests, game reserves, water catchment areas, national parks, and government animal sanctuaries 

as well as all minerals and mineral oils, and all rivers and 

lakes. Ownership of public land is divided between county 

government and national government as trustees for the 

people of Kenya. Government forests, government game 

reserves, national parks, and animal sanctuaries are vested 

in the national government. A national land commission 

established in the Constitution administers all public lands 

on behalf of the national and county governments.

Private land is held by individuals or corporations in 

the form of freehold or leasehold interests. These interests 

are established through a process of adjudication, consolidation, and titling of land. The only limitations to absolute 

proprietorship over land are when there are restrictions registered against the title, such as a right of way, caveats, or 

use agreements or land use regulations created by law (Laws 

of Kenya, 1963b, 2010). Absolute proprietorship over land 

is unlimited in time and can be transferred over generations. 

Private ownership of land is also achieved through leasehold tenure provided under the Registered Land Act (Laws 

of Kenya, 1963b). Leaseholds are generally issued for 99 

years and are renewable. Foreigners are not permitted to 

hold free-hold title in Kenya and they can only hold leases 

for up to 99 years (Laws of Kenya, 2010: Section 65(1)).

Most of the land outside protected areas that is used by 

wildlife is classified as “agricultural.” Ownership of agricultural land in Kenya is controlled by the Land Control Act 

(Laws of Kenya, 1967). Agricultural land transactions that 

are subject to the Land Control Act include sale, transfer, 

lease, mortgage, exchange, partition, or other disposal of 

or dealing with agricultural land, including shares in a private company or cooperative society that owns agricultural 

land (Laws of Kenya, 1967: section 6). The Act establishes 

Land Control Boards whose responsibility is to consider 

and grant or deny consent to any controlled transaction 

following application as set out in the Act. The Land Control Act limits dealings with agricultural land to citizens of 

Kenya, private companies, or cooperative societies whose 

membership is entirely comprised of Kenyan citizens, group 

representatives incorporated under the Land Group Representatives Act, or a state corporation established under the 

State Corporation Act (Laws of Kenya, 1986). Any other 

individuals or entities that wish to deal in agricultural land 

are required to seek exemption from the provisions of the 

Land Control Act, and this exemption can only be granted 

by the President of the Republic of Kenya following an application for exemption (Laws of Kenya, 1967: section 24). 

In the absence of an exemption, the Land Control Board is 

required to refuse consent to persons who do not meet the 

above criteria, and refusal of consent makes a controlled 

transaction for which the consent is sought void for all purposes (Laws of Kenya, 1967: section 9(2)). For example, if 
number 632 • 9 7

the African Wildlife Foundation (AWF) wishes to purchase 

agricultural land in Kenya, because it does not qualify as 

any of the above, AWF must seek exemption from the Land 

Control Board to own land. These controls on dealing with 

agricultural land are expected to change to reflect the provisions of the new Constitution on the rights to own property 

and the establishment of a national land commission. 

Lands held under group ranches and lands vested in 

county governments in trust for local communities living on 

the land are categorized as community lands. Group ranches 

are established under the Land (Group Representatives) Act 

(Laws of Kenya, 1968c) and apply mainly in the rangeland 

districts. Most of these rangelands are occupied by Maaspeaking communities that were largely nomadic pastoralists.

In 1968 the Land Adjudication Act (Laws of Kenya, 

1968b) and the Land (Group Representatives) Act (Laws 

of Kenya, 1968c) formalized the conversion of rangelands 

into group lands to communities with title vested in small 

groups. No individual group member can dispose of the 

land. The group members continue to access and use the 

land communally on the basis of customary laws (Laws of 

Kenya, 1968c). Group representatives are elected and, upon 

registration, become a body corporate. Members of the 

group have a right to vote and participate in decision making through meetings of the group. Over the past decade, 

many group ranches in southern Kenya have subdivided 

their land, changing the dynamics of group ranches by vesting ownership within parts of the ranch to individuals. For 

example, east of Amboseli National Park in southern Kenya, the Kimana Group Ranch has subdivided their land 

into 60 acre lots that are leased by individuals, who are now 

free to sell or lease their land for any use permitted by law. 

Land vested in county governments to hold in trust 

for the communities residing within them are referred to 

as trust lands. These were previously native reserves in 

the colonial period that were, at independence, vested in 

local government authorities as trustees for communities 

ordinarily resident in them (Laws of Kenya, 2010). These 

lands are now vested in county governments as per the 

new Constitution (Laws of Kenya, 2010: section 63). The 

communities residing in trust lands apply African customary law for purposes of land occupation, use, control, disposal, and succession. (Laws of Kenya, 1939: section 69).

COMPULSORY ACQUISITION AND 

LIMITATION OF LAND RIGHTS

None of Kenya’s land tenure systems as previously 

described grants an absolute right of land use without 

any limitations on use, nor does any land tenure preclude 

the government from acquiring such land. The Constitution provides the government two ways to limit property 

rights: through compulsory acquisition or through land 

use regulations. The Constitution states that the government may acquire property compulsorily for purposes of 

public interest and that this may be exercised for the following purposes: defence, public safety, public order, public morality, public health, or land use planning (Laws of 

Kenya, 2010).

In the case of compulsory acquisition, the landowner 

is entitled to full and prompt compensation under certain circumstances (Laws of Kenya, 2010: section 40(3)). 

When determining compensation, the government considers inter alia the market value of the land (Laws of Kenya, 

1968a: section 9). The Land Acquisition Act states that 

land may be acquired for purposes of the government or 

public body (Laws of Kenya, 1968a: section 6). Once the 

Minister for Lands makes a decision to acquire land for 

public purposes, a notice is published in the Kenya Gazette and served to the landowner and interested parties. 

The power of government to acquire land for purposes of wildlife conservation outside protected areas is 

also provided for within the Wildlife (Conservation and 

Management) Act, which empowers the government, 

through the minister responsible for wildlife, to declare an 

area a national park (Laws of Kenya, 1976: section 6). In 

the case of private land the process requires parliamentary 

approval before the requirements of the Land Acquisition 

Act (Laws of Kenya, 1968a) can apply. The end result of 

compulsory acquisition of land is that title is vested in the 

government free from all encumbrances (Laws of Kenya, 

1968a). When all the land has been acquired, documents 

of prior title are cancelled. When only part of the property 

has been acquired, the documents of title are amended to 

reflect the change against the title in the Register of Titles 

(Laws of Kenya, 1968a: section 20).

Although the potential to protect wildlife habitat 

through application of eminent domain is theoretically unlimited, any attempt to apply compulsory acquisition today 

would be met with resistance of a scale highly likely to defeat the objective. Kenya’s protected areas were established 

through a process of compulsory land acquisition largely 

belonging to communities, resulting in a great deal of bitterness and resentment that persists today. Any attempt to 

secure conservation land through eminent domain would 

be fought vigorously by community members and leaders. 

Apart from compulsorily acquiring land, the government has power to regulate and limit the rights of landowners to use land for the purpose of conserving wildlife 
9 8 • smithsonian contributions to zoology

habitat (Laws of Kenya, 2010: section 66). This is done 

mainly through land use planning and zoning and the use 

of regulations and sanctions to enhance compliance. The 

government has no obligation to compensate the landowner for land rights limited as a result of land use regulations.2

 Like eminent domain, forcefully limiting personal 

rights on land for wildlife conservation will be challenged 

because of the history of compulsory land acquisition for 

the creation and management of protected areas in Kenya. 

Thus, conservation via land use regulations alone is not a 

viable option for securing land for conservation. 

HABITAT CONSERVATION USING 

EXISTING STATUTES

Kenya has numerous statutes that are designed to 

conserve natural resources. This section highlights some 

of these statutes and how they may be used to protect land 

outside of protected areas. Currently, Kenya is reviewing 

and redrafting a number of its policies that pertain to natural resource conservation. Any new laws will need to be 

made fully compatible with the new Constitution.

Provisions of the Water Act (Laws of Kenya, 2002) 

relating to the protection of water bodies could, if implemented effectively, benefit wildlife habitat. The Act’s provisions can be used to ensure availability of clean water 

for wildlife and protect catchment habitat around water 

bodies. The Act prohibits various harmful activities relating to water bodies and requires permits for certain 

activities. For example, the Act criminalizes pollution of 

water bodies as well as obstruction or diversion of water 

from any water resource without authority under the Act, 

with stiff penalties for violations (Laws of Kenya, 2002: 

section 94).

The Agriculture Act (Laws of Kenya, 1963a) is aimed 

at promoting and maintaining agricultural production, 

stimulating the development of good land management and 

husbandry, developing agricultural land, conserving soil and 

soil fertility, and preventing soil erosion. “Agricultural land” 

refers here to all land that is used for purposes of agriculture 

and excludes any land that is by provision of any law relating to planning, such as the Physical Planning Act, and proposed for use other than agriculture (Laws of Kenya, 1963a: 

section 2). The term “agriculture” is used broadly under the 

Agricultural Act and includes cultivation, horticulture, dairy 

farming, beekeeping, and raising livestock. Much of the land 

used by wildlife outside of protected areas is arid, agricultural land primarily used by pastoralists. Agricultural land 

also includes land used for the keeping of game animals and 

birds as well as breeding and game ranching within the provisions of Kenya’s Wildlife (Conservation and Management) 

Act (Laws of Kenya, 1976: section 2).

The Agriculture Act gives the minister in charge of 

agriculture authority to issue sanctions to enforce the provisions of the Act. For example, the Minister may require 

a private landowner to carry out an activity that preserves 

soil (Laws of Kenya, 1963a: section 51). The Act makes 

it an offense for any landowner to fail to comply with an 

order, and if found guilty, the offender is liable to a fine, 

imprisonment, or the disposal of his land (Laws of Kenya, 

1963a: section 60). The Agriculture Act would benefit 

rangelands by preventing soil degradation; however, like 

most laws that require enforcement, these sanctions have 

not been effective because of lack of enforcement capacity. 

The Minister of Agriculture has wide powers for ensuring preservation of agricultural land, and these include 

powers to make regulations and rules for preservation and 

development of agricultural land (Laws of Kenya, 1963a: 

sections 51, 64). The Minister has wide powers for enforcing the provisions of the Agriculture Act as well as any 

regulations made under the Act, including the power to 

dispossess owners and occupiers of agricultural land and to 

purchase or compulsorily acquire the land from landowners who fail to comply with any regulations or orders made 

under the Agricultural Act. The Minister also has power to 

intervene in issues of land management on any agricultural 

land if, in the Minister’s opinion, the land is inadequately 

managed or has ceased to be managed; and the Minister can 

take measures to prevent or delay deterioration of such land, 

including issuing management orders against the landowner, 

which allows the Minister to manage the land to the exclusion of the landowner (Laws of Kenya, 1963a: section 187).

The Forests Act (Laws of Kenya, 2005) was intended 

to provide for conservation and management of Kenya’s 

forest resources. This is another statute that could have 

potential benefits to wildlife habitat outside protected areas; however, like the Water and Agricultural acts, this Act 

relies on enforcement and sanctions for compliance (Laws 

of Kenya, 2005).

The 1999 Environmental Management and Coordination Act (EMCA) provides a legal framework for 

management of the environment, which is defined to include the physical factors of the surroundings of human 

beings, including land, water, atmosphere, climate, sound, 

odor, taste, the biological factors of animals and plants, 

and the social factor of aesthetics, and includes both the 

natural and built environments (Laws of Kenya, 1999: 

section 2). The Environmental Management and Coordination Act contains several provisions that could be 
number 632 • 9 9

used to protect wildlife habitat outside protected areas. 

These provisions include environmental easements, environmental restoration orders, environmental impact assessment, and environmental impact licenses. 

The foregoing examples illustrate that Kenya has several strong statutes intended to protect natural resources 

and the environment. However, much of Kenya’s environmental legislation relies heavily on compliance, implementation, and enforcement. Because of low capacity of 

enforcing agencies, there is poor enforcement of Kenya’s 

environmental laws. Monitoring for compliance and prosecuting offences are challenging functions, especially where 

there is widespread default, and most government institutions have insufficient staff for these purposes. With few 

exceptions, the overall result is a general failure of existing 

legal measures for conserving wildlife habitat outside of 

protected areas, thus leaving the wildlife vulnerable to various threats and resulting in a continued decline in wildlife. 

MECHANISMS FOR HABITAT 

CONSERVATION BASED ON 

PROPERTY RIGHTS

If Kenya’s environmental laws were fully enforced and 

implemented, tools such as environmental easements, land 

acquisition, and conservation leases would still be necessary to adequately secure Kenya’s wildlife and landscapes. 

Conservation of natural resources is best served by a diversity of tools. Regulation may be appropriate in certain 

places, whereas different tools may be more appropriate in 

other areas. In addition, legislation is subject to change, so 

although Kenya may enjoy strong environmental legislation one year, that may change the following year, which 

jeopardizes the conservation of natural resources. 

Tools such as leases, described below, can provide 

direct benefit to landowners; otherwise, they may choose 

alternative land uses that are permitted by law yet are incompatible with wildlife and conservation. For example, 

Kitengela is a wildlife dispersal area south of Nairobi National Park. Development pressure is significant because of 

proximity to Nairobi. Much of the land in Kitengela has 

been subdivided and is held by title; thus, landowners can 

legally sell their land for development or agriculture. One 

conservation entity is leasing the land on an annual basis to 

keep the area open for wildlife movement. This land would 

have been fragmented by a number of legally permissible 

uses if not for the lease program. The challenge, however, 

is how to sustain the lease program as well as compete with 

escalating market prices. If someone can make four times 

as much by selling their land than they can from the lease 

program, they will be very tempted to sell.

The failure of coercive legal measures and existing 

statues for conserving wildlife habitat outside protected 

areas and the diversity of conservation needs and challenges call for additional approaches. There is an urgent 

need for incentive-based mechanisms that provide landowners with opportunities to conserve their land through 

the voluntary restriction of their property rights. These 

mechanisms include land purchase, environmental easements, and conservation leases. The successful use of 

conservation mechanisms based on property rights depends on clearly defined and articulated land rights. A 

landowner must have secure legal rights over the land and 

the power to convey rights and interests for the benefit of 

wildlife conservation in order to take advantage of these 

tools. Several such mechanisms are described below.

Land Purchase

Typically, when parcels of land come up for sale in 

Kenya, they are purchased, developed for agriculture, 

habitation, or other uses that are incompatible with wildlife conservation. These lands can be purchased by entities 

intending to manage them for conservation. For example, 

in 2004 a large-scale ranch in Laikipia District that was 

located in an area strategically important for conservation 

was purchased by a private company registered in Kenya 

and funded through an international conservation organization. Today, the land is managed successfully for conservation and wildlife, and it also benefits the surrounding 

communities. In this case, the provisions and requirements 

under the Land Control Act as to who can own land in 

Kenya were met because the purchaser of the above land 

is a company comprised of Kenyan shareholders. Buying land for conservation purposes, assuming compliance 

with the Land Control Act and other Kenyan statutes, is 

an effective tool that can be used by conservation entities 

to secure critical wildlife habitat. By owning the land, the 

conservation entity is in control of the land use and can 

manage specifically for wildlife purposes. The challenges 

to purchasing land are ensuring clear land rights of the 

seller, raising the funds to purchase land, and managing 

the land in the long term. 

Environmental Easements

An environmental easement is an agreement between 

a landowner and the holder of the easement, which restricts certain uses of a property to achieve conservation 
100 • smithsonian contributions to zoology

purposes. Easements in Kenya have their origin in English 

law (Gaunt and Morgan, 1997). They have also been used 

for non-conservation purposes, such as to create a right 

of way or grant water rights(Gaunt and Morgan, 1997). 

In Kenya, the 1999 EMCA makes provisions for use 

of easements to conserve wildlife habitat. The aspect of 

the environment that benefits from the restrictions placed 

on the land by an easement is called the “benefited environment,” and the land subject to the easement is called 

the “burdened land.” The essence of an environmental 

easement is to conserve target resources such as wildlife, 

watersheds, and habitat. 

Use of environmental easements to conserve wildlife 

habitat requires landowners to forego certain land use options on their land. Under EMCA, the process operates 

through the court and is not necessarily voluntary. If the 

court imposes an environmental easement, the landowner 

is entitled to compensation commensurate with the lost 

value of the use of the land (EMCA). The person awarded 

the easement pays the compensation, unless the court determines a national importance; then the government may 

be instructed to compensate the landowner. Under EMCA, 

section 112 (Laws of Kenya, 1999), anyone can hold a conservation easement. This is in contrast to the Land Control 

Act (Laws of Kenya, 1967), which controls dealings in agricultural land and limits the rights of noncitizens to own 

agricultural land without the consent of the President of the 

Republic of Kenya, which may be obtained following an 

application for exemption from the provisions of the Land 

Control Act. In a situation where an environmental easement is negotiated, the landowner and the easement holder 

would agree on the value of compensation due to the landowner. For example, if Mr. M. believes his neighbor’s property is important wildlife habitat, under EMCA he can file 

for an environmental easement. If the court approves the environmental easement, Mr. M. will hold the easement, and 

as per the recommendation of the court, he will compensate 

the neighbor. The neighbor has no choice in this matter.

To encourage landowners to place environmental easements on their land, EMCA and/or other legislation in Kenya should be amended to allow for the use of voluntary

environmental easements without being imposed through 

judicial process. Imposing an environmental easement 

through the court presumes an adversarial approach and 

could be viewed as punitive. Instead, landowners and appropriate organizations should be able to negotiate and agree 

to voluntary environmental easements. To ensure the longterm sustainability of easements, institutions that can monitor and uphold easements should be the easement holders. 

Moreover, EMCA makes no provision for determining the 

value of an environmental easement, leaving it to the courts 

to perhaps set the precedence. Less adversarial systems and 

a standardized appraisal process should be developed to 

determine the value of environmental easements (Farrier, 

1995). To date, an environmental easement has not been 

implemented through EMCA. In addition to proposing 

legislative change to provide for voluntary environmental 

easements, the African Wildlife Foundation is working with 

partners to bring a consensual environmental easement to 

the court to process through EMCA with the aim of setting 

a positive precedent for voluntary easements.

Community land presents a challenge to the use of 

environmental easements because rights of use over land 

among the communities that live on the land are governed 

by customary law, which is unregistered. This is expected 

to change as the Constitution provides that the national 

land commission would advise the national government 

on a comprehensive program for the registration of title in 

land throughout Kenya. The laws relating to governance 

of community land are also to be reviewed and rationalized, and an important outcome of this process would be 

provision for a process to define and quantify the nature 

of customary rights of use over land to enhance the security of tenure and enable the use of property-rights-based 

conservation mechanisms, especially easements and leases.

Leases

A lease provides an opportunity to obtain possession 

of a parcel of land for a given duration for specific uses. 

This mechanism is very common in land transactions and 

can be used for purposes of conserving land. For example, 

an appropriate entity may purchase a 99 year lease on a 

property and manage it for conservation purposes. Another variation is leasing specified rights on a property. 

Many safari and tourism companies lease land for tourism use, tented camps, and wildlife watching. The African 

Wildlife Foundation is using leases, for example, to conserve wildlife dispersal areas outside Amboseli National 

Park in southern Kenya, leasing land from landowners 

for habitat conservation. Through a lease, AWF is paying 

approximately 210 landowners to manage their land sustainably and not block wildlife movement with development, fencing, and farming. This program has protected 

approximately 12,500 acres of strategic wildlife habitat 

and is expanding by incorporating more adjacent landowners. This is another example that documents the need 

for tools such as leases because fragmentation by development, fencing, and agriculture is legally permissible in 

this area under Kenya law and landowners would choose 
number 632 • 101

alternative income-generating land uses that may not be 

compatible with wildlife. The challenge is finding a longterm funding source to sustain lease programs.

LACK OF INCENTIVES FOR WILDLIFE 

CONSERVATION ON PRIVATE LAND

Kenya landowners generally use land for purposes 

that are profitable, and land use that does not provide favorable returns is likely to be shunned. Landowners are 

required to absorb costs of keeping wildlife outside protected areas, and those not directly investing in ecotourism 

see little or no return from wildlife. In fact, many landowners face direct economic challenges from wildlife, for 

example, elephants raiding crops or predators killing livestock. Yet much of the wildlife in protected areas would 

not survive without seasonal access to adjoining private 

and community lands. A key challenge for wildlife conservation is therefore to find ways of making wildlife a profitable land use. One way to ensure that landowners benefit 

from wildlife is by compensating them for conservation of 

their land through conservation leases or environmental 

easements, such as the examples provided in the Amboseli 

region and in Kitengela.

One of the greatest challenges facing wildlife in Kenya

today is the shift in land use from grazing to cultivation. 

The government’s policy has promoted agricultural production through, for example, support for irrigation, financing for land subdivision and titling to establish farms, 

and subsidized loans to support cultivation of specific 

crops. Cultivation then becomes more profitable than 

keeping wildlife or cattle. In addition, there is significant 

international investment and focus on food security and 

large-scale agricultural development in Africa. This investment puts added pressure on land in Africa and further 

encourages governments to promote agricultural development without proper land use planning. 

At the same time, many group ranches are going 

through a process of subdivision into small, fenced units, 

which severely fragments the landscape and impedes wildlife movement. As a result, livestock production has become limited, with less room to roam. Many pastoralists 

have turned to cultivation, which further fragments the 

landscape, increases human-wildlife conflict, makes traditional pastoralism unsustainable, and leads to habitat 

degradation. 

Measures should be taken to encourage the government to prevent subdivision of rangelands below economically viable size to ensure that the rangelands continue to 

support livestock and wildlife and to provide tax incentives for landowners to establish wildlife-based tourism. 

Private land tools, such as conservation leases, can provide 

an added benefit to landowners and prevent conversion 

of open land to cultivated units. The greatest challenge 

herein is finding the funds to support conservation leases 

as donor funds cannot be relied upon in the long term. In 

some areas, tourism may directly support leases; however, 

this is not a viable option everywhere. Creative benefitsharing mechanisms from existing protected areas and 

tourism facilities must be explored. Payment for ecosystem services is another potential source of revenue to support conservation leases. 

KENYA LAND CONSERVATION TRUST

Many state institutions in Kenya have a mandate to 

conserve aspects of the environment; however, existing 

institutional frameworks for these agencies do not permit them to effectively conserve wildlife resources outside 

protected areas, especially through the use of propertyrights-based mechanisms. This is because they have a 

limited mandate to conserve wildlife outside protected areas, especially on private land. Therefore, effective use of 

mechanisms based on property rights for conservation of 

wildlife habitat in Kenya requires an enabling institutional 

framework that until recently has been lacking. Recognizing the need for a national organization that can pilot environmental easements, leases, and land acquisition and 

serve as a countrywide organization, the African Wildlife 

Foundation helped cofound the Kenya Land Conservation 

Trust (KLCT) in 2005 (Box 1).

The KLCT provides an institutional framework to 

support the development and use of property-rights-based 

mechanisms such as environmental easements and land 

purchase for habitat conservation. It aims to work in 

collaboration with landowners and partners to conserve 

wildlife habitat in areas such as the Ewaso ecosystem and 

to test and advance the use of legal and economic mechanisms for habitat conservation. The KLCT can serve as 

the institutional holder of environmental easements and 

promote the advancement of private land tools. 

AFRICAN WILDLIFE FOUNDATION

The African Wildlife Foundation (AWF) was founded 

in 1961 and works together with the people of Africa to 

ensure the wildlife and wildlands of Africa will endure 
102 • smithsonian contributions to zoology

forever. AWF’s land protection program had primarily focused on supporting protected areas and helping communities protect their land through management and land use 

plans and zoning. To successfully achieve its conservation 

goals, AWF recognized that in addition to its existing conservation tools and land conservation programs, it must 

put more emphasis on direct conservation measures on 

lands outside protected areas and to do so, it must utilize 

new conservation approaches.

The African Wildlife Foundation is working to identify strategic conservation priorities in the Ewaso ecosystem and throughout Kenya. AWF uses a scientific approach 

to identify its conservation targets, considering a suite of 

indicator species, access to water and other natural resources, wildlife movement patterns, connectivity, threat, 

and natural communities. Once a property is identified, 

AWF carefully evaluates the appropriate tool needed to 

conserve the parcel. AWF firmly believes that the advancement of conservation tools, as discussed in this paper, will 

help itself and other conservation organizations to achieve 

conservation success throughout Kenya. AWF has used a 

variety of these tools; it has purchased land, leased land, 

and is currently working with partners to secure land with 

an environmental easement through EMCA.

To advance the development of private land conservation tools, AWF convened a team of attorneys and conservation practitioners to explore environmental easements 

in Kenya, review existing statutes to determine how and 

if voluntary easements can be implemented under existing 

legislation or propose new legislation and amendments, 

and develop valuation methodologies for environmental 

easements and conservation leases. The results of this 

working committee were published by AWF (Watson et 

al., 2010).

CONCLUSION

Kenya’s wildlife is severely threatened by habitat 

fragmentation, loss of habitat, land use change, barriers 

to movement, and human population pressure. The current institutional frameworks for wildlife conservation 

have failed to effectively conserve wildlife resources outside protected areas in Kenya and provide landowners 

with an incentive to manage their land for wildlife. Most 

of Kenya’s wildlife is dependent upon habitat outside of 

protected areas. This is especially true for the Ewaso ecosystem, where wildlife thrives largely on privately and 

communally owned land. To adequately address the conservation challenge in Kenya, a wide diversity of tools 

must be developed, enhanced, and utilized. Kenya’s environmental statutes alone will not safeguard the country’s 

wildlife and landscapes. Tools such as leases, acquisition, 

and environmental easements play a vital role in conserving Kenya’s wildlife habitat. 

Kenya’s property rights regime provides substantial 

opportunities to use various mechanisms such as environmental leases to conserve wildlife habitat outside protected areas by restricting land use to activities that are 

BOX 1. 

The Kenya Land Conservation Trust.

Objectives

The Kenya Land Conservation Trust was established 

in 2005 as a private charitable trust, incorporated under 

the Trustees (Perpetual Succession) Act (Laws of Kenya, 

2009). The main objective of the trust is to ensure the 

viable function and integrity of natural habitat on land 

outside protected areas and the wildlife populations that 

they support while taking into account the social and economic interests of landowners and communities. The trust, 

in partnership with landowners and other stakeholders, 

will do this by (1) providing landowners with an opportunity to use land conservation options, including land 

acquisition, environmental easements, and leases for 

land conservation, and (2) supporting the formulation of 

policies relevant to biodiversity conservation in Kenya.

The Board of Trustees

The management of the trust is currently vested in 

the Board of Trustees. The founding trustees are representatives of African Wildlife Foundation, Kenya Wildlife 

Service, and the Ministry of Lands and Housing. One of 

the board seats is allocated to a member of the National 

Land Owner Forum. This is a fully Kenyan board.

Scope of the trust

The trust is focused on biodiversity conservation outside protected areas in Kenya. It will address aspects of 

conservation including wildlife migratory routes, dispersal areas, forest areas, and wetlands. With the support of 

the African Wildlife Foundation, the trust has undertaken 

a process of site selection and prioritization to determine 

the landscapes it will target for conservation intervention 

in Kenya. Overall, it is important that the trust will seek to 

engage with and respond to the landowners and stakeholders in designing and implementing habitat conservation mechanisms.
number 632 • 103

compatible to conservation of wildlife habitat as well as 

providing an opportunity to purchase critical parcels of 

land. These mechanisms not only protect important habitats but also provide landowners with the compensation 

they need to effectively manage their land for wildlife and 

ecological integrity. Environmental leases are already being used successfully, and this model can be replicated 

throughout Kenya. However, creative and collaborative 

funding mechanisms must be developed to support the 

large-scale conservation that is necessary to sustain Kenya’s wildlife and habitat.

The Environmental Management and Co-ordination 

Act provides an opportunity to use environmental easements, and this instrument has the potential to be used 

to conserve wildlife habitat on private land without the 

burden or cost of purchase. An amendment to EMCA and/

or other environmental legislation in Kenya to provide for 

voluntary easements would help catalyze the use of this 

tool and significantly enhance conservation in the Ewaso 

ecosystem and throughout Kenya.

Kenya boasts unique landscapes and wildlife; however, 

with only approximately 7.5% of the country formally protected, additional tools and strategies must be adopted to 

conserve additional lands. The use of environmental easements, leases, and land acquisition will complement Kenya’s 

existing conservation framework and tools to ensure the 

long-term sustainability of its wildlife and wildlands.

NOTES

1. Reference to the Constitution of Kenya refers to the 2010 constitution adopted and enacted through a national referendum on August 4, 

2010. This replaced the 1963 Constitution of Kenya. 

2. Land use control measures must be applied with caution as excessive land use regulations may be seen to amount to compulsory acquisition of the land and may be challenged on a constitutional basis as 

taking. In the United States, for example, land use regulations may be 

challenged on the grounds that they are so restrictive that the state has 

as good as acquired the land and may be declared unconstitutional, as 

was the case in Morris City Land Improv. Co. v. Parsippany-Troy, 40 

N.L. 539, 193 A.2d 233.242 (1963) and Lucas v. South Carolina Coastal 

Council, 112 S. Ct., 22 ELR (1992).

REFERENCES

Farrier, D. 1995. Conservation Biodiversity on Private Land: Incentives 

for Management or Compensation for Lost Expectations? The Harvard Environmental Law Journal, 19 19(2):304–305.

Gaunt, J., and P. Morgan. 1997. Gale on Easements. London: Sweet and 

Maxwell.

Laws of Kenya. 1908. Land Titles Act. Chapter 282. Nairobi.

———. 1915. Government Lands Act. Chapter 280. Nairobi.

———. 1920. Registration of Titles Act. Chapter 281. Nairobi.

———. 1939. Trust Land Act. Chapter 288. Nairobi.

———. 1963a. Agriculture Act. Chapter 318. Nairobi.

———. 1963b. Registered Land Act. Chapter 300. Nairobi.

———. 1967. Land Control Act. Chapter 302. Nairobi.

———. 1968a. Land Acquisition Act. Chapter 295. Nairobi.

———. 1968b. Land Adjudication Act. Chapter 284. Nairobi.

———. 1968c. Land (Group Representatives) Act. Chapter 287. Nairobi.

———. 1970. Trust Land Act. Chapter 288. Revised edition. Nairobi.

———. 1976. Wildlife (Conservation and Management) Act. Chapter 

376. Nairobi.

———. 1986. State Corporation Act. Chapter 446. Nairobi.

———. 1999. Environmental Management and Co-ordination Act. 

Number 8. Nairobi.

———. 2002. Water Act. Chapter 372. Nairobi.

———. 2005. Forests Act. Number 7. Nairobi.

———. 2009. Trustees (Perpetual Succession) Act. Chapter 164. Nairobi.

———. 2010. Constitution of Kenya. Nairobi.

Watson, R., K. H. Fitzgerald, and N. Gitahi. 2010. Expanding Options 

for Habitat Conservation Outside Protected Areas in Kenya: The 

Use of Environmental Easements. African Wildlife Foundation 

Technical Papers, 2. 

Western, D., and M. Wright, eds. 1994. Natural Connections: Perspectives in Community Based Conservation. Washington, D.C: Island 

Press.

Landscape-Scale Conservation Planning 

of the Ewaso Nyiro: A Model for 

Land Use Planning in Kenya?

Karl A. Didier, Alayne Cotterill, Iain 

Douglas-Hamilton, Laurence Frank, Nicholas J. 

Georgiadis, Max Graham, Festus Ihwagi, 

Juliet King, Delphine Malleret-King, Dan 

Rubenstein, David Wilkie, and Rosie Woodroffe

ABSTRACT. The unique wildlife of the Ewaso Nyiro and valuable services that the 

ecosystem provides for humans (e.g., clean water and productive grasslands) cannot be 

conserved by working solely on traditional conservation strongholds such as the national 

reserves and private ranches of central Laikipia. To reach objectives for conserving wildlife, stakeholders must work to preserve wildlife habitat and corridors in the surrounding 

human-dominated landscape—a daunting task considering the complexity of working 

at large spatial scales (e.g., many landowners, competing land uses) and limited conservation resources available. Systematic, landscape-scale conservation planning helps 

stakeholders set meaningful and transparent objectives, identify where to work to meet 

those objectives, and prioritize areas for immediate investment. We describe results and 

implications of an initial landscape-scale planning exercise for the Ewaso Nyiro that culminated in a workshop in January 2006. Forty participants selected nine focal features, 

set quantitative objectives for four of them (elephants, Grevy’s zebra, lions, wild dogs), 

and set spatial conservation priorities for the entire landscape on the basis of complementary needs of critical species. The modest objectives for these species (e.g., maintaining a 

population of 300 wild dogs) cannot be met by conservation focused solely on traditional 

strongholds. The exercise indicated that nearly 84% of the landscape needs conservation 

investment, and it identified three near-term priorities: (1) maintain current investments 

in conservation strongholds, (2) increase investment to secure the narrow corridor between Samburu and Laikipia Districts, and (3) increase investment to secure portions 

of Samburu District, including the Matthews Range. The results we describe represent 

the initiation of a land use planning process that, if continued, can help meet both biodiversity and livelihood development objectives. We recommend this process be carried 

forward in the Ewaso Nyiro and then in similar ecosystems in Kenya and eastern Africa.

Karl A. Didier, Global Conservation Programs, 

Wildlife Conservation Society, Gainesville, FL, 

USA. Alayne Cotterill, Living with Lions and 

the Wildlife Conservation Research Unit, University of Oxford, Nanyuki, Kenya. Iain DouglasHamilton, Save the Elephants and Department of 

Zoology, University of Oxford, Nairobi, Kenya. 

Laurence Frank, Living with Lions, Panthera, 

New York, NY, USA; and Museum of Vertebrate Zoology, University of California, Berkeley, 

CA, USA. Nicholas J. Georgiadis, Property and 

Environment Research Center, Bozeman, MT, 

USA. Max Graham, Department of Geography, 

University of Cambridge, Cambridge, UK. Festus Ihwagi, Save the Elephants, Nairobi, Kenya. 

Juliet King, Northern Rangelands Trust, Isiolo, 

Kenya. Delphine Malleret-King, Laikipia Wildlife 

Forum, Nanyuki, Kenya. Dan Rubenstein, Department of Ecology and Evolutionary Biology, 

Princeton University, Princeton, NJ, USA. David 

Wilkie, Global Conservation Programs, Wildlife 

Conservation Society, Waltham, MA, USA. Rosie 

Woodroffe, Department of Wildlife, Fish, and 

Conservation Biology, University of California, 

Davis, CA, USA. Correspondence: K. A. Didier, 

kdidier@wcs.org.

Expanded author information follows the Acknowledgments section. Manuscript received 

7 January 2009; accepted 18 May 2010.
106 • smithsonian contributions to zoology

INTRODUCTION

From previous chapters in this book, it seems clear 

that in the Ewaso Nyiro ecosystem, conservation goals 

cannot be achieved by focusing solely on the few parcels 

of land that are currently under conservation-friendly land 

uses and management. To improve chances of conserving wildlife in the Ewaso ecosystem and more generally 

throughout much of Kenya, we believe that stakeholders 

need to adopt a systematic approach to prioritizing conservation investments at the landscape level.

A landscape-scale framework for planning conservation is needed for two main reasons. First, many valued 

species cannot be effectively conserved on the few increasingly isolated patches of lands on which biodiversity conservation is the primary land use objective (Gardner et 

al., 2007). This is particularly true for species that require 

extremely large areas and access to many different habitats to persist, i.e., many of the large herbivores and carnivores common in Kenya. Landscape-scale approaches to 

conservation aim to explicitly address the needs of these 

area-demanding species and to ensure their long-term persistence (Nicholson et al., 2006; Rouget et al., 2006; Salomon et al., 2006). 

A second reason that a landscape-scale framework 

is needed is that many critical ecological processes (e.g., 

migration and gene flow) and ecosystem services (e.g., 

production of clean water) also cannot be effectively conserved by working solely with landowners whose property 

covers only a small percentage of the landscape (Poiani et 

al., 2000; Chan et al., 2006; Salomon et al., 2006; Rouget 

et al., 2006). These processes and services are often critical for both biodiversity elements (species and ecological 

communities) and people. Landscape-scale planning approaches help to determine the spatial scope of activities 

that are needed both to meet the needs of species and to 

maintain ecological processes and ecosystem services important for people.

A systematic approach to planning (Margules and 

Pressey, 2000; Groves et al., 2002), with clearly defined 

steps and procedures, is also useful because it instills transparency into the process. This is particularly valuable in 

places where governance is weak, contested, incomplete, or 

has been abdicated by the state. The systematic approach 

aims to define a clear and transparent process for arriving 

at conservation priorities—a process that allows for broad 

stakeholder participation. Systematic conservation planning 

also immerses people, often for the first time, in a process 

focused on biodiversity rather than economic development. 

It also allows a wide range of stakeholders to bring to bear 

their distinct and often underappreciated knowledge of the 

ecology of a landscape. Most importantly, the systematic 

approach tries to instill transparent, objective, and effective 

criteria into a process for prioritizing conservation actions 

and areas (Margules and Pressey, 2000). It is almost always 

the case in conservation that we cannot conserve, in the immediate sense, everything we need to, and prioritization is 

needed. Objective and rigorous criteria help ensure that our 

resources are directed most efficiently. 

Systematic, landscape-scale conservation planning 

now has a history of nearly two decades. It began when 

conservationists first recognized that existing protected 

areas were necessary but not sufficient to preserve biodiversity effectively (Rodrigues and Gaston, 2001; Brooks et 

al., 2006). Many hundreds of projects at various spatial 

scales (e.g., Younge and Fowkes, 2003) around the world 

now use systematic procedures for conservation planning 

(Sarkar et al., 2006). More recently, several international 

conservation organizations have created planning approaches designed particularly for the landscape scale and 

firmly rooted in a systematic framework (Sanderson et al., 

2002; Conservation International, 2004; Henson et al., 

2009; Pressey and Bottrill, 2009).

We present here the results and implications of a systematic, landscape-scale conservation planning exercise 

for the Ewaso ecosystem. The exercise was conducted over 

a six-month period that ended in a five-day workshop in 

2005–2006. The approach we initiated is novel in that it 

was designed to skirt some of the weaknesses of existing 

approaches, which often demand more time and resources 

than are available. The methods and its novelties are described in detail in Didier et al. (2009) and in the workshop proceedings (King and Malleret-King, 2006) and are 

presented here only briefly. 

Our objectives are to first present the results of the 

exercise for a larger audience not present at the workshop, 

including local stakeholders from the Ewaso region and 

the wider conservation and development community in 

Kenya. Second, we discuss briefly the implications of the 

exercise for the particular species considered and for conservation in the Ewaso Nyiro in general. Third, we discuss 

what we could not achieve during our time- and resourcelimited project and how we think the process should be 

carried forward in the future.

METHODS

The group proceeded through the basic steps of systematic conservation planning as outlined in Box 1 (full 
number 632 • 107

description of methods can be found in King and MalleretKing, 2006, and Didier et al., 2009). Nine focal biodiversity features (in this case all “features” were species) were 

identified whose conservation requirements not only were 

considered to be complementary to each other but also, 

when added together, would, if met, cover the conservation 

needs of most other features in the landscape. We did not 

complete these planning steps exhaustively for all species, 

rather, for a subset of four species (elephant, Grevy’s zebra, 

lion, and wild dog), aiming to produce useful products for 

making near-term decisions and to create a framework for 

proceeding with conservation priority setting in future. We 

proceeded in two phases. First, a six-month preparation 

phase by a small organizational committee and a group 

of ecologists who were each familiar with the landscape 

and particular biodiversity features produced draft plans 

for the four species. We completed all the planning steps 

for four of nine focal species, although conservation objectives (step 3; see Box 1) were considered preliminary and 

were not explicitly incorporated into the subsequent steps. 

The exercise was largely centered on the creation of five 

component maps for each species that were used to guide 

decisions about which areas were a priority for immediate 

conservation investment. For these maps, each 5 × 5 km 

grid square was classified as to (1) its current importance 

for supporting the population of the focal species, (2) the 

potential for recovering the population of the species, (3) 

the potential for future decreases in the population as a 

result of human activities, (4) the cost of conservation action, and (5) confidence in the information provided. The 

maps were then used as guides to make species-specific 

and cross-species maps of conservation priorities.

BOX 1. The steps of systematic, landscape-scale conservation planning.

The steps to systematic conservation planning at any scale (global, national, etc.) are fairly similar, are well accepted 

among organizations involved in conservation, and have been thoroughly described elsewhere. Our description is adapted from 

those of Margules and Pressey (2000), Groves et al. (2002), Groves (2003), and Sarkar et al. (2006). Further detail is also 

provided in the proceedings of our workshop (King and Malleret-King, 2006) and by Didier et al. (2009).

1. Define the context, compile critical information, and set overarching goals/vision. Basic information 

on the extent of the landscape, ecology, human influences/threats in the landscape, and stakeholders need to be summarized, 

and broad goals need to be set. Spatial information should be compiled on the current biodiversity value of areas throughout 

the planning region (i.e., presence or abundance of focal conservation targets), the vulnerability of biodiversity to future loss, the 

potential for recovery of biodiversity value (if locally relevant), and costs of conservation. 

2. Define a set of focal biodiversity features. Because it nearly impossible to plan conservation around all the 

possible elements of biodiversity (species, habitats, and ecosystem functions and services), it is necessary to select a suite of 

focal features on which to plan conservation efforts. The number and type will depend on the context (e.g., availability of data, 

complexity of environment). 

3. Set quantitative objectives for each focal conservation feature. It is important to set quantitative objectives 

for conservation feature (i.e., how much do you want?), so that the spatial extent of conservation actions necessary to meet these 

objectives can be considered and progress toward the objectives can be measured.

4. Assess the effectiveness of existing conservation activities/areas. This step should first involve assessing 

whether current activities and the extent of areas where activities are occurring are sufficient for meeting quantitative objectives 

for all focal features. If targets are currently at quantitative objectives across the landscape, are current activities and areas sufficient to maintain features at or above quantitative objectives (i.e., prevent future losses)? If features are currently below objectives, are current activities and areas sufficient to increase focal features to quantitative objectives?

5. Summarize the benefits and costs of continuing current activities/areas and starting new ones in 

order to meet or maintain quantitative objectives of conservation features. This step involves creating summary 

maps and indices highlighting where investment may be wise. It may include benefit:cost ratios, irreplaceability scores, or other 

results of reserve design algorithms (e.g., from Marxan, C-plan, etc.).

6. Negotiate a map of conservation priorities. Physically score planning units in terms of priority for investment, 

considering all information in previous maps, possible mistakes, and all relevant information not in the maps (e.g., political constraints, opportunities, etc.). May also include evaluating trade-offs, both among conservation features and between biodiversity 

features and human development objectives (e.g., increasing livestock numbers). Decision-support software can help (Marxan, 

C-plan, Vista).
108 • smithsonian contributions to zoology

RESULTS

Vision for Biodiversity Conservation

Workshop participants agreed on a simple, consensus 

vision statement for biodiversity conservation in the region: “To conserve the native biodiversity and integrity of 

the Ewaso Nyiro landscape.”

Focal Biodiversity Features

Participants discussed and then selected a subset of 

the biodiversity within the Ewaso landscape that would 

help focus conservation planning efforts. The focal suite 

comprised nine biodiversity features (Table 1). Elephants, 

Grevy’s zebra, lions, wild dogs, and reticulated giraffe 

(Giraffa camelopardalis) were selected because they represented major habitats or threats to biodiversity (see 

King and Malleret-King, 2006). Jackson’s hartebeest (Alcelaphus buselaphus) was added because it was viewed 

by participants as being unique to the Ewaso Nyiro and 

participants felt its conservation could not be guaranteed 

by focusing attention on the other features. The suite 

also included what participants felt were the two most 

important vegetation communities: the acacia-grassland 

mosaic, which covers most of the landscape, and dry upland/montane forests (Matthews Range, Kirisia Hills, and 

Mukogodo Forest in particular). Finally, the suite of focal features included one ecological system/service, the 

hydrological system, because maintaining water flow and 

water quality in the landscape was seen as central to both 

conservation of biodiversity and livelihoods. As a whole, 

participants agreed that the suite of features would act as a 

good surrogate for protecting the other native biodiversity 

of the region and that if these features were successfully 

conserved, most if not all of the other native biodiversity 

in the landscape would also be conserved. It was agreed 

that the steps of conservation planning would include only 

four species because of available time and resources.

Quantitative Objectives for Biodiversity Features

Workshop participants produced preliminary estimates of (1) the current population size across the landscape for each of the four species, (2) a range around 

that estimate, and (3) a preliminary conservation objective (Table 1). Estimates represent “best information” 

and were based on consensus among a subset of participants working on each particular species. In some cases, 

estimates were based partially on empirical data collected 

from parts of the landscape and extrapolated to remaining areas for which empirical data did not exist. Quantitative objectives for all four species represented an increase 

from the current population estimate, although the objective for Grevy’s zebra represented a nearly 165% increase, 

whereas for the three other species the objectives represented only 10%–25% increases. Although the objective 

for elephants represented an overall increase in the population, participants produced a more specific objective to 

TABLE 1. The suite of focal biodiversity features selected by participants at the Ewaso Nyiro Landscape Conservation Planning Workshop (January 2006) and quantitative conservation objectives for these features. Subsequent to selecting these features, participants 

concentrated on only the top four, although the other features should be considered in future exercises. The quantitative conservation 

objective represents the total population that participants would like to see across the landscape in 10 years to be considered successful. 

Information was not compiled during this exercise for the following features: Jackson’s hartebeest (Alcelaphus buselaphus), reticulated 

giraffe (Giraffa camelopardalis reticulate), acacia-grassland mosaic, dry upland/montane forest, and hydrological system.

Range of Quantitative

Estimate of current estimated conservation

Focal biodiversity feature abundance (January 2006) abundance objective a

African elephant (Loxodonta africana) 8,000 animals 7,000–9,000 10,000 b

Grevy’s zebra (Equus grevyi) 1,700 animals c 1,600–2,100 4,500 d

Lion (Panthera leo) 450 animals 400–500 500

African wild dog (Lycaon pictus) 300 animals in 17–18 packs 250–350 20 packs e

aTo reach and maintain the amounts below within 10 years.

bRedistributed from their current distribution to 2,000 in Laikipia District (decrease from current), 6,000 in Samburu (increase), and 2,000 in the Mount Kenya region. cA recent drought may have reduced the population from the more long-term mean of approximately 1,900 animals.

dRepresenting approximately a 10% increase/yr.

eExtrapolating from a range of 300 animals in 17–18 packs to 20 packs gives a range for the objective of 333–353 animals.
number 632 • 109

reduce populations in Laikipia District while raising them 

in Samburu District.

Component Maps

Participants completed, by consensus, a set of five 

maps for each of the four focal species (Figures 1–3). The 

maps were then used as guides for setting conservation 

priorities. Some comparative summaries of these maps are 

provided in Table 2, and a description of the maps for 

each species is provided below.

Elephants

Of all the focal species selected, elephants are considered the most resilient to many of the threats facing 

biodiversity, primarily because of their relatively large 

population in the region, high level of protection, and 

ecological adaptability. Workshop participants estimated 

that there were approximately 7,000–9,000 elephants in 

the landscape, between 5,000 and 6,000 in Samburu and 

Laikipia and possibly 2,000 in Mount Kenya, although no 

firm information exists for Mount Kenya. The elephant 

population appears to be increasing. Elephants currently 

use approximately 88% of the landscape (based on the 

current importance map in Figure 1). They are particularly 

abundant in and around the commercial ranches of Laikipia, Mount Kenya, and parts of Samburu, particularly 

in the Matthews Range, Kirisia Forest, and the national 

reserves. Key corridors were highlighted as being particularly important habitat features for maintaining connectivity for the elephant population across the landscape 

(Figure 4).

If conservation investment is not continued and in 

some cases increased, there is a high potential for reductions in the elephant population, particularly in known 

poaching hot spots that currently support high abundances 

of elephants (e.g., Kirisia Forest, east of Matthews Ranges, 

Mukogodo Forest, and the Laikipia Nature Conservancy). 

Failure to continue conservation investment in these areas 

would likely lead to increased poaching. Potential reductions are also high in the national reserves and along the 

Ewaso Nyiro River. The potential for future reductions in 

the population is generally lower in the private ranches 

in Laikipia, where conservation is more established, and 

higher in Samburu, where conservation is more tenuous. 

It is also possible to recover populations in much of 

the landscape through conservation investment, although 

there is little potential for range expansion (<1% of the 

planning units where recovery could occur are currently 

unused). Potential for recovery is particularly high in 

much of Samburu, especially in the north, if security is 

improved. Recovery potential is generally lower in the national reserves, southern Samburu along the Ewaso Nyiro 

River, and in much of Laikipia, where elephant numbers 

are currently near ecological, or at least social, carrying 

capacity. Recovery is not possible in approximately 25% 

of the elephants’ current range. There is also no potential 

for recovery in the agricultural areas of southern Laikipia 

and Meru Districts. 

Cost of elephant conservation is higher in the southern part of their range than in the north. This is primarily 

because elephant conservation in the south involves securing land that is more arable and therefore more valuable 

than land in the north. In addition, elephant conservation 

in southern Laikipia requires the mitigation of humanelephant conflict, particularly crop raiding, which could 

involve the construction of electrified elephant-proof 

fences. Cost for conservation in the Kirisia Forest is considered higher than elsewhere in Samburu since currently 

there is little conservation investment in this area and 

threats are high. Conservation for elephants in most of 

Samburu would involve improving security, which can be 

done for a relatively small investment in terms of cost per 

unit area, and would have an impact over a large area.

For elephants, participants were fairly or highly confident in the information they provided for nearly the entire 

landscape, except for in the Mount Kenya region and the 

northwestern and northeastern corners of the landscape 

(Figure 3).

Grevy’s Zebra

Grevy’s zebra currently use about 48% of the landscape (calculated at 5 × 5 km resolution) and number 

around 1,700 animals. Central Laikipia, the Laisamis area 

in the northeast of the landscape, Wamba to the west of 

the Matthews Ranges, Lewa, and the national reserves 

currently support relatively high abundances of Grevy’s 

zebra. Besides maintaining high abundances, certain areas are important because they contain water holes or are 

grazing areas, nursery areas, or corridors. Laikipia is the 

southern edge of the species range; the species only recently 

moved into this area and Lewa (since the early 1970s), and 

these areas are now critical for the species. Much of the 

information on Grevy’s zebra is based on radio-tracking 

data, data from community scouts, and aerial surveys. 

There is a high potential for future reductions in the 

population of Grevy’s zebra across much of their current range if investment in conservation is not continued. 
110 • smithsonian contributions to zoology

FIGURE 1. Component maps for elephants and Grevy’s zebra. The maps were produced by participants at the Ewaso 

Landscape Planning Workshop (January 2006) and were used to help set conservation priorities.
number 632 • 111

FIGURE 2. Component maps for lions and wild dogs. The maps were produced by participants at the Ewaso Landscape Planning Workshop (January 2006) and were used to help set conservation priorities.
112 • smithsonian contributions to zoology

Although the potential for reductions on the commercial 

ranches of Laikipia is low because landowners generally 

maintain wildlife-friendly land management practices, it is 

relatively high in community areas of Laikipia because of 

competition with livestock. There is high potential for loss 

in the national reserves, if they lose their integrity through 

lack of conservation investment, at key hot spots (e.g., 

water points), and in areas of Wamba, West Gate, and 

Laisamis.

It is possible to increase populations of Grevy’s zebra 

across nearly all of its current range as all existing subpopulations are below carrying capacity. There is also some 

TABLE 2. Summary of component maps for each of four biodiversity features (Figures 1 and 2). Maps were produced by participants in the Ewaso Nyiro Landscape Conservation Planning 

Workshop (January 2006) and were the basis for a conservation 

planning exercise.

 Current 

Planning units (%) distribution (%)

Focal Recovery Future

biodiversity Currently Potentially not losses

feature used useda possibleb possiblec

African elephant 87.6 87.8 19.6 97.7

Grevy’s zebra 47.4 49.7 2.3 99.5

Lion 60.3 71.0 6.3 99.9

African wild dog 28.7 61.9 1.6 99.3

a If new conservation actions were taken or current ones continued.

b

 In some proportion of current range, recovery is not possible because the population is at carrying capacity.

c If current conservation actions were abandoned and no new ones begun.

FIGURE 3. Confidence maps for the spatial data (component maps 

in Figures 1 and 2) for elephants, Grevy’s zebra, lions, and wild dogs. 

The maps were produced by participants at the Ewaso Landscape 

Planning Workshop (January 2006).

FIGURE 4. Elephant corridors. The map was produced by participants at the Ewaso Landscape Planning Workshop (January 2006) 

and was used to help set conservation priorities.
number 632 • 113

potential for expanding the current range to areas that are 

currently unused, but this is limited to ranches in Laikipia 

District (<3% of the landscape). Recovery is possible if 

conservationists can secure enough grassland habitat and 

safe water and limit the numbers of plains zebra and lions. 

The subpopulation in Laikipia is growing at only 1%–3% 

per year. Although the reason for the low increase is not 

known, it may relate to habitat quality or interspecific 

competition (i.e., with other herbivores) at the southern 

edge of their historical range. There is also a high potential 

for recovery in the community-managed areas of Samburu 

District. 

In general, the cost of conservation activities aimed 

at Grevy’s was scored lower in Laikipia and Lewa Conservancy, where there has already been substantial investment in direct conservation and conservation education 

with communities and where relationships have been built 

with local people. Cost is moderate in the national reserves and neighboring communities (West Gate, Kalama, 

and Wamba, where work is ongoing), and costs are highest farther north and in the Kipsing area of Isiolo. 

Through most of the landscape, participants indicated 

high confidence in the information they provided, based 

on the large amount of research (radio tracking, surveys, 

and community monitoring) that has been performed.

Lions

Approximately 450 lions currently use around 60% 

of the landscape (calculated at 5 × 5 km resolution). Abundance is highest on commercial ranches of Laikipia and 

the national reserves of Samburu, Buffalo Springs, and 

Shaba. Agricultural areas cannot support lions. Other 

areas that support subpopulations include the Matthews 

Range and Kirisia Forest. The buffer areas around the 

national reserves are important dispersal areas for lions. 

There are lions throughout the pastoralist areas but in low 

numbers, therefore those areas are not as important for 

the landscape population. 

In general, the areas with the highest potential reductions in the future are those where lions are currently most 

abundant. The highest potential for loss is on the commercial ranches in Laikipia and the national reserves and 

surrounding areas, where in the absence of conservation 

efforts, lions would likely be persecuted. In most of Samburu, threats are intense, although the current abundance 

of lions is low, and therefore, the potential future reductions are also low. 

Potential for recovery is low in the commercial ranches 

of Laikipia, where lions are nearer carrying capacity, and 

in pastoralist areas. Potential for recovery is higher where 

there are few people and where habitat is suitable, including northeast Samburu and the northern Matthews Range, 

assuming that conservation and incentives can improve the 

attitudes of local people toward lions. Lions could potentially expand out from their current range to an additional 

11% of the landscape, primarily in the northwestern extreme of the landscape, although participants indicated 

that they were uncertain about the current status of lions 

in this region and the possibility of recovery.

Cost for conserving lions is primarily influenced by 

attitudes toward conservation. Areas where communities 

have some conservation awareness were given lower cost 

than those with no exposure to or history of conservation. Costs were also low in areas where lion numbers 

are highest, even though lion populations in these areas 

will be artificially suppressed (e.g., Lewa, Solio, and Ol 

Pejeta). 

Participants indicated that they had high confidence 

in their conclusion that lions would continue to persist in 

Laikipia and southern Samburu, and especially within the 

national reserves. Participants indicated that their confidence was based on the intensive research that has been 

performed in these places. Confidence was somewhat 

lower in the northern parts of the landscape, where information is based on local knowledge. 

Wild Dogs

Wild dogs current use approximately 29% of the 

landscape and currently number around 300 animals in 

17–18 packs. Core areas that support higher abundances 

and have resident animals are the commercial ranches in 

central Laikipia and the group ranches on the edge of the 

Laikipia escarpment. The area of Isiolo District between 

Laikipia and Buffalo Springs/Shaba, although it may not 

support resident animals, is important for maintaining 

connectivity between these two subpopulations. Dogs 

move through northern Laikipia in the Kirimun area to 

the east of the Kirisia Hills and the Seiya Lugga. Another 

very important area is the Matthews Range, where there 

are several wild dog packs. Local knowledge suggests 

there are some packs in central Samburu and a pack newly 

established in Shaba National Reserve. 

Potential for reductions in the wild dog population 

is generally high where animals are resident and threats 

exist. Reductions could be high in the Suguta valley if 

more land on the Loroki Plateau is used for wheat, which 

would, in turn, force pastoralists to move more of their 

livestock into the Suguta valley. Areas where there is high 
114 • smithsonian contributions to zoology

potential for loss also include areas where a high level of 

conflict exists between human populations and wild dogs 

and where wild prey risk being depleted. 

The wild dog population across the landscape is in a 

recovery phase after becoming locally extinct, at least in 

Laikipia. With conservation action, dogs could potentially 

double the size of their range to nearly 60% of the landscape (based on 5 × 5 km resolution). Hilly areas have 

the highest potential for recovery as wild dogs tend to 

prefer that habitat. Potential recovery is lower where current abundances are high, as these are likely to be already 

at carrying capacity. Large areas of the landscape are not 

suitable for wild dogs and, therefore, have no potential 

for recovery. 

There are no areas where the cost of conservation actions aimed at wild dogs is zero, primarily because there is 

a major disease risk to wild dogs (primarily from domestic 

dogs), which has the potential to wipe out the entire population. The cost of conservation is lowest in community 

areas of Laikipia, on the Laikipia escarpment, and in the 

Matthews Range, where there is little conflict, prey species are abundant (dik-dik), wild dogs are currently doing 

well, and the only cost is due to disease risk. Areas where 

costs are high are either insecure, from the standpoint of 

operating a conservation project, or have depleted prey. 

The cost was considered prohibitive in agricultural areas 

where conservation would require conversion of habitat. Participants indicated that they had high confidence 

about the information they provided in Laikipia, where 

animals are radio collared, but that confidence was lower 

in other areas. 

Spatially Explicit Conservation Priorities

After completing the five component maps for each of 

the four focal species, a map reflecting the benefit to cost 

ratio for conservation action was calculated (Figures 1, 2). 

Conservation benefits were calculated by weighting then 

adding the maps of potential recovery and maps of potential reductions. In other words, benefits from conservation 

in any location include either helping to recover populations or preventing future reductions. For each species, 

participants assigned relative weights to these two maps 

(i.e., how important is prevention relative to recovery). 

For each species, the weights (prevention:recovery) were 

as follows: elephant, 1:1; Grevy’s zebra, 2:3; lions, 2:1; 

and wild dogs, 5:1. 

The benefit:cost maps, in combination with the five 

component maps, were then used by participants as guides 

for producing maps of conservation priorities for each 

species (Figure 5, Table 3). Finally, an integrated map of 

conservation priorities across all four conservation targets 

was created (Figure 6).

Species-Specific Priorities: Elephants

For elephants, the strong conservation investment 

currently being put into the ranches and conservancies 

of Laikipia, Buffalo Springs, Samburu and Shaba National Reserves, Mount Kenya National Park, and a few 

group ranches and conservancies (Namunyak) in Samburu 

should be maintained, as these are elephant strongholds. 

Additional investment over the next decade should be 

aimed at securing strong subpopulations that are vulnerable to decline, bolstering subpopulations in some places, 

and securing corridors. Areas where subpopulations are 

currently strong but increased investment is needed to 

maintain them include the following:

1. Laikipia Nature Conservancy requires improved community outreach and development of communityowned conservancies and associated tourism among 

the neighboring communities, particularly the Pokot, 

to the west and north of the property. Investment 

is also required to improve security and upgrade 

elephant-proof electrified fences.

2. Mount Kenya Forest Reserve requires the mitigation 

of human-elephant conflict on smallholder land surrounding the forest and the establishment of corridors 

to the Laikipia Plateau to maintain connectivity between subpopulations.

3. Rumuruti Forest is the highest conflict area in the landscape. Unless the Marmanet and Ol Arabel forests, 

located west of the Rumuruti Forest, can be secured 

and rehabilitated, there is no future for elephants in 

this area, and efforts should instead focus on removing them to the large-scale ranching matrix in central 

Laikipia.

High-priority areas where populations could be bolstered with increased investment include the following:

1. In north Laikipia and Samburu, investment aimed at 

the Kirisia Forest, Mukogodo Forest, the Matthews 

Range, and the Sera Conservancy should attempt to 

reduce poaching and increase security. 

2. In Lekurruki and the Livestock Marketing Division 

in Isiolo, activities should aim to maintain an area 

of low livestock densities and reduce insecurity (i.e., 

poaching).
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FIGURE 5. Conservation priorities for individual species as described by participants at the Ewaso Landscape Planning Workshop (January 2006).
116 • smithsonian contributions to zoology

Several critical corridors need to be maintained 

through increased investment, particularly those linking Mount Kenya to Laikipia and several through Isiolo, 

which link Laikipia and Samburu.

Species-Specific Priorities: Grevy’s Zebra

For Grevy’s zebra, current levels of conservation investment are sufficient in some areas, including the commercial ranches of Laikipia, Lewa, Il Ngwesi, Namunyak, 

West Gate, and Kalama and the reserves of Samburu and 

Buffalo Springs. These levels need to be maintained to ensure continued success.

High-priority areas needing additional investment 

include areas that are important for grazing, watering, 

and reproduction. Securing these areas is considered essential for both recovering the species and preventing any 

future declines. Specific areas that are a high priority for 

increased investment include Shaba National Reserve and 

the Livestock Marketing Division and Kipsing areas of 

Isiolo. These latter two areas are important for retaining 

the migration link between the Laikipia and the northern subpopulations of Grevy’s zebra. They are also important for recovery or recolonization, as they were good 

habitat in the past. Additional high-priority areas around 

Laisamis, Serolivi, and the northernmost subpopulation 

near Baragoi are areas that were historically the core of 

the species range and where there is high potential for 

recovery.

Conservation is not possible in areas that are outside 

the historical range of the species, unsuitable habitat (e.g., 

mountains and forest), and agricultural land.

Species-Specific Priorities: Lions

For lions, current levels of conservation investment 

are sufficient in the commercial ranches of Laikipia and 

the national reserves and should continue in these areas. 

Additional investment is critically needed in the community areas of Laikipia (e.g., Naibunga), through the 

Kipsing and Livestock Marketing Division area of Isiolo 

(particularly along luggas, which lions favor as habitat), 

and in a buffer zone around the reserves (to allow for safe 

dispersal of lions out of the reserves). There also needs to 

be more investment in education and awareness in these 

community areas to maintain connectivity for lions in the 

landscape. The Matthews Range is also a high priority 

for increased investment, as this is potentially important 

habitat that could support more lions and has low human 

population density. Additional investment is also strongly 

needed in the area north of Ol Pejeta to create a link to the 

rest of Laikipia. Areas where there are many people and 

high densities of livestock are considered to be areas where 

either priority is low or conservation is not possible. The 

level of knowledge on lions in much of the north of the 

landscape was considered insufficient to make a decision 

on priorities. More research is needed in that area as there 

may be potential for recovering lions. 

In general, educational efforts aimed at community 

residents are needed to explain the rationale for and practice of predator conservation. People are receptive to the 

idea when presented in terms meaningful to their own 

lives, but it is not always easy to convince them that they 

can benefit from wildlife, especially predators on their livestock. Development of tourism in the communities, with 

TABLE 3. Summary of the feature-specific priority maps for each of four biodiversity features. Maps were 

produced by participants in the Ewaso Nyiro Landscape Conservation Planning Workshop (January 2006).

Focal Planning units Range needing High-priority units

biodiversity needing conservation increased needing increased

feature investment (%) investment a

 (%) investment (%)

African elephant 80.1 83.4 54.8

Grevy’s zebra 47.3 80.2 59.5

Lion 69.5 84.9 48.6

African wild dog 60.1 100 100

a

 Range here is defined as any of the 5 × 5 km planning units needing conservation of some kind (continued investment at current level or 

increased). This column reflects the percentage of those units that need increased investment (low, medium, or high priority or for more 

information).
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FIGURE 6. Conservation priorities across four species. The map was produced by combining priority maps for individual species 

and creating useful categories and is a product of the Ewaso Landscape Planning Workshop (January 2006).
118 • smithsonian contributions to zoology

transparent and equable distribution of profits, would 

provide some incentive for people to tolerate carnivores. 

Trophy hunting would have great potential in northern 

Laikipia, especially given the good numbers of leopard 

and great kudu. 

Species-Specific Priorities: Wild Dogs

For wild dogs, investment in conservation is insufficient across the entire landscape. Risk of infectious diseases is a major threat to wild dogs, with highest risk 

in community areas. The priority for conservation is to 

secure the known populations of wild dogs. Community 

areas along the Laikipia escarpment (e.g., Naibunga Conservancy) and into Isiolo District (with linkages through 

Isiolo into Samburu) are important habitat for wild dogs 

and a high priority for conservation needing additional investment. The Seiya Lugga, Matthews Range, and Ndotos 

Range are also high conservation priorities needing additional investment. The commercial ranches of Laikipia 

are considered medium priority. Some additional areas are 

considered medium priority and are important for providing connectivity. Southern Laikipia and Meru Districts are 

not considered suitable habitat for wild dogs and are unlikely to ever become a part of the wild dog range. Therefore, conservation is not possible in those areas.

Cross-Species Priorities

Overall, participants indicated that if current levels of 

conservation investment are maintained, only about 3.6% 

of the landscape is secure (i.e., does not need additional 

investment) for all four species. Approximately 4.6% of 

the landscape is secure for three or four species but often needs additional investment for wild dogs (diagonally 

hatched areas shown in Figure 6). These areas are the most 

secure areas in the landscape and include the commercial 

ranches and conservancies of central Laikipia, Lewa Wildlife Conservancy, a small area in Samburu around West 

Gate and Wamba, and Samburu and Buffalo Springs National Reserves. For the remainder of the paper, we will 

refer to these areas as the “conservation strongholds.” 

An additional 13.4% of the planning units are “partially” secure, meaning that levels of investment are sufficient for one or two species that require it but up to three 

species may need increased investment (areas with small 

circles in Figure 6). An excellent example of this type of 

area is Shaba National Reserve, which is secure for elephants and lions but is a high priority for increased investment for Grevy’s zebra and wild dogs.

The analysis concluded that fully 84% of the landscape needs continued or additional conservation investment at low to high priority. Almost half (45%) of the 

landscape was given high-priority status for continued or 

additional conservation investment for at least one species. About one-third of the landscape was given a highpriority status for at least one species and was identified 

as needing increased investment (Figure 6, small map). 

Specifically, this land includes all of the national reserves 

and land west; Solio; Laikipia ranches, Naibunga Conservancy, and parts of southwestern Laikipia; elephant 

corridors between Ngare Ndare Forest to Mount Kenya 

and Solio to Mount Kenya; much of Samburu, including 

the Kirisia Forest and the Matthews Range; Sera; and the 

northeastern part of the landscape in Marsabit District. 

About 1% of the landscape is at high priority for additional investment for all four species, and about 4% is 

at high priority for at least three species. Increased investment in these high-priority locations would benefit not 

only all or most of these species but also, by association, 

much of the biodiversity of the region. The majority of 

planning units in this category are located in either the 

“peninsula” of Isiolo District between Laikipia and Samburu Districts or the Matthews Range in Samburu. 

DISCUSSION

Implications for Conservation in the Ewaso Ecosystem

We suggest there are three important implications of 

this exercise for conservation practitioners working in the 

Ewaso ecosystem. These implications are likely to be similar for ecosystems throughout Kenya and eastern Africa.

1. Conservation objectives in the Ewaso Nyiro cannot be 

achieved by focusing only on traditional strongholds 

of biodiversity conservation. Planning and investment 

strategies are needed across much of the landscape.

The traditional strongholds of conservation in the 

Ewaso Nyiro include about 9,500 km2 of land (~18% 

of the landscape), composed mostly of the commercial 

ranches and conservancies of central Laikipia, Lewa Conservancy, and two of the three national reserves (Buffalo 

Springs and Samburu; see the areas in Figure 6 with diagonal hatch-marks). The critical question is, can we meet our 

objectives for biodiversity conservation (see Table 1) by 

focusing conservation only on the traditional strongholds? 

The four species on which our exercise focused are excellent examples of why the answer is a resounding no.
number 632 • 119

Although it may seem possible to reach modest conservation objectives, such as reaching and maintaining a 

population of 10,000 elephants within 10 years, by simply 

trying to increase the number of animals within the traditional strongholds, this supposition is not true for two 

main reasons. First, for all four species, substantial portions of their populations currently reside outside of the 

conservation strongholds. For example, significant subpopulations of Grevy’s zebra are resident in and around 

Laisamis in the northeast of the landscape and Wamba to 

the west of the Matthews Range, and lions, elephants, and 

wild dogs all have significant subpopulations in the Matthews Range (see the current importance maps in Figures 

1 and 2). If conservation investment is not increased in 

those areas, these subpopulations could decrease dramatically in the next 10 years because of changing land use 

patterns, poaching, settlement, and disease (see the future 

reductions maps). These factors would make it difficult to 

maintain current numbers, let alone increase populations 

to the conservation objectives.

The second reason that it will be impossible to reach 

conservation goals by focusing only on strongholds is that 

even subpopulations within the strongholds rely on habitat 

in the surrounding landscape. This dependence is particularly evident for the 8,000 or so elephants currently living 

in the landscape. This population relies on an enormous 

proportion of the total area (nearly 88% of the planning 

units), far beyond the strongholds, including a large number of corridors connecting various subpopulations (Figure 4). A failure to increase investment in areas beyond the 

strongholds will likely lead to declines in the elephant population. Further, increasing the population to 10,000 animals will require that habitat outside of the strongholds is 

used more intensively (see the potential for recovery maps 

in Figures 1 and 2), especially considering that elephant 

subpopulations in most of the strongholds are currently 

at or near carrying capacity. The same is more or less true 

for the other species considered in our exercise: although 

their populations spend more time within the strongholds, 

they rely heavily on habitat and corridors outside of those 

strongholds (from 29% to 47% of the planning units in 

the landscape) and will need to spend more time in this 

habitat if populations are to increase. 

In addition to the four species considered here, we 

suspect that conservation focused only on strongholds 

will be insufficient for protecting many of the other species, ecological processes (such as herbivore migrations), 

and ecosystem services of the Ewaso Nyiro. Gardner et al. 

(2007) recently demonstrated that strictly protected areas 

created for large mammals in Tanzania were ineffective 

at conserving other taxa (e.g., butterflies and birds). The 

same is likely true for the Ewaso Nyiro. Although we 

do not yet have conservation objectives or maps for the 

other focal biodiversity features chosen by workshop participants (Table 1), many certainly require conservation 

at landscape scales and beyond the current strongholds. 

For example, reticulated giraffe populations suffer from 

poaching, mostly for meat and mostly in areas outside of 

the conservation strongholds. Finally, the hydrological 

system, which provides a critical service to both the wildlife and humans living in the Ewaso Nyiro, obviously cannot be effectively conserved with activities focused within 

the traditional strongholds. In fact, one can make the case 

that although the Ewaso River itself is the most critical 

resource to wildlife residing in the strongholds, its flow 

and water quality can only be assured by effective water 

management and conservation upstream, in the agricultural areas and at its source in the forests of Mount Kenya, 

where logging inside and outside of the national park is a 

problem. 

2. Thus far, the systematic, landscape-scale process suggests three priorities for conservation investment, 

which are, in order of priority, continue conservation 

investments at current levels to preserve the strongholds, increase investments to prevent the loss of the 

Isiolo corridor, and increase investments to improve 

the status of biodiversity features in portions of Samburu District.

Because there is so much land in the Ewaso Nyiro 

that requires some level of conservation investment (about 

84% of the landscape) and it is certainly beyond current 

resources to work in all of that land, a process for prioritizing future investments is needed. Thus far, although the 

planning process has focused on only four of nine species, 

three clear priorities have emerged for how conservation 

practitioners should invest their resources in the future. 

Although other priorities exist on an individual species 

basis (e.g., Kirisia Forest is a high priority for increased 

investment for elephants), the priorities below represent 

ones where the combined energy of the conservation 

community can be marshaled to benefit a large range of 

biodiversity. 

First, the planning process clearly demonstrates the 

importance of continuing the levels of investment currently being directed at conservation strongholds. With 

the possible exception of elephants, the majority of the 

populations of the four focal species reside within the 

conservation strongholds. Further, the potential for 
120 • smithsonian contributions to zoology

future reductions maps (Figures 1, 2) indicate that in the 

absence of current levels of investment by conservation 

practitioners, large reductions in the subpopulations residing on the strongholds would likely occur. Although 

the land use practices on many of the privately owned 

ranches and conservancies in Laikipia would probably 

remain wildlife friendly in absence of active conservation and research (assuming ownership did not change), 

the national reserves would likely lose their integrity. 

Continuing investment in active conservation on strongholds will ensure that subpopulations persist and, often, 

increase. In some cases, additional investment in the 

strongholds is needed in order to preserve and recover 

wild dogs, although these investments are often a lower 

priority for dogs than those needed outside of the strongholds (Il Ngewsi, Loisaba Conservancy, and the national 

reserves are a high priority for increased investment for 

wild dogs). The resources currently being invested in the 

strongholds should in no way be redirected toward the 

other priorities below.

Second, in the near future, practitioners should raise 

and invest resources to secure, as continued wildlife habitat, the Isiolo corridor: the peninsula of Isiolo District 

separating Laikipia and Samburu Districts. Increased 

investment is needed in several critical locations. Most 

important is a stretch of land in the center of the Livestock Marketing Division, along the Ewaso Nyiro River, 

which separates northeastern Laikipia (Il Ngwesi, Lukurruki Ranch, Mukogodo Forest, etc.) from Buffalo Springs 

National Reserve (Figure 6). Increased investment in this 

area is a high priority for all four species considered in our 

exercise and would certainly benefit the species by both 

preventing future declines of subpopulations using the region and helping to increase subpopulations. Most importantly, however, investments could help secure the area as 

a corridor between the strongholds of central Laikipia and 

the national reserves (and by extension, much of northern 

Samburu, including the Matthews Range). 

Increased investment should also be focused on other 

portions of Isiolo District, including Kipsing and Ol DoNgiro, especially the northern portions of these areas bordering Samburu, which are important for three species. 

Similarly, the group ranches along the Laikipia escarpment bordering Isiolo (e.g., Naibunga Ranch) and Lekurruki Ranch need increased investment. 

Concerns about the future of the Isiolo corridor center around (1) increased interest of local communities 

and government in livestock marketing and the possibility of creating an abattoir with extensive livestock holding grounds and (2) the lack of interest or perceived value 

by local communities in wildlife. Any development that 

did not consider wildlife would certainly be injurious to 

conservation since this vital wildlife corridor could be permanently closed. Engagement particularly with the Isiolo 

County Council along with the councils in Samburu and 

Laikipia is critical for success in the corridor, as these 

councils hold trust over most of the land. 

Finally, increased investment should be aimed at several portions of Samburu District. In particular, Matthews 

Range and surrounding areas represent an important forest and thicket habitat for elephants, wild dogs, and lions 

(but not for Grevy’s zebra, which are restricted to more 

open habitats). Increased investment should be aimed at 

securing this area as a vital “island” that would create a 

“source” for recovering populations across the northern 

half of the landscape. Other important areas for increased 

attention in Samburu include the following. The habitat in 

and around Sera Conservancy in western Samburu could 

help increase subpopulations of both Grevy’s zebra and 

elephants in addition to the two carnivores. Namunyak 

and Kalama conservancies, although they are secure for 

the herbivores, need increased investment for wild dogs 

and lions, and they form an important link between the 

national reserves and the Matthews Range. Finally, the 

group ranches south of Kirisia Forest are also important 

for both wild dogs and elephants as linkages from Laikipia 

to the Kirisia Forest and areas east. 

3. Some conservation objectives will conflict with human livelihood objectives. Part of a comprehensive 

landscape-scale planning process needs to be dedicated to negotiating these conflicts so that all objectives can be met to the greatest extent possible.

Assessment of the landscape-scale priorities for our 

four focal species highlighted a challenge of conservation 

in Ewaso Nyiro: it may be difficult to meet all conservation objectives and human livelihood development objectives on the same areas of land. For example, it will be 

difficult to maintain or increase livestock grazing and marketing on lands also important for meeting conservation 

objectives for conserving predators and large herbivores. 

We believe that, ultimately, conservation planning in the 

Ewaso landscape needs to evolve into comprehensive land 

use planning that incorporates both conservation and 

development objectives and negotiates trade-offs. Sometimes, the objectives of both may be partially met on the 

same lands (e.g., wildlife tourism), but often, this may not 

be the case, and priority may need to be given to one objective at the expense of another. 
number 632 • 121

Next Steps for Systematic, Landscape-­Scale

Planning in the Ewaso Nyiro

Although this exercise proved valuable for clarifying 

the scope of the conservation challenge in the Ewaso Nyiro, compiling useful information, and prioritizing some 

areas for conservation investment, it was only partially 

complete. We recommend that landscape-scale planning, 

as we have demonstrated here, continue and, in fact, 

become an integral part of conservation thinking in the 

Ewaso Nyiro. The process should be revisited and updated 

over time, possibly once every two to five years, depending 

on the changing conservation and development context. 

If substantial progress toward conserving biodiversity is 

made (e.g., the Isiolo corridor is secured), priorities should 

be reevaluated. The same should happen in the case of 

failure. We recommend that an inclusive and collaborative group of stakeholders concerned with landscapescale conservation and development, such as the recently 

formed Ewaso Conservation Group, take a leadership role 

in carrying the process forward. Important next steps for 

the planning process include the following:

1. Review the quantitative objectives for the elephants, 

Grevy’s zebra, lions, and wild dogs (Table 1) and incorporate these more explicitly into the spatially explicit priority setting.

The objectives shown in Table 1 represent preliminary 

and short-term (over the next 10 years) conservation objectives. A more formal and rigorous process for establishing quantitative objectives would be useful, such as 

those described in Groves (2003), Tear et al. (2005), and 

Sanderson (2006). Several possible quantitative objectives 

should be considered (e.g., minimum viable populations, 

ecologically functional populations, sustainable yield objectives, historical levels, several redundant populations), 

as this will help stakeholders set near- and long-term objectives and consider what is truly necessary for long-term 

conservation success (e.g., are 20 packs of wild dogs a sufficient long-term goal?). These objectives then need to be 

compared on a quantitative basis with the maps produced 

in this exercise to guide decisions about how much land is 

needed to reach the objectives (e.g., could areas marked 

as a high priority support sufficient number of animals to 

reach the objectives?) (Rodrigues and Gaston, 2001; Boshoff et al., 2002). If maps can be expressed in the same 

units as the quantitative targets (e.g., abundance), indices 

that are useful for prioritizing areas can be produced, including irreplaceability and marginal benefit:cost ratios 

(Cawardine et al., 2006), and “optimal” networks of conservation areas that meet objectives for all focal targets 

can also be produced.

2. Complete the planning steps for the remaining focal 

conservation features.

Thus far, the planning process has proceeded while 

considering only four of nine focal conservation features. 

Workshop participants agreed that planning based on all 

nine features was necessary to ensure that other forms of 

biodiversity, critical ecological functions (e.g., herbivore 

migration), and services (e.g., production of clean water) were well protected. Therefore, the planning process 

(especially steps 4–6; see Box 1) should include consideration of the remaining five features (i.e., reticulated giraffe, Jackson’s Hartebeest, acacia-grassland mosaic, dry 

upland/montane forest, and the hydrological system) and 

possibly examine how well these surrogates represent a 

wider range of taxa (Gardner et al., 2007). The process 

of setting quantitative targets, producing the component 

maps (e.g., current importance), and prioritizing actions 

for each feature doubtlessly would be enlightening and 

may significantly shift near-term priorities. 

3. Complete the planning steps for livelihood development goals and objectives and negotiate land use solutions that meet both conservation and development 

objectives.

As we noted above, we do not believe a landscapescale plan can be successfully implemented if it has not also 

incorporated, to some degree, objectives for human livelihood development and attempted to negotiate a network 

of land uses that meet both conservation and development 

objectives (Faith and Walker, 1996; Stewart and Possingham, 2005). If both sets of objectives are not considered 

in a transparent and relatively objective process, conservation priorities run the danger of being largely irrelevant in 

future land use decisions. Local stakeholders interested in 

pursuing development objectives, such as county councils 

and development nongovernmental organizations, need 

to be given at least equal time, so that they buy into the 

process as a whole and do not simply disregard it as an 

attempt to usurp their rights to the land. Development objectives should not be treated as “opportunity costs” of 

conservation (as in Stewart and Possingham, 2005) but, 

rather, as true objectives on an equal level with conservation objectives. In this way true conflicts, where conservation and development objectives are incompatible, as well 
122 • smithsonian contributions to zoology

as unique opportunities to meet both sets of objectives can 

be identified.

We recommend that first, conservation and development priorities be assessed separately, possibly in separate 

workshops, but using the same framework (such as the 

one in Box 1). The framework should ask participants to 

express their development objectives in quantitative and 

measurable terms and to consider the benefits and costs 

of various areas for meeting those objectives (e.g., what 

are the benefits and costs of increasing livestock grazing 

in particular areas). After independently setting priorities, development objectives and conservation objectives 

should be brought together, and solutions for meeting 

both negotiated. Producing future conservation and development scenarios (e.g., Baker et al., 2004; Hulse, 2004) 

may be a useful way for exploring options and their consequences. In general, land use planning processes that 

explicitly incorporate both conservation and development 

objectives are in their infancy, but the Ewaso Nyiro could 

easily lead in this arena. 

CONCLUSIONS

In many places in eastern Africa and, indeed, the world 

it is becoming apparent that the traditional approach of 

conservation focusing on protected-area creation and 

management, though essential, will likely be insufficient 

to meet conservation goals. Many species, especially the 

area-demanding ones that are so common in eastern Africa, and ecosystem services simply cannot be effectively 

conserved on small patches of land. Conservation practitioners need to increase their effort beyond the traditional 

conservation strongholds to the surrounding landscape 

dominated by human land uses. Unfortunately, taking 

conservation to sufficiently large scales is often a daunting undertaking, as was demonstrated in our exercise 

(84% of the landscape needs conservation investment), 

and requires conservationists to work with a huge range 

of landowners and stakeholders who may not be primarily 

concerned with conservation. Systematic, landscape-scale 

conservation planning approaches help practitioners to set 

ecologically meaningful and transparent objectives, to define the scale at which they need to work to meet these objectives, and to define priorities that help ensure that they 

invest their limited resources to make the most efficient 

progress toward their daunting objectives. The planning 

approach we have demonstrated here is an important step 

for scaling up conservation in the Ewaso Nyiro and setting clear conservation priorities for immediate action. It 

has produced concrete, near-term priorities for investment 

(e.g., secure the Isiolo corridor) that conservation practitioners in the region can work together to meet. It has 

catalyzed the creation of the Ewaso Conservation Group, 

a partnership of organizations and local governments 

that is charged with carrying the planning and prioritizing process forward and implementing actions based on 

the process. We encourage all those interested in creating 

a landscape that meets the needs of people and wildlife 

to help the Ewaso Conservation Group carry the process 

forward in the Ewaso Nyiro and to other areas of Kenya, 

such as Tsavo or the Mara, which face similar challenges. 

Acknowledgments

The authors thank E. McBean for her financial support of this project. We also thank all those who contributed time and information to the workshop and its 

results, including Anthony King, Shivani Bhalla, Geoffrey 

Chege, Ian Craig, Richard Hatfield, Onesmas Kahindi, 

Fred Kihara, Dominic Kilonzo, Anthony Leaduma, Moses 

Lenairoshi, Philip Lenaiyasa, Fabian Lolosoli, Belinda 

Low, Jonathan Moss, Philip Muruthi, Josephat Musyima, James Munyugi, Wycliffe Mutero, Kierna Mwandia, 

Chris Odhiambo, Nick Oguge, David Parkinson, Leslie 

Scott, Chris Thouless, Fritz Vollrath, Stuart Williams, and 

Phillip Winter. Finally, for logistical support, we thank 

K. Outram and the staff of the Mpala Research Station; 

J. Deutsch, M. Wrobel, and the staff of the Wildlife Conservation Society Africa Program; and D. Kelly.

Expanded Author Information

Karl A. Didier, Global Conservation Programs, Wildlife Conservation Society, 907 NW 14th Avenue, Gainesville, Florida 32601, USA. Alayne Cotterill, Living with 

Lions and the Wildlife Conservation Research Unit, University of Oxford, P.O. Box 555, 10400 Nanyuki, Kenya. 

Iain Douglas-Hamilton, Save the Elephants and Department of Zoology, University of Oxford, Post Office Box 

54667, 00200 Nairobi, Kenya. Laurence Frank, Living 

with Lions, Panthera, 8 West 40th Street, 18th Floor, New 

York, New York 10018, USA; and Museum of Vertebrate 

Zoology, University of California, Berkeley, California 

94720, USA. Nicholas Georgiadis, Property and Environment Research Center, 2048 Analysis Drive, Suite A, 

Bozeman, Montana 59718, USA. Max Graham, Department of Geography, University of Cambridge, Cambridge 

CB2 3EN, UK. Festus Ihwagi, Save the Elephants, Post 

Office Box 54667, 00200 Nairobi, Kenya. Juliet King, 
number 632 • 123

Northern Rangelands Trust, Private Bag, Isiolo, Kenya. 

Delphine Malleret-King, Laikipia Wildlife Forum, Post 

Office Box 764, 10400 Nanyuki, Kenya. Dan Rubenstein, 

Department of Ecology and Evolutionary Biology, Princeton University, Princeton, New Jersey 08544, USA. David 

Wilkie, Global Conservation Programs, Wildlife Conservation Society, 18 Clark Lane, Waltham, Massachusetts 

02451, USA. Rosie Woodroffe, Department of Wildlife, 

Fish, and Conservation Biology, University of California, 

Davis, California 95616, USA; present address: Institute 

of Zoology, Zoological Society of London, Regent’s Park, 

London NW1 4RY, UK.

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